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 

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Chlorine/Ammonia on Ecosystem Structure and Function in 

Experimental Streams. Report to the Qffice of Water, U.S. 

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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. 

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1983). 

16. Wang, W. 1990. Toxicity Assessment of Pretreated Industrial 

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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 
--- - - ____- 

3.1.3 General Consiftt?mtions in fffhmt Characterization 

There are two possible ways to characterize an effluent to determine the need for effluent limits for the protection of aquatic life 

and human health. First, an assessment may be made without 

generating effluent data; second, an assessment may be conducted after effluent data have been generated. Regulatory authorities must determine whether a discharge causes, has the 

“reasonable potential” lo cause, or contributes to an excursion 

above an applicable narrative or numeric water quality criterion. 

An analysis of “reasonable potential” determines an effluent’s 

capability to cause such excursions. 

In determining the need for a permit limit for whole effluent 

toxicity or for an individual toxicant, the regulatory authority is 

required to consider, at a minimum, existing controls on point and 

nonpoint sources of pollution, the variability of the pollutant or 

pollutant parameter in the effluent, the sensitivity of the involved 

species to toxicity testing (for whole effluent), and, where appropriate, the dilution of the effluent in the receiving water (40 CfR 

122.44(d)(ii)). 

The regulatory authority is also required by NPDES regulations to 

consider whether technology-based limits are sufficient to maintain State water quality standards. There are two possibilities that 

will need to be assessed. First, if the limits based on appropriate 

treatment technology have already been specified in a previous 

permit, and if the facility is operating at the required level, then 

historical effluent and receiving water information can be used. 

Second, if the facility has yet to achieve the required technology 

performance (best available technology or best conventional technology), the regulatory authority will need to assess the technology-based limit for reasonable potential for causing or contributing to an excursion above the water quality standard. 

In addition, the regulatory authority should consider all other 

available data and information pertaining to the discharger to 

assist in making an informed judgment. Where both effluent 

testing data and important other factors exist, the regulatory 

authority will need to exercise discretion in the determination of 

the need for a limit. The authority should employ the principle of “independent application” of the data and information that characterizes the effluent. In other words, effluent 

data alone, showing toxicity at the RWC, may be adequate to 

demonstrate the need for a limit for toxicity or for individual 

toxicants. Likewise, other factors may form an adequate basis for 

determining that limits are necessary. for example, where available dilution is low and monitoring information shows that toxic 

pollutants are frequently discharged at concentrations that have 

caused toxicity when discharged from similar facilities, the permitting authority may reason that a whole effluent toxicity limit is 

necessary even without whole effluent toxicity data from the 

specific facility. In all cases, the decision must be based upon 

consideration of factors cited in 40 CfR 122.44(d)(l)(ii). The 

regulatory authority will need to prioritize, on a case-by-case 

basis, the importance of all data and information used in making 

a determination. To assist in case-by-case determinations, recommended guidelines for characterizing an effluent for the need 

for a permit limit for whole effluent toxicity or individual toxicants are discussed below and summarized in Boxes 3-l through 

3-3. 

Box 3-1. Determining “Reasonable Potential” for Excursions Above Ambient Criteria Using 

Factors Other than Facility-specific Effluent Monitoring Data 

When determining the “reasonable potential” of a discharge to cause an excursion above a State water quality 

standard, the regulatory authority must consider all the factors listed in 40 CFR 122.44(d)(l)(ii). Examples of the 

types of information relating to these factors are listed below. 

Existina controls on mint and nonDoint sources of Dollution 

l Industry type: Primary, secondary, raw materials used, products produced, best management practices, 

control equipment, treatment efficiency, etc. 

l Publicly owned treatment work type: Pretreatment, industrial loadings, number of taps, unit processes, 

treatment efficiencies, chlorination/ammonia problems, etc. 

Variability of the oollutant or oollutant Darameter in the effluent 

l Compliance history 

l Existing chemical data from discharge monitoring reports and applications. 

Sensitivitv of the sDecies to toxicitv testing 

l Adopted State water quality criteria, or EPA criteria 

l Any available in-stream survey data applied under independent application of water quality standards 

l Receiving water type and designated/existing uses 

Dilution of the effluent in the receivina water 

l Dilution calculations 

49 
3.2 DETERMlWlNG THE NEED FOR PERYlT UMITS 

WITHOUT EFFLUENT MONITORING DATA FOR A 

SPECIFIC FACILITY 

If the regulatory authority so chooses, or if the circumstances 

dictate, the authority may decide to develop and impose a 

permit limit for whole effluent toxicity or for individual toxicants 

without facility-specific effluent monitoring data, or prior to the 

generation of effluent data. Water quality-based permit limits 

can be set for a single toxicant or for whole effluent toxicity based 

on the available dilution and the water quality criterion or the 

State standard in the absence of facility specific effluent monitoring data. However, in doing so, the regulatory authority must 

satisfy all the requirements of 40 CfR 122.44(d)(l)(ii). 

When determining whether or not a discharge causes, has the 

reasonable potential to cause, or contributes to an excursion of a 

numeric or narrative water quality criterion for individual toxicants or for toxicity, the regulatory authority can use a variety of 

factors and information where facility-specific effluent monitoring data are unavailable. These factors also should be considered 

with available effluent monitoring data. Some of these factors are 

the following: 

l Dilution-Toxic impact is directly related to available dilution for the effluent. Dilution is related to the receiving 

stream flow and the size of the discharge. The lower the 

available dilution, the higher the potential for toxic effect. 

If an effluent’s concentration at the edge of a mixing zone 

in a receiving water is expected to reach 1 percent or 

higher during critical or worst-case design periods, then 

such an effluent may require a toxicity limit (see discussion 

in Section 3.3.3). Assessment of the amount of stream 

dilution available should be made at the conditions required by the water quality standards or, if not specified in 

the standards, at the harmonic mean flow and the 7QlO 

flow. Figure 3-3 (Pg. 57) shows that, whereas a majority of 

NPDES permittees nationwide discharge to areas during 

annual mean flow ranging in dilution from 100 to 1,000, 

the majority of dischargers fall into the 1 to 10 dilution 

range during low-flow conditions. 

l Type of industry-Although dischargers should be individually characterized because toxicity problems are sitespecific, the primary industrial categories should be of 

principal toxicity concern. EPA’s treatment technology 

data base generally suggests that secondary industrial categories may have less potential for toxicity than primary 

industries. However, based on experience, it is virtually 

impossible to generalize the toxicity of effluents with any 

certainty. If two plants produce the same type of product, 

one effluent may be toxic while the other may not be toxic 

due to the type and efficiency of the treatment applied, 

general materials handling practices, and the functional 

target of the compound(s) being produced. 

. Type of POTW-POTWs with loadings from indirect dischargers (particularly primary industries) may be candidates for toxicity limits. However, absence of industrial 

input does not guarantee an absence of POTW discharge 

toxicity problems. For example, commercial pesticide applicators often discharge to POT%%, resulting in pesticide 

concentrations in the POTW’s effluent. Household disposal 

of pesticides, detergents, or other toxics may have a similar 

effect, The types of industrial users, their product lines, their 

raw materials, their potential and actual discharges, and 

their control equipment should be evaluated. POlWs should 

also be characterized for the possibility of chlorine and 

ammonia problems. 

l Existing data on toxic pollutants-Discharge monitoring 

reports (DMRs) and data from NPDES permit application 

forms 2C and 2A may provide some indication of the presence of toxicants. The presence or absence of the 126 

“priority pollutants” may or may not be an indication of the 

presence or absence of toxicity. There are thousands of 

“nonpriority” toxicants that may cause effluent toxicity. 

Also, combinations of several toxicants can produce ambient toxicity where the individual toxicants would not. EPA 

regulations at 40 CFR 122.21 (j) require POTWs with design 

flows equal to or greater than 1 MC0 and POTWs with 

approved pretreatment programs, or POlWs required to 

develop a pretreatment program, to submit the results of 

whole effluent toxicity tests with their permit applications. 

These regulations also provide discretion to the permitting 

authority to request such data from other POlWs at the 

time of permit application. 

l History of compliance problems and toxic impact-Reguiatory authorities may consider particular dischargers that 

have had difficulty complying with limits on toxicants or 

that have a history of known toxicity impacts as probable 

priority candidates for effluent toxicity limits. 

l Type of receiving water and designated use-Regulatory 

authorities may compile data on water quality. Examples of 

available data include fish advisories or bans, reports of fish 

kilts, State lists of priority waterbodies, and State lists of 

waters that are not meeting water quality standards. Regulatory authorities should use this information as a means of 

identifying point sources that discharge to impaired 

waterbodies and that thus may be contributing to this 

impairment. One source of this information is the lists of 

waters generated by states to comply with Section 304(l) 

regulations at 40 CFR 130.1 O(d)(6); 50 FR 23897-98, June 2, 

1989: 

1) Waters where fishing or shellfish bans and/or 

advisories are currently in effect or are anticipated; 

2) Waters where there have been repeated fish 

kills or where abnormalities (cancers, lesions, 

tumors, etc.) have been observed in fish or 

other aquatic life during the last ten years; 

3) Waters where there are restrictions on water 

sports or recreational contact; 

4) Waters identified by the state in its most recent state section 305(b) report as either “partially achieving” or “not achieving” designated 

uses; 

so 
5) 

6) 

7) 

8) 

9) 

10) 

11) 

12) 

13) 

Waters identified by the states under section 

303(d) of the Clean Water Act as waters needing water quality-based controls; 

Waters identified by the state as priority water 

bodies; 

Waters where ambient data indicate potential 

or actual excursions of water quality criteria 

due to toxic pollutants from an industry classified as a primary industry in Appendix A of 40 

CFR Part 122; 

Waters for which effluent toxicity test results 

indicate possible or actual excursions of state 

water quality standards, including narrative 

“free from” water quality criteria or EPA water 

quality criteria where state criteria are not available; 

Waters with primary industrial major dischargers where dilution analyses indicate 

exceedances of state narrative or numeric water quality criteria (or EPA water quality criteria 

where state standards are not available) fortoxic 

pollutants, ammonia, or chlorine; 

Waters with POlW dischargers requiring local 

pretreatment programs where dilution analyses indicate exceedances of state water quality 

criteria (or EPA water quality criteria where 

state water quality criteria are not available) 

for toxic pollutants, ammonia, or chlorine; 

Waters with facilities not included in the previous two categories such as major POlWs, and 

industrial minor dischargers where dilution 

analyses indicate exceedances of numeric or 

narrative state water quality criteria (or EPA 

water quality criteria where state water quality 

criteria are not available) for toxic pollutants, 

ammonia, or chlorine; 

Water classified for uses that will not support 

the “fishable/swimmable” goals of the Clean 

Water Act; 

Waters where ambient toxicity or adverse water quality conditions have been reported by 

local, state, EPA or other Federal Agencies, the 

private sector, public interest groups, or universities; 

14) Waters identified by the state as impaired in its 

most recent Clean Lake Assessments conducted 

under 314 of the Clean Water Act; and 

15) Surface waters impaired by pollutants from 

hazardous waste sites on the National Priority 

List prepared under section 105(8)(A) of 

CERCIA. 

16) Waters judged to be impaired as a result of a 

bioassessmenttbiosurvey. 

The presence of a combination of these factors, such as low 

available dilution, high-quality receiving water, poor compliance record, and clustered industrial and municipal discharges, 

could constitute a high priority for effluent limits. 

Regardless, the regulatory authority, if it chooses to impose an 

effluent limit after conducting an effluent assessment without 

facility-specific monitoring data, will need to provide adequate 

justification for the limit in its permit development rationale or 

in its permit fact sheet. A clear and logical rationale for the need 

for the limit covering all of the regulatory points will be necessary to defend the limit should it be challenged. In justification 

of a limit, EPA recommends that the more information the 

authority can acquire to support the limit, the better a 

position the authority will be in to defend the limit if necessary. In such a case, the regulatory authority may well benefit 

from the collection of effluent monitoring data prior to establishing the limit. 

If the regulatory authority, after evaluating all available information on the effluent, in the absence of effluent monitoring data, 

is not able to decide whether the discharge causes, has the 

reasonable potential to cause, or contributes to, an excursion 

above a numeric or narrative criterion for whole effluent toxicity 

or for individual toxicants, the authority should require whole 

effluent toxicity or chemical-specific testing to gather further 

evidence. In such a case, the regulatory authority can require 

the monitoring prior to permit issuance, if sufficient time exists, 

or it may require the testing as a condition of the issued/ 

reissued permit. 

Under these circumstances, the regulatory authority may find it 

protective of water quality to include a permit reopener for the 

imposition of an effluent limit should the effluent testing establish that the discharge causes, has the reasonable potential to 

cause, or contributes to excursion above a water quality criteria. 

A discussion of these options is provided later in this chapter. 

3.3 DETERMINING THE NEED FOR PERMIT 

UMITS WITH EFFLUENT MONITORING DATA 

3.3. I General Conslderatlons 

When characterizing an effluent for the need for a whole effluent toxicity limit, and/or an individual toxicant limit, the regulatory authority should use any available effluent monitoring 

data, together with any information like that discussed under 

Section 3.2 above, as the basis for a decision. The regulatory 

authority may already have effluent toxicity data available from 

previous monitoring, or it may decide to require the permittee 

to generate effluent monitoring data prior to permit issuance or 

as a condition of the issued permit. EPA regulations at 40 Cfi? 

122.21 (j) require POTWs with design flows equal to or greater 

than 1 MCD and POTWs with approved pretreatment programs, or POTWs required to develop a pretreatment program, 

to submit the results of whole effluent toxicity tests with their 

permit applications. These regulations also provide discretion 

to the permitting authority to request such data from additional 

POTWs at the time of permit application. 

51 
----------- __- __-__ ---- __--_ 

In the instance where the permittee is required to generate data in 

advance, data collection should begin 12 to 18 months in advance 

of permit development to allow adequate time for conducting 

toxicity tests and chemical analyses. The type of data, including 

toxicity testing data, should be specified by the regulatory authority at the outset so that decisions on permit actions will not be 

delayed. EPA recommends monitoring data be generated on 

effluent toxicity prior to permit limit development for the 

foIlowing reasons: (1) the presence or absence of effluent 

toxicity can be more clearly established or refuted and (2) 

where toxicity is shown, effluent variability can be more cleady 

defined. Several basic factors that should be considered in generating effluent monitoring data are discussed below. 

All toxic effects testing and exposure assessment parameters, for 

both effluent toxicity and individual chemicals, have some degree 

of uncertainty associated with them. The more limited the amount 

o! test data available, the larger the uncertainty. The least amount 

of uncertainty of an effluent’s impact on the receiving water exists 

where (1) a complete data base is available on the effects of acute 

and chronic toxicity on many indigenous species, (2) there is a 

clear understanding of ecosystem species composition and functional processes, and (3) actual measured exposure concentrations 

are available for all chemicals during seasonal changes and dilution 

situations. The uncertainty associated with such an ideal situation 

would be minimal. However, generation of these data can be very 

resource intensive. 

An example of uncertainty that results from limited monitoring 

data is if a regulatory authority has only one piece of effluent data 

(e.g., an LCso of 50 percent) for a facility. Effluent variability in 

such a case, given the range of effluent toxicity variability seen in 

other effluents, may range between 20 percent and 100 percent 

(see Appendix A). It is impossible to determine from one piece of 

monitoring data where in this range the effluent variability really 

falls. More monitoring data would need to be generated to 

determine the actual variability of this effluent and reduce this 

source of uncertainty. 

To better characterize the effects of effluent variability and reduce 

uncertainty in the process of deciding whether to require an 

effluent limit, EPA has developed the statistical approach described 

below and in Box 3-2. This approach combines knowledge of 

effluent variability as estimated by a coefficient of variation with 

the uncertainty due to a limited number of data to project an 

estimated maximum concentration for the effluent. The estimated 

maximum concentration is calculated as the upper bound of the 

expected lognormal distribution of effluent concentrations at a 

high confidence level. The projected effluent concentration after 

consideration of dilution can then be compared to an appropriate 

water quality criterion to determine the potential for exceeding 

that criterion and the need for an effluent limit. 

The statistical approach has two parts. The first is a characterization of the highest measured effluent concentration based on the 

desired confidence level. The relationship that describes this is the 

following: 

pn = (1 - confidence level)“” 

where pn is the percentile represented by the highest concentration in the data and n is the number of samples. The 

following are some examples of this relationship at a 99 

percent confidence level: 

l The largest value of 5 samples is greater than the 40 

percentile 

l The largest value of 10 samples is greater than the 63 

percentile 

l The largest value of 20 samples is greater than the 79 

percentile 

l The largest value of 100 samples is greater than the 96 

percentile. 

The second part of the statistical approach is a relationship 

between the percentile described above and the selected 

upper bound of the lognormal effluent distribution. EPA’s 

effluent data base suggests that the lognormal distribution 

well characterizes effluent concentrations (see Appendix E). 

For example, if five samples were collected (which represents a 40th percentile), the coefficient of variation is 0.6, 

and the desired upper bound of the effluent distribution is 

the 99th percentile, then the two percentiles can be related 

using the coefficient of variation (CV) as shown below: 

C99 exp(2.326o - 0.50~) 

= = 4.2 

c40 exp(-0.258o - 0.50~) 

where & = In ((X2+1) and 2.326 and -0.258 are the normal 

distribution values for the 99th and 40th percentiles, respectively. The use of the 99th percentile is for illustrative 

purposes here. Although it does represent a measure of the 

upper bound of an effluent distribution, other percentiles 

could be selected by a regulatory agency. The relationship 

shown above can be calculated for other percentiles and 

CVs by replacing the values in the equation. 

Tables 3-l and 3-2 show the combined effects of both parts 

for a 99percent confidence level and upper bounds of the 

99th and 95th percentiles, respectively. The factors shown 

in the tables are multiplied by the highest concentration in 

an effluent sample to estimate the maximum expected concentration. 

This procedure can be used for both single and multiple 

discharges to the same receiving waterbody. This is accomplished for multiple dischargers by summing the projected 

RWCs for the pollutant or pollutant parameter of concern 

from each individual discharger, and comparing it to the 

water quality standard. This involves an assumption of 

conservative additivity of the pollutant after discharge, which 

may not accurately reflect the true behavior of the toxicant. 

To overcome this, and to further refine the proportional 

contribution of each discharger and the resultant limits, the 

permitting authority should supplement this evaluation with 

multiple source WLA modeling and/or ambient water concentration monitoring. 

52 

A 
Box 3-2. Determining “Reasonable Potential” for Excursions Above 

Ambient Criteria Using Effluent Data Only 

EPA recommends finding that a permittee has “reasonable potential” to exceed a receiving water quality 

standard if it cannot be demonstrated with a high confidence level that the upper bound of the lognormal 

distribution of effluent concentrations is below the receiving water criteria at specified low-flow conditions. 

Step 1 Determine the number of total observations (“n”) for a particular set of effluent data (concentrations or 

toxic units PUS]), and determine the highest value from that data set. 

Step 2 Determine the coefficient of variation for the data set. For a data set where ~10, the coefficient of 

variation (CVJ is estimated to equal 0.6, or the CV is calculated from data obtained from a discharger. 

For a data set where n>lO, the CV is calculated as standard deviation/mean (see Figure 3-l). For less 

than 10 items of data, the uncertainty in the CV is too large to calculate a standard deviation or mean 

with sufficient confidence. 

Step 3 Determine the appropriate ratio from Table 3-l or 3-2. 

Step 4 Multiply the highest value from a data set by the value from Table 3-l or 3-2. Use this value with the 

appropriate dilution to project a maximum receiving water concentration (RWC). 

Step 5 Compare the projected maximum RWC to the applicable standard (criteria maximum concentration, 

criteria continuous concentration [CCC], or reference ambient concentration). EPA recommends that 

permitting authorities find reasonable potential when the projected RWC is greater than an ambient 

criterion. 

Example 

Consider the following results of toxicity measurements of an effluent that is being characterized: 5 TU,, 2 TU,, 9 TU,, 

and 6 TU,. Assume that the effluent is diluted to 2 percent at the edge of the mixing zone. Further assume that the 

CV is 0.6, the upper bound of the effluent distribution is the 99th percentile, and the confidence level is 99 percent. 

Step 1 There are four samples, and the maximum value of the sample results is 9 TU,. 

Step 2 The value of the CV is 0.6. 

Step 3 The value of the ratio for four pieces of data and a CV of 0.6 is 4.7. 

Step 4 The value that exceeds the 99th percentile of the distribution (ratio times xmax) after dilution is calculated as: 

[9 TU, x 4.7 x 0.021 = 0.85 TU,. 

Step 5 0.85 TU, is less than the ambient criteria concentration of 1 .O TU,. There is no reasonable 

potential for this effluent to cause an excursion above the CCC. 

3.3.3 Etfhnl CharacMzatiun tar Whh Efh?nt T~xWty 

Once an effluent has been selected for whole effluent toxicity 

characterization after consideration of the factors discussed above, 

the regulatory authority should require toxicity testing in accordance with appropriate site-specific considerations and the recommendations discussed below. In the past 5 years, significant 

additional experience has been gained in generating effluent 

toxicity data upon which to make decisions as to whether or not 

an effluent will cause toxic effects in the receiving water in both 

freshwater and marine environments. 

General Considerations and Assumptions 

EPA has revised its initial effluent toxicity data generation recommendations based on three observations made over the last 5 

years: 

1) Only rarely have effluents discharged by NPDES permittees 

been observed to have LCsos less than 1 .O percent or no 

observed effect concentrations (NOECs) less than 0.1 percent. However, there is always a chance that an effluent 

could be toxic at such low effluent concentrations. 

53 
-_-___- 

-- - -- -.- ------ 

Jumber o 

Samples 

1 

2 

3 

4 

5 

6 

7 

a 

9 

10 

11 

12 

13 I 

! :: 

16 I 

17 

18 

I 19 

20 

f 

I 

- 

Table 3-l. Reasonable Potential Multiplying Factors: 99% Confidence Level and 99% Probability Basis 

0.1 0.2 0.3 0.4 0.5 

1.6 2.5 3.9 6.0 9.0 

1.4 2.0 2.9 4.0 5.5 

1.4 1.9 2.5 3.3 4.4 

1.3 1.7 2.3 2.9 3.8 

1.3 1.7 2.1 2.7 3.4 

1.3 1.6 2.0 2.5 3.1 

1.3 1.6 2.0 2.4 2.9 

1.2 1.5 1.9 2.3 2.8 

1.2 1.5 1.8 2.2 2.7 

1.2 1.5 1.8 2.2 2.6 

1.2 1.5 1.8 2.1 2.5 

1.2 1.4 1.7 2.0 2.4 

1.2 1.4 1.7 2.0 2.3 

1.2 1.4 1.7 2.0 2.3 

1.2 1.4 1.6 1.9 2.2 

1.2 1.4 1.6 1.9 2.2 

1.2 1.4 1.6 1.9 2.1 

1.2 1.4 1.6 1.8 2.1 

1.2 1.4 1.6 1.8 2.1 

1.2 1.3 1.6 1.8 2.0 

- 

-I 

- 

Coefficient of Variation 

0.6 0.7 0.8 0.9 1.0 

13.2 18.9 26.5 36.2 48.3 

7.4 9.8 12.7 16.1 20.2 

5.6 7.2 a.9 11.0 13.4 

4.7 5.9 7.2 a.7 10.3 

4.2 5.1 6.2 7.3 8.6 

3.8 4.6 5.5 6.4 7.5 

T 

t 

1.1 1.2 1.3 1.4 1.5 1.6 1.7 1.8 1.9 2.0 

63.3 81.4 102.8 128.0 157.1 

24 9 30.3 36.3 43.0 50.4 

16.0 19.0 22.2 25.7 29.4 

12.2 14.2 16.3 18.6 21.0 

10.0 11.5 13.1 14.8 16.6 

8.6 9.8 11.1 12.4 13.8 

3.6 4.2 5.0 5.8 6.7 I 7.7 8.7 9.7 10.8 12.0 

3.3 3.9 4.6 5.3 6.1 1 6.9 7.8 8.7 9.6 10.6 

3.2 3.7 4.3 5.0 5.7 6.4 7.1 7.9 a.7 9.6 

3.0 3.5 4.1 4.7 5.3 ; 5.9 6.6 7.3 8.0 8.8 

2.9 3.4 3.9 4.4 5.0 5.6 6.2 6.8 7.4 8.1 

2.8 3.2 3.7 4.2 4.7 5.2 5.8 6.4 7.0 7.5 

2.7 3.1 3.6 4.0 4.5 I 5.0 5.5 6.0 6.5 7.1 

2.6 3.0 3.4 3.9 4.3 ) 4.8 5.2 5.7 6.2 6.7 

2.6 2.9 3.3 3.7 4.1 I 4.6 5.0 5.4 5.9 6.4 

2.5 2.9 3.2 3.6 4.0 I 4.4 4.8 5.2 5.6 6.1 

2.5 2.8 3.1 3.5 3.8 1 4.2 4.6 5.0 5.4 5.8 

2.4 2.7 3.0 3.4 3.7 ( 4.1 4.4 4.8 5.2 5.6 

2.4 2.7 3.0 3.3 3.6 4.0 4.3 4.6 5.0 5.3 

2.3 2.6 2.9 3.2 3.5 3.8 4.2 4.5 4.8 5.2 

1 

90.3 227.8 269.9 316.7 368.3 

58.4 67.2 76.6 86.7 97.5 

33.5 37.7 42.3 47.0 52.0 

23.6 26.3 29.1 32.1 35.1 

18.4 20.4 22.4 24.5 26.6 

15.3 16.8 18.3 19.9 21.5 

13.1 14.4 15.6 16.9 18.2 

11.6 12.6 13.6 14.7 15.8 

10.4 11.3 12.2 13.1 14.0 

9.5 10.3 11.0 11.8 12.6 

8.8 9.4 10.1 10.8 11.5 

a.1 8.8 9.4 10.0 10.6 

7.6 8.2 a.7 9.3 9.9 

7.2 7.7 8.2 a.7 9.2 

6.8 7.3 7.7 8.2 8.7 ; 

6.5 6.9 7.3 7.8 8.2 I 

6.2 6.6 7.0 7.4 7.8 1 

5.9 6.3 6.7 7.0 7.4 [ 

5.7 6.0 6.4 6.7 7.1 

5.5 5.8 6.1 6.5 6.8 ; 

Table 3-2. Reasonable Potential Multiplying Factors: 95% Confidence Level and 95% Probability Basis 

INumberof Coefficient of Variation 

Samples 0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8 0.9 1.0 

1 1 1.4 1.9 2.6 3.6 4.7 6.2 8.0 10.1 12.6 15.5 

2 1.3 1.6 2.0 2.5 3.1 3.8 4.6 5.4 6.4 7.4 

3 1.2 1.5 1.8 2.1 2.5 3.0 3.5 4.0 4.6 5.2 

4 ; 1.2 1.4 1.7 1.9 2.2 2.6 2.9 3.3 3.7 4.2 

5 Il.2 1.4 1.6 1.8 2.1 2.3 2.6 2.9 3.2 3.6 

6 j 1.1 1.3 1.5 1.7 1.9 2.1 2.4 2.6 2.9 3.1 

ii ; ::: ::; ::: ::: ::t ::i ::: :I': ::: :I", 

9 1 1.1 1.2 1.4 1.5 1.7 1.8 2.0 2.1 2.3 2.4 

10 ( 1.1 1.2 1.3 1.5 1.6 1.7 1.9 2.0 2.2 2.3 

11 I 1.1 1.2 1.3 1.4 1.6 1.7 1.8 1.9 2.1 2.2 

12 I 1.1 1.2 1.3 1.4 1.5 1.6 1.7 1.9 2.0 2.1 

:: ) :;; ::; :;; ;'I 1.54 ;;; 1.76 ;:; 1.98 2.0 1.9 

15 ' 1.1 1.2 1.2 1.3 1.4 1.5 1.6 1.7 1.8 1.8 

16 ' 1.1 1.1 1.2 1.3 1.4 1.5 1.6 1.6 1.7 1.8 c 

17 I 1.1 1.1 1.2 1.3 1.4 

18 I 1.1 7.1 1.2 1.3 1.3 

19 ' 1.1 1.1 1.2 1.3 1.3 

20 I 1.1 1.1 1.2 1.2 1.3 

F 1 

1.1 1.2 1.3 1.4 1.5 ' 1.6 1.7 1.8 1.9 2.0 ' 

8.7 22.3 26.4 30.8 35.6 40.7 46.2 52.1 58.4 64.9 

8.5 9.7 10.9 12.2 13.6 15.0 16.4 17.9 19.5 21.1 I 

5.8 6.5 7.2 7.9 8.6 9.3 10.0 10.8 11.5 12.3 ) 

4.6 5.0 5.5 6.0 6.4 6.9 7.4 7.8 a.3 8.8 I 

3.9 4.2 4.5 4.9 5.2 ; 5.6 5.9 6.2 6.6 6.9 

3.4 3.7 3.9 4.2 4.5 I 4.7 5.0 5.2 5.5 5.7 I 

3.1 3.3 3.5 3.7 3.9 ( 4.1 4.3 4.5 4.7 4.9 , 

2.8 3.0 3.2 3.3 3.5 1 3.7 3.9 4.0 4.2 4.3 I 

2.6 2.8 2.9 3.1 3.2 3.4 3.5 3.6 3.8 3.9 / 

2.4 2.6 2.7 2.8 3.0 1 3.1 3.2 3.3 3.4 3.6 ' 

2.3 2.4 2.5 2.7 2.8 ; 2.9 3.0 3.1 3.2 3.3 

2.2 2.3 2.4 2.5 2.6 \ 2.7 2.8 2.9 3.0 3.0 

2.1 2.2 2.3 2.4 2.5 2.5 2.6 2.7 2.8 2.9 ' 

2.0 2.1 2.2 2.3 2.3 ' 2.4 2.5 2.6 2.6 2.7 

1.9 2.0 2.1 2.2 2.2 2.3 2.4 2.4 2.5 2.5 

1.9 1.9 2.0 2.1 2.1 1 2.2 2.3 2.3 2.4 2.4 

1.4 1.5 1.6 1.7 1.7 1.8 1.9 1.9 2.0 2.0 ' 2.1 2.2 2.2 2.3 2.3 

1.4 1.5 1.6 1.6 1.7 1.7 1.8 1.9 1.9 2.0 2.0 2.1 2.1 2.2 2.2 

1.4 1s 1.5 1.6 1.6 , 1.7 1.8 1.8 1.9 

1.4 1.4 1.5 1.5 1.6 1.7 1.7 1.8 1.8 

54 
Long-term average 

0 0.5 1 1.5 2 2.5 3 3.5 4 4.5 5 

Value 

50 

Number 01 Samples 

100 

Figure 3-la. Frequency Distribution of Values for a Figure 3-1~. Relationship Between the Largest Value of n 

Lognormal Distribution with a Mean of 1 .O and a Samples and the Percentile It Exceeds 

Coefficient of Variation of 0.6 with 99 Percent Confidence 

I/ 

Long-term average 

0 0.5 1 1.5 2 2.5 3 3.5 4 4.5 5 

Value 

Figure 3-1 b. Comparison of Relative Frequencies of 

Lognormal Distributions with a Mean of 1 .O for 

Different Coefficients of Variation 

2) With the exception of a small number of “outliers” for 

which confirmation is not possible, acute-to-chronic ratios 

(ACRs) above 20 for effluents discharged by NPDES permittees have not been observed by EPA. The majority of 

observed ACRs are very seldom above 10. However, higher 

ACRs may be found for selected facilities. 

3) The use of the three commonly used freshwater species 

and of three of the five commonly used marine organisms 

has generally been sufficient to measure any effluent’s 

toxicity for the purposes of projecting effluent toxicity 

impact and making regulatory decisions. 

Percentile = (1 - 0.99) 1in 

8 

20 

I I 

40 60 80 100 

Number of Samples 

Figure 3-ld. Example of 90 Percent Confidence 

Intervals Around Coefficient of Variation Estimates for 

Numbers of Samples 

Figure 3-2 is a flow chart of EPA’s recommendations for data 

generation for three different dilution scenarios. It is divided 

into three basic steps: determining initial dilution, developing toxicity testing procedures, and developing decision 

criteria for permit limit. There are certain basic assumptions 

built into this flow chart. The basic principle used in making 

decisions is to compare available dilution to known or projected toxic effect concentrations in order to place an effluent into one of three categories: 

55 
STEP 1 

determination l 

STEP 2 

Conduct toxicity testing* based 

on dilution determination (3 species 

at a minimum of quarterly for 1 year) 

STEP 3 

Acute toxicity data or Chronic toxicity data or 

estimate based on ACR v estimate based on ACR 

YES A Develop permit 4 

limits 

4 NO 

NO 

NO 

/A 

NO 

Require 

C monitoring at -= 

reissuance 

Notes: 

‘Dilution determrnations should be performed for critical flows and any applicable mixing zones. 

2Toxicity testing recommendations 

a. Dilution > 1OOO:l: acute testing, check CMC only. 

b. 100: 1 < Dilution c 1000: 1: acute or chronic testing, check CMC and CCC with data or ACR. 

c. Dilution < 1 0o:l: conduct chronic testing, check CCC with data and CMC using acute data or ACR. 

3Reasonable potential: Use procedures in Box 3-3. 

Figure 3-2. Effluent Characterization for Whole Effluent Toxicity 

1) The effluent causes or contributes to an excursion of a 

numeric or narrative water quality criterion and the permit 

require5 a limit on toxicity. 

2) The effluent has a reasonable potential of causing or contributing to an excursion of a numeric or narrative water 

quality criterion and a limit is required. 

3) The effluent has a very low probability of causing or contributing to an excursion of a water quality standard and 

no limit is required. 

This categorization is accomplished by using dilution estimates in the first step and the results of the toxicity tests in 

the next steps. In addition, all these impact estimates 

assume discharge at critical conditions and imposition of 

any applicable mixing zone requirements. Therefore, a 

conservative assumption is used to determine whether or 

not an impact is projected to occur. Estimates of possible 

toxic impact are made assuming that the effluent is most 

toxic to the most sensitive species or lifestage at the time of 

lowest available dilution. 

56 
dilution toxicity tests are much more variable than multiple dilution tests. Therefore, the use of single concentration toxicity 

tests is strongly discouraged for this data generation process. 

Since the new data generation requirements are much less expensive than the previous requirements, tiered testing (less expensive, 

single-concentration, initial screening followed by increasingly 

expensive definitive data generation, using multiconcentration 

tests, as described in the September 1985 version of the technical 

support document) is unnecessary. However, elimination of the 

requirement to conduct toxicity testing on the basis of projections using dilution alone is not recommended. Although EPA’s 

data review suggests that an LCso of 1 percent and an NOEC of 

0.1 percent are the lower bounds on effluent toxicity, there may 

be other effluents that are presently unmeasured that are more 

toxic. Testing data are always desirable for fully characterizing 

discharges of concern. 

Steps in Whole Effluent Characterization Process 

The following is a detailed description of the major steps presented in Figure 3-2 and the rationale behind each. 

Step 1: Dilution Determination 

The initial step is to determine the dilution of the effluent at the 

edge of the mixing zone, assuming the State allows mixing zones. 

Figure 3-4 shows a schematic representation of typical mixing 

zone requirements for both acute and chronic toxicity. Calculating the dilution at the edges of mixing zones for site-specific 

situations can be complicated. Modeling can be employed using 

either steady-state or dynamic approaches to calculate the dilution (see Chapter 4). However, for complex situations, such as 

marine and estuarine waters or lakes, dye studies (or other techniques used to assess mixing zones) may still be required. 

Some State water quality standards do not allow the use of 

mixing in the control of acute toxicity. For these States, acute 

toxicity is often limited at the end of the pipe. Permit limits 

derived to enforce such requirements would be considered “water quality-based” because they would be based upon an ambient 

criterion (as opposed to an arbitrary test endpoint). Regardless, 

both chronic and acute toxicity must be assessed in these situations. 

Step 2: Toxicity Testing Procedures 

Where toxicity tests are required in order to make decisions 

regarding appropriate next steps in a screening protocol, EPA 

recommends as a minimum that three species (for example, a 

vertebrate, an invertebrate, and a plant) be tested quarterly 

for a minimum of 1 year. As discussed in Chapter 1, the use of 

three species is strongly recommended. Experience indicates that 

marine algae can be a highly sensitive test species for some 

effluents. Using a surrogate species of the plant kingdom adds 

another trophic level to the testing regimen. For both freshwater 

and marine situations, the use of three species is more protective 

than two species since a wider range of species sensitivity can be 

measured. EPA is continuing to develop toxicity test methods 

using additional organisms including plants. In addition, EPA has 

revised the test for Selenastnum, which has improved the test 

precision. 

2,467 2,771 

2,064 

1.626 

1,241 

2000 

0 

DiMon (stream/etfluent) 

(a) At Low Flow (7QlO) 

6000 

6000 

e al 

P 

2 

x 

a 

B 

4000 

$ 

E 

2 

2oQo 

0 

3.849 

2,477 

l-b 

Dilution (stream/effluent) 

(b) At Annual Mean Flow 

Figure 3-3. National Distribution of NPOES Dilution 

Conditions at 7410 and at Annual Mean Flow 

The changes to the EPA’s data generation recommendations 

eliminate the application of multiple sets of safety margins 

that was proposed in the 1985 version of this document. 

Rather, general observations on effluent toxicity described 

above now allow regulatory authorities to tighten the bounds 

of the initial dilution categorization, eliminate the species 

sensitivity uncertainty factor and target LC~I-JS of 1 percent 

and NOECs of 0.1 percent as the most extreme toxicity 

measurements that can normally be expected for the vast 

majority of effluents discharged by NPDES permittees for 

acute and chronic toxicity, respectively. The observation of 

toxicity was based on multiple dilution tests. The same 

observation may not hold for toxicity measured with single 

dilution tests (pass/fail). As reflected in Chapter 1, single 

57 
--. --. .-. ~ - - -.~ ~.- ~- 

-_ - - -.-__- - -.-. - --_ 

No initial mlxtng allowed 

ccc 

b 

mlxlng zone 

- 

I 

wer width 

flow - - - + 

lnlrlal mlxlng allowed: 

flow - - - + I 

where 

CMC = 0.3 TU, 

CCC = 1 .O TUc 

Figure 3-4. Schematic Representation of Mixing Zone Areas 

Where the CMC and CCC Apply 

EPA recommends against selecting a “most sensitive” species 

for toxicity testing. For one organism to consistently be the 

most sensitive in a battery of toxicity tests, two conditions must 

occur: (1) the toxicants causing toxicity must remain the same, 

and (2) the ratios of the toxicants in the effluent (if more than 

one) must remain the same. Based on EPA’s experience at the 

Duluth research laboratory, neither of these conditions is likely to 

occur. For example, the causes of effluent toxicity in POlWs can 

vary on a seasonal basis. Toxicity in the summer can be caused 

by pesticides to which invertebrates are most sensitive. However, 

the winter toxicity could be caused by ammonia to which fathead 

minnows will respond most sensitively. The most sensitive species for an effluent actually may not exist and at best is difficult to 

identify. 

Conducting toxicity tests using three species quarterly for 1 

year is recommended to adequately assess the variability of 

toxicity obserwd in effluents. Below this minimum, the chances 

of missing toxic events increase. The toxicity test result for the 

most sensitive of the tested species is considered to be the 

measured toxicity for a particular effluent sample. 

The data generation recommendations in Figure 3-2 represent 

minimum testing requirements. Since uncertainty regarding 

whether or not an effluent causes toxic impact is reduced with 

more data, EPA recommends that this test frequency be increased where necessary to adequately assess effluent variability. If less frequent testing is required in the permit, it is 

preferable to use three species tested less frequently than to test 

the effluent more frequently with only a single species whose 

sensitivity to the effluent is not well characterized. 

EPA recommends that a discharger conduct acute toxicity 

testing if the dilution of the effluent is greater than 1OOO:l at 

the edge of the mixing zone [3]. Such a discharger would be 

considered a low priority for chronic toxicity testing. The rationale 

for this is that the effluent concentration would be below 0.1 

percent at the edge of the mixing zone and thus incapable of 

causing an excursion above the CCC. A worst case NOEC of 0.1 

percent translates into 1,000 TU,, which would result in a concentration of less than 1 .O TU, at the edge of the mixing zone for this 

dilution category. The test results would be compared to the CMC 

after consideration of any allowable mixing. 

EPA recommends that a discharger conduct either acute or 

chronic toxicity testing if the dilution of the effluent falls 

between 1OO:l and 1,OOO:l at the edge of the mixing zone. 

Effluents have been shown to be both acutely and chronically toxic 

within this range of receiving water dilution. Under worst-case 

scenarios, LCsos of 1 .O percent and ACRs of 10 will result in 

excursions above both the CCC and CMC at the edge of the 

regulatory mixing zone. 

Although either acute or chronic testing can be required within 

this dilution range, acute testing would be more appropriateat the 

higher end of this dilution range (1,OOO:l or 0.1 percent). At the 

lower end of this dilution range (1OO:l or 1 .O percent), chronic 

tests may be more appropriate. Where other factors are equal, 

chronic testing may be preferable since the interim results in a 

chronic test gives data on acute toxicity as well. The acute 

endpoint data can then be used to compare directly to the CMC 

without the need for an ACR. 

Whichever type of toxicity test (either acute or chronic) is specified, the results from that test should be compared to the criterion 

associated with that type of test. For example, a chronic test 

would be compared to the CCC. Comparisons to the other criteria 

can be made by using the ACR or additional data generated to 

convert a chronic test result to an acute endpoint and vice versa. 

For example, a chronic NOEC of 5 percent effluent (or 20 TU,) 

represents an acute LCso of 50 percent (or 2 TU,) at an ACR of 10. 

EPA recommends that a discharger conduct chronic toxicity 

testing if the dilution of the effluent falls below 1OO:l at the 

edge of the mixing zone. The rationale for this recommendation 

is that chronic toxicity has been observed in some effluents down 

to the 1 .O percent effect concentration. Therefore, chronic toxicity 

tests, although somewhat more expensive to conduct, should be 

used directly in order to make decisions about toxic impact. 

There is a potential for acute toxicity within this dilution range, 

although this is less likely as the 1OO:l dilution level is approached. 

Thus, the recommended screening protocol shown in Figure 3-2 

includes a determination of whether excursions above the CMC 

are projected [4]. This analysis may be performed by assuming an 

ACR, applying this value to the chronic toxicity testing data, and 

allowing for any allowable initial mixing. Alternatively, the regulatory authority may use the interim results in the chronic test to 

calculate the acute toxicity. 

58 
Both the chronic and acute toxicity test data would be compared 

to their respective criterion. The chronic test results would be 

compared to the CCC, and the acute results, regardless of how 

calculated, would be compared to the CMC. 

Step 3: Decision Criteria for Permit Limit Development 

Once the toxicity data have been generated for a discharger, the 

regulatory authority must decide whether or not the results show 

that the permittee causes, has the reasonable potential to cause, or 

contributes to an excursion of an applicable numeric or narrative 

water quality criterion and therefore needs to limit effluent toxicity. To do this, these data should be used to project receiving 

water concentrations, which are then compared to the CCC and 

CMC. One of four outcomes will be reached when following the 

screening protocol shown in Figure 3-2: 

1) 

2) 

Excursion Above CMC or CCC-Where any one data point 

shows an excursion above the State’s numeric or narrative 

criterion for the parameter toxicity, EPA regulations require a 

permit limit be set for whole effluent toxicity (40 CFR 

122.44(d)(l)(iv or v)), unless limits on a specific chemical 

will allow the narrative water quality criterion to be attained 

or maintained. In the absence of a State numeric criterion 

for the parameter toxicity, EPA recommends that 1.0 TU, 

and 0.3 TU, be used as the CCC and CMC,.respectively. 

The decision to develop permit limits based upon an excursion above either the CMC or CCC will lead to protection 

against both acute and chronic toxicity if the permit derivation procedures in Chapter 5 are used to set effluent limits. 

Reasonable Potential for Excursion Above CMC or CCCEPA believes that “reasonable potential” is shown where 

an effluent is projected to cause an excursion above the 

CCC or CMC. This projection is based upon a statistical 

analysis of available data that accounts for limited sample 

size and effluent variability. EPA’s detailed recommendations for making a statistical determination based upon 

effluent monitoring data alone are shown in Box 3-2. Where 

a regulatory authority finds that test results alone indicate a 

“reasonable potential” to cause an excursion above a State 

water quality criterion in accordance with 40 CFR 

122,44(d)(l)(ii), a permit limit must be developed. 

A regulatory authority may select an alternative approach 

for assessing reasonable potential. For example, an authority may opt to use a stochastic dilution model that incorporates both ambient dilution and effluent variability for determining reasonable potential. Such an approach is analogous to the statistical approach shown in Box 3-2. Whatever 

approach selected by the authority, it must use all the 

factors that account for all the factors listed in 40 CFR 

122,44(d)(l)(ii). 

In some cases the statistical analysis of the effluent data may 

not actually project an excursion above the CMC or CCC 

but may be close. Under such conditions, reasonable potential determinations will include an element of judgment on 

the part of the regulatory authority. Other factors will need 

to be considered and given appropriate weight in the 

decisionmaking process, including value of waterbody (e.g., 

high-use fishery), relative proximity to the CCC or CMC, 

existing controls on point and nonpoint sources, information on effluent variability, compliance history of the facility, and type of treatment facility. These factors are 

summarized in Box 3-2 and are discussed in detail in 

Section 3.1. EPA recommends regulatory authorities 

establish a written policy and procedure for making 

determinations of “reasonable potential” under these 

circumstances. 

3) No Reasonable Potential for Excursions Above CMC or: 

a---In these situations, EPA recommends that the 

toxicity tests recommended above be repeated at a 

frequency of at least once every 5 years as a part of 

the permit application. Such testing is required for 

certain POTWs under 40 CFR 122.21 (j). 

4) lnadeauate Information-Where a regulatory authority 

has inadequate information to determine reasonable potential for an excursion of a numeric or narrative water 

quality criterion, there may still be a basis for concern on 

the part of the authority. The permit should contain 

whole effluent toxicity monitoring requirements and a 

reopener clause. This clause would require reopening of 

the permit and establishment of a limit based upon any 

test results, or other new factors, which substantiate that 

the effluent causes, has the reasonable potential of causing, or contributes to an excursion above the CCC or 

CMC. 

3.3.4 Use of TuxMy Tesfhg in Multiple-sourMI glsclrarge 

SItuBtI4RF 

Where more than one discharge to the same receiving waterbody 

contributes, or has the reasonable potential to contribute to an 

excursion of water quality standards, permit limits must be 

developed for each individual discharger on that waterbody. 

For the regulatory authority to make this assessment, additional 

testing may be needed to provide the authority with the information necessary to assess the relative impact of each source. 

For purposes of this discussion, a multiple-source discharge 

situation is defined as a situation where impact zones overlap, or 

where ambient receiving water concentrations of a pollutant 

are elevated due to upstream discharges. In multiple-source 

discharge situations, additivity, antagonism, and persistence of 

toxicity can be of concern. To collect additional data, the permit 

authority should employ the toxicity testing procedures for 

multiple dischargers described in Box 3-3. In addition, ambient 

toxicity testing, as described below, could be used. 

Assuming that screening has been conducted that reveals the 

need for permit limits, two options for controlling the discharges exist. The first option is for the permit authority to 

regulate each source separately using the procedures for individual point sources. In this option, the permitting authority 

would require use of upstream ambient water as a diluent in the 

toxicity test so as to be able to evaluate the contributions of 

upstream sources of toxicity. A second option is to treat each 

discharge as an interactive component of a whole system. In 

this option, the permit writer would determine a total maximum daily load for the receiving waterbody and develop individual wasteload allocations for each discharger using the procedures discussed in Chapter 4. 

59 
Box 3-3. Recommend Multiple-source Toxicity Testing Procedures 

Tests 

Where the combined effluents make up 1 percent or greater of the receiving waters, conduct chronic toxicity 

tests following the testing procedures described in Section 3.3.3. 

Where the combined effluents make up less than 1 percent of the receiving waters, conduct acute toxicity tests 

following the testing procedures described in Section 3.3.3 (see Figure 3-2) to determine if any of the effluents 

are exhibiting toxicity. 

An additional data requirement is the assessment of relative and absolute toxicity of each source so that 

appropriate permit conditions can be set for individual dischargers. The following procedure is suggested. 

1) Conduct one set of toxicity tests on the effluents using a control of reconstituted or uncontaminated dilution 

water. The set of tests will give an absolute toxicity measurement of the effluent. 

2) Run a parallel set of toxicity tests on the effluent using dilution water taken directly upstream from the point of 

discharge or, for estuarine waters, from an area outside of the immediate discharge impact zone (this will have 

to be determined by a dye study). This dilution water may be contaminated with upstream effluents or other 

toxicant sources. The purpose of this test is to project toxic impact of the effluent after it is mixed at its point 

of discharge. This is a relative effluent toxicity measurement. The relative testing procedure could result in a 

change in the standard concentration-effect curve generated by the testing. The dilution water for the relative 

toxicity test may cause significant mortality, growth, or reproductive effects at the lower effluent concentrations (including the 100 percent diluent control concentration) if the diluent from the receiving water is toxic 

(from an upstream discharge). Such mortality does not invalidate the test. Instead, analysis of toxicity trends 

resulting from the relative toxicity tests can be used to assess the effluent’s toxicity in relation to other sources 

and ambient receiving water conditions. However, a control dilution water with no toxicity must be used for 

quality assurance and determination of absolute toxicity of the effluent. 

3) Conduct ambient toxicity tests to (a) determine whether or not the effluent has a measurable toxicity after 

mixing, (b) measure persistence of toxicity from all sources contributing to receiving water toxicity, and (c) 

determine combined toxicity resulting from the mixing of multiple, point, and nonpoint sources of toxicity. 

See Appendix C for a discussion of ambient toxicity testing procedures. 

The ambient testing can be required of each discharger and conducted during low-flow or worst-case design 

periods. 

Frequency for Ambient Testing 

All testing should be conducted simultaneously by each discharger, if possible. At a minimum, the tests should 

be conducted concurrently starting within a short time period (1 to 2 days). Repeated ambient toxicity analyses 

will be desirable when variable effluents are involved. Effluent toxicity data showing variability can be used to 

assess what frequency will be most applicable. The level of repetition for variability analysis should be similar to 

that used in effluent variability analyses. 

Other Considerations 

Dye studies of effluent dispersion for rivers, lakes, reservoirs, and estuaries are strongly recommended. This 

allows analysis of effluent concentration at the selected sampling stations above and below the discharge points. 

The procedures suggested in this multiple source section are based on actual multiple source site investigations 

conducted under the Complex Effluent Toxicity Testing Program. Site reports from that study can be used to 

obtain further description of the toxicity testing procedures used to analyze multiple source toxic impact [l, 21. 

60 
3.3.5 Ambient Toxicity Testing 

Ambient toxicity testing also is useful in screening receiving 

water bodies for existing toxic conditions. The procedure described in Appendix C uses short-term chronic toxicity tests to 

measure the toxicity of samples of receiving water taken above, 

at, and below outfalls. It can be used in freshwater, marine, and 

estuarine systems. The procedure must be conducted during an 

appropriate low-flow or worst-case design period. 

The utility of the ambient toxicity screening approach is that 

actual receiving water toxicity is directly measured. No extrapolation from exposure or ACR is needed. Further, impact from 

multiple source discharge situations, which may not be apparent 

from individual discharger data, is identified. Finally, the technique can provide an assessment of the persistence of effluent 

toxicity. 

3.3.6 Special Considerations fur Dischaqes to Marine and 

Estuarine Environments 

Special problems are encountered when assessing and controlling impacts of toxic pollutants discharged to marine and estuarine waterbodies. These special problems include the following: 

l Determining the physical characteristics of estuaries and 

the complex mixing and effluent dilution situations for 

RWCs of effluents. 

l Generating toxicity data on nonsaline effluents that discharge to brackish or saline waters and establishing causeeffect relationships on that basis. 

l Assessing exposure and controlling impacts from persistent toxicants accumulating in fish and shellfish tissues 

and in sediments. These factors are particularly important 

in estuaries and near coastal waters because of high use of 

estuaries as breeding and fishing areas for important commercial seafood supplies and recreational fishing, and because many estuaries and near coastal waters act as sinks 

for pollutants that accumulate in sediments. 

Where these special problems are encountered, additional information may need to be gathered to better quantify dilution, to 

determine metals partitioning, and to identify potential interferences in whole effluent toxicity tests. 

To characterize the type of whole effluent toxicity that is most 

relevant for a particular discharge to marine and estuarine waters, the following questions should be considered [5]: 

l What is the salinity of the receiving water, and is this 

important in terms of the State standards? 

l What is the appropriate test organism to require for toxicity testing under differing salinity conditions? 

The answers to these questions will enable the permitting authority to determine what type of toxicity testing is most suitable 

for effluent characterization and whole effluent toxicity control. 

For most marine and estuarine discharges the choice of test 

species and dilution water should be made based on the characteristics of the receiving water at the critical conditions for flow, 

mixing, and salinity. Foremost in this determination should be 

the salinity of the receiving water and, to a lesser extent, the 

salinity of the effluent itself. 

The primary objective of whole effluent toxicity tests is to identify 

sources of toxicity that can potentially cause an excursion of a 

State’s narrative or numeric water quality criteria. For this reason, 

the toxicity tests should reflect the natural conditions of the 

receiving water so to be able to measure any effluent characteristic that could contribute to ambient toxicity. The marine toxicity 

test methods identify 1,000 mg/l as the point at which salinity 

begins to exert an effect on freshwater species. As a general 

rule, EPA recommends that freshwater organisms be used 

when the receiving water salinity is less than 1,000 mg/l, and 

that marine organisms be used when the receiving water 

salinity equals or exceeds 1,000 mg/l. 

Saline Effluent Discharges to Saltwater 

The dissolved salts in the effluent are pollutants. These salts may 

or may not be the same as those present in the receiving water. 

Also, the proportion of dissolved salts in the effluent may be 

different from that of the salts in the receiving water. In this case, 

the toxicity test needs to be able to determine if these salts 

contribute to ambient toxicity. For this reason, marine organisms are needed. 

Saline Effluent Discharged to Freshwater 

In this case, the dissolved salts in the effluent is a pollutant that 

does not exist in the receiving water. The toxicity test needs to 

determine whether the dissolved salts can be one of the toxicants 

that contribute to ambient toxicity. For this reason, freshwater 

organisms are needed. 

Freshwater Effluent Discharged to Saltwater 

In this instance, the lack of dissolved salts in the effluent can 

cause an apparent toxic effect to the marine organisms in the 

toxicity test. However, in contrast to the instances presented 

above, the toxicity test does not need to be able to measure this 

effect because a lack of salts is not a pollutant. The marine 

toxicity test methods account for this by requiring that the 

salinity of the effluent be adjusted to approximate the salinity of 

the receiving water. As an alternative to using a marine organism, 

a freshwater organism can be used if the test is being conducted 

only on a loo-percent effluent sample and if State water quality 

standards do not require that a marine organism be used. 

3.3.7 Using a Chemical-specific limit to Control Toxicity 

EPA regulations at 40 CFR 122.44(d)(l )(v) provide that 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 that chemical-specific limits for the effluent are 

sufficient to attain and maintain applicable numeric and narrative 

State water quality criteria. To make this demonstration that 

chemical-specific limits are sufficient, additional effluent information will be needed. EPA recommends that the discharger 

conduct a toxicity identification evaluation to identify the 

causative agent(s) in the effluent. Where the permitting authority determines that the demonstration required by 40 CFR 

122.44(d)(l)(v) has been made, limits on whole effluent toxicity 

61 
need not be imposed. Effluent limits on the controlling chemical 

with concurrent whole effluent monitoring will be sufficient. Where 

subsequent whole effluent toxicity testing reveals the presence of 

toxicity in the effluent, the above process will need to be repeated, 

or alternatively a whole effluent toxicity limit will be needed. If 

continued toxicity testing shows that additional chemical-specific 

effluent limits are insufficient to control whole effluent toxicity, 

then toxicity limits may be the only practical way to control 

toxicity. 

3.3.8 Eif/uent Characterizetion ior @edtic Chemkak 

The previous section discussed effluent characterization for whole 

effluent toxicity. This section will describe EPA’s recommendations 

for data generation to determine whether or not permit limits are 

needed to control specific chemical pollutants in effluents. While 

many of the same principles apply when developing chemicalspecific limits, there are some differences based upon regulatory 

and analytical considerations. 

Characterization of impacts due to specific chemicals do not require a determination of the type of testing as is required for whole 

effluent toxicity because there is generally only one type of test for 

specific chemicals. However, there are some antecedent steps that 

are unique to effluent characterization for specific chemicals: determination of the chemicals of concern and determination of 

acceptable ambient levels (RAC, CMC, or CCC) for these pollutants. 

Steps for Chemical-specific Effluent Characterization Process 

Figure 3-5 illustrates EPA’s recommendations for determining 

whether or not permit limits need to be developed according to 

an evaluation of a limited data set. The following discussion 

corresponds to the various activities shown in Figure 3-5. (Refer to 

the human health discussion in Section 3.3.9 for additional details 

on procedures to characterize the bioconcentration potential of 

effluents.) 

Step 7: ldentifj the Po//uCunCs of Concern 

This process should begin with an examination of existing data to 

determine the presence of specific toxicants for which criteria, 

standards, or other toxicity data are available. Sources of data 

include the following: 

l Permit application forms, DMRs, permit compliance systems 

(PCS), and permit files 

l Pretreatment industrial surveys 

l STORET for ambient monitoring data 

l SARA Title Ill Toxic Chemical Release Inventory 

l Industrial effluent guidelines development documents 

l The Treatability Manual [6] 

. Effluent bioconcentration assessment (see Section 3.3.9). 

Data on specific chemicals that are typically submitted with NPDES 

application forms will consist of a limited number of analytical test 

results for many of the reported parameters. Where the regulatory authority has reason to believe that additional data for key 

parameters of concern are needed in order to adequately characterize the effluent, this information should be requested as a part 

of the application or, in some cases, through the use of Section 

308 letters. It is recommended that 8 to 12 samples be analyzed for key parameters of concern. In some cases, special 

analytical protocols will need to be specified in order to gather all 

appropriate information. 

Step 2: Determine the Basis for fstablishing RAG, ChlCs, and 

CCCs for the Pollutants of Concern 

The second step is to identify the appropriate water quality standard, including designated or existing use, and criteria for use. 

Ideally, the State water quality standards include aquatic life and 

human health criteria for the pollutants of concern. If a State does 

not have a numeric water quality criterion for the pollutant of 

concern, then one of three options for using the narrative criterion may be used (40 CFR 122.44(d)(l)(vi)) to determine whether 

a discharge causes, has the reasonable potential to cause, or 

contributes to an excursion above a narrative criteria because of 

an individual pollutant. Although the provisions of 40 CFR 

122.44(d)(l)(vi) are presented in the regulation in the context of 

permit limit development, these same considerations should be 

applied in characterizing effluents in order to determine whether 

limits are necessary. The options available are as follows: 

l Option A allows the regulatory authority to establish limits 

using a “calculated numeric water quality criterion” that 

the regulatory authority demonstrates will attain and maintain applicable narrative water quality criteria and fully 

protect the designated use. This option allows the regulatory authority to use any criterion that protects aquatic life 

and human health. This option also allows the use of sitespecific factors, including local human consumption rates 

of aquatic foods, the State’s determination of an appropriate risk level, and any other current data that may be 

available. 

l Option B allows the regulatory authority to establish effluent limits using EPA’s Water Quality Criteria guidance documents, if EPA has published a criteria document for the 

pollutant supplemented where necessary by other relevant 

information. As discussed earlier, EPA criteria documents 

provide a comprehensive summary of available data on the 

effects of a pollutant. 

l Option C may be used to develop limits for a pollutant of 

concern based on an indicator parameter under limited 

circumstances. An example of an indicator parameter is 

total toxic organics (l-TO>; effluent limits on TTO are useful 

where an effluent contains organic compounds. However, 

use of this option must be justified to show that controls on 

one pollutant control one or more other pollutants to a 

level that will attain and maintain applicable State narrative 

water quality criteria and will protect aquatic life and human health (see 40 CFR 122.44(d)(l)(vi)(C)). Use of this 

option is restricted by regulation to those instances where it 

can be demonstrated that controls on indicator pollutants 

serve to control the toxicant of concern. Using Option A or 

Option B is a more direct and perhaps more defensible 

approach. 

62 
STEP 1 

STEP 2 Determine RAC and/or 

CMUCCC for pollutants 

RAC available 

CMC and/or 

CCC avarlable 

STEP 3 

rhiizkd Dilution determination 

for aquatic life 

NO A NO 

A 

NO 

I I 

NO 

I I 

Require 

monitoring at 

reissuance 

Notes: 

’ RAC and/or CMC/CCC: Use State numeric criterion or interpret State narrative criterion using one of three options specified under 40 CFR 

122.44(d). 

2 Dilution determination: Perform for critical flow and for any applicable mixing zones for aquatic life and human health protection procedures, 

respectively. 

3 Reasonable potential: Use procedures in Boxes 3-2 and 3-4. 

Figure 3-5. Effluent Characterization for Specific Chemicals 

Step 3: Dilution Determination Step 4: Decision Criteria for Permit Limit Development 

The third step is to calculate the effluent dilution at the edge of 

the mixing zone. The pertinent factors for consideration here are 

the same as were previously presented for whole effluent toxicity 

with one difference: there are two levels of dilution analysis for 

chemical data. The first level is to use simple fate models based 

on a dilution analysis and comparison with the RAC, CMC, or 

CCC. The second level of analysis is to use more complex fate 

models, including dynamic models to estimate persistence, and 

may be applied to lakes, rivers, estuaries, and coastal systems 

using a desktop calculator or microcomputer. EPA has supported development of a second level of analysis that estimates 

point source wasteload allocations and nonpoint source allocations and predicts the resulting pollutant concentrations in receiving waters [7]. 

After this dilution analysis has been performed, the projected RWC 

is compared to the RAC, CMC, or CCC (either the State numeric 

criteria or an interpretation of the narrative criteria as described 

earlier). Whereas analysis of aquatic impacts should include evaluations with respect to both the CCC and the CMC, analysis of 

human health impacts will only involve comparisons with the RAC. 

The four possible outcomes discussed above in the triggers for 

permit limit development discussion in Section 3.3.3 also apply 

here: 

l Excursion above the RAC, CMC, or CCC 

l Reasonable potential for excursion above the RAC, CMC, or 

ccc 

63 
- - - -. - 

- _. - -.- 

l No reasonable potential for excursion above the RAC, CMC, 

ccc 

. Inadequate information. 

If these evaluations project excursions or the reasonable potential 

to cause or contribute to an excursion above the RAC, CMC, or 

CCC, then a permit limit is required (40 CFR 122.44(d)(l)(iii)). 

The statistical approach shown in Box 3-2 or an analogous approach developed by a regulatory authority can be used to determine the reasonable potential. Effluents that are shown not to 

cause or that have a reasonable potential to cause or contribute to 

an excursion above an RAC, CMC, or CCC should be reevaluated 

at permit reissuance. 

Where chemical-specific test results do not show a reasonable 

potential but indicate a basis for concern after consideration of the 

other factors discussed in Section 3.2, or if there were inadequate 

information to make a decision, the permit should contain chemical testing requirements and a reopener clause. This clause would 

require reopening of the permit and establishment of a limit based 

upon any test results that show effluent toxicity at levels that cause 

or have a reasonable potential to cause or contribute to an excursion above the RAC, CCC, or CMC. 

3.3.9 Effluent Characteriwtion tar BiocunceMh? 

Pollutants 

The previous section discussed how to characterize effects of 

specific chemicals, including those that may threaten human health, 

to determine whether or not a discharge causes, has the reasonable potential to cause, or contributes to excursions above an 

water quality criterion. The primary disadvantage of this approach 

is that it does not identify all effluent chemicals of potential cqntern for human health. To help address this gap, EPA is developing a procedure for identifying pollutants with the propensity to 

bioconcentrate in fish tissue. This procedure is presently in draft 

form and should not be used for establishing NPDES permit limits 

until EPA releases the final document on the procedure. This 

section describes the outline of this procedure. 

The overall approach illustrated in Figure 3-6 is a seven-step procedure that starts with collecting samples and ends with developing 

permit effluent limits. The effluent characterization step unique to 

this approach lies in Step 3. There are two alternatives under this 

step: fish tissue residue and effluent assessment. An analytical 

chemistry laboratory with residue chemistry and gas chromatograph/mass spectometer (GUMS) capability is needed to conduct 

the analytical methods for both alternatives. A summary of the 

alternatives follows: 

l Tissue Residue Alternative: This alternative measures the concentrations of organic bioconcentratable chemicals in tissue 

samples of indigenous organisms from the receiving water. 

This analysis involves the collection of fish or shellfish samples, 

the extraction of the organic chemicals from the tissue and 

the analysis of these extracts with CC/MS to identify and 

quantify the bloconcentratable contaminants. The procedure 

provides recommendations to sort the results of this screening 

analysis in order to determine which of the contaminants pose 

a hazard and require regulatory action. The approach recommends that the identity of those contaminants then be confirmed prior to taking subsequent action. 

Select di6chargerr 

biOcOnCentr6tiOn 

etternatlve 

Ti66ue 

n6idUe 

akernative 

Develop 

RACIRTC 

I 

DWdOQ 

WastelOad 

allocalion 

I 

Develop 

permit limit6 

I 

Conduct 

monitodng 

Figure 3-6. Procedure for Assessment and Control of 

Bloconcentratable Contaminants in Surface Waters 

l Effluent Alternative: This alternative measures the concentrations of organic bioconcentratable chemicals in effluent 

samples from point source dischargers. This analysis involves the collection of effluent samples, the extraction of 

the organic chemicals from the effluent sample, and the 

separation of the chemicals that have characteristics known 

to result in bioconcentration from the other chemical components of the effluent sample. This separation is achieved 

by way of an analytical chemistry methodology called high64 
pressure liquid chromatography (HPLC). The HPLC 

also separates (fractionates) an effluent sample into 

three subsamples or “fractions.” These three fractions 

contain chemicals with increasing potential to 

bioconcentrate, with the third fraction containing those 

chemicals with the highest bioconcentration rates. 

Following HPLC fractionation, each fraction is then 

analyzed with CC/MS to identify and quantify the 

bioconcentratable contaminants. The effluent procedure also provides recommendations to sort the results of the initial screening analysis to determine 

which of the contaminants pose a hazard and require 

subsequent regulatory action. The approach then 

recommends that the identity of those contaminants 

then be confirmed prior to taking further regulatory 

action. 

While both of the assessment alternatives described above 

may be used for a given discharger, generally one of these 

alternatives may be preferred by the regulatory authority. 

The regulatory authority would select the assessment approach based on the available site- and facility-specific information and the objectives of the application. 

Although the approach provides a means to identify chemicals 

that can bioconcentrate, it does not identify all bioconcentratable 

chemicals. Chemicals that bioconcentrate include many organic 

compounds, and a small number of metals (e.g., mercury and 

selenium) and organometals (e.g., tributyltin). The new approach 

is limited to nonpolar organic chemicals that produce measurable 

chemical residues in aquatic organisms or that have log octanolwater partition coefficients greater than 3.5. 

3.3.10 Analytical Considerations for Chemicals 

Analysis of discharges for toxic substances requires special quality 

control procedures beyond those necessary for conventional parameters. Toxicants can occur in trace concentrations and are 

frequently volatile or otherwise unstable. An EPA publication entitled, Test Methods-Technical Additions to Methods for Chemical 

Analysis of Water and Wastes [8], contains sampling and handling 

procedures recommended by EPA for a number of toxic and 

conventional parameters. Additional methods for analyses for 

toxicants are described in Standard Methods of Water and Wastewater Analyses (ASTM, 17th edition, 1989, or most recent edition) 

and 40 CFR Part 136. Chapter 5 discusses detection limits and 

sampling requirements. 

65 
CHAPTER 3 

REFERENCES 

1. Mount, D., N. Thomas, M. Barbour, T. Norberg, T. Roush, and 

W. Brandes. 1984. Effluent and Ambient Toxicity Testing and 

Instream Community Response on the Ottawa River, Lima, Ohio. 

Permits Division, Washington, DC, Office of Research and 

Development, Duluth, MN. EPA-600/2-84-080, August 1984. 

2. Mount, D.I., and T.J. Norberg-King, eds. 1985. Validity of 

fffluent and Ambient Toxicity Tests for Predicting Biological 

Impact, Scippo Creek, Circleville, Ohio. U.S. EPA. EPAJ600/3- 

85/044, june 1985. 

3. Peltier, W., and C.I. Weber. 1985. Methods for Measuring the 

Acute Toxicity of Effluents lo Aquatic Organisms, 3d ed. Office of Research and Development, Cincinnati, OH. EPA600/4-85-013. 

4. Weber, C.I., et al., eds. 1989. Short-Term Methods far fstimating the Chronic Toxicity of Effluents and Receiving Waters 

to Freshwater Organisms, 2d ed. Office of Research and 

Development, Cincinnati, OH. EPA-600/4-89/001. 

5. Weber, C.I., et al., eds. 1988. Short-Term Methods for fstimating the Chronic Toxicity of Effluents and Receiving Waters to 

Marine and fstuarine Organisms. Office of Research and 

Development, Cincinnati, OH. EPA-600/4-87/02. 

6. U.S. EPA. 1983. The Treataability Manual, Volume 4. U.S. EPA 

Office of Research and Development. EPA 600/2-82-001 

(revised January 24, 1983). 

7. Mills, W., et al, 1982. Water Quality Assessment: A Screening 

Procedure Toxic and Conventional Pollutants: Parts 1 and 2. 

Office of Research and Development, Athens, CA. EPA 

600/6-82-004 A, September 1982. 

8. U.S. EPA. 1982. Test Methods - Technical Additions to Methods 

for Chemical Analysis of Water and Wastes. Office of Research and Development, Cincinnati, OH. EPA 600/4-82- 

055, December 1982. 

AIIDKIONAL REFERENCES 

Crane, J.L., A. Pilli, and R.C. Russa. 1984. CfT/S: Complex DiToro, D. 1985. Exposure Assessment for Complex 

Effluent Toxicity Information System. CfTiS Retrieval System Effluents: Principles and Possibilities. In Environmental 

User’s Manual. Office of Research and Development, Hazard Assessment of ffflvents. Eds. H. Bergman, R. 

Duluth, MN. EPA-60018-84-030, November 1984. Kimerle, and A. Maki. 

Crane, J.L., A. Pilli, and R.C. Russa. 1984. UT/S: Complex Macek, K. 1985. Perspectives on the Application of the 

Effluents Toxicity Information System. Data Encoding Cuide- Hazard Evaluation Process. In Environmental Hazard 

lines and Procedures. Office of Research and Development, Assessment of fMuents. Eds. H. Bergman, R. Kimerle, 

Duluth, MN. EPA-60018-84-029. November 1984. and A. Maki. 

66 
4. EXPOSURE AND WASTELOAD ALLOCATION 

4.1 INTRODUCTION 

At this point in the toxics control process, a water quality problem 

has been identified. Screening analyses may have been done to 

assess the extent of toxicity, or a wasteload allocation (WLA) 

based on an existing total maximum daily load (TMDL) may 

already have been established. A TMDL is the sum of the individual WLAs for point sources and load allocations (LAs) for 

nonpoint sources of pollution and natural background sources, 

tributaries, or adjacent segments. WLAs represent that portion of 

a TMDL that is established to limit the amount of pollutants from 

existing and future point sources so that surface water quality is 

protected at all flow conditions. 

The TMDL process uses water quality analyses to predict water 

quality conditions and pollutant concentrations. Limits on wastewater pollutant loads are set and nonpoint source allocations are 

established so that predicted receiving water concentrations do 

not exceed water quality criteria. TMDLs and WLAs/LAs should 

be established at levels necessary to attain and maintain the 

applicable narrative and numerical water quality standards, with 

seasonal variations and a margin of safety that takes into account 

any lack of knowledge concerning the relationship between point 

and nonpoint source loadings and water quality. Determination 

of WLAs/LAs and TMDLs should take into account critical conditions for stream flow, loading, and water quality parameters. 

Conditions that will protect the receiving water have been determined from State numeric or narrative water quality criteria. 

This chapter is divided into sections that explain the steps that 

precede establishment of a WLA and then the methods and tools 

(models) that can be used to determine the WLA. Section 4.2 

briefly discusses TMDLs and how they relate to waters identified 

as requiring a water quality-based approach for toxics control. 

The section also discusses different WLA schemes. Sections 4.3 

and 4.4 discuss mixing zones, areas described as allocated impact 

zones where acute and chronic water quality criteria may be 

exceeded. Section 4.3 provides background information on mixing zones and discusses EPA’s mixing zone policy and how this 

policy affects the allowable toxic load that can be discharged from 

a point source. State mixing zone dimensions and the determination of mixing zone boundaries are also discussed. 

Section 4.4 discusses mixing zone analyses for situations in which 

the discharge does not mix completely with the receiving water 

within a short distance. Included in Section 4.4 are discussions of 

outfall designs that maximize initial dilution in the mixing zone, 

critical design periods for mixing zone analyses, and methods to 

analyze and model near-field and far-field mixing. 

Section 4.5 discusses the calculations of the WLA and LA and the 

types of EPA-recommended mathematical models available to 

determine WLAs in completely mixed situations for both aquatic 

life and human health. The WLA models listed in Section 4.5 can 

be used to predict ambient concentrations and to calculate the 

effluent quality required to meet the criteria and protect designated and existing uses of the receiving water. The data requirements of each of these models are also described so that the 

effluent characterization procedures described in Chapter 3 can 

be designed to support the specific types of WLA modeling 

selected by the regulator. Section 4.6 discusses human health 

considerations and how to determine WLAs for human health 

toxicants. 

EPA is currently working on methods to develop sediment criteria. 

Once developed, point source discharges could be further limited 

to prevent accumulation of pollutants in the bed sediment; such 

accumulation impairs beneficial uses. Although the criteria are 

not yet available for this document, they will be addressed in 

future documents. In the meantime, some of the models discussed in Section 4.5 are capable of simulating interactions between 

the water column and sediment and between toxic transport and 

transformation in the sediment. EPA is encouraging the States to 

consider the role of sediments in WLA. 

4.2 TOTAL MAXIMUM DAILY LOADS AND WASTELOAD 

ALLOCATIONS 

4.2.1 Total Maximum daily Loads 

The Federal Clean Water Act (CWA), under Section 303(d), requires the establishment of TMDLs for “water quality limited” 

stream segments. In such segments, water quality does not meet 

applicable water quality standards and/or is not expected to meet 

applicable water quality standards even after the application of 

the technology-based effluent limitations. A TMDL includes a 

determination of the amount of a pollutant, or property of a 

pollutant, from point, nonpoint, and natural background sources, 

including a margin of safety, that may be discharged to a water 

quality-limited waterbody. Any loading above this loading capacity risks violating water quality standards. TMDLs can be expressed 

in terms of chemical mass per unit of time, by toxicity, or by other 

appropriate measures. Permits should be issued based on TMDLs 

where available. 

The establishment of a TMDL for a particular waterbody is dependent on the location of point sources, available dilution, water 

quality standards, nonpoint source contributions, background 

conditions, and instream pollutant reactions and effluent toxicity. 

All of these factors can affect the allowable mass of the pollutant 

in the waterbody. Thus, two issues must be determined in 

conjunction with the establishment of the TMDL: (1) the definition of upstream and downstream boundaries of the waterbody 

for which the TMDL is being determined, and (2) the definition of 

critical conditions. For the following discussion, the waterbody 

boundaries are delineated as the portion of the waterbody be67 
tween the pollutant source (whether point source or nonpoint 

source) that is farthest upstream and the downstream point at 

which water quality has recovered to the background quality 

found above the pollutant source that is farthest upstream. The 

delineation of critical conditions for stream flow, loading, and 

water quality parameters may be specific to the type of waterbody 

and is discussed in Section 4.4. 

TMDLs are established based on water quality criteria pertinent to 

the designated and existing uses for the waterbody in question. 

TMDLs are traditionally calculated using State water quality standards as applied to a specific waterbody. Such a fitting of the 

TMDL to desired water quality criteria requires information concerning the distribution of loadings within the waterbody, namely, 

thelocations and relative contributions of pollutant-specific loadings from point, nonpoint, and background sources during all 

flow conditions (40 CFR 130.2(f)). Low-flow TMDLs, by themselves, will not be adequate in situations where nonpoint source 

loadings (LAs) during high or intermediate flow conditions cause 

excursions above water quality standards (40 CFR 130.2(f)). 

The loading capacity of TMDLs have been determined in many 

ways, but the most common method is to find the pollutant 

loading that will attain and maintain applicable water quality 

criteria. For example, in the Tualatin River Basin in Oregon, 

loading capacity was determined by multiplying stream flow in 

critical flow periods by the pollutant water quality standard [1]. 

Another method of determining a loading capacity is by quantifying instream toxicity. This method was used in developing a 

TMDL for the Amelia River in Florida [2]. 

The allowable TMDL is defined as the sum of the individual WLAs 

and LAs; a margin of safety can be included with the two types of 

allocations to ensure that allocated loads, regardless of source, 

would not produce an excursion above water quality standards. 

The WLAs are those portions of the TMDL assigned to point 

sources; the LAs are those portions of the TMDL assigned to the 

sum of all nonpoint sources and background sources (40 CFR 

130.2(f)). The background sources represent loadings to the 

specified waterbody or stream segment that come from sources 

outside the defined segment. For example, loadings from regions 

upstream of the segment and estimated atmospheric deposition 

of the pollutant would constitute background sources. Sediments 

that are highly contaminated from upstream discharges or historical discharges might also act as a source of toxicants and contribute 

to the background levels; these sediments also may be part of the 

nonpoint sources. 

The TMDL represents a mass loading that may occur over a given 

time period to attain and maintain water quality standards. As a 

result, the design flows under which the TMDL is determined can 

significantly alter its value. This phenomenon results in a somewhat unusual dichotomy. The design flows for aquatic life protection most applicable to point source loadings (WLAs) usually 

involve low-flow events (e.g., 7Q10) because the volumes associated with the point sources generally do not decrease with decreased stream flow. As a result, the highest concentrations 

associated with specific point source loads would be expected 

under low flow conditions. Conversely, elevated nonpoint source 

pollutant loadings (i.e., urban, agricultural) generally correspond 

to storm events. In fact, agricultural and urban runoff are often 

minimal or nonexistent in the absence of precipitation (i.e., nonexistent under low-flow drought conditions), 

The TMDL is a composite of the allowable loads associated with 

point sources and nonpoint sources within the defined boundaries of the waterbody segment and the background loadings to 

that segment from upstream and from in-place sediments. 

Therefore, the TMDL should be evaluated under conditions that 

reflect worst-case (critical) conditions for both point and nonpoint 

source loadings (i.e., low-flow drought and high flow conditions). 

Determination of the TMDL under these two scenarios would 

identify the lower of the two loading capacities of the waterbody. 

This lower capacity is necessary to protect the waterbody in 

question. 

In the case of design flows for human health protection, the 

harmonic mean flow is recommended as the basis for TMDLs for 

carcinogens. Design flows for human health protection should 

consider worst-case conditions for both point and nonpoint source 

loadings under this flow condition (see Section 4.6). 

In many cases, LAs for nonpoint sources are difficult to assess 

because the information needed to describe the runoff associated 

with the high-flow storm events does not exist. This lack of 

information is due to the high variability of the events. Because of 

the importance of estimating the nonpoint contributions to the 

waterbody, site-specific models may be required to estimate 

nonpoint source loadings. Even then, detailed models are difficult 

to calibrate with accuracy without intensive monitoring studies, 

and simplistic correlations between loadings and rainfall can be, 

by their statistical nature, unreliable for estimating low-frequency 

events (e.g., worst 10-year storm). The uncertainties associated 

with nonpoint source loadings and background sources require 

that the TMDL be determined with a sufficient margin of safety to 

allow for significant variability in nonpoint source loadings. 

CWA Section 303(d) and EPA regulations (40 CFR Parts 35 and 130, 

January 11, 1985) require that TMDLs contain a margin of safety 

“which takes into account any lack of knowledge concerning the 

relationship between effluent limitations and water quality.” The 

margin of safety is to take into account any uncertainties related 

to development of the water quality-based control, including any 

uncertainties in pollutant loadings, ambient conditions, and the 

model analysis. The size of the required margin of safety can, of 

course, be reduced by collecting additional information, which 

reduces the amount of uncertainty. The margin of safety can be 

provided for in the TMDL process by one of the following: 

• Reserving a portion of the loading capacity to a separate 

margin of safety. 

• Including a margin of safety within the individual WLAs for 

point sources and within the LAs for nonpoint sources and 

background sources. 

Most TMDLs are developed using the second approach, most 

often through the use of conservative design conditions. 

In addition, all WLAs, LAs. and TMDLs must meet the State 

antidegradation provisions developed prusuant to the Water 

Quality Standards Regulation (Section 137.12 of 40 CFR Part 131, 

68 
November 8, 1983). This regulation establishes explicit procedures that must be followed prior to lowering existing water 

quality to a level that still supports the Section 101 (a)(2) “fishable/ 

swimmable” goal of the Act. WlAs, LAS, and TMDLs that allow 

such a decline in water quality cannot be established unless the 

applicable public participation and intergovernmental review 

requirements of the antigradation provisions have been met and 

all existing uses are fully maintained and protected. 

4.2.2 WastsoMAkation scrbemes 

WLAS for water quality-based toxics permits must be set in accordance with EPA regulations [3, 41. EPA has developed a number 

of WIA guidance documents to assist regulatory authorities in 

developing TMDLs and WtAs. The EPA Office of Water Regulations and Standards, Assessment and Watershed Protection Division, maintains the latest listing of all WLA guidance documents. 

Toxic WLA guidance documents are currently available for rivers 

and streams [5], lakes and reservoirs 161, and estuaries [7]. Cuidante for the determination of critical design conditions for steadystate modeling of rivers and streams also is available [8]. 

Table 4-l lists 19 allocation schemes that may be used by the 

States to develop WLAs. This is not intended to be a complete list 

of approaches; regulatory authorities may use any reasonable 

allocation scheme that meets the antidegradation provisions and 

other requirements of State water quality standards [3]. 

The most commonly used allocation methods have been equal 

percent removal, equal effluent concentrations, and a hybrid 

method. The equal percent removal approach can be applied in 

two ways: the overall removal efficiencies of each pollutant 

source must be equal, or the incremental removal efficiencies 

must be equal. The equal effluent concentration approach also 

can be applied in two acceptable ways-equal final concentrations or equal incremental concentration reductions. This method 

is similar to the equal percent removal method if influent concentrations at all sources are approximately the same. However, if 

one point source has substantially higher influent levels, requiring 

equal effluent concentrations will result in higher overall treatment levels for that source than the equal percent removal appreach . 

The final commonly used method of allocating wasteloads is a 

hybrid method in which the criteria for waste reduction may not 

be the same for each point source. One facility may be allowed to 

operate unchanged, while another may be required to provide 

the entire load reduction. More often, a proportionality rule that 

requires the percent removal to be proportional to the input 

loading can be assigned. In these cases, larger sources would be 

required to achieve higher overall removals. 

4.3 INCIMPLETELY MIXED, DISCHARGE RECEIVING 

WATER SITUATIOWS 

Mixing zones are areas where an effluent discharge undergoes 

initial dilution and are extended to cover the secondary mixing in 

the ambient waterbody. A mixing zone is an allocated impact 

zone where acute and chronic water quality criteria can be exceeded as long as a number of protections are maintained, including freedom from the following: 

l Materials in concentrations that settle to form objectionable deposits 

. Floating debris, oil, scum, and other matter in concentrations that form nuisances 

Table 41. Wasteload Allocation Methods [9] 

1. Equal percent removal (equal percent treatment) 

2. Equal effluent concentrations 

3. Equal total mass discharge per day 

4. Equal mass discharge per capita per day 

5. Equal reduction of raw load (pounds per day) 

6. Equal ambient mean annual quality (mg/l) 

7. Equal cost per pound of pollutant removed 

8. Equal treatment cost per unit of production 

9. Equal mass discharged per unit of raw material used 

10. Equal mass discharged per unit of production 

11 a. Percent removal proportional to raw load per day 

11 b. Larger facilities to achieve higher removal rates 

12. Percent removal proportional to community effective 

income 

13a. Effluent charges (dollars per pound, etc.) 

13b. Effluent charge above some load limit 

14. Seasonal limits based on cost-effectiveness analysis 

15. Minimum total treatment cost 

16. Best availability technology (BAT) (industry) plus some 

level for municipal inputs 

17. hsimilative capacity divided to require an “equal effort 

among all dischargers” 

18a. Municipal: treatment level proportional to plant size 

18b. Industrial: equal percent between best practicable technology (BPT) and BAT, i.e., Allowable wasteload allocation: 

(WIA) = BPT- 1 i. (BPT-BAT) 

19. Industrial discharges given different treatment levels for 

different stream flows and seasons. For example, a plant 

might not be allowed to discharge when stream flow is 

below a certain value, but below another value, the 

plant would be required to use a higher level of treatment than BPT. Finally, when stream flow is above an 

upper value, the plant would be required to treat to a 

level comparable to BPT. 

L 

69 
l Substances in concentrations that produce objectionable 

color, odor, taste, or turbidity 

l Substances in concentrations that produce undesirable 

aquatic life or result in a dominance of nuisance species. 

Acutely toxic conditions are defined as those lethal to aquatic 

organisms that may pass through the mixing zone. As discussed 

in Chapter 2, the underlying assumption for allowing a mixing 

zone is that a small area of concentrations in excess of acute and 

chronic criteria, but below acutely toxic releases, can exist without 

causing adverse effects to the overall waterbody. The State 

regulatory agency can decide to allow or deny a mixing zone on a 

site-specific basis. For a mixing zone to be permitted, the discharger should prove to the State regulatory agency that all State 

requirements for a mixing zone are met. 

When wastewater is discharged into a waterbody, its transport 

may be divided into two stages with distinctive mixing characteristics. Mixing and dilution in the first stage are determined by the 

initial momentum and buoyancy of the discharge. This initial 

contact with the receiving water is where the concentration of the 

effluent will be its greatest in the water column. The design of the 

discharge outfall should provide ample momentum to dilute the 

concentrations in the immediate contact area as quickly as possible. 

The second stage of mixing covers a more extensive area in which 

the effect of initial momentum and buoyancy is diminished and 

the waste is mixed primarily by ambient turbulence. In large 

rivers or estuaries, this second-stage mixing area may extend for 

miles before uniformly mixed conditions are attained. In some 

instances, such as larger lakes or coastal bays, completely mixed 

conditions are never reached in the waterbody. The general 

definition for a completely mixed condition is when no measurable difference in the concentration of the pollutant (e.g., does 

not vary by more than 5 percent) exists across any transect of the 

waterbody. 

This section provides background information on the policy of 

mixing zones and the means to characterize them for use in WLAs 

(Section 4.5). The first subsection discusses the concerns that 

must be addressed when the boundaries and restrictions of a 

mixing zone are determined. The second subsection discusses 

the guidelines for preventing lethal conditions in the mixing zone. 

4.3.1 Detlmnlnatian of Mlxlog Z&w Bmtdiwk 

Allowable mixing zone characteristics should be established to 

ensure the following: 

l Mixing zones do not impair the integrity of the waterbody 

as a whole. 

l There is no lethality to organisms passing through the 

mixing zone. 

l There are no significant health risks, considering likely pathways of exposure (see Section 2.2.2). 

The Water Quality Criteria-l 972 [lo] recommends that mixing 

zone characteristics be defined on a case-by-case basis after it has 

been determined that the assimilative capacity of the receiving 

system can safely accommodate the discharge. This assessment 

should take into consideration the physical, chemical, and biological characteristics of the discharge and the receiving system; 

the life history and behavior of organisms in the receiving system; 

and the desired uses of the waters. Nearly all States require such 

an analysis before they allow a mixing zone [l l]. Further, mixing 

zones should not be permitted where they may endanger critical 

areas (e.g., drinking water supplies, recreational areas, breeding 

grounds, areas with sensitive biota). 

EPA has developed a holistic approach to determine whether a 

mixing zone is tolerable [12]. The method considers all the 

impacts to the waterbody and all the impacts that the drop in 

water quality will have on the surrounding ecosystem and 

waterbody uses. It is a multistep data collection and analysis 

procedure that is particularly sensitive to overlapping mixing 

zones. It includes the identification of all upstream and downstream waterbodies and the ecological and cultural data pertaining to them; the collection of data on all present and future 

discharges to the waterbody; the assessment of relative environmental value and level of protection needed for the waterbody; 

and, finally, the allocation of environmental impact for a discharge 

applicant. Because of the difficulty in collecting the data necessary 

for this procedure and the general lack of agreement concerning 

relative values, this method will be difficult to implement in full. 

However, the method does serve as a guide on how to proceed in 

allocating a mixing zone. 

Most States allow mixing zones as a policy issue, but provide 

spatial dimensions to limit the areal extent of the mixing zones. 

The mixing zones are then allowed (or not allowed) after case-bycase determinations. State regulations dealing with streams and 

rivers generally limit mixing zone widths, cross-sectional areas, 

and flow volumes and allow lengths to be determined on a caseby-case basis. For lakes, estuaries, and coastal waters, dimensions 

are usually specified by surface area, width, cross-sectional area, 

and volume. 

Where a mixing zone is allowed, water quality standards are met 

at the edge of that regulatory mixing zone during design flow 

conditions and generally, (1) provide a continuous zone of passage that meets water quality criteria for free-swimming and 

drifting organisms and (2) prevent impairment of critical resource 

areas. Individual State mixing zone dimensions are designed to 

limit the impact of a mixing zone on the waterbody. Furthermore, 

EPA’s review of State WLAs should evaluate whether assumptions 

of complete or incomplete mixing are appropriate based on 

avaitable data. 

In river systems, reservoirs, lakes, estuaries, and coastal waters, 

zones of passage are defined as continuous water routes of such 

volume, area, and quality as to allow passage of free-swimming 

and drifting organisms so that no significant effects are produced 

on their populations. Transport of a variety of organisms in river 

water and by tidal movements in estuaries is biologically important in a number of ways: food is carried to the sessile filter 

70 
feeders and other nonmobile organisms, spatial distribution of 

organisms and reinforcement of weakened populations are enhanced, and embryos and larvae of some fish species develop 

while drifting [ll], Anadromous and catadromous species must 

be able to reach suitable spawning areas. Their young (and in 

some cases the adults) must be assured a return route to their 

growing and living areas. Many species make migrations for 

spawning and other purposes. Barriers or blocks that prevent or 

interfere with these types of essential transport and movement 

can be created by water with inadequate chemical or physical 

quality. 

As explained above, a State regulatory agency may decide to 

deny a mixing zone in a site-specific case. For example, denial 

should be considered when bioaccumulative pollutants are in the 

discharge. The potential for a pollutant to bioaccumulate in living 

organisms is measured by (1) the bioconcentration factor (BCF), 

which is chemical-specific and describes the degree to which an 

organism or tissue can acquire a higher contaminant concentration than its environment (e.g., surface water); (2) the duration of 

exposure; and (3) the concentration of the chemical of interest. 

While any BCF value greater than 1 indicates that bioaccumulation 

potential exists, bioaccumulation potential is generally not considered to be significant unless the BCF exceeds 100 or more. 

Thus, a chemical that is discharged to a receiving stream, resulting in low concentrations, and that has a low BCF value will not 

create a bioaccumulation hazard. Conversely, a chemical that is 

discharged to a receiving stream, resulting in a low concentration 

but having a high BCF value, may cause in a bioaccumulation 

hazard. Also, some chemicals of relatively low toxicity, such as 

zinc, will bioconcentrate in fish without harmful effects resulting 

from human consumption. 

Another example of when a regulator should consider prohibiting 

a mixing zone is in situations where an effluent is known to attract 

biota. In such cases, provision of a continuous zone of passage 

around the mixing area will not serve the purpose of protecting 

aquatic life. A review of the technical literature on avoidance/ 

attraction behavior revealed that the majority of toxicants elicited 

an avoidance or neutral response at low concentrations [13]. 

However, some chemicals did elicit an attractive response, but the 

data were not sufficient to support any predictive methods. Temperature can be an attractive force and may counter an avoidance 

response to a pollutant, resulting in attraction to the toxicant 

discharge. Innate behavior such as migration may also supersede 

an avoidance response and cause fish to incur a significant exposure. 

4.32 Minimkhg the She of Mixing Zones 

Concentrations above the chronic criteria are likely to prevent 

sensitive taxa from taking up long-term residence in the mixing 

zone. In this regard, benthic organisms and territorial organisms 

are likely to be of greatest concern. The higher the concentrations occurring within an isopleth, the more taxa are likely to be 

excluded, thereby affecting the structure and function of the 

ecological community. It is thus important to minimize the 

overall size of the mixing zone and the size of elevated concentration isopleths within the mixing zone. 

4.3.3 Prevention of Lethalfty to Passing lkgankms 

The Water Quo/ity Standards Handbook [14] indicates that whether 

to establish a mixing zone policy is a matter of State discretion, 

but that any State policy allowing for mixing zones must be 

consistent with the CWA and is subject to approval of the Regional Administrator. The handbook provides additional discussion regarding the basis for a State mixing zone policy. 

Lethality is a function of the magnitude of pollutant concentrations and the duration an organism is exposed to those concentrations. Requirements for wastewater plumes that tend to attract 

aquatic life should incorporate measures to reduce the toxicity 

(e.g., via pretreatment, dilution) to minimize lethality or any 

irreversible toxic effects on aquatic life. 

EPA’s water quality criteria provide guidance on the magnitude 

and duration of pollutant concentrations causing lethality. The 

criterion maximum concentration (CMC) is used as a means to 

prevent lethality or other acute effects. h explained in Appendix 

D, the CMC is a toxicity level and should not be confused with an 

LCso level. The CMC is defined as one-half of the final acute value 

for specific toxicants and 0.3 acute toxic unit (TU,) for effluent 

toxicity (see Chapter 2). The CMC describes the condition under 

which lethality will not occur if the duration of the exposure to the 

CMC level is less than 1 hour. The CMC for whole effluent toxicity 

is intended to prevent lethality or acute effects in the aquatic 

biota. The CMC for individual toxicants prevents acute effects in 

all but a small percentage of the tested species. Thus, the areal 

extent and concentration isopleths of the mixing zone must be 

such that the l-hour average exposure of organisms passing 

through the mixing zone is less than the CMC. The organism 

must be able to pass through quickly or flee the hiqh-concentration area. The objective of developing water quality recommendations for mixing zones is to provide time-exposure histories that 

produce negligible or no measurable effects on populations of 

critical species in the receiving system. 

Lethality to passing organisms can be prevented in the mixing 

zone in one of four ways. The first method is to prohibit concentrations in excess of the CMC in the pipe itself, as measured 

directly at the end of the pipe. As an example, the CMC should 

be met in the pipe whenever a continuous discharge is made to 

an intermittent stream. The second approach is to require that 

the CMC be met within a very short distance from the outfall 

during chronic design-flow conditions for receiving waters (see 

Section 4.4.2). 

If the second alternative is selected, hydraulic investigations 

and calculations indicate that the use of a high-velocity discharge with an initial velocity of 3 meters per second, or 

more, together with a mixing zone spatial limitation of 50 

times the discharge length scale in any direction, should 

ensure that the CMC is met within a few minutes under 

practically all conditions. The discharge length scale is defined 

as the square root of the cross-sectional area of any discharge 

pipe. 

A third alternative (applicable to any waterbody) is not to use a 

high-velocity discharge. Rather the discharger should provide 

71 
- 

-- 

data to the State regulatory agency showing that the most restrictive of the following conditions are met for each outfall: 

The CMC should be met within 10 percent of the distance 

from the edge of the outfall structure to the edge of the 

regulatory mixing zone in any spatial direction. 

The CMC should be met within a distance of 50 times the 

discharge length scale in any spatial direction. In the case 

of a multiport diffuser, this requirement must be met for 

each port using the appropriate discharge length scale of 

that port. This restriction will ensure a dilution factor of at 

least 10 within this distance under all possible circumstances, including situations of severe bottom interaction, 

surface interaction, or lateral merging. 

The CMC should be met within a distance of five times the 

local water depth in any horizontal direction from any 

discharge outlet. The local water depth is defined as the 

natural water depth (existing prior to the installation of the 

discharge outlet) prevailing under mixing zone design conditions (e.g., low flow for rivers). This restriction will prevent locating the discharge in very shallow environments or 

very close to shore, which would result in significant surface 

and bottom concentrations. 

A fourth alternative (applicable to any waterbody) is for the 

discharger to provide data to the State regulatory agency showing that a drifting organism would not be exposed to l-hour 

average concentrations exceeding the CMC, or would not receive 

harmful exposure when evaluated by other valid toxicological 

analysis, as discussed in Section 2.2.2. Such data should be 

collected during environmental conditions that replicate critical 

conditions. 

For the third and fourth alternatives, examples of such data 

include monitoring studies, except for those situations where 

collecting chemical samples to develop monitoring data would 

be impractical, such as at deep outfalls in oceans, lakes, or 

embayments. Other types of data could include field tracer 

studies using dye, current meters, other tracer materials, or detailed analytical calculations, such as modeling estimations of 

concentration or dilution isopleths. 

The Water Quality Criteria-l 972 [l l] outlines a method, applicable to the fourth alternative, to determine whether a mixing 

zone is tolerable for a free-swimming or drifting organism. The 

method incorporates mortality rates (based on toxicity studies for 

the pollutant of concern and a representative organism) along 

with the concentration isopleths of the mixing zone and the 

length of time the organism may spend in each isopleth. The 

intent of the method is to prevent the actual time of exposure 

from exceeding the exposure time required to elicit an effect [lo]: 

1 T 0-4 < , 

ET(X) at C(n) - 

where T(n) is the exposure time an organism is in isopleth n, and 

ET(X) is the “effect time.” That is, ET(X) is the exposure time 

required to produce an effect (including a delayed effect) in X 

percent of organisms exposed to a concentration equal to C(n), 

the concentration in isopleth n. ET(X) is experimentally determined; the effect is usually mortality. If the summation of ratios of 

exposure time to effect time is less than 1, then the percent effect 

will not occur. 

4.3.4 mvootho of Bio%cc~l~ioo Rvblmos for Muman 

States are not required to allow mixing zones. Where unsafe fish 

tissue levels or other evidence indicates a lack of assimilative 

capacity in a particular water-body for a bioaccumuiative pollutant, care should be taken in calculating discharge limits for this 

pollutant or the additivity of multiple pollutants. In particular, 

relaxing discharge limits because of the provision of a mixing 

zone may not be appropriate in this situation. 

4.4 MIXMG ZOHE ANALYSES 

Proper design of a mixing zone study for a particular waterbody 

requires estimation of the distance from the outfall to the point 

where the effluent mixes completely with the receiving water. 

The boundary is usually defined as the location where the concentrations across a transect of the waterbody differ by less than 5 

percent. The boundary can be determined based on the results of 

a tracer study or the use of mixing zone models. Both procedures, along with simple order-of-magnitude dilution calculations, are discussed in the following subsections. 

If the distance to complete mixing is insignificant, then mixing 

zone modeling is not necessary and the fate and transport models 

described in Section 4.5 can be used to perform the WLA. It is 

important to remember that the assumption of complete 

mixing is not a conservative assumption for toxic discharges; 

an assumption of minimal mixing is the conservative approach. If completely mixed conditions do not occur within a 

short distance of the outfall, the WLA study should rely on mixing 

zone monitoring and modeling. Just as in the case of completely 

mixed models, mixing zone analysis can be performed using both 

steady-state and dynamic techniques. State requirements regarding the mixing zone will determine how water quality criteria are 

used in the TMDL. 

This section is divided into five subsections. The first discusses 

recommendations for outfall designs and means to maximize 

initial dilution. The second provides a brief description of the four 

major waterbody types and the critical design period when mixing zone analysis should be performed for each. The third provides a brief description of tracer studies and how they may be 

used to define a mixing zone. The fourth and fifth subsections 

discuss simplified methods and sophisticated models to predict 

the two stages of mixing (i.e., discharge-induced and ambientinduced mixing). For a detailed explanation of the mechanisms 

involved in estimating both stages of mixing, two references are 

recommended, Halley and Jirka [15] and Fischer et al. [16]. 

Although the models presented in Sections 4.4.4 and 4.4.5 simplify the mixing process, the assessor should have an understanding of the basic physical concepts governing mixing to use these 

72 
models appropriately. (The U.S. EPA Center for Exposure Assessment Modeling [CEAM] in Athens, Georgia, provides an overview 

course that teaches the basics of mixing and how the basics 

should be used for water quality management.) 

It is important to note that the mixing zone models presented 

here attempt to predict the dispersion and dilution of the effluent 

plume. They do not attempt to predict any removal or transformation of the pollutants. In the near field, dispersion and dilution 

caused by discharge-induced mixing and then ambient-induced 

mixing will be the major cause of toxicity reduction. If incomplete 

mixing persists downstream (such as in the case of shore hugging 

plumes), then some far-field processes will become important. 

Some of the models described in Section 4.5 that have sophisticated hydrodynamic simulation routines coupled with fate simulation routines may be used for these far-field, incomplete mixing 

analyses. 

An important factor in maximizing the initial dilution of an effluent is the design of the effluent outfall. There are three major 

types of outfall designs: surface discharge from free flows in a pipe 

or canal, single-port submerged discharge, and multiport submerged discharge. The last type is often referred to as multiport 

diffusers. Of the three, the surface discharge type is the least 

favorable for toxic discharges since it offers the least initial mixing. 

In particular, surface discharges at the shoreline of a waterbody 

usually have an impact along the shoreline when there is significant cross-flow and thus yield high surface concentrations. 

Submerged discharges offer more flexibility in meeting the design 

goals for toxic discharges. Submerged discharges may be in the 

form of a single pipe outlet or of multiport discharges (diffusers) 

giving rise to one or several submerged discharge jets. A typical 

diffuser section is illustrated in Figure 4-1. Submerged discharges 

allow the effluent to be directed at different angles to the ambient 

flow to maximize the initial dilution. Diffusers are particularly 

effective in counteracting the buoyancy of the effluent. However, 

submerged muttiport discharges are only feasible in waterbodies 

that are of sufficient depth and are not subjected to periodic 

dredging or to considerable scour or deposition. 

0.20 m x 0.15 m 

90” reducer elbow 

0.15 m cast-iron pipe 

\ \ / F’ange 

Top of tremle 

encasement 

‘t7i / Flannta m,nl 

flange with 0.15 m casliron trap at Invert 

0.15 m 90” elbow 

Bottom of tremle / 

encasement 

Figure 4-l. A Typical Diffuser Section [17] 

Many of the complexities of submerged diffusers have been 

summarized by jirka [I 81, Halley and Jirka [15], and Roberts et al. 

[19, 20, 211. Submerged discharges should be designed to avoid 

direct surface impingement and bottom attachment of the submerged jet or jets. Surface and bottom impacts should be 

evaluated at critical design conditions (low flow or high stratification) and at off-design conditions (higher flow or lower stratification) to ensure the best placement and design of the diffuser. 

Diffusers provide more dilution than single outlets, but the alignment of the diffuser with the receiving water flow direction influences how much dilution will be provided. If the outlet structure 

is directed parallel to the direction of flow, dilution under high 

ambient velocities (off-design conditions) may be lower than 

under low velocities (critical design conditions). 

In rivers, the preferred arrangement for a submerged discharge is 

to direct the outlet into the current flow direction or vertically 

upward. To deal with the reversing currents of estuaries and 

coastal bays, the preferred arrangements for offshore discharges 

are parallel diffuser alignment (tee diffuser) and perpendicular 

diffuser alignment (staged diffuser) [18]. In lakes and reservoirs, 

the preferred arrangement for a negatively buoyant discharge is 

to direct the diffuser vertically upward. A positively buoyant, 

vertically directed jet could penetrate stratification, so the preference for this type of discharge is to orient the diffuser at a slight 

angle above the horizontal. For ocean outfalls, initial dilution is 

improved by longer (perpendicular to the shoreline) and deeper 

diffusers. Further, the ports of the diffuser should be sufficiently 

separated to minimize merging of the separate plumes [22]. 

4.4.2 Critical Des&n Periods for Watehdies 

This section provides a brief description of the four major waterbody 

types and defines the critical design periods that should be used 

when performing mixing zone analyses in each of these waterbody 

types. Appendix D provides a further discussion on the appropriate selection of design periods. 

1) Rivers and Run-of-River Reservoirs 

Rivers and run-of-river reservoirs are waterbodies that have a 

persistent throughflow in the downstream direction and do not 

exhibit significant natural density stratification. Recommendations for hydrologically based and biologically based design flows 

for completely mixed, steady-state modeling of rivers are described in Appendix D of this document. The biologically based 

design flows are determined using the averaging periods and 

frequencies specified in water quality criteria [8]. Also, the hydrologically based flows 1QlO and 7410 for the CMC and CCC, 

respectively, have been used traditionally and may continue to be 

used for steady-state modeling. Run-of-river reservoirs with residence times less than 20 days at critical conditions also should be 

analyzed using biologically or hydrologically based design flows 

(see below). Regulated rivers may have a minimum flow in excess 

of these toxicological flows. In such cases, the minimum flow 

should be used in TMDL modeling. 

2) Lakes and Reservoirs 

This receiving water category encompasses lakes and reservoirs 

with residence times in excess of 20 days at critical conditions 

73 
[23]. Seasonal variations in the water level, wind speed and 

direction, and seasonal solar radiation should be determined to 

define the critical period [23]. In the case of long and narrow 

reservoirs, areas above the plunge point (i.e., areas where no 

stream-like flow is present and waters are mixed or stratified by 

density) can be analyzed as rivers. The areas below can be 

analyzed as reservoirs. Since effluent density relative to the ambient water can vary over seasons, no one season or stratification 

condition can be selected as the most critical dilution situation for 

all cases. In general, all four seasons should be analyzed to 

determine the most critical periods for mixing zone analyses. All 

seasonal analyses should assume an ambient velocity of zero 

unless persistent currents have been documented. Special attention should be given to periods of rising water level since pollutants can move back into coves and accumulate under these 

conditions. Location of discharges in coves and dead-end 

embayments should be prevented whenever possible. 

3) Estuaries and Coastal Bays 

This receiving water category encompasses estuaries, which are 

defined as having a main channel reversing flow, and coastal 

bays, which are defined as having significant two-dimensional 

flow in the horizontal directions. For both waterbodies, the 

critical design conditions recommended here are based on astronomical, not meteorotogicat, tides. 

Determining the nature and extent of the discharge plume is 

complicated in marine systems by such conditions as differences 

in tides, riverine input, wind intensity and direction, and thermal 

and saline stratification. Because of the tidal nature of the estuaries and coastal systems and their complex circulation patterns, 

dilution of discharges cannot be determined simply by calculating 

the discharge rate and the rate of receiving water flow (i.e., the 

design flow). For example, tidal frequency and amplitude vary 

significantly in different coastal regions of the United States. 

Furthermore, tidat influences at any specific location have daily 

and monthly cycles. These and additional factors require that 

direct, empirical steps be taken to ensure that basic dilution 

characteristics of a discharge to salt water are determined. 

tn estuaries without stratification, the critical dilution condition 

includes a combination of low-water slack at spring tide for the 

estuary and design low flow for riverine inflow. In estuaries with 

stratification, a site-specific analysis of a period of minimum stratification and a period of maximum stratification, both at lowwater slack, should be made to evaluate which one results in the 

lowest dilution. In general, minimum stratification is associated 

with low river inflows and large tidal ranges (spring tide), whereas 

maximum stratification is associated with high river inflows and 

low tidal ranges (neap tide). 

After either stratified or unstratified estuaries are evaluated at 

critical design conditions, an off-design condition should be 

checked. The off-design condition (e.g., higher flow or lower 

stratification) recommended for both cases is the period of maximum velocity during a tidal cycle. This off-design condition 

results in greater dilution than the design condition, but it causes 

the maximal extension of the plume. Extension of the plume into 

critical resource areas may cause more water quality problems 

than the high-concentration, low-dilution situation. 

Recommendations for a critical design for coastal bays are the 

same as for stratified estuaries. The period of maximum stratification must be compared with the period of minimum stratification 

in order to select the worst case. The off-design condition of 

maximum tidal velocity should also be evaluated to predict the 

worst-case extent of the plume. 

4) Oceans 

Critical design periods for ocean analyses are described in two 

separate documents, the Section 301 (h) Technical Support Document [22] and the Section 301 (h) document, hitid Mixing Characteristics of Municipal Ocean Discharges [24]. The following subsection contains a summary from these documents. Like discharges to estuaries, discharges to ocean waters are subject to 

two-dimensional horizontal flows. Oceanic critical design periods 

must include periods with maximum thermal stratification, or 

density stratification. These periods shorten the distance of vertical diffusion that occurs in the zone of initial dilution. Thus, 

during these periods it is difficult to achieve the recommended 

loo-to-1 dilution that is to occur before the plume begins a 

predominantly horizontal flow as compared to vertical flow. Periods when discharge characteristics, oceanographic conditions 

(spring tide and neap tide currents), wet and dry weather periods, 

biological conditions, or water quality conditions that indicate 

that water quality standards are likely to be exceeded should also 

be noted. The 10th percentile value from the cumulative frequency of each parameter should be used to define the period of 

minimal dilution. 

4.4.3 lbwml Rwmmwht&as for Tracer 3iudk 

A tracer or dye study can be used to determine the areal extent of 

mixing in a waterbody, the boundary where the effluent has 

completely mixed with the ambient water, and the dilution that 

results from the mixing. Analysis of the mixing zone with a dye 

study that is supplemented with modeling should be performed 

at flow conditions that approach critical flow. Some of those 

design conditions are summarized above in the subsections dealing with specific waterbodies. Once the critical design condition 

has been selected for a waterbody, dye studies can be performed 

to provide data on the dimensions and dilution of the wastewater 

plume during this critical period. Tracer studies other than dye 

studies (e.g., chloride, lithium) can be performed for cases in 

which the receiving water is amenable to such tests. 

For WIA studies in which a discharge is already in operation, 

tracer studies can be used to determine specific concentration 

isopleths in the mixing zone that reflect both discharge-induced 

and ambient-induced mixing. The isopleth concentrations, with 

effluent toxic concentrations, should be superimposed over a 

map of Ihe various resource zones of the waterbody. The map 

will illustrate whether the State’s mixing zone dimensions are 

exceeded, whether the required zone of passage is provided, and 

whether the plume avoids critical resource areas. The WLA can 

then be calculated to provide the appropriate zone of passage 

and to prevent detrimental impacts on spawning grounds, nurseries, water supply intakes, bathing areas, and other important 

resource areas. 

Obviously, if the outfall is not yet in operation, it is impossible to 

determine discharge-induced mixing by tracer studies. Tracer 

74 
studies can be used in these situations to determine characteristics 

of the ambient mixing. For ambient mixing studies, the tracer 

release can be either instantaneous or continuous. Instantaneous 

releases are used frequently to measure longitudinal dispersion, 

but can also be used to determine lateral mixing in rivers [15] and 

lateral and vertical mixing in estuaries, bays, reservoirs, and lakes. 

For waterbodies with significant flow velocities, continuous releases of tracer are normally used to determine lateral and vertical 

mixing coefficients. Continuous releases can also be used to 

determine three-dimensional concentration isopleths for steadystate conditions. The tracer study must be made at critical design 

conditions in order to use the results directly for WlAs. If a tracer 

study for ambient mixing is conducted at near-to-design conditions, the observed data can be used to determine dimensionless 

mixing coefficients. These coefficients can then be extrapolated 

to critical conditions using hydraulic parameters [15]. A tracer 

study at near-to-critical conditions also can be used to determine 

the computer model required to predict critical-condition mixing 

and provide the coefficients needed for that TMDL model. 

A number of references provide information concerning the design, conduct, and analysis of tracer studies for mixing analyses. 

Techniques of Water-Resources Investigations of the USC.5 provides 

the best overview of how to conduct tracer studies [25, 26, 271. 

The fluorescent dyes (usually Rhodamine WT), measuring equipment, fluorometers, field and laboratory procedures, and calculation methods are all discussed. The procedures essentially consist 

of adding dye to the waterbody and recording concentrations of 

the dye at various stations at specific time intervals. Examples of 

tracer studies for river systems are presented in Fischer [28]; Kisiel 

[29]; Holley and )irka [l 51; and Yotsukura, Fisher, and Sayre [30]. 

Examples of tracer studies in tidal systems are presented in Wilson, 

Cobb, and Yotsukura [31] and Hetling and O’Connell [32], both 

of which are studies of the Potomac River estuary; Baily [33], a 

study of Suisun Bay in California; Fischer [34], a study of Bolinas 

Lagoon, a coastal bay in Marin County, California; and Cracker et 

al. [35], a study of Corpus Christi Bay, Texas. Methods to perform 

a tracer study in a reservoir are provided in johnson [36]. 

The dye study recommended for obtaining a quick saltwater 

dilution assessment is one in which Rhodamine WT dye is administered to a discharge and monitored in the receiving waters for 

not less than 24 hours. The basic goal of this study is to determine 

the near-field nature of the effluent dilution, not the steady-state 

or far-field dilution. The environmental and discharge conditions 

selected for the study should be those that would elicit “worstcase” conditions (i.e., highest ambient concentrations in the receiving water). These include low wind, neap tide (tide of minimum range occurring during the 1st and 3rd quarters of the 

moon), plume trapping by density stratification, low rainfall and 

low riverine input, and, if possible, high effluent discharge. 

The dye should be administered to the effluent before discharge 

to the receiving water in proportion to effluent flow rate. Dye 

should be maintained at a concentration in the effluent sufficient 

to permit detection of the dilution ratio of interest when the 

amount and variability of background fluorescence in the receiving water are taken into account. Measurements of dye concentration are made using a fluorometer and should be corrected for 

water temperature. 

A survey of background fluorescence and its variability in the 

anticipated mixing zone must be conducted just prior to the 

beginning of the study in order to permit correction of fluorescence data and to determine the dye concentration required in 

the effluent. Since Rhodamine Wl dye is bleached by free chlorine, a preliminary study of the degree of dye bleaching by the 

effluent should precede the study for chlorinated discharges to 

avoid underestimation of the extent of the mixing zone. Dye 

concentrations should be surveyed for two successive slack tides, 

and for any other conditions that could lead to concentration 

maxima. Surveys should extend from the point of discharge to a 

distance at which the effluent dilution ratio of interest is attained. 

The dye fluorescence at this point should be at least twice the 

variability in background fluorescence. 

EPA has completed two TMDL studies to test the procedures 

outlined in the previous version of this document. Both studies 

used dye to determine the mixing zone and the dilution within it. 

The first study was performed on the Amelia River, an estuarine 

system in Florida [2]; the second was performed on the Creenwith Cove, an embayment of Narragansett Bay in Rhode Island 

[37]. In both studies, Rhodamine Wl dye was introduced continuously into the effluent and numerous stations were set up to 

measure the spatial and temporal distribution of the dye. Both 

studies are good examples of how to perform a dye study in 

complex tidal systems. 

4.4.4 Discharge-induced Mixing 

The first stage of mixing is controlled by discharge jet momentum 

and buoyancy of the effluent (see Figure 4-2). This stage generally covers most of the regulatory or near-field mixing zone. It is 

particularly important in lakes and reservoirs and slow moving 

rivers since ambient mixing in those waterbodies is minimal. 

In shallow environments, it is important to determine whether 

near-field instabilities occur. These instabilities, associated with 

surface and bottom interaction and localized recirculation cells 

extending over the entire water depth, can cause buildup of 

effluent concentrations by obstructing the effluent jet flow. There 

are no simple means to estimate dilution in these cases. Criteria 

for these instabilities and specialized predictive models have been 

developed to address these problems [13]. 

In the absence of near-field instabilities, horizontal or nearly horizontal discharges will create a clearly defined jet in the water 

column that will initially occupy only a small fraction of the 

available water depth. The following equations and models are 

designed to describe mixing under stable near-field conditions. 

1) Use of a Simplistic Screening Equation 

A minimum estimate of the initial dilution available in the vicinity 

of a discharge can be made using the following equation derived 

from information in Holley and lirka (1986) [15]: 

s=o.3 ; 

where 

5 = flux-averaged dilution 

X = distance from outlet 

d = diameter of outlet. 

75 
recelvlng waler surface 

/ 

recewmg water bottom 

c) Deep-wafer, high-buoyancy, nonvertical discharge 

recelvlng water surface 

/ 

receiving 

water 

bottom 

d) St-tallow-water. low-buoyancy, nonvertical discharge 

IFigure 42. Example of Discharge-Induced Mixing [7] 

The coefficient 0.3 represents the average of two values derived 

from the literature, 0.28 [16] and 0.32 [38]. 

The equation provides a minimum estimate of mixing because it 

is based on the assumptions that outlet velocity is zero and the 

discharge is neutrally buoyant. Dilution may be underestimated 

for partially fuH pipes because the equation assumes a fully flowing pipe. The equation can be used in inverse form to solve for 

the discharge x at which a desired solution-for example, that 

corresponding to the CMC-has been achieved. The equation is 

valid only close to the discharge, up to a distance corresponding 

to several (two to three) water depths. At longer distances, other 

factors are of increasing importance in jet mixing and must be 

included. 

Mixing graphs that include the effects of discharge buoyancy, 

ambient velocity, and stratification can be found in Halley and 

Jirka [15], Fischer et al. [16], and Wright [39]. They are useful to 

account for these other initial dilution factors and can aid in 

determining whether criteria will be met at the edge of the 

regulatory mixing zone. 

2) Use of Detailed Computer Models 

More detailed design data for the mixing zone can be obtained 

from the use of computer models based on integral jet techniques. It is important to note that most models represent an 

idealization of actual field conditions and must be used with 

caution to ensure that the underlying model assumptions hold for 

the site-specific situation being modeled. In general, these buoyant jet models require the following input data: discharge depth, 

effluent flow rates, density of effluent, density gradients in receiving water, ambient current speed and direction, and outfall characteristics (port size, spacing, and orientation). Model output 

includes the dimensions of the plume at each integration step, 

time of travel to points along the plume centerline, and the 

average dilution at each point. 

Described below are six mixing zone models that are available 

through EPA. All of the models require a user who is well versed in 

mixing concepts and the data necessary to run the models. The 

first model, CORMIX [40, 411, may be the most useful to regulators since it is an expert system that guides the user in selecting an 

appropriate modeling strategy for rivers or estuaries. It is available 

from the National Technical Information Service (NTIS), and user 

support is available from the U.S. EPA CEAM. The other models 

were developed and designed for ocean discharges. All but one 

can be used on rivers, lakes, and estuaries with appropriate input 

modifications; UPLUME is restricted to stagnant water environments where the ambient water current velocity is zero (e.g., 

lakes, reservoirs). 

These five models were designed for submerged discharges in 

oceans. They all report dilution, and all terminate execution 

when the vertical ascent of the plume is zero (e.g., when the 

plume reaches the surface or when plume density is equal to 

ambient density in some stratified systems). With the exception 

of CORMIXl, they all assume that there is a “deep” receiving 

stream (i.e., no bottom interference). They too are available from 

NTIS, and user support is provided by the U.S. EPA Hatfield 

Marine Science Center in Newport, Oregon [24]. These five 

models have been modified such that the user inputs the data 

into a universal data format that allows the user to apply any of 

the five models with only minor input changes. 

l CORMIX is a series of software elements for the analysis and 

design of a submerged buoyant or nonbuoyant discharge 

containing conventional or toxic pollutants and entering 

into stratified or unstratified watercourses, with emphasis 

on the geometry and dilution characteristics of the initial 

mixing zone. Subsystem CORMlXl deals with single-port 

discharges, and subsystem CORMIXZ addresses multiport 

diffusers. The system operates on microcomputers with the 

MS-DOS operating system. CORMlXl can summarize dilution characteristics of the proposed design, flag undesirable 

designs, give dilution characteristics at specified boundaries 

(i.e., legal and toxic mixing zones) and recommend design 

alterations to improve dilution characteristics. The CORMlXl 

program guides the user, based on the user’s input, to 

76 
appropriate analyses of design conditions and mixing zone 

dimensions. 

l UPLUME is an initial dilution model that can be used for 

stagnant waterbodies, such as lakes and reservoirs, where 

the ambient currents can be assumed to be zero. The 

model simulates a submerged single-port discharge. The 

bouyancy between the effluent and ambient water can be 

accounted for, and the discharge can be given a vertical 

angle. UPLUME calculates flux-averaged dilutions and, for 

one output option, a centerline dilution. 

l UOUTPLM can be used in flowing and stagnant waterbodies. 

The user specifies the current speed of the ambient water, 

and this speed is assumed to be constant with depth. The 

model simulates a submerged single-port discharge. Buoyancy between the effluent and ambient water can be modeled, as well as the discharge vertical angle. The ambient 

current is assumed to be perpendicular to the diffuser. 

l UMERCE is a model that can also be used for both flowing 

and stagnant waters. It has capabilities that UOUTPLM 

does not have: it considers multiple submerged ports, and 

the user can specify arbitrary ambient current speed variations with depth. The ports are assumed to be equally 

spaced. The model accounts for adjacent plume interferences over the course of the plume trajectory and in the 

subsequent dilution calculation. Positive buoyancy is accounted for, and the discharge vertical angle can be modified. The ambient current is assumed to be perpendicular 

to the diffuser. 

l UDKHDEN is a three-dimensional model that can be used 

for flowing and stagnant waterbodies. It has all the capabilities of UMERCE plus the ability to simulate instances 

where the ambient current flow is not perpendicular to the 

diffuser. 

l ULINE models a vertical slot jet discharge into a flowing 

waterbody. The discharge angle is assumed to be perpendicular to ambient current. The ambient current may vary 

with depth, and the axis of the diffuser may range from 

parallel to perpendicular to the ambient current. The buoyancy of the effluent can also be modeled. 

An evaluation and comparison of all these models can be found in 

the Technical Guidance Manual for Performing Wastelood Allocations-&ok 3, Estuaries [ 71. 

4.4.5 Ambieut-induced Mixing 

The equations for discharge-induced mixing can be used to predict concentrations in the regulatory mixing zone where strong 

jet mixing predominates over ambient mixing. Beyond this point, 

the mixing is controlled by ambient turbulence. Thus, ambient 

mixing models must be used to predict the pollutant concentration distributions up to the stage of complete lateral mixing to 

provide boundary conditions for the completely mixed fate and 

transport models described in Section 4.5. This information also 

may be needed to estimate concentrations encountered at important resource areas or at subsequent downstream dischargers. 

If there is no discharge-induced vertical mixing associated with 

the jet action of the discharge, then mixing over the depth of the 

waterbody must be accomplished by ambient mixing. For a 

neutrally buoyant, soluble effluent discharged with low velocity at 

the surfaceor at the bed of a stream, the flow distance required to 

achieve complete vertical mixing is on the order of 50 to 100 

times the depth of water in that portion of the channel where the 

effluent is discharged [42]. For a discharge that is either lighter 

(positively buoyant) or heavier (negatively buoyant) than the 

ambient water, but still has no excess momentum, the flow 

distance for mixing over the depth will be greater. In the normal 

case with a high-velocity jet designed to prevent lethality in the 

mixing zone, mixing over the depth will be accomplished primarily by jet action, and the distance required for this vertical mixing 

will be much shorter. 

In general, ambient mixing must also accomplish mixing over the 

width of a waterbody to bring the effluent to the completely 

mixed condition. For situations where the width of the zone that 

is mixed by the discharge-induced mixing is much smaller than 

the width of the river, the flow distance (X,) required to achieve 

the completely mixed condition may be estimated from an equation of the form [16]: 

x,= 

mW2u 

DY 

where 

W = width of the river 

U = flow velocity for the critical design flow 

Dy = lateral dispersion coefficient as discussed below 

m = a parameter whose value depends on the degree of 

uniformity used to define “complete mixing” and 

on the transverse location of the outfall in the 

stream. 

If completely mixed conditions are defined as a S-percent variation in concentration across the stream width, the value of m 

would be approximately 0.1 for a discharge near the center of 

river flow (not the center of river width) and approximately 0.4 for 

a discharge near the edge of the river. If, because of other 

uncertainties, a 25percent variation across the width is accepted 

as being completely mixed, then the corresponding values for m 

would be approximately 0.06 for a discharge near the center of 

river flow and approximately 0.24 for a discharge near the edge of 

the river. For a very small stream, X, may be only a few hundred 

feet; for medium and large streams, X, is normally several miles 

to several tens of miles. 

The lateral dispersion coefficient (Dy) for most rivers can be 

calculated with the following equation [16]: 

where 

Dy = 0.6 du* f 50% 

d = water depth at design flow 

u’ = shear velocity. 

The coefficient (0.6) can vary from 0.3 to above 1 .O depending 

on the type and degree of irregularity of the channel crosssections. The more straight and uniform the flow, the tower the 

77 
value; the more irregular the flow (resulting from curves, sidewall 

interference, etc.), the higher the value. Values approaching and 

exceeding 1 .O are normally associated with significant channel 

meandering [42]. The following equation for shear velocity should 

be used [16]: 

u* = (gds)‘/2 

where 

9 = acceleration due to gravity 

S = slope of the channel 

d = water depth. 

For diffusers that initially spread the discharge across a significant 

part of the river width or for cases where the discharge-induced 

mixing causes mixing across a significant part of the river width, 

the values of m and X, can be smaller than the ones indicated 

here. For distances greater than X,, the models for completely 

mixed effluents discussed in Section 4.5 can be used to calculate 

concentrations at these distances. For shorter distances, maximum concentrations can be much greater than those predicted 

by “completely mixed” models and should be estimated using 

the following equation: 

c, = CeQeW 

where 

C, = maximum pollutant concentration distance x from 

the outlet 

C, = effluent concentration 

Qe = design effluent flow 

Qs = design stream flow 

Dy = lateral dispersion coefficient 

X = distance from the outlet 

W = stream width 

U = flow velocity for the design flow. 

It should be noted that this estimate of C, is a worst-case prediction since the equation assumes no significant discharge-induced 

mixing and a neutrally buoyant effluent. A more accurate way to 

predict concentrations within this second stage of mixing is to use 

the methods of Yotsukura and Sayre [42]. To use this approach, 

however, the value of Dy and pollutant concentrations after discharge-induced mixing must be known from tracer studies and/ 

or from the use of one of the discharge-induced models. 

The PSY model can be used to predict ambient mixing in shallow, 

freshwater streams where water depth is small in proportion to 

the width. PSY is a steady-state, two-dimensional plume model 

that predicts dilution of a surface discharge into a shallow receiving water where the plume attaches to both bottom and nearshore 

[43]. Uniform vertical mixing is assumed to occur at the point of 

discharge. 

Ambient mixing is minor for lakes and reservoirs because flow 

velocity is assumed to be minimal and mixing is accomplished by 

means of the discharge momentum and buoyancy. For estuaries 

that are completely mixed with regard to salinity, the equations 

presented above can be used to estimate concentrations between 

the outlet and the point of complete mixing with a slight modification of shear velocity. The above equations will be applicable to 

only unstratified estuaries since the time required to mix across 

the estuary must be significantly less than the time required for 

the effluent to pass out of the unstratified part of the estuary, the 

time required for the effluent to pass into a segment of greatly 

changed cross-section, or the time required for the substance to 

decay. When the above equations for estuaries are used, the 

velocity of the design flow should include the velocity associated 

with the inflow of freshwater as well as the tidal velocity; thus ut, 

which is based on an average total velocity; is substituted for u in 

the equations and shear velocity becomes 

lJ* = 0.10 u t. 

The CORMlX expert system model can also be used to obtain 

predictions for the ambient-induced mixing. In addition to the 

routines for discharge-induced mixing, this model also includes 

predictiveelements that apply to ambient mixing in riverine, lake, 

or coastal situations. 

4.5 CDMPl.ElElY MIXED DtSCHARGE RECEIVING WATER 

SITIJATIDNS 

At the present time, most States and EPA Regions use steady-state 

models that assume the wastewater is completely mixed with the 

receiving waters in order to calculate WLAS for contaminants. 

This approach is appropriate for conventional contaminants where 

critical environmental effects are expected to occur far downstream from the source. WLAs for toxic chemicals require a 

different approach, however, because critical environmental conditions occur near the discharge before complete mixing with the 

receiving water occurs. Consequently, mixing analyses should be 

performed because many of these toxicants can exert maximal 

toxicity in a variety of regions spanning from the discharge point 

to significant distances downstream. 

If complete mixing occurs near the discharge point, such as in 

effluent-dominated receiving streams, then steady-state models 

may be used to calculate TMDLs. Recent EPA developments in 

the identification of critical design flows based on toxicological 

concerns provide for better use of steady-state models in calculating toxic WLAS. However, if complete mixing does not occur near 

the discharge point and the effluent plume is discernible downriver, 

then modeling techniques that can simulate and predict mixing 

conditions are more appropriate. The mixing zone models presented in the previous section may be used to define the mixing 

zone. However, they only determine the dispersion and dilution 

of the effluent and do not account for chemical or biological 

processes in the mixing zone. TMDL models are available that 

can simulate mixing processes and predict areas of maximal 

concentrations in the receiving stream based on chemical, biological, and physical processes. 

4.5. I Wastelapd Metkli~ Techuhptes 

1) Steady-State Modeling Techniques 

A steady-state model requires single, constant inputs for effluent 

flow, effluent concentration, background receiving water concentration (RWC), receiving water flow, and meteorological conditions (e.g., temperature). The frequency and duration of ambient 

concentrations predicted with a steady-state model must be assumed to equal the frequency and duration of the critical receiv78 
ing water conditions used in the model. The variability in effluent 

flows and concentrations also affects RWCs, but these effects 

cannot be predicted with constant inputs. Steady-state models 

can be improved for toxic WlAs by means of the following: 

l Using design flows that will ensure criteria compliance at 

the appropriate duration and frequency. 

l Calculating both acute and chronic WlAs. 

EPA is encouraging the States to adopt two-number aquatic life 

water quality criteria and is using them in WIA studies. Ambient 

water quality criteria have been established for numerous toxic 

pollutants. These criteria specify an acute concentration (CMC) 

and a chronic concentration (criteria continuous concentration, 

or CCC) for each toxicant, as well as durations and frequencies of 

exposure for the two concentration levels. The design flows used 

in steady-state modeling should be reflective of the CCC and 

CMC durations and frequencies. The duration of the design flow 

is based on the maximum exposure time that will prevent acute 

and chronic effects. The duration of flow is assumed to apply to 

the duration of the allowable effluent concentration or load. For 

example, if the flow used is a 7-day average value, the allowable 

load is considered to be a 7-day average. The return frequency is 

based on the number of years required for biological population 

recovery after criteria have been exceeded. Appendix D describes 

the toxicological basis for selecting receiving stream design flows 

for steady-state modeling and recommends specific design flows 

for CCC and CMC calculation of TMDLs for rivers and streams. 

In summary, there are two types of design flows, hydrologically 

based and biologically based. The hydrologically based design 

flows are those traditionally used by the States, in which the 7410 

flow is used as the CCC design flow and the 1 QlO is used as the 

CMC design flow. The biologically based method uses the 1 -day, 

3-year duration-frequency for determining the CMC design flow 

and the 4day, 3-year duration-frequency for determining the 

CCC design flow. Consequently, the biologically based design 

flows are based on specific toxicological effects of a pollutant and 

biological recovery times from localized stresses [6]. The advantages of both types, as well as how they may be calculated, also 

are described in Appendix D. 

A 4-day, 3-year biological design flow does not equate to a 443 

hydrological design flow. EPA has determined that a 443 design 

flow would result in an excessive number of water quality criteria 

exceedances. As explained in Appendix D, a hydrologically based 

7QlO will, for most streams, be similar to a biologically based 4- 

day, 3-year design flow. 

At the present time, there are no recommended toxicological 

flows for steady-state modeling of lakes, reservoirs, or estuaries. 

The design conditions recommended for these waterbodies in 

Section 4.4.2 are based on hydrological and meteorological conditions rather than on toxicological duration and frequency data. 

These conditions should be used until further guidance is provided. 

Another improvement in steady-state toxics modeling can be 

realized by performing two separate WLAS, one for the CMC and 

one for the CCC. Steady-state WIA models should be used to 

calculate the allowable effluent load that will meet the CMC at the 

acute design flow and the allowable load that will meet the CCC 

at the chronic design flow. Calculation of these values will enable 

the permit writer to calculate the more limiting long-term average 

(LTA) for the treatment system and develop permit limits protective of both WLAs (see Chapter 5). 

In addition to stream design flow, steady-state models require 

design temperature, pH, alkalinity, and hardness, depending on 

the pollutants modeled at site-specific conditions. To determine 

stream design temperature, pH, alkalinity, and hardness, a program called DESCON was developed. (See Appendix D for 

additional information.) DESCON is a computer program that 

estimates design conditions for WIA modeling. These conditions 

are based on maintaining a desired limit on the frequency of 

water quality excursions in a receiving water. DESCON considers 

the effect that daily fluctuations in stream flow and water quality 

conditions, such as temperature and pH, have on the variability of 

the capability of a receiving water to accept pollutant loadings. It 

specifically accounts for the within-year correlations observed 

between such variables as stream flow, temperature, pH, alkalinity, hardness, and dissolved oxygen. DESCON determines design 

conditions using a four-step process (see Figure 4-3): 

1) A long-term record of observed stream flows and pertinent 

water quality data are assembled or synthesized. 

2) The maximum allowable pollutant load that the receiving 

water can accept without causing a water quality excursion 

is computed for each day of this record. 

3) This synthesized record of allowable loads is searched for 

the critical load, i.e., the load whose frequency of not being 

exceeded matches the desired water quality excursion frequency. 

4) Design conditions are then derived from receiving water 

conditions realized during the period of record when the 

computed allowable load was closest to the critical load. 

DESCON provides the same advantages as continuous simulation 

by considering the joint occurrences of stream flow and other 

water quality parameters as observed in the historical record. In 

addition, it is more computationally efficient; it contains a facility 

for extracting and analyzing flow and water quality data from 

STORET; it can use both the extreme value and the biologically 

based methods of calculating of water quality excursions; and it is 

specifically designed to handle such pollutants as ammonia, heavy 

metals, pentachlorophenol, and biochemical oxygen demand 

(BOD) for which water quality criteria are functions of such design 

condition variables as temperature, pH, alkalinity, hardness, and 

dissolved oxygen. The main limitations of DESCON are that it 

requires at least 10 years of historical daily flow data and it can 

only analyze a single discharger, edge-of-mixing zone situations 

(or a simplified Streeter-Phelps dissolved oxygen response for 

BOD). 

2) Dynamic Modeling Techniques 

Steady-state modeling considers only a single condition; effluent 

flow and loading are assumed to be constant. The impact of 

receiving water flow variability on the duration for which and 

frequency with which criteria are exceeded is implicitly included 

79 
Allowable Stream Loading 

Days 

Days 

Design Temperature 

Days 

--- 

Final 

Critical 

Load 

Trial Critical Load 

1 AllOWed 

I number 

I 

Number of Excursions 

Figure 4-3. Computational Scheme for Deriving Design Conditions 

in the design conditions if these conditions reflect the desired 

toxicological effects regime. Dynamic modeling techniques explicitly predict the effects of receiving water and effluent flow and 

of concentration variability. The three dynamic modeling techniques recommended by EPAfor WLAs are continuous simulation, 

Monte Carlo simulation, and lognormal probability modeling. 

These methods calculate a probability distribution for RWCs rather 

than a single, worst-case concentration based on critical conditions. Prediction of complete probability distributions allows the 

risk inherent in alternative treatment strategies to be directly 

quantified. 

The use of probability distributions in place of worst-case conditions has been accepted practice for years in water resource 

engineering, where it was found to produce more cost-effective 

design of bridge openings, channel capacities, floodplain zoning, 

and water supply systems. The same cost-effectiveness can be 

realized for pollution controls if probability analyses are used. 

The dynamic modeling techniques have an additional advantage 

over steady-state modeling in that they determine the entire 

effluent concentration frequency distribution required to produce 

the desired frequency of criteria compliance. Maximum daily and 

monthly average permit limits can be obtained directly from the 

effluent LTA concentration and coefficient of variation (CV) that 

characterize this distribution. Generally, steady-state modeling 

has been used to calculate only a chronic WIA. Steady-state 

modeling generates a single allowable effluent value and no 

information about effluent variability. If the steady-state model is 

used to calculate both acute and chronic wasteloads, limited 

information will be provided and the entire effluent distribution 

will not be predicted. Steady-state WLA values can be more 

difficult to use in permits and enforcement because of the variable 

nature of the receiving waterbody and the effluent. The outcome 

of probabilistic modeling can be used to ensure that permit limits 

are determined based on best probability estimates of RWCs 

rather than a single, worst-case condition. As a result, maximum 

daily and monthly average permit limits, based on compliance 

80 
with water quality criteria over a 3-year period, can be obtained 

directly from the probability distribution. 

Continuous Simulation Models. As shown in Figure 44, a 

continuous simulation model uses daily effluent flows (Q) and 

concentration data (Ce) with daily receiving water flow (QJ and 

background concentration data (C,) to calculate downstream 

RWCs (44). The model predicts these concentrations in chronological order with the same time sequence as the input variables 

(Ch versus time). The daily RWCs can then be ranked from the 

lowest to the highest without regard to time sequence. A probability plot can be constructed from these ranked values, and the 

occurrence frequency of any 1 -day concentration of interest can 

be determined (Cb versus frequency). Running average concentrations for 4 days (i.e., the chronic design flow), or for any other 

averaging period, also can be computed from the daily concentrations (Figure 4-5). 

The probability plot generated by the continuous simulation model 

using existing effluent data will indicate whether criteria are predicted to be exceeded more frequently than desired. Appendix D 

discusses how to select the appropriate allowed frequency of 

excursions based on the biological recovery period required for a 

specific waterbody. If recurrence intervals of 10 or 20 years are 

desired, at least 30 years of flow data should be available to 

provide a sufficient record to estimate the probability of such rare 

events. Of the 30 years of required flow data, at least 20 to 25 

years should be continuous daily data, with the remaining years 

represented with only intermittent data. The data should be 

examined to verify that the receiving stream has not undergone 

significant hydrological modification. The data also should be 

examined to determine if there were any long-term changes due 

to technology-based treatment or periodic changes due to industrial or municipal plant closings or expansions. The same data 

requirements are also true for the lognormal probabilistic and 

Monte Carlo methods. However, except for the continuous 

simulation models, other nonsteady-state models in this section 

I I 

A c I ’ dw / I 

01 : : : : : : : ; : ’ . -4 

99 99.5 100 

Percent of Time Concenlra!~on Is Less Than or EqUal 

IO Concemt~on Plotted 

0.5 1 * 510.2 

I I I w 20 

Figure 4-4. Frequency of Occurrence of Concentrations in 

Receiving Waters and Recurrence Intervals Generated 

by a Continuous Simulation Model 

cannot be used to account for the duration and frequency provision of the two-number water quality criteria. Users are cautioned 

about the specific limitations of some of the dynamic models 

included here. Continuous simulation models have the following 

advantages compared to steady-state formulations: 

l The frequency and duration of toxicant concentrations in a 

receiving water can be predicted. 

l The cross-correlation and interaction of time-varying pH, 

flow, temperature, pollutant discharges, and other parameters are incorporated. 

l The effect that the serial correlation of daily flows and other 

parameters has on the persistence of criteria excursions is 

incorporated. 

l Long-term stream flow records for ungauged rivers using 

precipitation and evapotranspiration data can be synthesized. 

l Long simulation times can prevent the initial conditions 

used in the model from affecting the calibration of fate and 

transport processes. 

Unlike steady-state models, continuous simulation models require 

significantly more data to apply, to calibrate, and/or to verify a 

specific problem and require that input information for the application of the model be time-series data. Also, the model results 

need manipulation to calculate the effluent LTA concentration 

and CV for use in developing effluent limits. 

Monte Carlo Simulation Models. Monte Carlo simulation combines probabilistic and deterministic analyses since it uses a fate 

and transport mathematical model with statistically described 

inputs. Monte Carlo simulations have been the most frequently 

used approach in stochastic water quality studies [45-511. The 

probability distributions of effluent flow, effluent concentration, 

and other model input must be defined using the appropriate 

duration for comparison to the CMC and CCC. If 1 -day average 

RWCs must be predicted for CMC comparisons, probability distributions of daily model input data are needed for Monte Carlo 

simulation. If 4-day average concentrations must be predicted for 

CCC comparisons, the probability distributions of 4-day average 

input data are required. The computer selects input values from 

these distributions using a random generating function. The fate 

and transport model is repetitively run for a large number of 

randomly selected input data sets. The result is a simulated 

sequence of RWCs. These concentrations do not follow the 

temporal sequence that is calculated with the continuous simulation model, but they can be ranked in order of magnitude and 

used to form a frequency distribution. Monte Carlo analyses can 

be used with steady-state or continuous simulation models (521. 

The approach for calculating the allowable pollutant load distribution using Monte Carlo simulation is the same as that described 

for the continuous simulation model. The advantages of Monte 

Carlo simulation are the following: 

l It can predict the frequency and duration of toxicant concentrations in a receiving water. 

81 
It can be used with steady-state or continuous simulation 

models that include fate processes for specific pollutants. 

It can be used with steady-state or continuous simulation 

models that include transport processes for rivers, lakes, 

and estuaries. 

It can be used with steady-state or continuous simulation 

models that are designed for single or multiple pollutant 

source analyses. 

It does not require time series data. 

It does not require model input data to follow a specific 

statistical distribution or function. 

It can incorporate the cross-correlation and interaction of 

time-varying pH, flow, temperature, pollutant discharges, 

and other parameters if the analysis is developed separately 

for each season and the results are combined. 

The primary disadvantages of Monte Carlo simulation are that it 

requires more input, calibration, and verification data than do 

steady-state models, and the model results need manipulation to 

calculate the effluent LTA concentration and CV to develop effluent limits. 

Lomormal Probabilistic Dilution Model. Without resorting to 

the continuous simulation method of computing RWCs in temporal sequence, this probabilistic method uses the lognormal prob 

ability distributions of the input variables to calculate probability 

distributions of output variables [53]. As a result, the method 

requires only the relevant statistical parameters of the input vatiables (medians and coefficients of variation) rather than the actual 

time series data needed for continuous simulation, If l-day 

average RWCs must be predicted for comparisons with the CMC, 

lognormal probability distributions of daily input data are needed. 

If 4-day average concentrations must be predicted, the lognormal 

probability distributions of 4-day average input data are required. 

Because this probabilistic model cannot, as yet, incorporate fate 

and transport processes, it can be used to predict the concentration of a substance only after complete mixing and before degradation or transformation significantly alters the concentration. 

The lognormal probabilistic dilution model has the following 

advantages: 

l It can predict the frequency and duration of toxicant concentrations in riverine environments. 

l It does not require time series data, 

l It can incorporate the cross-correlation and interaction of 

time-varying pH, flow, temperature, pollutant discharges, 

and other parameters if the analysis is developed separately 

for each season and the results are combined. 

The lognormal probability dilution model has the following disadvantages: 

l It requires more input than a steady-state model. 

c b 

Time 

cb 

,- _,-,’ 

,__A-” -- 

,/’ Standard 

/’ 

L---- - - - 

Frequency 

Figure 4-5. Concentration Frequency Curves 

l It does not include instream fate processes. 

l It applies only to rivers and streams. 

l It analyzes multiple pollutant sources inaccurately. 

l It requires model input data to be lognormally distributed. 

4.5.2 Ca/cu/ath# t&s A//uwab/e Etttumt Cuncentrath 

Information concerning effluent concentration means and variabilities can be obtained from data bases on existing treatment 

plants and from development documents for specific industrial 

point source categories. This information is available from the 

Industrial Technology Division of the Office of Water Regulations 

and Standards. These effluent data can be used with dynamic 

models to determine what the effluent concentration distribution 

must be to meet water quality standards. Two possible approaches can be taken to determine this distribution regardless of 

the type of dynamic modeling technique (i.e., continuous, Monte 

Carlo, or lognormal probabilistic). One approach is based on the 

simplifying assumption that treatment will change only the mag82 
nitude of effluent concentrations; no changes are assumed to 

occur in effluent flows or in the relative variability of effluent 

concentrations. With these assumptions, no additional model 

runs are needed to determine the allowable distribution for effluent concentrations. The other approach assumes that the required effluent concentration distribution is the same as the existing distribution except that it is reduced in magnitude by whichever is greater-the percentage necessary for the 1 -day average 

concentrations to meet the CMC, or the 4-day average concentrations to meet the CCC at the desired recurrence interval. 

Chapter 5 includes details on how permit limits are derived from 

the mean and coefficient of variation of effluent concentrations 

determined from this analysis. 

The second approach for determining the allowable effluent concentration distribution is based on the assumption that effluent 

concentrations after treatment will not have the same CV as 

concentrations before treatment. Studies have documented that 

advanced secondary treatment increases the CV of BOD and total 

suspended solids concentrations compared to secondary treatment. Where feasible, investigations should be conducted to 

evaluate how treatment processes for heavy metals, organic chemicals, and effluent toxicity will change the variability of these 

constituents. The development documents mentioned above 

also provide some variability data for treatment processes. To 

account for a change in variability, an alternative approach should 

be used to determine the allowable effluent distribution. Iterative 

model runs can be performed using different concentration means 

with the effluent “future treatment” variance until a mean is 

found that meets the criteria at the desired recurrence intervals. 

These iterative model runs require stochastic generation of effluent input data since daily effluent concentrations will not be 

available for the hypothetical treatment schemes. The required 

“future treatment” mean and CV of effluent concentration can 

then be used to set permit limits (see Chapter 5). 

EPA’s Office of Water Regulations and Standards developed an 

interactive preprocessor for DYNTOX that automatically creates 

input for continuous simulation models, randomly selects the sets 

of input data required for Monte Carlo simulations, and performs 

the numerical integration calculation for the lognormal probabilistic model. DYNTOX is available from the EPA CEAM, Environmental Research Laboratory (ERL) [54]. If the observed data base 

is fairly complete but missing a few points, a linear interpolation 

scheme is used to fill in the missing data. If data are scarce, a lagone Markov method is used to generate daily data stochastically. 

The lag-one Markov method uses the mean, standard deviation, 

and daily correlation coefficient of the observed data to create 

random sequences of data having the same statistical properties. 

The interactive program is written in FORTRAN and is available for 

use on mainframe or IBM PC-compatible computers. 

Two common methods exist to calculate the return period for a 

given concentration from probabilistic modeling: the percentile 

method and the extrema method. The percentile method used 

by DYNTOX ranks a listing of all individual daily concentrations. 

The return period for a concentration is then calculated based on 

the percentile occurrence. In the extrema method, only annual 

extrema values are used in the ranking. The return periods 

calculated from these two methods are equally valid statistical 

representations. When using the percentile method, results express an average return period and multiple occurrences within 

any year. The extrema method describes the return period for an 

annual extreme and includes only the extreme of multiple occurrences within a year. 

4.5.3 6uutwalRec aPnm8twb88 for Model&?ktitm 

The reliability of the predictions from any of the modeling techniques depends on the accuracy of the data used in the analysis. 

The minimum data required for model input include receiving 

water flow, effluent flow, effluent concentrations, and background 

concentrations. In many locations, stream flow data should be 

sufficient for both steady-state and dynamic models. At least 30 

years of flow data should be available if excursions of the CMC 

and CCC must be evaluated at rare frequency of once in 10 or 20 

years. Measurements of effluent toxicity or individual toxicity can 

be much more limited. 

If only a few toxicant or effluent toxicity measurements are available, steady-state assessments should be used. Modeling also 

should be limited to steady-state procedures if a daily receiving 

water flow record is not available; however, in effluent-dominated 

situations, critical flow may be used to characterize the receiving 

stream. Appendix D describes how to select appropriate design 

flows if State regulations do not require a specific design flow for 

river WI-As. Fate and transport models or dilution calculations can 

be used for individual toxicants. At the present time, only dilution 

calculations or first-order decay equations are recommended for 

effluent toxicity analyses. Chapter 1 discusses the conservative/ 

additive assumption for toxicity. 

If adequate receiving water flow and effluent concentration data 

are available to estimate frequency distributions, one of the dynamic modeling techniques should be used to develop more 

cost-effective treatment requirements. If the effluent data exhibit 

significant seasonal differences or batch process trends, the continuous simulation approach may be the easiest dynamic modeling method to use. The best results will, of course, be obtained if 

daily effluent flows and concentrations are available for model 

input for an entire year. The lag-one Markov technique can be 

used to generate daily effluent data for the entire simulation as 

long as adequate measurements for the site-specific facility (or a 

similar one) are available to estimate a day-to-day correlation 

coefficient and to determine when seasonal or batch process 

changes in effluent quality occur. 

If adequate receiving water flow and effluent concentration data 

are available and if effluent data exhibit no seasonal or batch 

process trends, lognormal and Monte Carlo methods may be 

easier and require less computer time than the continuous simulation approach. 

The following section recommends models for toxicity and individual toxicants for each type of receiving water-rivers, lakes, 

and estuaries. Detailed guidelines on the use of fate and transport 

models of individual toxicants are included in the toxic TMDL 

guidance available from the Monitoring Branch of EPA’s Office of 

Water Regulations and Standards [5, 6, 71 and Office of Research 

and Development [55]. These manuals describe in detail the 

83 
-- --. 

--~.- - - ~. 

transport and transformation processes involved in water quality 

modeling. Transport processes include the dispersion and advection of a contaminant once it enters the receiving stream; its 

volatilization from the water; and its sorption to suspended sediment, eventual settling, and possible resuspension and diffusion 

from the sediment. Transformation processes include the oxidation, hydrolysis, photolysis, biodegradation, and bioaccumulation 

of the chemical. 

Most water quality models were developed with an emphasis on 

the dynamics in the water column and the eventual water column 

concentrations. Several models, including some of those listed 

below (EXAMS-II, WASP4) are now capable of simulating water 

column-sediment interactions (resuspension, settling, and diffusion), however, additional work needs to be completed on the 

mechanisms of sediment-water column exchange before themodels can be validated for predictive applications involving sediments. With the advent of sediment criteria in the next few years, 

it will be necessary to use models that predict concentrations in 

both receiving water and bed sediment. This will be of particular 

importance in areas where the sediments are contaminated to the 

point at which they act as the source of a pollutant to the water 

column. Table 4-2 lists and summarizes models that may be used 

for predicting the fate and transport of toxicants and that are 

supported by the EPA CEAM [56]. All the models, plus two 

bioaccumulation models, briefly are described below. 

l DYNTOX 1541 is a WLA model that uses a probabilistic 

dilution technique to estimate receiving water chemical 

concentrations or whole effluent toxicity fractions. The 

model considers dilution and net first-order loss, but not 

sorption and benthic exchange. The net loss rate must be 

determined empirically on a case-by-case basis and cannot 

be extrapolated to different conditions of flow, temperature, solids, pH, or light. 

. EXAMS-II [57] is a compartment model that can be used as 

either a steady-state or quasi-dynamic model designed for 

evaluation of the behavior of synthetic organic chemicals in 

aquatic ecosystems. It simulates a toxic chemical and its 

transformation products using second-order kinetics for all 

significant organic chemical reactions. EXAMS-II does not 

simulate the solids with which the chemical interacts. The 

concentration of solids must be user-specified for each 

compartment. The model accounts for sorbed chemical 

transport based on solids concentrations and specified transport fields. Sediment exchanges with the water column 

include pore-water advection, pore-water diffusion, and 

solids mixing. The last describes a net steady-state exchange associated with solids that is proportional to porewater diffusion. 

l WASP4 [58] is a generalized modeling framework for contaminant fate in surface waters. Based on the flexible 

compartment modeling approach, WASP4 can be applied 

in one, two, or three dimensions, given the transport of 

fluxes between segments. WASP4 can read output files 

from the link-node hydrodynamic model DYNHYD4, which 

predicts unsteady flow rates in unstratified rivers and estuaries, given variable tides, wind, and inflow. TOX14, a subset 

of WASP4, simulates up to three interacting toxic chemicals 

and up to three sediment size fractions in the bed and 

overlying waters. First- or second-order kinetics can be 

used for all significant organic chemical reactions. Sediment exchanges include pore-water advection, pore-water 

diffusion, and deposition/scour. Net sedimentation and 

burial rates can be specified or calculated. The output can 

be used with the two bioaccumulation models FCETS and 

FCMZ, which are described below. 

HSPF [59] simulates watershed hydrology and water quality 

for both conventional and toxic organic pollutants. HSPF 

incorporates the watershed-scale ARM and NPS models 

into a basin-scale analysis framework that includes transport and transformation in one-dimensional stream channels. The simulation provides a time history of the runoff 

flow rate, sediment load, and nutrient and pesticide concentrations, along with a time history of water quantity and 

quality at any point in a watershed. HSPF simulates three 

sediment types (sand, silt, and clay) in addition to specific 

Table 4-2. Toxicant Fate and Transport Models 

Model 

DY NTOX 

EXAMS-II 

WASP4 

HSPF 

SARAH2 

M lNTEQA2 

Environment 

river 

lake, river, 

estuary 

lake, river, 

estuary 

river 

river 

lake, river, 

estuary 

Time Domain 

dynamic 

steady-state, 

quasi-dynamic 

steady-state, 

dynamic 

dynamic 

steady-state 

steady-state 

Spatial Domain 

far field, 

1 -dimensional 

far field, 

3-dimensional 

far field, 

3-dimensional 

far field 

1 -dimensional 

treatment plant, 

near field, 

2-dimensional 

- 

Chemical 

organic, 

metal 

organic 

organic, 

metal 

organic, 

metal 

organic 

metal 

84 
--- ---- - --- __-----------.--.--.__~____ - -_- - -- -- ------- --.-.- _ -- - - -----___ 

organic chemicals and transformation products of those 

chemicals. The reaction and transfer processes included are 

hydrolysis, oxidation, photolysis, biodegradation, volatilization, and sorption. Sorption is modeled as a first-order 

kinetic process in which a desorption rate and an equilibrium partition coefficient for each of the three solid types 

must be specified. Resuspension and settling of silts and 

clays (cohesive solids) are defined in terms of shear stress at 

the sediment-water interface. For sands, the system’s capacity to transport sand at a particular flow is calculated 

and resuspension or settling is defined by the difference 

between the sand in suspension and the calculated capacity. Sediment exchanges with surficial benthic sediments 

are modeled as sorption/desorption and deposition/scour. 

Underlying sediment and pore water are not modeled. 

l SARAH2 [60] is a steady-state, near-field model for calculating acceptable concentrations of hazardous organic chemicals discharged to land disposal or wastewater treatment 

facilities. Acceptable leachate or treated industrial waste 

discharge constituent concentrations are estimated by a 

“back calculation” procedure starting from chemical safety 

criteria in surface water, drinking water, or fish. For steady 

or batch waste streams, SARAH.2 considers the following 

concentration reductions: dilution and loss during treatment, initial Gaussian mixing at the edge of a stream, 

lateral and longitudinal diffusion in the mixing zone, sorption, volatilization, hydrolysis, and bioaccumulation in fish. 

The user must specify appropriate concentrations for protection of the aquatic community and of humans exposed 

through consumption of fish and water. The benthic community is not presently considered. Treatment loss is handled 

empirically. SARAH2 contains data sets for three disposalwatershed scenarios that can be easily modified and employed. The model is designed for screening analysis and 

contains numerous assumptions that should be verified 

before the model is used in actual cases. 

l MINTEQA2 is an equilibrium metals speciation model for 

dilute aqueous systems [61]. It does not have any transport 

and transformation processes and must be run with one of 

the above models. It can be used to calculate the mass 

distribution at equilibrium among dissolved, absorbed, and 

solid phases and the species distribution within each phase. 

MlNTEQA2 contains a chemical component data set for 

major ions commonly found in aqueous systems (e.g., Ca, 

Fe, and S), trace metals/metalloids of pollution interest 

(e.g., Cd, Cr, Ni, Pb, and Zn), and organic ligands of 

significant affinity for metal complexation. The model can 

be used to calculate the concentrations of adsorbed metals 

via any of seven different adsorption algorithms. 

l FCETS is a toxicokinetic model that simulates the 

bioaccumulation of nonpolar organic chemicals by fish from 

both water and food [62]. Both of these routes of exchange are modeled as diffusion processes that depend 

upon physicochemical properties of the pollutant and morphological/physiological characteristics of the fish. FCETS 

contains a moderately sized data base of allometric relationships for gill morphology with which it can simulate the 

direct gill/water exchange of organic chemicals for essentially any fish species, assuming certain default values. FCETS 

also contains a limited data base of physiological/morphological relationships that are used to set parameters for food 

exchange. In addition to simulating bioaccumulation of 

organic toxicants, FCETS can calculate time to death from 

chemicals whose mode of action is narcosis. This calculation is based on the existence of a single, lethal, internal 

chemical activity for such chemicals. The concentrations of 

toxic chemical to which the food chain is exposed may be 

specified by the user or may be taken directly from the 

values calculated by the exposure concentration model 

WASP4. Thus FCETS may be executed as a separate model 

or as a postprocessor to WASP4. 

l FCM2 is a generalized model of the uptake and elimination 

of toxic chemicals by aquatic organisms (631. It generates a 

mass balance calculation in which the rates of uptake and 

elimination are related to the bioenergetic parameters of 

the species. A linear food chain or a food web may be 

specified. Fish tissue concentrations are calculated as a 

function of time and age for each species included. Exposure to the toxic chemical in food is based on a consumption rate and predator-prey relationships that are specified 

as a function of age. Exposure to the toxic chemical in 

water is functionally related to the respiration rate. Steadystate concentrations also may be calculated. The concentrations of the toxic chemical to which the food chain is 

exposed may be specified by the user or may be taken 

directly from the values calculated by the exposure concentration model WASP4. Thus FCM2 may be executed as a 

separate model or as a postprocessor to WASP4. Migratory 

species, as well as nonmigratory species, may be considered. Separate nonmigratory food chains may be specified, 

and the migratory species is exposed sequentially to each 

food chain based on its seasonal movements. 

To apply the steady-state, continuous simulation, or probabilistic 

methods to effluent toxicity modeling, the percent effluent measurements should be converted to toxic units (TUs). As discussed 

in Chapters 1, 2, and 3, it is necessary to convert toxicity to units 

that can be directly related to mass. When comparing toxicity 

among chemicals, the relationship between toxicity and concentration is inverse; chemicals that have toxic effects at low concentrations have a greater “toxicity” than chemicals that have toxic 

effects at higher concentrations. The modeling of toxic effluents 

is based on mass balance principles; therefore, toxicity needs to 

be in units that increase when the percent of the effluent of the 

receiving stream increases. Thus, a TU is the reciprocal of the 

dilution that produces the test endpoint, i.e., acute toxicity endpoint (ATE) or chronic toxicity endpoint (CTE). An acute toxic 

unit (TU,) is the reciprocal of an ATE. A chronic toxic unit (TU,) is 

the reciprocal of a CTE. The TMDL must ensure that the CMC 

and the CCC are met in the receiving water at the desired 

duration and frequency. The CMC for toxicity is recommended 

as 0.3 TU,. This is a value that should prevent lethality unless the 

duration of exposure exceeds 1 hour. 

The CCC for toxicity measured with chronic tests is recommended 

as the following: 

CCC = 1 .O TU,. 
The first step in the TMDL process is to calculate the allowable 

acute effluent toxicity that meets the CMC in the receiving water 

at the duration and frequency discussed in Appendix D. 

The next step in the TMDL process is to calculate the allowable 

chronic effluent toxicity that meets the CCC in the receiving water 

at the duration and frequency discussed in Appendix D. To 

compare the allowable acute toxicity value to the allowable chronic 

toxicity value, the numbers must be converted to the same units 

as follows: 

TU, = (ACR)(TlJ,) 

where the acute-to-chronic ratio (ACR) is determined from tests 

on the effluent. It is important that the ACR used for TMDL 

purposes be based on actual data and not be assumed to be 10 or 

20, as in the screening procedure (Chapter 3). The value of this 

ratio will influence whether the acute or chronic TMDL is more 

stringent and is used to calculate the permit limit using the 

methods described in Chapter 5. 

At the present time, the fate of effluent toxicity in a receiving 

water is not fully understood. Even if a decay rate for toxicity can 

be measured on a given day in a site-specific situation, there is no 

way as yet to know how this rate is affected by temperature, pH, 

or other environmental conditions. There is also no way to know 

how this rate may change when new treatment is installed. 

lnstream measurements of toxicity should be made at least once 

per season to identify any time-varying trends in site-specific fate 

processes. These monitored decay rates can then be used in 

steady-state or continuous simulation fate and transport models 

to predict receiving water toxicity, assuming that the rates will not 

change with future treatment. 

Without specific information concerning the persistence of toxicity, it is recommended that effluent toxicity be limited to dilution 

estimates and that toxicity be assumed to be additive and conservative. Toxicity is expected to be additive even when the toxicity 

of one effluent affects selected biota while the toxicity of a downstream discharge affects different biota. For rivers and run-of-river 

reservoirs with a detention time of less than 20 days, the following 

ditution equation should be used, assuming completely mixed 

conditions: 

where 

c = 

c, = 

Qs = 

c, = 

Qe= 

c = C@s f C&e 

Qe + Qs 

downstream concentration (TU, or TU, ) 

upstream concentration (TU, or TU, ) 

upstream flow (cfs) 

effluent concentration (TV, or TU, ) and 

effluent Row (cfs). 

For multiple dischargers, this equation must be applied sequentially to find the concentration as a function of distance downstream. The equation can be used for a steady-state analysis if Qs 

is set equal to the design flow, & is set equal to the historical 

plant flow, and C, is calculated to meet the CMC and CCC. This 

equation can also be used with the continuous simulation, lognormal probabilistic, or Monte Carlo methods. For these dynamic analyses, a series of C, Qe, C, and Qs values would be 

used. 

If instream toxicity measurements are available and a first-order 

decay rate for toxicity can be estimated, the following equation 

should be used: 

c = C,e-K(X/U) 

where 

C = downstream concentration (TU, or TU,) 

Co = concentration after the point source discharge has 

mixed completely with the river (TU, or TU,,) 

X = distance downstream of complete mix point 

U z velocity of river 

K = measured decay rate. 

Additional statistical approaches are available that might provide 

better statistical fits to the available data. However, these models 

are somewhat more limited than the example provided above. 

The same equations used for toxicity analyses in rivers can also be 

used in steady-state, continuous simulation, or probabilistic analysis of long, narrow, shallow impoundments with high inflow 

velocities. Wider, deeper lakes require more complicated analyses since prolonged detention times (>20 days) and stratification 

exert a significant impact on water quality. The prolonged detention times make it essential that receiving water measurements of 

toxicity be available to estimate decay factors. These measurements should be made at least once per season to identify any 

time-varying trends in toxicity fate processes. Steady-state or 

continuous simulation fate and transport models for lakes can 

then be run with monitored decay rates for toxicity. A simple 

steady-state analysis can be performed using the following equations [64]: 

where 

T, = 

v = 

Q = 

c = 

Gin = 

K = 

T,V = V/Q 

C = Ci,/(l +TwK) 

mean hydraulic residence time 

lake volume at design conditions 

mean total inflow rate at design conditions 

steady-state lake concentration (TV, or TU,) 

steady-state inflow concentration (TU, or TU,) 

first-order decay rate. 

If effluent is discharged into a stratified lake and mixes only with 

the hypolimnion or epilimnion, the volume of the layer should be 

used only to calculate mean hydraulic residence time (T,,.,). The 

mean total inflow rate (Q) and the inflow concentration (Gin) 

should be calculated as the sum of all sources to the lake, including point source, nonpoint source, and tributary inputs. 

Dilution calculations for effluent toxicity discharges to an estuary 

are complicated by the oscillatory motion of the tides and possible stratification of the estuary. The prolonged detention times 

make it essential that field measurements of toxicity be available 

to estimate decay factors. These measurements should be made 

at least once per season to identify any time-varying trends in 

toxicity rate processes. Steady-state or continuous simulation fate 

and transport models for estuaries can then be run with monitored decay rates for toxicity. A simple steady-state analysis can 

be performed using the following equations for each 

nonconservative pollutant entering from the river at the head of 

an estuary [64]: 

86 
where 

(fi> 

Bi = 1 _ (li$e’kt 

ri = exchange ratio for segment i as defined by modified 

tidal prism method 

t = flushing time 

fi = fraction of freshwater in segment i 

Ci = nonconservative pollutant concentration in segment 

i (TU, or TU,) 

k = decay rate of pollutant. 

The following equations should be used for each nonconservative 

pollutant entering along the side of an estuary: 

For segments downstream of outfall: 

fi ri 

ci = 'O A c 1- (l-ri)e-kt 

i= 1 

For segments upstream of outfall: 

where 

Ci = 

co = 

ri = 

; I 

fb = 

Si = 

so = 

k = 

t = 

nonconservative pollutant mean concentration in 

segment i (TU, or TU,) 

nonconservative pollutant mean concentration in 

segment of discharge 

exchange ratio for segment i as defined by the 

modified tidal prism method 

number of segment away from outfall 

fraction of freshwater in segment i 

fraction of freshwater in segment with discharge 

salinity in segment i 

salinity in segment of discharge 

decay rate 

flushing time. 

The details of how to calculate exchange ratios and flushing times 

for estuaries are included in Part 2 of EPA’s water quality assessment manual [64]. This manual also describes how to perform 

these calculations for stratified estuaries using a two-dimensional 

box model analysis. 

4.6 HUMAN HEALTH 

4.6.1 Mman kal# considerations 

Human exposure to pollutants should be evaluated as completely 

as available information will allow. Exposure information is used 

in calculating the human health reference ambient concentration 

(RAC) from the formulas in Chapter 2, Water Quality Standards. 

This information should be used to estimate exposures due to fish 

consumption and drinking water ingestion, background concentrations, and other exposure routes, such as recreational, occupational, drinking water, dietary (other than fish), and inhalation. 

Factors in the formulas for which information is not available can 

be omitted from the calculation. If States choose, bioaccumulation 

factors also can be modified. 

4.6.2 lbtem?inim.gthe TMDLBasedonlllurnanhaM~ Rdants 

TMDLs are typically necessary only where mixing is allowed. 

Mixing zones are used at the discretion of the States. If a State 

does not allow a mixing zone or the assumption of complete 

mixing, then the RAC is applied at the end of pipe and no TMDL 

determination is typically necessary. 

With persistent or bioconcentlatable pollutants, special mixing 

zone considerations apply. Bioconcentratable pollutant criteria 

exceedances within the mixing zone can potentially result in 

tissue contamination of organisms directly or indirectly through 

contamination of bed sediments with subsequent incorporation 

into the food chain. For discharge situations with incomplete 

mixing (e.g., large rivers, lakes, estuaries, oceans), States need to 

carefully consider whether mixing zones for persistent or 

bioconcentratable pollutants are appropriate. Where a mixing 

zone is allowed, one TMDL should be calculated to achieve the 

RAC or criterion selected above [65]. Because most human health 

criteria are chronic only, a TMDL to protect against acute effects 

will usually not be needed, although EPA’s Office of Drinking 

Water does have acute criteria for some pollutants. 

For the purpose of the following discussion, use of simple, steadystate dilution models is assumed. l-lowever, these models may be 

inappropriate for certain situations where sediments serve as a 

sink for bioconcentratable pollutants and where additional factors 

need to be considered. Dynamic models, where available, are 

useful tools for accounting for an array of variables that may have 

an impact on the fate of bioconcentratable pollutants in the food 

chain. These models may be used by States for surface waters in 

appropriate instances. 

In simple situations, the TMDL is determined from the RAC and 

the design flow of the receiving water. In more complicated 

situations, e.g., where mixing is not rapid or where lakes or 

estuaries are involved, a spatial averaging scale must be chosen. 

Selection of the spatial scale must be consistent with reasonable 

assumptions about the behavior of aquatic organisms and the 

target human population. 

In some cases, it may be necessary to apply the chronic human 

health criterion within a mixing zone if it is reasonable to assume 

that the bioconcentrating aquatic organisms have little mobility, 

thus spending most of their time within the mixing zone; and the 

target human population consistently consumes fish from the 

mixing zone (over a 70-year lifetime, for carcinogenic risks). 

The procedure for developing TMDLs/WLAs generally requires 

determining values for the following parameters, based upon 

water quality considerations: (1) the duration of the averaging 

period applicable to the WlA; (2) design considerations, e.g., 

flow; (3) the discharge &VIA) concentration that will result in 

meeting the ambient water quality criterion during the design 

condition; and (4) the allowable probability (or frequency) of the 

discharge’s exceeding the WlA, averaged over the appropriate 

87 
duration. The technical basis for setting these values is discussed 

in the following sections. 

1) Averaging Periods 

The duration of the averaging period for the WLA should be 

selected to be consistent with the assumptions used to derive the 

water quality criteria. Two categories of pollutants should be 

recognized: carcinogens and noncarcinogens. 

The human health criteria for carcinogens are derived assuming 

lifetime exposure. The upper-bound risk is directly proportional 

to the lifetime arithmetic mean dose. The criteria thus apply to 

the ambient water concentrations averaged over a 70-year period. 

The duration of exposure assumed in deriving criteria for 

noilcarcinogens may be ambiguous, particularly where a criterion 

is derved from animal studies. Furthermore, the duration may be 

highly variable, ranging as high as 20 to 30 years for cadmium. 

2) Dilution Design Conditions 

u) Carcinogens: River and Stream Discharge Situations 

In well-mixed situations, the RWC, C, is determined by the pollutant load, W (mass/time), and the combined receiving water plus 

effluent flow, Q, such that, C = W/Q. 

The long-term harmonic mean flow IS recommended as the 

design flow for carcinogens. The recommendation of long-term 

harmonic mean flow has been derived from the definition of the 

human health criteria (HHC) for carcinogenic pollutants. The 

adverse impact of carcinogenic pollutants is estimated in terms of 

receptors (human) lifetime intakes. To be within the acceptable 

level of life-time body-burden of any carcinogen, such intakes 

should not exceed the HHC during the average life-time of the 

receptor. A life-time for exposure to carcinogenic pollutants is 

defined as 70 years, or approximately 365 (days/year) multiplied 

by 70 years. 

The HHC for carcinogenic pollutants can be numerically expressed 

as: 

HHC = C (design) = (Cl + C2 + C3 + ----- + C, )/n 

where 

n = (365 days/year) x 70 years 

C = concentrations 

Based on an assumption of a constant daily load from a treatment 

facility, the fully mixed instream concentration will go up or down 

inversely with the ups and downs of receiving water flows. Therefore, instream concentration is a function of, and inversely proportional to, the streamflow downstream of the discharge. Using 

this concept, 1 /Q can be substituted for C, as follows: 

1 /Q (design) = (1 /QT + 1 /Q2 + 1 /Q3 + ----- + 1 /Qn)/n. 

The stream design flow (Q design) can then be shown as follows: 

Q (design) = ni(1 /QT + 1 IQ2 + 1 tQ3 + ----- + 1 /Qn) 

The harmonic mean is expressed as follows: 

Q (design) = n/k (1 /Qi) 

i=l 

where 

n = the number of recorded flows. 

The harmonic mean is always less than the arithmetic mean. The 

harmonic mean is the appropriate design flow for determining 

long-term exposures using steady-state modeling of effluents. 

The arithmetic mean flow is not appropriate as the design flow 

since it overstates the dilution available. Extreme value statistics 

(such as 7410 or 30QS) are also not appropriate since they have 

no consistent relationship with the long-term mean dilution. 

However, for situations involving seasonably variable effluent discharge rates, hold-and-release treatment systems, and effluentdominated sites, the harmonic mean may not be appropriate. In 

these cases, the effluent load and downstream flow are not independent (i.e., they are correlated). Modeling techniques that can 

calculate an average daily concentration over a long period of 

time are more appropriate to determine the long-term exposure 

in these cases. 

The harmonic mean flow may be estimated by any of several 

methods [8], assuming that flows are approximately lognormally 

distributed: 

where 

Qhrn = 

Qgl-l12 

Q 

am 

q, is the geometric mean flow 

Q am is the arithmetic mean flow. 

For U.S. Geological Survey flow records, summaries of the statistical parameters needed to estimate the harmonic mean can be 

quickly obtained from STORET, through a user-friendly procedure 

for permit writers, as described in Appendix D. 

WQA8 DFLOW is a software package available for computation 

of harmonic mean flow. The DFLOW program (as discussed 

below and described in Appendix D) should be used with data 

that are not lognormally distributed. 

To develop some quantitative sense of how a long-term harmonic 

mean flow of any stream compares with its 7410 flow, the 

Assessment and Watershed Protection Division and the Risk Reduction Engineering Laboratory at Cincinnati, Ohio, analyzed 

flow records of 60 streams selected at random throughout the 

United States. These are the same stream flow records that had 

been analyzed for stream design flow condition for aquatic life 

protection as listed in EPA guidance [8]. Based on the long-term 

harmonic flow and 7-day, 1 O-year low-flow estimates for these 60 

streams, the long-term harmonic mean flows of all 60 streams 

were equal to or greater than two times the 7410 low flow. Fiftyfour of the streams’ harmonic mean flows were equal to or 

greater than 2.5 times their 7QlO low flows. Finally, 40 of the 60 

streams’ harmonic mean flows were equal to or greater than 3.5 

times the 7410. 

Based on the above observations, permit authorities may choose 

a multiplication factor of 3 x 7QlO to estimate stream design flow 

for human health protection for carcinogenic pollutants. How88 
ever, it is recommended that the harmonic mean flow be calculated directly from the historical daily flow record, if possible. 

Alternatively, the following equation might be used to estimate 

harmonic mean flow [66]: 

Qhm = [l .194 l (Qam)0.473] l [(7Ql O)“.552], rz = 0.99. 

In this equation, Qam and 7QlO are estimated using the U.S. 

Geological Survey computer program, FLOSTAT. 

b) Noncarcinogens: River and Stream Discharge Situations 

The choice of average period represents a level-of-protection 

consideration inherent in the risk management decision to be 

made by the permitting agency. If a short-term duration of 

exposure is chosen (i.e., 90 days or less), design flows may be 

appropriately based on extreme value statistics. Because the 

effects from noncarcinogens are more often associated with shortened exposures, EPA suggests the use of 3045. However, in the 

comparisons of flows for smaller rivers (i.e., low flow of 50 cfs), the 

3045 flow was, on the average, only 1 .l times that of the 7QlO. 

For larger rivers (i.e., low flow of 600 cfs), the factor was, on the 

average, 1.4 times. If the effects from certain noncarcinogens 

are manifested after a lifetime of exposure, then a harmonic 

mean flow may be appropriate. 

3) Point of Application of the Criteria 

The point at which the chronic criteria are to be met in the 

receiving water may be fixed by existing State standards or may 

be determined by considerations for managing individual and 

aggregate risks. The several possibilities include the following: 

l Where State standards allow no mixing zone and no spatial 

averaging, the criterion would be met at the end of the 

pipe. 

l Where State standards specify that the criterion must be 

met at the end of the mixing zone, the criterion would be 

applied at that point. 

l Where State standards allow consideration of spatial averaging, the criterion may be met as an average within a 

specified area, as appropriate for the individual and aggregate risk scenarios underlying the application. 

89 
- 

CRAMER 4 

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Document. U.S. EPA, Office of Water Program Operations, 

Washington, DC. EPA 430/9-82-011. 

23. Ford, D.E., and L.S. johnson. 1986. An Assessment of Reservoir Mixing Processes. Technical Report 86-7 prepared by 

Ford, Thornton, Norton and Associates, ltd., Little Rock, 

AR, for the U.S. Army Engineer Waterways Experiment 

Station, Vicksburg, MS. 

24. Muellenhoff, W.P., A.M. Soldate, D.J. Baumgartner, M.D. 

Schuldt, L.R. Davis, and W.E. Frick. 1985. Initial Mixing 

Characteristics of Municipal Ocean Discharges: Volume I, 

Procedures and Applications. U.S. EPA, Environmental Research Laboratory, Narragansett, RI. PB 86-l 37478. 

25. Hubbard, E.F., F.A. Kilpatrick, L.A. Martens, and J.F. Wilson. 

“Techniques of Water-Resources investigations of the 

USGS.” Measurement of Time of Travel and Dispersion in 

Streams by Dye Tracing. Book 3, A@‘cation of Hydraulics. 

U.S. Department of the Interior, Reston, VA. 

26. Kilpatrick, F.A., and E.D. Cobb. “Techniques of Water-Resources Investigations of the USGS.“ Measurement of Discharge Using Tracers. Book 3, Application of fiydmulics. U.S. 

Department of the Interior, Reston, VA. 

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27. Wilson, J.F., E.D. Cobb, and F.A. Kilpatrick. “Techniques of 

Water-Resources Investigations of the USGS.” fluorometric 

Procedures for Dye Tracing. Book 3, Application of Hydraulics. 

U.S. Department of the Interior, Reston, VA. 

28. Fischer, H.B. 1968. Methods for Predicting Dispersion Coefficients in Natural Streams with Applications to Lower Reaches 

of the Green and Duwamish Rivers, Washington. U.S. Geological Survey Professional Paper 582-A. U.S. Geological 

Survey, U.S. Department of the Interior, Reston, VA. 

29. Kisiel, CC. 1964. Dye Studies in the Ohio River. Proceedings 

of Symposium on Diffusion in Oceans and Fresh Waters. 

Columbia University. 

30. Yotsukura, N., H.B. Fisher, and W.W. Sayre. 1970. Measurement of Mixing Characteristics of the Missouri River Between 

Sioux City, Iowa, and Plattsmouth, Nebraska. Geological 

Survey Water-Supply Paper 1899-G. U.S. Geological Survey, U.S. Department of the Interior, Reston, VA. 

31, Wilson, J.F., E.D. Cobb, and N. Yotsukura. 1969. Movement 

of o Solute in the Potomac Estuary ut Washington, D.C. ot 

Low Flow Conditions. U.S. Geological Survey Circular 5298. 

U.S. Geological Survey, U.S. Department of the Interior, 

Reston, VA. 

32. Hetling, L.]., and R.L. O’Connell. 1966. A Study of Tidal 

Dispersion in the Potomac River. Water Resources Research 

2(4): 825. 

33. Baily, T.E. 1966. Fluorescent-Tracer Studies of an Estuary. ). 

of the Water Pollution Contra/ Federation 38(12):1988. 

34. Fischer, H.B. 1972. A Longrangian Method for Predicting 

Pollutant Dispersion in Bolinas lagoon, Morin County, California. Dispersion in Surface Water. U.S. Geological Survey 

Professional Paper 582-B. U.S. Geological Survey, U.S. 

Department of the Interior, Reston, VA. 

35. Cracker, P.A., G. Morrison, S.R. Rives, F.S. Shipley, R. Palachek, 

D. Neleigh, and S. Schimmel. 1989. Case Study to 

Evaluate Effluent and Receiving Water Toxicity for an Estuarine Discharger: U.S. Naval Air Station, Corpus Christi, Texas. 

Technical Section, Water Quality Management Branch, 

U.S. EPA, Region 6. 

36. Johnson, M.C. 1984. fluorometric Techniques for Tracing 

Reservoir Inflows. Instruction Report E-84-l. U.S. Army 

Engineer Waterways Experiment Station, Vicksburg, MS. 

37. Dettmann, E.H., J.F. Paul, 1.5. Rosen, and C.J. Strobel. 1988. 

Transport, fate, and Toxic Effects of a Sewage Treatment 

Plant Effluent in a Rhode Island Estuary. U.S. EPA, Environmental Research Laboratory, Narragansett, RI. EPA/600/X87-366. 

38. Albertson, M.L., Y.B. Dai, R.A. Jensen, and H. Rouse. 1950. 

Diffusion of Submerged Jets. Trans. Am. Sot. Civ. Eng. 

(15):639-64. 

39. Wright, S.]. 1977. Mean Behavior of Buoyant Jets in a 

Crossflow. 1. of Ihe Hydruulics Division. Proc. ASCE, volume 103. 

40. Donekar, R.L., and G.H. lirka. 1988. CORMIX 1. An Expert 

System for Mixing Zone Analysis of Toxic and Conventional 

Single Port Aquatic Discharges. U.S. EPA, Environmental 

Research Laboratory, Athens, GA. PB 88-220 S&+/AS. 

41. Akar and Jirka. 1990. An Expert System for Mixing Zone 

Analysis of Toxic and Conventional Single Port Aquatic Discharges (CORMIXZ). Report under peer review. 

42. Yotsukura, N., and W.W. Sayre. 1976. Transverse Mixing in 

Natural Channels. Water Resources Research 12(4): 695- 

704. 

43. Pailey, P.P., and W.W. Sayre. 1978. Model for ShoreAttached Thermal Plumes in Rivers. I. of the Hydraulics 

Division. Proc. ASCE, volume 104. 

44. Danigan, A., and W. Waggy. 1984. Simulation-A Tool for 

Water Resource Management. Water Resources Bulletin 

1 O(2). 

45. Tiwari, J.L., and J.E. Hobbie. 1976. Random Differential 

Equations as Models of Ecosystems: Monte Carlo Simulation Approach. Math Bioscience, volume 28. 

46. Malone, R.F., D.S. Bowles, W.J. Grenney, and M.P. Windham. 

1979. Stochastic Analysis of Water Quality. Utah State 

University. UWRL/Q-79/01. 

47. Hornberger, G.M., and R.C. Spear. 1980. Eutrophication in 

Peel Inlet-l. The Problem-Defining Behavior and a 

Mathmatical Model for the Phosphorous Scenario. Water 

Research 14:29-42. 

48. Ford, D.E., K.W. Thornton, A.S. Lessem, and J.L. Norton. 

1981. A Water Quality Management Model for Reservoirs. 

ASCE Proc. of Symposium on Surface Water Impoundments, June 2-5, 1980, at Minneapolis, MN. 

49. Scavia, D., W.F. Powers, R.P. Canale, and J.L. Moody 1981. 

Comparison of First-order Error Analysis and Monte Carlo 

Simulations in Time-dependent Lake Eutrophication Models. Water Resources Research 17(4):1051-59. 

50. Rose, K.A., J.K. Summers, and R.I. McLean. 1989. A Monte 

Carlo Bioaccumulation Model of Silver Uptake by the Oyster, Crassostrea Virginia. Ecol. Mod. 42:l O-21. 

51. Versar. 1989. Prediction of Uncertainties Associated with 

Water Qua/ity Models. U.S. EPA, Office of Research and 

Development and Office of Toxic Substances. Washington, DC. 

52. Thornton, K.W., D.E. Ford, and J.L. Norton. 1982. Use of 

Monte Carlo Water Qua/iv Simulations to Evaluate Reservoir 

Operational and Managemenl Alternatives. Presented at 

Third International Conference on State-of-the-Art Ecological Modeling, Colorado State University. 

53. DiToro, D. 1984. Probability Model of Stream Quality Due 

to Runoff. /. of Environ. Engineering. Proc. ASCE, vol. 110, 

no. 3. 

54. Limno Tech, Inc. 1985. Dynamic Toxics Wasteload Allocution 

Mode/(DYNTOX): User’s Manual. U.S. EPA. Office of Water 

Regulations and Standards, Washington, DC. 

91 
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55. McKeon, T.J., and ].I. Segna. 1987. Selection Criteria for 

Mathematical Models Used in Exposure Assessments Surface 

Water Models. U.S. EPA, Office of Health and Environmental Assessment, Washington, DC. EPA/600/8-87/042. 

56. Ambrose, R.B., and T.O. Barnwell. 1989. Environmental 

Softwore at the U.S. EPA Center for Exposure Assessment 

Modelrng. U.S. EPA, Environmental Research Laboratory, 

Athens, GA. 

57. Burns, L.A., and D.M. Cline. 1985. Exposure Analysis Modeiing System Reference Manual for EXAMS II. U.S. EPA, Environmental Research Laboratory, Athens, GA. EPA 600/3- 

85-038. 

58. Ambrose, R.B., T.A. Wool, J.P. Connolly, and R.W. Shantz. 

1988. WASP4, A Hydrodynamic and Water Quality Mode/- 

Model Theory, User‘s Manual, and Programmer’s Guide. U.S. 

EPA, Center for Exposure Assessment Modeling, Athens, 

GA. EPAf600/3-87f039. 

59. Johansen, R.C., J.C. Imhoff, J.L. Kittle, and AS. Donigian. 

1984. Hydrologic Simulation PrograHORTRAN (HSPF): 

Users Manual for Release 8.0. U.S. EPA, Environmental 

Research Laboratory, Athens, GA. EPA/600/3-84-066. 

60. Vandergift, S.B., and R.B. Ambrose. 1988. SARAH2 A Near 

Field Exposure Assessment Model for Surface Waters. U.S. 

EPA, Environmental Research Laboratory, Athens, GA. EPA/ 

600/3-881020. 

61. Brown, D.S., and 1.0. Allison. 1987. MINTEQAl Equilibrium 

Metal Speciation Model: A Users Manual. U.S. EPA, Environmental Research Laboratory, Athens, GA. EPA/BOO/ 

3-871012. 

62. Barber, M.C., L.A. Suarez, and R.R. Lassiter. 1988. FCETS 

(Food and Gill Exchange for Toxic Substances): A Simulation 

Model for Predicting Bioaccumulation of Nonplor Organic 

Pollutants by fish. U.S. EPA, Center for Exposure Assessment Modeling, Athens, GA. EPA/600/3-87/038. 

63. Connolly, J,P., and R.V. Thomann. 1985. WASTOX, A Fmmework for Modeling the Fate of Toxic Chemicals in Aquatic 

Environments. Port 2: food Chain. U.S. EPA, Environmental 

Research Laboratory, Gulf Breeze, FL. EPA/600/4-85/040. 

64. Mills, W-B., I.D. Dean, D.B. Porcella, S.A. Cherini, R.J.M. 

Hudson, W.E. Frick, G.L. Rupp, and G.L. Bowie. 1982. 

Water Quality Assessment: A Screening Procedure for Toxic 

and Conventional Pollutants. U.S. EPA, Environmental Research Laboratory. Athens, GA. EPA 600/6-82-004 

a and b. 

65. U.S. EPA. 1987. Permit Writers Guide to Water Quo&y-based 

Permitting for Toxic Pollutants. U.S. EPA, Office of Water 

Enforcment and Permits, Washington, DC. EPA/440/4-87/ 

005. 

66. Rossman, L.A. Design Stream Flows Based on Harmonic 

Means. /, of Hydraulic Engineering 116(7):946-50. 

92 
5. PERMIT REQUIREMENTS 

5.1 INTRODUCTION 

As the final step in the “standards-to-permits” process, development of permit requirements is often the culmination of the 

activities discussed in the preceding chapters. This chapter 

describes the basic principles of effluent variability and permit 

limit derivation and provides recommendations for deriving limits 

from various types of wasteload allocation outputs such that 

water quality standards are protected. It also addresses important 

considerations in the expression of limits and other types of 

permit requirements, including toxicity reduction evaluations. 

The first portion of the chapter deals principally with aquatic life 

protection. Permitting for protection of human health is found in 

Section 5.4.4. 

5.1.1 Regulatory Requirements 

There are both mandatory and discretionary elements associated 

with the development of water quality-based permit limits to 

control toxic pollutants and toxicity. The mandatory elements are 

described in the revisions to the National Pollutant Discharge 

Elimination System (NPDES) Surface Water Toxics Control Program regulations (54 FR 23868, June 2, 1989). The regulations at 

40 CFR 122.44(d)(1) require that regulatory authorities first determine whether a discharge causes, has the reasonable potential to 

cause, or contributes to an excursion above water quality standards (narrative or numeric). In making these determinations, 

regulatory authorities must use a procedure that accounts for 

effluent variability, existing controls on point and nonpoint sources 

of pollution, available dilution, and (when using toxicity testing) 

species sensitivity. Each of these regulations were previously 

discussed in Chapter 3. 

There is a degree of flexibility in the specific procedures a regulatory authority uses in determining whether an excursion occurs or 

is reasonably expected to occur and in the weight given to the 

various factors in conducting the evaluation of a specific discharger. The Environmental Protection Agency’s (EPA) guidance 

for making these determinations is contained in the recommendations in Chapter 3. 

There are also several EPA policies that reflect these regulatory 

requirements, including the “National Policy for the Development 

of Water Quality-Based Limits for Toxic Pollutants” (Appendix B2) and EPA’s “Whole Effluent Toxicity Permitting Principles and 

Enforcement Strategy,” (Appendix B-4). This strategy states that 

“all major permits and minors of concern must be evaluated for 

potential or known toxicity (chronic or acute if more limiting).” In 

addition, the strategy states that “[final whole effluent toxicity 

limits must be included in permits where necessary to ensure that 

State Water Quality Standards are met. These limits must properly account for effluent variability, available dilution, and species 

sensitivity.” 

There is an element of judgment inherent in the specific permit 

limit derivation procedures used for an individual discharger once 

a decision has been made to develop a specific type of limit. 

Case-specific considerations will usually dictate the most appropriate approach to be taken in individual situations. Nevertheless, 

the various assumptions used in the permit limit development 

process should be consistent with the assumptions and principles 

inherent in the effluent characterization and exposure assessment 

steps preceding permit limit development. The permit limit 

derivation procedure used by the permitting authority should 

be fully enforceable and should adequately account for effluent variability, consider available receiving water dilution, 

protect against acute and chronic impacts, account for compliance monitoring sampling frequency, and protect the 

wasteload allocation (WLA) and ultimately water quality standards. To accomplish these objectives, EPA recommends that 

permitting authorities use the statistical permit limit derivation procedure discussed in Section 5.4 with the outputs from 

either steady state or the dynamic wasteload allocation modeling. 

5.2 BASIC PRINCIPLES OF EFFLUENT 

VARIABILITY 

An understanding of the basic principles of effluent variability is 

central to water quality-based permitting. Many of the concepts 

are the same as those considered in the development of technology-based limits. However, the process for applying the principles is substantially different, as explained below. 

5.21 Variations in Effluent Quality 

Effluent quality and quantity vary over time in terms of volumes 

discharged and constituent concentrations. Variations occur due 

to a number of factors, including changes in human activity over 

a 24-hour period for publicly owned treatment works (POTWs), 

changes in production cycles for industries, variation in responses 

of wastewater treatment systems to influent changes, variation in 

treatment system performance, and changes in climate. Very few 

effluents remain constant over long periods of time. Even in 

industries that operate continuous processes, variations in the 

quality of raw materials and activities, such as back-washing of 

filters, cause peaks in effluent constituent concentrations and 

volumes. 

If effluent data for a particular pollutant or pollutant parameter for 

a typical POTW are plotted against time, the daily concentration 

variations can be seen (see Figure 5-1, left-hand graphs). This 

behavior can be described by constructing frequency-concentration plots of the same data (see Figure 5-1, right-hand graphs). 

93 
Figure 5-1. Data Relative Frequency Distributions for Ceriodaphnia Toxicity, Daphnia Toxicity, 

and Zinc Concentrations for Three Different Effluents 

94 
5.2.2 Statistical Parameters and Relathnsh(p to Pemdt 

Lhdts 

Based upon the shape of the curve of a frequency-concentration 

plot, the data can be described in terms of a particular type of 

statistical distribution. The choices for statistical distributions 

include normal (bell-shaped), lognormal (positively skewed), or 

other variations on the lognormal distribution. From the vast 

amount of data that EPA has examined, it is reasonable to assume 

(unless specific data show otherwise) that treated effluent data 

follow a lognormal distribution. This is because effluent values 

are non-negative and treatment efficiency at the low end of the 

concentration scale is limited, while effluent concentrations may 

vary widely at the high end of the scale, reflecting various degrees 

of treatment system performance and loadings. These factors 

combine to produce the characteristically positively skewed appearance of the lognormal curve when data are plotted in a 

frequency histogram. Appendix E discusses the basis for concluding that effluent data are typically lognormally distributed, as well 

as recommendations for handling data sets from treatment plants 

that follow some other type of distribution. 

Effluent data from any treatment system may be described using 

standard descriptive statistics, such as the mean concentration of 

the pollutant or pollutant parameter (i.e., the long-term average 

[LTA] and the coefficient of variation [CVj). The CV is a standard 

statistical measure of the relative variations of a distribution or set 

of data, defined as the ratio of the standard deviation to the 

mean. Using a statistical model, such as the lognormal, an entire 

distribution of values can be projected from limited data, and 

limits can be set at a specified probability of occurrence. Figure 5- 

1 shows the frequency-concentration curve and the relative positions of the concentrations corresponding to the mean for the 

data. 

:x----- 

--___---__----__----_ 

, Wasteloed Akxat~on cv=o9 

s 12 

= 10 

!z 0 

Es 

s 4 

2 

n 

Days 

Figure S-2a. Relationship Between a Single Wasteload 

Allocation and Two Long-Term Averages 

for Different Coefficients of Variation 

All permit limits, whether technology-based or water qualitybased, are set at the upper bounds of acceptable performance. 

The purpose of a permit limit is to specify an upper bound of 

acceptable effluent quality. For technology-based requirements, 

the limits are based on proper operation of a treatment system. 

For water quality-based requirements, the limits are based on 

maintaining the effluent quality at a level that will comply with 

water quality standards, even during critical conditions in the 

receiving water. These requirements are determined by the WM. 

The WIA dictates the required effluent quality which defines the 

desired level of treatment plant performance or target LTA. 

In the development of technology-based effluent limits guidelines, the operating records of various wastewater treatment facilities for a particular category of discharger are examined. Based 

on the effluent data for the treatment facilities, a composite mean 

or LTA value for the parameter is determined. This LTA value, 

with relevant estimates of variability, is then used to derive effluent limit guidelines, which lead directly to permit limits. 

In contrast, the process operates in reverse for water quality-based 

permit limits. The WV\, determined from water quality standards, defines the appropriate discharge level, which in turn 

determines the requisite target LTA for the treatment facility in 

order to meet that WIA. Permit limits may then be derived from 

this targeted LTA and CV. Figure 5-2 illustrates the relationship 

among the various statistical parameters. As these figures show, 

highly variable effluents require a much lower targeted LTA to 

meet the WLA and account for the variability that occurs in 

effluent concentration above the LTA. 

It is extremely important to recognize that the various statistical 

principles and relationships discussed above operate in any dis05 10 1.5 20 

Coeffmeni of Vanatlon 

Figure S-2b. Long-Term Average Per Unit Wasteload 

Allocation as a Function of the Coefficient of Variation 

95 
charge situation-whether or not they are specifically recognized 

or accounted for. Where a permit limit derivation procedure does 

not address these principles specifically, the permit writer will be 

implicitly assuming that there are enough conservative assumptions built into other steps in the process (e.g., water quality 

models, “buffer” between permit limits and actual operating 

conditions) to ensure that there will be no reasonable potential for 

excursions above water quality standards. 

5.23 &m?ssho of Pemit lhits 

The NPDES regulations at 40 CFR 122.45(d) require that all permit 

limits be expressed, unless impracticable, as both average monthly 

and maximum daily values for all discharges other than POlWs 

and as average weekly and average monthly limits for POTWs. 

The maximum daily permit limit (MDL) is the highest allowable 

discharge measured during a calendar day or 24-hour period 

representing a calendar day. The average monthly permit limit 

(AML) is the highest allowable value for the average of daily 

discharges obtained over a calendar month. The average weekly 

permit limit (AWL) is the highest allowable value for the average 

of daily discharges obtained over a calendar week. 

EPA believes that a maximum daily permit limit can be directly 

used to express an effluent limit for all toxic pollutants or pollutant 

parameters except chronic whole effluent toxicity. The typical 

toxicity test used to measure chronic toxicity consists of samples 

collected from at least 3 different days over a 7-day period. 

Therefore, the test does not measure toxicity in any given 24-hour 

period or calendar day, but rather measures toxicity over a 7-day 

period. The toxicity could be caused by any one sample or a 

combination of samples. To address this situation, EPA recommends that the permit contain a notation indicating that 

when chronic toxicity tests are required in a permit, the MD1 

should be interpreted as signifying the maximum test result 

for the month. 

Additionally, in lieu of an AWL for POTWs, EPA recommends 

establishing an MDL (or a maximum test result for chronic toxicity) for toxic pollutants and pollutant parameters in water quality 

permitting. This is appropriate for at least two reasons. First, the 

basis for the 7-day average for POTWs derives from the secondary 

treatment requirements. This basis is not related to the need for 

assuring achievement of water quality standards. Second, a 7-day 

average, which could comprise up to seven or more daily samples, 

could average out peak toxic concentrations and therefore the 

discharge’s potential for causing acute toxic effects would be 

missed. A MDL, which is measured by a grab sample, would be 

toxicologically protective of potential acute toxicity impacts. 

5.3 EHSURIWG CONSISTENCY WITH THE 

waslEl.oao au.ocaTloH 

The WLA provides a definition of effluent quality that is necessary 

to meet the water quality standards of the receiving water. The 

WIA is based on ambient criteria and the exposure of the resident 

aquatic community or humans to toxic conditions. Once a WIA 

has been developed, accounting for all appropriate considerations, a water quality-based permit limit may be derived to 

enforce the WLA. The method used to derive the permit limits 

must be consistent with the nature of the WLA. 

The WIA addresses variability in effluent quality. For example, a 

WLA for human health pollutants is typically expressed as a single 

level of receiving water quality necessary to provide protection 

against long-term or chronic effects. On the other hand, a WLA 

for toxic pollutants affecting aquatic life (with corresponding 

duration and frequency requirements) should describe levels necessary to provide protection against both short-term and longterm effects. 

5.3. I 3tat&tlcol &onsi~~os of bus 

Direct use of a WLA as a permit limit creates a significant risk that 

the WLA will be enforced incorrectly, since effluent variability and 

the probability basis for the limit are not considered specifically. 

For example, the use of a steady state WIA typically establishes a 

level of effluent quality with the assumption that it is a value never 

to be exceeded. The same value used directly as a permit limit 

could allow the WLA to be exceeded without observing permit 

violations if compliance monitoring was infrequent, Confusion 

can also result in translating a longer duration WlA requirement 

(e.g., for chronic protection) into maximum daily and average 

monthly permit limits. The permit writer must ensure that permit 

limits are derived to implement a WIA requirement correctly. 

Potential problem areas are as follows: 

The WL4 must be enforced in a regulatory context by 

translating it into MDLs and AMLs; then and only then, will 

compliance monitoring associated with permit limits allow 

the regulatory authority to determine whether or not such 

permit limits are violated. 

The WLA that assumes that the discharge is steady state 

(i.e., not changing over time) requires a limit derivation 

assumption regarding how the effluent may vary. 

MDLs and AMLs average monthly limits must be developed 

so that they are consistent with each other and mandate 

the required level of wastewater treatment facility performance. 

If the acute WLA is used alone directly as the MDL, the limit 

will not necessarily be protective against chronic effects. If 

the acute WLA is used alone directly as the AML, the limit 

can allow excursions above the WIA within each month. 

If the chronic WLA is used alone as an MDL, the limit will be 

protective against acute and chronic effects but at the 

expense of being overly stringent. If the chronic WLA is 

used alone as the AML, the limit may be protective against 

acute and chronic effects depending upon effluent variability. 

The objective is to establish permit limits that result in the effluent 

meeting the WL4 under normal operating conditions virtually all 

the time. It is not possible to guarantee, through permit limits, 

that a WLA will never be exceeded. It is possible, however, using 

the recommended permit limit derivation procedures, to account 

for extreme values and to establish low probabilities of exceedence 

96 
of the WIA in conformance with the duration and frequency 

requirements of the water quality standards. This is not to suggest that permit writers should assume a probability of exceeclence 

of the WIA, but rather, that they should develop limits that will 

make an exceedance a very small likelihood. 

Since effluents are variable and permit limits are developed based 

on a low probability of exceedence, the permit limits should 

consider effluent variability and ensure that the requisite loading 

from the WIA is not exceeded under normal conditions. In effect 

then, the limits must “force” treatment plant performance, which, 

after considering acceptable effluent variability, will only have a 

low statistical probability of exceeding the WIA and will achieve 

the desired loadings. 

figure 5-3 shows a number of important aspects of the relationships among the various statistical parameters. In this illustration, 

the most limiting LTA (after comparing the LTAS derived from 

both acute and chronic WLAs) has been chosen for the chronic 

limiting condition. The more restrictive LTA will automatically 

meet both WIA requirements. If the effluent “fingerprint” for this 

LTA (and associated CV) is projected, it can be seen that the 

distribution of daily effluent values will not exceed the acute or 

chronic wasteload allocations for unacceptable periods of time. 

The duration and frequency requirements of the acute and chronic 

criteria for the pollutant or pollutant parameter will not be exceeded. This figure also illustrates permit limits derived from the 

more limiting LTA. (Note that for the scenario depicted in Figure 

5-3, the MDL is lower than the acute WIA and the average 

monthly limit is lower than the chronic WIA. This scenario will 

occur when a 99-percent probability basis is used to calculate the 

LTA and a 95percent probability basis is used to calculate the 

permit limits from the lower of the acute and chronic LTA. For 

other probability assumptions, these relationships will differ.) 

5.3.2 Tvpes of Water QuaMy Models and ModsI lli@wts 

Each of the two major types of water quality models, steady-state 

and dynamic, and their WlA outputs ha\‘? specific implications 

I I I I I I 

0 10 20 30 40 50 

Days 

Figure 5-3. Relationship Between Daily Concnetrations, 

Long-Term Average, Wasteload Allocations, 

and Permit Limits 

for the subsequent permit limit development process. These 

implications are discussed in detail below. EPA recommends 

that steady-state WLA analyses generally be used by permitting authorities in most cases and especially where few or no 

whole effluent toxicity or specific chemical measurements are 

available, or where daily receiving water flow records are not 

available. Two-value, steady-state models, although potentially 

more protective than necessary, can provide toxicologically protective results and are relatively simple to use. If adequate 

receiving water flow and effluent concentration data are available to estimate frequency distributions, EPA recommends 

that one of the dynamic WLA modeling techniques be used to 

derive WiAs that will more exactly maintain water quality 

standards. 

Steady-State Modeling 

Traditional single-value or two-value steady-state WIA models 

calculate WlAs at critical conditions, which are usually combinations of worst-case assumptions of flow, effluent, and environmental effects. For example, a steady-state model for ammonia 

considers the maximum effluent discharge to occur on the day of 

lowest river flow, highest upstream concentration, highest pH, 

and highest temperature. Each condition by itself has a low 

probability of occurrence; the combination of conditions may 

rarely or never occur. Permit limits derived from a steady-state 

WL4 model will be protective of water quality standards at the 

critical conditions and for all environmental conditions less than 

critical. However, such permit limits may be more stringent than 

necessary to meet the return frequency requirements of the water 

quality criterion for the pollutant of concern. 

On the other hand, a steady-state model approach may involve 

simplifying assumptions for other factors, such as ambient background concentrations of a toxicant, multiple source discharges 

of a toxicant, number of pollutants causing toxicity, incorrect 

effluent variability assumptions, and infrequent compliance monitoring. The effect of these types of factors, especially if unaccounted 

for in the WLA determination, can reduce the level of protectiveness provided by the critical condition assumptions of the steadystate model approach. Therefore, when using a steady-state WIA 

model, the permitting authority should be aware of the different 

assumptions and factors involved and should consider these assumptions and factors adequately consideration when developing permit limits. 

In general, steady-state analyses tend to be more conservative 

than dynamic models because they rely on worst case assumptions. Thus, permit limits derived from these outputs will generally be lower than limits derived from dynamic models. 

o) Sing/e Value From a Steady-Stote Analysis 

Some single-value, steady-state modeling has been used to calculate only chronic WlAs. These models produce a single effluent 

loading value and no information about effluent variability. Single 

value WLAs are typically based upon older State water quality 

standards that do not specify levels for both acute and chronic 

protection but only include one level of protection. Such outputs 

also would be found where a model is based upon protection of 

human health, since only a single long-term ambient value is of 

concern. 

97 
b) Two Values from Steady-State Analysis 

Steady-state modeling for protection of aquatic life can specify 

two sets of calculations-one for protection against acute effects 

and one for protection against chronic effects. These models 

must use water quality criteria specifying two levels of protection. 

In addition, these models include considerations of mixing zones 

when developing WlAs to afford two levels of protection. Like 

the single-value, steady-state models, these models do not produce any information about acceptable effluent variability and 

may require additional calculations to be translated into permit 

limits. 

For complex discharge situations (i.e., multiple dischargers or 

complex environmental factors needing consideration), water quality models and associated WLAS are typically developed by specialized water quality analysts in the regulatory authority. However, the permit writer is often required to develop a water quality 

model and WLA prior to permit limit derivation. In the latter 

situation, water quality modeling usually consists of simple steadystate dilution models using worst-case assumptions. 

Dynamic Modeling 

Dynamic models use estimates of effluent variability and the 

variability of receiving water assimilation factors to develop effluent requirements in terms of concentration and variability. The 

outputs from dynamic models can be used to base permit limits 

on probability estimates of receiving water concentrations rather 

than worst-case conditions. The advantages and disadvantages 

of various types of dynamic models are provided in Chapter 4. 

In general, dynamic models account for the daily variations of and 

relationships between flow, effluent, and environmentat conditions and therefore directly determine theactual probability that a 

water quality standards exceedence will occur. Because of this, 

dynamic models can be used to develop WLAs that maintain the 

water quality standards exactly at the return frequency requirements of the standards. Since this return frequency is usually one 

event in 3 years, Wk developed by dynamic models are typically 

higher than those developed by steady-state models. 

A targeted long-term average performance level and coefficient of 

variation can be derived from each type of dynamic model output, but some of the outputs require some additional manipulation of the data to develop the LTA and the CV. These parameters 

are also the starting point for the statistical permit limit derivation 

procedures discussed in the next section. Continuous Simulation models offer an array of effluent data that require further 

manipulation to develop an LTA and a CV. Both Monte Carlo 

and Lognormal Probabilistic models produce an LTA and CV, 

which can be used directly in developing permit limits. Chapter 4 

details the different dynamic models. Specific instructions for the 

use of dynamic models are available in the references listed at the 

end of Chapter 4. 

5.4 PERMlTlJMlTDERlVATlOW 

There are a number of different approaches currently being used 

by permitting authorities to develop water quality-based limits for 

toxic pollutants and toxicity. Differences in approaches are often 

attributable to the need for consistency between permit limit 

derivation procedures and the assumptions inherent in various 

types of water quality models and WlA outputs. In addition, 

permitting authorities also are constrained by legal requirements 

and policy decisions that may apply to a given permitting situation. In some instances, however, permitting procedures have 

been adopted without careful consideration of the toxicological 

principles involved or the advantages and disadvantages of the 

procedure. 

To avoid this problem, EPA recommends that the statistical 

permit limit derivation procedure described in this chapter be 

used for the derivation of both chemical-specific and whole 

effluent toxicity limits for NPDES permits. The type of WLA 

chosen from which to derive the limits is a matter of case-by-case 

application, as determined by the permitting authority. Although 

there are advantages and disadvantages associated with each of 

the procedures, EPA believes that the statistical derivation procedures will result in the most defensible and protective water 

quality-based permit limits for both specific chemicals and whole 

effluent toxicity. 

The following section explains EPA’s recommended permitting 

procedures and highlights advantages and disadvantages of various other approaches. With this information, permitting authorities will be better informed when deciding on the most appropriate permit limit derivation approach. For example, permitting 

authorities may decide to derive water quality-based permit limits 

for all dischargers using a steady-state WIA model as a baseline 

limit determination. If time and resources are available or if the 

discharger itself takes the initiative (after approval by the regulatory authority), dynamic modeling could be conducted to further 

refine the WV\ from which final permit limits would be derived. 

Box 5-l presents example permit limit calculations for each of the 

principal types of WL4 outputs dlscussed in Section 5.4.1. Permit 

limits derived from dynamic modeling are usually higher than 

those based upon steady-state modeling. The difference is reflected in Box 5-l and has been observed in actual applications [l, 

2, 31. In addition, the case studies in Chapter 7 illustrate how 

water quality-based permit limits are derived and compare the 

results of limits derived from steady state and dynamic wasteload 

allocations. 

54.1 EPA Rec~n~tituts for Pmritthrg for Aquatic 

Uf&? mtRctiun 

Permit Limit Derivation from Two-Value, Steady-State Outputs for Acute and Chronic Protection 

A number of WL4s have two results: acute and chronic requirements. These types of allocations will be developed more often as 

States begin to adopt water quality standards that provide both 

acute and chronic protection for aquatic life. These WLA outputs 

need to be translated into MDLs and AMLs. The following 

methodology is designed to derive permit limits for specific chemicals as well as whole effluent toxicity to achieve these WLAs. 

l A treatment performance level (LTA and CV) that will allow 

the effluent to meet the WLA requirement is calculated. 

98 
Box B-1. Sample Calculations of Permit Limits for Whole Effluent Toxicity 

from Different Wasteload Allocation Data 

r---- ----- .- - - --- 

L------------ Available Data 

1 

Two Value wasteload Dynamic model Single warteload j ------ --- - ---~-~_ 

allocation ~-----------.- output allocation - ---.--.-.- --A 

, Wastefoad Allocation (WLA) ____ _--_ 14.3 

Acute Wasteload Allocation (WlAa) 2.60 ---_ -__- 

Chronic Wasteload Allocation (WlAc) 14.3 __-_ __-_ 

I Acute-Chronic Ratio 4.62 ____ I-.- 

1 Coefficient of Variation (CV) 0.8 0.8 0.8 

) Number of Samples per Month (n) 4 4 4 

I Long Term Average (LTA) 

I 

---- 9.44 -_.- 

----- --i 

I- -.- - 

From two-value steady state wasteload allocatlon 

---7 I- - - - - - 

From dynamk model output I 

i I 

WLA,,,= WlA,*ACR = 2.6094.62 = 12.0 

LTA, = WfApe (0.504~-2.326041 = 14.300.440 (from Table 5-1) = 6.29 MDL = LTAc*e [2.326o-O.So*]= 9.4404.01 (from Table 5-2)= 37.9 

LTAa,c = WIAa,c*e [O.S$-2.3260]= 12.0*0.249 

II 

I 

(from Table S-l) = 2.99 

MDL = LTA,,c*e [2.326o-O.Sa*] = 2.99*4.01 (from Table 5-2) = 12.0 II AML = LTApe [2.3260n-O.Son*]= 9.44.2.27 (from Table S-2)= 21.4 

AML = LTAa,,*e [2.326o,-O.So,*~= 2.9902.27 (from Table 5-2) = 6.79 

I I 

! ! i__ -~---.- - - - - - - -I 

r---- -- ---- -.--------~ 

From slngk warteload allocation 

------- ------------J 

lOntion 1 I 

I . 

LTA = WLA*e [0.502-2.32601 = 14.3*0.440 (from Table 5-l) = 6.29 ’ 

MDL = LTA-e [2.3260-O.Sa*] = 6.29*4.01 (from Table 5-2) = 25.2 ’ 

Note: All calculations use the 99th 

AML = LTA*e [2.3260,-OSo,*) = 6.29-2.27 (from Table 5-2) = 14.3 

percentile z statistic for calculation 

I of long-term averages and permit 

: Option 2 ’ limitsMDL =wlA = 14.3 I 

1 AML = MDLI2 = 7.15 

I - .- - .- - -J 

Where two requirements are specified based on different 

duration periods, two performance levels are calculated 

(Box 5-2, Step 2). 

l For whole effluent toxicity only, the acute WV\ is converted 

into an equivalent chronic WL4 by multiplying the acute 

WLA by an acute-to-chronic ratio (ACR). This ratio should 

optimally be based on effluent data, but also can be estimated as 10, based on the information presented in Chap 

ter 1 and Appendix A. 

l Permit limits are then derived directly from whichever performance level is more protective (Box 5-2, Steps 3 and 4). 

Figure 54 presents a flow chart summarizing the various steps in 

this procedure. In addition, the equations used in Box 5-2 are 

based on the lognormal distribution, which is explained in more 

detail in Appendix E. The principal advantages of this procedure 

are described below. 

l This procedure provides a mechanism for setting permit 

limits that will be toxicologically protective. A steady-state 

WL4 uses a single value to reflect the effluent loading and 

thus is an inherent assumption that the actual effluent will 

not exceed the calculated loading value. If the WIA is 

simply adopted as the permit limit, the possibility exists for 

exceedance of the WLA due to effluent variability. Clearly, 

however, effluents are variable. Therefore, permit limits are 

established using a value corresponding to a percentile of 

the selected probability distribution of the effiuent (e.g., 

95th or 99th percentile). 

. It allows comparison of two independent WLAs (acute and 

chronic) to determine which is more limiting for a discharge. The WLA output provides two numbers for protection against two types of toxic effects, each based upon 

different mixing conditions for different durations. Acute 

effects are limited based upon 1 -hour exposures at critical 

conditions, close to the point of discharge, or where necessary, at the end of the pipe. Chronic effects are limited 

based on 4-day exposures after mixing at critical conditions. These requirements yield different effluent treatment 

requirements that cannot be compared to each other without calculating the LTA performance level the plant would 

need to maintain in order to meet each requirement. Without this comparison (or in the absence of procedures that 

address this comparison), the WLA representing the more 

critical condition cannot be determined. A treatment system will only need to be designed to meet one level of 

99 
Box 5-2. Calculating Permit Limits Based on Two-Value Wasteload Allocation 

To set maximum daily and average 

monthly permit limits based on 

acute and chronic wasteload 

Step 1 (for whole effluent toxicity only) 

allocations, use the following four 

WLA,, (in TU,) = WLA, (in TU,) l ACR 

steps: 

Convert the acute wasteload 

1 allocation to chronic toxic Step 2 (start here for chemical specific limits) 

units. Skip to Step 2 for 

chemical-specific limits. LTA, c = Wu, c l e [o.52 - “1 

Calculate the long-term where o’ = fn(CV’+l) 

2 average wasteload that will z = 1.645 for 95th percentile probability basis, and 

satisfy the acute and chronic z = 2.326 for 99th percentile probability basis 

wasteload allocations. 

Determine the lower (more 

LTA, = W& l e [“.50z - z”4] 

3 limiting) of the two long-term where 04* = ln(CV2/4 +l) 

averages z = 1.645 for 95th percentile probability basis, and 

z = 2.326 for 99th percentile probability basis 

Calculate the maximum daily 

Step 3 

Term Meaning 

LTA = min (LTA,, LTA,,) 

cv 

0 

w%,c 

w% 

W”c 

LTAsc 

LTA, 

T”a 

T”c 

ACR 

MDL 

AML 

Z 

Coefficient of variation 

Standard deviation 

Acute wasteload allocation 

in chronic toxic units 

Acute wastelaad allocation 

in acute toxic units 

Chronic wasteload 

allocation in chronic toxic 

units 

Acute tong-term average 

wasteload in chrunic units 

Chronic long-term average 

wasteload 

Acute toxic units 

Chronic toxic units 

Acute-to-chronic ratio 

Maximum daily limit 

Average monthly limit 

2 statistic 

Step 4 

MDL = LTA l e iza -o.52l 

where d = In(CV2+1) 

z = 1.645 for 95th percentile probability basis, and 

z = 2.326 for 99th percentile probability basis 

AML = LTA l e Izon - o.50,21 

where on2 = ln(CV*/n +l ) 

z = 1.645 for 95th percentile probability basis, and 

z = 2.326 for 99th percentile probability basis 

‘Full &tails of this procedure are found in Appendix E. 

100 
I 

1 CakulateChmnic 1 

Calculate Maximum 

Dally Umlt 

/ 

I Calculate Avaraga 

Monthly Llmh I 

Figure S-4. Flowchart for Calculating Permit Llmits From 

Two-Value, Steady-State Wasteload Allocatlon 

for Aquatic Life Protection 

treatment for effluent toxicity-treatment needed to control 

the most limiting toxic effect. 

l The actual number of samples can be factored into permit limit 

derivation procedures. The procedure provides the means to 

accurately determine the AML based on the number of observations that will be taken. 

The principal disadvantages of this approach are: 

l Some permit writers have indicated that additional mathematical calculations associated with these procedures increase 

the burden for the permit writer and add what is perceived to 

be an unnecessary step. 

l The use of a steady-state WLA may result in permit limits that 

are more conservative due to the assumption of critical conditions. However, these limits are still protective of water quality 

criteria. The level of conservatism may be necessary in those 

instances where limited data prevent a more precise evaluation 

of a WLA. 

This procedure provides a toxicologically sound approach. To 

help the permit writer, EPA has developed tables (see Tables 5-l 

and 5-2) to be used to quickly determine the necessary values. In 

addition, some permit authorities have developed their own computer programs to readily compute the necessary information 

from the appropriate inputs. 

Permit Limit Derivation From Dynamic Model Outputs 

The least ambiguous and most exact way that a WIA for specific 

chemicals or for whole effluent toxicity can be specified by using 

dynamic modeling from which the WIA is expressed as a required 

effluent performance in terms of the LTA and CV of the daily 

values. When a WLA is expressed as such, there is no confusion 

about assumptions used and the translation to permit limits. A 

permit writer can readily design permit limits to achieve the WLA 

objectives. The types of dynamic exposure analyses that yield a 

WLA in terms of required performance are the continuous simulation, Monte Carlo, and lognormal probabilities analyses. Chapter 

4 provides a general discussion of these models. Guidance manuals for developing WLAS are listed in the references at the end of 

Chapter 4. Once the WL4 is determined, the permit limit derivation procedure which can be used for both whole effluent toxicity 

and specific chemicals, is as follows: 

The WL4 is first developed by iteratively running the dynamic model with successively lower LTAs until the model 

shows compliance with the water quality standards. 

The effluent LTA and CV must then be calculated from the 

model effluent inputs used to show compliance with the 

water quality standards. This step is only necessary for the 

Monte Carlo and continuous simulation methods. 

The permit limit derivation procedures described in Box 5 

2, Step 4 are used to derive MDLs and AMLs from the 

required effluent LTA and CV. Unlike these procedures for 

steady-state WIAS, there is only a single LTA that provides 

both acute and chronic protection, and, therefore, the 

comparison step indicated in Figure 5-4 and Box 5-2 is 

unnecessary. 

The principal advantages of this procedure are: 

l It provides a mechanism for computing permit limits that 

are toxicologically protective. As with the procedure summarized below for two-value, steady-state WLA outputs, 

the permit limit derivation procedures used with this type 

of output consider effluent variability and derive permit 

limits from a single limiting LTA and CV. 

l Actual number of samples is factored into permit limit 

derivation procedures. This procedure has the same elements as discussed for the statistical procedures in Option 2 

below. 

l Dynamic modeling determines an LTA that will be adequately protective of the WLA, which relies on actual flow 

data thereby reducing the need to rely on worst case critical 

flow condition assumptions. 

101 
Table S-l. Back Calculations of Long-Term Average 

---_----_---_ 

WIA Multipliers 

.-- - --- 

(-J/ I e[o.5 a2-201 I 

i--~-------- 

__ ---I 

95th 

Percentile 

____t----L------! 

0.1 

0.2 

0.3 

0.4 

0.5 

0.6 

0.7 

0.0 

0.9 

1.0 

1.1 

1.2 

1.3 

1.4 

1.5 

1.6 

1.7 

1.8 

1.9 

2.0 

0.853 

0.738 

0.644 

0.571 

0.514 

0.468 

0.432 

0.403 

0.379 

0.360 

0.344 

0.330 

0.319 

0.310 

0.302 

0.290 

0.290 

0.285 

0.2m 

0.277 

0.797 ; 

0.643 ( 

0.527 

0.440 

0.373 

0.321 

0.281 

0.249 

0.224 

0.204 1 

0.187 

0.174 

0.162 

0.153 

0.144 

0.137 

0.131 

0.126 1 

0.121 

0.117 I 

I_ - -.---~ -----------____-- - - -- -- ------ _ - - --- ---- ---, 

AC&? 

LTA,,c = WLA,, l e[“” 02- ’ = ’ 

here u*= /n[CV* + 11. 

2=1.645for95thpercentileoaxrnen~probebili(y,and I 

z= 2.326 for9Qth percentileoccurrerxe probability I 

--- ------------ 

I 

------- T--- - 

---_-------, 

WIA Multipliers I 

Chronic 

(4-day average) 

where u,2= ffl[CV2/4+ 11. 

z = 1.645foc95thpercentileoccunenceprobability,and 

z =2,326for991trpercentile~~eprohabiliOl 

cv 

0.; 

0.2 

0.3 

0.4 

0.5 

0.6 

0.7 

0.8 

0.9 

1.0 

1.1 

1.2 

1.3 

1.4 

1.5 

1.6 

1.7 

1.8 

1.9 

2.0 

j e[o.s u4’- L u4 ] I 

7---- 

_-- ----- 

0.922 

0.853 

0.791 

0.736 

0.687 

0.644 

0.606 

0.571 

0.541 

0.514 

0.490 

0.460 

0.449 

0.432 

0.417 

0.403 

0.390 

0.379 

0.369 

0.380 

I 0.891 

0.797 

’ 0.715 

’ 0.643 

, 0.581 

/ 0.527 

0.481 

I 0.440 

1 0.404 

’ 

0.373 

’ 

0.345 

0.321 

, 0.300 

0.261 

0.264 

; 0.249 

0.236 

, 0.224 

i 0.214 

0.204 _L_ _--- - 

The principal disadvantages of this procedure are: Permit Limit Derivation From Single, Steady-State Model 

Output 

l Necessary data for effluent variability and receiving water 

flows may be unavailable, which prevents the use of this 

approach. 

Some State water quality criteria and the corresponding WlAs are 

reported as a single value from which to define an acceptable 

l The amount of staff resources needed to explain how the level of effluent quality. For example, “copper concentration 

limits were developed and to conduct the WLA also is a 

must not exceed 0.75 milligrams per liter (mg/l) instream.” Steadyconcern. The permit documentation (i.e., fact sheet) will 

state analyses assume that the effluent is constant and, therefore, 

need to clearly explain the basis for the LTA and CV and this the WL4 value will never be exceeded. This presents a problem in 

can be resource intensive. 

deriving permit limits because permit limits need to consider 

effluent variability. 

102 
Table 5-2. Calculation of Permit Limits 

LTA multipliers 

cv eIzo-o.502] 

95th 99th 

Percentile Percentile 

0.1 1.17 1.25 

0.2 1.36 1.55 

0.3 1.55 1.90 

0.4 1.75 2.27 

0.5 1.95 2.66 0.8 2.13 3.11 

0.7 2.31 3.56 

0.8 2.40 4.01 

0.9 2.64 4.46 

1.0 2.78 4.90 

1.1 2.91 5.34 

1.2 3.03 5.76 

1.3 3.13 6.17 

1.4 3.23 6.56 

1.5 3.31 6.93 

1.6 3.38 7.29 

1.7 3.45 7.63 

1.8 3.51 795 

1.9 3.56 6.26 

2.0 360 8.55 

Maximum Daily Limit 

MDL=LTAme[Zu-o.5621 

where u2 = In [ CV2 + 1 1. 

z = 1.645 for 95th percentile occurrence probability, and 

z = 2.326 for 99th percentile occurrence probability 

LTA Multipliers 

e’ 2 a, -0.5 CT"21 

cv . 95th 99th 

Average Monthly Limit Percentile Percentile 

rkl n=2 n=4 n=lO n=30 n=l n=2 n=4 n=iO n=30 

0.1 1.17 1.12 1.06 106 1.03 1.25 1.18 1.12 106 1.04 

0.2 1.36 1.25 1 17 1.12 1.06 1.55 1.37 1.25 116 109 

0.3 1.55 1.38 1.26 1.18 1.09 1.90 159 1.40 124 113 

0.4 1.75 t.52 1.36 1.25 1 12 2.27 1.83 1.55 1.33 1 16 

0.5 

JML=LTAee [zo,-0’5n”21 0.6 

1.95 1.86 1.45 1.31 1.16 2.68 2.09 1.72 1.42 1.23 

2.13 1.60 1.55 1.38 1.19 3.11 2.37 190 152 1.26 

0.7 2.31 1.94 1.65 1.45 1.22 3.56 2.66 2.06 162 1.33 

wherean2=In[CV2/n+1], 0.8 2.48 2.07 1.75 1.52 1.26 4.01 2.96 2.27 173 1.39 

09 2.64 2.20 1.65 1.59 1.29 4.46 3.28 2.48 184 1.44 

z = 1.645 for 95th percentile, 1.0 2.78 2.33 1 95 1.66 1.33 4.90 3.59 2.68 1 96 150 

z = 2.326 for 99th percentile, and t t 2.91 2.45 2.04 '.73 1.36 5.34 3.91 2.90 2.07 156 

n = number of sampletimonth 12 3.03 2.56 2.13 1.80 1.39 5.76 4.23 3.11 2.19 1.62 

1.3 3.13 2.67 2.23 1.87 1.43 6.17 4.55 3.34 2.32 166 

14 3.23 2.77 2.31 1.94 1.47 6 56 4.86 3.56 2.45 1 74 

1.5 3.31 2.66 2.40 2.00 1.50 6.93 5.17 3.78 2.58 160 

1.6 3.38 2.95 2.48 2 07 1.54 7.29 5.47 4.01 2.71 1.67 

17 3.45 3.03 2.56 2.14 1.57 7.63 5.77 4.23 2.84 193 

1.8 3.51 3.10 2.64 2.20 1.61 7.95 6.06 4.46 2.98 2.00 

1.9 3.56 3.17 2.71 2.27 1.64 8.26 6.34 4.68 3.12 2.07 

2.0 3.60 3.23 2.78 2.33 1.66 8.55 6.61 4.90 326 2 14 

The proper enforcement of this type of WIA depends on the 

parameter limited. For nutrients and biochemical oxygen demand (BOD), the WlA value generally has been used as the 

average daily permit limit. However, the impact associated with 

toxic pollutants is more time dependent, as reflected in the 4-day 

average duration for the criteria continuous concentration (CCC) 

(see Chapter 2). Where there is only one water quality criterion 

and therefore only one WL4, permit limits can be developed 

using the following procedure: 

The principal advantages and disadvantages of this procedure are 

similar to those for the two-value permit limit derivation method 

discussed previously except that it does not examine two WL4s. 

5.4.2 Uthor A#mdo.s to Pamittin~ for Aquatfc life 

Other approaches for translating WLA outputs into permit limits 

have been used by some permitting authorities. These methods 

may combine elements of the statistical procedures discussed 

earlier with specific technical and policy requirements of the 

permitting authority to derive limits that may be protective of 

water quality and consistent with the requirements of the WLA. 

Such approaches may use simplified statistical procedures. 

l Consider the single WLA to be the chronic WLA and derive 

an chronic LTA for this WLA using the procedures in Box S2 (Step 2, Part 2). 

l Derive MDLs and AMLs using the procedures ‘in Box 5-2 

(Step 4). 

103 
For example, some permitting authorities assume a value for the 

CV and an acute to chronic ratio above which the chronic WIA 

will always be more limiting. Where such simplifying assumptions 

are used, the need to compare LTAs derived from acute and 

chronic steady-state models is unnecessary. Similarly, for assumed values for n, CV, and exceedence probability, the various 

equations shown in Box S-2 can be simplified further, such that 

the AML will always be a constant fraction of the MDL. 

These approaches allow the permit writer to rapidly and easily 

translate the results of WL4.s into permit limits. However, the 

permit writer clearly should understand the underlying procedures and carefully explain the basis for the chosen assumption. 

Appropriate State or regional guidance documents also should be 

referenced. 

Another approach used by some permit authorities involves the 

direct use of the WIA as a permit limit. This approach sometimes 

involves the following steps: 

l The WlA value for toxic pollutants is used as the MDL. 

. In the absence of other information, permit writers typically 

divide the MDL by 1.5 or 2.0 to derive an AML (depending 

on the expected range of variability). 

The principal advantage of this approach is that it is very straightforward to implement and requires minimal resources. The disadvantage of this option is that the average monthly limits must be 

derived without any information about the variability of the effluent parameter; therefore, the permit writer cannot be sure that 

these procedures are protective of water quality criteria. Conversely, limits derived from this approach may be overly stringent 

and subject to challenge. 

The direct application of both the acute and chronic WLAS as 

permit limits is another approach that has been used. The WLA 

developed for protection against chronic effects becomes the 

average monthly limit and the acute WIA becomes the MDL. 

EPA discourages the use of this approach. Since effluent variability has not been specifically addressed with this approach, 

compliance with the monthly average (30-day) effluent limit 

during critical conditions could exceed the chronic (4-day) WLA. 

Whether standards are violated with excessive frequency under 

such conditions would depend upon whether the conditions 

represented by the worst-case assumptions of the model also 

were occurring at the same time. By contrast, compliance with 

limits that were developed using statistical procedures have a low 

chance of leading to WIA excursions before effluent variability is 

accounted for in deriving the limits (see Figure S-3). 

Another permitting approach is to use a narrative “no toxicity” 

limit that is measured using a toxicity testing method that employs only a control and a single exposure at the receiving water 

concentration (RWC). This is sometimes referred to as a “pass/ 

fail” toxicity test. Although these tests can be less expensive than 

full dilution series testing, they provide no knowledge as to the 

extent of toxicity present during the test and therefore no data 

concerning the seriousness of the impact or the amount of toxicity reduction necessary. The death of a single test animal can 

occur at any concentration level beyond the lethality threshold for 

the test organism; therefore, such a test is much less powerful 

from a statistical standpoint. In addition, it is not possible to 

determine dose-response relationships for the test organisms without using multiple effluent concentrations. Dose-response curves 

are useful in determining quality assurance of the tests and in 

defining threshold dosages for regulatory purposes. Because the 

drawbacks of the approach generally outweigh the benefits, EPA 

recommends that whole effluent toxicity limits be established 

using a statistical derivation procedure that adequately accounts for effluent variability and that monitoring for compliance with whole effluent toxicity limits be conducted using a 

full dilution series. 

When setting a whole effluent toxicity limit to protect against 

acute effects, some permitting authorities use an end-of-pipe 

approach. Typically, these limits are established as an LCsdl OOpercent effluent at the end of the pipe. These limits are routinely 

set without any consideration as to the fate of the effluent and the 

concentrations of toxicant after the discharge enters the receiving water. Limits derived in this way are not water quality-based 

limits and suffer from significant deficiencies since the toxicity of a 

pollutant depends mostly upon concentration, duration of exposure, and repetitiveness of the exposure. This is especially true in 

effluent dominated waters. For example, an effluent that has an 

LCso=lOO percent contains enough toxicity to be lethal to up to 

50 percent of the test organisms. If the effluent is discharged to a 

low-flow receiving waterbody that provides no more than a threefold dilution at the critical flow, significant mortality can occur in 

the receiving water. Furthermore, such a limit could not assure 

protection against chronic effects in the receiving waterbody. 

Chronic effects could occur if the dilution in the receiving water 

multiplied by the acute to chronic ratio is greater than 100 

percent. Therefore, in effluent dominated situations, limits set 

using this approach may be severely underprotective. In contrast, 

whole effluent toxicity limits set using this approach in very high 

receiving water flow conditions may be overly restrictive. Because of these problems, EPA recommends that all whole 

effluent toxicity limits be set as water quality-based limits and 

that to do so, the statistical permit limit derivation procedures 

discussed in Section 5.4.1 be followed. 

5.4.3 special Pemlntin# lh@mmt?& 

Water quality-based permit limit development for discharges to 

marine and estuarine waters follows the same basic steps as the 

water quality-based approach for freshwater discharges. There 

are some differences in the water quality criteria used as the basis 

for protection, the designation of mixing zones, and the water 

quality models used to develop WL4s; however these differences 

are addressed in the WL4. (See discussions of these elements in 

previous chapters.) In addition, there are some special regulatory 

considerations associated with these types of dischargers, including special reviews of permits with such programs as the Coastal 

Zone Management Program. Some discharges also require an 

Ocean Discharge Criteria Evaluation under Section 403(c) of the 

Clean Water Act (CWA). 

5.4.4 EPA ltmmnmmt8tions for PemMting fbr Muman 

Healtlr Ntection 

Permit development to protect against certain routes of exposure 

is another key consideration. Ingesting contaminated fish and 

shellfish is a toxic chemical exposure route of serious potential 

104 
human health concern for which there is no intervening treatment process, unlike the drinking water route of exposure. Effluent limits designed to meet aquatic life criteria for individual 

toxicants and whole effluent toxicity are not necessarily protective 

of toxic pollutant residue formation in fish or shellfish tissue. 

Developing permit limits for pollutants affecting human health is 

somewhat different from setting limits for other pollutants because the exposure period is generally longer than 1 month, and 

can be up to 70 years, and the average exposure rather than the 

maximum exposure is usually of concern. Because compliance 

with permit limits is normally determined on a daily or monthly 

basis, it is necessary to set human health permit limits that meet a 

given WLA for every month. If the procedures described previously for aquatic life protection were used for developing permit 

limits for human health pollutants, both MDLs and AMLs would 

exceed the WIA necessary to meet criteria concentrations. Thus, 

even if a facility was discharging in compliance with permit limits 

calculated using these procedures, it would be possible to constantly exceed the WL4. This approach clearly is unacceptable. In 

addition, the statistical derivation procedure is not applicable to 

exposure periods more than 30 days. Therefore, the recommended approach for setting water quality-based limits for human health protection with statistical procedures is as follows: 

l Set the AML equal to the WLA 

l Calculate the MDL based on effluent variability and the 

number of samples per month using the multipliers provided in Table 5-3. 

This approach ensures that the instream criteria will be met over 

the long-term and provides a defensible method for calculating a 

MDL. Both an MDL (weekly average limit for POTWs) and a 

monthly average limit are required by EPA regulations, unless 

impracticable (40 CfR 122.45(d)) and are applicable for human 

health protection. The MDL sets an upper bound on effluent 

values used to determine the monthly average and provides a 

measure of effluent compliance during operational periods between monthly sampling. 

5.5 SPEClAlCONSlDERATlONSlHlJSEOF 

STATlSTlCAlPERMlTUMfTDERlVATlON 

TECHNIQUES 

The following discussion summarizes the effect of changes in the 

various statistical parameters on the permit limits that are derived. 

An understanding of these relationships is important for the permit writer. Additional considerations of each of these parameters 

with respect to the statistical methods for permit limit derivation 

also are discussed below. 

5.5. I Eflect of Chaqes of 3fotisticol Panmreters on Pmnit 

Lfmlts 

l Effect of changes in CV on derivation of LTA from WLA: 

IG the CV increases, the LTA decreases; and conversely, as 

the CV decreases, the LTA increases (see Figure 5-S). 

Reason: The LTA must be lower relative to the WLA to 

account for the extreme values observed with high CVs. An 

LTA with a zero CV equals the WLA. 

l Effect of changes in CV on derivation of permit limits for 

a fixed probability basis: As the CV increases, the permit 

limits increase (become less stringent); and conversely, as 

the CV decreases, the permit limits decrease (become more 

stringent; see Figure S-6). 

Reason: A higher value for the permit limit is produced for 

the same LTAs as the CV increases in order to allow for 

fluctuations about the mean. Following the steps in Box S2 to derive the LTA will account for such fluctuations. 

l Effect of changes in number of monthly samples on 

permit limits: As the value for “n” (number of observations) increases in the average monthly permit limit derivation equations, the average monthly permit limit decreases 

to a certain point. The effect on the average monthly limit 

is minimal for values of n greater than approximately 10. 

Conversely, as the value for “n” decreases, the AML increases until n=l, at which point the AML equals the MDL 

(see Figure S-7). 

Reason: As n increases, the probability distribution of the 

n-day average values becomes less variable (narrower) 

around the LTA. Therefore, the 95th or 99th percentile 

value for an n-day average decreases in absolute value as n 

increases. (See additional discussion in Section 5.5.3.) 

l Effect of changes in probability basis for permit limits: 

As the probability basis for the permit limits expressed in 

percentiles (e.g., 95 percent and 99 percent) increases, the 

value for the permit limits increases (becomes less stringent). The converse is true as the probability basis decreases (see Figure 5-6). 

0 I 

0.0 0.5 10 1.5 2. 

Coefficient of Vanation 

Figure 5-5. Long-Term Average as a Function of the 

Coefficient of Variation 

105 
Reason: There is a higher probability that any randomly 

chosen effluent sample will be in compliance with its permit 

limits, if those limits are statistically designed to be greater 

than a high percentage (e.g., 99 percent) of all possible 

values for a given LTA and CV. 

illustrates how the CV, number of samples and probability basis 

affect the derivation of the AML. Figure 5-9 illustrates the combined effect of the CV and the probability basis on the derivation 

of the MDL. 

The overall combination of the coefficient of variation, number of 5.52 Coemcknt af vwaml 

samples, and the assumed probability basis for calculating the LTA 

from the WLA, and the most limiting LTA, has different effects on 

the derived limits depending upon the selection made for each. 

To help illustrate the combined effect of these factors, Figure 5-8 

Use of the statistical method of permit limit derivation requires an 

estimate of the CV of the distribution of the daily measurements 

of the parameter after the plant complies with the requirements. 

Table 5-3. Multipliers for Calculating Maximum Daily Permit Limits From Average Monthly Permit Limits 

To obtain the maximum daily permit limit (MDL) for a bioconcentratable pollutant, multiply the average monthly permit limit 

(AML) (the wasteload allocation) by the appropriate value in the following table. 

Each value in the table is the ratio of the MDL to the AML as calculated by the followinq relationship derived from Step 4 of the 

statistically based permit limit calculation procedure. 

MDL = exp [z,o - 0.502] 

AML exp [zaan - 0.50n2] 

where 

un2= In (CV2/n + 1) 

o2 = In(CV2+1) 

CV = the coefficient of variation of the effluent concentration 

n = the number of samples per month 

Z m= the percentile exceedance probability for the MDL 

za = the percentile exceedance probability for the AML. 

Ratio Between Maximum Daily and Average Monthly Permit Limits 

Maximum = 99th percentile Maximum = 99th percentile 

Average = 95th percentile Average = 99th percentile 

cv n=l n=2 n=4 n=a n=30 n=l n=2 n=4 n=8 

0.1 1.07 

0.2 1.14 

0.3 1.22 

0.4 1.30 

0.5 1.38 

0.6 1.46 

0.7 1.54 

0.8 1.61 

0.9 1.69 

1.0 1.76 

1.1 1.83 

1.2 1.90 

1.3 1.97 

1.4 2.03 

1.5 2.09 

1.6 2.15 

1.7 2.21 

1.8 2.27 

1.9 2.32 

2.0 2.37 

.13 

.2s 

.37 

so 

.622 

.73 

.a4 

1.94 

2.03 

2.11 

2.18 

2.25 

2.31 

2.37 

2.42 

2.42 

2.52 

2.56 

2.60 

2.64 

.oo 

1.16 1.18 

.oo 

1.22 1 .oo 

1.33 1.39 

.oo 

1.46 1.00 

1 .so 1.60 

.oo 

1.74 / 1 .oo 

1.67 

.oo 

1.82 2.02 - 1.00 

.oo 

1.84 2.04 2.32 I 1 .OO 

2.01 2.2s 2.62 1.00 

2.16 2.45 2.91 1 .oo 

2.29 2.64 3.19 1 .oO 

2.41 2.81 3.45 , 1 .oo 

2.52 2.96 3.70 1.00 

2.62 3.09 3.93 , 1 .oo 

2.70 3.20 4.13 1.00 

2.77 3.30 4.31 1 .oo 

2.83 3.39 4.47 

2.89 3.46 4.62 

2.89 3.46 4.62 

2.98 3.57 4.85 

3.01 3.61 4.94 

3.05 3.65 5.02 

1.07 

1.3s 

1.12 

1.84 

1.16 

1.13 

2.46 

1.24 

1.34 

1.32 

1.19 

1.83 

1.36 

2.46 

1.49 

1.24 

1.33 

1.46 

1 .a2 2.46 

1.66 

I .2a 

1.32 

1.56 

1.80 

1.81 

1.31 

2.45 

1.64 

1.31 

1.9s 

1.34 

1.78 

1.71 

2.43 

2.08 

1.30 

1.3s 1.76 

1.76 

2.19 

2.41 

1.36 1.80 2.27 

1.37 1.83 2.34 

1.37 1 .a4 2.39 

1.36 1.85 2.43 

1.36 1.85 2.45 

3.07 3.67 5.09 1 .oo 1.29 1.74 2.38 

n=30 j 

\ 

1.20 1 

1.43 , 

1.67 ’ 

1.92 

2.18 

2.43 / 

2.67 1 

2.89 1 

3.09 

3.27 I 

3.43 i 

3.56 1 

3.68 ; 

3.77 I 

3.84 

3.90 

3.94 

3.97 j 

3.99 1 

4.00 ( 

106 
99% ProbabdMy Basrs , 

0.5 1.0 

Coefficient of Variation 

1.5 2.0 

Figure S-6. Maximum Daily Permit Limit as a Function of the 

1 

0.6 

!i 

2 

zi 

06 

z 

.s 2 0.4 

02 

0 

Coefficient of Variation 

10 20 

Number of Samples Per Month 

30 

Figure S-7. Relationship Between Average Monthly Permit 

Limits and Number of Samples Per Month 

If variability is mostly related to production, current data may be 

used to estimate the CV. If future variability is expected to be 

substantially different, the CV must be estimated. Discharges of 

toxic pollutants are generally more variable than discharges of 

conventional pollutants. It is important to use the best estimate of 

the CV that can be reasonably achieved. As explained in Chapter 

3, EPA’s review of the uncertainty associated with effluent variability suggests that a minimum of 10 samples is needed to reasonably quantify the CV. 

One concern with respect to using an appropriate CV in the 

statistical limit derivation procedures is that CVs of regulated 

systems may be quite different from nonregulated systems. In 

other words, after permit limits are in place and the permittee is 

operating to achieve the requisite limits, the variability associated 

with the parameter of concern may change considerably. Where 

the permit writer has reason to believe that the CV of the regulated system may behave differently from the nonregulated system (e.g., where changes in the treatment facility are planned), 

information concerning effluent concentration means and variability can be obtained from effluent guideline documents for 

individual chemical parameters. 

Variability associated with effluent levels of both individual chemicals and whole effluent toxicity is difficult to predict for any 

individual situation. However, it is important to recognize that 

failure to assign any CV to an individual toxicant or the parameter 

toxicity involves an implicit assumption that there is no effluent 

variability present. Based upon analyses of a wide variety of data 

from various types of plants, EPA recommends a value of 0.6 as 

a default CV, if the regulatory authority does not have more 

accurate information on the CV for the pollutant or pollutant 

parameter. Permit limits are usually not extremely sensitive to 

small changes in the 0’. The value of 0.6 is typical of the range of 

variability of effluents measured by EPA (see Appendix A) and 

represents a reasonable degree of relative variability. However, 

wherever possible, it is recommended that data on effluent variability for the pollutant of concern be collected to define a CV 

rather than selecting a default value. 

5.53 Nmbef of samples 

The statistically based method for permit limit derivation results in 

an MDL that does not depend on monitoring frequency. However, the AML decreases as the monitoring frequency increases, 

and a greater number for “n“ is inserted in the relevant equations. 

Some permit writers are concerned with this outcome because 

facilities with more frequent sampling requirements appear to 

receive more stringent permit limits than those with less frequent 

monthly sampling requirements. 

The AML decreases as the number of monthly samples increases 

because an average of 10 samples, for example, is closer to the 

LTA than an average based on 4 samples. This phenomenon 

makes AMLs based on 10 samples appear to be more stringent 

than the monthly limit based on 4 samples. However, the stringency of these procedures is constant across monitoring frequencies because the probability basis and the targeted LTA performance are the same regardless of the number of samples taken. 

Thus, a permittee performing according to the LTA and variability 

associated with the wasteload allocation will, in fact, meet either 

of these AMLs when taking the corresponding number of monthly 

samples. 

For water quality-based permitting, effluent quality is determined 

by the underlying distribution of daily values, which is determined 

by the LTA associated with a particular WLA and by the CV of the 

effluent concentrations. Increasing or decreasing monitoring 

frequency does not affect this underlying distribution or treatment performance, which should, at a minimum, be targeted to 

comply with the values dictated by the WLA. Therefore, it is 

recommended that the actual planned frequency of monitoring normally be used to determine the value of n for calculating the AML. However, in situations where monitoring frequency is once per month or less, a higher value for n must be 

assumed for AML derivation purposes. This is particularly 

applicable for addressing situations such as where a single criterion is applied at the end of the pipe and a single monthly sample 

107 
-.- 

3 

2.5 

2 

1.5 

1 

0.5 

a. LTA at 95% Probability Basis 

AML at 99% Probability Basis 

0.5 1 .o 1.5 2.0 

Coefficient of Variation 

3.5 - 

c. LTA at 95% Probabllity Basis 

3- 

AML at 95% Probabllity Basis 

s 

6 2.5 - 

2 

; 2- 

E 

3 1.5- 

& 

a 

J 5 l0.5 - / 

n=30 

0 “““‘I”“J1’-“’ 

0.0 0.5 1 .o 1.5 2.0 

Coefficient of Variation 

3.5 

b. LTA at 99% Probability Basis 

AML at 95% Probability Basls 

n=l 

0”““““““““” 

0.0 0.5 1.0 1.5 

Coefficient of Variation 

2.0 

-.- ^I d. LTA at 99% Probability Basis 

AML at 99% Probability Basis 

0”““““““““” 

0.0 0.5 1.0 1.5 

Coefficient of Variation 

Figure 5-8. Effect of Coefficient of Variation on Average Monthly Limits 

2.0 

108 
8. LTA at 9f 

o”“““l”““““‘j 

0.0 0.5 1 .o 1.5 2.0 

Coefficient of Variation 

c. LTA at 95% Probabiilty Basis 

3 

MDL at 95% Probablllty Basls 

2.5 

0.5 1.0 1.5 

Coefficient of Variation 

3.5 

3- 

b. LTA at 99% Probability Basis 

MDL at 95% Probability Basis 

2.5 - 

2- 

1.5 

Ol”““I”““““‘IJ 

0.0 0.5 1 .o 1.5 2.0 

Coeff bent of Variation 

3.5 

3 

2.5 

2 

t .5 

1 

0.5 

0 

d. LTA at 99% Probability Basis 

MDL at 99% Probabfllty Basis 

0.0 0.5 1.0 I.5 

Coefficient of Variation 

2.0 

Figure 5-9. Effect of Coefficient of Variation on Maximum Daily Limits 

109 
-- 

-- - 

is contemplated for compliance monitoring purposes, or where 

monitoring frequency is only quarterly. In this case, both the 

average monthly and the MDL would exceed the criterion. (For 

example, for a CCC of 1 .O chronic toxic unit [TU J applied as a 

WLA at the end of the pipe, both the MDL and AML would be 1.6 

TU,; assuming CV=O.6, n=l, and a 99percent probability basis.) 

A discharger could thus comply with the permit limit but routinely exceed the criterion. Under these circumstances, the 

statistical procedure should be employed using an assumed 

number of samples of at least fm for the AML derivatlon. 

5.5.4 ProQaQlllly nasls 

Selection of the probability basis for use in the equations in Boxes 

5-l and 5-2 is a permitting authority decision necessary for establishing statistically derived permit limits. Where a permitting 

authority does not have specific guidance for the probability 

basis, EPA recommends the following: 

For calculation of the LTAs from the WLAs (Box 5-2): 

l Both acute and chronic WLA-.Ol probability (99th percentile level). 

For calculation of permit limits from the most limiting LTA 

1): 

l MDL-.Ol probability basis (99th percentile level) 

l AML--.05 probability basis (95th percentile level). 

(Box 5- 

The probability levels for deriving permit limits have been used 

historically in connection with development of the effluent limits 

guidelines and have been upheld in legal challenges to the guidelines [4). It is important to note that these levels are statistical 

probabilities used as the basis for developing limits. The goal in 

establishing these levels is to allow the regulatory agency to 

distinguish between adequately operated wastewater treatment 

plants with normal variability from poorly operated treatment 

plants and to protect water quality criteria. 

The level for the calculation of the LTA from the WLA is based 

upon EPA’s interpretation of the steady state model used to 

develop the WI-A. EPA considers the WLA to produce an effluent 

condition that should never be exceeded whenever the critical 

design conditions occur. To characterize this effluent condition, 

EPA uses the 99th percentile concentration from the upper tail of 

the effluent probabilistic distribution curve. The selection of this 

value is one which can have a significant influence on the level of 

conservatism in the permit limits. Permit authorities should consider Figures 5-8 and 5-9 to understand the effect of this decision 

along with other decisions on the AMls and MDLs. 

5.6 PERMK GOCUMENTATION 

The fact sheet and supporting documentation accompanying the 

permit must clearly explain the basis and the rationale for the 

permit limits. When the permit is in the draft stage, the supporting documentation will serve to explain the rationale and assumptions used in deriving the limits to the permittee and the general 

public in order to allow public comment on the draft permit. 

When the permit is issued, the administrative record for the 

facility (particularly the fact sheet) will be the primary support for 

defending the permit in administrative appeals including 

evidentiary hearings. This information also will serve to alert 

compliance/enforcement personnel to any special considerations 

that were addressed at the time of permit issuance. In addition, 

the accompanying documentation will be extremely important 

during permit reissuance and will assist the permit writer in developing a revised permit. 

In 40 CFR Part 124.56, a fact sheet containing “[a]ny calculations 

or other necessary explanation of the derivation of specific effluent limitations” for many draft permits is required. Accordingly, 

the WlAs along with the required LTA and CV used and the 

calculations deriving them must be included or referenced in the 

fact sheet. The permit limit derivation method used must also be 

explained in the permit documentation. Where a permitting 

authority develops a standardized and simplified method for permit limit development as discussed in Section 5.4.2, the permitting authority may not need to document all of the underlying 

assumptions in the fact sheet, provided that the fact sheet references a written permit limit development protocol. Any other 

guidance used must also be cited. 

5.7 EXPRESSHi LIMITS AND DEVELOPING 

MONlTORING REQUIREMENT!3 

Limits must be expressed clearly in the NPDES permit so that they 

clearly are enforceable and unambiguous. Chapter 6 discusses 

compliance monitoring and enforcement problems that can result from improperly expressed limits. All limits, both chemicalspecific and whole effluent, should appear in Part 1 of the permit. 

Special considerations in the use of both chemical-specific and 

whole effluent toxicity limits are discussed below. 

5.7.1 Atmdawd Eftbtmt Llmtb 

Mass-based effluent limits are required by NPDES regulations at 

40 CFR 122.450. The regulation requires that all pollutants 

limited in NPDES permits have limits, standards, or prohibitions 

expressed in terms of mass with three exceptions, including one 

for pollutants that cannot be expressed appropriately by mass. 

Examples of such pollutants are pH, temperature, radiation, and 

whole effluent toxicity. Mass limitations in terms of pounds per 

day or kilograms per day can be calculated for all chemicalspecific toxics such as chlorine or chromium. Mass-based limits 

should be calculated using concentration limits at critical flows. 

For example, a permit limit of 10 mg/l of cadmium discharged at 

an average rate of 1 million gallons per day also would contain a 

limit of 38 kilograms/day of cadmium. 

Mass-based limits are particularly important for control of 

bioconcentratable pollutants. Concentration-based limits will not 

adequately control discharges of these pollutants if the effluent 

concentrations are below detection levels. For these pollutants, 

controlling mass loadings to the receiving water is critical for 

preventing adverse environmental impacts. 

However, mass-based effluent limits atone may not assure attainment of water quality standards in waters with low dilution. In 

110 
these waters, the quantity of effluent discharged has a strong effect 

on the instream dilution and therefore upon the RWC. At the 

extreme case of a stream that is 100 percent effluent, it is the effluent 

concentration rather than the effluent mass discharge that dictates 

the instream concentration. Therefore, EPA recommends that permit limits on both mass and concentration be specified for 

effluents discharging into waters with less than 100 fold dilution 

to ensure attainment of water quality standards. 

5.7.2 EmugytZtmsmdon 

Water quality-based permit limits by themselves do not provide any 

incentive to dischargers to reduce wastewater flows. The reverse is 

true; a more dilute effluent means water quality-based limits are 

more easily achieved. However, increased flow translates into increased power consumption for treatment facilities. Significant power 

usage stems from pumping and mixing of volumes of wastewater in 

treatment systems. If the volume of wastewater can be reduced, 

power consumption can be reduced and less fossil fuel burned. Such 

reductions can be expected to result in concomitant decreases in air 

pollution. 

Therefore, EPA recommends that flow reductions and energy savings 

be specifically encouraged where appropriate (usually in dilutions 

greater than 1OO:l) by allowing water quality-based permit limits to 

be mass-based and by allowing concentration-based limits to vary in 

accordance with flow reduction requirements. The permit also could 

include an energy savings analysis subject to approval by the permitting authority. 

5.7.3 lZWsi&afions intbelkwof~icd-wlfk Umits 

Metals 

Another common problem encountered in expressing permit limits 

occurs for metals. Some water quality standards express numeric 

criteria for metals in terms of the dissolved or acid soluble phase of 

the metal. NPDES regulations at 40 CFR 122.45(c) require permit 

limitations for metals to be expressed in terms of total recoverable 

metal unless (1) an effluent guideline requires the use of another 

form, (2) technology-based limits are established on a case-by-case 

basis, or (3) the approved analytical method measures only the 

dissolved form. 

Where State water quality standards are expressed directly as total or 

total recoverable metals, the permit limit can be established directly. 

Where the water quality standards are expressed as dissolved or acid 

soluble metal, the permit writer will need to reconcile the different 

expressions of metals when establishing the permit limits. Some 

State water quality standards implementation policies or procedures 

provide the requirements for this conversion. In instances where a 

State has no policy or procedure, the permit writer can take one of 

four approaches. First, the permit writer could assume no difference 

between the dissolved or acid soluble phases and the total recoverable phase. This is the most stringent approach and would be most 

appropriate in waters with low solids, where the discharged form of 

the metal was mostly in the dissolved phase, or where data to use the 

other options are unavailable. Second, the permit writer could 

develop a site-specific relationship between the phases of metals by 

developing a relationship through review of information on instream 

metal concentrations. This approach requires concurrent sampling 

of both metal phases during periods reflective of the environmental 

conditions used to determine the WL4. Third, the permit writer 

could use a relationship developed by EPA from national data; 

this relationship is described in the national guidance for determining WlAs for toxic metals in rivers. This relationship requires knowledge of instream concentrations of total suspended 

solids at the environmental conditions used to determine the 

WIA. Fourth, the permit writer could use a geochemical 

model, such as the equilibrium metal speciation model 

MINTEQA2 (see Chapter 4). However, the input data requirement of this model are equivalent to collecting site-specific 

data under Option 2. These options will be expressed in more 

detail in subsequent guidance issued by EPA. 

Updute: The Agency has issued “interim Guidance on Interpretotion and lmplementution Aquatic Life Criteria for Metals. ” See the 

update notice in front of this document for ovoilobility. 

Detection Level Limits 

A commonly encountered problem is the expression of calculated limits for specific chemicals where the concentration of 

the limit is below the analytical detection level for the pollutant 

of concern. This is particular’ true for pollutants that are toxic 

in extremely low concentrations or that bioaccumulate. 

The recommended approach for these situations is to include in Part 1 of the permit the appropriate permit limit 

derived from the water quality model and the WLA for the 

parameter of concern, regardless of the proximity of the 

limit to the analytical detection level. The limit also should 

contain an accompanying requirement indicating the specific 

analytical method that should be used for purposes of compliance monitoring. The requirement should indicate that any 

sample is analyzed in accordance with the specified method 

and found to be below the compliance level will be deemed to 

be in compliance with the permit limit unless other monitoring 

information (as discussed below) indicates a violation. Sample 

results reported at or above the compliance level should be 

reported as observed whereas samples below the compliance 

level should be reported as less than this level. 

The level of compliance cited in the permit must be clearly 

defined and quantifid. For most NPDES permitting situations, EPA recommends that the compliance level be defined in the permit as the minimum level (ML). The ML is 

the level at which the entire analytical system gives recognizable mass spectra and acceptable calibration points. 

This level corresponds to the lowest point at which the calibration curve is determined based on analyses for the pollutant of 

concern in a reagent water. The ML has been applied in 

determinations of pollutant measurements by gas chromatography combined with mass spectrometry. The concept of a 

minimum level recently was used in developing the Organic 

Chemicals, Plastics, and Synthetic Fibers effluent guidelines 

PI. 

The minimum level is not equivalent to the method detection 

level, which is defined in 40 CFR Part 136 Appendix 6 as the 

minimum concentration of a substance that can be measured 

and reported with 99-percent confidence that the analyte concentration is greater than zero and is determined from the analysis of a sample in a given matrix containing the analyte. EPA is 

not recommending use of the method detection level because 

quantitation at the method detection level is not as precise as at 

111 
- 

the ML. It is not similar to the practical quantitation limit (PQL), 

which is typically set as a specific (and sometimes arbitrary) 

multiple of the method detection level. Because the PQL has no 

one definition, EPA is not recommending its use in NPDES permitting. Nor is it similar to other terms such as the limit of detection, 

limit of quantitation, estimated quantitation limit, or instrument 

detection limit. 

The permitting authority may choose to specify another level at 

which compliance determinations are made. Where the permitting authority so chooses, the authority must be assured that the 

level is quantifiable, defensible, and close as possible to the permit 

level. 

Where water quality-based limits below analytical detection 

levels are placed in permits, EPA recommends that special 

conditions also be included in the permit to help ensure that 

the limits are being met and that excursions above water 

quality standards are not occurring. Examples of such special 

conditions include fish tissue collection and analyses, limits and/or 

monitoring requirements on internal waste streams, and limits 

and/or monitoring for surrogate parameters. This information 

can be used to help support reopening the permit to establish 

more stringent effluent limits if necessary. 

5.7.4 ConsMevatioa in the Use of wlrole Effhnt ToxMy 

lhifs 

Test Methods 

NPDES regulations at 40 CFR 122.44(i)(l)(iv) require that methods approved under 40 CFR Part 136 be used for compliance 

monitoring, and in the absence of an approved method, the 

permit must specijl the method to be used. The permit should 

also carefully consider any other case-specific aspects of the whole 

effluent toxicity test method that should be designated in the 

permit. Such aspects as the dilutions at which testing will be 

conducted, the different species to be used, the specific endpoints, the statistical procedures for analyzing the data, quality 

assurance, and other factors should be clearly stated as a permit 

condition to assure that the whole effluent toxicity testing that is 

performed to ascertain compliance with a limit or monitoring 

requirement is the test procedure the regulatory authority desires. 

In some instances, promulgated methodologies allow significant 

flexibility and choice in how the method is actually conducted. A 

simple reference to the methodology in the permit may not result 

in the test being conducted as intended. 

Units of Expression and Detection Levels 

The permit limit for toxicity itself and the detection levels, or 

sensitivity levels, associated with the various types of toxicity tests 

determine the type of monitoring requirement, which should be 

specified with the limit. It is a misconception to think, for example, that only acute toxicity tests should be used where the 

WLA for acute protection is used to derive the more limiting LTA 

or should always be used to monitor for the MDL. It is a similar 

misconception to think that only chronic tests should be used 

where chronic LTA is limiting or should always be used to monitor 

for the average monthly limit. The MDLs and AMLs are derived 

from the more limiting of the two LTAs. Therefore, either acute or 

chronic tests might apply to a given situation depending upon 

the test detection levels or test sensitivity. 

For example, a limit of 5 TU, (no observed effect concentration 

[NOEC] of 20 percent or greater) would require chronic toxicity 

testing where the ACR is 20 for that effluent, An acute test would 

not be sensitive enough to measure effluent toxicity in this instance, since 5 TU, would be equivalent to 0.25 TU,. Conversely, 

if the ACR was 2, then an acute test could be used because 5 TU, 

would be equal to 2.5 TUa. Generally, there is no reason to mix 

two types of monitoring requirements for the same limit when 

limits are derived from the most limiting LTA. Doing so will 

confuse the results and complicate assessments of average monthly 

limits where sampling frequency is greater than once per month. 

The acute toxicity test, when using an LCso as the test endpoint, 

has an upper sensitivity level of loo-percent effluent, or 1 .O TU,. 

If less than 50 percent of the test organisms die at loo-percent 

effluent an LCso cannot be determined from the test data, and 

the true LCso value for the effluent cannot be measured. In this 

situation, an acute test could still be used for compliance monitoring purposes but the endpoint would need to be changed to a 

greater level of sensitivity. The endpoint could be specified in 

terms of “no statistically significant difference in acute toxicity 

between 100 percent effluent sample and the control.” This is 

the most sensitive application of an acute test and could be used 

for monitoring compliance with a limit that, because of lack of 

available dilution, applies the EPA recommended acute criterion 

of 0.3 TU, at the end of the pipe. 

However, these tests would not accurately quantify any level of 

chronic toxicity present. For chronic testing, an effluent with an 

NOEC of greater than 100 percent presents a similar test sensitivity problem. An effluent with an NOEC of greater than 100 

percent contains less than 1.0 TU, and would meet the EPA 

recommended chronic criterion for toxicity at the edge of the 

mixing zone, if dilution were available, as well as at the end of the 

pipe if no dilution were available. 

Description of Limits 

When toxicity limits are used, additional description of the limit is 

required. The limit should be stated in Part 1 as “effluent toxicity” 

in the parameter column with “maximum TUs,” “minimum ATE 

[acute toxicity endpoint],” or “minimum NOEC” in parentheses 

underneath. The numerical values should be placed in the appropriate concentration column followed by TU or a percent sign. A 

footnote should direct the reader to Part 3 for specific requirements on how to conduct the tests. The description in Part 3 

should accomplish the following: 

l Explain how the limit is expressed (e.g., the limit is the 

minimum ATE expressed as percent effluent or the limit is 

the maximum TU,) 

l Specify the test species and the test methods for compliance monitoring purposes 

l Describe any special reporting or followup requirements 

(e.g., requirements to conduct a toxicity reduction evaluation). 

112 
- -~ 

-- 

The language in Part 3 should be modified as needed to suit the 

situation. The following example language is provided only for 

purposes of illustration: 

l “The effluent toxicity limit contained in Part 1 is the allowable chronic toxicity to the most sensitive of three test 

species. It is expressed as the allowable NOEC in percent 

effluent. The required test species and the procedures to 

follow are described in Short Term Methods for Estimating 

the Chronic Toxicity of Effluents and Receiving Waters to Freshwater Organisms, EPAf600/4-89/001, March 1989.” 

l “The permittee shall conduct monitoring of effluent toxicity 

once per month. One 24-hour composite sample shall be 

collected and tested within 24 hours of collection. Results 

shall be reported as the NOEC. Any test that does not meet 

quality control requirements as described in the above 

referenced methods shall be repeated using a freshly collected sample as soon as practicable.” 

57.5 sekction of MoRiforillg freguencles 

There is no fixed guidance on establishment of monitoring frequencies. The decision on the monitoring frequency is casespecific and needs to consider a number of factors, including 

those listed below: 

l Type of treatment process, including retention time 

l Environmental significance and nature of the pollutant or 

pollutant parameter 

l Cost of monitoring relative to the discharger’s capabilities 

and benefit obtained 

l Compliance history 

l Number of monthly samples used in developing the permit 

limit 

l Effluent variability. 

Based upon an array of data analyzed for both individual chemicals and whole effluent toxicity, and independent of other considerations, EPA has observed that ideally 10 or more samples per 

month provides the greatest statistical likelihood that the average 

of the various monthly values will approach the true monthly LTA 

value. In practice, however, selection of monitoring frequencies 

will need to consider the previously mentioned factors and arrive 

at a reasonable compromise of the appropriate considerations. 

Permits require monitoring to establish whether a facility is discharging at a level that complies with the permit limits. All 

monitoring includes analytical variability. The true concentration 

in a sample can be higher or lower than the measured one due to 

this variability; however, there is no way to predict which way it 

will go. 

Historically, EPA has not directly considered analytical variability 

from monitoring methods when establishing permit limits. If the 

upper bound of the analytical variability was added to the limit, 

there would be a higher potential that the permit limit would fail 

to protect the wasteload allocation. This would not be consistent 

with 40 CfR 122.44(d)(l). On the other hand, if the lower bound 

of the analytical uncertainty was subtracted from the limit, there 

would be better assurance that the limit achieved the WLA. This 

approach could be overly conservative given the other factors 

used to develop permit limits. EPA believes that its recommended 

approach provides a balance between these two extremes. 

57.7 kMmksi~dhg 

CWA Section 402(o) establishes express statutory language prohibiting the relaxation of permit limits based on water quality. 

Under the statute, relaxation of water quality-based limits is permissible only if either the requirements of Sections 402(o)(2) or 

303(d)(4) are met. These two provisions constitute independent 

exceptions to the prohibition against relaxation of permit limits. tf 

either is met, relaxation is permissible. 

Relaxation of Water Quality-based Limits Under 

Section303(d)(4) 

Section 402(o)(l) prohibits the establishment of less stringent 

water quality-based effluent limitations “except in compliance 

with Section 303(d)(4).” Section 303(d)(4) has two parts: Paragraph (A), which applies to “nonattainment waters” and Paragraph (B), which applies to “attainment waters.” 

l Nonattainment waters: Section 303(d)(4)(A) allows establishment of less stringent water quality-based effluent limitations in a permit for discharge into a nonattainment 

water only if (1) the existing permit limitation must have 

been based on a total maximum daily load (TMDL) or other 

WLA established under Section 303, and (2) attainment of 

water quality standards must be assured. 

l Attainment waters: Section 303(d)(4)(8) allows establishment of less stringent water quality-based effluent limitations in a permit for discharge into an attained water as 

long as the revised permit limit is consistent with a State’s 

antidegradation policy. This is not restricted to limits based 

on a TMDL or WLA. 

Relaxation of Water Quality-based Limits Under 

Section 402 

Section 402(o)(2) also outlines exceptions to the general prohibition against establishment of less stringent water quality-based 

permit limits in a permit. Under Section 402(o)(2), the establishment of less stringent limits based on water quality may be 

allowed where: 

1) There have been material and substantial alterations or 

additions to the permitted facility which justify this relaxation. 

2) Good cause exists due to events beyond the permittee’s 

control (e.g., acts of God) and for which there is no reasonably available remedy. 

3) The permittee has installed and properly operated and 

maintained required treatment facilities but still has been 

unable to meet the permit limitations (relaxation may only 

be allowed to the treatment levels actually achieved). 

113 
4) New information (other than revised regulations, guidance, 

or test methods) justifies relaxation of water quality-based 

permit limitations. 

This last exception applies to water quality-based permit 

limitations only where the revised limitations result in a net 

reduction in pollutant loadings and are not the result of 

another discharger’s elimination or substantial reduction of 

its discharge for reasons unrelated to water quality (e.g., 

operation termination). 

Although Paragraph 402(o)(2) lists two additional exceptions, 

one for technical mistakes and mistakes of law and one for permit 

modifications or variances, the statute provides that these exceptions do not apply to water quality-based effluent limitations. As a 

result, these exceptions do not provide a basis for relaxing water 

quality-based limitations. 

Relaxation of Water Quality-Based Permit Conditions or Standards 

The provisions in Section 402(o) discussed previously only address the relaxation of effluent limits based on water quality. The 

relaxation of other permit conditions or standards based on water 

quality are governed by EPA’s existing antibacksliding regulations 

at 40 CfR 122,44(l)(l). Under these regulations when a permit is 

renewed or reissued, interim effluent limitations, standards, or 

conditions must be at least as stringent as the final effluent 

limitations, standards, or conditions in the previous permit “unless the circumstances on which the previous permit was based 

have materially and substantially changed since the time the 

permit was issued and would constitute cause for permit modiftcation...“. In other words, unless cause for permit modification is 

present, relaxed conditions or standards are not permissible. EPA 

regulations setting forth cause for permit modification can be 

found at 40 CFR 122.62. 

Restrictions of Backsliding 

Even if any of the backsliding exceptions outlined in the statute or 

regulations are applicable and met, Section 402(o)(3) acts as a 

floor and restricts the extent to which water quality-based permit 

limitations may be relaxed. Paragraph (o)(3) prohibits the relaxation of water quality-based permit limitations in all cases if there 

will be a violation of applicable effluent limitation guidelines or 

water quality standards, including antidegradation requirements. 

This requirement affirms existing provisions of the CWA that 

require permit limits, standards, and conditions to ensure comptiante with applicable technology-based limits and water quality 

standards. 

5.8 T'OXICITY IlEwlcTltM EVAUJATlWS 

Where monitoring indicates unacceptable effluent toxicity, one 

principal mechanism for bringing a discharger into compliance 

with a water quality-based whole effluent toxicity requirement is a 

toxicity reduction evaluation (TRE) [6]. The purpose of a TRE is to 

investigate the causes and to identify corrective actions for difficult effluent toxicity problems. The permitting authority may 

require that the permittee conduct a TRE in those cases where the 

discharger is unable to explain adequately and immediately correct exceedances of a whole effluent toxicity permit limit or 

requirement. 

A TRE is a site-specific study conducted in a stepwise process to 

narrow the search for effective control measures for effluent toxicity. TREs are designed to identify the causative agents of effluent 

toxicity, isolate the sources of the toxicity, evaluate the effectiveness of toxicity control options, and then confirm the reduction in 

effluent toxicity. The ultimate objective of a TRE is for the discharger to achieve the limits or permit requirements for effluent 

toxicity contained in the permit and thereby attain the water 

quality standards for receiving waters. 

The requirement for a permittee to conduct a TRE may be written 

into the special conditions section of a permit, which contains 

whole effluent toxicity limits. In some cases, the permit issuing 

authority may also use other legally binding mechanisms, including Section 308 letters, Administrative Orders, or Consent Decrees, to require a TRE. 

5.8.1 RE 8uMance aOeur#rrrls 

To assist permittees in conducting TREs and achieving compliance 

with whole effluent toxicity limits, EPA has developed a series of 

three guidance documents [6, 7, 81: 

1) Generalized Methodology for Conducting Industrial Toxicity 

Reduction Evaluations (EPAl600/2-88/070) 

2) Toxicity Reduction Evaluation Protocol for Municipal Wostewater Treatment Plants (EPA/600/2-88/062) 

3) Methods for Aquatic Toxicity identification Evaluations: 

Phase 1 Toxicity Characterization Procedures (EPA/600/3- 

88/034) 

Phase 2 Toxicity Identification Procedures (EPA/600/ 

3-88/035) 

Phase 3 Toxicity Confirmation Procedures (EPA/600/ 

3-881036). 

These guidance documents describe the methods and procedures for conducting TREs and Toxicity Identification Evaluations 

(TIES). They are based on the results of EPA’s continuing efforts in 

TRE methods research and case study applications. Separate TRE 

guidance has been developed for industrial dischargers and municipal wastewater treatment plants to better address the circumstances of each type of facility. Procedures for the characterization, identification, and confirmation of the causative agents of 

effluent acute toxicity have been developed and are described in a 

three-phased TIE methods manual. These TIE methods are applicable to both industrial and municipal effluents and are an integral part of the protocols for TREs described in the industrial and 

municipal TRE guidance documents. TIE methods using chronic 

toxicity tests for identifying toxicants will soon be developed and 

available in a draft guidance document. 

5.8.2 Ret- &mJ8chflulA%nwmgl7r& 

To ensure the successful completion of a TRE, the guidance 

documents recommend a systematic, stepwise approach that 

114 
eliminates the possible causes or sources of toxicity until a solution 

or control method is determined. The guidance documents 

discourage “playing hunches” or implementing extensive control 

measures solely on the basis of unsubstantiated conclusions (e.g., 

selecting and implementing a treatment plant upgrade without 

adequate information). Experience shows that unnecessary delays 

and expenditures in achieving the objective of the evaluation are 

avoided by building a sound scientific and engineering basis for 

selection of a control method. This can best be done by the 

logical interpretation of the information and data collected in a 

systematic approach to a TRE. The causes or control methods 

identified should then go through a confirmation stage. This is 

especially important in cases where the control method selected 

requires the construction of additional treatment. A flow chart, 

generalized from the guidance documents, for this approach to 

TREs is presented in Figure 5-l 0. The steps in this flow chart are 

summarized in the following discussion. 

Determination of TRE Objectives and Development of the TRE 

Plan 

Obviously, the success of any study is dependent on a clear 

understanding of what is to be achieved and how these objectives 

are to be demonstrated and measured. Typically, TRE objectives 

are set by the regulatory authority in terms of a toxicity test 

endpoint (ATE or chronic toxicity endpoint [CTE]) in order to 

I I 

TRE Regulatory 

Requirements 

r-l lnlorrnatlon and 

Data Acqulsltlon 

Facility Operation 

and Maintenance 

IdenUficatlon 

Control Method 

Selectctlon and 

Implemefltatlon 

FoHowup and 

Conflrmatlon 

I I 

meet a limit or permit condition. TRE plans should be submitted 

by the discharger as soon as possible. In some cases, this could be 

30 to 60 days following notification that a TRE is required. In 

other instances, this period could be longer. These plans are 

important for ensuring that the TRE objectives are well understood and that the TRE to be conducted is thorough and represents a reasonable effort to achieve the required reduction in 

effluent toxicity. An implementation schedule should also be 

developed describing the timeframe for completion of the specific 

components of the TRE plan by the required TRE completion 

date. This schedule should be submitted for review in conjunction with the TRE plan. EPA recommends that the TRE schedule 

should be set or approved by the regulatory agency. Approval 

of the schedule and the completion date should not imply approval of the TRE plan itself or the procedures and methods 

outlined in the plan. Instead, the TRE plan should only be 

reviewed and any comments provided to the permittee as needed. 

To assist in this review, Box 5-3 provides evaluation criteria for TRE 

plans. The permitting authority should review the TRE plan and 

inform the discharger of any apparent shortcomings or potential 

problems. The TRE should not be delayed pending completion of 

the review of the plan. The specified completion date for the TRE 

must still be met and the permittee should be expected to begin 

steps to investigate and alleviate the effluent toxicity as soon as 

possible following notification that a TRE is required. bring the 

course of the TRE, the regulatory agency should provide oversight, as time permits, to make the TRE as effective as possible. 

Evaluation of Existing Site-specific Information 

The next step involves the collection of any information and 

analytical data relevant to the effluent toxicity. The permittee 

should begin collecting and evaluating this information as soon as 

possible following notification that a TRE is required. In some 

cases, this step may be conducted concurrently with accelerated 

toxicity testing as part of the development of a TRE plan. For an 

industrial discharger, this part of the evaluation would include 

information such as plant and process information, influent and 

effluent physical and chemical monitoring data, effluent toxicity 

data, and material use. For a POlW, additional information, such 

as industrial waste survey applications, local limits compliance 

reports, and monitoring data, should be collected. This information is used to supplement the data generated in the later steps of 

the TRE and may be useful at that stage to point to potential 

sources or treatment options. 

Evaluation of Facility Operations and Maintenance 

Practices 

This part of the evaluation is performed in order to ascertain 

whether the facility is consistently well operated and whether the 

effluent toxicity is Ihe result of periodic treatment plant upsets, 

bypass, or some other operational deficiency that may be causing 

or contributing to the effluent toxicity. This part of the TRE should 

be initiated immediately after notification that a TRE is required. 

Alternatively, the permittee may begin to conduct this step at the 

same time that any accelerated toxicity testing is required. At 

both municipal and industrial facilities, this step would involve the 

evaluation of “housekeeping,” treatment system operation, and 

chemical use. In some cases, best management practices (BMPs) 

Figure 5-10. Generalized TRE Flow Chart 

115 
may be identified, which would improve operations and effluent 

quality. However, the effectiveness of BMPs in reducing effluent 

toxicity should be carefully confirmed, and it will usually be 

necessary to test a number of samples and perhaps to conduct 

Phase 1 of the TIE to develop this level of certainty. The results of 

this evaluation may lead to preliminary strategies for source reduction and pollution prevention, including spill or leak prevention, improvements in material handling and disposal practices, 

or substitution or re-use of a compound known to be highly toxic. 

Toxicity Identification Evaluation 

TIE procedures are performed in three phases: characterization, 

identification, and confirmation [7]. In each phase, aquatic organism toxicity tests are used to track toxicity at each step of the 

procedure. In most cases, these are abbreviated or shortened 

nature of the causative agents of effluent toxicity or toxicants is 

determined. This is done by conducting a battery of tests to 

characterize the physical/chemical characteristics of the toxicity: 

solubility, volatility, decomposability, complexibility, filterability, 

and sorbability. This information can then be used to decide 

which chemical analytical methods will to use in Phase 2 or it can 

be used to design treatability studies. 

The results of Phase 1 also may be used to provide additional 

confirmation of the effectiveness of any BMP that was implemented in the previous step of the TRE to reduce the effluent 

toxicity. This would require conducting at least one Phase 1 

analysis prior to implementation of the BMP (i.e., any source 

control method implemented as a result of the evaluation of 

facility operation and maintenance). The results of this analysis 

would then be compared with Phase 1 results from samples taken 

toxicity tests. In the toxicity characterization phase, the general after BMP implementation. 

Box 5-3. Evaluation Criteria for TRE Plans 

. Are the objectives or targets of the TRE stated clearly and accurately? 

. Are the schedule and milestones for accomplishing the tasks described in the study plan? 

. Are the final TRE report, progress reports, and meetings with the regulatory authority included as part 

of the schedule? 

. Are the approaches or methods to be used described to the extent possible prior to beginning the 

TRE? 

. Has available EPA guidance been used in designing the TRE and developing the TRE plan (or if other 

methods are proposed, are these sufficiently documented)? 

1 Does the TRE plan specify what results and data are to be included in the interim and final reports? 

l Does the TRE plan provide for arrangements for any inspections or visits to the facility or laboratory 

that are determined to be necessary by the regulatory authority? 

. Are the toxicity test methods and endpoints to be used described or referenced? 

. Does the approach described build on previous results and proceed by narrowing down the possibilities in a logical progression? 

. Does the plan provide for all test results to be analyzed and used to focus on the most effective 

approach for any subsequent source investigations, treatability studies, and control method evaluations? 

. Are optimization of existing plant/treatment operations and spill control programs part of the initial 

steps of the TRE? 

l Does the TRE plan allow a sufficient amount of time and appropriate level of effort for each of the 

components of the study plan? 

. Does the TIE use broad characterization steps and consider quantitative and qualitative effluent 

variability? 

. Is toxicity tracked with aquatic organism toxicity tests throughout the analyses? 

. Is the choice of toxicity tests for the TRE logical and will correlations be conducted if the species used 

are different from those used for routine biomonitoring? 

. Is the laboratory analytical capability and the expertise of the investigator broad enough to conduct 

the various components of the evaluation? 

116 
In Phase 2 of the TIE, the results of Phase 1 are built upon, and the 

TIE proceeds to chemical analyses designed to identify the specific 

chemicals causing effluent toxicity. In Phase 3, the identified 

toxicants are confirmed using a number of procedures, including 

correlation of toxicity with chemical concentration, spiking experiments, toxicity mass balance, and additional test species and 

their symptoms. 

The current version of the TIE methods uses acute toxicity tests to 

characterize and identify the toxicants. In some cases, these 

methods may also be used for TREs where the objective is to 

reduce chronic toxicity. In order for these methods to be applicable, however, there must be some measurable acute toxicity in 

the effluent samples that are to be characterized in Phase 1 and 

analyzed in Phase 2. If this approach is used, the appropriate 

chronic toxicity test, as specified in the TRE objectives and permit 

requirements, should then be used in the Phase 3 confirmation 

procedures. This will confirm that the toxicant identified using 

acute tests in Phases 1 and 2, are indeed causing the whole 

effluent chronic toxicity, which must be reduced. 

It is possible to use the methods and procedures described in the 

other components of the overall TRE with either acute or chronic 

toxicity tests. The fact that the previous version of the EPA TIE 

methods use acute toxicity tests should not be construed to mean 

that TREs cannot be required or conducted for the reduction of 

chronic toxicity. These methods provide additional tools to assist 

permittees in the reduction of whole effluent chronic toxicity. 

Phase 1 procedures that use chronic toxicity tests will soon be 

available in draft EPA guidance. These TIE methods are applicable 

to freshwater discharges to either saltwater or freshwater receiving waters. The use of these methods for saltwater receiving 

waters may require their adaption for use with marine test species 

or, preferably, an initial correlation of the recommended freshwater TIE test species to the marine species used for monitoring. 

Source Investigation 

Based on the results of the TIE, a decision is made on whether to 

conduct treatability studies on the final effluent and/or conduct a 

source investigation. A source investigation is most readily performed when the specific toxicants have been identified and 

influent samples can be analyzed for the presence of these compounds or when potential source streams can be selected for 

chemical analysis (based on the results of the initial data acquisition step). However, in some cases where the specific causative 

agents of effluent toxicity have not been identified in the TIE, it 

may be possible to conduct a source investigation by “treating” 

influent samples in bench-scale models of the facility treatment 

plant, measuring the toxicity of the treated sample and then 

tracking this toxicity to its source. 

Source investigations will lead to control methods, such as chemical substitution, process modification, treatment of process or 

influent streams (pretreatment), and possible elimination of the 

process. For POTWs, source investigations may lead to the development of local limits or to the requirement that an indirect 

discharger evaluate and control their effluent so as to reduce its 

toxicity and prevent passthrough at the POTW. The implementation of source control methods can effectively reduce effluent 

toxicity and also can avoid any cross-media transfer of pollutants 

to air or sludge, which may occur as a result of end of pipe 

treatment. Types of source control methods that have proven to 

be effective in reducing effluent toxicity are improvements in 

facility housekeeping, chemical substitution, process optimization, reclamation/reuse, and pretreatment. 

Toxicity Treatability Evaluation 

Toxicity treatability evaluations are conducted to identify possible 

treatment methods that can effectively reduce effluent toxicity 

and may involve modifications or additions to the existing system. 

Treatability studies generally use the same type of information on 

the nature of the chemicals to be removed as is generated by 

Phase 1 of the TIE. These treatability tests should be conducted 

on a bench-scale initially and then a pilot scale prior to construction of additional treatment or substantial modification of the 

existing plant. The use of these bench- and pilot-scale tests, 

coupled with aquatic organism toxicity tests, should be used to 

confirm the effectiveness of the treatment option. Confirmation 

of the results of treatability studies is equally important as it is for 

the TIE. Skipping this confirmation step is an invitation for 

unwarranted expense. 

Toxicity Control Method Selection and Implementation 

After the investigative steps of the TRE are completed, it is not 

unusual for a number of possible control options to have been 

identified. At this point, a site specific selection must be made by 

the discharger based on the technical and economic feasibility of 

the various alternatives. Following this selection, the toxicity 

control method is implemented or a compliance plan is submitted if construction of additional treatment requires a substantial 

amount of time. 

Followup and Confirmation 

After the control method is implemented and the final TRE report 

is submitted, the permitting agency should direct the permittee 

to conduct followup monitoring to confirm that the reduction in 

effluent toxicity is attained and maintained. Normally, this monitoring should follow an accelerated schedule, weekly or biweekly 

toxicity tests, for a period of 2 to 3 months to confirm the 

effectiveness of the controls implemented and the continued 

attainment of the TRE objective. This followup monitoring should 

use the same species as were specified for routine toxicity testing 

in the permit. The test endpoints of these toxicity tests should be 

the same as those which were calculated by the water qualitybased permit limit derivation procedure used when the permit 

was issued. Once the discharger has demonstrated the successful 

completion of the TRE, the permitting agency should direct the 

discharger to return to the routine permit monitoring schedule. 

5.8.3 l%wmstmcas Wmattting a TRE 

It is the responsibility of the permitting authority to determine if 

the permit limits and/or the State water quality criteria have been 

threatened or violated and to notify the permittee if a TRE is 

required. It is appropriate for the permitting authority to require 

additional toxicity testing following the initial exceedance or violation. This additional testing may precede notification that a TRE 

will be required or it may be considered as the initial part of the 

117 
TRE and be conducted simultaneously with TRE plan development and the evaluation of other existing site-specific information. 

It is important to recognize that the purpose of this additional 

toxicity testing is to determine the continued presence or absence 

of effluent toxicity and the magnitude of that toxicity. This 

information can then be used to determine the continued compliance or noncompliance with the limit or permit conditions for 

effluent toxicity. These tests do not serve to verify or confirm the 

initial test results from an earlier sample. Instead, the permit 

authority shall use the results of these tests to determine if a TRE or 

some other action is the appropriate response to the initial occurrence of toxicity. 

If the permit has a limit for whole effluent toxicity, then generally, 

the permit should not include any specific conditions for accelerated toxicity testing or for triggering a TRE or some other action 

(e.g., exceedances in two consecutive tests or exceedances in any 

three out of five tests). CWA Section 309 requires that any single 

violation of a permit limit may be subject to enforcement. The 

EPA Compliance Monitoring and Enforcement Strategy for Toxics 

Control (January 19, 1989, Appendix 8-4) states that, “Each 

exceedance of a directly enforceable whole effluent toxicity limit 

is of concern to the regulatory agency and therefore qualifies as 

meeting the VRAC [violation review action criterion] requiring 

professional review.” Accelerated monitoring should only be 

used to assist in this professional review to determine what, if any, 

enforcement response is necessary, including the need for the 

permittee to conduct a TRE. It will be necessary for the Region or 

State regulatory authority to determine this on a case-by-case 

basis. This must be done in a manner consistent with the priorities established in their respective toxics control strategies and 

permitting procedures. 

In situations where it is determined that accelerated testing is 

appropriate, a maximum of weekly tests for a minimum period of 

2 months is recommended. This would result in eight tests, plus 

the routine monitoring toxicity test that initially indicated the 

exceedence or violation, for a total of nine tests in the series. As a 

practical approach for determining if a TRE is an appropriate 

response, EPA recommends if toxicity is repeatedly or periodically present at levels above the effluent limits more than 20 

percent of the time, a TRE should be required. With toxicity 

present at this rate, the TRE protocols will be useful. 

In most cases, any one additional exceedance (beyond the initial 

routine monitoring toxicity test result) in the accelerated toxicity 

tests could result in notification of the permittee that a TRE is 

required. Exceptions to this guideline might include cases where 

the permittee is able to adequately demonstrate that the cause of 

the exceedances is known and corrective actions have been immediately implemented or cases where additional test quality 

assurance/quality control (QA/QC) is necessary or desirable. The 

submittal of QC fact sheets for self-biomonitoring (e.g., Appendix 

6-2) should always be recommended to avoid QA/QC problems. 

If the test results indicate that toxicity is not consistently or 

repeatedly present in the test series, previous discharge monitoring reports (DMRs) should be examined to ascertain if a recurrent 

problem exists. If the problem is recurrent, a TRE should be 

required, and the TRE plan should explain how the design of the 

evaluation will address this periodic or recurrent effluent toxicity 

problem. In these cases, more elaborate sampling design and 

influent or process stream monitoring may be needed. It should 

be expected that TREs conducted under these circumstances will 

probably require a more flexible schedule and perhaps additional 

time before the required completion date. 

If the accelerated testing and previous DMRs show the continued 

absence of effluent toxicity, then the initial exceedance would be 

considered an episodic event and a TRE should not be required. A 

TRE is not an appropriate response to a single, episodic effluent 

toxicity event (e.g., a spill or a plant upset). By conducting 

accelerated testing following a violation or exceedance of a permit condition, unnecessary TREs can be avoided. Similarly, conducting accelerated testing as part of the initial steps of a TRE will 

allow for the TRE to be ended in its very early stages if the toxicity 

is immediately controlled or determined to be episodic or nonrecurrent. By following the TRE guidance and incorporating accelerated testing into the TRE, unnecessary analyses and expense can 

be avoided. 

It also is important to note that for the practical purposes of 

conducting a TRE (as opposed to the purpose of determining if a 

TRE should be required or not), the magnitude of the effluent 

toxicity needed to conduct a TRE may be less than the magnitude 

or level set as the permit limit or permit monitoring condition. 

This is because if the limit or monitoring condition is water 

quality-based then some amount of dilution will usually be incorporated in determining the unacceptable level of effluent toxicity. 

In some cases, it may be possible for the TRE procedures to be 

carried out even if the toxicity does not actually exceed this 

permitted level. This will be the case as long as the effluent 

toxicity is periodically or consistently present in measurable 

amounts in samples of loo-percent effluent. 

It also is reasonable for a discharger to initiate a TRE prior to the 

establishment of a permit limit for toxicity if unacceptable levels of 

toxicity are found in the effluent through routine monitoring or 

through inspection and compliance sampling by the regulatory 

authority. Under these circumstances the regulatory authority 

will need to identify what constitutes unacceptable levels of toxicity since this will not be defined by a permit limit (see Chapter 3 

on determining the reasonable potential for excursions of water 

quality standards). It also is not unreasonable for the discharger 

to voluntarily initiate a TRE under these circumstances. 

There are a number of mechanisms that can be used to require a 

TRE. In most cases, the TRE should be required by a Section 308 

letter or by an enforcement action, such as a Section 309 Administrative Order or a Consent Decree. The permittee should receive 

notification from the permit authority of what response is required. This enables the permit authority to assess whether a TRE 

is the appropriate action to pursue. If effluent toxicity reappears 

following the successful completion of a TRE, then the permit 

authority should be able to review this type of situation to determine if an additional TRE is appropriate or if some other action is 

required. In general, when the permit is issued with whole 

effluent toxicity limits in Part 1 of the permit, TRE requirements 

should be used where necessary to bring the permittee into 

compliance with those limits. Box 5-4 provides example lan118 
guage for effluent toxicity limits, developed as part of the Whole 

Effluent Toxicity Basic Permitting Principles and Enforcement Strategy (Appendix B-4). 

Box 5-5 presents sample language for use in requiring TREs by a 

Section 308 letter or a Section 309 Order. This sample language, 

especially the reporting dates, should be tailored to fit the specific 

permittee. The completion date should be specified on a case-bycase basis. Factors to consider in setting this completion date 

include the type of facility, the variability of the effluent, and the 

previous compliance history. In order to conduct a TRE, reasonable timeframes are 6 to 18 months for an industrial discharger 

and 12 to 24 months for a municipal wastewater treatment plant. 

For POlWs, it may take longer to conduct a TRE due to lengthy 

government contracting procedures, large sewer collection systems, and less influent constituent control. It should be recognized that extensions to these initial timeframes may be granted if 

the progress reports demonstrate that this is warranted. In situations where reductions in chemical concentrations to meet chemical-specific limits are needed as well as reductions in effluent 

toxicity, the timeframes may be adjusted to enable those efforts 

to proceed simultaneously. 

Box 54. Model Permit Language for Effluent Toxicity Limits 

Part 1 .A. Final Effluent Limits and Monitoring Requirements 

During the period beginning on the effective date of this permit and lasting until the expiration date, the 

permittee is authorized to discharge in accordance with the following limits and monitoring requirements 

from the following outfall(s): 001. 

Effluent Characteristic Discharge Limit Concentration 

Reporting Daily Monthly 

Code/Units Parameter Maximum Average 

Monitoring Requirement 

Measurement Sample 

Frequency Type 

-TUc Toxicity 10.0 5.0 x/month composite 

The permittee shall use the toxicity testing and data assessment procedures described in Part 3.8 of this 

permit. 

Box 5-5. Example Language for Requiring Toxicity Reduction Evaluations 

The discharger shall demonstrate that effluent toxicity-based permit limits described in Part 1 .A. of the 

permit are being attained and maintained through the application of all reasonable treatment and/or 

source control measures. Upon identifying noncompliance with those limits the discharger shall initiate 

corrective actions according to the following schedule: 

Task Deadline 

1. Take all reasonable measures necessary to Within 24 hours 

reduce toxicity immediately. 

2. Submit a plan and schedule to attain continued Within 30 days 

compliance with the effluent toxicity-based permit 

limits in Part I.A.,where source of toxicity is known, 

if immediate compliance is not attained. 

3. Submit a TRE study plan detailing the toxicity Within 45 days 

eduction procedures to be employed where source is 

unknown and toxicity cannot be immediately controlled 

through operational changes. EPA’s Toxicity Reduction 

Procedures, Phases 1, 2, and 3 (EPA-600/3-8Bi034, 035, 

and 036) and TRE protocol for POTWs (EPA-600/2-88i062) 

shall be the basis for this plan and schedule. 

119 
Box 5-5. Example Language for Requiring Toxicity Reduction 

Evaluations (continued) 

4. Initiate TRE plan. Within 45 days 

5. Comply with approved TRE schedule. Immediately upon approval 

6. Submit results of the TRE, including summary of Per approved schedule 

findings, corrective actions required, and data generated. 

7. Implement TRE controls as described in the final report. On due date of final report 

per approved schedule 

8. Complete TRE implementation to meet permit limits Per approved schedule, but 

and conditions. in no case later than XX 

months from initial noncompliance. 

120 
CHAPTER 5

REFERENCE5 

1. Marr, J.K., and R.P. Canale. 1988. Load Allocation for Toxics 5. U.S. EPA. 1987. Development Document for Effluent CuideUsing Monte Carlo Techniques. /ourno/ WPCF 60(5):659- lines and Standards for the Organic Chemicals, Plastics and 

66. Synthetic fibers Point Source Cateqow, volumes 1 and 2; EPA 

2. Freedman, P.L., 1.F. Penderqast, C. Wilber, and S.C. Chanq. 440/l -87/009. 

- , 

1988. Seasonal Changes and Effluent Limits. )ouma/ WP<f 6. U.S. EPA. 1988. Toxicity Reductions Evaluation Protocol for 

60(3):317-23. Municipal Wastewater Treatment Plants. EPA/600/2-88/ 

3. Parkerton, T.F., S.M. Stewart, K.L. Dickson, J.H. Rodgers, and 062. 

F.Y. Saleh. 1989. Derivation of Site-Specific Water Quality 7. U.S. EPA. 1988. Methods for Aquatic Toxicity ldentificotion 

Criteria for Zinc: Implications for Wasteload Allocation. Evaluations: Phase 1, Toxicity Characterization Procedures 

Research journal WPCF 61(11 ,12):1636-44. (EPA/600/3-88/034); Phase 2, Toxicity identification Proce4. Shell Chemical Company NPDES Permit Nos. LAOO05762, 

LAOO50962, TXOO048663, Appeal No. 85-l 4, 85-15, 85 

dures (EPA/600/3-88/035); and Phase 3, Toxicity Confirmation Procedures (EPA/600/3-88i0360. 

16 before the Administrator,’ U.S. EPA. U.S. .EPA Judicial 8. U.S. EPA. 1988. Generalized Methodology for Conducting InOfficer decision, October 20, 1987. dustrio/ Toxicity Reduction Evaluations. EPAJ600/2-88/070. 

ADDlTlONAl REFERENCES 

Aitchinson, J., and J.A.C Brown. 1963. The Lognormal Distributions. London: Cambridge University Press. 

Cilliom, R.H., and D.R. Helse. 1986. Estimation of Distributional Parameters for Censored Trace Level Water Quality 

Data 1 and 2. Water Resources Research 22:135-55. 

Kahn, H.D., and M.B. Rubin. 1989. Use of Statistical Methods in Industrial Water Pollution Control Regulations in the 

United States. Environmental Monitoring and Assessment 

12:129-48. 

Shumway, R.H., A.S. Azari, and P. Johnson. 1989. Estimating Mean Concentrations Under Transformation for Environmental Data with Detection Limits. Technometrics 

31(3):347-56. 

121 
6. COMPLIANCE MONITORING AND ENFORCEMENT 

6.1 INTRODUCTION 

Once a water quality-based permit containing limitations and 

conditions to control effluent quality is issued, the permittee is 

responsible for attaining, monitoring, and maintaining compliance with the requirements of that National Pollutant Discharge 

Elimination System (NPDES) permit. Failure to comply with any 

requirements stated in the permit is a violation of the Clean Water 

Act (CWA). 

The Environmental Protection Agency (EPA) and authorized State 

agencies are responsible for tracking compliance with and enforcing NPDES permit requirements in the enforcement of the CWA. 

Section 308 of the CWA and equivalent State statutes enable the 

regulatory agency to verify compliance with permit conditions 

(including water quality-based toxics limitations and compliance 

schedules) by authorizing the agency to impose on permittees 

requirements for sampling and analysis, record-keeping, and reporting. Section 308 also authorizes access by EPA or State 

agencies to facilities and records for verifying compliance with 

permit conditions. All records associated with monitoring must 

be maintained by the facility and available for a 3-year inspection 

period in conformance with 40 CFR Part 122.41. 

The CWA establishes the authority to enforce water quality-based 

permit conditions. The ability to enforce water quality-based 

permit conditions, however, relies on well-written, clearly stated 

permits. The enforcement official must be familiar with the 

process by which permit requirements were derived, including 

the procedures used to determine the wasteload allocation based 

on applicable water quality standards and the procedures used to 

derive limitations from the wasteload allocation. 

6.2 PERMIT REQUIREMENTS 

The conditions that are to be included in NPDES permits are 

described in 40 CFR Part 122 Subpart C. In general, permits 

include effluent limitations, schedules of compliance, and accompanying reporting requirements. Permits should prescribe the 

self-monitoring procedures, frequency of analysis, sampling location and procedures, acceptable or required analytical techniques, 

and frequency of reporting. Permits often require that analytical 

methods referenced in 40 CFR 136 be used for analysis, but may 

specify methodology not included in Part 136 for pollutants with 

no approved methods or where the approved method is inappropriate for a particular permit limitation. Permits should define any 

effluent limitations and explain specific procedures for calculating 

averages of data if different from arithmetic averaging. Permits 

should identify what information must be retained by the permittee, and what data must be submitted to EPA or the State. Results 

from self-monitoring required by the permit are reported on 

discharge monitoring reports (DMRs) that generally are submitted monthly. Sampling and analysis that is done more frequently 

than required by the permit must be included in the DMR. 

6.3 COMPLIANCE MONITORING 

Since most of the routine information gathered in compliance 

monitoring results from permittee self-monitoring, quality assurance (QA) is as important as compliance with limits. It is essential 

that permittees develop and adhere to a QA plan consistent with 

the required monitoring and analyses. The permittee is responsible 

for maintaining data to demonstrate compliance with QA procedures established in the test methodology or as specified in the 

permit. 

The regulatory agency generally has three ways of determining 

compliance with an NPDES permit and assuring adequate QA: 

self-monitoring reports, DMR/QA results, and inspections. Each 

of these methods is discussed below. 

6.3.1 Self-monitoring Reports 

Self-monitoring reports provide much of the compliance data 

used by the regulatory authority in the review of permittee compliance. These reports include DMRs and reports of progress on 

compliance schedules. DMRs contain information on the sampling method, frequency and location, and analytical results of 

permittee self-monitoring. These data and data from progress 

reports on major schedule milestones must be entered into the 

Permit Compliance System (PCS), a computerized data base, by 

the State or EPA [1]. When the required data are entered into the 

system, PCS will automatically “flag” violations of permit limitations, 

compliance schedules, and reporting requirements. 

In order to detect any problems with the quality of the sample 

analysis, it is often desirable to obtain QA information with the 

self-monitoring data. For this reason, several States and Regions 

have developed additional QA forms to accompany permittee 

self-monitoring reports, This additional information may be required through the permit or through a Section 308 order. The 

QA data are compared to a reference QA data sheet that can be 

completed by the regulatory authority to indicate acceptable 

ranges of values for the required protocol. Appendix B-5 provides 

an example of a reference QA data sheet for a whole effluent 

toxicity test. Once completed, this QA data sheet can be included 

in the compliance file for quick reference by compliance personnel. 

It is important to note that poor QA is a violation if the permit 

explicitly specifies adequate QA or references an acceptable protocol with corresponding QA procedures. It also is important to 

note that the signatory’s certification of effluent data certifies 

123 
compliance with the specified protocols. Any problems with QA 

should be reported at the time of DMR submission and the testing 

repeated. 

6.3.2 Discharge Monitoring Report/Quality Assurance 

(DMR/QA) 

The DMR/QA program evaluates a permittee’s ability to analyze 

and report accurate data. This program is intended to improve 

overall laboratory analytical performance for self-monitoring data. 

Authority for requiring participation is granted in CWA Section 

308. In the DMR/QA program, permittees are required to analyze 

“blind” samples with constituents and concentrations that can be 

found in their industrial or municipal wastewaters. The permittees’ results are compared to the known content of the sample, 

and an evaluation of the reported data is sent to the permittees. 

Permittees are expected to use the same personnel and methods 

employed for reporting NPDES data to analyze the samples. 

Permittees are required to follow the instructions for reporting 

results and include a signed certification statement in accordance 

with 40 CFR 122.22. 

Regulatory agencies conduct followup investigations to address 

poor or incomplete DMR/QA results, failure to participate, or late 

submittal of DMR/QA results. DMR/QA performance results are 

compiled annually. 

In the past, only chemical-specific analyses were tested in the 

DMR/QA program. The Environmental Monitoring and Support 

Laboratory (EMSL) in Cincinnati has developed a reference toxicant 

DMR/QA sample for permittees with whole effluent toxicity 

monitoring requirements. National implementation is occurring 

in 1991. 

6.3.3 Inspections 

Inspections are conducted by the regulatory authority or its contractors to address specific violations or problems and to verify 

permittee compliance with permit conditions and QA procedures. 

Inspections may include reviewing records, inspecting treatment 

facilities, assessing progress with compliance schedules, evaluating 

laboratory facilities and performance, and collecting samples for 

analysis or “splitting” samples taken by the permittee for concurrent analyses. EPA has defined several types of inspections based 

on the tasks that are included in the NPDES Compliance inspection 

Manual[2]. Because regulatory authorities are expected to inspect 

all major permittees annually regardless of compliance status, 

nonsampling inspections (which are generally less resource-intensive) are encouraged for routine evaluation of permittee performance. However, sampling inspections are still encouraged to 

address permitting and enforcement priorities. For that reason, 

the regulatory agency must have the full capability to assess 

effluent compliance through inhouse resources or contract support. 

Inspections that focus on toxics control can provide useful information for water quality assessment and permit reissuance in 

addition to compliance data. Procedures for inspecting facilities 

with toxicity testing requirements and measuring effluent toxicity 

are detailed in the NPDES Compliance inspection Manual, Chapter 

7[2]. 

6.4 VIOLATION REVIEW 

Review of permittee self-monitoring data to determine appropriate enforcement response generally involves a two-tiered review. 

The first tier is a preliminary review for timely, complete data that 

indicates compliance with permit requirements. Minor violations 

of requirements are often handled through informal phone calls 

or warning letters that do not require extensive review or oversight. As violations increase in magnitude, duration, or frequency, 

they generally are assigned to personnel who are responsible for 

the second-tier review (determining what enforcement action, if 

any, is appropriate). The guidelines for this process are presented 

in the Enforcement Management System (EMS) [3], but the basic 

concepts of responsible compliance tracking of water qualitybased requirements are discussed below. Section 6.5 discusses 

the enforcement decision process. 

When the initial review of effluent monitoring data indicates that 

unacceptable analytical methods were used by a permittee or its 

contract laboratory, the results should be assigned for review by 

personnel qualified to determine the significance of the results. If 

the monitoring is insufficient to determine compliance with effluent limitations, a warning letter or Section 308 letter requiring 

that the tests be repeated using acceptable procedures would be 

an appropriate response. 

Tracking a permit or Section 308 letter that contains “monitor 

only” requirements requires both a compliance review (e.g., to 

determine if results of acceptable quality were submitted on 

time), and an action review (e.g., to determine if the permit 

should be modified or re-issued to include a limitation). This 

second review should be assigned to personnel who are qualified 

to make this regulatory decision. 

In addition to the guidelines for reviewing monitoring data in the 

absence of a specific effluent limitation, EPA also has recommended 

a criterion for determining which effluent violations must be assigned for review by a professional who will determine if a formal 

enforcement action is needed, or if a phone call, warning letter, or 

Section 308 letter is more appropriate. These criteria are known 

as the Violation Review Action Criteria and are listed in the EMS. 

In the case of a whole effluent toxicity limitation, any violation 

must be reviewed by a qualified professional responsible for the 

enforcement decision. EPA makes this recommendation to ensure 

that adequate attention is given to QA and to ensure that additional 

testing is required if permitted testing frequency is less than once 

per month. 

In the case of a violation of a chemical-specific permit limitation, 

EPA recommends that monthly average limitation violations be 

reviewed by a professional for potential enforcement response 

whenever two or more violations occur in a 6-month period. 

Seven-day average and daily maximum violations should likewise 

be reviewed if a minimum of two or four, respectively, occur 

during the course of 1 month. Although there is no delineation 

between technology-based versus water quality-based limitations 

in these Violation Review Action Criteria, Regions and States may 

wish to adopt a criteria of “any violation” for all water quality124 
..-_ 

based, chemical-specific limitations as these criteria are solely to 

determine the level of review and do not prescribe enforcement 

action. 

6.5 EWFORCEMENT 

Effective enforcement of toxic controls depends upon clearly 

expressed requirements in NPDES permits. These controls are 

generally in the form of numeric limits on specific toxic chemicals 

or whole effluent toxicity and schedules to initiate construction or 

other compliance measures. 

Exceeding a permit limitation is a violation subject to enforcement. Some members of the regulated community have expressed 

concerns that single violations of stringent water quality-based 

limitations will result in unreasonable enforcement actions. EPA’s 

guidance outlines a systematic review of all violations to determine 

the appropriate level of response. This guidance generally suggests 

an informal response for minor or infrequent violations, escalating 

to formal enforcement and perhaps penalties for more frequent 

and environmentally harmful violations. 

In evaluating appropriate response to violations, EPA’s “Enforcement Response Guide” of the EMS should be used for guidelines 

on the minimum acceptable response [3]. 

Further guidance on addressing violations of whole effluent toxicity 

limitations in particular is presented in the Compliance Monitoring 

and Enforcement Strategy for Toxics Control [4] (see Appendix t3- 

4). This strategy expects that all available avenues to compliance 

will be explored by the permittee, that the treatment facility is 

designed, constructed, maintained, and operated to achieve all 

water quality-based, chemical-specific or best available technology/ 

secondary treatment limitations, that chemical or process substitutions have been attempted and pretreatment explored, and 

that, in the case of publicly owned treatment works (POTWs), 

pretreatment program requirements and local limits have been 

established and enforced. The strategy further expects that the 

permittee will pursue a Toxicity Reduction Evaluation (TRE) as 

discussed in Chapter 5 in compliance with enforcement requirements or under its own initiative. If all of these expectations have 

been met and the facility is unsuccessful in identifying the cause, 

source, or treatability of toxicity despite making good-faith efforts 

to do so, the strategy allows for relief from civil penalties. The 

underlying responsibility to achieve compliance with the permit 

limitation remains in effect. 

Some members of the regulated community have requested EPA 

and several State agencies to define more clearly enforcement 

discretion with respect to violations of whole effluent toxicity 

limitations. To define enforcement discretion would in effect 

make it no longer discretionary. Furthermore, the purpose of 

such guidance would be questionable as individual enforcement 

responses by EPA and the States are open to review by the public 

and the courts. In lieu of such additional guidance on enforcement discretion, it is recommended that Regions and States 

adhere to the principles presented in the EMS, the strategy, and in 

this document. 

EPA also has developed a policy [S] on the assessment of appropriate civil penalties in both administrative and civil judicial actions in response to any CWA violation. This policy bases the 

penalty amount on the seriousness of the violation, the economic 

benefit enjoyed as a result of delayed compliance, any history of 

such violations, any good-faith efforts to comply, and the violator’s 

ability to pay. In no instance can this calculated penalty exceed 

the statutory maximum penalties defined in CWA Section 309. 

If any violation occurs, the permittee has the responsibility of 

informing the regulatory agency. If the violation potentially 

endangers health or the environment, the violation must be 

verbally reported to the regulated agency within 24 hours and the 

permittee must submit a noncompliance report within 5 days of 

violation detection. If there is no danger to health or the environment, the written report must be submitted at the time 

monitoring reports are submitted. These reports must include a 

description of the violation, its cause, the period of noncompliance, 

and if the noncompliance has not been corrected, the anticipated 

time when compliance will be achieved. 

As with other NPDES permit limitation violations, violation of a 

water quality-based toxics limit should prompt immediate action 

on the part of the permittee. Permittee response should include 

evaluation of the cause of the violation, correction of operational 

deficiencies or improvement of treatment efficiency, and any 

other initial steps necessary to resolve the violation and mitigate 

the environmental effects. These immediate investigatory and 

corrective steps also should provide information that may be used 

in developing a compliance schedule if the violation is not resolved 

quickly. 

When a water quality-based toxicity limit is violated, the regulatory 

agency may require additional monitoring to determine the frequency and duration of the violation. If the permit limit is not met 

quickly through improved housekeeping, operation, or raw waste 

control (e.g., POW enforcement of pretreatment requirements, 

or chemical substitution by industries), requiring a TRE as discussed 

in Chapter 5 may be appropriate. Where toxicity-based limitations 

are in effect, the enforcement response must require expeditious 

compliance with the limit. 

Available enforcement mechanisms include Section 308 orders, 

Section 309 Administrative Orders, Administrative Penalty Orders 

with Administrative Orders, or judicial action. Enforcement action 

must be tailored to the specific violation and type of remedial 

action required. Enforcement actions must be worded carefully 

so that they clearly are understood, easily tracked, and expeditiously 

enforced. 

Violating limitations of pollutants at concentrations that pose a 

threat to human health should receive immediate enforcement 

attention to prompt rapid resolution of the noncompliance. The 

regulatory agency should consider the pollutant concentration, 

exposure route, and whether or not the pollutant exhibits a 

threshold response in determining if a schedule may be allowed. 

Immediate injunctive relief (such as a temporary restraining order 

or preliminary injunction) should be sought when necessary to 

protect public water supplies and fish and sheltfish areas from 

imminent or substantial impairment. 

125 
6.6 REPORTlWGOFVlCHAllOS 

The regulatory authority is responsible for reporting to the public 

on permittees in violation. Reporting requirements for the Quarterly Noncompliance Report (QNCR) of major permittees in violation of their NPDES permits are established in 40 CFR 123.45. 

Reporting of violations of water quality-based monitoring, limitations, schedules, and reporting requirements by major facilities 

must be consistent with 40 CfR 123.45. Violations of permit or 

enforcement order conditions by major permittees must be reported as follows [6]: 

l Effluent violations (chemical-specific and whole effluent toxicity) must be reported on the QNCR if the violation has the . 

potential to have caused a water quality problem (40 Cf!7 

123.45(a)(z)(iii)(A)(l)). 

l Chemical-specific toxic permit limit violations must be reported on the QNCR if two or more monthly average 

measurements in a 6-month period exceed the limit by a 

l 

factor of 1.2 for a Group I parameter or 1.4 for a Group II 

parameter as defined in the Regulation, or if four or more 

monthly average measurements in a 6-month period exceed the limit by any amount (40 CFR 123.45(a)(2)(ii)(C)). 

Any violation during the quarter of an interim monthly 

average chemical-specific toxic limit established in an administrative order or court order/consent decree must be 

reported on the QNCR (40 CFR 123.45(a)(2)(ii)(A)). (Note: 

Whole effluent toxicity is not characterized as a Group I or 

Group II parameter, and as such, must be evaluated on a 

professional judgement basis under 40 CFR 

123.4S(a)(2)(iii)(A)(l).) 

Compliance schedule milestones that are not met within 90 

days of the scheduled date must be reported on the QNCR 

(40 CFR 123.45(a)(Z)(ii)(B)). 

Failure to submit a report within 30 days of the due date 

must be reported on the QNCR (40 CFR 123.45(a)(Z)(ii)(D)). 

126 
CHAPTER 6 

REFERENCE6 

1. Jensen, L.]. Permit Compliance System (PCS) Policy Statement. 

2. NPDES Compliance Inspection Manual, May 1988. 

3. Enforcement Management System for the National Pollutant 

Discharge Elimination System, September 1986. 

4. Hanmer, R.W. 1989. Who/e Effluent Toxicity Basic Permitting 

Principles and Enforcement Strategy. 

5. Jensen, L.J. 1986. C/eon Water Act Penalty Policy for Civil 

Settlement Negotiations. 

6. Hanmer, R.W. 1986. Guidance for Prepurotion of Quarterly 

and Semi-Annual Noncompliance Reports. 

127 
7. CASE EXAMPLES 

7.1 INTRODUCTION 

This chapter presents examples of the development of water 

quality-based discharge limits to illustrate the integration of the 

guidance of the previous chapters. There are three examples: an 

industrial discharge with ample dilution, a publicly owned treatment works (POTW) with moderate dilution, and the combination of an industrial facility and a POTW discharge to the same 

reach. 

7.2 CASE:1 INDUSTRIAL DISCHARGE 

The first example is the Jaybird Corporation, a metal finishing 

firm. The NPDES permit for the facility is about to expire, and the 

corporation has submitted an application for a new permit. The 

example shows the steps that a permitting authority would take 

to determine if a water quality-based effluent limit is necessary 

and then to establish such a limit. The example also illustrates 

when best available technology (BAT) limits are applied instead of 

water quality-based limits, the use of human health criteria, and 

the variations in the limits derived by different wasteload allocation methods. 

• 

• 

• 

7.2.1 General Site Description and Information 

The Jaybird Corporation facility discharges into the Locapunct 

River. The river is approximately 60 miles long and its banks are 

occupied by small towns separated by woodland and farmland. 

The river is classified by the State in the water quality standards as 

having designated uses of a fish habitat, primary contact recreation, and a drinking water supply. For these uses, the State has 

adopted the federal water quality criteria into the water quality 

standards to protect aquatic life and human health. The State 

standards also includes a narrative criterion of “no toxics in toxic 

amounts” for other toxic materials. 

Water quality monitoring indicates some infrequent excursions 

above water quality criterion for copper and nickel. These pollutants have been found in measurable quantities in the effluents of 

several facilities. 

The Jaybird Corporation is a metal finishing facility that specializes 

in copper plating of lead shells for a nearby military installation, 

As a metal finisher, the Jaybird Corporation is relatively small with 

a discharge of 0.034 cfs (0.022 mgd). The effluent at the Jaybird 

Corporation is treated by precipitation and settles before discharge through a multiport diffuser. The corporation is subject to 

BAT and best practicable technology (BPT) effluent limits for the 

metal finishing industry. 

7.2.2 Effluent Characterization for Specific Chemicals 

The permitting authority has adopted a procedure in which pollutants concentrations in each facility are evaluated for the potential to cause, have the reasonable potential to cause, or contribute 

to an excursion of the water quality standards. The authority used 

the effluent characterization process for specific chemicals described in Chapter 3 in this evaluation. In general, the procedures 

are designed to determine which pollutants are of concern and 

which require effluent limits. 

Step 1: Identify Pollutants of Concern 

Data were obtained from a number of sources to identify and 

quantify the pollutants of concern in the Jaybird Corporation 

effluent: 

Effluent chemical concentrations were taken from the Permit Application Form 2C, Discharge Monitoring Reports 

(DMRs), EPA’s Permit Compliance System (PCS), and permit files. 

EPA’s STORET data base was used to obtain U.S. Geological 

Survey flow data and ambient monitoring data for the river. 

BAT limits for the metal finishing industry were obtained 

from 40 CFR 433 Subpart A. 

The permitting authority noticed in review of these data that the 

information in Form 2C replicated the information in the DMRs, 

and therefore decided to use the DMR data as the primary basis 

for characterizing the effluent. These data for toxicants DMRs are 

shown in Table 7-1. For those parameters currently not covered 

by the permit, Form 2C data indicated that pollutant concentrations were below detection limits. The permitting authority requested information from the facility showing the detection levels 

used; these levels were consistent with the detection levels listed 

in the National Pollutant Discharge Elimination System (NPDES) 

regulations at 40 CFR 136. 

The effluent from the Jaybird Corporation is regulated by the 

Metal Finishing Point Source Category effluent guidelines at 40 

CFR 433 Subpart A. These guidelines regulate the following toxic 

pollutants: cadmium, chromium, copper, cyanide, lead, nickel, 

silver, zinc, and total toxic organics. 

Although these parameters were regulated at the Jaybird Corporation, the only toxic pollutants evident in the discharge were 

lead, copper, and nickel. The facility’s treatment system reduced 

concentrations of other pollutants to below detection. 

Step 2: Determine the RAC, CMC, and CCC for Pollutants of 

Concern 

The State has adopted numeric water quality criteria for acute 

toxicity (criterion maximum concentration [CMC]), chronic toxicity (criterion continuous concentration [CCC)), and protection of 

human health (reference ambient concentration [RAC]). The 

water quality standards present the CMC and CCC criteria as 

equations based on ambient hardness concentrations. The standards require that the 85th percentile lowest hardness be used. 

This value is 100 mg/l as CaC03 for the Locapunct River. 

129 
Table 7-1. Effluent Data for the Jaybird Corporation 

n 

1 

2 

3 

4 

5 

6 

7 

8 

9 

10 

11 

12 

Copper Lead Nickel Toxicity 

μg/l μg/l μg/l TUC 

1,317 187 223 5 

1,092 230 261 10 

1,073 258 464 5 

1,059 423 341 20 

1,072 227 369 

1,677 275 1,058 

2,664 364 199 

1,058 170 259 

3,439 259 437 

6,596 264 773 

1,211 267 300 

1,082 175 356 

Mean 1,945 258 420 10 

SD 1,650 74 252 7.1 

CV 0.8 0.3 0.6 0.7 

Max 6,596 423 1,058 20 

Min 1,058 170 199 5 

N 12 12 12 4 

Source: DMR data for chemicals; 308 request for whole effluentoxicity. 

Notes: 

Metals reported as total recoverable metals; toxicity reported in chronic toxic 

units (100/NOEC). 

The permittee did not use a geometric dilution series for the toxicity tests. The 

results are the highestoxic units for any of the test organisms used. 

The aquatic toxicity criteria for metals in the standards are expressed as the acid soluble form of the metal. The State has 

adopted a ratio to express the acid soluble form of metals as the 

total recoverable form for the purposes of developing NPDES 

permit limits. This ratio is based on historical data that the State 

has collected for rivers in the basin where the Locapunct lies. The 

values of the ratio are 0.35 for lead, 0.70 for copper, and 0.85 for 

nickel. The standards consider the criteria for human health 

protection to be in the total recoverable form of the metal. 

Based on the hardness and acid soluble-to-total recoverable ratios, the applicable state water quality criteria are the following: 

Pollutant 

Lead 

Copper 

Nickel 

CCC CMC RAC 

μg/l μg/l μg/l 

9.1 235 50 

17.1 25.7 NA 

188 1,647 13.4 

Step 3: Determine Dilution for Aquatic life and Human Health 

Impacts 

The State water quality standards require that compliance with 

water quality criteria be achieved at the edge of the mixing zone. 

The standards specify the minimum dilution at which the criteria 

apply. These are the 7Q10 flow for the CCC, the 1Q10 flow for 

the CMC, and the harmonic mean flow for human health criteria 

(RAC). The U.S. Geological Survey operates a gaging station on 

the river; the flow statistics were calculated using the data from 

this station: 

• Harmonic mean flow = 38.0 cfs 

• 7Q10flow = 13.0 cfs 

• 1Q10 flow = 10.1 cfs. 

The facility provided a study of the outfall that showed that the 

multiport diffuser quickly achieved complete mixing across the 

width of the river. Dilution at the edge of the mixing zone could 

therefore be characterized by the complete mixing equation: 

C = (CemQe + CsQs/(Qe + Qs) 

where 

C = the receiving water concentration 

Cem = the maximum effluent concentration 

Qe = theeffluent flow 

Cs = the receiving water background concentration 

Qs = the appropriate receiving water flow. 

Step 4: Determine Reasonable Potential for Excursions 

To determine if the facility discharge was expected to cause or 

have the reasonable potential to cause the CMC, CCC, or RAC to 

be exceeded in the receiving water, the maximum receiving 

water concentration of each pollutant was first compared to the 

appropriate receiving water criterion. If the criteria were exceeded, then this was considered evidence that a water qualitybased limitation must be developed. 

Maximum expected concentrations were calculated using the 

average effluent flow, maximum effluent concentrations, background receiving water concentrations, and the relevant receiving water flow: the 1Q10 for the CMC, the 7Q10 for the CCC, or 

the harmonic mean for the RAC. The background receiving water 

concentrations for total recoverable metals were obtained from 

STORET data: 

Lead 1.6 μg/l 

Copper 4.8 μg/l 

Nickel 13.2 μg/l 

The maximum effluent concentration was estimated using the 

statistical approach in Chapter 3. There were 12 concentrations of 

each metal reported in the DMRs. For lead, these concentrations 

had a maximum value of 423 μg/l, an arithmetic mean of 258 μg/ 

I, an arithmetic standard deviation of 74, and an arithmetic 

coefficient of variation of 74/258, or 0.3. This coefficient of 

variation and the number of observations determined which multiplier was selected from Table 3-1. In this case, the multiplier 

value for 12 observations and a CV of 0.3 was interpolated from 

the values for 12 observations and CVs of 0.2 and 0.4. The 99th 

percentile multiplier was estimated to be 1.7. Similar calculations 

were conducted for copper (multiplier of 2.8) and nickel (multiplier of 3.7). 

130 
The receiving water concentration for lead for comparison with 

the CCC was calculated using data from Table 7-1: 

C = I(1.7 x 423 us/l x 0.034 cfs) + (1.6 us/l x 13 cfsjj 

(0.034 cfs + 13 cfs) 

= 3.5 pg/l 

where 

13cfs = the receiving water flow at 7410 

0.034 cfs = the mean effluent flow 

423 ug/l = the maximum effluent concentration 

1.7 = the statistical effluent multiplier to estimate the 

99th percentile concentration 

1.6 ug/l = the background receiving water concentration. 

The value of the calculated receiving water concentration, 

3.5 ug/l, was less than the chronic water quality standard of 9.1 

Pg/l for lead, and therefore there is no reasonable potential for the 

CCC to be exceeded. 

Using the effluent data presented in Table 7-1, the receiving water 

concentration is compared to the CMC as: 

C = I(1.7 x 423 us/l x 0.034 cfs) + (1.6 us/l x 10.1 cfs)] 

(0.034 cfs + 10.1 cfs) 

= 4.0 fig/l 

where 10.1 is the receiving water 1 QlO flow and the other values 

are identical to those for the CCC comparison. The resulting 

concentration of 4.0 ug/l was less than the acute standard of 234 

pg/l for lead. There is no reasonable potential for the CMC to be 

exceeded. 

For human health criterion evaluation, the receiving water concentration for compared to the RAC was calculated as: 

C = r(l.7 x 423 us/l x 0.034 cfs) + (1.6 us/l x 38 cfsIJ 

(0.034 cfs + 38 cfs) 

= 2.2 ug/l 

where 38 cfs is the harmonic mean flow and other values are the 

same as above. This value was less than the human heath criteria 

value of SO Pg/l for lead, so there is no reasonable potential for the 

RAC to be exceeded. 

Similar calculations were done for copper and nickel: 

Copper 

ccc 

CMC 

Nickel 

ccc 

CMC 

RAC 

Criterion Receiving Water 

cPg/l) Concentration @g/l) 

17.1 22.0 

25.7 26.9 

188 15.9 

1,647 16.6 

13.4 14.1 

The effluent characterization showed the reasonable potential for 

excursions above the CCC for copper and above the RAC for 

nickel. Therefore, permit limits are necessary for these two pollutants. 

7.2.3 E#oont Chwacto?lzotlon far mlo ERhnwt rO%iCifJ’ 

Whole effluent toxicity also was evaluated since there was a 

potential for excursions above the narrative water quality criterion 

due to the combination of effluent toxicants with other toxicants 

in the receiving water and in the effluent but below the detection 

level. The procedures used below follow those presented schematically in Figure 3-2, Chapter 3. 

Step 1: Dilution Determination 

The initial dilution determination was used to establish the types 

of toxicity tests that are conducted to characterize the effluent. 

The dilution at the low-flow characteristics for the facility is the 

following: 

At the 7Q10, dilution = (0.034 cfs + 13 cfs)/0.034 cfs 

= 383 

At the 1 Ql 0, dilution = (0.034 cfs + 10.1 cfs)/O.O34 cfs 

= 298. 

Step 2: Conduct Toxicity Testing 

EPA recommends that a discharger having a dilution between 

100 and 1,000 be required to conduct either chronic or acute 

toxicity testing. The permitting authority decided to require 

chronic testing but required the permittee to report the test 

results at the 48-hour endpoint so that acute toxicity could be 

measured. One year before the permit was due to expire, the 

permitting authority requested, under the authority of the Clean 

Water Act (CWA) Section 308, that the permittee test his effluent 

for toxicity to provide effluent information in order to write the 

next NPDES permit. In this case, the permitting authority specified that the discharger submit quarterly chronic toxicity data for 

1 year using the EPA toxicity tests for Selenostrvm, Cerioduphnia, 

and Pimephales. The permitting authority also specified that upstream ambient water be used as the diluent in the tests so as to 

allow the tests to measure additive effects from ambient toxics. In 

response to the Section 308 request, the discharger submitted 

the whole effluent toxicity data shown in Table 7-l. 

Step 3: Determine Reasonable Potential for Excursions 

The State interprets its narrative criteria for whole effluent toxicity 

to require that the technical support document recommendations of 0.3 TU, and 1 .O TU, be used as numeric values for acute 

and chronic toxicity, respectively. In accordance with the State 

standards, the CMC applies under the 1 Ql 0 flow and the CCC 

applies under the 7QlO flow. 

The determination of exceedance of the CMC or the CCC was 

simplified by the way in which the tests were conducted. Since 

the upstream ambient water was used as a diluent, the test results 

131 
already include an assessment of contributions from background 

toxicity. Therefore, the upstream receiving water concentration 

was set to zero. 

The maximum effluent concentration was again estimated by 

using the statistical approach in Chapter 3. As shown in Table 7- 

1, there were four observations of whole effluent toxicity. Based 

on the guidance of Box 3-4, these are insufficient to determine 

the CV accurately; therefore, the default CV of 0.6 was used. The 

effluent multiplier of 4.7 was obtained from Table 3-l using the 

number of observations, the CV, and the 99percent probability 

basis. 

The receiving water concentration for chronic toxicity for comparison with the CCC was calculated using data from Table 7-l : 

C = (4.7 x 20 TU, x 0.034 cfs) + (0 TU, x 13 cfs)] 

(0.034 cfs + 13 cfs) 

= 0.25 TU, 

where 

13cfs = the receiving water flow at 7410 

0.034 cfs = the mean effluent flow 

4.7 = the statistical effluent multiplier 

20 TU, = the maximum effluent concentration. 

The value of the calculated receiving water concentration, 0.25 

TU,, was less than the chronic water quality standard of 1 .O TU,, 

and therefore there is no reasonable potential for the CCC to be 

exceeded. 

To calculate the receiving water concentration for acute toxicity, 

the permitting authority first converted the chronic toxicity data 

into equivalent acute toxicity units by applying the acute-tochronic ratio (ACR) of 5 obtained from the monitoring data. The 

receiving water concentration for acute toxicity was then calculated: 

C = [(4.7 x 20 TU, / 5 ACR x 0.034 cfs) + (0 TU, x 10.1 cfs)] 

(0.034 cfs + 10.1 cfs) 

= 0.06 TU, 

where 10.1 cfs is the receiving water flow at 1 Ql 0, 5 is the acute 

to chronic ratio, and the other values are the same as above. The 

calculated value of 0.06 TU, is below the criterion of 0.3 TU,; 

therefore, there is no reasonable potential for the CMC to be 

exceeded. Since there was no reasonable potential for exceedances 

above either the acute or chronic criterion, permit limits were not 

developed for whole effluent toxicity. 

7.2.4 &!temhe wisstaa Aw8tl~ 

The wasteload allocation @VIA) was used to determine the level 

of effluent concentration that would comply with water quality 

standards in the receiving waters. A WlA will only be determined 

for those parameters that have a reasonable potential to cause 

exceedances of water quality standards. Therefore, WLAs were 

determined for copper and nickel. Since there was no reasonable 

potential for excursions above the CMC or CCC for nickel, only 

the WLA for human health was calculated. 

To determine WLAs, the numeric criteria in the water quality 

standards and background concentrations were used to calculate effluent concentrations that would result in compliance 

with those standards. The calculation of WlAs used receiving 

water flows that were appropriate to each standard: chronic 

WLAs were calculated using the 7410 flow, acute WlAs were 

calculated using the 1 QlO flow, and human health WLAs were 

calculated using the harmonic mean flow. Since the effluent 

was mixed rapidly by the multiport diffuser, the complete mix 

equation was used: 

MA = WQC x (Q + QJ - Q&I/Q, 

where 

Qe = the effluent flow 

QS = the receiving water flow 

cs = the background receiving water concentration 

WQC = the water quality criterion. 

The chronic and acute WLA for copper were calculated at the 

7410 and 1 Ql 0 flows, respectively: 

WLA, = [17.1 ug/l x (0.034 cfs + 13 cfs) - 13 cfs x 

4.8 pg/l] / 0.034 cfs 

= 4,720 pg/l 

WLAa = [25.7 ug/l x (0.034 cfs + 10.1 cfs) - 10.1 cfs x 

4.8 pg/l] IO.034 cfs 

= 6,234 pg/l. 

The human health WIA for nickel was calculated at the harmonic 

mean flow: 

w,m/,h = [13.4 pg/l x (0.034 cfs + 38 cfs) - 38 cfs x 

13.2 ug/l / 0.034 cfs 

= 237 ug/l. 

7.2.5 Dsrelbp PennIl Udfs 

Permit limits were developed using a steady-state, two-value WL4 

model as described in Box 5-2, Chapter 5. Values for constants 

were obtained from Table 5-2, Chapter 5. 

Step 1: Calculate LTA (note: this is Step 2 in Box 5-2) 

The chronic long-term average (LTA) for copper was calculated 

using the following formula: 

LTh = WLA x exp [OS o 2 - z o] 

= 4,720 ug/l x 0.440 

= 2,077 pg/l 

where values of exp [ 0.5 o2 - z ou ] are presented in Table S-1 

(see Chapter 5). The CV of 0.8 was used, and following the 

132 
guidance of Section 5.5.4, the z value for the 99th occurrence With a CV of 0.6, four samples per month for sampling, and a 

probability was used. 99th percentile used for the MDL, the factor is 1.64: 

The acute LTA for copper was calculated, again using the 99th 

percentile occurrence probability values from Table 5-1 as the 

multiplier: 

MDL = AMLx1.64 

= 237 Pg/l x 1.64 

= 389 Pg/l. 

LTAa = 6,234 Pg/l x 0.249 

= 1,552 Pg/l. 

lines 

The LTA for nickel human health permitting is considered to be 

the same as the WLA because the 70-year averaging period is 

used for human health evaluations (see Section 5.4.4). The LTA is 

calculated as: 

LTAf, = WL$, 

= 237 Pgg/l. 

The NPDES regulations require that effluent limits require treatment characteristic of the appropriate treatment technology and 

also achieve water quality standards. If water quality-based limits 

are more stringent than BAT limits, then the water quality-based 

limits become the basis for the effluent limits. Conversely, if the 

treatment technology (BAT) limits are more stringent, then they 

become the basis of the limits. 

Step 2: Determine the More Limiting LTA 

The limiting LTA for each pollutant was the minimum of the 

chronic, acute and human health LTA. The limiting LTA value was 

used in the next step to calculate maximum daily limits and 

average monthly limits. The limiting LTA for copper was found to 

be the acute LTA (1,552 kg/l) and the limiting LTA for nickel was 

found to be the human health LTA (237 ug/l). 

The comparison between the water quality-based and technology-based effluent limits are shown below. The more stringent 

limits are different for different pollutants: for nickel, water quality-based limits are more stringent whereas for copper, the BAT 

limits are the more stringent. 

Coowr Nickel 

Step 3: Calculate Maximum Daily and Average Monthly Limits 

The maximum daily limit (MDL) for copper was calculated using 

the expression: 

MDL = LTA x exp [z o - 0.5 a21 

= 1,552 ug/l x 4.01 

= 6,224 ug/l 

Water quality MDL 6,224 389 

AML 2,716 237 

BAT MDL 3,380 3,980 

AML 2,070 2,380 

Limit to use MDL 3,380 389 

AMC 2,070 237 

where the appropriate value for exp [ z o - 0.5 02] was taken from 

Tabfe S-2 using the row with the CV for copper (0.8) and the 

column for the 99th percentile probability basis. 

In accordance with NPDES regulations, the effluent limits were 

expressed in the permit as mass (pounds per day) by multiplying 

the concentrations above by the effluent flow of 0.034 cfs and the 

conversion factor of 5.394: 

The average monthly limit (AML) for copper was calculated using 

the expression: 

AML = LTA x exp [ z on - 0.50,-,~] 

= 1,552 fig/l x 1.75 

= 2,716 ug/l 

MDL 

AML 

Copper 

(lb/d) 

0.62 

0.38 

Nickel 

(lb/d) 

0.071 

0.043 

where the value for exp [ z CT,, - 0.5 o,~] was taken from Table 5- 

2 and, for this case, the number of samples per month was four. 

following the recommendations in Section 5.5.4, the z value for 

the 95th percentile probability basis was used. 

The effluent limits for nickel were determined by using the recommendations in Section 5.4.4, Chapter 5. The AML was considered 

to be identical to the W&, whereas the MDL was calculated from 

the AML by using the appropriate multiplier factor in Table 5-3. 

7.2.7 Cm#ariingD~lindtDsveloprnenttet~ds 

Permit limits for copper also were developed using a Monte Carlo 

simulation in order to compare the results to the permit limits 

derived from the two-value, steady-state model. A Monte Carlo 

simulation was used to generate receiving water concentrations 

to determine the effluent LTA for each of the pollutants such that 

the water criteria are achieved at the required frequency in the 

water quality standards. 

133 
Monte Carlo simulation used the same completely mixed dilution 

equation as was used for the steady state calculation: 

C = (Q&e + QsWQ + 9s) 

where C is the receiving water concentration (in Pg/l); C, and C, 

are the effluent concentration and the background concentration 

of the receiving water, respectively (in Pg/l); and Qe and Qs and 

effluent and receiving water flows, respectively (in cfs). Effluent 

flows were held constant at the mean effluent flow, and river 

flows were read from a computer file containing 60 years of daily 

flow data provided by the U.S. Geological Survey. The effluent 

concentrations were characterized by a lognormally distributed 

random variate. The random variate had a coefficient of variation 

that matched the CV of the pollutant in the effluent. 

The Monte Carlo simulation was run using 22,276 iterations. 

Once 22,276 receiving water concentrations had been calculated, receiving water concentrations were sorted, highest first. 

The 20th value (corresponding to the maximum concentration 

expected for 1 day in 3 years) was compared with the appropriate 

criterion. The 1 -day in 3-year return frequency is recommended 

by EPA for criteria (see Chapter 2). lf this value was higher than 

the criterion, the effluent LTA was reduced, and a new set of 

22,276 numbers was generated. When the receiving water concentration of the 20th value was just under the water quality 

criterion (and the 19th value was just over the same value), then 

the LTA effluent concentration generating these results was sufficient to achieve the water quality criterion; this LTA was then used 

in permit limit determinations. 

For chronic criteria, 4-day average concentrations were generated 

by taking the 4-day running average of modeled daily concentrations. The recurrence concentration was calculated in the same 

way as the l-day calculations described in the previous paragraph. Calculations were not made for the human health criterion. 

The permit limits were calculated according to the procedures 

given in Box S-3. Each LTA was multiplied by the 99th percentile 

multiplier from Table 5-3 for the MDL, and by the 95th percentile 

multiplier from Table 5-3 for the AML. For the AML, the same 

number of samples were used for the steady state and Monte 

Carlo permit limits (n=4). Thus, the resulting permit limits are 

directly comparable. The results of the Monte Carlo simulation for 

copper compared to the steady state calculations in units of 

micrograms/liter are shown below: 

Monte Carlo 

Steady State 

Maximum Average 

Daily Monthly 

8,618 3,761 

6,224 2,716 

7.3 CASE2 POTWDISCIIASE 

The second example is 01’ 3 fictitious POlVV that discharges to the 

same reach as the Jaybird C’orporation. The NPDES permit for this 

facility also is up for reissuance. The example highlights the use of 

background receiving water concentrations, and demonstrates 

the differences between industrial and POTW permit limits. In 

developing permit limits for the POlW in this example, the 

potential impacts from the Jaybird Corporation discharge were 

considered in the use of background receiving water concentrations. The interrelationships between the two facilities are discussed explicitly in Section 7.4. 

7.3.1 6l?nmai slle Ddscripuon %mi IThe Locapunct River receives discharges from a POlW serving the 

city of Auburn, a small city of about 10,000 people. The POlW 

treats a mixture of household and industrial waste with an activated sludge process. The mean effluent flow from the POTW is 

1.23 cfs. The POTW has no pretreatment program, but the 

municipality generally is aware of the small industries that are 

indirect dischargers because of research conducted by a local 

university. Generally, the plant is well operated. 

7.3.2 Em& C~c~~oll for !qt?cmc clbidmk 

The permitting authority’s approach for determining which pollutants cause, have the reasonable potential to cause, or contribute to excursions above water quality standards applies to POTWs 

as well as industries. The authority used the procedures described 

for the jaybird Corporation in the evaluation of the Auburn POTW. 

Step 1: Identify Pollutants of Concern 

At the time of the last permit issuance, there was evidence of a 

number of toxic pollutants in the POTW’s effluent, including 

copper, chlorine, and ammonia. These pollutants had monitoring requirements in the previous permit. Because there were 

metals in the effluent and, due to the industries discharging into 

the POTW sewer system, the permitting authority requested the 

POTW to conduct a complete priority pollutant scan of the effluent. The data received following the Section 308 letter request 

indicated that the concentrations of all priority pollutants except 

copper were below detection limits. The POTW’s primary toxic 

pollutants of concern were copper, chlorine, and ammonia (see 

Table 7-2). 

Step 2: Determine RAC, CMC, or CCC for Pollutants of Concern 

As described in the example of the industrial discharge, the water 

quality standards include numeric criteria for copper. The State 

also has adopted a numeric criterion for ammonia that is a 

function of the river 85th percentile pl-l and temperature; these 

values are 8.25”C and 25’C, respectively. Finally, the State interprets its narrative criterion of “no toxics in toxic amounts” to 

require use of the federal water quality criteria in the absence of a 

numeric state criterion. As a result, the permitting authority uses 

the federal criteria for chlorine. The applicable water quality 

criteria for the river are as follows: 

Copper 

Chlorine 

Ammonia 

ccc 

W/l) 

17.1 

11 

540 

CMC 

@3/l) 

25.7 

19 

4,000 

134 
Table 7-2. Effluent Data for the Auburn POlW 

n 

1 268 185 11,009 

2 115 301 13,025 

3 228 881 12,201 

4 59 372 24,548 

5 53 245 9,700 

6 213 244 15,645 

7 68 123 21,358 

8 200 343 3,976 

9 262 153 22,307 

10 519 448 7,427 

11 53 1,022 11,834 

12 474 347 8,430 

13 115 130 4,382 

14 2.59 128 9,330 

15 404 271 6,137 

16 57 451 6,448 

17 101 701 37,772 

18 187 582 14,307 

19 103 178 16,848 

20 76 436 28,205 

21 198 347 12,119 

22 265 475 11,778 

23 60 153 3,109 

24 112 268 4,474 

Mean 185 366 13,182 

SD 133 235 8,491 

cv 0.7 0.6 0.6 

Max 519 1,022 37,772 

Min 52.6 123 3,109 

Copper 

INI 

Chlorine Ammonia Toxicity 

l-u‘ 

1.5 

0.6 

0.4 

2 

1 

Source: DMR data for chemicals; 308 request for whole effluent toxicity. 

Notes: 

Metals as total recoverable; toxicity in toxic units (1001NOEC). 

The results are the highest toxic units for any of the test organisms used. 

Step 3: Determine Dilution for Aquatic Life and Human Health 

Impacts 

The State water quality standards requires that compliance with 

water quality criteria be achieved at the edge of the mixing zone. 

The standards specify the minimum dilution at which the criteria 

apply. These are the 7410 flow for the CCC, the 1 Ql 0 flow for 

the CMC, and the harmonic mean flow for human health criteria 

(RAC). The U.S. Geological Survey operates a gaging station on 

the river. The Row statistics were calculated using the data from 

this station: 

l Harmonic mean flow = 38.0 cfs 

l 7410 flow = 13.0 cfs 

l 1Q10flow=10.1 cfs. 

The POTW is located at a bend of the river where mixing is rapid. 

Therefore, the permitting authority used the complete mixing 

equation to calculate the receiving water concentrations. This is 

the same equation used for the industrial example. 

Step 4: Determine Reasonable Potential for Excursions 

The determination of possible exceedances in the CMC or CCC 

was based on a calculation of the maximum receiving water concentration of each pollutant, followed by a comparison to the 

appropriate receiving water criterion. The calculation of the maximum receiving water concentrations were made using the statistical estimate of the 99th percentile concentration of each pollutant 

in the effluent, the same flow used in the industrial example, and 

considered background receiving water concentrations of: 

Copper 4.8 pg/l 

Chlorine 0 w/l 

Ammonia 120 fig/l. 

The maximum effluent concentration was estimated using the 

statistical approach in Chapter 3. There were 24 concentrations 

of each chemical reported in the DMRs. For copper, these 

concentrations had a maximum value of 519 Pg/l, an arithmetic 

mean of 185 pg/l, an arithmetic standard deviation of 133, and 

an arithmetic coefficient of variation of 133/l 85, or 0.7. The 

multiplier was calculated to be 2.4 based on the CV of 0.7, 24 

observations, and a 99percent confidence level (see Section 3.3.2). 

Similar calculations were conducted for chlorine (multiplier of 

2.2) and ammonia (multiplier of 2.2). 

The receiving water concentrations for each pollutant were calculated. An example calculation for the comparison of copper to 

the CCC is shown below: 

C = J(2.4 x 519 us/l x 1.23 cfs) + (4.8 us/l x 13 cfs)l 

(1.23 cfs + 13 cfs) 

= 112pg/l 

where 

519 pg/l = the maximum measured effluent concentration 

2.4 = the statistical multiplier 

1.23 cfs = the average effluent flow 

4.8 Pg/l = the upstream receiving water concentration 

13cfs = the 7410 flow. 

The maximum receiving water concentrations for comparison to 

applicable standards for all pollutants were calculated to be: 

Copper 

ccc 

CMC 

Chlorine 

ccc 

CMC 

Ammonia 

ccc 

CMC 

Criterion 

w-II0 

17.1 

25.7 

11 

19 

540 

4,000 

Receiving Water 

Concentration 

@g/l) 

112 

140 

194 

244 

7,292 

9,128 

135 
--- 

The effluent characterization showed the reasonable potential for 

excursions above the CCC and CMC for copper, chlorine, and 

ammonia. Therefore, permit limits were developed for these 

pollutants. 

7.3.3 Eiflumt B fw ml@ Ef&rsnt foxicit~’ 

Step 1: Dilution Determination 

The initial dilution determination was used to establish the types 

of toxicity tests that must be conducted to characterize the effluent. The dilution at the low flow characteristics for the facility is 

the following: 

At the 7410, dilution = (1.23 cfs + 13 cfs)/l.23 cfs 

= 11.6 

At the lQ10, dilution = (1.23 cfs + 10.1 cfs)/l.23 cfs 

= 9.2. 

Step 2: Conduct Toxicity Testing 

EPA recommends that a discharger having a dilution less than 100 

be required to conduct chronic testing. The permitting authority 

requested through a Section 308 letter that the POTW provide 

quarterly chronic toxicity data for the year prior to permit 

reissuance. Tests using Selenastrum, Ceriodaphnio, and Pimephaies 

were required. The permitting authority also required the permittee to report the test results at the 48-hour endpoint so that acute 

toxicity also could be measured. Table 7-2 summarizes the results 

of the whole effluent toxicity testing. 

Step 3: Determine Reasonable Potential for Excursions 

As explained in the industrial example, the State interprets its 

narrative criteria for whole effluent toxicity to require that the 

technical support document recommendations of 0.3 TU, and 1 .O 

Tu, be used as numeric values for acute and chronic toxicity, 

respectively. In accordance with the State standards, the CMC 

applies under the lQ10 flow and the CCC applies under the 

7410 flow. 

The reasonable potential determination of exceedance of the 

CMC or the CCC was conducted in the same way as described in 

the industrial example. Upstream ambient water was used as a 

diluent to assess contributions directly from background toxicity; 

therefore, the upstream receiving water concentration was set to 

zero. The maximum effluent concentration was again estimated 

by using the statistical approach in Chapter 3. For the same 

reasons as were expressed in the industrial example, a multiplier 

of 4.7 was used. 

The receiving water concentration for chronic toxicity for comparison with the CCC was calculated using data from Table 7-2: 

C = (4.7 x 2 TU, x 1.23 cfs) + (0 TU, x 13 cfs) 

(1.23&+ 13cfs) 

= 0.8TUc 

where 

13 cfs = the receiving water flow at 7410 

1.23 cfs = the mean effluent flow 

4.7 = the statistical effluent multiplier 

4 TU, = the maximum effluent concentration. 

The value of the calculated receiving water concentration, 0.8 

TU,, is less than the chronic water quality standard of 1 .O TU,, and 

therefore there is no reasonable potential for the CCC to be 

exceeded. 

To calculate the receiving water concentration for acute toxicity, 

the permitting authority first converted the chronic toxicity data 

into equivalent acute toxicity units by applying the ACR of 2 

obtained from the monitoring data. The receiving water concentration for acute toxicity was then calculated: 

C= [(4.7x2TUc/2ACRx1.23cfs)+(OTU,x10.1 cfs)] 

(1.23 cfs + 10.1 cfs) 

= 0.5 TU, 

where 10.1 cfs is the receiving water flow at 1 Ql 0, 2 is the acute 

to chronic ratio, and the other values are the same as above. The 

calculated value of 0.5 TU, is greater than the criterion of 0.3 TU,, 

Therefore, there is reasonable potential for the CMC to be exceeded and permit limits were developed for whole effluent 

toxicity. 

7.3.4 lhthmnim wasteload All&iWLAs for chemicals and whole effluent toxicity were determined 

using information on the available dilution at the edge of the 

mixing zone. The calculation of WIA using the steady-state model 

was described in Section 7.2.4. The WLAs for the POTW using the 

equation discussed in Section 7.2.4 are: 

wLAa 

WLA, 

Toxicity Copper Chlorine Ammonia 

t-w WI) WD (WI/l) 

2.8 197 175 35,860 

11.6 147 127 4,979 

7.3.5 ik7nT&pPlwnlltlhMs 

The permit limit development process described in Box 5-2, 

Chapter 5 was applied to all pollutants. This process is identical to 

that explained in Section 7.2.5 except that (1) the WLA for acute 

toxicity needs to be expressed in equivalent chronic toxic units by 

multiplying by the ACR of 2, and (2) daily sampling of chlorine is 

required in the permit. The calculated LTA and permit limits are: 

LTA, 

LW 

MDL 

AML 

Toxicity Copper Chlorine Ammonia 

TUC 019/l) &l/b &l/l) 

1.8 55.4 56.2 11,511 

6.1 70.7 66.9 2,625 

5.6 197 175 8,162 

2.8 91 87 4,067 

136 
7.3.6 llotmdning and Expssia6 Llre Gontrollin6 Effluent 

llndt 

The treatment technology for POTWs is secondary treatment and 

is characterized by effluent limits for biochemical oxygen demand, total suspended solids, and pti. There are no BAT limits for 

toxics for POlWs, so there was no need to compare these water 

quality-based limits with other limits to determine which were 

more stringent. 

The permitting authority decided to use acute toxicity tests rather 

than chronic tests to measure compliance with the toxicity effluent limits. The appropriate effluent limits in terms of TU, were 

calculated by dividing the above calculation for TU, by the ACR of 

2 that was obtained from effluent monitoring. 

In accordance with NPDES regulations, the effluent limits for 

chemicals were expressed in the permit as mass (pounds per day) 

by multiplying the concentrations above by the effluent flow of 

1.23 cfs and the conversion factor of 5.394. Because there is no 

equivalent mass based unit for toxicity, toxicity mass limits are 

impractical under the regulation. 

MDL 

AML 

Toxicity Copper Chlorine Ammonia 

TUa (lb/d) (I b/d) (lb/d) 

2.8 1.31 1 .16 54.2 

1.4 0.64 0.58 27.0 

7.3.7 ~/ng DhYwent limit &?vetopnn?nt Moth& 

Permit limits also were developed using a Monte Carlo simulation 

to compare the results to the steady-state permit limits. A Monte 

Carlo simulation was used to generate receiving water concentrations for determining the appropriate LTA for each of the pollutants. The methodology for the Monte Carlo simulation is presented 

in Section 7.2.7. The results for this case are presented below. 

MOLs in TU, and ug/l 

Toxicity Copper Chlorine Ammonia 

Monte Carlo 3.9 264 249 9,657 

Steady State 2.8 197 175 8,162 

AMLs in TU, and pg/l 

Toxicity Copper Chlorine Ammonia 

Monte Carlo 2.7 171 170 6,614 

Steady State 1.4 91 87 4,067 

7.4 CASE 3: MDLTlPlE DlSCHARDES INTO THE SAME 

REACH 

Permit development for water quality-based toxics control has 

been illustrated for two single dischargers. This process increases 

in complexity in cases of multiple dischargers into a reach. The 

development of permit limits for multiple dischargers is based on 

the degradation in water quality resulting from the combined 

discharges, the development of total maximum daily loads (TMDls) 

for the river reach before generating WI& and the allocation of 

discharges to each discharger. The following example describes 

the permit development process when two dischargers release 

effluent into the same reach of a river. The dischargers are the 

Jaybird manufacturing plant described in Case 1 and the Auburn 

POTW described in Case 2. These facilities discharge into the 

Locapunct River, whose flow characteristics previously were described. 

7.4.1 Effluent Chamcterbafion 

The major differences in the effluent characterization for one 

facility and for multiple facilities is to identify those pollutants that 

are common to more than one facility, and to determine whether 

the combined discharges cause or are likely to cause water quality 

standards excursions. 

Step 1: Identify Pollutants of Concern 

Based on the data in Form 2C, the DMRs from the laybird 

Corporation and the data in the DMRs and Section 308 request 

from the Auburn POTW, the permitting authority found two 

contaminants common to both discharges: copper and whole 

effluent toxicity. Lead and nickel were found to be a problem at 

the Jaybird Corporation, but since there were no complicating 

discharges from the POTW, it was dealt with as a pollutant only at 

the metal finishing facility. Similarly, chlorine and ammonia were 

discharged solely by the POTW, so it was not necessary to provide 

effluent limits for the metal finishing facility for these chemicals. 

Step 2: Determine the CMC and CCC for Pollutants of Concern 

The numerical standards adopted by the State already have been 

presented. The relevant values for copper and whole effluent 

toxicity are: 

Copper 

Toxicity 

ccc CMC 

17.1 f.lg/l 25.7 ug/l 

1 .O TU, 0.3 Tu, 

Step 3: Determine Dilution for Aquatic Lie and Human Health 

Impacts 

Since this example is concerned with potential excursions above 

standards resulting from the collective discharge of two dischargers, the calculation of dilution includes the combined effluent flow 

from both facilities. The combined dilution can be characterized 

by the complete mixing equation: 

C = (&I Qel + C,zQez + CsQsMQet + Qe2 + Qs) 

where 

Qel and Qez = the flows of the two facilities 

C,l and Cc2 = the effluent concentrations of the two facilities 

CS = the upstream receiving water concentration 

Qs = the receiving water flow. 

137 
Step 4: Determine Reasonable Potential for Excursions 

To determine if the CMC or CCC were exceeded as a result of the 

combined discharges into the river, the receiving water concentration of each pollutant was calculated and compared to the 

appropriate criterion. The receiving water concentration calculation was based on the maximum value of the effluent concentralions (obtained from effluent data and multiplied by the appropriate statistical factor), average effluent flows, background receiving 

water concentrations, and appropriate river flows. All this information has been presented previously in the separate examples. 

The following results were obtained: 

Copper 

ccc 

CMC 

Toxicity 

ccc 

CMC 

Criterion 

01gA) 

17.1 

25.7 

1.0 

0.3 

Receiving Water 

Concentration 

@s/l) 

156 

194 

0.57 

0.45 

These calculations demonstrated exceedances of the copper CCC 

and CMC criteria and the toxicity CMC criterion. Permit limits 

were required. 

7.4.2 TIII#LsstnlHZ& 

WLAs were calculated to develop permit limits. WLAs for each 

discharger and chemical were based on calculated TMDLs, the 

total load to the Locapunct River that would not result in water 

quality standards exceedances. TMDLs are comprised of a load 

allocation for nonpoint sources, WlAs for point sources, and, if 

required by the State, a reserve capacity. TMDLs are further 

described in Section 4.2, Chapter 4. 

Step 1: Calculate TMDL 

The first step in developing individual WlAs for the two dischargers was to develop TMDLs for each pollutant of concern. TMDLs 

were developed in the same way as an individual WLA with the 

total load of a pollutant from the two dischargers being considered as a single discharge. 

The calculation of TMDLs used the following formula: 

TMDL=WQSx(Qt+Q,) 

where 

WQS = the water quality standard 

Qt = the combined flow of both effluents 

QS = the appropriate receiving water flow. 

The acute copper TMDL was calculated by using the data presented in the previous hyo examples as: 

TMDL = 25.7 lg/l x (0.034 cfs + 1.23 cfs + 10.1 cfs) 

= 292 pg-&/I 

where 

25.7 pg/l = theCMC 

0.034 cfs and 

1.23 cfs = the average effluent flows 

10.1 = the 1QlO. 

Similar calculations were made for chronic copper and acute 

toxicity. A TMDL was not calculated for chronic toxicity because 

the information presented in Chapter 1 indicates that chronic 

toxicity does not demonstrate additivity. The results are summarized below. 

Copper bg-cfs/l) 

Toxicity (TU,-&/I) 

Total Maximum Daily Loads 

Chronic Acute 

244 292 

NA 3.4 

Step 2: Develop WlAs 

The State had adopted an approach into the water quality management plan that described how WLAS were to be calculated. 

The approach required that existing upstream concentrations be 

used to determine the load allocation part of the TMDL and that 

10 percent of the TMDL had to be reserved and unavailable for 

allocation. The remainder of the TMDL could be apportioned to 

point sources in the WLA. 

The permitting authority decided to allocate the wasteloads based 

on the proportion of the existing load of each parameter that was 

attributed to each of the existing discharges. Based on the 

information shown in Tables 7-l and 7-2 and the average effluent 

flows, the pollutant loads from each facility are shown below. 

Parameter 

COP per 

@g-ds/l) 

Toxicity 

0%6s) 

Auburn POTW Jaybird Corporation 

Load Proportion Load Proportion 

227.6 0.77 66.1 0.23 

1.23 0.90 0.14 0.10 

Individual WLAs were then determined using the following 

equation: 

WfA = (TMDL - LA - 10% TMDL) x proportion/Q 

where the chronic TMDL was used to determine the chronic WLA, 

and the acute TMDL was used to determine the acute WLA for 

each facility. The WLAs for each pollutant and for each facility are 

presented on the following page. 

138 
Jaybird 

Copper 6.0) 

Toxicity UU,) 

Acute WLA 

134 1,450 

2.2 9.0 

Chronic WLA 

98.4 1,063 

NA NA 

7.4.3 PfmIdtumnlhvf~ 

Once the WLAs had been determined, permit limit development 

proceeded as in the previous examples. LTAs were calculated 

from the WLAS, and the limiting LTA was selected for calculating 

permit limits. For the metal finisher, where BAT limits were more 

restrictive than the water quality-based limits, the BAT limits 

applied. For the POTW, permit limits for toxic materials were 

required only to prevent exceedances of water quality standards. 

This process is summarized below. 

Step 1: Calculate LTAs 

The LTA was calculated for each discharger and pollutant as 

described in Step 2, Box 5-2, Chapter 5; the LTAs are shown 

below. 

Parameter 

Copper (c1gN 

Toxicity (TU,) 

Acute LTA 

POIW Jaybird 

37.7 361 

0.71 2.9 

Chronic LTA 

POW Jaybird 

47.3 468 

NA NA 

Step 2: Determine the More Lmiting LTA 

The minimum LTA was used to calculate MDLs and AMLs. The 

acute LTA was the lower LTA for both pollutants. 

Step 3: Calculate the Maximum Daily and Average Monthly 

Limits 

The MDL and AML were calculated as described in Box 5-2, 

Chapter 5: 

Average Monthly Limit Maximum Daily Limit 

Parameter POTW Jaybird POTW jaybird 

Copper @g/l) 62 632 134 1,448 

Toxicity (TU,) 1.1 4.5 2.2 9.0 

Step 4: Express the Limits 

The final step is to compare the water quality-based limits to the 

BAT limits to ensure that the more restrictive of the two are used, 

and to express the copper limits in terms of mass. The copper 

water quality-based limits for Jaybird Corporation are lower than 

the BAT ones (see Section 7.2.6). Therefore, the water qualitybased limits are required by the permit. In addition, the limits are 

lower than those calculated when only one of the facilities were 

considered. The final permit limits are listed below. 

Average Monthly Limit Maximum Daily Limit 

Parameter POlw Jaybird Polw Jaybird 

Copper (lb/d) 0.41 0.12 0.89 0.27 

Toxicity (TU,) 1 .l 4.5 2.2 9.0 

139 
Acute toxicity endpoints (ATEs) 

lethal concentration (LC) 4 

Acute toxicity testing 59 

Acute-chronic ratio (ACR) 17 

Additivity 24 

Allowable Effluent Concentration Distribution 82 

Ambient toxicity testing 61 

Ambient-induced mixing 77 

lateral dispersion coefficient 77 

shear velocity 77 

Amelia River 68 

Ames test 25 

Analytical considerations for chemicals 65 

Antidegradation policy 29 

Aquatic community 18 

Aquatic Lie Protection 34 

ARM 84 

BAF 38 

Bioaccumulation 37, 38, 72 

Bioaccumulation consideration 38 

Bioconcentration 38, 64 

Biocriteria 41,42 

Biological assessment/bioassessment 18, 20 

Biological criteria 

biological integrity 1, 18 

Biological surveylbiosurvey 18, 19 

Calculations 

CCC for toxicity 85 

CMC for toxicity 85 

Concentration (multiple dischargers) 86 

Concentration (nonconservative pollutant) 87 

Harmonic mean flow 88 

Lateral dispersion coefficient 77 

Carcinogenic&y 25 

Carcinogens 68 

Carcinogens, calculating l?ACs 40 

CCC 34,85 

CCC. See Criterion continuous concentration 79, 85 

Center for Exposure Assessment Modeling (CEAM) 73 

Chronic toxicity endpoints 

effective concentration 4 

lowest observed effect concentration 4 

no observed effect concentration 4 

Chronic toxicity testing 59 

Clean Water Act (CWA) 1, 67 

CMC 34, 71, 85 

CMC. 71 

Coefficient of variation (CV) 95 

in permit limit derivation 105 

Completely mixed discharge-receiving water situations 

Complex Effluent Toxicity Testing Program (CElTP) 7, 9 

Compliance monitoring 123 

discharge monitoring report/quality assurance 124 

inspections 124 

self-monitoring report 123 

Compliance problems 53 

Concentration 

flow distance 77 

Continuous Simulation 80 

CORMIXl 76 

Criterion continuous concentration (CCC) 48 

Criterion maximum concentration (CMC) 48, 71, 85 

Critical conditions 67 

Critical design periods 

estuaries and coastal bays 74 

oceans 74 

rivers and run-of-river reservoirs 73 

lakes, reservoirs 73 

Design periods. See Critical design periods 

Designated use 54 

Detection level 

minimum level 111 

practical quantitation limit 112 

Detection levels 111 

Determining the need for a limit 

statistical approach 56 

when is a chemical-specific limit sufficient? 62 

with effluent data 55 

without effluent data 50 

Dilution 52-53, 61 

Dilution determination 58, 63 

Discharge monitoring reports 54 

Discharge-Induced Mixing 75 

Duration 31-32 

Duration for single chemicals and WET 35 

Dye study 51, 58 

Dynamic modeling. See Modeling 

Modeling techniques; Models 76, 78, 79 

DYNHYD4 89 

DYNTOX 83,84 

Effluent bioconcentration evaluation 42, 64-65 

Effluent characterization 47 

addressing uncertainty in 56 

for aquatic life effects 48 

for human health effects 48, 62 

or multiple dischargers 60-61 

for specific chemicals 61-63 

for whole effluent toxicity 56-60 

process 53, 63 

78 

141 
Effluent characterization (cont.) 

purpose 48 

special considerations for marine and estuarine 

systems 61 

Effluent variability 

basic principles 93 

Enforcement 125 

enforcement discretion 125 

enforcement mechanisms 125 

whole effluent toxicity enforcement 124, 125 

EXAMS-II 84 

Excursion above CMC or CCC 60 

Excursions above ambient criteria 50 

Expressing limitations and developing monitoring require 

ments 110 

Expressing permit limits 

mass-based limit 110 

maximum daily 96 

FCM2 85 

FCETS 85 

Fish consumption values 37 

Fishable/swimmable 29 

FLOSTAT 89 

Food chain 87 

Frequency 31, 32 

Frequency for single chemicals and WET 36 

CC/MS 64 

Generating Effluent Data 56 

Cenotoxic pollutants 25 

Health effects 25 

nonthreshold effects 25 

threshold effects 25 

HPLC 65 

HSPF 84 

Human exposure 

Background concentrations 37 

drinking water ingestion 37 

fish consumption 37 

Human health criteria 

Ql l 39,40 

RfD 39, 40 

updating 38 

Human health protection 25 

Human health protection (WQC) 36 

Human health/Human exposure 37 

Implementation methods for state antidegradation policies 

Tier I 29 

Tier II 30 

Tier Ill (ONRWs) 30 

Independent application 22, 31, 49 

Integrated approach 1 

bioassessment approach 1, 22 

chemical-specific approach 1, 20 

whole effluent approach 1,4, 21 

Integrated Risk Information System (IRIS) 38 

Isopleths, concentration 72 

LA. See Load allocation 67 

Lateral dispersion coefficient 77 

LC50 4, 57, 58, 71 

Limits for metals 111 

Load allocation 67 

Lognormal Probabilistic Dilution Model 82 

Long term average (LTA) 95 

Magnitude 3 1, 32 

Magnitude for single chemicals 34 

Magnitude for whole-effluent toxicity 35 

Magnitude, duration, and frequency (Criteria) 31, 32 

Margin of safety 67 

Marine and estuarine discharges 61 

Marine and estuarine permitting 104 

Maximum daily permit limits 

chronic toxicity 96 

Metals 111 

MINTEQA2 84,85 

Mixing zone 58, 72 

Mixing Zones 33 

Modeling techniques 

continuous simulation 81 

lognormal probabilistic dilution model 82 

Monte Carlo simulation 81 

steady state 78 

Models 

CORMIXl 76 

DYNHYD4 84 

DYNTOX 84 

EXAMS-II 84 

FCM2 85 

FCETS 85 

FLOSTAT 89 

HSPF 84 

MINTEQA2 85 

Mixing zone 70 

PSY 78 

SARAH2 85 

Selection of 83 

STORET 79 

TOX14 84 

UDKHDEN 77 

ULINE 77 

UMERGE 77 

UOUTPLM 77 

UPLUME 76, 77 

WASP4 84 

142 
Monte Carlo simulation 80, 81 

Monticello Ecological Research Station (MERS) 2 

Multiple source toxicity testing procedures 60-61 

Multiple-source discharge 60-61 

Narrative criteria 29 

Noncarcinogens 89 

NPDES program 1 

Number of samples 

in permit limit derivation 105 

Numeric criteria 29 

Outfall design recommendations 

in lakes and reservoirs 73 

in oceans 74 

in rivers 73 

multiport submerged 73 

single-port submerged 73 

surface discharge 73 

Outstanding national resource waters (ONRWs) 30 

Penalties 124 

Permit documentation 110 

Permit limit derivation 98 

aquatic life protection 98 

average monthly 96 

average weekly 96 

basic objective 96 

detection levels 111 

direct application of both the acute and 

chronic WlAs 104 

dynamic 101 

human health protection 104-l 05 

maximum daily 96 

other approaches 103 

selection of monitoring frequencies 113 

single value steady state 102 

two value steady state 98 

use of a WLA as a permit limit 96-97 

Pollution prevention 111 

Potential for excursion above CMC or CCC 58 

POlw 53 

Precision 2, 11, 12 

coefficient of variation (CV) 12 

inter-laboratory precision 2, 11 

intra-laboratory precision 11 

variability 11 

Priority toxic pollutants 30 

Probability basis 

average monthly 110 

daily maximum 110 

in permit limit derivation 110 

LTAforWlA 110 

Quality assurance 12, 123 

discharge monitoring report/quality assurance 124 

Reasonable potential 48,49,50, 58 

as a trigger 58 

for multiple dischargers 60 

for whole effluent toxicity 58 

regulatory basis 49 

with effluent data 50, 56 

without effluent data 49 

Receiving water concentration (RWC) 48 

Reference ambient concentration (RAC) 48 

Reference toxicants 12 

Regulation requirements 4849 

Reporting violations 126 

to the public 126 

to the regulatory agency 124 

Return period 82 

Rhodamine WT. See Tracer (dye) studies 75 

SARAH2 85 

Screening protocol 53 

Sediment 42,67 

Sediment criteria 42 

Shayler Run, Ohio 2 

Single value wasteload allocations 

use in permit limit derivation 102 

Species sensitivity 59 

Species sensitivity differences 16 

Statistical considerations of effluent limits 105 

changes in CV on limits 106 

changes in CV on LTA 105 

changes in monthly samples 105, 107 

changes in probability basis on limits 105, 110 

Statistical considerations of WLAS 96 

Statistical distributions of effluent data 

lognormal (positively skewed) 95 

normal (bell shaped) 95 

STORET 79 

TMDL. See Total maximum daily load 

Total maximum daily load 

calculation of 78 

margin of safety 67 

Total maximum daily load (TMDL) 1, 67, 78 

TOXl4 84 

Toxic units (TVs) 6 

acute toxicity units (TU,) 7 

chronic toxicity units (TV,) 7 

Toxicity 

acute toxicity 4 

CCC for 85 

chronic toxicity 4 

CMC for 85 

whole effluent toxicity 4, 71 

143 
Toxicity persistence 23 

Toxicity reduction evaluations 

308 letters 118-l 19 

additional testing 117-l 18 

administrative orders 114 

approach 114 

circumstances warranting a TRE 117 

consent decrees 114 

evaluation criteria 115 

guidance documents 114 

requiring TREs 118 

toxicity identification evaluation (TIE) 116 

TRE plans 115 

Toxicity test endpoint concentrations 

acute toxicity endpoints (ATEs) 4 

chronic toxicity endpoints (CTEs) 4 

Toxicity testing 11 

composite sample 13 

flow-through toxicity test 13 

grab sample 13 

off-site tests 11 

on-site tests 11 

static toxicity test 1 3 

Toxicity testing procedures 58 

Treatment plant performance 97 

Triggers for permit limit development 58, 63-64 

Tualatin River Basin 68 

UDKHDEN 77 

ULINE 77 

UMERCE 77 

UOUTPLM 77 

UPLUME 76, 77 

Variability 

effluent variability 16 

exposure variability 16 

species sensitivity differences 16 

Violation review 124 

WASP4 84 

Wasteload 67 

Wasteload allocation 

schemes 67, 69 

Wasteload allocation (WL4) 1 

Water quality criteria 1, 29, 32 

aquatic life 34 

criterion continuous concentration (CCC) 32 

criterion maximum concentration (CMC) 34, 35 

RACs for non-carcinogens 39 

reference ambient concentration (RAC) 36 

reference tissue concentration (RTCs) 37 

Water quality models 

continuous simulation 81, 98 

dynamic 98 

lognormal probabilistic 82 

Monte Carlo 81, 98 

single value steady state 97 

two value steady state 98 

Water quality standards 1 

biological criteria 1 

narrative water quafity criterion 1 

numeric criteria 1 

water quality criteria 1 

Water quality standards regulatory considerations 30 

Section 303(c)(2)(8) of CWA 30 

40 CFR 122.44 & 40 CFR Part 131 31 

Section 307(a) of CWA 30 

Section 131 .l 1 Standards Regulation 2.1 29 

When is a chemical-specific limit sufficient? 62 

Whole effluent approach 4 

Whole effluent toxicity 

“pass/fail” tests 104 

acute endpoint sensitivity 112 

chronic endpoint sensitivity 112 

description of limits 112 

detection levels 112 

end-of-pipe approach 104 

test methods 112 

use of acute versus chronic tests 112 

Whole effluent toxicity data generation 58 

WfA 67 

* This is not a comprehensive index. Only topics of importance 

are highlighted. 

144 
APPENDICES 

A. Raw Data 

B. EPA Regulations and Policies 

C. Ambient Toxicity Testing and Data Analysis 

D. Duration and Frequency 

E. Lognormal Distribution and Permit Limit Derivations 

F. Sampling 

c. The Development of A Biological Indicator Approach to Water Quality-based Human Health Toxics Control 

H. Reference Dose (RFD): Description and Use in Health Risk Assessment 

I. Chemicals Available on IRIS 

145 
APPENDIX A-1 

TOXICITY TEST PRECISION DATA 
MARINE/ESTUARINE SHORT-TERM CHRONIC TOXICITY 

TESTS 
SHEEhiEAD MINNOW (Cyprinodon variegatus) 

Seven-day Larval Survival and Growth Test 

Single Laboratory Precision Data 

Table A-l-l. Single laboratory precision of test performed in 

40 fathoms artificial seawater, using larvae from fish 

maintained and spawned in 40 fathoms artificial seawater, 

using copper as the reference toxicant [l]. 

Test 

Number ; 

, Most 1 

Go 1 

1 0.05 0.1133 

2 I <0.05* I 0.0543 

3 <0.05’ 0.0418 

4 I 0.05 0.0632 

5 

i 

<“‘05’ 0.0577 

6 I 0.05 0.0483 

7 0.05 0.0796 

8 j 0.05 j 0.1235 

n: 5 

Mean: I 0.05 

CV(o/o): NA 

1 i.0727 

41.82 

(m$l) 

0.1523 

0.0975 

0.0714 

0.0908 

0.0998 

0.1325 

0.1597 

0.2364 

Sensitive 

Endpoint 

S 

c 

G 

S 

s 

c 

G 

G 

40.77 i I 

* The lowest concentration tested was 0.05 mg./l 

NOEC Range: ~0.05’ - 0.05 mg/l. 

Copper concentrations in Tests l-6 were 0.050, 0.10, 0.20, 0.40, and 0.80 mgil 

and Tests 7-8 were 0.025, 0.050, 0.10, 0.20, and 0.40 mgll. 

Prepared by Florence Kessler, TAI, Cincinnati, OH, fanuary 11, 1990 (ICp 

Program, versron 1 .l b). 

Table A-l-2. Single laboratory precision of test performed in 

40 fathoms artificial seawater, using larvae from fish 

maintained and spawned in 40 fathoms artificial seawater, 

using sodium dodecyl sulfate (SDS) as the reference toxicant 

PI. 

2 ,l.O 

3 1.0 ( 2.3051 1 2.8367 ’ S 

4 0.5 1.9855 2.6237 ; G 

5 1.0 1.1901 ; 1.4267 5 

6 0.5 

---- 

, 1 1.1041 1.4264 1 c 

fl: 6 ! 6 I 6 I 

Mean: , 0.8 1.5456 1.9595 ’ 

CVl%): NA 31.44 I 31.82 ! 

NOEC Range: 0.5 - 1 .O mg!l (this represents a difference of one exposure 

concentration). 

SDS concentrations in Tests l-2 were 1 .O, 1.9, 3.9, 7.7, and 15.5 mg/l and in 

Tests 3-6 were 0.20, 0.50, 1.0, 1.9, and 3.9 mgif. 

Prepared by Florence Kessler, TAI, Cincinnati, OH, January 11, 1990 (ICp 

Program, version 1 .l b). 

Table A-l-3. Single laboratory precision of test performed in 

natural seawater, using larvae from fish maintained and 

spawned in natural seawater, using copper as the reference 

toxicant [ 11. 

I Most 

Test ’ Sensitive 

Number 1 Endpoint 

1 ; 125 1 320.3 j 437.5 j S - 

4 125 

5 ’ 125 

n: T 5 

Mean: I 106.2 

182.3 

I 

323.0 ’ C 

333.4 , 483.4 I 5 

228.4 343.8 ’ S 

300.4 I 396.9 , 

1 CW%k 1 NA I 33.0 : 19.2 ’ I 

* No linear interpolation estimate could be calculated from the data, since none 

of the group response means were less than 50 percent of the control response 

mean. 

NOEC Range: 31 - 125 mg/l (this represents a difference of two exposure 

concentrations). 

Prepared by Elise Torello, SAIC, Narragansett, RI, and Margarete Heber, EPA, 

Washington, DC, February 1990 (ICp Program, version 1 .l b). 

Table A-l-4. Single laboratory precision of test performed in 

natural seawater, using larvae from fish maintained and 

spawned in natural seawater, using sodium dodecyl sulfate 

(SDS) as the reference toxicant [l]. 

Test ’ INOEC 

Number 1 (mg/l) 

I 

1 ’ 2.5 

2 1.3 

3 

I 

/ 1.3 

4 1.3 

5 , 1.3 

n: 5 

Mean: 1.5 

CW%l: ; NA 

2s 

; (:9/u 

2.9 

NC’ 

1.9 

I 2.4 1.5 

i 

I 4 2.2 

! 27.6 

j tSI, 

3.6 

NC2 

I 

2.4 

NC2 

-I 1.8 

----.- - 

3 

I 2.6 35.3 

’ Most 

, Sensitive 

I Endpoint 

f------ 

I s 

I G 

5 

I c 

5 

i---- --- 

NOEC Range: 1.3 - 2.5 mg/l (this represents a difference of one exposure 

concentration). 

‘No Linear interpolation estimate could be calculated from the data, smce none 

of the group response means were less than 75 percent of the control response 

-. 

2No linear interpolation estrmate could be calculated from the data, smce none 

of the group response means were less than 50 percent of the control response 

mean. 

Prepared by Elise Torelfo, SAIC, Narragansett RI, and Margarete Heber, EPA, and 

Washington, DC, February 1990 (ICp Program, version 1 .I b). 

A-l-l 
SHEEPSHEAD MINNOW (Cyprinodon variegatus) 

Seven-day Larval Survival and Growth Test 

Interlaboratory Precision Data 

Table A-1-5. Interlaboratory precision of test using an industrial effluent as 

the reference toxicant [ 11. 

Most Sensitive Endpoint 

Test 1 NOEC 

;i$ 

GO 

.~_____-~-~ .~ _ -. Number .- - -- A- -!%,--,----- -‘--P?-- 

Laboratory A I 

1 , 3.2 (S,G) / 7.4 (5) I 7.4 (G) 

2 1 3.2 (S,G) 7.6 (S) 14.3 (G) 

Laboratory B 

1 3.2 (S,G) 5.7 (C} 9.7 (C) 

2 3.2 (S,C) 5.7 (C) 8.8 (C) 

Laboratory C I 

1 1 .o (5) 4.7 (5) 7.2 (5) 

Laboratory II 

1 ’ 3.2 (S,G) 7.4 (G) 24.7 (G) 

2 ’ 1.0(G) -----.-F.-..e---1 ----- 

5.2 (5) 7.2 (5) 

tl: 7 7 7 

Mean: j 2.6 5.5 11.3 

CV(%): NA 44.2 56.9 

NOEC Range: 1 .O - 3.2 percent (this represents a difference of one exposure concentration). 

Prepared by Elise Torelto, SAIC, Narragansett, RI, and Margarete Heber, EPA, and Washington, DC, 

February (Itp Program, version 1 .l b). 

SHEEPSHEAD MINNOW (Cyprinudm variagatus) 

Embryo-larval Survival and Teratogenicity Test 

Single Laboratory Precision Data 

Table A-l-6. Single laboratory precision of test performed in HW Marinemix 

artificial seawater, using embryos from fish maintained and spawned in HW 

Marinemix artificial seawater, using copper as the reference toxicant [l]. 

Test 

Number 

I 

ECIO 60 7 NOEC 

bJg/u OJgm wi/J) 

1 173 189 198 234 240 

2 l * t * 240 

3 l l t t I 240 

4 182 197 206 ’ 240 , 240 

5 171 187 197 234 240 

6 . * l * / <200 

l l l c 220 

195 203 : 208 I 226 , 220 

n: 4 4 4 4 7 

Mean: 180 194 202 233 234 

CV(%): 6.1 3.8 2.8 2.5 NA 

l Data do not fit the Probit model. 

NOEC Range: 200 - 240 (this represents a difference of two exposure concentrations). 

A-l -2 
SHEEPSHEAD MINNOW (Cyprinodun variegatus) (continued) 

Embryo-larval Survival and Teratogenicity Test 

Single Laboratory Precision Data 

Table A-l-7. Single laboratory precision of test performed in HW Marinemix 

artificial seawater, using embryos from fish maintained and spawned in HW 

Marinemix artificial seawater, using sodium dodecyl sulfate (SDS) as the 

reference toxicant [l], 

Mean: 1.3 

11.7 ’ NA 

l Data do not fit the Probit model. 

NOEC Range: 2.0 - 4.0 ug/l (this represents a difference of one exposure concentration) 

INLAND SILVERSIDE (Menldia beryllfna 

Seven-day Larval Survival and Growth Test Single Laboratory Precision Data 

Table A-l-8. Single laboratory precision of the inland 

silverside (Menidio beryllina) larval survival and growth test 

performed in natural seawater, using larvae from fish 

maintained and spawned in natural seawater, using copper as 

the reference toxicant [I]. 

Test ; 

Number 

1 

2 ’ 

3 

4 

5 

NOEC 

(WU 

----- 

63 

125 

63 

125 

31 

n: 5 

Mean: 81.4 

CV(%j: ’ NA 

I 

/ (ItgL;1) 

1 96.2 

207.2 

j 2189 

! - 

177.5 

1 350.1 

I 43.7 

I 

I (i$i) A----- 

148.6 

NC 

493.4 

! 241.4 

+Y1 340.8 

I 50.7 

I Most 

1 Sensitive 

1 Endpoint 

I s 

I i 

! s t”- 

l No linear interpolation estimate could be calculated lrom the data, since none 

of the group response means were less than 50 percent of the control response 

mea”. 

NOEC Range: 31 - 125 ug/l (this represents a difference of two exposure 

concentrations). 

Prepared by Elise Torello. SAIC, Narragansett, RI, and Margarete Heber, EPA, and 

Washington, DC, February 1990 (ICp Program, version 1 .lb). 

Table A-l-9. Single laboratory precision of the inland 

silverside (Menidia beryllina) larval survival and growth test 

performed in natural seawater, using larvae from fish 

maintained and spawned in natural seawater, using sodium 

dodecyl sulfate (SDS) as the reference toxicant [l]. 

I I 

Test 1 NOEC ( ICL, i I& 

Most 

1 Sensitive 

Number ; (mg/l) ) @q/l) (mg/l) Endpoint 

NOEC Range: 1.3 mg/l 

Prepared by Eke Torello, SAIC, Narragansett, RI, and Margarete Heber, EPA, and 

Washington, DC, February 1990 (ICp Program, version 1.1 b). 

A-l-3 
Seven-day Survival, Growth, and Fedundlty Test Single 

Laboratory Precision Data 

Table A-1-10. Single laboratory precision of the mysid (Mysidopsis 

bahia) survival, growth and fecundity test performed in natural seawater, 

using juveniles from mysids cultured and maintained in natural 

seawater, using copper as the reference toxicant [I]. 

5 i125 I 157.7 j 200.3 1 S I 

n: 

Mean: 5 i 112.6 

CV(%): / NA 

l No linear interpolation estimate could be calculated from the data, since none of the 

group response means were less than 50 percent of the control response mean, 

NOEC Range: 63 125 ugll (this represents a difference of two exposure concentrations). 

Prepared by Eke Torello, SAIC, Narragansett, RI, and Margarete Heber, EPA, and 

Washington, DC, February 1990 (ICp Program, version 1 .lb). 

Table A-l-11. Single laboratory precision of the mysid (Mysidopsis 

bohia) survival, growth, and fecundity test performed in natural 

seawater, using juveniles from mysids cultured and maintained in 

natural seawater, using sodium dodecyl sulfate (SDS) as the reference 

toxicant [l). 

__--- 

‘No linear interpolation estimate could be calculated from the data, since none of the 

group response means were less than 75 percent of the control response mean. 

‘No linear interpolation estimate could be calculated from the data, since none of the 

group response means were less than SO percent of the control response mean. 

NOEC Range: ~0.3 _ 5.0 mg/l (this represents a difference of four exposure concentrations). 

Prepared by Elise Torello, SAIC, Narragansett, RI, and Margarete Heber, EPA, and 

Washington, DC, February 1990 (ICp Program, version 1 .l b). 

A-l-4 
SEA URCHIN (Arhcia punctrrlata) 

Fertilization Test Single Laboratory Precision Data 

Table A-1-12. Single laboratory precision of the sea urchin Table A-1-14. Single laboratory precision of the sea urchin 

(Arbucio punctuloto) fertilization test performed in natural (Arbacia punctulata) fertilization test performed in natural 

seawater, using gametes from sea urchins maintained and seawater, using gametes from sea urchins maintained and 

spawned in artificial seawater (40 Fathoms), using copper spawned in natural seawater, using copper as the reference 

as the reference toxicant [l]. toxicant [l]. 

Test ! NOEC 

CV(%): ( NA 

T 

c 

! 

1 

(!iil) 

8.92 

26.35 

11.30 

34.28 

36.67 

5 

23.51 

54.60 

UOEC Range: ~5.0 - 12.5 ug/l (this represents a difference of one exposure NOEC Range: ~6.1 _ 24.4 ug/l (this represents a difference of two exposure 

roncentration). concentrations). 

kpper concentrations in Test 1 were 2.5, 5.0, 10.0, 20.0, and 40.0 ug/l and in 

rests 2-S were 6.25, 12.5, 25.0, 50.0, and 100.0 ug/l. 

Prepared by Florence Kessler, TAI, Cincinnati, OH, lanuary 11, 1990 (ICp 

Program, version 1 .l b). 

Prepared by Elise Torello, SAIC, Narragansett, RI, and Margarete Heber, EPA, and 

Washington, DC, February 1990 (ICp Program, version 1 .l b). 

Table A-1-13. Single laboratory precision of the sea urchin Table A-l-15. Single laboratory precision of the sea urchin 

(Arbacio punctulota) fertilization test performed in natural (Arbocio punctulato) fertilization test performed in natural 

seawater, using gametes from sea urchins maintained and seawater, using gametes from sea urchins maintained and 

spawned in artificial seawater (40 Fathoms), using sodium spawned in natural seawater, using sodium dodecyl sulfate 

dodecyl sulfate (SDS) as the reference toxicant [l]. (SDS) as the reference toxicant [l]. 

Test 

Number 

--- 

1 

2 

3 

4 

5 

n: 4 5 5 

Mean: 1.4 1.73 2.38 

CV(%): NA 1 29.7 23.3 

NOEC Range: 1.2 - 3.3 mg/l (this represents a difference of one exposure 

concentration). 

SDS concentrations for all tests were 0.9, 1.8, 3.6, 7.2, and 14.4 mg/l. 

Prepared by Florence Kessler, TAI, Cincinnati, OH, January 11, 1990 (ICp 

Program, version 1 ,l b). 

NOEC Range: 0.9 - 1.8 n-g/l (this represents a difference of one exposure 

concentration). 

Prepared by Elise Torello, WC, Narragansett, RI, and Margarete Heber, EPA, and 

Washington, DC, February 1990 (ICp Program, version 1 .l b). 

Test 

Number 

- 

1 

2 

3 

4 

5 

n: 

Mean: 

cv(%): 

: 

I 

NOEC 

b%lm 

12.2 

12.2 

24.4 

~6.1 

6.1 

4 

13.7 

NA 

30.3 46.3 

26.2 34.1 

11.2 17.2 

-,- 

5 5 

22.8 29.9 

41.9 48.2 

A-l -5 
RED MACROALGAE(C~~~~~~ga~~~~ 

Reproduction Test Single Laboratory Precision Data 

Table A-l-16. Single laboratory precision of the red 

macroalga (Champio porvu/u) reproduction test performed in 

SO/SO natural seawater and GP-2 artificial seawater. Copper is 

the reference toxicant [l]. 

Test NOEC ] lCLs 

Number , WI/u WIiiu 

1 I 1.0 1.67 

c - 

NOEC Range: 0.5 - 1 .O ug/l (this represents a difference of one exposure 

concentration). 

Prepared by Elise Torello, SAIC, Narragansett, RI, and Margarete Heber, EPA, and 

Washington, DC, February 1990 (ICp Program, version 1.1 b). 

Table A-1-17. Single laboratory precision of the red 

macroalga (Chompio porvulu) reproduction test performed in 

50/50 natural seawater and GP-2 artificial seawater. Sodium 

dodecyl sulfate (SDS) is the reference toxicant [l] (personal 

communication with G. Thursby, SAIC, Narragansett, RI). 

Test 

Number 

1 

2 

3 

4 

5 

6 

7 

8 

9 

n: 

Mean: 

CV(%): 

NOEC 

+ 

co.48 0.4 

co.48 0.2 

0.26 

; 

0.2 

0.09 I 0.1 

0.16 0.2 

0.09 i 0.1 

<0.29 I 0.3 

5 9 

0.22 I 0.31 

NA I 69.0 

1 

0.3 

0.6 

0.2 

0.4 

0.5 

0.3 

0.3 

0.2 

0.4 

9 

0.36 

37.0 

NOEC Range: 0.09 - 0.48 mg/l (this represents a difference of two exposure 

concentrations). 

Prepared by Elise Torelfo, SAIC, Narragansett, RI, and Margarete Heber, EPA, and 

Washington, DC, February 1990 (IQ Program, version 1.1 b). 

Table A-1-18. Single laboratory precision testing of the red macroal a (Chompia porvulo) reproduction test in natural seawater (30 

OJOO salinity). The reference toxicants used were copper sulfate (Cu) P g2 and sodium dodecyl suffate (SDS)2t3 [7]. 

cu @g/l) SDS (mg/l) 

__---- 

Test r -NOEC 1 Gs I ‘CSO NOEC GS GO 

CV(%): 1 NA 1 61.09 ’ 34.45 1 NA 1 62.29 1 22.92 

‘Copper concentrations were OS, 1 .O, 2.5, 5.0, 7.5, and 10 ugll. Concentrations of Cu were made from a 100 ug/ml CuSO4 standard obtained from Inorganic Ventures, 

Inc., Brick, NJ. 

2All tests were conducted at 23 + l°C in natural seawater with irradiance set at 40 uE/m’,‘s. 

%DS concentrations were 0.0375, 0.075, 0.15, 0.30, 0.60, and 1.20 mg/l. Concentrations of SDS were made from a 44.64 + 3.33 mg/ml standard obtained from U.S. EPAEMSL, Crncinnati, OH. 

Prepared by Steven H. Ward and Glen Thursby, EPA, Narragansett, RI (ICp Program, version 1 .l b). 

A-l -6 
FRESH WATER SHORT-TERM CHRONIC TOXICITY TESTS 
FATHEAD MIWWDW (Pimephales promelas) 

Seven-day Larval Survival and Growth Test and 

Embryo-larval Survival and Teratogenicity Test 

Single Laboratory Precision Data 

Table A-1-19. Single laboratory precision of the fathead 

minnow (Pimepholes promelos) embryo-larval survival and 

teratogenicity test performed in using Diquat as the reference 

toxicant [2]. 

I Test I LCl I 

2 2.31 

3 1 so 

4 1.71 

5 1.43 ~ 

n: 5 

Mean: 1 Sl 

CV(%): 41.3 

Table A-1-20. Single laboratory precision of the fathead 

minnow (Pimephales promelos) embryo-larval survival and 

teratogenicity test performed in using cadmium chloride as the 

reference toxicant (21. 

Test 

Number 

1 --4 

2 

3 

4 I 

0.014 

0.006 

0.005 

0.003 

5 

-- -- ~~..OEL 

n: 5 

Mean: 0.0068 0.012 

CV(%): 62 NA 

NOEC Range: 0.011 _ 0.013 mg/l (this represents a difference of one exposure 

concentration). 

A-l-7 
FATHEAD MINNOW (Pimephaks prume/as) 

Seven-day Larval Survival and Growth Test 

Single Laboratory Precision Data 

Table A-l-21. Single laboratory precision of the fathead 

minnow (Pimephales promelos) larval survival and growth test 

performed in using NAPCP as the reference toxicant [2]. 

Test NOEC* 

Number (WI) 

1 256 

2 128 

3 256 

4 128 

5 128 

n: 5 

Mean: 

cv(%): 

I 179.2 

NA 

*Raw data unavailable, It25 and Itso values could not be calculated. 

NOEC Range: 128 _ 256 ug/l (this represents a difference of one exposure 

concentration). 

Table A-l-22. Results of the performance evaluation for contract laboratories conducted for the California Regional Water Quality 

Control Board. All tests were conducted using potassium chromate (expressed as Cr+6) and testing the fathead minnow (Pimephales 

promelos) in the 7-day subchronic tests [3]. 

Lab 

- 

l ! Tap’ ! 2X 

2 

3 

4 

5 

6 I 

7 ’ 

8 

9 MHRW ! 3X 

10 

- - _. - --..- .- 

Mean: 

cv (%): I 

’ Moderately hard tap water 

2 Control with three replicates and all concentrations with two replicates. 

3 Value is extrapolated and is not included in coefficient of variation calculation. 

4 Weight measurements made with questionable techniques. 

’ Dechlorinated Lake Ontario tap water. 

6 Not reported. 

’ Well water mixed with spring water, moderately hard. 

’ Value may be skewed as middle concentration had 45 percent survival but no weights reported. 

A-l-8 
Seven-day larval Survival and Growth Test 

Interlaboratory Precision Data 

Table A-l-23. lnterlaboratorv orecision data of the fathead minnow 1Pimeohale.s oromelas) 7-dav larval survival and orowth test. 

Combined frequency distribution for survival NOECs for all participatkg kboratdries [2]. ’ ’ 

a 

NOEC Frequency (%) Distribution 

.- 

Tests with 2 Reps Tests with 4 Reps 

Sample 

Sodium Pentachlorophenate ’ 

l--.--.-L 

Median 1 +1(a) 

35 53 

Sodium Pentachlorophenate 2 42 

Potassium Dichromate 1 47 

Potassium Dichromate 2 41 

Refinery Effluent 301 26 

Refinery Effluent 401 37 

Utility Waste 501 56 

42 

47 

41 

68 

53 

33 

’ Percent of values with one concentration intervals of the median. 

’ Percent of values within two or more concentrations intervals of the median. 

11 56 33 11 

Table A-l-24. Interlaboratory precision data of the fathead minnow (Pimephales promelos) 7-day larval survival and growth test. 

Combined frequency distribution for weight NOECs for all participating laboratories [2]. 

NOEC Frequency (46) Distribution 

I ---~--~~-- Tests with 2 Reps rtTests with 4 Reps 

- Median ~ r *I(a) .- >Z(b) 

-~. - 

*l(a) >W) 

41 0 

63 f 0 

47 18 

47 41 

53 12 

47 16 : 33 I 56 1 11 

61 I 28 33 I 56 ’ 11 

I 

57 43 0 

22 45 33 

88 0 63 25 j :: 

75 I 25 0 

Sample Median 

- 

__~ .~ .-~ 

Sodium Pentachlorophenate 1 59 

Sodium Pentachlorophenate * 37 

Potassium Dichromate 1 35 

Potassium Dichromate 2 d--- 12 

Refinery Effluent 301 

I ;; Refinery Effluent 401 

Utility Waste 501 11 

’ Percent of values with one concentration intervals of the median. 

’ Percent of values within two or more concentrations intervals of the median. 

A-l -9 
CERIODAPHMA (Ceriodaphnia dubia] 

Seven-day Reproduction Test 

Single Laboratory Precision Data 

Table A-1-25. Single laboratory precrsion of (Cerioduphnio dubio) 

reproduction test performed in using sodium pentachlorophenol as 

the reference toxicant [2]. 

Table A-1-26. Single laboratory precision, from six discrete laboratories, of the (Ceriodophnio dubia) reproduction test performed using sodium 

chloride (NaCI) as the reference toxicant. Tests were conducted in 1989 [4]. 

Test Number NOEC 

@W) 

lC25 IGO 

(w/U (mg/l) 

-- -~ .- -- 

19 0.30 

I 

0.3754 0.4508 

46A 0.20 0.0938 0.2608 

460 0.20 0.2213 0.2897 I 

49 0.20 0.2303 0.2912 

55 0.20 0.2306 0.3177 

56 0.10 0.1345 0.1744 

57 0.20 0.2241 ' 0.2827 

n: 7 7 7 

Mean: 0.20 I 0.2157 0.2953 

CV(%): , NA , 41.1 27.9 

NOEC Range: 0.25 - 0.30 mgil (these values all fell within the same 

concentration range). 

Prepared by Florence Kessler, TAI, Cincinnati, OH, fanuary 11, 1990 (ICp 

Program, version 1 .l b). 

laboratory 

Test NOEC 

Number 

k!5 

, (mg/l) 1 (mg/l) 1 (rug/l) 

‘Go 

A 0.77 

1.34 

1.32 

4 l.ooR ’ 0.67 1.28 

--F 

n: 

Mean: 

4 ! 0.50~ 

5 1 .oOR 

-.- -. 6- 

n: 

-. t +!L 

/ 

Mean: 

CV(%): 

--c -T 

i y:“’ 

; 1 .OOs 

o.50s 

i 3 

4 

I I.$ 

5 I 1 .oos 

1.31 

1.12 

6 

1.04 

24.11 

1.23 

0.46 

1.25 

1.13 

1.22 

0.69 0.88 

0.81 ' 1.16 

1.84 

1.57 ..- - - 

! 6 

1 

1.43 

24.37 -. - 

1.49 

1.02 

1.50 

1.44 

1.49 

6 1 .oos 1.21 ! 1.51 - -. - - _-. _ - ._ -*- - -.-- 

n: 

Mean: ( ! i.83 Y.13 ! Y.41 

CV(%) : NA 16.54 1 13.62 

R = Reproduction was the most sensitrve endpoint. 

5 = Survival was the most sensitive endpoint. 

Test NOEC 

Laboratory Number (w/l) 

IQ5 

(w/l) 

D 1 0.50R 

2 0.2SR 

3 1 .oos 

4 1 .oOs 

5 1.00s 

0.58 

0.30 

0.84 

1.04 

I 1.04 4 -- 6 -L - OSOR -r-L. 0 76 

n: 6 

Mean: 0.71 1 60.76 

-. - "'A 37.55 

1 1 .oosq 0.44 

4 ' 1.00~ 

5 1 .oos I 

- 

1.04 

1.06 

1.13 

1.13 

1.19 

-- 

6 

1.00 

27.96 

0.61 

0.63 

0.66 

0.65 

IC50 

(mg/O 

0.84 

0.60 

1.22 

1.38 

1.37 

1.14 --. 

6 

1.08 

28.56 

0.74 

1.37 

1.37 

1.42 

1.42 

1.46 -- 

6 

1.30 

21 20 -L 

1.13 

1.20 

0.83 

0.81 

Prepared by William Peltier, EPA, Region IV, November 28, 1990 (ICp Program, version 1 ,l b). 

A-1-10 
CERlODAPHulA (Curiodaplinia dubid 

Seven-day Larval Reproduction Test Interlaboratory Precision Data 

Table A-l-27. Interlaboratory precision of (Cerioduphnio dubia) 

reproduction test, using sodium chloride (NaCI) as the 

reference toxicant. The single lab precision data are presented 

in the preceding table [4]. - 

Laboratory 

A 

B 

C 

~ 

D 

E 

F 

n: 

Mean: 

a/(%): 

NOEC 

ml/l) 

0.83 

0.83 

0.83 

0.71 

1.00 

0.50 

6 

0.80 

NA 

0.89 1.26 

1.04 1.43 

1.13 1.41 

0.76 1.09 

1 .oo 1.30 

0.63 0.95 

- 

6 I 6 

0.91 ! 1.24 

20.53 15.17 

Prepared by William Peltier, EPA, Region IV, November 28, 

1990 (ICp Program, version 1 .l b). 

Table A-l-28 Interlaboratory precision of (Ceriodo 

reproduction test, using an industrial effluent as t f: 

hnia dub) 

e reference 

toxicant and sodium chloride (NaCI) as a reference toxicant. 

Tests were conducted in May 1987 [ 31. 

Effluent Reference Toxlcant 

Lab Go w lC25 w ICm (%,I ii, (96) 

A 6.20 4.9 33.0 -- il.8 - 

B 8.40 6.2 38.8 30.8 

D 7.69 5.8 36.3 29.4 

E 6.34 5.0 36.6 28.0 

F 4.00 1.2 8.1* 1.21* 

I 2.84 1.9 35.1 25.2 

K 6.89 5.3 18.4 13.2 

M 5.70 1.9 38.1 31 .o 

N 7.43 5.9 27.8 10.4 

0 0.04’ 0.02* 35.1 27.3 -^.-. 

t-l: 9 9 

Mean: ’ z.17 I.4 32.8 24.1 

CV(%): 29 67 21 31 

‘Values were excluded from mean calculations because they fell outside of + 2 

standard deviations. For this reason, these values are considered statistical 

outliers and, according to EPA’s toxicity methods guidance [2] on reference 

toxicant control charts, are excluded. 

Table A-l-29. Results of the performance evaluation for contract laboratories conducted for the California Regional Water Quality 

Control Board. All tests were conducted using sodium chloride and testing Cerioduphnio dobiu in the 7-day chronic 

tests 131. 

I / X Young/ h5 KU NOEC 

Lab Water ’ Age Control ’ (g/l NaCI) Endpt 

1 ’ Tap1 YCT/Algae 0-4;<24 1 7.80.202 (0.14-0.35) -- co.25 R 

2 Hard W3 TF/Algae O-4 26.51.3 (0.78-l .7) 1.0 R 

3 DMW4 YCT/Algae O-6 24.90.21 2 (0.17-0.54) <0.25 R 

4 Tap5 YCT o-4 17.20.49 (0.35-l .O) 0.5 R 

5 HRW YCT 0-4;<24 19.80.42 (0.20-I .I) 0.5 R 

6 Surface6 YCT/Algae O-6 14.80.90 (0.66-l .l) 0.25 R 

7 MHRW YCT/Algae 4-8 17.20.56 (0.24-0.64) 0.25 R 

8 MHRW YCT ~24 16.80.21 2 (0.11-o. 32) 0.25 R 

9 MHRW YCT O-4 12.80.71 (0.56-0.81) 0.50 R 

10 DMw4 YAT/Algae o-4 4 31 so.91 (0.45-l .l ) 1.0 R 

Mean: O.-j6 

CV(%): j 40 

’ Moderately hard tap water. MHRW = Moderately hard reconstituted water 

’ Dose response curve limited. HRW = Hard reconstituted water 

3 Hard well water. WV = Well water 

4 Ten percent diluted mineral water. YCT = Yeast-Cerophyl-Trout chow 

5 Dechlorinated Cake Ontario tap water. YAT = Yeast-Alfalfa-Trout chow 

6 Briones reservoir water. TF = Trout food suspension 

R = Reproductive endpoint Algae = S&nostrum capricornotum 

A-l -11 
CERIODAPHRIA (Cerhdapnnia dubiaj (contlnued) 

Seven-day Larval Reproduction Test Interlaboratory Precision Data 

Table A-1-30. Interlaboratory precision data for Ceriodophniu dubiu summarized for 

eight materials, including reference toxicants and effluents [S]. 

Test Mean 

Material ’ I& 

Sodium chloride 1.34 

Industrial 

Sodium chloride 

Pulp & Paper 

Potassium dichromate 

Pulp & Paper 

Potassium dichromate 

Industrial 

Mean 

cv 46 GS cv % / 

29.9 1 1.00 34.3 

3.4 : 83.3 

I I I 

3.2 78.1 

0.96 57.4 0.90 44.4 

60.0 28.3 47.3 27.0 

35.8 23.4 32.7 

70.2 55.7 f2.2 

53.2 29.3 46.8 

69.8 67.3 36.7 ~-. I 8 I 8 37.5 39.0 

1 Standard Deviation: 1 23.0 19.1 

SELEHASTRUM CA PRICORUUTUM 

Growth Test 

Single Laboratory Precision Data 

Table A-1-31. Single laboratory precision of (Selenastrum 

cupricorrrufum) growth test performed in using cadmium as 

the reference toxicant [2]. 

Test I 

Number 1 

1 2.3 

2 2.4 

3 2.3 

4 2.8 

5 2.6 

6 2.1 

7 2.1 

8 2.1 

1 ii;: ; iI;7 8.42 

44.1 

‘Outlier values are excltided from mean because they fell outside of a QA 

control table for these reference toxicants. 

Control Variation 

(%Cv) 

4.8 

9.6 

5.5 

13.3 

4.4 

8.2 

14.4 

7.1 

11.9 

5.0 

36.4’ 

77.8’ 

Note: Sodium chloride concentrations were 1, 2, 4, 8, and 16 g/l in all tests. 

Prepared by Dr. Cornelius Weber, EPA, Cincinatti, OH, March 15, 1991. 

A-1-12 
APPENDIX A-2 

EFFLUENT VARIABILITY DATA 
Table A-2-1. Percent mortality in 100 percent collected 1989 

by grab method (personal communication W. Pettier, EPA, 

Athens, GA). Results indicate variability over 24 hours and 

differences in species sensitivity over time. Tests were 

1830 

0030 

0630 

1230 

1830 

0030 

7/26/89 0630 0 

1830 0 

0030 0 

1230 0 

Date 

3/07/89 

3/07/89 

3/08/89 

3/08/89 

3/20/89 

3/20/89 

3/21/89 

3/21/89 

6/19/89 

6/19/89 

6/20/89 

D. pulex 

15 

85 

65 

30 

0 

100 

95 

70 

5 

40 

100 

100 

0 

100 

100 

55 

Time P. promelas 

1230 0 

conducted according to methods described in Reference 6. 

% Mortality in 100% Effluent 

C. dubia 

100 

100 

100 

80 

100 

100 

100 

100 

100 

100 

100 

100 

100 

100 

100 

100 

0 

0 

0 

0 

0 

0 

6/20/89 0630 0 

7125189 1230 0 

7/25/89 1830 0 

7126189 0030 0 

Table A-2-2. LC50s for a POTW effluent over 17 months. All 

tests were conducted using Ceriodaphnia dubio and tests were 

run for 48 hours. All tests were conducted according to the 

methods described in Reference 6. Dates with roman numeral 

notation mean that more than one sample was collected at 

different times over a short interval (1 to 2 days). (Data 

source: [8].) 

Sample Date 48 hours LC50 (%) 

08/23/86-I 71 

03/09/87-I 71 

05/02/87-I 35 

05/03/87-I 65 

05/04/87-I 71 

05/23/87-I 71 

05/23/87-II 71 

05/23/87-III 61 

06/27/87-I 36 

06/27/87-II 41 

06/27/87-III 18 

09/22/87-I 71 

12/18/87-I 87 

01/05/88-I 68 

01/05/88-II 63 

Mean LC50: 60.0 

cv (%): 31.1 

n: 15 

Table A-2-3. LC50s for a POTW effluent over 7 months. All 

tests were conducted using Ceriodaphnia dubia and tests were 

run for 48 hours. All tests were conducted according to the 

methods described in Reference 6. Dates with roman numeral 

notation mean that more than one sample was collected at 

different times over a short interval (1 to 2 days). (Data 

source: [8].) 

Sample Date 

10/06/87-I 

10/06/87-II 

10/06/87-III 

10/30/87-I 

12/03/87-I 

12/03/87-II 

01/12/88-I 

01/13/88-I 

02/03/88-IX 

02/03/88-X 

03/03/88-III 

03/03/88-IV 

03/23/88-I 

03/23/88-II 

04/28/88-I 

04/28/88-II 

05/17/88-I 

05/17/88-II 

Mean LC50: 

cv (%): 

n: 

48 hours LC50 (%) 

71 

71 

71 

87 

61 

35 

61 

58 

50 

50 

87 

81 

25 

35 

50 

55 

61 

35 

58.0 

31.4 

18 

A-2-1 
Table A-2-4. LCsOs for a POlW effluent over 12 months. 

Tests were conducted using either Ceriodaphnia dubia or 

fathead minnows or both. Ceriodaphnia tests were conducted 

for 48 hours while fathead minnow tests were 96 hours. Both 

the 4%hour and 96-hour fathead minnow results are shown in 

order to evaluate how the LCses for the two species compare. 

All tests were conducted according to the methods described 

in Reference 6. Dates with roman numeral notation mean that 

more than one sample was collected at different times over a 

short interval (1 to 2 days). (Data source: [8].) 

Table A-2-5. LCsOs for a POTS effluent over 4 months. 

Tests were conducted using either Ceriodaphnia dubia or 

fathead minnows or both. Ceriodophnio tests were conducted 

for 48 hours while fathead minnow tests were 96 hours. Both 

the 48-hour and g&hour fathead minnow results are shown in 

order to evaluate how the LCsOs for the two species compare. 

All tests were conducted according to the methods described 

in Reference 6. Dates with roman numeral notation mean that 

more than one sample was collected at different times over a 

short interval (1 to 2 days). (Data source: (El].) 

i _______ LCEFL-------- 

Sample Date ! C. dubia P. promelas 

- 48 hours 48 hours ’ 96 hours 

-.- --- 

, 1 

03/16/88-1 , 62 

+--s---i--T--- 

06/09/88-l 18 

* ’ -* 

09/08/88-l I 68 >lOO ; >lOO 

10/04/88-l ’ 61 -* l I -1 

10/04/88-II 63 

12/l 4/88-l , /70 58 1 34 

12/l 4/88-II 17 I 60 41 

02/17/89-l ) 35 I 61 39 I 

02/l 7/89-II 1 35 I 61 I 37 

03/22/89-l 35 81 64 

03/22/89-II ! 47 I 61 I 40 

Mean LCs0: 46 I 59.6 ’ 40.0 

CV (%): ! 42 22.4 I 29.7 

n: 

I 11 ; 7 I 7 

l Data not available. 

Note: Greater than (>) values were excluded from the mean LCc,o calculation. 

c __-__ &cl (W -__ 

Sample Date, C. dubicl , P. promelos 

&#fy; I !!;oclz;7z. 

1 l/l 5/88-f ’ 92’ / 67 1 37 

11/16/88-l 61 >lOO I 100 

1 l/30/88-11 >lOO >lOO I 33 

11/30/88-Ill 95 >lOO , >lOO 

12/08/88-l 100 I 87 54 

12/08/8&ll , >lOO ’ 87 / 53 

12/13/88-l j 90 >I00 

I 

77 

12/l 3/88-II 87 , 85 51 

01 /10/89-l 75 58 ; -* 

01 /10/89-h 61 ’ 41 

l 

-- 

I 

01 /19/89-l I 100 88 I 68 

01/19/89-II I 87 84 69 

01/25/89-l >lOO 

01/25/89-II 1 95 

87 I 64 

85 I 56 

01/31/89-l , 90 70 I 60 

01/31/89-II 

I 

63 70 I 60 

---- ----- ---------- 

-Mean: 

_---- 

78.4 75.8 - - 61.3 

cv (%): 33.1 19.6 I 27.7 

It 14 12 14 

*Not obtained. 

Note: Greater than (>) values were not included in the mean LCso calculation. 

A-2-2 
Table A-2-6. NOECs for a POlW effluent conducted 20 times 

over 1 year. All tests were conducted using Chompia parvulo 

according to methods described in Reference 1. All effluent 

samples were 24-hour composites collected post-chlorination. 

(personnal communication-Glen Thursby, SAIC, 

Narragansett, RI). 

Table A-2-7. NOECs for a POTW effluent over 1 year. All 

tests used Arbocia punctulata according to methods described 

in Reference 1. All effluent samples were 24-hour composites 

collected post-chlorination. (Data source: ERL-Narragansett, 

RI.) 

% Effluent 

-- -- -. .~-~ 

ICE I& i NOEC 

.~. - ___ .- .- 

12/09/85 

12/l O/85 

12/l l/85 

12/l 2/85 0.41 

12/l 3/85 3.09 

12/l 5185 2.16 

07/l 6186 2.99 

07/l 7186 3.59 

07/l B/86 .44 

07/l 9186 .47 

07/20/86 .24 

07/21/86 .ll 

07/22/l-36 

9/09/86 

09/l O/86 

09/l l/86 

09/l 2186 

09114186 ! .25 ! 1.76 ~. .~. 1 

0.65 1.23 

0.38 I 0.76 

0.69 1.50 

0.82 

I 4.12 

4.09 

4.33 

4.68 

4.76 

3.41 

3.98 

3.20 

.84 5.19 

.07 3.02 

.l 7 4.13 

.73 3.62 

.57 1.89 

l-i: 

Mean: 

I 18 18 18 

2.2 3.1 I 4.2 

1.25 

1.25 

2.50 

1.25 

5.00 

5.00 

5.00 

5.00 

5.00 

5.00 

7.50 

5.00 

5.00 

2.50 

7.50 

7.50 

1.25 

2.50 

cv (46): / 52.8 / 46.8 1 NA 

Table A-2-8. NOECs for a POlW effluent over 1 year. All 

tests used Mysidopsis bahia according to methods described in 

Reference 1. All effluent samples were 24-hour composites 

collected post-chlorination. (Data source: ERL-Narraqansett, 

RI.) . 

016 Effluent 

12/09 - 12/l 6/85 

n: 4 ,4 

R-Reproductive endpoint 

S-Suwval endpoint 

G-Growth endpoint 

Test Date 

--I 12fO9f 85 

12/l Of85 

12/l l/85 

12/l 2185 

12/l 3185 

12/14/85 

12/l 5185 

07/l 6/86 

07/l 7186 

07/l 8186 

07/l 9/86 

07/20/86 

07/21 I86 

07/22/86 

09/09/86 

09/l l/86 

09/l 2186 

09/l 3f86 

09f 14186 

09/l 5/86 

11/11/86 

11 /13/86 

1 l/14/86 

11/15/86 

Mean: 

cv (96): 

n: , 24 , 24 11 

I.- - --- -- 

bi 

1.09 4 

1.41 

0.75 

3.28 

2.65 

1.11 

1.29 

0.17 

0.21 

0.63 

1.09 

0.54 

0.40 

0.40 

0.31 

0.47 

0.21 

3.30 

0.23 

0.10 

0.27 

0.88 

0.82 

0.34 

0.91 

101.3 

% Effluent 

kl 

1.71 

2.84 

1.09 

4.06 

5.32 

1.60 

1.84 

0.35 

0.46 

0.86 

1.68 

1.13 

0.58 

0.56 

0.41 

0.79 

0.48 

5.42 

0.35 

0.15 

0.54 

1.48 

1.61 

0.56 __- 

1.49 

96.9 

.-. - 

NOEC 

0.65 

0.65 

0.65 

1.30 

2.50 

0.65 

0.65 

<0.30 

<0.30 

<0.30 

<0.30 

<0.30 

<0.30 

<0.30 

<0.30 

<0.60 

<0.20 

1.30 

<0.20 

<0.20 

1.30 

0.30 

0.60 

<0.30 

I / 0.95 

NA 

Note: Less than (<) values wete excluded from CV and mean NOEC calculations. 

A-2-3 
Table A-2-9. NOECs for a POTW effluent over 1 year. All 

tests used Menidiu beryllino according to methods described 

in Reference 1. All effluent samples were 24-hour composites 

collected post-chlorination. (Data source: ERL-Narragansett, 

RI.) 

ll/ll - 11/18/86 

--- 

Mean: 

cv (%): 

n: 

NC - Value is not calculable. 

NC 10.0 

14.9 20.8 10.0 

4.3 3.0 NA 

3 3 4 

Table A-2-10. LC5Os for a refinery effluent over 14 months. 

Tests were conducted using either Ceriodaphnio dubia or 

fathead minnows Pimepholes promelas or both. Ceriodophnia 

tests were conducted for 48 hours while fathead minnow tests 

were 96 hours. Both the 48-hour and 96hour fathead minnow 

results are shown in order to evaluate how the LC5Os for the 

two species compare. All tests were conducted according to 

the methods described in Reference 6. Dates with roman 

numeral notation mean that more than one sample was 

collected at different times over a short interval (1 to 2 

days). (Data source: [8].) 

01/05/88 

02/09/88-l 

02/09/88-l 

03/02/88-l 

03/02/88-II 

03/24/88-l 

05/06/88-l 

07/l 4/88-l 

I <12 38 

35 35 

35 

t 

55 61 

07/28/88-l , 37 I 35 22 

- - --___ 

.- -.- 

12/01/87 / 15 

19 

<12 

cl2 

l 

l 

t 

l 

25 

07/28/88-II 

09/29/88-l 

12/01/88-l 

12/07/88-l 

01/27/89-l 

01/27/89-/l 

28 31 I 

41 

75 

’ 

39 

56 

18 67 

I 

~25 

25 

34 

13 

100 61 

71 60 

03/23/89-l 1 58 ’ 54 20 

--------+- 

Mean LCsO: 43 I 45 24 

cv (%): 

1 

54 28 I 32 

n: 16 15 10 

l Data not available 

Note: Less than values excluded from mean LCSO calculations 

Table A-2-11. LC5Os for a refinery effluent conducted over 6 

months using fathead minnows (Pimephales promeias), 

Ceriodophnia dubio, and mysids (Mysidopsis bahio), according 

to methods described in Reference 6. (Data source: Dorn, 

1989.) 

4/l 7186 71.2 37.2 

4123186 , 71.8 35.9 38 -. 0 

5/14/86 82.0 38.7 I 35.8 

5/28/86 65.4 22.0 

6/l l/86 82.0 I - 24.7 

Mean NOLC: 66.3 32.9 ’ 32.8 

cv (%): ! 18.7 19.5 21.6 

n: 1 11 10 3 

Table A-2-12. NOECs for a refinery effluent conducted over 6 

months using fathead minnows (Pimepholes promelos), 

Ceriodaphnio dubia, and mysids (Mysidopsis bahia), according 

to methods described in References 1 and 2. (Data source: 

Dorn, 1989.) 

3/ 19186 10.1 14.1 

4/02/86 18.0 14.1 - 

A-2-4 
Table A-2-13. LCsos for a manufacturing effluent conducted over 2 

ears. All tests were conducted using Dophnio mo 

methods described in Reference 6. (Data source: [8 4 

no according to 

.) 

Test Date (% Siient) 

1982 (1 st quarter) 56 

1982 (4th quarter) 90 

1982 (4th quarter) 70 

1983 (3rd quarter) 69 

1983 (3rd quarter) 36 

1983 (3rd quarter) 36 

1983 (3rd quarter) 32 

1983 (3rd quarter) <ia 

1983 (3rd quarter) 28 

1983 (3rd quarter) 67 

1983 (3rd quarter) <lO 

1983 (4th quarter) 46 

1983 (4th quarter) 75 

1983 (4th quarter) 78 

1983 (4th quarter) 24 

1983 (4th quarter) 26 

1983 (4th quarter) 32 

1983 (4th quarter) 19 

Mean LCso: 45.1 k24.3 

cv (%): 53.9 

n: la 

Note: Less than (<) values were excluded from the mean LCso calculations. 

Table A-2-14. LCgy for a manufacturing effluent conducted over 8 

All tests were conducted using Pimephales promelas according 

r?%thods described in Reference 6. (Data source: [8].) 

Test Date 

1979 (1st quarter) 72.0 

1979 (1 st quarter) 62.0 

1979 (1 st quarter) 52.0 

1979 (3rd quarter) 39.0 

1981 (2nd quarter) 64.0 

1981 (4th quarter) 70.0 

1982 (2nd quarter) 44.0 

1982 (2nd quarter) 66.0 

I985 (1 st quarter) 59.6 

I985 (4th quarter) >lOO.O 

I986 (2nd quarter) I 49.2 

I986 (2nd quarter) 63.8 

I986 (2nd quarter) 50.0 

I986 (3rd quarter) 75.7 

I986 (3rd quarter) 80.0 

I986 (3rd quarter) 79.0 

1986 (4th quarter) 71 .o 

Wean LCso: 64.5 + 15.1 

zv (%): 23.5 

1: 17 

Note: Greater than (>) values were excluded from the mean LCso calculations. 

Table A-2-15. LC5Os for a manufacturing effluent conducted 

over 5 years. All tests were conducted using Pimephales 

promelos according to methods described in Reference 6. 

(Data source: [a].) 

I 

Test Date (% Zent) 

1980 (1 st quarter) 18.0 

1980 (2nd quarter) 11 .o 

1980 (3rd quarter) 32.0 

1980 (4th quarter) 16.0 

1981 (1 st quarter) 32.0 

1981 (2nd quarter) 

1981 (3rd quarter) 

1981 (4th quarter) 

1982 (1 st quarter) 

1982 (2nd quarter) 

1982 (3rd quarter) 

1982 (4th quarter) 

1983 (1 st quarter) 

1983 (2nd quarter) 

1983 (3rd quarter) 

23.0 

17.0 

46.0 

9.0 

32.0 

28.0 

52.0 

34.0 

33.0 

20.0 

A-2-5 

1983 (4th quarter) 43.0 

1984 (1 st quarter) 45.0 

1984 (2nd quarter) 19.0 

1984 (3rd quarter) 61 .O 

Table A-2-16. LC5os for a manufacturing effluent conducted 

over 5 years. All tests were conducted using Daphnio magna 

according to methods described in Reference 6. (Data source: 

PI.) 

Test Date I 

--._B .~.. 

981 (2nd quarter) 100.0 

981 (3rd quarter) I >lOO.O 

982 (3rd quarter) >lOO.O 

984 (4th quarter) 80.0 

985 (1 st quarter) 75.0 

1986 (1 st quarter) 25.0 

1986 (2nd quarter) 82.0 

1987 (1 st quarter) 75.0 

1987 (1 st quarter) I 24.0 

1987 (1 st quarter) ,100.o 

,100.o 

65.9 + 29.5 

44.8 

n: I 11 

Note: Greater than (>) values were excluded from the mean LCc,o calculations. 
Table A-2-17. LC5Os for a manufacturing effluent conducted 

over 7 years. Ail tests were conducted using Uaphnio pulex 

according to methods described in Reference 6. (Data source: 

Table A-2-18. LC5Os for a manufacturing effluent conducted 

over 3 months. All tests were conducted using Daphnio 

mogno according to methods described in Reference 6. (Data 

source: [8].) 

Test Date (% kkent) 

1980 (izquarter) -- - 

-. .-- - ~ 

55.0 

1980 (4th quarter) 33.0 

1981 (I st quarter) 60.0 

1981 (1 st quarter) 24.0 

1981 (1 st quarter) >lOO.O 

1981 (2nd quarter) >lOO.O 

1981 (3rd quarter) >lOO.O 

1982 (3rd quarter) 

1 

1986 (3rd quarter) 

1986 (3rd quarter) -1. 

2100.0 

>lOO.O 

>lOO.O 

- - .- - 

Mean LCse: 43.0 + 17.3 

cv (%): 40.2 

n: 10 

Note: Greater than (>) values were excluded from the mean CC50 calculations. 

Test Date I (96 EL) 

982 (4th quarter) 

982 (4th quarter) 

982 (4th quarter) 

982 (4th quarter) 

982 (4th quarter) 

982 (4th quarter) 

982 (4th quarter) 

982 (4th quarter) 

982 (4th quarter) 

982 (4th quarter) 

982 (4th quarter) 

1982 (4th quarter) 

Mean LCso: 

cv (96): 

n: 

I 

>lOO.O 

81 .O 

I 57.0 

61 .O 

87.0 

90.0 

90.0 

>lOO.O 

z-1 00.0 

54.0 

74.0 

I >lOO.O 

74.3 * 15.1 

20.3 

12 

Note: Greater than (P) values were excluded from the mean LCso calculations. 

A-2-6 
APPENDIX A-3 

ACUTE-TO-CHRONIC RATIO DATA 
Table A-3-1. Example of Acute-to-Chronic Ratios 

Oil Refinery1 

Mean ACR: 

n: 

Range: 

Fathead 

Minnow 

1.89 

3.47 

2.60 

2.87 

2.33 

2.43 

5.26 

5.08 

2.74 

3.11 

5.1 

Ceriodaphnia 

9.09 

3.89 

6.58 

3.63 

6.91 

7.05 

7.11 

3.63 

2.59 

5.5 

4.4 

> 10.0 

> 7.1 

> 3.3 

> 2.0 

> 3.0 

2.8 

5.42 

3.3 5.3 

11 13 

1.89-5.26 2.59->10.0 

Mysids 

1.58 

1.49 

1.84 

1.64 

3 

1.49-1.84 

1 Personal communication P. Darn. 

2 Personal communication M.L.C. Ramos and E. Bertoletti (Sao Paulo, Brazil). 

Note: Greater than (>) values were excluded from mean calculations, 

Table A-3-2. Examples of Acute-to-Chronic Ratios 

Chemical Manufacturers 

Mean ACR: 

n: 

Range: 

Fathead 

Minnow 

0.17 

0.07 

8.4 

7.6 

>3.0 

3.9 

>3.0 

1.8 

3.7 

6 

0.07 - 8.4 

Ceriodaphnia 

> 1.0 

> 1.0 

>10.0 

>50.0 

> 2.9 

> 1.4 

1.4 

1.4 

3.9 

2.8 

> 2.0 

> 4.0 

4.0 

1.4 

5.5 

1.8 

> 3.3 

> 3.3 

> 3.3 

1.4 

> 2.0 

5.5 

1.5 

1.4 

5.0 

> 10.0 

> 2.0 

> 3.3 

3.111 

14.01 

4.31 

2.51 

1.81 

5.51 

5.41 

3.72 

20 

1.4->50 

1 Personal communication M.L.L.C. and E. Bertoletti (Sao Paula, Brazil). 

2 Greater than (>) values were excluded from the mean calculation. 

A-3-1 
Table A-3-3. Example of Acute-to-Chronic Ratios 

POTWs 

Fathead I 

Minnow 1 Ceriodaphnia 

1 ::; 5.5 14

I > 

i 13.0 1.8 > 1.0 

1.5 ! 1.0 

2.6 1.8 I 

I 9.3 1.4 

> 1.0 2.0 

> 3.0 2.4 

I 5.3 3.0 

I 3.3 ; 3.0 

5.4 5.5 

I I 

> 3.0 4.9 I 3.0 I > 2.0 

I > 8.0 2I 

/ > 1.0 

> 3.3 

I > 2.0 

4.4 

16.1 

> 4.0 

> 3.3 

>lO.O 

I 2.6 

I 5.7 

2.8 

I 

I 

I >lO.O 

> 2.0 

I 1.4 

I I 2.6 

I 

> 3.3 

I I 5.5 18

I 1.5 

I I > 3.3 

5.5 

Mean ACR: ; 4.9 ----/- 3.8* 

il: / 11 21 

Range: I 1.5 - 9.3 I 1.4 - 16.1 

l Greater than (>) values were excluded from mean calculations. 

A-3-2 
APPENDIX A 

REFERENCES 

1. 

2. 

3. 

4. 

5. 

6. 

7. 

8. 

Weber, Ct. et. al., eds. 1988. Short-Term Methods for Estimating the Chronic Toxicity of Effluents and Receiving Waters 

to Marine and Estuarine Organisms. EPA 600/4-87/028. Office of Research and Development, Cincinnati, OH. 

Weber, C.I. et. al., eds. 1989. Short-Term Methods for Estimating the Chronic Toxicity of Effluents and Receiving Waters 

to Freshwater Organisms, 2d ed. EPA 600/4-89/001. Office of Research and Development, Cincinnati, OH. 

Norberg-King, T.j. Memorandum: Summary of Effluent Toxicity Data for the Technical Support Document for Water 

Quality-based Toxics Control, Chapters 2 and 3. April 11, 1989. U.S. EPA, Environmental Research Laboratory, Duluth, 

MN, to M.A. Heber, U.S. EPA, Enforcement Division, Washington, DC. Memorandum: Technical Support Document for 

Water Quality-based Toxics Control Data: Fathead Minnow Round Robin Data Analyzed to Obtain the IC25s. May 6, 

1990. U.S. EPA, Environmental Research Laboratory, Duluth, MN, to M.A. Heber, U.S. EPA, Enforcement Division, 

Washington, DC, 

Peltier, W. January 23, 1991. Memorandum: Intra- and Interlaboratory Precision Data from Six Laboratories Conducting 

Short-Term Chronic Toxicity Tests Using Ceriodaphnia dubia. U.S. EPA, Region IV, Environmental Services Division, 

Athens, GA, to M.A. Heber, U.S. EPA, Enforcement Division, Washington, D.C. 

DeCraeve, GM., 1.0. Cooney, B.H. Marsh, T.L. Poltock, and N.G. Reichenbach. 1989. Intra- and Interlaboratory Study to 

Determine the Reproducibility of the Seven-day Ceriodaphnia dubia Survival and Reproduction Tests. Battelle, 

Columbus Division, Columbus, OH (in preparation). 

Peltier, W., and Cl. Weber. 1985. Methods for Measuring the Acute Toxicity of Effluents to Aquatic Organisms, 3d ed. 

EPA 600/4-85/013. Office of Research and Development, Cincinnati, OH. 

Ward, S.H. 1990. Technical Report: Procedures for Toxicant Testing and Culture of the Marine Macroalga, Champia 

parvula. U.S. EPA, Region II. 

Norberg-King, T.J. May 4, 1989. Memorandum: Additional Data for the Technical Support Document for Water Qualitybased Toxics Control, Chapters 2 and 3. U.S. EPA, Environmental Research Laboratory, Duluth, MN, to M.A. Heber, U.S. 

EPA, Enforcement Division, Washington, DC. 

A-3-3 
APPENDIX B-1 

SUMMARY OF CLEAN WATER ACT PROVISIONS 
CLEAN WATER ACT (33 U.S.C. 1251 SEQ.) 

Statutory Authority for the Use of Toxicity Testing and Whole Effluent Toxicity 

Limitations in NPDES Permits: 

Over the years, a developmental process has occurred regarding the use of biological techniques to assess 

effluent discharges and set permit limits. The acquisition of data and the development of new techniques has 

contributed to the refinement of toxicity testing methods, thus enabling EPA to more fully act in accordance 

with its mandates to implement statutory requirements relating to the attainment and maintenance of water 

quality. 

Toxicity testing of Whole Effluents and Whole Effluent toxicity limitations in National Pollutant Discharge 

Elimination System (NPDES) permits are essential components in the control of the discharge of toxic 

pollutants to the nation’s waters. The use of toxicity testing and Whole Effluent toxicity limitation in the 

NPDES program is clearly authorized by the Clean Water Act (CWA). 

Relevant provision of the CWA that provide the statutory authority for using toxicity testing and Whole 

Effluent toxicity limitations include the following: 

• 

• 

• 

• 

• 

• 

• 

• 

Section 101 (a) sets forth not only the goal of restoring and maintaining the “chemical, physical, and 

biological integrity of the Nation’s waters” (emphasis added), but also in Section 101(a)(3) the 

national policy of prohibiting the “discharge of toxic pollutants in toxic amounts” (emphasis added). 

As defined at Section 502(15), biological monitoring means that “determination of the effects on 

aquatic life, including accumulation of pollutants in tissue, in receiving waters due to the discharge 

of pollutants (A) by techniques and procedures, including sampling of organisms representative of 

appropriate levels of the food chain appropriate to the volume and the physical, chemical, and 

biological characteristics of the effluent, and (B) at appropriate frequencies and locations.” 

Section 304(a)(8) requires EPA to develop information on methods, including biological monitoring 

and assessment methods, to establish and measure water quality criteria for toxic pollutants on bases 

other than pollutant by pollutant criteria. 

Section 303(c)(2)(8) states, “Nothing in this section shall be construed to limit or delay the use of 

effluent limitations or other permit conditions based on or involving biological monitoring or 

assessment methods...” (emphasis added). 

Section 302(a) provides the authority to establish water quality-based effluent limitations on 

discharges that interfere with the attainment or maintenance of that water quality which shall assure 

protection of public health, public water supplies, and the protection and propagation of a balance 

population of shellfish, fish and wildlife, among other uses. The effluent limitations established 

must reasonably be expected to contribute to attainment or maintenance of such water quality. 

Under Section 301(b)(1)(C) and Section 402, all NPDES permits must comply with any more stringent 

limitations necessary to meet applicable water quality standards, whether numeric or narrative. 

CWA Section 308(a) and Section 402 provide authority to EPA or the Sate to require that NPDES 

permittees/applicants use biological monitoring methods and provide chemical, toxicity, and 

instream biological data when necessary for the establishment of effluent limits, the detection of 

violations, or the assurance of compliance with water quality standards. 

Section 510 provides the authority for states to adopt or enforce any standards or effluent limitations 

for the discharge of pollutants only on the condition that such limitations or standards are no less 

stringent than those in effect under the CWA. 

B-1-1 
APPENDIX B-2 

POLICIES FOR TOXICS CONTROL 
9016 Federal Register / Vol. 49, No. 48 / Friday, March 9, 1984 / Notices 

[OW-FRL-2533-1] 

Development of Water Quality-Based 

Permit Limitations for Toxic Pollutants; 

National Policy 

AGENCY: Environmental Protection 

Agency (EPA). 

ACTION: Notice. 

SUMMARY: EPA has issued a national 

policy statement entitled “Policy for the 

Development of Water Quality-Based 

Permit Limitations for Toxic Pollutants.” 

This policy addresses the technical 

approach for assessing and controlling 

the discharge of toxic substances to the 

Nation’s waters through the National 

Pollutant Discharge Elimination System 

[NPDES) permit program. 

FOR FURTHER INFORMATION CONTACT: 

Bruce Newton or Rick Brandes, Permits 

Division (EN-336). Office of Water 

Enforcement and Permits. U.S. 

Environmental Protection Agency, 

Washington. D.C. 20460-7010. 

B-2-1 
Federal Register / Vol. 49, No. 48 / Friday, March 9, 1984 1 Notices 9027 

~UPPLENENTMY INFORMA~~OPK As the 

water pollution control effort in the 

United State8 progresses and the 

“traditional” pollutants (oxygen 

demanding and eutrophying materials) 

become sufficiently treated to protect 

water quatity. attention is shifting 

towards pollutants that impact water 

quality through toxic effects. Compared 

with the traditional pollutants, 

regulation of toxic pollutants is 

considerably more difficult. The 

difficulties include (I) the great number 

of toxic chemicals that may potentially 

be discharged to receiving waters and 

the difficulties in their analysis: [z) the 

changes in the toxic effects of a 

chemical resulting from reactions with 

the matrix of constituents in which it 

exists; and [3) the inability to predict the 

effects of exposure to combinations of 

chemicals. 

To overcome some of these problems. 

EPA and the States have begun to use 

aquatic toxicity tests and various human 

health assessment techniques to 

complement chemical analyses of 

effluents and receiving water samples. 

Because these techniques or their 

application to effluent testing are new, 

EPA and the States have been cautious 

in their use. Based on EPA’s evaluation 

of these techniques and the experiences 

of several States, EPA is now 

recommeding the use of biological 

techniques as a complement to 

chemical-specific analyses to assess 

effluent discharges and express permit 

limitations. EPA has issued these 

recommendations through a statement 

of policy and is developing a technical 

guidance document to help implement 

I he policy. 

The complete test of the national 

policy statement follows: 

Policy for the Development of Water 

Quality-Based Permit Limitations for 

Toxic Pollutants 

Stoten?eni $poIic;, 

To confrr~i pollutants beyond Best 

Available Technology Economically 

Achievable (BAT). secondary treatment. 

and other Clean Water Act technologybased requirements in order to meet 

water quality standard% the 

Environmental Protection Agency (EPA) 

will use an integrated strategy 

consisting of both biological and 

chemical methods to address toxic and 

nonconventional pollutants from 

industrial and municipal sources. Where 

State standards contain numerical 

criteria for toxic pollutants, National 

Pollutent Discharge Elimination System 

(NPDES) permits will contain limits as 

necessary to assure compliance with 

these standards. In addition to enforcing 

specific numerical criteria, EPA and the 

States will use biological techniques and 

avaiiabte data on chemical effects to 

assess toxicity impacts and human 

health hazards based on the general 

standard of “no toxic materials in toxic 

amounts.” 

EPA, in its oversight role, will work 

with States to ensure that these 

techniques are used wherever 

appropriate. Under section 308 and 

section 402 of the Clean Water Act (the 

Act), EPA or the State may require 

NPDES permit applicants to provide 

chemical, toxicity. and instream 

biological data necessary to assure 

compliance with standards. Data 

requirements may be determined on a 

case-by-case basis in consultation with 

the State and the discharger. 

Where violations of water quality 

standards are identified or projected, 

the State will be expected to develop 

water quality-based effluent limits for 

inclusion in any issued permit. Where 

necessary, EPA will develop these limits 

in consultation with the State. Where 

there is a eignificant likelihood of toxic 

effects to biota in the receiving water, 

EPA and the States may impose permit 

limits on effIuent toxicity and may 

require an NPDES permittee to conduct 

a toxicity reduction evaluation. Where 

toxic effects are present but there is a 

significant likelihood that compliance 

with technology-based requirements will 

sufficiently mitigate the effects, EPA and 

the States may require chemical and 

toxicity testing after installation of 

treatment and may reopen the permit to 

incorporate additional limitations if 

needed to meet water quality standards. 

(Toxicity data, which are considered 

“new information” in accordance with 

40 CFR 122.6z(a)(Z]. could constitute 

cause for permit modification where 

necessary.) 

To carry out this policy, EPA Regional 

Administrators will assure that each 

Region has the capability to conduct 

water quality assessments using both 

biological and chemical methods and 

provide technical assistance to the 

States. 

Background 

The Clean Water Act establivhes two 

principal bases for effluent limitations. 

First, existing dischargers are required 

to meet technology-based effluent 

limitations that reflect the best controls 

available considering economic impacts. 

New source dischargers must meet the 

best demonstrated technology-based 

controls. Second, where necessary, 

additional requirements are imposed to 

assure attainment and maintenance of 

water quality standards established by 

the State8 and approved by EPA. In 

0-2-2 

establishing or reviewing NPDES permit 

limits, EPA must ensure that the limits 

will result in the attainment of water 

quality standards and protect 

designated water uses, including an 

adequate margin of safety. 

For toxic and nonconventional 

pollutants it may be difficult in some 

situations to determine attainment or 

nonattainment of water quality 

standards and set appropriate limits 

because of complex chemical 

interactions which affect the fate and 

ultimate impact of toxic substances in 

the receiving water. In many cases, all 

potentially toxic pollutants cannot be 

identified by chemical methods. In such 

situations, it is more feasible to examine 

the whole effluent toxicity and instream 

impacts using biological methods rather 

than attempt to identify all toxic 

pollutants, determine the effects of each 

pollutant individually, and then attempt 

to assess their collective effect. 

The scientific basis for using 

blological techniques has advanced 

significantly in recent yearn. There is 

now a general consensus that an 

evaluation of effluent toxicity, when 

adequately related to instream 

conditions, can provide a valid 

indication of receiving system impacts. 

This information can be useful in 

developing regulatory requirements to 

protect aquatic life, especially when 

data from toxicity testing are analyied 

in conjunction with chemical and 

ecological data. Generic human health 

effects methods, such as the Ames 

mutagenicity test. and structure-activity 

relationship techniques are showing 

promise and should be used to identify 

potential hazards. However, poliutantspecific techniques are the best way to 

evaluate and control human health 

hazards at this time. 

Biological testing of effluents is an 

important aspect of the water qualitybased approach for controlling toxic 

pollutants. Effluent toxicity data in 

conjunction with other data can be used 

to establish control priorities, assess 

compliance with State water quality 

standards, and set permit limitations to 

achieve those standards. All States have 

water quality standards which include 

narrative statements prohibiting the 

discharge of toxic materials in toxic 

amounts. A few State standards have 

criteria more specific than narrative 

criteria [for example, numerical criteria 

for specific toxic pollutants or a toxicity 

criterion to achieve designated uses). In 

States where numerical criteria are not 

specified, a judgment by the regulatory 

authority is required to set quantitative 

water quality-based limits on chemicals 

and effluent toxicity to assure 
9018 Federal Register 1 Vol. 49, No. 48 1 Friday, March 9, 1984 I Notices 

. __ 

compliance with water qualit> 

standards. 

Note.-Section 308 of the Act trnd 

corresponding State statutes authorizca WA 

and the Stales to require of the owner/ 

operator any information reasonably required 

lo determine permit limits and to determine 

compliance with standards or permit linlik 

Kological methods are specifically 

mentioned. Toxicity permit limits arp 

authorized under Secllon 301 and 4(32 nnrl 

supported by Section 101. 

Application 

This policy applies to EPA and the 

States. The policy addresses the use of 

chemical and biological methods for 

assuring that effluent discharges are 

regulated in accordance with Federal 

and State requirements. This policy was 

prepared. in part, in response to 

concerns raised by litigants to the 

Consolidated Permit Regulations [see FR 

52079, November 18.1982), Use of these 

methods for developing water quality 

standards and trend monitoring are 

discussed elsewhere (see 48 FR 51400. 

November 8,1983 and Basic Water 

Monitoring Program EPA-440/9-76-025). 

This policy is part of EPA’s water 

quality-based control program and does 

not supersede other regulations, policy. 

and guidance regarding use 

attainability, site-specific criteria 

modification, wasteload allocation. and 

water quality management. 

fmplementotion 

Slate Role 

The control of toxic substances to 

protect water quality must be done in 

the context of the Federal-State 

partnership. EPA will work 

cooperatively with the States in 

identifying pbtential water quality 

standards violations. assembling 

relevant data, developing appropriate 

testing requirements. determining 

whether standards are being violated, 

and defining appropriate permit limits. 

No&.-Under sections 303 and 401 of the 

Act. States are given primary responsibility 

for developing water quality standards and 

limits to meet those standards. FPA’R role is 

to review the Stale standards and limits and 

develop revised or additional standards or 

limits as needed to meet the requirements of 

the Act. 

Integration of Approaches 

The type of testing that is most 

appropriate for assessing water quality 

impacts depends on the type of effluent 

and discharge situation. EPA 

recommends that an integrated 

approach, including both biologictll and 

chemical techniques. be used to assess 

and control water quality. The principal 

advantages of chemical-specific 

techniques are that (I) chemical 

analyses are usually less expensive than 

biological measurements in simple 

cases; (21 treatment systems are more 

easily designed to meet chemical 

requirements than’toxicity requirements: 

and (3) human health hazards and 

bioaccumulative pollutants can best be 

addressed at this time by chemicalspecific analysis. The principal 

advantages of biological techniques are 

Ihal (t J the effects of complex 

discharges of many known and 

unknown constituents can be measured 

only by biological analyses; (21 

bioavailability of pollutants after 

discharge is best measured by toxicity 

testing: and (3) pollutants for which 

there are inadequate chemical analytical 

methods or criteria can be addressed. 

Pollutant-specific chemical analysis 

techniques should be used where 

discharges contain few, well-quantified 

pollutants and the interactions and 

effects of the pollutants are known. In 

addition, pollutant-specific techniques 

should be used where health hazards 

are a concern or bioaccumulation is 

suspected. Biological techniques should 

be used where effluents are complex or 

where the combined effects of multiple 

discharges are of concern. EPA 

recognizes that in many cases both 

types of analysis must be used. 

Testing Requirements 

Requirements for dischargers* to 

collect information to assess attainment 

or nonattainment of State water quality 

standards will be imposed only in 

selected cases where the potential for 

nonattainment of water quality 

standards exists. Where water quality 

problems are suspected but there is a 

strong indication that complying with 

BCT/BAT will sufficiently mitigate the 

impacts. EPA recommends that 

applicable permits include testing 

requirements effective after BCT/BAT 

compliance and reopener clauses 

allowing reevaluation of the discharge. 

The chemical, physical, and biological 

testing to be conducted by individual 

dischargers should be determined on a 

case-by-case basis. In making this 

determination, many factors must be 

considered, including the degree of 

impact. the complexity and variability of 

the discharge, the water body type and 

hydrology, the potential for human 

health impact. the amount of existing 

data, the level of certainty desired in the 

water qualily assessment, other sources 

of pollutants, and the ecology of the 

receiving water. The specific data 

needed to measure the effect that a 

discharger has on the receiving water 

will vary according to these and other 

factors. 

82-3 

An assessment of water quality 

should, to the extent practicable, include 

other point and nonpoint sources of 

pollutants if the sources may be 

contributing to the impacts. Special 

attention should be focused on Publicly 

Owned Treatment Works (POTW’s) 

with a significant contribution of 

industrial waste-water. Recent studies 

have indicated that such POTW’s are 

cJftel1 SignifiCank SOUrCeS Of bXiC 

materials. When developing monitoring 

requirements. interpreting data, and 

determining limitations. permit 

engineers should work closely with 

water quality staff at both the State and 

Federal Irvels. 

A discharger may be required to 

provide data upon request under section 

308 of the Act, or such a requirement 

may be included in its NPDES permit. 

The development of a final assessment 

may require several iterations of data 

collection. Where potential problems are 

identified, EPA or the State may require 

monitoring to determine whether more 

information is needed concerning water. 

quality effects. 

Use of Data 

Chemical, physIcal. and biological 

data will be used to determine whether, 

after compliance with BCT/BAT 

requirements. there will be violations of 

State water quality standards resulting 

from the discharge(s). The narrative 

prohibition of toxic materials in toxic 

amounts contained in all State 

standards is the basis for this 

determination taking into account the 

designated use for the receiving water. 

For example. discharges to waters 

classified for propagation of cold water 

fish should be evaluated in relation to 

acute and chronic effects on cold water 

organisms. potential spawning areas, 

and effluent dispersion. 

Setting Permit Limikions 

Where violations of water quality 

standards exist or are projected, the 

State and EPA will determine pollution 

control requirements that will attain the 

receiving water designated use. Where 

effluent toxicily is an appropriate 

control parameter, permit limits on 

effluent toxicity should be developed. 4 

such cases, EPA may also iequire a 

permittee to conduct a toxicity reduction 

evaluation. A toxicity reduction 

evaluation is an investigation conducted 

within a plant or municipal system to 

isolate the sources of effluent toxicity, 

specific causative pollutants if possible, 

and determine the effectiveness of 

pollution control options in reducing the 

effluent toxicity. If specific chemicals 

are identified as the cause of the water 
F&al Register / Vol. 49, No. 49 :/ Friday, March 9, 1984 } Notices 

- 

9019 

quality standards violaliori. these 

individual poWants should be limited. 

If a toxicity reduction evaluation 

demonstrates that limiting ap indicator 

parameter will ensure attainment of the 

water quality-based effluent toxicity 

requirement. limits on the indicator 

parameter should be considered in lieu 

of limits on effluent toxicity. Such 

indicator limits are not limits on 

causative pollutants but limits 

demonstrated to result in a specific 

toxicity reduction. 

Monitoring 

Where pollution control requirements 

are expressed in terms of a chemical or 

toxicological piirameter, compliance 

monitoring must include monitoring for 

that parameter. If an indicator 

parameter is used based on the results 

of a toxicity reduction evaluation, 

periodic toxicity testing may be required 

to confirm the adequacy of the indicator. 

Where biological data were used to 

develop a water quality assessment or 

where the potential for water quality 

standards violations exist. biological 

monitoring (including instream 

monitoring) may be required to ensure 

continuing compliance with water 

quality standards. 

EPA believes that the intelligent 

application of an integrated strategy 

using both biological and chemical 

techniques for water quality assessment 

will facilitate the development of 

appropriate controls and the attainment 

of water quality standards. W’A looks 

forward IO working with the States in a 

spirit of cooperation lo further refine 

these techniques. 

Policy sipned February 3. 1984 by jack E. 

Ravan. As&ant administrator for Water. 

Dated: Febrwry 16. 1%. 

jack E. Raven. 

As~r’stonf .4hm,nisfmtcrfor Litter 

IFII 0~ M-twS Fllkd 3-844 1.45 ,ml 

lll.UYG CODE sso4o-Y 

B-24 
APPENDIX B-3 

REGULATIONS FOR TOXICS CONTROL 
UNITED STATES ENVIRONMENTAL PROTECTION AGENCY 

WASHINGTON, D.C. 20460 

AUG 21 I989 OFFICE OF 

WATER 

MEMORANDUM 

SUBJECT: New Regulations Governing Water Quality-Based 

Permitting in the NPDES Permitting Pro

FROM: James R. Elder, Director 

of Water Enforcement 

TO: Water Management Division Directors 

Regions I - X 

On May 26, 1989 the Deputy Administrator signed regulations 

that implement section 304(1) of the CWA. The regulations became 

effective upon his signature and were published in the Federal 

Register on June 2, 1989 (54 Fed. Req. 23868). This rulemaking 

also clarified and reinforced EPA's existing regulations 

governing water quality-based permitting. The purpose of this 

memorandum is to describe the significance of these 

clarifications to EPA's baseline water quality-based permitting 

regulations. 

CHANGES TO 40 C.F.R. PART 122 

Section 122.44 covers the establishment of limitations, 

standards, and other permit conditions in NPDES permits. 

Subsection (d) covers water quality standards and state 

requirements. Prior to the promulgation of these new regulations 

the subsection was non-specific, requiring only that NPDES 

permits be issued with requirements more stringent than 

promulgated effluent guidelines as necessary to achieve water 

quality standards. We have strengthened considerably the 

requirements of §122.44(d). The new language is very specific 

and requires water quality-based permit limits for specific 

toxicants and whole effluent toxicity where necessary to achieve 

state water quality standards. The following is a section-bysection description of the new requirements. 

B-3-1 
2 

1. $122.44(d)(l)(i) 

This new paragraph provides that all pollutants that cause, 

have the reasonable potential to cause, or contribute to an 

excursion above a water quality standard must be controlled to 

achieve all applicable water quality standards, including 

narrative water quality criteria. We added this paragraph so 

that our regulations would reflect EPA's approach to water 

quality-based permitting. 

2. $122.44(d)(l)(ii) 

Subparagraph (ii) of the new regulations requires the states 

to use valid procedures to determine whether a discharge causes, 

has the reasonable potential to cause, or contributes to an 

excursion above a water quality standard. These procedures must 

account for existing controls on point and nonpoint sources of 

pollution, the variability of the pollutant in the effluent, the 

sensitivity of the test species (when evaluating whole effluent 

toxicity), and where allowed by state water quality standards, 

the dilution of the effluent in the receiving water. The purpose 

of this new regulation is to require the states to use 

technically valid procedures when determining whether a discharge 

is exceeding a numeric or narrative water quality criterion. 

When the permitting authority determines, using these procedures, 

that a discharge causes, has the reasonable potential to cause, 

or contributes to an excursion above a water quality criterion, 

that permit must include one or more water quality-based effluent 

limits established under subparagraphs (iii) - (vi). 

Subparagraphs (iii) and (iv) deal with water quality-based 

limitations where the state has numeric water quality criteria: 

subparagraphs (v) and (vi) deal with a state's narrative water 

quaity criteria. 

3. $122.44(d)(l)(iii) 

This paragraph requires NPDES permits to include effluent 

limitations for every individual pollutant that causes, has the 

reasonable potential to cause, or contributes to an excursion 

above a numeric water quality criterion. Thus, when a state has 

adopted a water quality criterion for an individual pollutant and 

the state determines under subparagraph (ii) that an effluent 

limit is necessary, subparagraph (iii) requires an effluent limit 

for that individual pollutant. 

4. 3122.44(d)(l)(iv) 

Subparagraph (iv) requires effluent limitations on whole 

effluent toxicity when a discharge is exceeding a state numeric 

criteria for whole effluent toxicity. This paragraph is applied 

B-3-2 
where a state has adopted a numeric criterion for whole effluent 

toxicity (e.g. a discharge must achieve an LC50 of 50% or 

higher). 

5. $122.44(d)(l)(v) 

When the state determines that a discharge exceeds a 

narrative water quality criterion, subparagraph (v) requires 

effluent limitations on whole effluent toxicity. If, however, 

chemical-specific effluent limitations are demonstrated to be 

sufficient to achieve all applicable water quality standards, 

then subparagraph (v) allows the permitting authority to forego a 

limitation on whole effluent toxicity. It may be necessary for 

you to work with an individual state to ensure that they have the 

necessary protocols to support whole effluent toxicity limits. 

6. $122.44(d)(l)(vi) 

Where an actual or projected excursion above a narrative 

water quality criterion is attributable to a particular 

pollutant, but the state has not adopted a water quality 

criterion for the pollutant of concern, this new regulation 

requires water quality-based effluent limitations which will 

control the pollutant of concern. Subparagraph (vi) establishes 

three options for developing such limitations. Under these 

options, a state may: 1) calculate a numeric criterion for the 

pollutant: 2) use EPA's water quality criterion for the pollutant 

of concern; or 3) establish effluent limits on an indicator 

parameter. 

By an indicator parameter we mean a pollutant or pollutant 

parameter for which control of this indicator will result in 

control of the pollutant of concern. For example, it may be 

shown that a more stringent control on total suspended solids 

will reduce discharge of a metal to a level which achieves the 

water quality standard. Subparagraph (vi) also sets out four 

provisions which must be met to allow the use of an indicator: 

1) The permit must identify which pollutants are intended 

to be controlled by a limit on the indicator parameter. 

2) The fact sheet must set forth the basis for the limit, 

including a finding that compliance with the limit will 

result in controls on the pollutant of concern that are 

sufficient to achieve the water quality standard. 

3) The permit must require all monitoring necessary to 

show continued compliance with water quality standards. 

0-3-3 
4 - 

4) The permit must contain a reopener clause allowing for 

changes in the permit as needed to achieve water 

quality standards. 

A state's narrative water quality criterion serves as the legal 

basis for establishing such effluent limits. 

7. #122.44(d)(l)(vii) 

Subparagraph (vii) requires that all water quality-based 

effluent limitations adhere to two fundamental principles: 1) the 

effluent limitations must be derived from and comply with all 

applicable water quality standards: and 2) the effluent 

limitations are consistent with the assumptions and requirements 

of an applicable wasteload allocation (WLA) if a WLA is available 

for the pollutant. 

CHANGES TO 40 C.F.R. PART 123 

We amended the permit objection regulations at 40 C.F.R. 

$123.44 to reflect the amendments to $122.44(d)(l). Under 

$123.44(c)(8) EPA may now object to a state-issued permit if the 

permit does not meet the requirements of $122.44(d)(l). Thus, if 

a state does not use technically sound procedures for evaluating 

the need for water quality-based effluent limitations then EPA 

may object to the permit. Also, if a state fails to include 

chemical-specific or whole effluent toxicity limitations in a 

permit as required by paragraphs (iii) - (vi), then EPA may 

object to the permit. Finally, if a water quality based effluent 

limitation is not derived according to the principles in 

subparagraph (vii) then EPA may object to the permit. 

If a state's surface water toxics control program is not 

adequate to implement these requirements, the new regulations at 

40 C.F.R. $123.63 expand EPA's criteria for withdrawing a state's 

NPDES program. Under the new regulations ($123.63(a)(S)), EPA 

may withdraw a state's NPDES program if the state fails to 

develop an adequate regulatory program for developing water 

quality-based effluent limitations. In November 1987, 

Headquarters provided procedural and technical guidance to the 

Regions on conducting state toxic8 control program reviews to 

assess the adequacy of state water quality-based control 

programs. This guidance sets guidelines for assessing whether or 

not a state's regulations, policies, and technical guidance 

constitute an adequate program. 

The significance of these additions to Part 123 is twofold. 

First, the Regions must issue permits which comply with these 

requirements and must work with the NPDES states to insure they 

also issue permits which comply with these regulations. If the 

states do not issue permits consistent with Part 123, the Region 

must veto insufficient permits and work with the states to 

B-34 
reissue the permits with water quality-based effluent limitations 

which achieve water quality standards. The specific requirements 

in $122.44(d) are structured in a way that implements EPA's 

Policy for the Development of Water Quality-Based Permit 

Limitations for Toxic Pollutants (49 Fed. Req. 9016 March 9, 

1984). Second, Regions will need to look closely at each state's 

surface water toxics control program to ensure that the state's 

regulations, policies and technical guidance result in the 

consistent and comprehensive development of NPDES permits which 

achieve the state's water quality standards. Where this does not 

occur, each Region should work with the state to rectify the 

problem and, after these negotiations and where necessary, 

investigate the possibility of withdrawing the NPDES program. 

I hope these regulations will assist you in developing water 

quality-based effluent limits and will support your efforts to 

implement surface water toxic8 control programs. If you have 

questions or need more information about these requirements, 

please contact Cynthia Dougherty at FTS 475-9545 or have your 

staff contact Rick Brandes at FTS 475-9537. 

cc: Permits Branch Chiefs, Regions I - X 

Martha Prothro, OWRS 

B-3-5 
23868 Federal Register / Vol. 54, No. 105 / Friday, June 2, 1989 / Rules and Regulations 

ENVIRONMENTAL PROTECTION 

AGENCY 

[FRL-35574 

40 CFR plrtr 122,123 d 130 

Nstlonal Pollutant Dkchrr@ 

Ellminatlon System; Sum Wdar 

Toxic8 Control Program 

AOENCY: JZnviranmentaI Protection 

Agency. 

ACTION Final Rule. 

and Standards, (WH-553), U.S. 

Environmental Protection Agency, 4.01 M 

Street SW., Washington, DC 2OM0, (202) 

382-7056. The Public record for this 

regulation ie available at the EPA 

library, M~~KM, U.S. Environmental 

Protection Agency, 401 M Street SW., 

Washington, DC 20480. 

SUMMARY: Today’s action amends Parts 

122,123. and 130 of EPA’s regulations. 

The regulations clarify EPA’s surface 

water toxica control program, and 

incorporate section 308(a) of the Water 

Quality Act of 1987 into EPA’s toxics 

control program. Section 308[a) of the 

Water Quality Act added section 304(l) 

to the Clean Water Act [hereafter 

referred to as section 304(l)). Section 

304(l) requires the states to identify 

those waters that are adversely affected 

by toxic, conventional, and 

nonconventional pollutants, and 

requires the states to prepare individual 

control strategies that will control point 

source discharges of toxic pollutants. 

The states must submit lists of waters 

and individual control strategies to EPA 

for review, and if EPA disapproves a 

state’s decision with respect to a list or 

an individual control strategy, then EPA 

must implement the requirements of 

section 304(l) in cooperation with the 

state. EPA and the states must 

accomplish the tasks in section 304(l) 

according to an ambitious series of 

deadlines. Today’s regulations will 

strengthen State and Federal controls 

over discharges to toxic pollutants, and 

will assist EPA and the states in 

satisfying the requirements of section 

304(l) of the CWA. 

EFFECTIVE DATE These regulations shall 

be effective on May 243,1989 at 1:fXl p.m. 

Eastern Daylight Savings Time. In 

accordance with 40 CFR 23.2, EPA 

hereby specifies that these regulations 

shall be considered final agency action 

for purposes of judicial review at 1:OO 

p.m. Eastern Daylight Savings Time on 

May 26.1989. 

FOR FURTHER INFOAYATION COWACE 

Paul Connor, Program Development 

Branch, Office of Water Enforcement 

and Permits, @N-338), US. 

Environmental Protection Agency, 401 M 

Street, SW., Washington, DC 20480, 

(202) 475-9537, or Judith Leckrone. 

Assessment and Watershed Protection 

Division, Office of Water Regulations 

B-3-6 
Faded RQW- / Vol. 5% No. 106 / Friday, June Z,~SSS / Rules and Reguiatiorm 2x95 

PART In--EPA ADMINISTERED 

Pf%MlT PfWGRAMS THE NATIONAL 

1. The authority citation for Part 122 

continuer to read au follows: 

AutbaitY: Tba Clean Water Act. 33 USC. 

1251 et req. 

2. Section 1222 is amended by adding 

inunw~betical order a new definition aa 

. 

om.2 D@fhhma 

l l l l t 

Whole effruent toxicity mean0 the 

aggregate toxk effect of an effluent 

measured directly by a toxicity test. 

3. Paragraph [d)(l) of 0 IZZ.M is 

revised to read as follows: 

8w2U Ed@bkhgIlmltr(knr, 

rtmdudr,~-P.mlto#lditkn 

~ww&--msproomy8e. 

. l l . l 

(4 . l . 

(11 Achieve wrter quality standards 

established under rection 303 of the 

CWA including State narrative criteria 

for water qualitv. . 

(I) Limitition~ must control all 

pollutants or pollutant parameters 

(either conventional. nonconventional 

or toxic pollutants) which the Director’ 

determines are or may be discharged at 

a level which will cause, have the 

reasonable potential to catrae. or 

contribute to an excursion above any 

State water quality standard, including 

State narrative criteria for water qualtty. 

(ii) When determining whether a 

discharge causes, has the reasonable 

potential to cause, or contributes to an 

in-&ream excursion above a narrative or 

numeric criterie within a State water 

quality standard. the permitting 

authority shell use procedures which 

account for existing controle on point 

and nonpoint sources of pollution, the 

variability of the pollutant or pollutant 

parameter in the effluent, the Beneitivity 

of the species to toxicity testing (when 

evaluating whole effluent toxicity), and 

where appropriate, the dilution of the 

effluent in the receiving water. 

(iii) When tbe permitttng authority 

determines, uain~ the procedures in 

paragraph (d)(l)(ii) of this section. thst a 

discharge causes, hae the reasonable 

potential to cause, or contributes to an 

in&ream excursion above the allowable 

ambient concentration of a State 

numeric criteria witbin a State water 

quality standard for an individual 

pollutant, the permit must contain 

effluent limits for that pollutant. 

(iv) When the permitting authority 

determines; wingt the prodedurea in 

paregraph (d)(l)(ii) of this section, that a 

discharge causea, haa the reasonable 

potential to cause, or contributes to an 

in-dream excursion above the numeric 

criterion for whole effluent toxicity, the 

permit must contain effluent limit8 for 

whole effluent toxicity. 

Iv1 ¢ aa Drovided in thie 

rubpamgr~ph, Ghen the permitting 

authority deterrntnea, using the 

procedurea in paragraph (d)(l)(ii) of this 

recttoa toxicity testing data, or other 

information, tbat a discharge causes, has 

the reasonable potential to catme, or 

contributes to an in&ream excursion 

above a narrative criterion within an 

applicable State water quality standard, 

the permit must contain effluent hmita 

for whole effluent toxicity. Limits on 

whole effluent toxicity are not necessary 

where tbe permitting authority 

demonstrates in the fact sheet or 

statement of basis of the NPDES permit, 

using the procedure8 in paragraph 

(d)(l)@) of this section, that chemicalspecific limita for the effluent are 

sufficient to attain and maintain 

applicable numeric and narrative State 

water quality etandarda. 

(vi) Where a State has not established 

a water quality criterion for a specific 

chemical pollutant that ia present in an 

effluent at a concentretton that cnuae8, 

has the reaeonable potential to catme, or 

contributes to an excursion above a 

narrative criterion within an applicable 

State water quality standard, the 

0-3-7 

permitting authority muat establish 

effluent hmits ueing one or more of the 

following optione: 

(A) Establteh effluent limita ustng a 

calculated numeric water quality 

criterion for the pollutant which the 

permitting authority demonstrates will 

attain and maintain applicable narrative 

water quality criteria and will fully 

protect the designated use. Such a 

criterion may be derived wing a 

proposed State criterion, or an explicit 

State policy or regulation interpreting its 

narrative water quality criterion, 

supplemented with other relevant 

information which may include: EPA’s 

Water Quality Standards Handbook, 

October ‘166% risk assessment data, 

exposure data, information about the 

pollutant from the Food and Drug 

Administration, and current EPA criteria 

documents; or 

(B) Establish effluent limits on a caaeby-case basis, using EPA’s water quality 

criteria, published under section m(a) 

of the CWA, supplemented where 

necessary by other relevant information; 

or 

(C) Establish effluent limitatione on an 

indicator parameter for the pollutant of 

concern, provided: 

(1) Tbe permit identifies which 

pollutanta are intended to be controlled 

by the use of the effluent limitation; 

(2) The fact sheet required by P 1~~6 

sets forth the basis for the limit, 

including a finding that compliance with 

the effluent limit on the indicator 

parameter will result in controls on the 

pollutant of concern which are sufficient 

to attain and maintain applicable water 

quality standards; 

(3) The permit requires ill effluent and 

ambient monitoring necessary to show 

that during the term of the permit the 

limit on the indicator parameter 

continuea to attain and maintain 

applicable water quality standards; and 

(4) The permit contains a reopener 

clause allowing tbe permitting authority 

to modify or revoke and reissue the 

Permit if the limit8 on the indicator 

parameter no longer attain and maintain 

applicable water quality standards. 

(vii) When developing water qualitybased effluent limits under this 

EEqh$e permitting authority shall 

(A) The level of water quality to be 

achieved by limite on point sources 

established under this paragraph ie 

derived from, and complies with all 

applicable water quality standards; and 

(B) Effluent bmtte developed to 

protect a narrative water quality 

criterion, a numeric water quality 

criterion, or both, ere consistent with the 

aesumptiona and requirements of any 

available wasteload allocation for the 
Feded Register / Vol. 64, No. 105 / Friday, June 2. 1989 / Rules and Regulations 

discharge prepared by the Stete and 

approved by EPA pursuant to 40 CFR 

130.7. 

. l . . l 

4. The title of paragraph (eB] of 8 12244 

is revised to read as follows: 

. t . . . 

(e) Technology-based controls for 

toxic pollutants. l l l 

PART 123-STAlE PROORAM 

REOWREYENTS 

1. The authority citation for Part 123 

continues to read aa follows: 

Authority: Clean Water Act, 33 U.S.C. 1251 

et seq. 

2. Section 123.4 ie amended by 

adding paregraph [C)(S) to read as 

follows: 

0122.44 EPAmfhwofmldoqecthsto 

stat. pBmtl8. 

l + . . l 

# ;ha kffluent limits of a permit fail 

to satisfy the requirement8 of 40 CFR 

122.&(d). 

. . . t l 

3. In 0 123.46 paragraph (a) is revised 

and paragraphs (cl, (d). (e) and (fj are 

added. as follows: 

0 122A6 IndlvmJsl clbbor 8tratoglaa 

(a) Not later than February 4,1989. 

each State shall submit to the Regional 

Administrator for review, approval, and 

implementation an individual control 

strategy for each point source identified 

by the State pursuant to section 

304(1)(1)(C) of the Act which will 

produce a reduction in the dkharge of 

toxic pollutants from the point source8 

identified under section 304(l)(l)(C) 

through the establishment of effluent 

limi ta tione under section 402 of the 

CWA and water quality standards 

under section 303(c)(2)(B) of the CWA, 

which reduction is sufficient, in 

combination with existing controla on 

point and nonpoint sources of pollution, 

to achieve the applicable water quality 

standard as soon as possible. but not 

later than three years after the date of 

the establishment of such strategy. 

l l . . . 

(cl For the purposes of this section the 

term individual control strategy, as set 

forth in section 304(l) of the CWA. 

means a final NPDES permit with 

supporting document& on showing that 

effluent limits acs consistent with an 

approved wasteload allocatfon, or other 

documentation which shows that 

applicable water quality standards wlll 

be met not later than three years after 

the individual control strategy ia 

established. Where a State is unable to 

issue a final permit on or before 

Febmary ~,ISICI, au individual kntrol 

strategy may be a draft per&t with an 

attached schedule (provided the State 

meets the schedule for iesu@ the final 

permit) irdlcating that the permit will be 

issued on or before February 4.1990. If a 

point source is eubject to section 

304(I)(l)(C) of the CWA and is also 

subject to an on-eite response action 

under section8 101 or 108 of the 

Comprehensive Environmental 

Re8pon8e. Compensation, and Liability 

Act of 1980 (CERCLA), (42 U.S.C. 8801 et 

seq.), an individual control strategy may 

be the decision document (which 

incorporates the applicable or relevant 

and appropriate requirements under the 

CWA) prepared under sections 104 or 

106 of CERCLA to addree8 the release or 

threatened release of hazardow 

substances to the environment. 

(d] A petition submitted pursuant to 

section 304(l)(3) of the CWA must be 

submitted to the appropriate Regional 

AdminMrator. Petitions must identify a 

waterbody in sufficient detail so that 

EPA is able to determine the location 

and boundaries of the waterbody. The 

petition must also identify the list or 

lists for which the waterbody qualifiis. 

and the petition must explain why the 

waterbody satisfies the criteria for 

listing under CWA section X&%(l) and 40 

CFR 13mO(d)(8). 

(e) If the Regional Administrator 

disapproves one or more individual 

control strategies, or if e State fails to 

provide edequate public notice and an 

opportunity to comment on the ICSs. 

then, not later than June 4.lseS. the 

Regional Administrator shall give a 

notice of approval or disapproval of the 

individual control strategies submitted 

by each State pursuant to this section es 

follows: 

(11 The notice of approval or 

dlsapproval given under this paragraph 

shall include the following: 

(i) The name and address of the EPA 

office that reviews the State’8 

submittals. 

(ii) A brief description of the section 

304(l) process. 

(Iii) A list of ICSs disapproved under 

this section and a finding that the ICES 

will not meet all applicable review 

criteria under this section and section 

304(l) of tha CWA. 

(iv) If the Regional Administrator 

determines that a State did not provide 

adequate pubklc notice and an 

opportunity to cqmment on the waters, 

point m or ICSs prepared pursuant 

to section 304(I). or if the Regional 

Administrator choosea to exercise his or 

her discr8tioa a lbt of the ICS8 

approved under this section, and a 

B-3-8 

5nding that the ICSs satisfy all 

applicable review rsiteria. 

iv) The location where interested 

per---y -8 WA’S RCOrdS Of 

approval end disepprvvaL 

(vi) The name. address, and telephone 

number of the person at the Regional 

Office from whom interested persons 

may obtain more informaff on. 

(vii) Notice that written petitions or 

comments are due within 120 days. 

(2) The Regional Administrator shall 

provide the notice of approval or 

disapproval given under this paragraph 

to the eppropriate State Director. The 

Regional Administrator ahalI publish a 

notice of availability, ln a daily or 

weekly newspaper with State-wide 

circulation or In the Fetied Register, for 

the no&e of approval or disapproval. 

The Regional Administrator shall also 

provide written notice to each 

discharger identified under section 

304(1)(1)(C), that EPA has listed the 

discharger under section 3w(l)(l](C). 

(3) A8 soon a8 practicable but not 

later than June 4.lSaa the Regional 

Offices shall issue a response to 

pet&ions or comments received under 

section 304(f). The response to 

comments shall be given in the same 

manner as the notice described in 

paragraph (e) of this section except for 

the following changes: 

(i) The liata of ICSs reflecting any 

changes made pursuant to comments or 

petitions received. 

(ii) A brief description of tbe 

subsequent steps in the section ~~(11 

process. 

(fj EPA 8haIl review, and approve or 

disapprove, the individual control 

strategies prepared under section 304(l) 

of the CWA, using the applicable 

criteria set forth in saction 304(l) of the 

CWA, and in 40 CFR Part 122, including 

P 12214(d). At any time after the 

Regional Administrator disapproves an 

ICS (or conditionally aproves a draft 

permit as an ES). the Regional Office 

may submit a written notification to the 

State that the Regional Office intends to 

issue the KS. Upon mailing the 

notification, and notwithstanding any 

other reguIation. exclusive authority to 

issue the permit passes to EPA. 

4. Section 123.63 is amended by 

eddlng paragraph (a)(5) to read as 

followa: 

(a) ’ l l 

(51 Where the State fail8 to develop an 

adequate regulatory program for 

developing watsr qua&y-based effluent 

lmts in NPDEspermit8. 
APPENDIX B-4 

WHOLE EFFLUENT TOXICITY PERMITTING PRINCIPLES 

AND ENFORCEMENT STRATEGY 
UNITED STATES ENVIRONMENTAL PROTECTION AGENCY 

WASHINGTON. D.C. 20460 

January 25, 1989 

OFFICE OF 

WATER 

MEMORANDUM 

SUBJECT: 

Enforcement Strategy 

Whole Effluent Toxicity Basic Permitting Principles and 

FROM: Rebecca W. Hanmer, Acting Assistant Administrator 

Office of Water 

TO: Regional Administrators 

Since the issuance of the "Policy for the Development of 

Water Quality-based Permit Limitations for Toxic Pollutants" in 

March of 1984, the Agency has been moving forward to provide 

technical documentation to support the integrated approach of 

using both chemical and biological methods to ensure the 

protection of water quality. The Technical Support Document for 

Water Quality-based Toxics Control [September, 1985) and the 

Permit Writer's Guide to Water Quality-based Permitting for Toxic 

Pollutants (July, 1987) have been instrumental in the initial 

implementation of the Policy. The Policy and supporting 

documents, however, did not result in consistent approaches to 

permitting and enforcement of toxicity controls nationally. When 

the 1984 Policy was issued, the Agency did not have a great deal 

of experience in the use of whole effluent toxicity limitations 

and testing to ensure protection of water quality. We now have 

more than four years of experience and are ready to effectively 

use this experience in order to improve national consistency in 

permitting and enforcement. 

In order to increase consistency in water quality-based 

toxicity permitting, I am issuing the attached Basic Permitting 

Principles for Whole Effluent Toxicity (Attachment 1) as a 

standard with which water quality-based permits should conform. 

A Workgroup of Regional and State permitting, enforcement, and 

legal representatives developed these minimum acceptable 

requirements for toxicity permitting based upon national 

experience. These principles are consistent with the toxics 

control approach addressed in the proposed Section 304(1) 

regulation. Regions should use these principles when reviewing 

draft State permits. If the final Section 304(1) regulations 

include changes in this area, we will update these principles as 

necessary. Expanded guidance on the use of these principles will 

be sent out shortly by James Elder, Director of the Office Of 

B-4-1 
-2- 

Water Enforcement and Permits. This expanded guidance will 

include sample permit language and permitting/enforcement 

scenarios. 

Concurrent with this issuance of the Basic Permitting 

Principles, I am issuing the Compliance Monitoring and 

Enforcement Strategy for Toxics Control (Attachment 2). This 

Strategy was developed by a Workgroup of Regional and State 

enforcement representatives and has undergone an extensive 

comment period. The Strategy presents the Agency's position on 

the integration of toxicity control into the existing National 

Pollutant Discharge Elimination System (NPDES) compliance and 

enforcement program. It delineates the responsibilities of the 

permitted community and the regulatory authority. The Strategy 

describes our current efforts in compliance tracking and quality 

assurance of self-monitoring data from the permittees. It 

defines criteria for review and reporting of toxicity violations 

and describes the types of enforcement options available for the 

resolution of permit violations. 

In order to assist you in the management of whole effluent 

toxicity permitting, the items discussed above will join the 1984 

Policy as Appendices to the revised Technical Support Document 

for Water Quality-based Toxics Control. To summarize, these 

materials are the Basic Permitting Principles, sample permit 

language, the concepts illustrated through the permitting and 

enforcement scenarios, and the Enforcement Strategy. I hope 

these additions will provide the needed framework to integrate 

the control of toxicity into the overall NPDES permitting 

program. 

I encourage you and your staff to discuss these documents 

and the 1984 Policy with your States to further their efforts in 

the implementation of EPA's toxics control initiative. 

If you have any questions on the attached materials, please 

contact James Elder, Director of the Office of Water Enforcement 

and Permits, at (FTS/202) 475-8488. 

Attachments 

cc: ASWIPCA 

Water Management Division Directors 

B-4-2 
EASIC PERMITTING PRINCIPLES FOR WHOLE EFFLUENT TOXICITY 

1. Permits must be protective of water quality. 

a. At a minimum, all major permits and minors of 

concern must be evaluated for potential or known 

toxicity (chronic or acute if more limiting). 

b. Final whole effluent toxicity limits must be 

included in permits where necessary to ensure 

that State Water Quality Standards are met. 

These limits must properly account for effluent 

variability, available dilution, and species 

sensitivity. 

2. Permits must be written to avoid ambiguity and ensure 

enforceability. 

a. Whole effluent toxicity limits must appear in Part I 

of the permit with other effluent limitations. 

b. Permits contain generic re-opener clauses which 

are sufficient to provide permitting authorities 

the means to re-open, modify, or reissue the 

permit where necessary. Re-opener clauses covering 

effluent toxicity will not be included in the 

Special Conditions section of the permit where 

they imply that limit revision will occur based 

on permittee inability to meet the limit. Only 

schedules or other special requirements will be 

added to the permit. 

C. If the permit includes provisions to increase 

monitoring frequency subsequent to a violation, it 

must be clear that the additional tests only determine the continued compliance status with the limit; 

they are not to verify the original test results. 

d. Toxicity testing species and protocols will be 

accurately referenced/cited in the permit. 

3. Where not in compliance with a whole effluent toxicity 

limit, permittees must be compelled to come into compliance 

with the limit as soon as possible. 

a. Compliance dates must be specified. 

b. Permits can contain requirements for corrective 

actions, such as Toxicity Reduction Evaluations 

(TRW, but corrective actions cannot be delayed 

pending EPA/State approval of a plan for the 

corrective actions, unless State regulations 

require prior approval. Automatic corrective 

actions subsequent to the effective date of a final 

whole-effluent toxicity limit will not be included 

in the permit. 

B-4-3 
ATTACHMENT 1 

Explanation of the Basic Permitting Principles 

The Basic Permitting Principles present the minimum 

acceptable requirements for whole-effluent toxicity permitting. 

They begin with a statement of the goal of whole-effluent 

toxicity limitations and requirements: the protection of water 

quality as established through State numeric and narrative Water 

Quality Standards. The first principle builds on the Technical 

Support Document procedures and the draft Section 304(l) rule 

requirements for determining potential to violate Water Quality 

Standards. It requires the same factors be considered in setting 

whole-effluent toxicity based permits limits as are used to 

determine potential Water Quality Standards violations. It 

defines the universe of permittees that should be evaluated for 

potential violation of Water Quality Stand-ids, L.Ad the,tfore 

possible whole-effluent limits, as all majors and minors of 

concern. 

The second permitting principle provides basic guidelines 

for avoiding ambiguities that may surface in permits. Wholeeffluent toxicity limits should be listed in Part I of the permit 

and should be derived and expressed in the same manner as any 

other water quality-based limitations (i.e., Maximum Daily and 

Average Monthly limits as required by Section 122.45(d)). 

In addition, special re-opener clauses are generally not 

necessary, and may mistakenly imply that permits may be re-opened 

to revise whole-effluent limits that are violated. This is not 

to imply that special re-opener clauses are never appropriate. 

They may be appropriate in permits issued to facilities that 

currently have no known potential to violate a Water Quality 

Standard; in these cases, the permitting authority may wish to 

stress its authority to re-open the permit to add a wholeeffluent limit in the event monitoring detects toxicity. 

Several permittees have mistakenly proposed to conduct 

additional monitoring subsequent to a violation to "verify" their 

results. It is not possible to verify results with a subsequent 

test whether a new sample or a split-sample which has been stored 

(and therefore contains fewer volatiles) is used. For this 

reason, any additional monitoring required in response to a 

violation must be clearly identified as establishing continuing 

compliance status, not verification of the original violation. 
- 2- 

The second principle also deals with the specification of 

test species and protocol. Clearly setting out the requirements 

for toxicity testing and analysis is best done by accurately 

referencing EPA's most recent test methods and approved 

equivalent State methods. In this way, requirements which have 

been published can be required in full, and further advances in 

technology and science may be incorporated without lengthy permit 

revisions. 

The third and final permitting principle reinforces the 

responsibility of the permittee to seek timely compliance with 

the requirements of its NPDES permit. Once corrective actions 

have been identified in a TRE, permittees cannot be allowed to 

delay corrective actions necessary to comply with water qualitybased whole effluent toxicity limitations pending Agency review 

and approval of voluminous reports or plans. Any delay on the 

part of the permittee or its contractors/agents is the 

responsibility of the permittee. 

The final principle was written in recognition of the fact 

that a full-blown TRE may not be necessary to return a permittee 

to compliance in all cases, particularly subsequent to an initial 

TRF,. As a permittee gains experience and knowledge of the 

operational influences on toxicity, TREs will become less 

important in the day to day control of toxicity and will only be 

required when necessary on a case-specific basis. 

B-4-5 
ATTACHMENT 2 

to the Cwce Watorba UEniorcm 

StratecnriQlc,hToxicsControl 

The me Modtormnt Strateuv for 

Contrd sets forth the Agency's strategy for tracking 

compliance with and enforcing whole-effluent toxicity monitoring 

requirements, limitations, schedules and reporting requirements. 

The Strategy delineates the respective responsibilities of 

permittees and permitting authorities to protect water quality 

through the control of whole-effluent toxicity. It establishes 

criteria for the review of compliance data and the quarterly 

reporting of violations to Headquarters and the public. The 

Strategy discusses the integration of whole-effluent toxicity 

control into our existing.fnspection and quality assurance 

efforts. It provides guidelines on the enforcement of wholeeffluent toxicity reguirements. 

The Strategy also addresses the concern many permittees 

share as they face the prospect of new requirements in their 

permit - the fear of indiscriminate penalty assessment for 

violations that they are unable to control. The Strategy 

recognizes enforcement discretion as a means of dealing fairly 

with permittees that are doing everything feasible to protect 

water quality. As indicated in the Strategy, this discretion 

deals solely with the assessment of civil penalties, however, and 

is not an alternative to existing procedures for establishing 

relief from State Water Quality Standards. The Strategy focuses 

on the responsibility of the Agency and authorized States to 

require compliance with Water Quality Standards and thereby 

ensure protection of existing water resources. 

B-4-6 
01/19/89 

COMPLIANCE MONITORING AND ENFORCEMENT STRATEGY 

FOR TOXICS CONTROL 

I. Background 

Issuance of NPDES permits now emphasizes the control of toxic 

pollutants, by integrating technology and water quality-based 

permit limitations, best management practices for toxic discharges, 

sludge requirements, and revisions to the pretreatment implementation requirements. These requirements affect all major permittees 

and those minor permittees whose discharges may contribute to 

impairment of the designated use for the rec'eiving stream. The 

goal of permitting is to eliminate toxicity in receiving waters 

that results from industrial and municipal discharges. 

Major industrial and municipal permits will routinely contain 

water quality-based limits for toxic pollutants and in many cases 

whole effluent toxicity derived from numerical and narrative 

water quality standards. The quality standards to establish NPDBS 

permit limits are discussed in the "Policy for the Development of 

Water Quality-based Permit Limits for Toxic Pollutants," 49FR 9016, 

March 9, 1984. The Technical Support Document for Water Qualitybased Toxics Control, EPA X440/44-85032 September, 1985 and the 

.Permit Writer's Guide to Water Quality-Gased Permitting for Toxic 

Pollutants, Office of Water, May, 1987, provide guidance for interpreting numerical and narrative standards and developing permit 

limits. 

The Water Quality Act (WQA) of 1987 (PL 100-4, February 4, 

1987) further directs EPA and the States to identify waters that 

require controls for toxic pollutants and develop individual 

control strategies including permit limits to achieve control of 

toxics. The WQA established deadlines, for individual control 

strategies (February 4, 1989) and for compliance with the toxic 

control permit requirements (February 4, 1992). This Strategy 

will support the additional compliance monitoring, tracking, evaluation, and enforcement of the whole effluent toxicity controls 

that will be needed to meet the requirements of the WQA and EPA's 

policy for water quality-based permitting. 

It is the goal of the Strategy to assure compliance with 

permit toxicity limits and conditions through compliance inspections, compliance reviews, and enforcement. Water quality-based 

limits may include both chemical specific and whole effluent toxicity limits. Previous enforcement guidance (e.g., Enforcement 

Management System for the National Pollutant Discharge Elimination 

System, September, 1986: National Guidance for Oversight of NPDES 

Programs, May, 1987: Guidance for Preparation of Quarterly and 

Semi-Annual Noncompliance Reports, March, 1986) has dealt with 

0-4-7 
-2- 

chemical-specific water quality-based limits. This Strategy will 

focus on whole effluent toxicity limits. Such toxicity limits may 

appear in permits, administrative orders, or judicial orders. 

II. Strategy Principles 

This 

1) 

2) 

3) 

4) 

strategy is based on four principles: 

Permittees are responsible for attaining, monitoring, 

and maintaining permit compliance and for the quality 

of their data. 

Regulators will evaluate self-monitoring data quality 

to ensure program integrity. 

Regulators will assess compliance through inspections, 

audits, discharger data reviews, and other independent 

monitoring or review activities. 

Regulators will enforce effluent limits and compliance 

schedules to eliminate toxicity. 

III. Primary Implementation Activities 

In order to implement this Strategy fully, the following 

activities are being initiated: 

A. Immediate development 

1. The NPDES Compliance Inspection Manual was 

revised in May 1988 to include procedures for 

performing chronic toxicity tests and evaluating 

toxicity reduction evaluations. An inspector 

training module was also developed in August 

1988 to support inspections for whole effluent 

toxicity. 

2. The Permit Compliance System (the national NPDES 

data base) was modified to allow inclusion 

of toxicity limitations and compliance schedules 

associated with toxicity reduction evaluations. 

The PCS Steering Committee will review standard 

data elements and determine if further modifications are necessary. 

3. Compliance review factors (e.g., Technical 

Review Criteria and significant noncompliance 

definitions) are being proposed to evaluate 

violations and appropriate response. 

4. A Quality Assurance Fact Sheet has been developed 

(Attached) to review the quality of toxicity test 

results submitted by permittees. 

04-a 
-3- 

5. The Enforcement Response Guide in the Enforcement 

Management System will be revised to cover the use 

of administrative penalties and other responses to 

violations of toxicity controls in permits. At 

least four types of permit conditions are being 

examined: (1) whole-effluent toxicity monitoring 

(sampling and analysis), (2) whole effluent 

toxicity-based permit limits, (3) schedules to 

conduct a TRE and achieve compliance with water 

quality-based limits, and (4) reporting requirments. 

B. Begin development in Spring 1989 

With the assistance of the Office of Enforcement and 

Compliance Monitoring (OECM), special remedies and model forms 

will be developed to address violations of toxicity permit 

limits (i.e., model consent decrees, model complaints, revised 

penalty policy, model litigation reports, etc.) 

IV. Scope and Implementation of Strategy 

A. Compliance Tracking and Review 

1. Compliance Tracking 

The Permits Compliance System (PCS) will be 

used as the primary system for tracking limits and 

monitoring compliance with the conditions in NPDES 

permits. Many new codes for toxicity testing have 

already been entered into PCS. During FY 89, headquarters will provide additional guidance to Regions 

and States on PCS coding to update existing documentation. The Water Enforcement Data Base (WENDB) 

requirements as described in the PCS Policy Statement 

already require States and Regions to begin 

incorporating toxicity limits and monitoring information 

into PCS. 

In addition to guidance on the use of PCS, 

Headquarters has prepared guidance in the form 

of Basic Permitting Principles for Regions and 

States that will provide greater uniformity 

nationally on approaches to toxicity permitting. 

One of the major problems in the tracking and 

enforcement of toxicity limits is that they differ 

greatly from State-to-State and Region-to-Region. 

The Permits Division and Enforcement Division in 

cooperation with the PCS Steering Committee will 

establish standard codes for permit limits and 

procedures for reporting toxicity results based on 

this guidance. 

B-4-9 
-4- 

Whole effluent toxicity self-monitoring data 

should undergo an appropriate quality review. (See 

attached checklist for suggested toxicity review 

factors.) All violations of permit limits for 

toxics control should be reviewed by a professional 

qualified to assess the noncompliance. Regions and 

States should designate appropriate staff. 

2. Compliance Review 

Any violation of a whole effluent toxicity 

limit is of concern to the regulatory agency and 

should receive an immediate professional review. 

In terms of the Enforcement Management System (EMS), 

any whole effluent violation will have a violation 

review action criterion (WAC) of 1.0. However, the 

appropriate initial enforcement response may be to 

require additional monitoring and then rapidly 

escalate the response to formal enforcement if the 

noncompliance persists. Where whole effluent 

toxicity is based on a pass-fail permit limitation, 

any failure should be immediately targeted for 

compliance inspection. In some instances, assessment 

of the compliance status will be required through 

issuance of Section 308 letters and 309(a) orders to 

require further toxicity testing. 

Monitoring data which is submitted to fulfill 

a toxicity monitoring requirement in permits that do 

not contain an independently enforceable whole-effluent 

toxicity limitation should also receive immediate 

professional review. 

The burden for testing and biomonitoring is on 

the permittee; however, in some instances, Regions and 

States may choose to respond to violations through 

sampling or performance audit inspections. When an 

inspection conducted in response to a violation identifies noncompliance, the Region or State should 

initiate a formal enforcement action with a compliance 

schedule, unless remedial action is already required 

in the permit. 

B. Inspections 

EPA/State compliance inspections of all major permittees 

on an annual basis will be maintained. For all facilities 

with water quality-based toxic limits, such inspections should 

include an appropriate toxic component (numerical and/or 

whole effluent review). Overall the NPDES inspection and 

data quality activities for toxics control should receive 

greater emphasis than in the present inspection strategy. 

B-4-10 
-5- 

1. Regional/State Capability 

The EPA’s "Policy for the Development of Water 

Quality-based Permit Limits for Toxic Pollutants" 

(March 9, 1984 Federal Register) states that EPA 

Regional Administrators will assure that each 

Region has the full capability to conduct water 

quality assessments using both biological and chemical methods and provide technical assistance to the 

States. Such capability should also be maintained 

for compliance biomonitoring inspections and toxics 

sampling inspections. This capability should include 

both inspection and laboratory capability. 

2. Use of Nonsampling Inspections 

Nonsampling inspections as either compliance 

evaluations (CEIs) or performance audits (PAIs) can 

be used to assess permittee self-monitoring data 

involving whole effluent toxicity limits, TREs, and 

for prioritization of sampling inspections.* As 

resources permit, PAIs should be Used to verify 

biomonitoring capabilities of permittees and 

contractors that provide toxicity testing selfmonitoring data. 

3. Quality Assurance 

All States are encouraged to develop the 

capability for acute and chronic toxicity tests 

with at least one fish and one invertebrate species 

for freshwater and saltwater if appropriate. NPDES 

States should develop the full capability to assess 

compliance with the permit conditions they establish. 

EPA and NPDES States will assess permittee 

data quality and require that permittees develop 

quality assurance plans. Quality assurance plans 

must be available for examination. The plan should 

include methods and procedures for toxicity testing 

and chemical analysis: collection, culture, maintenance, and disease control procedures for test 

organisms; and quality assurance practices. The 

* Due to resource considerations, it is expected that sampling 

inspections will be limited to Regional/State priorities in 

enforcement and permitting. Routine use of CEIs and PAIs should 

provide the required coverage. 

B-4-11 
-6- 

permittee should also have available quality control 

charts, calibration records, raw test data, and 

culture records. 

In conjunction with the QA plans, EPA will 

evaluate permittee laboratory performance on EPA 

and/or State approved methods. This evaluation is 

an essential part of the laboratory audit process. 

EPA will rely on inspections and other quality 

assurance measures to maintain data quality. However, 

States may prefer to implement a laboratory certification program consistent with their regulatory 

authorities. Predetermined limits of data acceptability will need to be established for each test 

condition (acute/chronic), species-by-species. 

c. Toxicity Reduction Evaluations (TREs) 

TREs are systematic investigations required of permittees 

which combine whole effluent and/or chemical specific testing 

for toxicity identification and characterization in a planned 

sequence to expeditiously locate the source(s) of toxicity and 

evaluate the effectiveness of pollution control actions and/or 

inplant modifications toward attaining compliance with a permit 

limit. The requirement for a TRE is usually based on a 

finding of whole effluent toxicity as defined in the permit. 

A plan with an implementation schedule is then developed to 

achieve compliance. Investigative approaches include 

causative agent identification and toxicity treatability. 

1. Requiring TRE Plans 

TRE's can be triggered: 1) whenever there is a 

violation of a toxicity limit that prompts enforcement 

action or 2) from a permit condition that calls for a 

toxicity elimination plan within a specified time 

whenever toxicity is found. The enforcement action 

such as a 309(a) administrative order or State 

equivalent, or judicial action then directs the 

permittee to take prescribed steps according to a 

compliance schedule to eliminate the toxicity. This 

schedule should be incorporated into the permit, an 

administrative order, or judicial order and compliance 

with the schedule should be tracked through PCS. 

2. Compliance Determination Followup 

Compliance status must be assessed following the 

accomplishment of a TRE plan using the most efficient and effective methods available. These methods 

include site visits, self-monitoring, and inspections. 

B-4-12 
-7- 

Careful attention to quality assurance will assist in 

minimizing the regulatory burden. The method of 

compliance assessment should be determined on a 

case-by-case basis. 

D. Enforcing Toxic Control Permit Conditions 

Enforcement of toxic controls in permits depends upon a 

clear requirement and the process to resolve the noncompliance. In addition to directly enforceable whole effluent 

limits (acute and chronic, including absolute pass-fail 

limits), permits have contained several other types of 

toxic control conditions: 1) "free from" provisions, 

2) schedules to initiate corrective actions (such as TREs) 

when toxicity is present, and/or 3) schedules to achieve 

compliance where a limit is not currently attained. 

Additional requirements or schedules may be developed 

through 308 letters, but the specific milestones should be 

incorporated into the permit, administrative order or 

State equivalent mechanism, or judicial order to ensure 

they are enforceable. 

1. The Quarterly Noncompliance Report (QNCR) 

Violations of permit conditions are tracked and 

reported as follows: 

a. Effluent Violations 

Each exceedance of a directly enforceable whole 

effluent toxicity limit is of concern to the 

regulatory agency and, therefore, qualifies 

as meeting the VRAC requiring professional 

review (see section IV.A.2.). 

These violations must be reported on the QNCR 

if the violation is determined through professional review to have the potential to have 

caused a water quality impact. 

All QNCR-reportable permit effluent violations 

are considered significant noncompliance (SNC). 

b. Schedule Violations 

Compliance schedules to meet new toxic controls 

should be expeditious. Milestones should be 

established to evaluate progress routinely and 

minimize delays. These milestones should be 

tracked and any slippage of 90 days or more 

must be reported on the QNCR. 

B-4-13 
-8- 

The following milestones are considered SNC when 

90 days or more overdue: submit plan/schedule 

to conduct TRE, initiate TRE, submit test results, 

submit implementation plan/schedule (if appropriate), start construction, end conrtruction, 

and attain compliance with permit. 

c. Reporting/Other Violations 

Violation of other toxic control requirements 

(including reports) will be reported using 

criteria that are applied to comparable NPDES 

permit conditions. For example, failure to 

submit a report within 30 days after the due 

date or submittal of an inaccurate or inadequate 

report will be reportable noncompliance (on 

the QNCR). 

Only failure to submit toxicity limit selfmonitoring reports or final TRE progress reports 

indicating compliance will be SNC when 30 days 

or more overdue. 

Resolution (bringing into compliance) of all three 

types of permit violations (effluent, schedule, 

and reporting/other) will be through timely and 

appropriate enforcement that is consistent with 

EPA Oversight Guidance. Administering agencies 

are expected to bring violators back into compliance 

or take formal enforcement action against facilities 

that appear on the QNCR and are in SNC: otherwise, 

after two or more quarters the facility must be 

listed on the Exceptions List. 

2. Approaches to Enforcement of Effluent Limitations 

In the case of noncompliance with whole effluent 

toxicity limitations, any formal enforcement action 

will be tailored to the specific violation and remedial 

actions required. In some instances, a Toxicity 

Reduction Evaluation (TRE) may be appropriate. However, 

where directly enforceable toxicity-based limits are 

used, the TRE is not an acceptable enforcement response 

to toxicity noncoqiance if it requires only additional 

monitoring without a requirement to determine appropriate 

remedial actions and ultimately compliance with the 

limit. 

If the Regions or States use administrative 

enforcement for violations of toxic requirements, 

such actions should require compliance by a date 

certain, according to a set schedule, and an 

B-4-14 
-9- 

administrative penalty should be considered.1 

Failure to comply with an Administrative Order 

schedule within 90 days indicates a schedule delay 

that may affect the final compliance date and a 

judicial referral is the normal response. In instances 

where toxicity has been measured in areas with potential 

impacts on human health (e.g., public water supplies, 

fish/shellfish areas, etc.), regions and states 

should presume in favor of judicial action and seek 

immediate injunctive relief (such as temporary 

restraining order or preliminary injunction). 

In a few highly unusual cases where the permittee has implemented an exhaustive TRE plan2, applied 

appropriate influent and effluent controls3, maintained 

continued compliance with all other effluent limits, 

compliance schedules, monitoring, and other permit 

requirements, but is still unable to attain or maintain 

compliance with the toxicity-based limits, special 

technical evaluation may be warranted and civil penalty 

relief granted. Solutions in these cases could be 

pursued jointly with expertise from EPA and/or the 

States as well as the permittee. 

Some permittees may be required to perform a 

second TRE subsequent to implementation of remedial 

action. An example of the appropriate use of a 

subsequent TRE is for the correction of new violations 

of whole effluent limitations following a period of 

lFedera1 Administrative penalty orders must be linked to violations 

of underlying permit requirements and schedules. 

2See Methods for Aquatic Toxicity Identification Evaluations, 

Phase I, Toxicity Characterization Procedures, EPA-600/3-88/035, 

Table 1. An exhaustive TRE plan covers three areas: causative 

agent identification/toxicity treatability; influent/effluent 

control; and attainment of continued compliance. A listing of 

EPA protocols for TREs can be found in Section V (pages 11 and 

12). 

3For industrial permittees, the facility must be well-operated 

to achieve all water quality-based, chemical specific, or BAT 

limits, exhibit proper 0 & M and effective BMPs, and control 

toxics through appropriate chemical substitution and treatment. 

For POTW permittees, the facility must be well-operated to 

achieve all water quality-based, chemical specific, or secondary 

limits as appropriate, adequately implement its approved pretreatment program, develop local limits to control toxicity, and 

implement additional treatment. 

B-4-15 
- 10 - 

sustained compliance (6 months or greater in duration) 

indicating a different problem from that addressed 

in the initial TRE. 

3. Enforcement of Compliance Schedule and Reporting 

Requirements 

In a number of instances, the primary 

requirements in the permits to address toxicity 

will be schedules for adoption and implementation 

of biomonitoring plans, or submission of reports 

verifying TREs or other similar reporting requirements. Regions and States should consider any 

failure (1) to conduct self-monitoring according 

to EPA and State requirements, (2) to meet TRE 

schedules within 90 days, or (3) to submit reports 

within 30 days of the specified deadline as SNC. 

Such violations should receive equivalent enforcement follow-up as outlined above. 

4. Use of Administrative Orders With Penalties 

In addition to the formal enforcement actions 

to require remedial actions, Regions and States 

should presume that penalty AO's or State equivalents can be issued for underlying permit violations 

in which a formal enforcement action is appropriate. 

Headquarters will also provide Regions and States 

with guidance and examples as to how the current 

CWA penalty policy can be adjusted. 

5. Enforcement Models and Special Remedies 

OWEP and OECM will develop standard pleadings 

and language for remedial activities and compliance 

milestones to assist Regions and States in addressing violations of toxicity or water quality-based 

permit limits. Products will include model litigation reports, model complaints and consent decrees, 

and revised penalty policy or penalty algorithm 

and should be completed in early FY 1989. 

54-16 
- 11 - 

v. Summary of Principal Activities and Products 

A. 

B. 

c. 

D. 

E. 

Compliance Tracking and Review guidance 

1. PCS Coding Guidance - May, 1987; revision 

2nd Quarter 1989 

2. Review Criteria for Self-monitoring Data (draft 

attached) 

Inspections and Quality Assurance 

1. Revised NPDES Compliance Inspection Manual - 

May 1988. 

2. Quality Assurance Guidance - 3rd Quarter FY 1989. 

3. Biomonitoring Inspection Training Module - 

August 1988. 

4. Additions of a reference toxicant to DMRQA program - 

(to be determined) 

Toxic8 Enforcement 

1. Administrative and Civil Penalty Guidance - 4th 

Quarter FY 1989 

2. Model Pleadings and Complaints - 2nd Quarter 1989 

3. EMS Revision - 2nd Quarter FY 1989 

Permitting Consistency 

1. Basic Permitting Principles - 2nd Quarter FY 1989 

Toxicity Reduction Evaluations 

1. Generalized Methology for Conducting Industrial 

Toxicity Reduction Evaluations - 2nd Quarter 

FY 1989 

2. Toxicity Reduction Evaluation Protocol for 

Municipal Wastewater Treatment Plants - 2nd Quarter 

FY 1989 

B-4-17 
- 12 - 

3. Methods for Aquatic Toxicity Indentification 

Evaluations 

a. Phase I. 

b. Phase II. 

C. Phase III. 

Toxicity Characterization 

Procedures, EPA-600/3-88/034- 

September 1988 

Toxicity Identification 

Procedures, EPA-600/3-88/035- 

2nd Quarter 1989 

Toxicity Confirmation ProceduresEPA-600/3-88/036 d Quarter 

FY 1989 

84-l 8 
APPENDIX B-5 

QUALITY CONTROL FACT SHEETS 
Attachment 

Quality Control Fact Sheet for Self-Biomonitoring 

Acute/Chronic Toxicity Test Data 

Permit No. 

Facility Name 

Facility Location 

Laboratory Investigator 

Permit Requirements 

Sampling Location 

Limit 

Type of Test 

Test Results 

LC50/EC50/NOEC/IC25 

Quality Control Summary 

Date of Sample 

Control Mortality % 

Type of Sample 

Test Duration 

Test Organism Age 

95 Percent Confidence Interval 

Dates of Test 

Control Mean Dry Weight 

Temperature maintained within ±2°C of test temperature? Yes No 

Dissolved oxygen levels always greater than 40 percent saturation? Yes No 

Loading factor for all exposure chambers less than or equal to maximum allowed for the test type 

and temperature? Yes No 

Do the test results indicate a direct relationship between effluent concentration and response of 

the test organism (i.e., more deaths occur at the highest effluent concentrations)? Yes 

No 

B-5-1 
APPENDIX B-6 

CASE DECISIONS ON WHOLE EFFLUENT TOXICITY 
CASE SUMMARY 

Natural Resources Inc. v. EPA, Defense 859 F.2d 159 Council, (D.C. Cir. 1988). 

This consolidated case, which arose from EPA’s promulgation of various National Pollutant Discharge 

Elimination System regulations, addresses a multitude of issues. The following paragraphs note issues 

particularly relevant to this document. 

• The Court held that EPA has the authority to express permit limitations in terms of toxicity as long as 

the limits reflect the appropriate requirements of the Clean Water Act (CWA), as provided in 40 CFR 

125.3(c)(4). The Court concluded that although toxicity appears to be an attribute of pollutants 

rather than a pollutant itself, the CWA (by means of the broad definition of “pollutant” in section 

502(6)) authorizes the use of toxicity to regulate effluents. 

• Industry asked the Court to address several other issues related to setting toxicity limitations (whether 

EPA failed to demonstrate the existence of a reliable methodology for setting toxicity limits and 

whether EPA’s use of toxicity to set water quality-based limitations to meet “narrative” State water 

quality standards represents an impermissible trespass on the State’s right to set water quality 

standards). However, the Court did not regard these issues to be adequately developed (“ripe”) for 

review in this case. 

• 

• 

The Court disagreed with industry’s assertions that EPA’s 1984 policy statement (“Development of 

Water Quality-Based Permit Limitations for Toxic Pollutants: National Policy,” 49 Federal Register 

9016 [March 9, 1984]) and draft Technical Support Document (“TSD”) were “rules” requiring notice and 

comment under the Administrative Procedure Act, 5 USC 553. The Court noted that informal 

rulemaking regarding 40 CFR 125.3(c)(4), which was pending between 1980 and 1984, did not limit 

Agency information gathering to the issuance of new or revised notices of proposed rulemaking, and 

the two documents did not have independent legal value. (In other words, the EPA national policy 

and the TSD were not binding norms but general statements of policy/guidance.) 

Industry also challenged EPA’s refusal to provide an affirmative upset defense to noncompliance with 

water quality-based limits. The Court indicated that the CWA does not expressly allow such an upset 

defense, and, upon considering the Act’s structure and legislative history, it could discern no 

congressional intent to provide for the defense in water quality permitting. Significantly, in 

reaching this position, the Court relied heavily upon the language and legislative history of CWA 

Section 301(b)(1)(C), by which Congress clearly did not relate compliance with water quality-based 

limitations to the capabilities of technology. In the Court’s view, “Congress had a deep respect for 

the sanctity of water quality standards and a firm conviction of the need for technology-forcing 

measures.” 895 F.2d at 208-09. However, the Court concluded that EPA had acted arbitrarily in 

dismissing the defense as impracticable, and directed EPA to conduct further proceedings on the issue. 

l Finally, the Court rejected challenges to EPA’s regulations governing State public participation 

requirements and penalty levels. In deciding these issues, the Court noted Congressional desire for 

nationally uniform effluent limitations as reflected in the legislative history of the 1972 CWA. The 

Court stated : 

Uniformity is indeed a recurrent theme in the Act, a direct manifestation of concern that the permit 

program be standardized to avoid the “industrial equivalent of forum shopping” and the creation of 

“pollution havens” by migration of dischargers to areas having lower pollution standards (859 F.2d at 

174 [footnotes omitted] and see accompanying footnotes 17-20 citing various provisions of the 

legislative history of the 1972 CWA). 

B-6-1 
APPENDIX C 

AMBIENT TOXICITY TESTING AND DATA ANALYSIS 
Ambient Toxicity Analysis 

Ambient toxicity testing procedures are useful where measurement of toxicity levels after discharge is important 

in the assessment of toxic effluent impact. This is particularly true where impact is caused by the presence of 

multiple point sources. The purpose of this testing is to provide an analysis of toxicity levels instream from 

whatever sources of toxicity are affecting the receiving water. 

Procedures 

The basic ambient toxicity testing procedure is to expose test organisms to receiving water samples taken from 

selected sampling stations above, at, and below the discharge point(s). Since effluent concentrations after 

discharge are often relatively low, chronic toxicity tests should be conducted so that the tests are sensitive 

enough for the purpose. 

The methods available for chronic testing of sufficiently short duration are limited. Two organisms for which 

short-term chronic toxicity tests are available are Pimephales promelas and Ceriodaphnia sp. 

The following procedures are used: 

• 

• 

• 

• 

• 

Select instream sampling stations based on the mixing characteristics involved in the specific 

discharge situation. 

Collect a daily grab sample or a daily composite sample of receiving water from each station. 

Use a renewal testing method to expose test organisms to the daily samples collected at each station. 

Use an appropriate number of replicates (10 for Ceriodaphnia) for each sampling station. No dilution 

series is required where screening is the primary goal. 

Conduct testing at a low-flow period, although it is not necessary to conduct the tests at the critical 

low-flow period. Testing is best when relatively stable flow occurs during the test period. 

Record the results of the testing in the format shown in Table C-1. The survival of the test organisms 

and the effect on their growth or reproduction are used as endpoints. Figure C-l plots the results in 

graphic form so that the pattern of ambient toxicity can be observed. 

C-l 
Table C-l. Young Production and Percent Survival of Cerioduphniu in Ambient 

Toxicity Tests at Ottawa River, Lima, Ohio 

River Young Final Dailv Survival 

Station Station Description Mile Female SD Survival 1 2 3 4 5 6 7 

1 Above Lima 46.0 15.5 8.0 90 100 100 100 90 90 90 90 

2 Above STP 37.7 14.1 2.1 0 100 100 100 100 90 10 0 ; 

3 Below STP 37.4 0 - 0 100 100 10 0 0 0 0 

3A Midway between 

STP and refinery 37.3 0 - 0 100 100 10 0 0 0 0 

36 Above refinery 37.1 0.4 - 0 90 90 40 0 0 0 0 

4 Above chemical plant 36.9 7.5 3.6 10 4A Below chemical plant 36.3 11.1 4.6 30 100 100 100 100 100 50 10 I 100 100 100 100 100 40 30 

5 Shawnee Bridge 36.4 5.7 4.0 0 9090909090600~ 

6 Route 117 32.5 12.6 3.8 10 100 100 100 100 100 100 10 

i 8’ Allentown 28.8 16.8 6.1 100 100 100 100 100 100 100 100 1 

Rimer 16.0 17.4 9.5 100 90909090 908080 

BA “Boonie” Station 8.0 25.0 3.3 100 100 100 100 100 100 100 100 

9 Kalida 1.0 25.6 5.5 100 100 100 100 100 100 100 100 

I - I I I t 

60 65 45 32 16 0 

River Kilometers 

Figure C-l. Cerioduphniu Young Production in Water from Various Stream 

Stations on the Ottawa River, Lima, Ohio 

c-2 
Selecting Sampling Stations 

The selection of sampling stations is determined by the characteristics of the site. When determining stations, 

consider the following factors: 

l Mixing and flow-The mixing characteristics of the discharge site are useful to determine the 

placement of sampling stations. Knowledge of concentration isopleths allows the regulatory 

authority to place stations at locations instream that correspond to concentrations measured in the 

dilution series in the effluent tests. For example, where effluent testing shows the effluent no 

observed effect concentration is 10 percent, an instream station should be placed where dilution is 

estimated to create a lo-percent instream waste concentration. In this way, the size of a toxic plume 

can be measured. Sampling stations should be placed where the effluents exist at relatively constant 

and relatively specific concentrations. Test at specific low-flow conditions, if possible. Presence of 

tributaries or other sources of dilution will influence positions and numbers of stations. Where 

smaller tributaries have several point sources on them, treat the tributary as a point source. Obvious 

nonpoint source areas also should be used to set stations. 

l Existing biological data-Where biosurvey data are available, sampling station location should be 

influenced by the more obvious trends in impact. In particular, control stations and recovery stations 

can be determined by biosurvey data. 

. Single point sources-Single point source situations should be bracketed with an above station, an 

immediate mixing station, several intermediate stations corresponding to different instream 

concentrations, and a recovery station. Of course, a control station should be established. 

l Presence of other point sources-Multiple point source situations require the ptacement of more 

stations between discharge points. Each source should be bracketed by sampling stations. 

There are four areas or zones that can be recognized when establishing the sampling stations for ambient 

toxicity testing: 

Zone l-An upstream zone before the effluent enters 

Zone 2-A zone of mixing 

Zone 3-A zone after mixing and before additional dilution water enters 

Zone 4-A zone where additional dilution occurs either from effluents or tributaries. 

All possible combinations of occurrences are not practical to discuss but must be sorted out for each site. Some 

generalizations are important to mention: 

l Any upstream sources of contaminants, such as other discharges, will confound the individual effects 

of a downstream discharge. For example, Zone 3 of the downstream discharge may occur in Zone 4 of 

an upstream discharge. This does not invalidate the measurement of ambient toxicity. It only makes 

it difficult to attribute amounts of response to each individual discharge. Response to the instream 

mixture is what is measured. 

l Careful location of sampling stations in Zone 3 is critical. Zone 3 is the only place where toxicity 

decay rates of any one discharger can be measured and then only if there are no upstream discharges, or 

if there are, only if that upstream effluent is stable in that reach. 

l In Zone 4, not only is degradation of the effluent toxicity occurring, but there is dilution of it by 

other effluents and tributaries. Depending on the site circumstances, one may not be able to learn 

anything about the ambient toxicity characteristics of the effluent of concern in this zone. 

l To emphasize, what can be measured in each zone depends on the above considerations. In the more 

complex situation, only an estimate of ambient toxicity at each station can be obtained. No 

information about one effluent’s toxicity decay rate will be available where several toxic effluents 

mix. In the most simple situation of one discharge and no dilution downstream for a long distance, 

Zone 3 will be large enough to get a good measure of toxicity decay. 

c-3 
Analysis of Ambient Toxicity Measurement 

When used in screening, the ambient toxicity data can identify areas in receiving waters where 

ambient toxicity exists instream. Attributing such impact to specific point sources (particularly where 

several sources discharge) may require effluent toxicity testing. 

Except when used for screening purposes, ambient toxicity measurements must be interpreted with 

effluent toxicity test data if conclusions are to be drawn concerning changes in toxic effect after 

discharge. The same species must be used in both the ambient and the effluent toxicity tests. 

When analyzing the data, the performance of the animals at each station downstream is compared to 

that oftbe animals exposed to receiving water without the effluent of concern in it but containing 

all other upstream additions. The result is an integration of effects from all contaminants and 

components and represents not only the toxicity of the effluent of concern but also the interactions 

of it with other effluents. 

Where the downstream stations show toxic effect at the concentrations measured as toxic in the 

effluent toxicity tests, effluent toxicity can be considered to be occurring instream, after discharge. 

Where the toxic effect decreases from station to station downstream in the absence of further dilution, 

the effluent toxicity is degrading. If the decay rate is rapid (e.g., no toxicity at the closest instream 

station to the discharge point), the effluent has a nonpersistent toxicity. Where the decay rate is 

more gradual, toxicity is more persistent. The rate of decay of toxicity together with mixing data 

allows the regulatory authority to approximate a receiving water toxicity impact area. That impact 

area can then be compared to the appropriate State water quality standards when establishing control 

requirements. 

In some cases, ambient toxicity may increase in relation to effluent toxicity measurements. Either 

upstream sources of toxicity exist or some factor in the receiving water is reacting with the effluent to 

increase its toxicity. Again, the pattern and magnitude of change in toxicity should be analyzed. 

Differences in toxicity levels between stations will reveal what is happening to the effluent as it is 

mixed instream and interacts with the constituents of the receiving water. 

Trend analysis in the raw test data is important when interpreting ambient toxicity data. As used in 

this context, trend analysis means observing toxic effect as it occurs in the test itself and relating it 

to what is occurring instream (plug flow, intermittent discharge, peak toxicity of effluents). Using 

time-of-travel data or receiving water flow rates and patterns, observe effects on the test organisms 

from day to day. There may be a pattern of mortality that can be linked to discharge events. For 

example, in the table the data indicate late mortality at downstream stations on Days 6 and 7. Flow 

rates for the river in this example correlated this mortality to the downstream movement of a toxic 

slug illegally discharged upstream. 

c-4 
APPENDIX D 

DURATION AND FREQUENCY 
DURATION AND FREQUENCY 

As discussed on pages 7 through 13 of the Guidelines for Deriving Numerical National Water Quality Criteria for 

the Protection of Aquatic Organisms and Their Uses [1], the format used to express water quality criteria for aquatic 

life should take into account toxicological and practical realities. Because of variation in the flows of the 

effluent and the upstream receiving water as well as variation in the concentrations of pollutants in the 

effluent and in the upstream receiving water, a simple format, such as specifying a concentration that must not 

be exceeded at any time or place, is not realistic. Furthermore, such a simple format does not take into account 

the fact that aquatic organisms can tolerate higher concentrations of pollutants for short periods of time than 

they can tolerate throughout a complete life cycle. The format that was selected for expressing water quality 

criteria for aquatic life consists of recommendations concerning concentrations, durations of averaging periods, 

and average frequencies of allowed excursions. Use of this concentration-duration-frequency format allows 

water quality criteria for aquatic life to be adequately protective without being as overprotective as would be 

necessary if criteria were expressed using a simpler format. In addition, this format can be applied directly to 

hydrological data and to the flow of, and concentrations of pollutants in, effluents using both dynamic and 

steady-state modeling [2, 3]. 

In aquatic life criteria for both individual chemicals and Whole Effluents, the recommended concentrations are 

the criterion maximum concentration (CMC) and the criterion continuous concentration (CCC). For individual 

chemicals the CMC and CCC are derived using the procedures described by Stephan et al. [1]. As described in 

Chapter 3 of this TSD, the CMC and CCC for Whole Effluents can be specified generically in terms of toxic 

units. Alternatively, for a particular effluent the CMC is specified in terms of an acute toxicity endpoint (ATE), 

which is either an LC50 or an EC50, and the CCC is specified in terms of a chronic toxicity endpoint (CTE), 

which is either a no observed effect concentration (NOEC) or an ICxx, if the LC50, EC50, NOEC, and ICxx, 

were obtained from appropriate toxicity tests conducted on the effluent with sensitive species. 

The CCC is intended to be the highest concentration that could be maintained indefinitely in a receiving 

water without causing an unacceptable effect on the aquatic community or its uses. Any concentration above 

the CCC, if maintained indefinitely, is expected to cause an unacceptable effect. Due to the four sources of 

variation mentioned above, the concentration in the receiving water will not be constant. Because organisms 

can tolerate higher concentrations for short periods of time, it is expected that the concentration of a pollutant 

in a body of water can exceed the CCC without causing an unacceptable effect if (a) the magnitudes and the 

durations of exceedances are appropriately limited and (b) there are compensating periods of time during 

which the concentration is below the CCC. These goals are accomplished by specifying a duration of an 

averaging period over which the average concentration should not exceed the CCC. For example, if the 

concentration is twice the CCC for one-half the specified averaging period, it must be zero for the rest of the 

averaging period if the average concentration is not to exceed the CCC. Thus, both the magnitude and 

duration of an exceedance are limited and there must be a compensating period of time during the averaging 

period when the concentration is below the CCC. Because exceedences are defined to be due to usual variation, 

most exceedences will be small, with larger exceedances becoming increasingly rare [1, 2]. 

Although an exceedance is defined to occur whenever the instantaneous concentration is above the CCC, an 

excursion is defined to occur only when the average concentration over the duration of the averaging period is 

above the CCC. It is expected that excursions can occur without causing unacceptable effects if (a) the 

frequency of such excursions is appropriately limited and (b) all other average concentrations are below the 

CCC. The recommended average frequency of allowed excursions is intended to appropriately limit the 

frequency of excursions. Because excursions are the highest average concentrations that occurred due to usual 

variation, all other average concentrations will be less than the CCC. As for exceedances, excursions that are 

defined to be due to usual variation will be small, with larger excursions becoming increasingly rare. The 

duration of the averaging period is intended to limit the impact of exceedances, whereas the average frequency 

of allowed excursions is intended to limit the impact of excursions. (Note: The words “exceedance” and 

“excursion” are used slightly differently here than in References 1 and 2.) 

Although spills can impact aquatic communities, they are not considered exceedances or excursions because 

they are not part of the usual variation in the concentrations of pollutants in receiving water. In the Complex 

Effluent Toxicity Testing Program, eight field studies were conducted to evaluate the use of toxicity tests to 

diagnose the cause of biological impact. Ambient toxicity measurements were taken over a 7-day period. 
During two of these studies [4, 51 spills of pollutants resulted in acute toxicity. This suggests that the impacts 

caused by spills might be as important as impacts caused by variation in the compositions and flows of the 

effluent and the receiving water. 

The primary purpose of this appendix is to present the rationale for the recommendations of the U.S. EPA 

concerning duration and frequency in national water quality criteria for aquatic life. The recommended 

duration is based on data from laboratory toxicity tests, whereas the recommended frequency is based on field 

data. With the concurrence of the U.S. EPA, States may adopt site-specific criteria, rather than national criteria, 

in their State standards. Such site-specific criteria may include not only site-specific concentrations, but also 

site-specific, and possibly pollutant-specific, durations of averaging periods and average frequencies of 

allowed excursions. If adequate justification is provided, site-specific and/or pollutant-specific 

concentrations, durations, and frequencies may be higher or lower than those given in national water quality 

criteria for aquatic life. A secondary purpose of this appendix is to discuss rationales that might be used as a 

basis for selecting alternative durations of averaging periods and average frequencies of allowed excursions. 

Duration 

In order for this concentration-duration-frequency format to allow water quality criteria for aquatic life to be 

adequately protective without being unnecessarily overprotective, the duration of the averaging period must 

allow some exceedances above the CCC without allowing unacceptable effects. Thus, the averaging period 

must appropriately limit the magnitude and duration of exceedances and provide compensating periods of time 

during which the concentration is below the CCC. 

Even though only a few tests have compared the effects of a constant concentration with the effects of the 

same average concentration resulting from a fluctuating concentration, nearly all the available comparisons 

have shown that substantial fluctuations result in increased adverse effects [6-161. Thus, the duration of the 

averaging period must be shorter than the duration of the chronic tests on which the CCC is based so that the 

averaging period does not allow substantially more adverse effect than would have been caused by a continuous 

exposure to the same average concentration. Life-cycle tests with species such as mysids and daphnids and early 

life-stage tests with warmwater fishes usually last for 20 to 30 days, whereas life-cycle tests with Ceriodaphnids 

usually last for 7 days. If the duration of the averaging period is too short, however, it will not allow any 

meaningful exceedances and will, in effect, defeat the purpose of the concept of the averaging period. For 

example, because few effluents are monitored more often than once a day, an averaging period of 24 hours 

would effectively mean that for most effluents each individual sample that was above the CCC would be 

considered an excursion. 

For the following reasons, a 4-day averaging period is recommended for application of the CCC in aquatic-life 

criteria for both individual pollutants and Whole Effluents: 

l It is substantially shorter than the 20- to 30-day duration of most chronic tests and is somewhat 

shorter than the 7-day duration of the Ceriodaphnio life-cycle test. 

l The results of some chronic tests apparently are due to an acute effect on a sensitive life stage that 

occurs at some time during the test, rather than being caused by either long-term stress or long-term 

accumulation of the test material in the organisms. Horning and Neiheisel [17] documented one such 

situation, and others are probably the cause of at least some of the acute-chronic ratios that are not 

much greater than unity. 

l For both endrin and fenvalerate, Jarvinen et al. [18] found that a 72-hour exposure caused about the 

same amount of effect on the growth of fathead minnows in early life-stage tests as did a 30-day 

exposure to the same concentration. 

l In some life-cycle tests on effluents with Ceriodaphnids, concentrations of effluents that were a 

factor of 1.8 greater than the CCC caused unacceptable effects in 4 or 5 days [S, 19, 201. 

l It is not so short as to effectively defeat the purpose of the concept of the averaging period. 

As discussed below, other averaging periods might be acceptable on a site-specific or pollutant-specific basis. 

D2 
Just as the concept of exceedances can be applied to the CCC, it also can be applied to the CMC. As with the 

CCC, the CMC averaging period should be substantially less than the lengths of the tests on which the CMC is 

based, i.e., substantially less than 48 to 96 hours. Because 4- to 8-hour LC5gs are about the same as the 96-hour 

LC50 for some materials [21-271, the duration of the averaging period for the CMC should be less than 4 hours. 

One hour is probably an appropriate duration of the averaging period for the CMC because concentrations of 

some materials that are only a factor of two higher than the 96-hour LC50 cause death in one to three hours 

[25]. Even when organisms do not die within the first hour or so, it is not known how many organisms might 

have died due to the delayed effects of the short exposure [28-311. If the l-hour average does not exceed the 

CMC, it is unlikely that the concentration of the pollutant in the receiving water can fluctuate rapidly enough 

during the hour to cause additional adverse effect. Thus, it seems inappropriate to apply the CMC to 

instantaneous concentrations. 

With adequate justification, the CMC and/or CCC averaging periods may be increased or decreased on a sitespecific or pollutant-specific basis. A possible site-specific justification for increasing the duration of the 

CCC averaging period would be that the variation in the concentration of the pollutant in the receiving water 

is low. Where variation is demonstrated to be consistently low, a longer CMC averaging would be acceptable 

because the magnitudes and durations of exceedances above the CCC would be limited. A possible pollutantspecific justification for a longer averaging period would be that the LC50 decreases substantially as the 

length of the exposure increases. For example, an B-hour averaging period might be justified for the CMC if it 

were shown that 24-hour exposures of a variety of sensitive species resulted in 96-hour LC5gs that were 

substantially above the 96-hour LCjos obtained from continuous exposure to a constant concentration for 96 

hours. 

In some situations the duration of the averaging period does not have to be stated explicitly because one can 

be implicitly defined using an uptake rate and a depuration rate. For example, if it is known that a specific 

concentration of a pollutant in the whole body or in a particular tissue of an important aquatic species will 

result in an unacceptable effect on the survival, growth, and/or reproduction of that species, and if applicable 

that species or tissue, the only additional information needed to allow calculation of an excessively high 

estimate of the total maximum daily load from the record of daily flows is the allowed frequency of 

exceedances of the concentration in the aquatic species. Thus, this approach can be used whenever the 

following are available: 

l A record of daily flows of the body of water, preferably for more than 10 years 

l A maximum acceptable concentration in the whole body or in a particular tissue of an aquatic species 

l Uptake and depuration rates that are applicable to that pollutant in the whole body or tissue of that 

species 

l An allowed frequency of exceedances of the maximum acceptable concentration. 

This approach is likely to be especially useful when an exposure causes delayed effects that are considered 

unacceptable. For example, it might be found in a test that no fish die during a 2-day exposure of rainbow trout 

to a pollutant but 50 percent of the fish die within 4 weeks of being transferred to clean water, whereas no 

comparable control fish die. If values are available for the concentration of the pollutant in the fish at the 

end of the 2-day exposure and for the uptake and depuration rates, these data could be used with a flow record 

for a river to determine how often a specified constant daily input of the pollutant to the river would have 

resulted in exceedances of this concentration and therefore the death of rainbow trout. 

Regardless of what averaging periods are used, exact calculation of the number of excursions would require 

continuous monitoring of the concentration in the receiving water, which is not feasible in most cases. A 

valid alternative would be to use a statistically designed monitoring program and a statistical interpretation of 

the measured concentrations. The l-hour averaging period for the CMC would imply that the samples analyzed 

should be l-hour composites; the 4-day averaging period would imply that concentrations in all samples 

obtained within any 4-day period should be averaged, preferably using a time-weighted average. If information 

is available concerning the discharge pattern of a particular effluent, it might be possible to design a 

monitoring program that is specifically appropriate for that effluent. 

D3 
Unless critical species are especially sensitive to particular toxicants, most excursions of criteria should have 

minor impacts on aquatic communities. However, whereas excursions above the CCC will probably reduce 

growth and reproduction, excursions above the CMC will probably cause death and other severe acute effects. 

In addition, special care should be exercised when many outfalls exist in a small segment of a receiving water, 

because if low flow causes an excursion for one discharge, that same low flow will probably also cause 

excursions for other discharges at the same time. Several “minor” excursions might thus add up to a “major” one. 

Frequency 

The purpose of the average frequency of allowed excursions is to provide an appropriate average period of time 

during which the aquatic community can recover from the effect of an excursion and then function normally for 

a period of time before the next excursion. The average frequency is intended to ensure that the community is 

not constantly recovering from effects caused by excursions of aquatic-life criteria. Because most regulated 

discharges are to flowing water (lotic) systems, this discussion will emphasize discharges to rivers and streams 

rather than to lakes, ponds, reservoirs, and estuaries. 

6emral Cunsideratlons for Settln# Frequency wttb Wbicb Excursions of Criteria May Occur 

Not long ago ecological communities were thought to be largely in equilibrium and their structure and function 

determined primarily by internal interactions between species, such as competition and predation. 

Communities were considered to be analogous to “super-organisms,” with close parallels to organisms in their 

response to stress and in “health.” Current understanding is that external factors, including disturbances, often 

play a major role in the structure of communities [32, 331. The frequency of disturbance affects a community not 

only by decreasing the fitness of component species, but also by causing a natural selection of species or 

phenotypes having characteristics that allow them to tolerate or even thrive under the disturbance regime. 

Natural disturbances such as floods and droughts are common in lotic systems [32] and vary in intensity not only 

between headwater streams and large rivers, but also between similar sized lotic communities in different 

climatic regions. Rather than requiring more time to recover from the effects of additional anthropogenic 

disturbances, lotic communities with high natural background disturbance frequencies are actually predisposed 

to recover more rapidly because only species that are able to recolonize and reproduce quickly, or perhaps to 

avoid disturbances, can persist there [34-371. This does not imply that they also are more resistant to novel 

anthropogenic disturbances with which they have had no previous evolutionary experience; it only implies that 

they are predisposed to recover quickly once the disturbance is gone. The question then is how frequently can 

aquatic communities experience these additional disturbances (excursions of criteria) without being 

unacceptably affected. 

In an extensive review of the published literature, Niemi et al. [38] reviewed the published literature and 

identified more than 150 case studies of freshwater systems in which some aspect of recovery from the impact of 

a disturbance was reported. A case study was used only if the disturbance caused a death or displacement of 

organisms. This restriction was necessary because it was rarely possible to determine if an event was outside the 

normal intensity range (a common alternate definition of disturbance), mainly because it is usually difficult to 

define the normal intensity range. It also permitted the inclusion of natural as well as anthropogenic events. 

Approximately 80 percent of these systems were lotic, and the remainder were lentic (lakes and ponds). The 

impacts were due to such disturbances as persistent and nonpersistent chemicals, logging, flooding, 

channelization, dredging, and drought. Reported endpoints for recovery were sparse for phytoplankton, 

periphyton, and macrophytes, but were numerous for macroinvertebrates and fishes. Because more than one 

recovery endpoint was reported for most studies, the number of endpoints greatly exceeded the number of case 

studies. For short-term (nonpersistent) disturbances, approximately 85 percent of all macroinvertebrate 

endpoints indicated recovery in less than 2 years. Macroinvertebrate biomass, density, and taxonomic richness 

recovered in less than 1 year for approximately 95 percent of reported endpoints. Dipterans (flies, mosquitos, 

midges, etc.), which generally have short generation times or high dispersal ability, recovered most rapidly, 

whereas stoneflies and caddisflies recovered least rapidly. Fishes recovered in 2 years or less for over 85 percent 

of reported endpoints. However, as discussed below, important exceptions did occur. 

Most excursions of criteria will be minor and their impacts will therefore be difficult to detect. Although most 

disturbances in the above case studies caused more severe impacts than most criteria excursions are expected to 
cause, CMC excursions will result in death of some organisms. These data indicate that as a general rule, the 

purpose of the average frequency of allowed excursions will be achieved if the frequency is set at once every 3 

years on the average. Excursions of the CCC are more difficult to evaluate because nonlethal excursions could 

not be evaluated from the data used by Niemi et al. [38]. It is reasonable to expect, however, that cumulative 

effects from too frequent excursion of the CCC also will result in unacceptable degradation of lotic 

communities. 

Considerations fer PK~OS/R# Sit@-specific Incmases or Decreases in tb8 Average Frequency of Allowed 

Excursions 

Although an average frequency of one criterion excursion every 3 years should usually be protective of lotic 

communities, more frequent excursions might be acceptable in certain situations. Sedell et al. [39] have shown 

that lotic systems with refugia (areas of refuge) such as well-developed riparian zones, connected flood plains 

and meanders, snags, etc., recover more rapidly from disturbances than segments without such refugia, because 

organisms are better able to avoid disturbances and return or repopulate. However, many of these refugia are 

likely to be most restricted and vulnerable during the low-flow periods when criteria excursions also are most 

likely to occur. Evidence of action to preserve refugia, particularly during low-flow periods, or to create or 

restore them, might be grounds for demonstrating that an excursion frequency of more than once every 3 years on 

the average is acceptable. Schlosser [36] found that lower-order (i.e., headwater) streams, because of their 

natural high variability, contain communities consisting of species that have short life cycles and/or high 

dispersal ability and can recover from major disturbances in a year or even less. Thus, many lower-order streams, 

particularly those for which refugia are available, may be able to tolerate somewhat higher excursion 

frequencies, unless other considerations are important. For example, discharges to lower-order streams sometimes 

constitute a large fraction of the stream flow for most of the year. 

Although lower-order streams are naturally highly variable and can therefore tolerate higher disturbance 

frequencies, the converse is true for higher-order lotic streams for at least two partially related reasons: (1) 

segments with tributaries draining a large watershed will be buffered from the effects of localized droughts in a 

portion of the watershed, and will therefore experience a less severe natural disturbance regime, and (2) 

organisms inhabiting these segments will therefore not be adapted to disturbances that are as frequent or severe 

as those in lower-order segments. Fish in particular will be larger and have longer generation times in larger 

streams and rivers. Consequently, it will take longer for these populations to reproduce and regain 

predisturbance densities and size class distributions. Schlosser [36] suggests that, based on such life-history 

characteristics, fish communities in larger rivers might take 20 to 25 years to re-establish the predisturbance age 

and size structure of their component populations after a severe disturbance such as a major drought or spill. 

Extreme cases in which recovery has taken much longer than 3 years usually involve spills of persistent 

chemicals or severe habitat modification, such as stream channelization or clear-cutting of a watershed [38]. If 

the chemical contaminant is not widespread, recovery is limited primarily by the rate of disappearance of the 

chemical rather than by strictly ecological processes. Widespread contamination can affect recovery by 

increasing the distance over which recolonizers must travel. Watershed clear-cutting reduces the input of 

organic matter that provides the food base of streams in forested watersheds and also provides woody debris 

and snags that serve as refugia. Channelization and dredging reduce the in-stream habitat diversity and thereby 

decrease refugia. In addition to these anthropogenic disturbances, multiple excursions during a drought, due to 

low-flow conditions, can result in a severe cumulative impact on sensitive species even if the individual 

excursions are small. Special measures, such as plant shutdowns, might be required in extreme cases. Finally, 

severe chemical spills, which cannot be regulated but which will occur in any highly industrialized river 

segment, will affect aquatic life over a large area. If maintenance of long-lived fish species in these segments 

is desired, recovery periods up to 25 years may be necessary. 

Based on the above considerations, recovery periods longer than 3 years may be necessary after multiple minor 

excursions or after a single major excursion or spill during a low-flow period in medium-to-large rivers, and up 

to 25 years where long-lived fish species are to be protected. Even longer times may be necessary as the size of 

the affected area or the persistence of the pollutant increases. 
Calculation of Design Conditions 

The use of aquatic-life criteria for developing water quality-based permit limits and for designing waste 

treatment facilities requires the selection of an appropriate wastetoad allocation model. Dynamic models are 

preferred for the application of aquatic-life criteria in order to make best use of the specified concentrations, 

durations, and frequencies. If dynamic models cannot be used, then an alternative is steady-state modeling. 

Because steady-state modeling is based on various simplifying assumptions, it is less complex, and might be 

less realistic, than dynamic modeling. 

An important step in the application of steady-state modeling to streams is calculating the design flow. The 

procedures outlined in the EPA document Technical Guidance Manual for Petforming Waste Load Allocation, Book 6, 

Design Conditions: Chapter I, Steam Design Flow for Steady-State Modeling. (U.S. EPA 1986) are recommended for 

calculating design flows for rivers and streams. States may use other methods so long as the methods are 

technically defensible. The document discusses and recommends two methods for determining design flows, 

the hydrologically based method and the biologically based method, and the flows that should be used for the 

CCC and CMC for both methods. 

The hydrologically based design flow method is presently used by many States. It is based on selecting and 

identifying an extreme value, e.g., the 7QlO flow. The underlying assumption of this method is that the 

design flow will occur X number of times in Y years. Thus, this method limits the number of years in which one 

or more excursions below the design flow can occur. The method has two advantages: (1) the log-Pearson Type 

III flow estimating technique or other extreme value analytical techniques that are used to calculate flow 

statistics from daily flow data are consistent with past engineering and statistical practice, and (2) the U.S. 

Geological Survey provides technical support for this method. The disadvantage of this method is that it is 

essentially independent of biological considerations. Design flows calculated using this method might allow 

more or fewer excursions than once every 3 years on the average. tn addition, it is difficult to use site-specific 

durations and frequencies with this method. For toxic wasteload allocation studies in which the 

hydrologically based method is used, EPA recommends the use of the 1 QlO flow as the design flow for the 

CMC and the 7410 as the design flow for the CCC. 

The biologically based design flow method was developed by the U.S. EPA Office of Research and 

Development and directly uses the averaging periods and frequencies specified in the aquatic-life water quality 

criteria for individual pollutants and Whole Effluents for determining design flows. The method is an 

empirical iterative convergence procedure that includes the calculation of harmonic means of the flow to 

determine the total number of excursions. The method makes exact use of whatever duration and frequency are 

specified for the CMC and CCC. These might be 1 day and 3 years for the CMC and 4 days and 3 years for the 

CCC or site-specific durations and frequencies. 

The two methods were used on approximately 60 different rivers to compare the hydrologically based 1 QlO and 

7410 design flows with the biologically based 1 -day/3-year and 4-day/3-year design flows. For most of the 

rivers the hydrologically based design flows resulted in more than the allowed number of excursions. For some 

of the rivers, the 1 QlO and 7410 allowed substantially more or fewer excursions than the intended number of 

excursions. Because the biologically based method calculates the design flow directly from the national or 

site-specific duration and frequency, it always provides the maximum allowed number of excursions and never 

provides more excursions than allowed. 

EPA provides software tools to calculate both types of design flows via the STORET environment on its NCCIBM mainframe. Biologically based design flows can be calculated using the program DFLOW [40]. The 

hydrologically based design flows can be calculated using FLOSTAT or DFLOW; the latter uses a simplified 

version of the log-Pearson Type III method. Both programs access the STORET Flow file that contains daily 

flow records for U.S. Geological Survey gaging stations. They are easy to use and the user simply needs to know 
the identification number of the gaging station. To obtain further information on the STORET environment 

and the programs, contact: 

Mr. Thomas Pandolfi 

U.S. Environmental Protection Agency 

Office of Water Regulations and Standards (WH-553) 

401 M Street, S.W. 

Washington, D.C. 20460 

(202) 382-7030 

The methods described above use daily flow data to determine design ftow, but they do not consider any other 

physical or chemical condition that might affect toxicity. EPA has prepared a supplementary method and a 

software tool named DESCON that incorporate such supplemental water quality parameters as temperature, pH, 

alkalinity, hardness, and dissolved oxygen to determine design conditions. Note that DESCON takes into 

account such things as effluent variability, which DFLOW does not take into account. The method and software 

are described in two documents available from the Assessment and Watershed Protection Division of the Office 

of Water Regulations and Standards-Technical Cuidonce on Supplementary Stream Design Conditions for Steady 

Store Modeling [3] and DESCON Users Manual [40]. 

The supplementary method is consistent with the hydrologically and biologically based methods described 

above. It simply extends them to include other conditions besides streamflow. The advantage of considering 

multiple conditions is that the worst-case conditions necessary to protect water quality criteria might not occur 

when the streamflow is low; e.g., low DO or high temperatures might occur at times other than when the flow is 

low. 

This supplementary method can be used for five pollutant categories with the physical-chemical parameters 

described above. The pollutant categories are general toxicant, ammonia, heavy metals (Cd, Cr+3, Cu, Pb, Ni, 

Zn), pentachlorophenol, and ultimate oxygen demand. 

The software tool to facilitate this method is called DESCON. It is on EPA’s IBM mainframe and is available 

through the STORET environment. DESCON accesses the STORET flow file for the daily flow record and the 

water quality file for data on the physical-chemical parameters. Options are available to the user if the area of 

concern has no flow record or if no water quality data are available. 

D7 
1. 

2. 

3. 

4. 

5. 

6. 

7. 

8. 

9. 

10. 

11. 

12. 

13. 

14. 

15. 

16. 

17. 

18. 

APPENDIX D 

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Their Uses. PB85-227049. National Technical Information Service, Springfield, VA. 

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Mount, D.I., N.A. Thomas, T.J. Norberg, M.T. Barbour, T.H. Roush, and W.F. Brandes. 1984. Effluent und 

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Mount, D.I., T.J. Norberg-King, and A.E. Steen, eds. 1986. Vulidity of Effluent ond Ambient Toxicity Tests for 

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and G.J. Niemi. In Recovery of Lotic Communities and Ecosystems from Disturbonce: Theory and 

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Niemi. In Recovery of Lotic Communities and Ecosystems from Disturbance: Theory and Applications. 

Environ. Management (submitted). 

Poff, N.L., and J.V. Ward. 1989. “The Physical Habitat Template of Lotic Systems: Recovery in the 

Context of Historical Pattern of Spatio-Temporal Heterogeneity.” Ed. J.D. Yount and G.J. Niemi. In 
Recovery of iotic Communities and Ecosystems from Disturbance: Theory and Applications. Environ. 

Management (submitted). 

38. Niemi, C.J., P. DeVore, N. Detenbeck, D. Taylor, I.D. Yount, A. Lima, j. Pastor, and R.J. Naiman. 1989. 

“An Overview of Case Studies on Recovery of Aquatic Systems from Disturbance.” Ed. J.D. Yount and 

C.J. Niemi. In Recovery of Lotic Communities and Ecosystems from Disttirbance: Theory and Applications. 

Environ. Management (submitted). 

39. Sedell, J.R., F.R. Hauer, C.P. Hawkins, and j.A. Stanford. 1989. “The Role of Refugia in Recovery from 

Disturbance: Modern Fragmented and Disconnected River Systems.” Ed. J.D. Yount and C.J. Niemi. In 

Recovery of Lotic Communties and Ecosystems from Disturbance: Theory and Applications. Environ. 

Management (submitted). 

40. Rossman, Lewis A. DFLOW Users Manual. 

DlO 
APPENDIX E 

LOGNORMAL DISTRIBUTION AND PERMIT LIMIT 

DERIVATIONS 
LOGNORMAL DISTRIBUTION AND PERMIT LIMIT 

DERIVATIONS 

Introduction 

This appendix provides supporting information for the statistical methodology used in permit limit 

calculations. The methodology described in this appendix applies to many types of data including data that 

are used to develop both technology-based and water quality-based permit limits. The appendix is divided 

into two sections. The first section gives an overview of permit limits: the derivation of water quality-based 

limits and the consistency among different permit limits. The second section describes the statistical 

methodology for the normal distribution, the lognormal distribution, the delta-lognormal distribution, 

methods of checking distributional assumptions, and correlation. This section also provides guidance on the 

application of each distribution to permit limits. Tables E-1, E-2, and E-3 at the end of the appendix summarize 

the permit limit calculations. This appendix describes the statistical methodology for three distributions that 

are often used in determining permit limits. Other distributions can be used, and this topic is discussed in the 

subsection, Other Distributions. 

Section 1: Overview of Permit Limits 

Two types of permit limits are contained in the effluent guidelines regulations: daily maximum limits and 

monthly average limits. The daily maximum permit limit is the maximum allowable value for any daily sample. 

The daily maximum limits are usually based on the 99th percentile of the distribution of daily measurements. 

The monthly average permit limit is the maximum allowable value for the average of all daily samples obtained 

during 1 month. Monthly average limits are in most cases based on the 95th percentile of the distribution of 

averages of daily values. 

The following two subsections discuss the derivation of water quality-based limits and the consistency among 

different permit limits. 

Derivation of Water Quality-based limits 

Water quality-based limits are derived from the required treatment system performance necessary to comply with 

the wasteload allocation (WLA). Technology-based effluent limits are derived from treatment system 

performance. The mathematical expressions for water quality-based limits are the same as those for technologybased effluent limits; the major difference is that the means and standard deviations in those expressions are 

derived from the WLA. This topic is discussed in Chapter 5. 

Consistency Among Different Permit limits 

The current Technical Support Document for Water Quality-based Toxics Control (TSD) procedures provide 

consistency among different permit limits. The stringency of permit limits is independent of monitoring 

frequency and is determined entirely by the WLA and permit limit derivation procedures. The daily maximum 

limit is constant regardless of monitoring frequency. The numerical value of the monthly average limit 

decreases as monitoring frequency increases only because averages become less variable as the number of values 

included in the average increases. For example, an average based on 10 samples is less variable than an average 

based on 4 samples. This phenomenon makes monthly average permit limits based on 10 samples appear to be 

more stringent than the monthly limit based on 4 samples. A permittee performing according to the WLA 

specifications will in fact be equally capable of meeting either of these monthly average limits when taking 

the corresponding number of samples. The stringency of the TSD procedures, accordingly, is constant across 

monitoring frequencies. 

E-l 
Section 2: Statistical Methodology 

The statistical procedures that are used in permit limit development involve fitting distributions to effluent 

data. The estimated upper percentiles of the distributions form the basis of the limits. This section describes 

the statistical methodology applied to permit limits in the following subsections: the normal distribution, the 

lognormal distribution, the delta-lognormal distribution, methods of checking distributional assumptions, and 

correlation. Before discussing these topics several definitions are made for notation, assumptions, coefficients 

of variation, and variability factors. 

In the calculations in this appendix, natural logarithms (i.e., logarithms to the base e), denoted by In(x), are 

used. The calculations can be modified to use logarithms to the base 10 by replacing loglO for In(x) in the 

formulas. 

Assmptlons 

The distribution fitting methods assume that the daily measurements are independent, uncorrelated 

observations. 

The fundamental assumptions underlying the discussion on calculating limits are: 

l Daily pollutant measurements are approximately lognormally distributed for values above the 

detection limit 

l Maximum n-day monthly averages for n 5 10 are approximately lognormally distributed above the 

detection limit 

l Maximum n-day monthly averages for n > 10 are normally distributed. 

Recommendation of the use of the lognormal distribution for daily pollutant measurements is based on 

practical rather than theoretical consideration. Usually environmental data sets possess the basic lognormal 

characteristics of positive values and positive skewness. In addition, the lognormal distribution is flexible 

enough to model a range of nearly symmetric data. Furthermore, in comparison to other positive valued, 

positively skewed distributions that could be used to model environmental data, the lognormal is relatively 

easy to use. 

When lognormal data are log transformed, the properties of the normal distribution apply to the transformed 

data. The section on statistical methodology describes the properties of the normal distribution and its 

relationship to the lognormal distribution. The delta-lognormal distribution is a generalization of the 

lognormal distribution and may be used to model data that are a mixture of non-detect measurements with 

measurements that are lognormally distributed. In delta-lognormal procedures, nondetect values are weighted 

in proportion to their occurrence in the data. 

In determining permit limits based on averages (e.g., monthly average permit limits), a distribution should be 

used that approximates the distribution of an average of pollutant measurements. The lognormal distribution 

can be used for approximating the distribution of averages for small sample sizes where the individual 

measurements are approximately lognormally distributed. For larger sample sizes, a powerful statistical result, 

called the Central Limit Theorem, provides theoretical support for determining limits based on averages of 

individual measurements. According to the Central Limit Theorem, when the sample size n is large enough, the 

average of the n sample values will be approximately normally distributed regardless of the distribution of the 

individual measurements. The section on statistical methodology provides procedures and guidance for 

calculating averages for both small and large samples sizes where the individual measurements are lognormally 

distributed. 

E-2 
The shape of the observed data is the key factor in evaluating a distributional model. For environmental data 

the lognormal distribution is usually appropriate. The critical question in a given situation is how well a 

particular distribution models the shape of observed data. Although the lognormal does not provide an exact 

fit in all cases, it usually provides an appropriate and functional fit to observed environmental data. Graphical 

displays and goodness-of-fit tests, as described in the subsection, Other Distributions, may be used as a guide in 

verifying assumptions and selecting a distribution. 

Coefflcimts uf V8ffatiae 

The coefficient of variation (denoted by “CV”) is the ratio of the standard deviation to the mean. Thus, the CV 

is a dimensionless measure of the relative variability of a distribution. Estimates of the CV can be used when 

the actual CV cannot be calculated or if the available data sets for calculating the CV are small. In such cases, 

different values for the CV should be used in the permit calculations to assess the effect of the CV on the final 

permit limit. Typical values of the CV for effluent data usually range from 0.2 to 1.2. The CV is a measure of 

the relative variation in observed data. In many cases, changes in the CV will have little impact on the final 

permit limit. In assessing the sensitivity of the permit limit to the CV, the calculations may include CV = 0.6 

as a conservative estimate (assumes relatively high variability). If the final permit values vary greatly with 

different CV values either of two approaches may be used. The first approach is to use a conservative estimate 

of the CV that assumes relatively high variability (e.g., CV = 0.6) in the final permit limit. The second 

approach is to collect additional data to obtain a more definitive value for the CV. 

viffi8biiitjf Factors 

An important component of the process used by the Environmental Protection Agency (EPA) for developing 

technology-based limits are variability factors. The variability factor is the ratio of a large concentration level 

of a pollutant to the average level determined from that particular plant. The ratio expresses the relationship 

between the average treatment performance level and large values that would be expected to occur only on rare 

occasions in a well-designed and operated treatment system. Such factors are useful in situations where little 

data are available to characterize the long-term performance of a plant. 

In cases where only a small number of observations are available from a plant, EPA has been reluctant to 

estimate a variability factor. In the Organic Chemicals, Plastics, and Synthetic Fibers (OCPSF) rulemaking [l], a 

minimum of seven daily observations from a plant, with at least three of the seven above the detection limit, 

was established for calculation of a plant level priority pollutant variability factor. However, EPA has not 

established a minimum number of observations required for calculating variability factors for all pollutants in 

all industries. 

The calculations for variability factors for the daily maximum and the monthly average are included in the 

discussion of the different distributions below. 

Normal Distribution 

The normal distribution plays a central role in the methods described in this appendix. In most cases, the 

normal distribution is not an appropriate model for individual pollutant measurements; however, the normal 

distribution is related to the lognormal distribution that is used to establish many permit limits. In most cases, 

the simple logarithmic transformation of effluent and water quality data results in data distributions that are 

normally distributed. Such data are referred to as being lognormally distributed. When lognormal data are log 

transformed, the properties of the normal distribution apply to the transformed data. Since the normal and 

lognormal distributions are related in a straightforward manner, the methods of analysis for normal and 

lognormal data also are easily related. The normal distribution is described below and is followed by a 

discussion of the lognormal distribution and its relationship to the normal distribution. 

E-3 
Figure E-l. Normal Probability Distribution 

The normal probability distribution is encountered in a number of applications. The bell-shaped curve of the 

normal distribution is shown above in Figure E-l. Excellent introductions and reviews of the normal 

distribution are found in numerous statistical, engineering, and scientific texts, as for example in Reference 2. 

Only a brief review is given here. 

A sample of independent observations, denoted by xl, x2 ,...,Xk, from a normally distributed population can be 

used to estimate the mean, p, and variance, cr2, according to the formulas below: 

6 = estimated mean 

= z[Xi] / k, 1 I i 5 k 

3 = estimated variance 

= Z[(xi-b)2]/(k-l),1 <ilk 

6 = estimated standard deviation 

= (&2)1/2 

c^v = estimated coefficient of variation 

The characteristics of the normal distribution are the range is defined for positive and negative values, and the 

frequency curve is bell-shaped and symmetric about the mean. In most cases, the normal distribution is not an 

appropriate model for the distribution of individual pollutant measurements. Environmental data rarely are 

symmetric, which is a fundamental property of the normal distribution. In addition, the normal distribution is 

defined over a range that includes negative values while pollutant measurements are restricted to nonnegative 

values. Thus, fitting a data set to a normal distribution allows for the possibility, however small, of observing 

negative values. The lognormal distribution, or any positive valued distribution, is not defined for negative 

values and thus avoids assigning any probability to negative values. 

Daily MaMaa Pem/t LMts Based on tbe Normal Distrihtiun 

For data sets which have the characteristics of the normal distribution, the daily maximum permit limits can be 

calculated. The upper percentile daily maximum permit limits for the normal distribution are calculated using 

the quantity zp, the standardized Z-score for the pth percentile of the standardized normal distribution (i.e., 

normal distribution with mean = 0, and variance = 1). For example, the Z-score for the 95th percentile is 1.645. 

Z-scores are listed in tables for the normal distribution (in most statistical textbooks and references). The pth 

percentile daily maximum limit is estimated by: 

4, = pth percentile daily maximum limit 

= Ji + zp 6. 

E-4 
For example: 

g.95 = 95th percentile daily maximum limit 

= i+1.645& 

R.99 = 99th percentile daily maximum limit 

= r; + 2.326 6 

Note: 

+,, = 1.645 

k = 2.326 

The daily variability factors (denoted by if,) are estimated by: 

Daily maximum 95th percentile VF1 = A.95 / i 

Daily maximum 99th percentile VF1 = A.99 / c 

HWbty Av#ra#o h&f LWfs 8md on fb# #omwl DistrlbutIm 

The normal distribution can be used to model the averages of the individual measurements for a wide range of 

circumstances. Although the normal ,distribution usually is not an appropriate model for individual pollutant 

measurements, the averages of those individual measurements can often be modeled by the normal distribution. 

This subsection explains the theory behind using the normal distribution for averages and provides the general 

formulas. 

A powerful statistical result, called the Central Limit Theorem, provides theoretical support for determining 

limits based on averages of individual measurements. According to the Central Limit Theorem, when the 

sample size n is large enough, the average of the n sample values will be approximately normally distributed 

regardless of the distribution of the individual measurements. In determining permit limits, the calculations 

incorporate the number of samples that will be required for monitoring purposes during the specified time 

period (usually a month). For the purposes of permit writing, monitoring sample sizes greater than 10 are 

recommended to be sufficiently “large enough” to assume the sample average is approximately normally 

distributed. The above formulas can be modified for finding the estimated mean and variance for the average 

from a sample of size n (e.g., for 14-day monthly average, n = 14 samples during the month for monitoring 

purposes). The parameters p,, and CJ~ denote the mean and variance, respectively, of the distribution of the 

average of n values. The estimates of the n-day average and the variance of the n-day average are denoted by 

fin and $?,, respectively. 

r; = 

62 = 

in = 

= 

.2 

0” = 

= 

i& = 

= 

CGn = 

= 

estimated mean of distribution of X 

estimated variance of distribution of X 

mean of distribution of the n-day monthly average 

3 

variance of distribution of the n-day monthly average 

i52 / n 

standard deviation 

(62,)‘/2 

coefficient of variation 

47 / fin. 

E-5 
The upper percentile limits are: 

A., = pth percentile n-day monthly average limit 

= in + Zp 6n 

where zp is the pth percentage point of the standard normal distribution. 

For example: 

R .95 = 95th percentile n-day monthly average limit 

= jIi,, + 1.645 &, 

k +99 = 99th percentile n-day monthly average limit 

= (l, + 2.326 b,, 

Note: 

Zg5 = 1.645 

Zgg = 2.326. 

The monthly average variability factors (denoted by VF,) are estimated by: 

Monthly average 95th percentile VF, = 2.95 / b 

Monthly average 99th percentile VF, = k.99 / $ 

The above discussion of the normal distribution can be modified for data from the lognormal distribution. The 

next subsection explains the modifications. 

Lognormal Distribution 

Experience has shown that daily pollutant discharges are generally lognormally distributed. The distributional 

fit of the data varies somewhat from application to application, but not enough to alter the conclusion that 

effluent pollutant discharges are generally lognormally distributed. Ambient water quality data also are often 

lognormally distributed. Figure E-2 displays the positively skewed shape of the lognormal distribution. 

The distribution fitting methods assume that the daily measurements are independent, uncorrelated 

observations. Although, in general, this assumption is not satisfied exactly, the lognormal distribution has 

been used in the effluent guidelines program primarily because it consistently provides a reasonably good fit to 

observed efftuent data distributions. Figure E-3 shows the lognormal distribution applied to data used in the 

development of the OCPSF effluent guidelines regulation [l]. 

E-6 
4.5 

4 

3.5 

3 

2.5 

2 

1.5 

1 

0.5 

0 

150 

50 

0 

- 

0 

I 

I 

3 

0.5 1 1.5 2 2.5 3 3.5 4 

Value 

Figure E-2. Examples of Lognormal Densities 

1 

9 15 21 27 33 39 45 51 57 63 69 

I I I I I I I I I I 

Concentration in ugl 

Figure E-3. BOD Frequency Distribution - Plant C 

E-7 
The logarithmic transformation of the random variable X, Y = In(X) results in a random variable Y that is 

normally distributed. Therefore, the analysis procedures for analyzing lognormal data are similar to those for 

the normal distribution. The mean and variance from the normal distribution of the random variable Y are oy 

and ;ry’ respectively. These parameters can be estimated by: 

$Y = C&i) / k 

and 

.* 2 

=Y = C[(yi - jQ2] / (k - l), respectively 

where - 

Yi = ln(xi) for i=1,2,...k. 

When data are lognormally distributed, these values from the normal distribution can then be used to calculate 

the mean, variance, and coefficient of variation for the random variable X that is lognormally distributed. The 

mean, variance, and coefficient of variation of the random variable X may be estimated by c(X), V(X), and 

c;(X), respectively. 

i?(X) = 

= 

V(X) = 

= 

h(X) = 

= 

daily average 

exp( ;ly + 6: / 2) 

variance 

exp(2 jiy + 6;) [exp(+ - l] 

coefficient of variation 

[exp($) - 1]li2. 

Dally Maximmu Petm/t LimMs Bawd on the Lo~nu/ma/ D&tribntion 

The upper percentile limits for the random variable X (which is lognormally distributed) are: 

R +P = pth percentile daily maximum limit 

= exp[ iiy + zp tSy 21 

where zp is the pth percentage point of the standard normal distribution. 

For example: 

k-95 = 95th p ercentile daily maximum limit 

= exp[ fIy + 1 A45 Gy] 

899 = 99th percentile daily maximum limit 

= exp[ ii,, + 2.326 &y]. 

Note: 

Zg5 = 1.645 

Z99 = 2.326. 

E-8 
The daily maximum variability factors (denoted by VFT) are estimated by: 

Daily maximum 95th percentile VFT = X.95 / k(X) 

Daily maximum 99th percentile VFT = X.99 / t(X). 

mkfth/y An?f8#@ Pem?it LhRits Based OR the 10#mn8/ Distfib#tioR 

This subsection contains the formulas required to approximate the distribution of the average of a small number 

of lognormally distributed values with another lognormal distribution. Although, the Central Limit Theorem 

holds that the average of a sample of independent measurements is normally distributed provided that the 

number of measurements, n, is sufficiently large, the minimum value for n required in specific cases may vary 

considerably. In cases where the individual values are lognormally distributed, the minimum required for the 

average to be normally distributed may be quite large. As a consequence, the distribution of the average of a 

small number of lognormally distributed values may be better approximated by another, related lognormal 

distribution [3]. For sample sizes larger than 10 when the data are lognormally distributed, it is recommended 

that the calculations given in Table E-3 should be used. For the purposes of permit writing, monitoring sample 

sizes of 10 or less are recommended to be “small enough” to assume the sample average is approximately 

lognormally distributed. The mean, variance, and coefficient of variation of the distribution of the average of 

n daily values are &,, k’,, and&, estimated by: 

.2 

nn = variance 

= In (V(X) / [n[~(X)121 + 1) 

fin = n-day monthly average 

= In&X)) - 0.5 &Z, 

4-t = standard deviation 

= (&y/2 

CGn = coefficient of variation 

= [exp($) - 1]1/2 

where 

k(X) = exp( by + 6; / 2) 

V(X) = exp(2 cy + 6;) [exp($ - 11. 

The upper percentile limits of the maximum n-day monthly average are: 

k *P = pth percentile n-day monthly average limit 

= exp[ in + Zp &n] 

where zp is the pth percentage point of the standard normal distribution. 

For example: 

A.95 = 95th percentile n-day monthly average limit 

= exp[ &, + 1.645 6t-J 

R.99 = 99th p ercentile n-day monthly average limit 

= exp[ tin + 2.326 ;n] 

E-9 
Note: 

Zg5 = 1.645 

Zgg = 2.326. 

The variability factors are: 

Monthly average 95th percentile VF, = R-95 / j.&, 

Monthly average 99th percentile VFn = R-99 / in. 

Delta-Lognormal Distribution 

The delta-lognormal distribution is a generalization of the lognormal distribution. The delta-lognormal 

distribution may be used when the data contain a mixture of nondetect values and values above the detection 

limit and can be used to model nondetects in water quality-based limits. In delta-lognormal procedures, 

nondetect values are weighted in proportion to their occurrence in the data. The values above the detection 

limit are assumed to be lognormally distributed values. The delta-lognormal distribution can be used in setting 

daily maximum limits and for setting limits on monthly averages with the recommended number of monitoring 

samples being 10 or less. 

The delta-lognormal distribution models data as the combination of two distributions: the lognormal 

distribution and a distribution with discrete probability of obtaining observations at or below the detection 

limit. The lognormal distribution models the observations above the detection limit. The nondetect values 

are modeled by the distribution with discrete probability of obtaining observations at or below the detection 

limit. The organic priority pollutant data set shown in Figure E-4 contains a number of observations that were 

reported as “nondetect.” These detection limit measurements are observations that are censored at the detection 

limit and are represented by the left-most bar in the histogram. Data sets of this form are fairly typical of 

organic chemicals in wastewater. The delta-lognormal distribution often provides an appropriate and 

computationally convenient model for analyzing such data. 

The estimation procedure for the delta-lognormal distribution assumes that a certain proportion, 6, of values are 

at the detection limit, which is denoted by D. (The estimation procedure when D = 0 is detailed in Reference 4. 

These values set to D are observations that can only by quantified as nondetect (ND) at some minimum level. 

This minimum level is the detection limit as estabtished by the laboratory performing the chemical analysis. 

Let XT ,x2 ,~*.nxpxr+l,-.-, Xk denote a random sample of size k, with r observations recorded as nondetects, and k-r 

observations greater than the detection limit. The k-r positive observations are assumed to follow a lognormal 

distribution. The entire data set is assumed to follow the delta-lognormal distribution with censoring point 

equal to the detection limit D. Let iy and r$ be the sample mean and variance of the distribution of the 

logarithmic transformation Y = In(X) of the observations greater than the detection limit. Let 8 be the sample 

proportion of nondetects. Then the estimates of the mean and variance of the delta-lognormal distribution are 

estimated by: 

w = daily average 

= SD + (1 - 8 ) exp( i$, + 0.5 6; ) 

W) = variance 

= (1 - 8 ) exp(2 4 + 8: ) [exp($) - (1 - 8 )] + 8 (1 - 8 ) D [ D - 2 exp( fiy + 0.5 c$ )] 

&(X’) = coefficient of variation 

= p(x*)p* /ax*> 

E-10 
160 

120 

40 

0 

t10 30 70 110 150 190 230 270 310 350 

Concentration in ugll 

Figure E-4. Organic Priority Pollutant Frequency Distribution-Plant B 

E-l 1 
where 

k = number of samples 

D = detection limit 

k-r 

= number of nondetect values in sample 

= number of values greater than the detection limit 

Yi = In(Xi) r+l 5 i 5 k, r < k 

= C&i) / (k - r) r+l I i I k, r c k 

= C(yi - iiy12 / (k - r - 1) r+l <ilk, r<k 

= r/k 

Dally Maxhum Pemtt limits Dased on the Delta-Lopomal DlstribiMn 

The 95th and 99th upper percentile limits for the random variable X (which is delta-lognormally distributed) 

are given by the following formulas: 

The estimated 95th percentile daily maximum limit is: 

rD 

k .95* = 1 

8 2 0.95 

Lmax [D, exp( jiy + ~‘6; ] $5 < 0.95 

where 

z* = o-1 [(0.99 - 8 ) / (1 - 8 )]. 

The estimated 99th percentile daily maximum limit is: 

I-D 8 2 0.99 

keg,* = I 

where 

Lmax [D, exp( iy + z*by)] 8 < 0.99 

z* = o-1 [(0.99 - s ) / (1 - 8 )I. 

W1 [ ] is the mathematical notation for Z-scores. For example, when 8 = 0, then the corresponding value is 

W1 [.99] = Z99 = 2.326. Values of Q-l [ ] are available from tables of the normal distribution (available in most 

statistical textbooks and references). 

The variability factors (denoted by VF) are estimated by: 

Daily maximum 95th percentile VF = R-95 / k(X) 

Daily maximum 99th percentile VF = A.99 / k(X). 

DebLopnmna~ Dlstributlon of Averages 

The derivation of the formulas for the averages computationally is difficult and beyond the scope of this 

appendix. However, the formulas for n-day averages are included in Table E-2. The derivation of 4-day monthly 

averages using the detta-lognormal distribution is available in Appendix VII-F of the Development Document 

for the OCPSF regulation [l]. For the purpose of permit writing, it is recommended that data sets of greater than 

10 samples be assumed to fit the normal distribution and the averages be calculated using the formulas given in 

Table E-3. 

E-l 2 
Checking Distributional Assumptions 

Two methods of checking distributional assumptions are goodness-of-fit and probability plots. When checking 

distributional assumptions, the sample size must be large enough. Small sample sizes may lead to erroneous 

conclusions. 

6oodmss-of-Fit Tests 

In some cases, statistical goodness-of-fit tests may indicate that a particular distribution provides a reasonable 

fit to a data set of pollutant measurements. Such cases should be evaluated carefully to verify that the frequency 

curve for the data also show the shape characteristic of the distribution. 

Probability Plots 

Use of probability plots is one method of determining whether a normal distribution is appropriate for 

modeling a population using only a limited set of measurements. The set of measurements should have at least 

20 observations [5]. Consider an independent sample of size k, labeled X1,X2,...,Xk. Let Ul,U2,...,Uk be the 

ordered sample of x-values in ascending order in which U1sU2s,...,<Uk. Now for each Ui, find Zi from the normal 

table (in any statistical reference or textbook) such that P[Z I Zi] = i/(k+l) and plot each pair (Zi,Ui) on linear 

graph paper (or use a computer graphics software package). If the data are from a normal distribution, they will 

fall approximately along a straight line. 

This same method can be adapted to check the assumption of lognormality. Log-probability plots are similar to 

probability plots used for the normal distribution. To construct a log-probability plot, set yf = In(xi) for 

i=l,2,..., k and then prepare a probability plot for the yi, first by ordering the data as described in the previous 

section. If the data are from a lognormal distribution, they will fall approximately along a straight line, as 

illustrated by Figure E-S. 

Other Dlstributfons 

If the probability plots or the log-probability plots show serious deviation from straight lines, other 

distributions should be considered. Nonparametric methods, which do not require the assumption that the data 

follow a particular distributional form, are often useful for this type of data. Further details are available in 

many statistical references (e.g., Reference 6). 

Up to this point, we have assumed that all the observed pollutant levels are independent, i.e., uncorrelated 

with one another. This subsection is not intended to address correlation between observed pollutant levels 

and plant operating factors that influence and control treatment performance. 

E-l 3 
-1.5 0.0 

2 Score 

1.5 3.0 

Figure E-S. Example of a Log-Probability Plot with a Normal Dlstributlon 

E-14 
In the case of the monthly average limit derivation, the assumption that observed pollutant levels are 

independent can be quite important. If the effluent levels are correlated, the actual monthly average limit can 

be substantially higher than that derived from the analysis based on the independence assumption. However, 

correlation has essentially no effect on the calculated daily permit limits. This sub-section provides guidance 

on determining when levels may be correlated, and adjusting the sample size. 

A mafor factor that determines whether effluent levels are highly correlated is the retention time of the 

wastewater treatment system. If the retention time is large relative to the time between effluent samples, then 

those samples will tend to be correlated with each other in most cases. In municipal systems, for example, the 

retention time is frequently a matter of days, and sampling is often conducted on a daily basis. The effluent 

levels, consequently, may be substantially correlated. However, in many industrial systems, for instance a 

physical/chemical treatment system for electroplating wastewaters, the treatment system retention time is 

relatively short 4 to 8 hours. Daily effluent levels from these kinds of systems are generally uncorrelated, i.e., 

statistically independent. These general patterns are the same irrespective of the kind of pollutant in question. 

Significant correlation between observed pollutant levels, when present, should be factored into monthly 

average permit limits. 

Several different methods can be used to account for correlation in determining limits. One general approach 

involves time series modeling. Another possible approach is to use a direct computation of the covariance 

among the observed data to adjust the variance of the average used in determining the limit. Help in adjusting 

the sample size for correlation is available from the OW Statistics Section (phone number [202] 382-5397). 

Table E-l. Daily Maximum Permit Limit Calculations 

The daily maximum permit limit is usually the 99th upper percentile value of the pollutant distribution. In 

:ertain cases the 95th percentile value may be allowable. The following gives the formulas: 

JVlTH ALL MEASUREMENTS > DETECTION LIMIT (based on lognormal distribution) 

8.95 = 

= 

A.99 = 

= 

ivtiere 

Xi = 

Yi = 

k = 

by = 

.2 

=Y = 

k(X) = 

cl(X) = 

&(X) = 

95th percentile daily maximum limit 

exp[&+ 1.645 $1 

99th percentile daily maximum limit 

exp[ j$ + 2.326 $1 

daily pollutant measurement i 

In(Xi> 

sample size of data set 

Uy$ I k 1 <ilk 

Z[(yi-by)‘]/(k-1) 1 <irk 

exp( f$ + 0.5 6; ) 

exp(2 f!iy + 6;) [exp($) - l] 

[exp(+ - 1]li2 

E-15 
Table E-l. Daily Maximum Permit Limit Calculations (continued) 

VITH SOME MEASUREMENTS < DETECTION LIMIT (based on delta-lognormal distribution) 

k.95 = 95th percentile daily maximum limit 

rD 8 2 0.95 

495 = I 

L max [D, exp( $y + z* $1 8 < 0.95 

vith z’ = a-l [(0.95 - 8 ) / (1 - 8 )] 

A .99 = 99th percentile daily maximum limit 

rD 8 2 0.99 

k I .99 = 

Lmax [D, exdy + ~~$1 8 < 0.99 

with z* = Q-l [(0.99 - 8 ) / (1 - 8 )] 

yhere 

xi = daily pollutant measurement i 

k = sample size of data set 

D = detection limit (as established by the laboratory) 

r = number of nondetects (xl ,x2,...,xr are 2 D) 

k - r = number of detects (Xr+ltXr+2,...tXk are > W 

3’ 

= Ill(Xi) for r+l I i 5 k 

= r/k 

LY = E&i) / (k - 0 r+l 5 i 5 k (exclude values I D from sum) 

.2 

=Y = E[(yi - i$21 / tk - r - 1) r+l <ilk 

&x*) = 80 + (1 - 8 ) exp( i$, + 0.5 6;) 

0(x*) = (1 - 8 )exp(Z i$, + 8;) [exp(&$ - (1 - 8 )] + 8 (1 - 8 )D[D - 2 exp( i$, + 0.5 SC)] 

E-l 6 
Table E-2. Monthly Average Permit Limit Calculations for 

Ten Samples or Less 

The monthly average permit limit is usually based on the estimates of the 95th percentile of the distribution 01 

:he average of the daily effluent values. For sample sizes less than or equal to 10, the data are assumed to br 

ognormally distributed (or delta-lognormally distributed if the data includes nondetects). 

4LL MEASUREMENTS > DETECTION LIMIT (based on lognormal distribution) 

A.95 = 95th percentile n-day monthly average limit 

= exp[i, + 1.645 kn] 

k-99 = 99th percentile n-day monthly average limit 

= exp[ j$., + 2.326 ;m] 

Nhere 

xi = daily pollutant measurement i 

Yi = Ill(Xi) 

k = sample size of data set 

r; 

3 

= C&i) / k l<ilk 

&Y = E[(yi-b) ]/(k-1) 2 

Y 1 li<k 

k(X) = exp( jYiy + 0.5 6; ) 

V(X) = exp(2 iy + B$[exp($ - I] 

A2 

on = InCOoQ / (n[ h12> + 1) 

I& = In&x)) - 0.5 62, 

C"Vn .2 

= [exp(o,) - 1]lj2 

E-l 7 
Table E-2. Monthly Average Permit Limit Calculations for 

Ten Samples or Less (continued) 

,OME MEASUREMENTS < DETECTION LIMIT (based on delta-lognormal distribution) 

k 95 = 95th percentile n-day monthly average limit 

rD s 2 0.95 

495 = 1 

Lmax [D, exp( cn + z*&,)] 8 < 0.95 

vith z* =@-I[ (0.95 - !Zi ) / (1 - $)I. 

k 99 = 99th percentile n-day monthly average limit 

rD 6 2 0.99 

899 = 1 

L max [D, exp( j$, + z*&r,)] 8 < 0.99 

vith z* = W1 [(0.99 - 8 ) / (1 - 8 )] 

vhere 

xi = daily pollutant measurement i 

k = sample size of data set 

D = detection limit (as established by the laboratory) 

r = number of nondetects (x1,x2,..., xrare<D) 

k - r = number of detects (Xr+l?r+2s-.tXk are > D) 

Yi = In(Xi) for r+l 5 i I k 

8 = r/k 

fiY = Z(yi) / (k - r) r+l 5 i I k (exclude values ID from sum) 

.2 

“Y = Z[(yi-jiy)211(k-r-l) r+l I i I k 

ik(X*) = 8 D + (1 - 8 ) exp( j?ty + 0.56;) 

v(X*) = (1 - 8 ) exp(2 by + 6:) [exp( 6;) - (1 - 8 )] + 8 (1 - 8 )D[D - 2 exp( fiy + 0.5 $1 

.2 

% = In{(l - 8”) [l + A + B + C]) 

vith 

A = 0(X*) / [n&X*) - 8” D)2] 

6 = - 16” D2(1 - 8”)] / &X*) - 6” D)2 

c = (2 8 D) / &X*) - 8’ D) 

fin = In[(l!(X*) - 8” D) / (1 - 8”)] - 0.5 6: 

E-18 
Table E-3. Monthly Average Permit Limit Calculations for More Than Ten Samples 

‘he monthly average permit limit usually is based on the estimates of the 95th percentile of the distribution 

>f the average of the daily effluent values. These daily values are assumed to be lognormally distributed. For 

#ample sizes larger than 10, the averages (represented by the random variable X,) are assumed to be normally 

Sstributed. 

k.95 = 95th percentile n-day monthly average limit 

= ^E(X,) + 1.645 fl&)]‘/2 

499 = 99th percentile n-day monthly average limit 

= i(Xn) + 2.326 yC/(Xn)]‘/2 

vhere 

xi = daily pollutant measurement i 

Yi = In (Xi) 

k = sample size of data set 

F-Y = Ii / kt l<i<k 

*2 

“Y = Z[(yi - fiyj2] / (k - 1) 1 5 i 5 k 

&)o = exp( $y + 0.5 ire ) 

w> = exp(2 jI$ + 6:) [exp($) - 1 ] 

‘bn) = E(X) 

0(X,) = V(X) / n 

C”v(Xn) = V(Xn)‘j2 / (Xn) 

E-l 9 
APPENDIX E 

REFERENCES 

1. 

2. 

3. 

U.S. EPA. 1987a. Development Document for Effluent Limitations Guidelines and Standards for the 

Organic Chemicals, Plastics and Synthetic Fibers. EPA 440/l -87/009. 

Mendenhall, W., 9. Scheaffer, and D. Wackerly. 1981. Mathematical Statistics with Applications. 2d 

ed. Wadsworth, MA. 

Barakat, R. 1976. Sums of Independent Lognormally Distributed Random Variables. J. Optical Sot. Am. 

66:21 l-l 6. 

4. 

5. 

Aitchison, J., and J. Brown. 1963. The Lognormal Distribution. Cambridge University Press. 

Johnson, R., and D. Wichern. 1982. Applied Multivariate Statistical Analysis. New Jersey: PrenticeHall. 

6. Hollander, M., and D. Wolfe. 1973. Nonparametric Statistical Methods. New York: Wiley. 

ADDITIONAL REFERENCES 

Kahn, H. 1989. Memorandum: Response to Memorandum from Dr. Don Mount of December 22, 1988. 

August 30, 1989. U.S. EPA, Washington, DC, to J. Taft, U.S. EPA, Permits Division, Washington, DC. 

Kahn, H., and M. Rubin. 1989. Use of Statistical Methods in Industrial Water Pollution Control 

Regulations in the United States. Environmental Monitoring ond Assessment 12:129-48. 

E-20 
APPENDIX F 

SAMPLING 
SAMPLING 

The objective of an effluent or instream sampling program is to obtain a sample (or samples) from which a 

representative measure of the parameters of interest can be obtained. Unfortunately, many of the industrial and 

municipal National Pollutant Discharge Elimination System sampling protocols presently in use are carryovers 

from schemes used for calculating loadings of nutrients and oxygen-demanding substances, or were developed 

to evaluate treatment plant operational efficiency. Sampling for individual toxicants and particularly for 

effluent toxicity can require more specific (and thus different) sampling procedures. 

Wastewater variability is an important consideration in selecting the method and frequency of sampling for 

both chemical analysis and toxicity testing. Industrial waste characteristics have been shown to vary in 

frequency, intensity, and duration [1]. As noted by Bender [2], the sources of effluent variability include both 

random and systematic components that influence both daily and annual characteristics of waste discharges. 

Although toxic pollutant loading may be of primary concern in assessing human health impact or 

bioaccumulation, loading may be of lesser importance in toxicity assessment than frequency, intensity, and 

duration of peak toxic discharge. Sampling must be tailored to measure the type of toxicity of importance for 

that discharge: either long-term (chronic) impact, which is a more constant effect, or short-term (acute) impact, 

which is more variable and subject to peaks of intensity. 

There are several chemical parameters for which continuous analysis is possible. These include pH, temperature, 

dissolved oxygen, and other parameters involving instantaneous measurement. All other types of measurement 

involve some time period over which the analysis is conducted. Toxicity tests require an exposure period. 

Chemical tests require sample preparation and analysis. There is no continuous analysis method for toxicity. 

It should be noted that although it is difficult to design a representative sampling program for toxicity 

analysis, the problems are of no greater magnitude than similar problems associated with obtaining a 

representative sample for conventional pollutants. 

Sampling Methods 

Continuous Flow Samples 

For toxicity testing, the test organisms may be exposed to serial dilutions of a sample continuously pumped 

from the effluent pipe or ditch. In the case of effluents, if optimum accuracy is desired, then the ratio of 

effluent flow to test chamber volume can be scaled to simulate the time-varying concentration at the mixing 

zone boundary. 

Although flowthrough methods can provide a realistic simulation of time-varying exposure, they are relatively 

expensive and are usually conducted on site. Therefore, flowthrough methods may only be practical where the 

goals of the analysis of impact require this type of testing or where treatment costs are sufficiently high that 

this type of analysis can be required. A flowthrough exposure method is not a continuous analysis because only 

one result or data point is obtained at the end of the test. However, the continuous exposure does provide some 

measure of time-varying exposure effects. 

Discrete Samples 

Grab or flow composited sampling provide a discrete sample for chemical analysis or toxicity testing. Static or 

renewal toxicity tests using discrete samples result in exposure of test organisms to a constant effluent 

composition over the period of the tests, or for the period between renewals. 

If discrete samples are collected during peaks of effluent toxicity then constant concentration exposure static 

tests provide a measure of maximum effect. 

Depending on the duration of a peak and the compositing period, composited samples may not be useful for 

examining toxicity peaks because the compositing process tends to dilute the peaks. Cornposited samples are 

F-l 
usually appropriate for chronic tests where peak toxicity of short duration is of less concern. The averaging 

effect of compositing may be misleading when testing for acute toxicity. 

Grab samples must be collected at sufficiently frequent intervals to provide a high probability of sampling 

daily peaks. Fortunately static toxicity tests are relatively inexpensive and can be done on shipped samples; 

thus, it may be cost effective to conduct individual tests on a series of grab samples collected over a 24-hour 

period. 

Sampling Frequency 

Nonrandom effluent variability, resulting from batch processing, variable loadings, etc., is often known or can 

be determined. Therefore, the first step in designing a sampling program for chemical analysis or toxicity 

testing is to select the annual sampling frequency based on available site-specific operational information. 

This is important in selecting sampling periods for both continuous flow and discrete sampling methods. 

If discrete sampling methods (grabs or composites) are used, then random variations between and within days 

for each sampling period must be considered. It is important to recognize the tradeoff between the long-term 

(between days) frequency and short-term (within days) frequency of sample collection and analysis for toxics. 

At present, the permit requirements for sampling and analyzing chemical parameters are site specific and 

generally involve a single grab or 24-hour composite sample collected at daily, weekly, or monthly intervals. 

Unfortunately, a sampling scheme involving a single daily grab or a 24-hour composite sample can conceal the 

presence of those daily extreme values that may be of importance. To optimize sampling cost and 

effectiveness, it may be desirable to reduce long-term frequency so that daily frequency can be increased. 

For example, a weekly grab or composite involves 52 analyses per year. It may be more efficient to reduce the 

annual frequency to monthly or bimonthly, but collect and analyze four or eight grabs daily. Either scheme (12 

x 4 or 6 x 8) would involve 48 analyses per year versus 52 for the weekly single sample approach. Assuming 

that daily toxic events of environmentally significant intensity and duration would not be masked by shortterm composites, it might be more efficient to collect eight samples each composited over a 3-hour interval. 

If costs or other constraints prohibit satisfactory daily and annual replication of sampling, then a level of 

uncertainty must be introduced into the calculations used to evaluate waste toxicity (see Section 3, Table 3-l). 

Box F-l presents EPA’s recommendations on sampling methods. 

F-2 
Box F-l. Recommendations 

The initial sampling design step should involve stratification of sampling periods to account 

for nonrandom sources of variation (e.g., batch processing). The second step includes 

selection of the frequency and the method of sampling to be conducted within each sampling 

period. Depending on site-specific considerations, several options are available. 

l Flowthrough Methods - Ideally, for both acute and chronic effluent toxicity tests, the 

exposure of biota should simulate the time-varying concentration at a predetermined point in 

the receiving water. For regulatory purposes, the critical point is often the edge of the mixing 

zone where the waste should exhibit neither acute nor chronic toxicity. Therefore, if 

warranted by site-specific factors, it is recommended that test biota be exposed to a 

continuously collected flowthrough sample of serially diluted effluent. If no systematic 

annual variations (e.g., batch processing) are known or suspected, flowthrough testing can be 

conducted at a minimum of quarterly intervals for at least 1 year. 

l Grab Sample Methods - Grab samples are recommended for chemical analyses and for acute 

and chronic toxicity tests where site conditions (such as wastewaters that are known to have 

relatively constant composition) do not require use of continuous flow methods. Grab 

samples of effluent or receiving water may be used for static or renewal acute toxicity tests, 

which may be conducted onsite or at a remote lab. The design of a toxics grab-sampling 

program must take into account the tradeoff between long-term and short-term sampling 

intensity. Where there is no ponding of wastes or retention time is insufficient for thorough 

mixing, it is important to collect or analyze a sufficient number of samples to provide a 

measure of daily spikes. Therefore, to minimize analytical costs where daily fluctuations are 

known or suspected, the annual sampling frequency should be reduced in favor of more 

intensive daily sampling. It is recommended that on an annual cycle, grab sampling and 

analysis include a minimum of four to six daily grabs collected monthly. An option could 

include the use of short-term (4-hour) composites rather than grabs. If site-specific data are 

available to indicate that treatment system retention time is adequate to minimize daily 

variations, then the daily replicates may be omitted in favor of more frequent annual sampling 

(e.g., weekly or semimonthly rather than monthly). If, to minimize costs in screening tests, 

only single samples are collected at infrequent intervals (e.g., quarterly) an uncertainty factor 

for variability should be used in the toxicity evaluation (see Section 3). 

l Composite Sample Methods - If static or renewal methods are used for evaluation of toxicity, 

it is recommended that 24-hour, continuous-flow composite samples be collected. 

Considerations of annual frequency are the same as those for grab samples. 

F-3 
APPENDIX F 

REFERENCES 

1. Nemetz, P.N., and H.D. Dreschler. 1978. The Role of Effluent Monitoring in Environmental Control. /. 

Water, Air, and Pollution 10:477-97. 

2. Bender, E.S. 1984. “Sources of Variations in Effluent Toxicity Tests.” In Environmental Hazard Assessment of 

fffluents. Ed. H. Bergman, R. Kimerle, and A.W. Maki. Proceedings of the Fifth Pellston 

Environmental Workshop, Cody, WY, August 1982. 

F-4 
APPENDIX G 

THE DEVELOPMENT OF A BIOLOGICAL INDICATOR 

APPROACH TO WATER QUALITY-BASED HUMAN HEALTH 

TOXICS CONTROL 
THE DEVELOPMENT OF A BIOLOGICAL INDICATOR 

APPROACH TO WATER QUALITY-BASED HUMAN HEALTH 

TOXICS CONTROL 

Current Approach 

With one exception (New Jersey), the chemical-specific approach to protecting human health is currently the 

only method used to regulate human health toxicants in effluents. The chemical-specific approach identifies 

the individual chemicals in an effluent and regulates them based upon health risk assessment information for 

each individual chemical. Where data are available for such human health toxicants, the chemical-specific 

approach can be used to develop permit limits. 

However, the complex characteristics of effluent mixtures limit the effectiveness of the single-chemical 

approach. When used as the sole basis for identifying effluents of human health concern, the chemical-specific 

approach can overlook wastewaters potentially toxic to humans for the following reasons: 

1. Analytical methods may not be sensitive enough to detect extremely small quantities of chemicals 

which may exert their effects on human health after a long latency period. 

2. Human health data are limited or lacking for many of the §307(a) “priority” pollutants. Moreover, 

the number of human health toxicants discharged far exceeds the “priority” pollutants list. 

3. The various chemical constituents of an effluent may resulting in synergistic, additive or 

antagonistic chemical effects. 

As a result of these limitations, biological indicator tests have been developed for human health impact 

effluent screening, including both in vitro and in vivo tests. Though not yet widely implemented, biological 

indicator test results can be important supplements to a chemical-by-chemical effluent characterization. 

Short-term biological indicator tests for human health impact screening are based on cellular-level responses, 

indicating whether the substances being tested are biologically active, and providing some measure of that 

activity. While these tests do not quantify the degree of toxicity to humans, they can be used to identify 

effluents with potential human health impacts, and regulatory priority-setting and targeting of dischargers for 

further chemical-specific analyses. Research is currently underway within EPA and in the private sector to 

evaluate various biological indicator test batteries for whole effluent analysis. 

Biological Indicator Tests 

Biological indicator tests include in vitro (test tube) and in vivo (whole animal) tests which can help form the 

first tiers of a single chemical evaluation process. A battery of simple biological tests can be used to test for 

the major types of effects which are underlying causes of potential health impact, since each biological test 

measures a different type of response. The results of these tests can be used to decide whether more definitive 

(and more resource-intensive) testing is needed to identify actual problem pollutants. 

Test results can serve as triggers to additional chemical-specific analysis or more sophisticated definitive 

biological tests. Where results of these screening tools indicate potential health hazards, further 

characterization of the effluents, and regulation based upon toxicological data and/or chemical structureactivity relationships can proceed. If an effluent is extremely variable in other parameters, screening assays 

should be repeated periodically to ensure that potentially hazardous discharges are detected. Two types of 

biological indicator tests are discussed below: tests for non-threshold (no safe level exists) chemicals and tests 

for threshold (a safe level is presumed to exist) chemicals. 

G-1 
Genotoxicity Tests for Non-Threshold Chemicals 

Cenotoxicity is the ability of a substance to damage an organism’s genetic material (its DNA). Certain 

positively-charged compounds tend to bind to DNA and may lead to permanent changes in the genetic 

information. Such damage to the DNA of reproductive (germ) cells can impair reproductive ability or can 

produce a change in the DNA structure that could be passed on to offspring as a heritable mutation. Alterations 

in the DNA of somatic cells can result in cancer or other diseases. 

Interpretation of genotoxicity test results assumes that DNA damage in nonhuman cells may be predictive of 

latent diseases in humans such as genetic disorders, birth defects, and cancer. EPA believes that genotoxicity 

tests for point mutations, numerical and structural chromosome aberrations, DNA damage/repair and in vitro 

transformation provide supportive evidence of carcinogenicity [U.S. EPA, 1979 and 1987~1. In addition, 

wastewater mutagenicity tests could be used to detect genotoxic activity which can adversely affect aquatic 

biota [Black, et. al., 19801. Several short-term assays have been developed which can assess genotoxic effects 

(discussed below). 

For example, a correlation has been established between animal carcinogens and positive mutagenic responses 

in the Ames Test. The Ames test is often used to assess point mutation effects. The original correlation study 

revealed that 90% of tested carcinogens were detected as mutagens, while 87% of noncarcinogens were 

identified as nonmutagens. Other studies have determined that between 77% and 91% of tested carcinogens 

produce positive responses in the Ames test. The Ames Test has been used in over 2,000 laboratories worldwide 

for drug and food additive screening, product development, and environmental testing [New jersey DEP, 19831. 

To assess clastogenic effects (chromosomal breakage) either the mammalian sister chromatid exchange test or a 

mammalian ceil chromosomal aberrations test can be conducted. Both of these tests typically use Chinese 

hamster ovary cell cultures and involve cytologic examination after exposure to determine if chromosomal 

effects are evident. The Organization of Economic Cooperation Development (OECD) test methodology is 

recommended [OECD]. EPA’s Office of Toxic Substances and Office of Pesticides Programs also have published 

test methods [U.S. EPA 1982a and 1982b] that are consistent with the OECD tests. 

Most effluent samples need special preparation (for example, concentration) to produce a measurable biological 

indicator test response for human health effects. When samples are concentrated, the response is calculated in 

terms of the pre-concentration sample. In addition, for genotoxicity tests, because many chemicals are not 

actively mutagenic in humans until they enter the body and are metabolized, many in vitro tests are 

supplemented with extracts from mammalian livers which act as a source of enzymes. The extract enzymes act to 

mimic metabolic activation of procarcinogens and promutagens in humans, providing a more realistic picture 

of potential effects [U.S. EPA, 19791. 

A number of genetic toxicity assay batteries have been suggested in order to address the many potential effects 

produced by nonthreshold chemicals (for which no safe level exists) [U.S. EPA, 1979; Lave and Omenn; 

Environment Canada]. In addition to providing assays that detect different endpoints, a battery of tests can 

also be structured to minimize effort at the screening level while supplying more definitive data for samples 

failing the initial tier of testing. Positive results can lead to further effluent characterization, including 

priority and other pollutant chemical analyses, or mutagenicity testing of specific processes or effluent 

fractions. Another approach would be to evaluate the effects of various treatment or waste segregation 

techniques on mutagenicity [McGeorge, et. al,, 19851. 

Many of the proposed test batteries utilize the Ames Assay as a screening level test because of its relatively 

high degree of sensitivity (i.e. a high percentage of carcinogens are Ames positive) and specificity (i.e. a high 

percentage of noncarcinogens are Ames negative) [Tennant, et. al., 19871. One study of 28 selected industrial 

discharges revealed that 11 of the 28, or 39%, produced positive results using the Ames Test (described below). 

Other test endpoints frequently covered in the initial tier of testing include mammalian cell chromosomal 

effects, mammalian gene mutation and microbial and mammalian cell DNA damage. 

Results of a recent National Toxicology Program project suggest that combinations of four of the most 

commonly used short-term tests covering these endpoints did not show significant differences in individual 

G2 
concordance with rodent carcinogenicity results for pure chemicals [Tennant, et. al., 19871. This suggests that 

if a sample causes only one type of endpoint as measured by several screening level tests, its potential to cause 

human health effects should not be disregarded. 

TO assess the potential carcinogen hazard, subsequent tests focusing on effluent-induced malignant changes in 

mammalian cells in vitro can be conducted. Higher levels of testing may include in vivo rodent testing or the 

Medaka (fish) tumor assay, for example. It should be noted that under existing guidelines, in vivo mammalian 

tumor assays are necessary to establish a material as a possible human carcinogen. Results from short term tests 

alone are considered as inadequate to establish human carcinogenicity [U.S. EPA, 1986~1. Guidelines for risk 

assessment of individual compounds are covered in U.S. EPA, 1986b and 1987~. 

In vivo tests on complex mixtures are extremely complicated and expensive given the variability intrinsic to 

effluents. As a result, it is recommended that after each tier of biological indicator testing, the cost of further 

refining the weight of evidence for carcinogenesis or mutagenesis be balanced against the cost of conducting a 

causative agent identification evaluation. Given the identity of the substance leading to positive results in 

short term in vitro tests, it should not be necessary to generate in vivo dose-response data for risk 

characterization if these data are already available in the literature for the specific chemical. 

In addition, causative agent identification studies may be unnecessary if information on the physical and/or 

chemical characteristics of the toxicant is obtained. Such information may provide clues to appropriate 

effluent treatment technologies needed to reduce effluent mutagenicity. 

In weighing the need for more definitive biological assays against causative agent evaluation, the frequency 

(i.e., how often the effluent tests positive) and intensity (e.g., revertants/liter) of the effluent’s mutagenicity 

must be considered. As a default assumption, a high dose of a carcinogen received over a short period of time is 

equivalent to a low dose spread over a life-time [U.S. EPA, 1986~1. While effluents which are highly variable 

in their mutagenicity are of concern, they will be more difficult and costly to deal with in subsequent phases of 

study. 

Accordingly, the initial tier of qualitative tests for human health effects assessment can be relatively 

inexpensive, rapid, and have a low rate of false negative results. Subsequent tests can be designed to increase 

confidence in the predictive nature of the results. Additional levels of testing may also provide diagnostic 

information on the characteristics of the causative agent(s) in the effluent. 

Subsequent tiers of testing should focus on a more concise assessment of risk. Such an assessment can be used to 

delineate hazard type; in effect, to separate germ cell mutations (heritable genetic risk) from carcinogen risk. 

Thus, to assess heritable mutation, subsequent testing should focus on mammalian germ cells, ultimately tested 

in vivo [U.S. EPA, 1986b]. To assess potential carcinogen hazard, subsequent tests focusing on effluent-induced 

malignant changes in in vitro mammalian systems should be conducted. Ultimately, testing must result in a 

dose-response assessment to be used with an exposure assessment in characterizing risk [U.S. EPA, 1987a]. 

EPA’s Region V (Chicago), New Jersey, and Environment Canada have been conducting mutagenicity testing at 

selected facilities. In Region V Ames test results are used to suggest the need for more intensive chemicalspecific analyses of the effluent. New Jersey has incorporated a prohibition against discharging mutagenic 

compounds in amounts that are mutagenic into its “New Jersey Administrative Code” [N.J.S.A. Section 7:9-4.5 

(a)4, May 19851. 

For both types of endpoints (genotoxicity and carcinogenesis), hazard identification should be followed by 

quantitative risk assessment which includes assessment of dose response (requiring in vivo data) and human 

exposure. Human exposure assessment typically considers the composition and size of the population exposed 

and the types, magnitude, frequency and duration of exposures [U.S. EPA, 1986d]. 

Evaluation of Effluent Genotoxicity Screening Results 

Control of human health hazards depends upon assessment of both the toxicological properties of the 

pollutants and the level of exposure. The permit authority should review the results of a human health toxicant 

c3 
effluent screening program and establish the actions triggered by each level of potential risk indicated. For 

example, a discharger with either a high exposure risk or a high effects risk might automatically be required to 

conduct a detailed assessment or institute controls. A medium risk in both exposure and effects might require 

further review of the data and a case-specific decision about whether to require additional assessment. A 

medium effects risk and a low exposure risk might indicate the need for limited testing to ensure that the low is 

really indicative of the risk. Low risk in both exposure and effects should receive low priority for further 

assessment. The bioconcentration evaluation procedures can be used to aid in defining exposure risk, as well as 

determining receiving water concentration. 

One possible tool for evaluating results of biological indicator effluent screening is the “relative potency 

approach,” a concept used rather widely in radiation biology and chemical pharmacology. The relative 

potency of an effluent is the dose of a reference agent needed to produce an effect of a given magnitude in a 

particular bioassay, divided by the dose of the effluent needed to produce the same magnitude of the same 

effect in the same bioassay. A predictive battery of several short-term biological tests, when standardized to a 

reference agent, could provide a rank or comparative estimate of the hazard posed by an effluent in the context 

of measures of other known hazards [Glass, 19881. It should be recognized that this approach does not consider 

exposure through bioaccumulation. 

When screening has indicated a high potential for health hazard, further assessment should be required. A 

chemical-specific approach is recommended to evaluate and regulate the discharge constituents. The first half 

of this process involves characterizing the composition of the effluent. Typically, only a small fraction of the 

total organic carbon (TOC) can be accounted for as individual chemicals. Therefore, effort should be placed on 

identifying constituents through means other than chemical analysis, such as through a detailed process 

evaluation and/or toxicant characterization evaluation. 

A process evaluation is a study in which components in the wastewater are determined from an analysis of 

feedstocks, manufacturing processes, products, by-products, and pollution control in place. The result is a list 

of compounds or classes of compounds with a high probability of being present in the wastewater. Chemical 

analysis can also be conducted for not only the priority pollutants but also nonpriority pollutant peaks and 

bioconcentratable chemicals [EPA/6OO/XX-XX]. IRIS and SAR can be used to determine the likelihood that a 

given compound is causing positive results in the bioassay. The toxicant characterization evaluation can 

provide information on the physical/chemical nature of the chemical producing positive bioassay results. 

Summary of Current Blological Indicator Tests for Non-Threshold Human Health Toxicants 

The following tests are currently in use or under development for assessing carcinogenicity or mutagenicity: 

. Salmonella tvphimurium Assay (Ames Test) [U.S. EPA, 1985 and 19831 

Backoround: Strains of Salmonella requiring the amino acid, histidine, are exposed to a solvent 

extract of the effluent. Tests are performed with and without added rat liver enzyme for activation 

of indirect mutagens. The bacteria are grown on histidine-free medium; colony formation 

indicates the effluent contains mutagenic compounds capable of genetically altering the bacteria. 

Endboint: Gene mutation; response measured in revertant colonies/L effluent. 

Advantaou: Test is rapid, relatively inexpensive. The Ames Test has been shown to have broad 

application for the assessment of the mutagenic activity of a diversity of industrial effluent types 

(McCeorge, et. al., 19851. Test sensitivity and specificity are documented [Ashby and Tennant, 

19881. 

CA 
Disadvantaaes: Requires metabolic activation and several different strains of Salmonella to detect 

a broad range of compounds, requires extrapolation from prokaryot, use of effluent extract may 

exclude certain types of compounds, epigenetic carcinogens not detected. 

Cost: Approximately $1200 [Lave and Omenn, 19861 

l Escherichia coli SOS Assay (SOS Chromotest) [Quillardet, et.al., 19851. 

Backaround: All cells contain an “SOS” enzymatic system for detecting and correcting errors in 

their genetic material. A strain of E. coli has been genetically engineered so that DNA damage 

ultimately results in production of an enzyme which reacts with test reagents to form a blue color. 

Bacteria are exposed to effluent or an extract of the effluent, with or without added rat liver 

enzyme for activation indirect mutagens. The intensity of color produced indicates the extent to 

which the effluent contains mutagenic compounds capable of damaging bacterial DNA. 

Endpoint: DNA damage; response measured as the change in optical density. 

Advantaaes: Simple kit commercially available, test requires <8 hrs to perform, relatively 

inexpensive. Test sensitivity, specificity documented [Quillardet, et.al., 19851. 

Disadvantaaes: Requires metabolic activation, extrapolation from prokaryot, use of effluent 

extract may exclude certain types of compounds, epigenetic carcinogens not detected, measurement 

of effect must be referenced to known genotoxic compound. 

Cost: ?? 

l Sister-Chromatid Exchange Assay (SCE) [Eckl, et. al., 19871 

Backaround: 

division. 

Sister chromatid exchange occurs when damaged DNA is replicated during cell 

Recent advances allow the use of cultured rat hepatocytes in detecting SCE formation, 

thus precluding the need to add rat liver enzyme for metabolic activation. Hepatocyte exposure to 

the sample is effected by using filter sterilized effluent in preparing the cell culture medium. 

Exposed cells are lysed and genetic material fixed in order to count SCEs. 

Endpoint: DNA damage; response measured in SCE per chromosome/L effluent. 

Advantages: Test is rapid, relatively inexpensive, does not require metabolic activation (therefore 

more realistic). Uses mammalian cells, therefore results more readily applicable to humans. 

Disadvantaaes: Sensitivity, specificity not well documented, test more complex relative to 

prokaryotic systems, filter sterilization may remove some genotoxic compounds from the sample, 

epigenetic carcinogens not detected. 

Cost: $5000 [Jirtle, 19891 

l HCPRT Assay with Chinese Hamster Ovary Cells (HCPRTKHO) [Hsie, et. al., 19811 

Backaround: Strains of Chinese Hamster Ovary cells in culture are exposed to the effluent or an 

extract of the effluent, with or without added rat liver enzyme. Mutagen interactions with certain 

sections of the DNA make the cell resistant to toxicants like Gthioguanine. Cell survival is used 

to indicate both cytotoxicity (cell death) and genetic mutations resulting from effluent 

components. 

Endpoint: Gene mutation; response measured in % survival/L. 

Advantaaes: Test is rapid and uses a mammalian system. 

Disadvantaaes: Sensitivity, specificity not well documented, use of effluent extract may exclude 

certain types of compounds, epigenetic carcinogens not detected, requires metabolic activation. 

-. Cost. $6500 

l Medaka Tumor Assay [U.S. EPA, 1988; U.S. EPA, 1989b.l 

Backaround: Larval fish are exposed to nonlethal concentrations of effluent for one month, this 

period is followed by a S-month grow out period in clean water. At six months, fish are sacrificed 

and submitted for histopathological studies. 

Endooint: Tumor formation, response measured in frequency of tumors at a given site/effluent 

concentration. 

Advantaaes: Use of whole effluent, whole organism, oncogenic endpoint 
Disadvantaaes: Carcinogen levels in unconcentrated effluent may not be high enough to produce 

tumors at a detectable frequency in exposed populations, effluent must not be toxic to Medaka, 

requires extrapolation from non-mammalian system, relatively expensive, length of test, endpoint 

requires pathologist experienced in fish cancers, method still in developmental stages. 

Cost: $20,000 [Johnson, 19891. 

Other Human Health Effects 

Toxicants present in effluents may produce a variety of effects in humans besides genotoxicity or 

carcinogenicity via exposure through ingestion of water and/or contaminated fish and shellfish. Potential 

health effects could include suppression of the immune system, neurotoxicity, specific organ toxicity, or 

developmental toxicity. These effects occur after exposure above a presumed safe (threshold) level and are 

referred to as “systemic.” 

Formerly, the only means to assess systemic effects was by using subchronic toxicity procedures designed to 

determine the effects that may occur with repeated exposure over a part of the average life span of an 

experimentat animal. However, such studies are expensive ($100,000 and over) and beyond the cost constraints 

for most effluent analyses. As an alternative, a number of short-term in vitro tests utilizing mammalian cells 

have been developed [U.S. EPA, 1978; Wilson, 1978; Kimmel, et. al., 1982; Brown and Fabro, 1982; 

Borenfreund and Puerner, 19851. Test endpoints include cytotoxicity, effects on cell growth, division, structure, 

metabolism and function, alterations in enzyme activities, and metabolite formation. 

As with the nonthreshhold assays previously discussed, these in vitro assays only serve to qualify potential 

human health hazards. In the case of positive in vitro results, tests on intact mammals can be pursued in order 

to confirm screening test findings and establish a dose-response relationship. Alternatively, causative agent 

evaluations resulting in either the identity of the toxicant or toxicity treatability data may be pursued. 

Current limitations of the Biological Regulatory Approach 

At present, the use of biological indicator tests as a regulatory tool is limited for a number of reasons. First, 

biological indicator information must be linked to human exposure to wastewater components. To date, no 

definitive mechanism exists for interpreting the human health hazard implications of the biological test 

results. While many in vitro (i.e. test tube) human health assays provide data about cellular changes relative to 

the dose delivered to the target tissue, they do not provide the information necessary to correlate 

environmental exposure to target tissue dose or cellular change to ultimate human health effects (e.g., cancer). 

The higher animal testing necessary to quantify the dose-response relationship (or “potency” of the effluent) 

would be extremely costly. 

Second, as with aquatic organism toxicity tests, a human health hazard test must be capable of dealing with 

intra- and interspecies sensitivity variability. This concern is particularly relevant for those effluents 

containing chemicals which only become carcinogenic upon metabolism by mammalian systems (i.e. 

procarcinogens). The use of cultured human liver cells (hepatocytes), currently being tested, would eliminate 

the need for interspecies extrapolation. 

Finally, whole effluent testing to assess potential human health impacts presents several unique practical 

problems such as the continual change in composition typical for most effluents, the need to concentrate 

samples to obtain a dose-response curve, and the need to compensate for or eliminate interferences from 

cytotoxic (toxic to ceils) components of the effluent. Only those components which occur in the relatively 

nonvolatile, nonpolar organic fraction of the effluent sample are conventionally measured. [AndersonCarnahan, article in preparation]. 

Until additional research resolves these difficulties, biological indicator tests will be most useful as screening 

tools, with actual regulation of effluents posing potential health hazards likely to remain on a chemical-bychemical basis. 

CA 
1. 

2. 

3. 

4. 

5. 

6. 

7. 

8. 

9. 

16. 

18. 

APPENDIX G 

REFERENCES 

Anderson-Carnahan, L., P. M. Eckl, and R. L. Jirlle. Sister Chromatid Exchange in Screening Wastewaters. 

In preparation. 

Ashby, J. and R.W. Tennant, 1988. Chemical Structure, Salmonella Mutagenicity and Extent of 

Carcinogenicity as Indicators of Genotoxic Carcinogens Among 222 Chemicals Tested in Rodents by 

Ihe U.S. NCI/NTP. Mutation Research 204:17-l 15. 

Black, J., P. Dumerski, and Zapisek, W. 1980. “Fish Tumor Pathology and Aromatic Hydrocarbon 

Pollution in a Great Lakes Estuary.” In B. Afghan and D. MacKay (eds.), Hydrocarbons and 

Halogenated Hydrocarbons in the Aquatic Environment. Plenum Press, New York, 1980. 

Borenfreund, E. and J. Puerner. 1985. Toxicity Determined in vitro by Morphological Alterations and 

Neutral Red Absorption. Toxicology Letters 24. pp. 119-l 24. 

Brown, N.A. and S.E. Fabro. 1982. The in vitro Approach to Teratogenicity Testing. In: K. Snell. ed. 

Developmental Toxicology. London, England: Croom-Helm, p.p. 31057. 

Eckl, P.M., S.C. Strom, C. Michalopoulos, and R.L. Jirtle. 1987. Induction of Sister Chromatid Exchanges 

in Cultured Hepatocytes by Directly and Indirectly Acting Mutagens/Carcinogens. Carcinogenesis 

8:1077-l 083. 

Environment Canada. 1986. Guidelines on the Use of Mutagenicity Tests in the Toxicological 

Evaluation of Chemicals. Health and Welfare Canada. Ottawa, Canada 

Glass, L.R. “Background and Rationale for Relative Potency Framework for Evaluating Hazards Associated 

with Waste Water Samples.” Appendix B in “Health Hazard Evaluation of Waste Water Using 

Bioassays: Preliminary Concepts”. C. E. Easterly, et. al. Oak Ridge National Laboratory, Oak Ridge, 

TN 37831-6101 and U.S. EPA Office of Research and Development, Health Effects Research 

Laboratory, Cincinnati, Ohio 45268. July 1988. 

Hsie, A.W., D.A. Casciano, D.B. Coach, D.F. Krahn, J.P. O’Neill and B.L. Whitfield. 1981. The Use of 

Chinese Hamster Ovary Cells to Quantify Specific Locus Mutation and to Determine Mutagenicity of 

Chemicals. Mutation Research 86:193-214. 

Jirtle, R., Duke University Medical School, Durham, NC. Personal communication, April 24, 1989. 

Johnson, Rodney. Office of Research and Development, Environmental Research Laboratory, Duluth, MN. 

Personal communication, June 1, 1989. 

Kimmel, C.L., K. Smith, D.M. Kochhar, and R.M. Pratt. 1982. Overview of in vitro Teratogenicity 

Testing: Aspects of Validation and Application to Screening. Teratogenesis. Carcinog. Mutagen. 

2. pp. 221-229. 

Lave, L. 8. and G. 5. Omenn. 1986. Cost-Effectiveness of Short-Term Tests for Carcinogenicity. Nature 

324, 6:29-34. Note: Costs based on 1981 figures for pure chemicals. 

Marx, J.L. 1989. Detecting Mutations in Human Genes. Science 243. pp. 737-738. 

McGeorge, Leslie J., Judith 8. Louis, Thomas B. Atherholt, and Gerard J. McCarrity. 1985. “Mutagenicity 

Analyses of Industrial Effluents: Results and Considerations for Integration into Water Pollution 

Control Programs,” in Short-Term Bioassays in the Analysis of Complex Environmental Mixtures, IV. 

Edited by Michael D. Waters, Shahbeg 5. Sandhu, Joellen Lewtas, Larry Claxton, Gary Strauss and 

Stephen Nesnow. (Hearst Publishing Corp., 1985). 

Miller, ].A., and E.C. Miller. 1977. Ultimate Chemical Carcinogens as Reactive Mutagenic 

Electrophiles, in H.H. Hiatt, J.D. Watson and J.A. Winston (Eds.), Origins of Human Cancer, Cold 

Springs Harbor Laboratory, pp. 605-628. 

New Jersey Department of Environmental Protection. Mutagenicity Analyses of Industrial Effluents: 

Background and Results to Date. Office of Science and Research. August 1983. 

Quillardet, P., C. de Bellecombe, and M. Hofnung. 1985. “The SOS Chromotest, a Calorimetric Bacterial 

Assay for Genotoxins: Validation Study with 83 Compounds”. Mutation Research, 147. pgs. 79-95. 
19. 

20. 

21. 

22. 

23. 

24. 

25. 

26. 

27. 

28. 

29. 

30. 

31. 

32. 

33. 

34. 

3.5. 

Organization of Economic Cooperation and Development (OECD). 1984. Guidelines for Testing 

Chemicals. Section 4 -- Health Effects. Director of Information, OECD, 2, rue Andre-Pascal 75775 

Paris CEDEX 16, France. 

Tennant, R.W., 8.H. Margolin, M.D. Shelby, E. Zeiger, J.K. Haseman, J. Spalding, W. Caspary, M. Resnick, 

5. Stasiewicz, 6. Anderson, and R. Minor. 1987. Prediction of Chemical Carcinogenicity in Rodents 

From in vitro Genetic Toxicity Assays. Science 236:943-941. 

U.S. Environmental Protection Agency. 1978. Directory of Short Term Tests for Health and Ecological 

Effects. Health Effects Research Laboratory. EPA 600/l -78-052. 

U.S. Environmental Protection Agency. 1979a. Environmental Assessment: Short-term Tests for 

Carcinogens, Mutagens and other Genotoxic Agents. Health Effects Research Laboratory. Research 

Triangle Park, N.C. 

U.S. Environmental Protection Agency. 1979b. Short Term Tests for Carcinogens, Mutagens, and other 

Genotoxic Agents. Health Effects Research Laboratory, Research Triangle Part. EPA 625/9-79-003. 

U.S. Environmental Protection Agency. 1982a. Pesticide Assessment Guidelines, Office of Pesticide 

Programs. EPA/9-82-018 through 028. 

U.S. Environmental Protection Agency. 1982b. Toxic Substances Test Guidelines, Office of Toxic 

Substances. EPA/6-82-001 through 003. 

U.S. Environmental Protection Agency. 1983. Interim Procedures for Conducting the 

Salmonella/MicrosomaI Mutagenicity Assay - Ames Test. EPA 600/4-82-068. 

U.S. Environmental Protection Agency. 1985. Guidelines for Preparing Environmental and Waste 

Samples for Mutagenicity (Ames) Testing: Interim Procedures and Panel Meeting Proceedings. 

Office of Research and Development. EPA 600/4/85-058. 

U.S. Environmental Protection Agency. 1986a. Guidelines for the Health Risk Assessment of Chemical 

Mixtures. Federal Register 51 (185). pp. 34014-34025. 

U.S. Environmental Protection Agency. 1986b. Guidelines for Mutagenicity Risk Assessment. Federal 

Register 51 (185). pp. 34006-34012. 

U.S. Environmental Protection Agency. 1986~. Guidelines for Carcinogen Risk Assessment. Federal 

Register 51(185). pp. 33932-34003. 

U.S. Environmental Protection Agency. 1986d. Guidelines for Exposure Assessment. Federal Register 

51(185). pp. 34042-34054. 

U.S. Environmental Protection Agency. 1988. Validation of the Medaka Assay for Chemical 

Carcinogens: A Progress Report (Deliverable #I 8095A). Office of Research and Development, 

Environmental Research Laboratory, Duluth, MN. August 1988. 

U.S. Environmental Protection Agency. 1989a. Draft Guidance on Assessment, Criteria Development, and 

Control of Bioconcen-tratable Contaminants in Surface Waters. 

U.S. Environmental Protection Agency. 1989b. The Medaka Caarcinogenesis Model: A Progress Report 

(Deliverable # 8094A). Office of Research and Development, Environmental Research Laboratory, 

Duluth, MN. February 1989. 

Wilson, J. G. 1978. Survey of in vitro Systems: Their Potential Use in Teratogenicity Screening. In J.G. 

Wilson and F.C Fiaser, eds. Handbook of Teratology. Vol. 4. New York, NY. Plenum Press. pp. 135- 

153. 
APPENDIX H 

REFERENCE DOSE (RFD): DESCRIPTION AND USE IN 

HEALTH RISK ASSESSMENTS 
REFERENCE DOSE (RFD) DESCRIPTION AND USE IN 

HEALTH RISK ASSESSMENTS (REVISED 02/10/88) 

Principal Author: 

Donald Barnes, Ph.D. (OPTS) 

RfD Work Group: 

Donald Barnes, Ph.D. (OPTS) 

Judith Bellin, Ph.D. (OSWER) 

Christopher DeRosa, Ph.D. (ORD) 

Michael Dourson, Ph.D. (ORD), Co-Chair 

Reto Engler, Ph.D. (OPTS) 

Linda Erdreich, Ph.D. (ORD) 

Theodore Farber, Ph.D. (OPTS) 

Penny Fenner-Crisp, Ph.D. (OW) 

Elaine Francis, Ph.D. (OPTS) 

George Ghali, Ph.D. (OPTS) 

Richard Hill, M.D., Ph.D. (OPTS) 

Stephanie Irene, Ph.D. (OPTS) 

William Marcus, Ph.D. (OW) 

David Patrick, P.E., B.S. (OAR) 

Susan Perlin, Ph.D. (OPPE) 

Peter Preuss, Ph.D. (ORD), Co-Chair 

Aggie Revesz, B.S. (OPTS) 

Reva Rubenstein, Ph.D. (OSWER) 

Jerry Stara, D.V.M., Ph.D. (ORD) 

Jeanette Wiltse, Ph.D. (OPTS) 

Larry Zaragosa, Ph.D. (OAR) 

H-1 
REFERENCE DOSE (RFD): DESCRIPTION AND USE IN 

HEALTH RISK ASSESSMENTS 

Introduction 

This concept paper describes the U.S. Environmental Protection Agency’s (U.S. EPA) principal approach to and 

rationale for assessing risk for health effects other than cancer and gene mutations from chronic chemical 

exposure. By outlining principles and concepts that guide EPA risk assessment for such systemic effects the 

paper complements the new risk assessment guidelines (U.S. EPA, 1987), which describe the Agency’s approach 

to risk assessment in other areas, specifically carcinogenicity, mutagenicity, developmental toxicity, exposure, 

and chemical mixtures. (In this document the term “systemic toxicity” refers to an effect other than 

carcinogenicity or mutagenicity induced by a toxic chemical.) 

Backgroaad and Sunnnary 

Chemicals that give rise to toxic endpoints other than cancer and gene mutations are often referred to as 

“systemic toxicants” because of their effects on the function of various organ systems. In addition, chemicals 

that cause cancer and gene mutations also commonly evoke other toxic effects (i.e., systemic toxicity). Based on 

our understanding of homeostatic and adaptive mechanisms, systemic toxicity is treated as if there is an 

identifiable exposure threshold (both for the individual and for populations) below which there are no 

observable adverse effects. This characteristic distinguishes systemic endpoints from carcinogenic and 

mutagenic endpoints, which are often treated as nonthreshold processes. 

Systemic effects have traditionally been evaluated using such terms as “acceptable daily intake (ADI),” “safety 

factor (SF),” and “margin of safety (MOS),” concepts that are associated with certain limitations described 

below. The U.S. EPA established the Reference Dose (RfD) Work Group to address these concerns. 

In preparing this report, the RfD Work Group has drawn on traditional report on risk assessment (NRC, 1983), to 

more fully articulate the use of noncancer, nonmutagenic experimental data in reaching regulatory decisions 

about the significance of exposures to chemicals. In the process, the Work Group has coined less value-laden 

terminology -- “reference dose (RfD),” “uncertainty factor (UF)“; “margin of exposure (MOE)“; and “regulatory 

dose (RgD)” -- to clarify and distinguish between aspects of risk assessment and risk management. These 

concepts are currently in general use in many parts of U.S. EPA. 

Traditional Approach to Assessing Systemic Toxicity 

The U.S. EPA’s approach to assessing the risks associated with systemic toxicity is different from its approach 

to assessing the risks associated with carcinogenicity, because of the different mechanisms of action thought to 

be involved in the two cases. 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 theoretically no level of exposure for such a 

chemical that does not pose a small, but finite, probability of generating a carcinogenic response. In the case 

of systemic toxicity, however, organic homeostatic, compensating, and adaptive mechanisms exist that must be 

overcome before a 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 depleted before the effect is seen. 

The threshold concept is important in the regulatory context. The individual threshold hypothesis holds that a 

range of exposures from zero to some finite value can be tolerated by the organism with essentially no chance 

of expression of the toxic effect. Further, it is often prudent to focus on the most sensitive members of the 

population; therefore, regulatory efforts are generally made to keep exposures below the population threshold, 

which is defined as the lowest of the thresholds of the individuals within a population. 

H-2 
&script/on of the Tradifioaal Amroach 

In many cases, risk decisions on systemic toxicity have been made by the Agency using the concept of the 

“acceptable daily intake (ADI)” derived from an experimentally determined “no-observed-adverse-effect level 

(NOAEL).” The ADI is commonly defined as the amount of a chemical to which a person can be exposed on a 

daily basis over an extended period of time (usually a lifetime) without suffering a deleterious effect. The ADI 

concept has often been used as a tool in reaching risk management decisions (e.g., establishing allowable 

levels of contaminants in foodstuffs and water.) 

A NOAEL is an experimentally determined dose at which there was no statistically or biologically significant 

indication of the toxic effect of concern. In an experiment with several NOAELs, the regulatory focus is 

normally on the highest one, leading to the common usage of the term NOAEL as the highest experimentally 

determined dose without a statistically or biologically significant adverse effect. The NOAEL for the criticaf 

toxic effect is sometimes referred to simply as the NOEL. This usage, however, invites ambiguity in that there 

may be observable effects that are not of toxicological significance (i.e., they are not “adverse”). For the sake 

of precision, this document uses the term NOAEL to mean the highest NOAEL in an experiment. In cases in 

which a NOAEL has not been demonstrated experimentally, the term “lowest-observed-adverse-effect level 

(LOAEL)” is used. 

Once the critical study demonstrating the toxic effect of concern has been identified, the selection of the 

NOAEL results from an objective examination of the data available on the chemical in question. The ADI is 

then derived by dividing the appropriate NOAEL by a safety factor (SF), as follows: 

ADI (human dose) = NOAEL (experimental dose)/SF (Equation 1) 

Generally, the SF consists of multiples of 10, each factor representing a specific area of uncertainty inherent in 

the available data. For example, a factor of 10 may be introduced to account for the possible differences in 

responsiveness between humans and animals in prolonged exposure studies. A second factor of 10 may be used 

to account for variation in susceptibility among individuals in the human population. The resultant SF of 100 

has been judged to be appropriate for many chemicals. For other chemicals, with databases that are less 

complete (for example, those for which only the results of subchronic studies are available), an additional 

factor of 10 (leading to a SF of 1000) might be judged to be more appropriate. For certain other chemicals, 

based on well-characterized responses in sensitive humans (as in the effect of fluoride on human teeth), an SF as 

small as 1 might be selected. 

While the original selection of SFs appears to have been rather arbitrary (Lehman and Fitzhugh, 1954), 

subsequent analysis of data (Dourson and Stara, 1983) lends theoretical (and in some instances experimental) 

support for their selection. Further, some scientists, but not all, within the EPA interpret the absence of 

widespread effects in the exposed human populations as evidence of the adequacy of the SFs traditionally 

employed. 

Same Diffice/t/es in Ut/Mn# the Traditimwl Approach 

Scientific Issues 

While the traditional approach has performed well over the years and the Agency has sought to be consistent in 

its application, observers have identified scientific shortcomings of the approach. Examples include the 

following: 

a. Too narrow a focus on the NOAEL means that information on the shape of the dose-response curve is 

ignored. Such data could be important in estimating levels of concern for public safety. 

b. As scientific knowledge increases and the correlation of precursor effects (e.g., enzyme induction) 

with toxicity becomes known, questions about the selection of the appropriate “adverse effect” 

arise. 

C. Guidelines have not been developed to take into account the fact that some studies have used larger 

(smaller) numbers of animals and, hence, are generally more (less) reliable than other studies. 

H-3 
These and other “scientific issues” are not susceptible to immediate resolution, since the database needed is not 

yet sufficiently developed or analyzed. U.S. EPA work groups are presently considering these issues. 

Management-related Issues 

The use of the term “safety factor” - The term “safety factor” suggests, perhaps inadvertently, the notion of 

absolute safety (i.e., absence of risk). While there is a conceptual basis for believing in the existence of a 

threshold and “absolute safety” associated with certain chemicals, in the majority of cases a firm experimental 

basis for this notion does not exist. 

The implication that any exposure in excess of the ADI is “unacceptable” and that any exposure less than 

the ADI is “acceptable” or “safe” - In practice, the ADI is viewed by many (including risk managers) as an 

“acceptable” level of exposure, and, by inference, any exposure greater than the ADI is seen as “unacceptable.” 

This strict demarcation between what is “acceptable” and what is “unacceptable” is contrary to the views of 

most toxicologists, who typically interpret the ADI as a relatively crude estimate of a level of chronic exposure 

which is not likely to result in adverse effects to humans. The ADI is generally viewed by risk assessors as a 

“soft” estimate, whose bounds of uncertainty can span an order of magnitude. That is, within reasonable limits, 

while exposures somewhat higher than the ADl are associated with increased probability of adverse effects, that 

probability is not a certainty. Similarly, while the ADI is seen as a level at which the probability of adverse 

effects is low, the absence of all risk to all people cannot be assured at this level. 

Possible limitations imposed on risk management decisions - Awareness of the “softness” of the ADI 

estimate, as discussed above, argues for careful case-by-case consideration of the toxicological implications of 

individual situation, so that ADls are not given a degree of significance that is scientifically unwarranted. In 

addition, the ADI is only one factor in a risk management decision and should not be used to the exclusion of 

other relevant factors. 

Development of different ADls by different programs - In addition to occasionally selecting different 

critical toxic effects, Agency scientists have reflected their best scientific judgments in the final ADI by 

adopting factors different from the standard factors listed in Table 1. For example, if the toxic endpoint for a 

chemical in experimental animals is the same as that which has been established for a related chemical in 

humans at similar doses, one could argue for an SF of less than the traditional 100. On the other hand, if the 

total toxicologic data base is incomplete, one could argue that an additional SF should be included, both as a 

matter of prudent public policy and as an incentive to others to generate the appropriate data. 

Such practices, as employed by a number of scientists in different programs/agencies, exercising their best 

scientific judgment, have in some cases resulted in different ADls for the same chemical. The fact that 

different ADls were generated (for example, by adopting different SFs) can be a source of considerable 

confusion when the ADls are used exclusively in risk management decisionmaking. The existence of different 

ADls need not imply that any of them is more “wrong”--or “right”--than the rest. It is more nearly a reflection of 

the honest difference in scientific judgment. 

However, on occasion, these differences in judgment of the scientific data, can be interpreted as differences in 

the management of the risk. As a result, scientists may be inappropriately impugned, and/or perfectly 

justifiable risk management decisions may be tainted by charges of “tampering with the science.” This 

unfortunate state of affairs arises, at least in part, from treating the ADI as an absolute measure of safety. 

EPA Assessment of Risks Assoclatsd with Systemic Toxlclty 

The U.S. EPA approach to analyzing systemic toxicity data follow the general format set forth by NRC in its 

description of the risk assessment process (NRC, 1983). The determination of the presence of risk and its 

potential magnitude is made during the risk assessment process, which consists of hazard identification, doseresponse assessment, exposure assessment, and risk characterization. Having been apprised by the risk assessor 

that a potential risk exists, the risk manager considers control options available under existing statutes and 

other relevant non-risk factors (e.g., benefits to be gained and costs to be incurred). All of these considerations 

go into the determination of the regulatory decision (Figure 1). 
Table 1. Guidelines for the Use of Uncertainty Factors in Deriving Reference 

Doses and Modifying Factors 

Standard Uncertainty Factors (Ws): 

Use a lo-fold factor when extrapolating from valid experimental results in studies using 

prolonged exposure to average healthy humans. This factor is intended to account for the 

variation in sensitivity among the members of the human population and is referenced as 

“1 OH”. 

Use an additional IO-fold factor when extrapolating from valid results of long-term studies 

on experimental animals when results of studies of human exposure are not available or are 

inadequate. This factor is intended to account for the uncertainty involved in 

extrapolating from animal data to humans and is referenced as “1 OA”. 

Use an additional lo-fold factor when extrapolating from less than chronic results on 

experimental animals when there are no useful long-term human data. This factor is 

intended to account for the uncertainty involved in extrapolating from less than chronic 

NOAELs to chronic NOAELs and is referenced as “10s”. 

Use an additional lo-fold factor when deriving an RfD from a LOAEL, instead of a NOAEL. 

This factor is intended to account for the uncertainty involved in extrapolating from 

LOAELs to NOAELs and is referenced as “1OL”. 

ModlfyinRg Factor (MF): 

Use professional judgment to determine the MF, which is an additional uncertainty factor 

that is greater than zero and less than or equal to 10. The magnitude of the MF depends 

upon the professional assessment of scientific uncertainties of the study and data base not 

explicitly treated above; e.g., the completeness of the overall data base and the number of 

species tested. The default value for the MF is 1. 

*Source: Adapted from Dourson and Stara, 1983 

Hazard fdmtlficatlon 

Evidence 

Type of effect - Exposure to a given chemical, depending on the dose employed, may result in a variety of toxic 

effects. These may range from gross effects, such as death, to more subtle biochemical, physiologic, or 

pathologic changes. In assessments of the risk posed by a chemical, the toxic endpoints from all available 

studies are considered, although primary attention usually is given to the effect (the “critical effect”) 

exhibiting the lowest NOAEL. In the case of chemicals with limited data bases, additional toxicity testing 

may be necessary before an assessment can be made. 

Principal studies - Pnncrpal studies are those that contrtbute most srgnrfrcantly to the qualrtatrve assessment 

of whether or not a particular chemical is potentially a systemic toxicant in humans. In addition, they may be 
Figure 1. 

Conceptual Framework for Risk Assessment and Risk Management* 

Risk Assessment Risk Management 

Hazard Identification Control Options 

Dose-Response Assessment 

(e.g., RfD) 

Exposure Assessment 

Non-risk Analyses 

V V V V V 

Risk Characterization 

(e.g., Criterion) 

>Regulatory Decision 

(e.g., RgD, Standard) 

*Source: Adapted from NRC, 1983 

used in the quantitative dose-response assessment phase of the risk assessment. These studies are of two types: 

studies of human populations (epidemiologic investigations) and studies using laboratory animals. 

1. Epidemiologic studies - Human data are often useful in qualitatively establishing the presence of 

an adverse effect in exposed human populations. When there is information on the exposure level 

associated with an appropriate endpoint, epidemiologic studies can also provide the basis for a 

quantitative dose-response assessment. The presence of such data obviates the necessity of 

extrapolating from animals to humans; therefore, human studies, when available, are given first 

priority, with animal toxicity studies serving to complement them. 

In epidemiologic studies, confounding factors that are recognized can be controlled and measured, 

within limits. Case reports and acute exposures resulting in severe effects provide support for the 

choice of critical toxic effect, but they are often of limited utility in establishing a quantitative 

relationship between environmental exposures and anticipated effects. Available human studies on 

ingestion are usually of this nature. Cohort studies and clinical studies may contain exposureresponse information that can be used in estimating effect levels, but the method of establishing 

exposure must be evaluated for validity and applicability. 

2. Animal studies - For most chemicals, there is a lack of appropriate information on effects in 

humans. In such cases, the principal studies are drawn from experiments conducted on nonhuman 

mammals, most often the rat, mouse, rabbit, guinea pig, hamster, dog, or monkey. 

Supporting studies - These studies provide supportive, rather than definitive, information and can include 

data from a wide variety of sources. For example, metabolic and other pharmacokinetics studies can provide 

insights into the mechanism of action of a particular compound. By comparing the metabolism of the chemical 

exhibiting the toxic effect in the animal with the metabolism found in humans, it may be possible to assess the 

potential for toxicity in humans or to estimate the equitoxic dose in humans. 

Similarly, in vitro studies can provide insights into the chemical’s potential for biological activity; and under 

certain circumstances, consideration of structure-activity relationships between a chemical and other 
structurally related compounds can provide clues to the test chemical’s possible toxicity. More reliable in vitro 

tests are presently being developed to minimize the need for live-animal testing. There is also increased 

emphasis on generating mechanism-of-action and pharmacokinetics information as a means of increasing 

understanding of toxic processes in humans and nonhumans. 

Route of exposure - The U.S. EPA often approaches the investigation of a chemical with a particular route of 

exposure in mind (e.g., an oral exposure for a drinking water contaminant or an inhalation exposure for an air 

contaminant). In most cases, the toxicologic data base does not include detailed testing on all possible 

routes of administration, with their possibly significant differences in factors such as mechanism-of-action and 

bioavailability. In general, the U.S. EPA’s position is that the potential for toxicity manifested via one route 

of exposure is relevant to considerations of any other route of exposure, unless convincing evidence exists to 

the contrary. Consideration is given to potential differences in absorption or metabolism resulting from 

different routes of exposure, and whenever appropriate data (e.g., comparative metabolism studies) are 

available, the quantitative impacts of these differences on the risk assessment are delineated. 

Length of exposure - The U.S. EPA is concerned about the potential toxic effects in humans associated with 

all possible exposures to chemicals. The magnitude, frequency, and duration of exposure may vary considerably 

in different situations. Animal studies are conducted using a variety of exposure durations (e.g., acute, 

subchronic, and chronic) and schedules (e.g., single, intermittent, or continuous dosing). Information from all 

these studies is useful in the hazard identification phase of risk assessment. For example, overt neurological 

problems identified in high-dose acute studies tend to reinforce the observation of subtle neurological changes 

seen in low-dose chronic studies. Special attention is given to studies involving low-dose, chronic exposures, 

since such exposures can elicit effects absent in higher dose, shorter exposures, through mechanisms such as 

accumulation of toxicants in the organisms. 

Quality of the study - Evaluation of individual studies in humans and animals requires the consideration of 

several factors associated with a study’s hypothesis, design, execution, and interpretation. An ideal study 

addresses a clearly delineated hypothesis, follows a carefully prescribed protocol, and includes sufficient 

subsequent analysis to support its conclusions convincingly. 

In evaluating the results from such studies, consideration is given to many other factors, including chemical 

characterization of the compound(s) under study, the type of test species, similarities and differences between 

the test species and humans (e.g., chemical absorption and metabolism), the number of individuals in the study 

groups, the number of study groups, the spacing and choice of dose levels tested, the types of observations and 

methods of analysis, the nature of pathologic changes, the alteration in metabolic responses, the sex and age 

of test animals, and the route and duration of exposure. 

Weight-of-Evidence Determinatlon 

As the culmination of the hazard identification step, a discussion of the weight-of-evidence summarizes the 

highlights of the information gleaned from the principal and supportive studies. Emphasis is given to 

examining the results from different studies to determine the extent to which a consistent, plausible picture of 

toxicity emerges. For example, the following factors add to the weight of the evidence that the chemical poses 

a hazard to humans: similar results in replicated animal studies by different investigators; similar effects across 

sex, strain, species, and route of exposure; clear evidence of a dose-response relationship; a plausible relation 

between data on metabolism, postulated mechanism-of-action, and the effect of concern; similar toxicity 

exhibited by structurally related compounds; and some link between the chemical and evidence of the effect of 

concern in humans. 

Dose-Response Assessment 

Concepts and Problems 

Empirical observations have generally revealed that as the dosage of a toxicant is increased, the toxic response 

(in terms of severity and/or incidence of effect) also increases. This dose-response relationship is well- founded 

in the theory and practice of toxicology and pharmacology. Such behavior is observed in the following 

instances: in quanta1 responses, in which the proportion of responding individuals in a population increases 

with dose; in graded responses, in which the severity of the toxic response within an individual increases 

H-7 
with dose; and in continuous responses, in which changes in a biological parameter (e.g., body or organ 

weight) vary with dose. 

In evaluating a dose-response relationship, certain difficulties arise. For example, one must decide on the 

critical endpoint to measure as the “response.” One must also decide on the correct measure of “dose.’ tn 

addition to the interspecies extrapolation aspects of the question of the appropriate units for dose, the more 

fundamental question of administered dose versus absorbed dose versus target organ dose should be considered. 

These questions are the subject of much current research. 

Selection of the Critical Oata 

Critical study - Data from experimental studies in laboratory animals are often selected as the governing 

information when performing quantitative risk assessments, since available human data are usually insufficient 

for this purpose. These animal studies typically reflect situations in which exposure to the toxicant has been 

carefully controlled and the problems of heterogeneity of the exposed population and concurrent exposures to 

other toxicants have been minimized. In evaluating animal data, a series of professional judgments are made 

which involve, among others, consideration of the scientific quality of the studies. Presented with data from 

several animal studies, the risk assessor first seeks to identify the animal model that is most relevant to humans, 

based on the most defensible biological rationale (e.g., for instance using comparative pharmacokinetics data). 

In the absence of a clearly most relevant species, the most sensitive species (i.e., the species showing a toxic 

effect at the lowest administered dose) is used by risk assessors at U.S. EPA, since there is no assurance that 

humans are not at least as innately sensitive as the most sensitive species tested. This selection process is more 

difficult when the routes of exposure in the animal tests are different from those involved in the human 

situation under investigation. In order to use data from controlled studies of genetically homogeneous 

animals, the risk assessor must also extrapolate from animals to humans and from high experimental doses to 

comparatively low environmental exposures, and must account for human heterogeneity and possible concurrent 

human exposures to other chemicals. 

Although for most chemicals there is a lack of well-controlled cohort studies investigating noncancer 

endpoints, in some cases an epidemiologic study may be selected as the critical data (e.g., in cases of 

cholinesterase inhibition). Risk assessments based on human data have the advantage of avoiding the problems 

inherent in interspecies extrapolation. In many instances, use of such studies, as is the case with animal 

investigations, involves extrapolation from relatively high doses (such as those found in occupational 

settings) to the low doses found in the environmental situations to which the general population is more 

likely to be exposed. In some cases, a well-designed and well-conducted epidemiologic study that shows no 

association between known exposures and toxicity can be used to directly project an RfD (as has been done in 

the case of fluoride). 

Critical data - In the simplest terms, an experimental exposure level is selected from the critical study that 

represents the highest level tested in which “no adverse effect” was demonstrated. This “no-observed-adverseeffect-level” (NOAEL) is the key datum gleaned from the study of the dose-response relationship and, 

traditionally, is the primary basis for the scientific evaluation of the risk posed to humans by systemic 

toxicants. This approach is based on the assumption that if the critical toxic effect is prevented, then all 

toxic effects are prevented. 

More formally, the NOAEL is defined in this discussion as the highest experimental dose of a chemical at which 

there is no statistically or biologically significant increase in frequency or severity of an adverse effect in 

individuals in an exposed group when compared with individuals in an appropriate control group. As noted 

above, there may be sound professional differences of opinion in judging whether or not a particular response is 

adverse. In addition, the NOAEL is a function of the size of the population under study. Studies with a small 

number of subjects are less likely to detect low-dose effects than studies using larger numbers of subjects. Also, 

if the interval between doses in an experiment is large, it is possible that the experimentally determined 

NOAEL is lower than that which would be observed in a study using intervening doses. 

Critical endpoint - As noted under “Traditional Approach to Assessing Systemic Toxicity”, a chemical may 

elicit more than one toxic effect (endpoint), even in one test animal, or in tests of the same or different 

duration (acute, subchronic, and chronic exposure studies). In general, NOAELs for these effects will differ. The 

critical endpoint used in the dose- response assessment is the effect exhibiting the lowest NOAEL. 

Ha 
Reference Doss (RfD) 

The reference dose (RfD) and uncertainty factor (UF) concepts have been developed by the RfD Work Group in 

response to many of the problems associated with ADls and SFs, as outlined under “Traditional Approach to 

Assessing Systemic Toxicity” above. The RfD is a benchmark dose operationally derived from the NOAEL by 

consistent application of generally order-of-magnitude uncertainty factors (UFs) that reflect various types of 

data sets used to estimate RfDs. For example, a valid chronic animal NOAEL is normally divided by an UF of 

100. In addition, a modifying factor (MF), is sometimes used which is based on a professional judgment of 

the entire data base of the chemical. These factors and their rationales are presented in Table 1. 

The RfD is determined by use of the following equation: 

RfD = NOAEL / (UF x MF) 

which is the functional equivalent of Equation 1. In general, the RfD is an estimate (with uncertainty 

spanning perhaps an order-of-magnitude) of a daily exposure to the human population (including sensitive 

subgroups) that is likely to be without an appreciable risk of deleterious effects during a lifetime. The RfD is 

generally expressed in units of milligrams per kilogram of body weight per day (mg/kg/day). 

The RfD is useful as a reference point from which to gauge the potential effects of the chemical at other doses. 

Usually, doses less than the RfD are not likely to be associated with adverse health risks, and are therefore less 

likely to be of regulatory concern. As the frequency and/or magnitude of the exposures exceeding the RfD 

increase, the probability of adverse effects in a human population increases. However, it should not be 

categorically concluded that all doses below the RfD are “acceptable” (or will be risk-free) and that all doses 

in excess of the RfD are “unacceptable” (or will result in adverse effects). 

The U.S. EPA is attempting to standardize its approach to determining RfDs. The RfD Work Group has 

developed a systematic approach to summarizing its evaluations, conclusions, and reservations regarding RfDs 

in a “cover sheet” of a few pages in length. The cover sheet includes a statement on the confidence (high, 

medium, or low) the evaluators have in the stability of the RfD. High confidence indicates the judgment that 

the RfD is unlikely to change in the future because there is consistency among the toxic responses observed in 

different sexes, species, study designs, or in dose-response relationships, or that the reasons for existing 

differences are well understood. High confidence is often given to RfDs that are based on human data for the 

exposure route of concern, since in such cases the problems of interspecies extrapolation have been avoided. 

Low confidence indicates the judgment that the data supporting the RfD may be of limited quality and/or 

quantity and that additional information could result in a change in the RfD. 

The third step in the risk assessment process focuses on exposure issues. For a full discussion of exposure 

assessment, consult U.S. EPA’s guidelines on the subject (U.S. EPA, 1987). In brief, the exposure assessment 

includes consideration of the size and nature of the populations exposed and the magnitude, frequency, 

duration and routes of exposure, as well as evaluation of the nature of the exposed populations. 

R/s& Cbafact8rlzathm 

Risk characterization is the final step in the risk assessment process and the first input to the risk management 

(regulatory action) process. The purpose of risk characterization is to present the risk manager with a synopsis 

and synthesis of all the data that should contribute to a conclusion with regard to the nature and extent of the 

risk, including: 

a. The qualitative (“weight-of-evidence”) conclusions as to the likelihood that the chemical may pose 

a hazard to human health. 

b. A discussion of the dose-response information considered in deriving the WD, including the UFs 

and MFs used. 

C. Data on the shapes and slopes of the dose-response curves for the various toxic endpoints, 

toxicodynamics (absorption and metabolism), structure-activity correlations, and the nature and 

severity of the observed effects. 

l+9 
d. Estimates of the nature and extent of the exposure and the number and types of people exposed. 

e. Discussion of the overall uncertainty in the analysis, including the major assumptions made, 

scientific judgments employed, and an estimate of the degree of conservatism involved. 

In the risk characterization process, a comparison is made between the RfD and the estimated (calculated or 

measured) exposure dose (EED). The EED should include all sources and routes of exposure involved. If the EED 

is less than the RfD, the need for regulatory concern is likely to be small. 

An alternative measure that may be useful to some risk managers is the margin of exposure (MOE), which is the 

magnitude by which the NOAEL of the critical toxic effect exceeds the estimated exposure dose (EED), where 

both are expressed in the same units: 

MOE = NOAEL (experimental dose) / EED (human dose). 

When the MOE is equal to or greater than UF x MF, the need for regulatory concern is likely to be small. 

“Hypothetical, Simplified Example of Determining and Using WD” contains an example of the use of the 

concepts of NOAEL, UF, MF, RfD, EED, and MOE. 

Application In Risk Management 

Once the risk characterization is completed, the focus turns to risk management. In reaching decisions, the risk 

manager utilizes the results of risk assessment, other technological factors, and legal, economic and social 

considerations in reaching a regulatory decision. These additional factors include efficiency, timeliness, 

equity, administrative simplicity, consistency, public acceptability, technological feasibility, and nature of 

the legislative mandate. 

Because of the way these risk management factors may impact different cases, consistent -- but not necessarily 

identical -- risk management decisions must be made on a case-by-case basis. For example, the Clean Water Act 

calls for decisions with “an ample margin of safety”; the Federal Insecticide, Fungicide and Rodenticide Act 

(FIFRA) calls for “an ample margin of safety,” taking benefits into account; and the Safe Drinking Water Act 

(SDWA) calls for standards which protect the public “to the extent feasible.” Consequently, it is entirely 

possible and appropriate that a chemical with a specific RfD may be regulated under different statutes and 

situations through the use of different “regulatory doses (RgDs).” 

That is, after carefully considering the various risk and nonrisk factors, regulatory options, and statutory 

mandates in a given case (i), the risk manager selects the appropriate statutory alternative for arriving at an 

“ample” or “adequate” margin of exposure [MOE(i)]. As shown in Equation 2 below, this procedure establishes 

the regulatory dose, RgD(i) (e.g., a tolerance under FIFRA or a maximum contaminant level under SDWA), 

applicable to the case in question: 

RgD(i) = NOAEL / MOE(i) (Equation 2) 

Note that different RgDs are possible for a given chemical with a single RfD. Note also that comparing the RfD 

to a particular RgD(i) is equivalent to comparing the MOE(i) with the UF x MF: 

RfD/RgD(i) = MOE(i) / (UF x MF). 

In assessing the significance of a case in which the RgD is greater (or less) than the RfD, the risk manager 

should carefully consider the case- specific data compiled by the risk assessors, as discussed under “Risk 

Characterization”. In some cases, additional explanation and interpretation may be required from the risk 

assessors in order to arrive at a responsible and clearly articulated final decision on the RgD. 

It is generally useful to the risk manager to have information regarding the contribution to the RfD from various 

environmental media (e.g., air, water and food). Such information can provide insights that are helpful in 

H-10 
choosing among available control options. However, in cases in which site-specific criteria are being 

considered, local exposures through various media can often be determined more accurately than exposure 

estimates based upon generic approaches. In such cases, the exposure assessor’s role is particularly important. 

For instance, at a given site, consumption of fish may clearly dominate the local exposure routes, while, on a 

national basis, fish consumption may play a minor role compared to ingestion of treated crops. 

Work is underway in the U.S. EPA to apportion the RfD among the various environmental media. For example, 

consider the case of a food-use pesticide which is a contaminant in drinking water. In selecting among risk 

management actions under the Safe Drinking Water Act, it might be prudent to assume an RfD for drinking water 

purposes which is some fraction of the total RfD. Such an apportionment would explicitly acknowledge the 

possible additional exposure from ingestion of treated crops. The apportionment of the RfD would, in general, 

provide additional guidance for risk managers of the various media- specific programs. 

Other Directions 

In addition to the development of reference doses, the U.S. EPA is pursuing other lines of investigation for 

systemic toxicity. For example, the Office of Air Quality Planning and Standards is using probabilistic risk 

assessment procedures for criteria pollutants. In this procedure, the population at risk is characterized, and the 

likelihood of the occurrence of various effects is predicted through the use of available scientific literature and 

of scientific experts’ rendering their judgments concerning dose-response relationships. This dose-response 

information is then combined with the results of the exposure analysis to generate population risk estimates for 

alternative standards. Through the use of these procedures, decisionmakers are presented with ranges of risk 

estimates in which uncertainties associated with both the toxicity and exposure information are explicitly 

considered. The Office of Policy, Planning and Evaluation is investigating similar procedures in order to 

balance health risk and cost. In addition, scientists in the Office of Research and Development have initiated 

a series of studies designed to increase the reliability of risk assessments. They are investigating the use of 

extrapolation models as a means of estimating RfDs, taking into account the statistical variability of the 

NOAEL and underlying UFs. ORD is also exploring procedures for conducting health risk assessments that 

involve less- than-lifetime exposures. Finally, they are working on approaches to ranking the severity of 

different toxic effects. 

Hypothetical, Simplified Example of Determining and Using RfD 

Experimenttal Results 

Suppose the U.S. EPA had a sound 90-day subchronic gavage study in rats with the data in Table 2. 

Analysis 

Determination of the Reference Dose (RfD) 

Using the NOAEL - Because the study is on animals and of subchronic duration, 

UF = 10H x 10A x 10s = 1000 (Table 1) 

In addition, there is a subjective adjustment (MF) based on the high number of animals (250) per dose group: 

MF = 0.8. These factors then give UF x MF = 800, so that 

RfD = NOAEL/(UF x MF) = S/800 = 0.006 (mg/kg/day). 

H-l 1 
Table 2. Hypothetical Data to Illustrate the Reference Dose Concept 

Dose 

mg/kg/day 

Observation Effect Level 

0 Control-no adverse effects observed - 

1 No statistically or biologically significant differences NOEL 

between treated and control animals 

5 % decrease* in body weight gain (not considered to be NOAEL 

of biological significance); increased ratio of liver 

weight to body weight; histopathology indistinguishable 

from controls; evaluated liver enzyme levels 

25 20% decrease* in body weight gain; increased’ ratio of 

liver weight to body weight; enlarged, fatty liver with 

vacuole formation; increased* liver enzyme levels 

LOAEL 

‘Statistically significant compared to controls. 

Using the LOAEL - If the NOAEL is not available, and if 25 mg/kg/day had been the lowest dose tested that 

showed adverse effects, 

UF = 10H x 10A x 10s x 1OL = 10,000 (Table 1). 

Using again the subjective adjustment of MF = 0.8, one obtains 

RfD = LOAEL/(UF x MF) = ZS/SOOO = 0.003 (mg/kg/day). 

Risk Characterization Considerations 

Suppose the estimated exposure dose (EED) for humans exposed to the chemical under the proposed use pattern 

were 0.01 mg/kg/day (i.e., the EED is greater than the RfD). Viewed alternatively, the MOE is: 

MOE = NOAEL/EED = S (mg/kg/day) / 0.01 (mg/kg/day) = 500. 

Because the EED exceeds the RfD (and the MOE is less than the UF x MF), the risk manager will need to look 

carefully at the data set, the assumptions for both the RfD and the exposure estimates, and the comments of the 

risk assessors. In addition, the risk manager will need to weigh the benefits associated with the case, and other 

non-risk factors, in reaching a decision on the regulatory dose (RgD). 

H-l 2 
APPENDIX H 

REFERENCES 

1. Dourson, M.L. and J.F. Stara. 1983. Regulatory Toxicology and Pharmacology. 3: 224-238. 

2. Lehman, A.J. and O.G. Fitzhugh. 1954. Association of Food Drug Officials. USQ Bull. 18: 33-35. 

3. NRC (National Research Council). 1983. Risk Assessment in the Federal Government: Managing the 

Process. NAS Press, Washington, DC. 

4. U.S. EPA. 1987. The Risk Assessment Guidelines of 1986. Office of Health and Environmental 

Assessment, Washington, DC. EPA/600/8-871045. 

H-l 3 
APPENDIX I 

CHEMICALS AVAILABLE IN IRIS 
Page No. 1 

09/28/90 

CAS Chemical RfD q1* 

number Name mg/kg/day /mg/kg/day 

USE ONLY FOR SCREENING Values from IRIS 9/1/90 

Consult IRIS for Update and Whenever Possible, Site Specific 

Lipid, Consumption and Bioaccumulation Factors Should be 

Used in Application of RAC in Regulatory Action. 

50000 Formaldehyde 0.2 

50293 p,p1-DDT 0.0005 

50328 Benzo[a]pyrene ** 

51285 2,4-Dinitrophenol 0.002 

55185 N-Nitrosodiethylamine ** 

56235 Carbon tetrachloride 0.0007 

56359 Tributyltin oxide 0.00003 

56382 Parathion ** 

57125 Cyanide, free 0.02 

57249 Strychnine 0.0003 

57749 Chlordane 0.00006 

58899 gamma-Hexachlorocyclohexane 0.0003 

58902 2,3,4,6-Tetrachlorophenol 0.03 

60297 Ethyl ether 0.2 

60515 Dimethoate 0.0002 

60571 Dieldrin 0.00005 

62384 Phenylmercuric acetate 0.00008 

62533 Aniline ** 

62737 Dichlorvos 0.0008 

62759 N-Nitrosodimethylamine ** 

63252 Carbaryl 0.1 

64186 Formic acid 2 

65850 Benzoic acid 4 

67561 Methanol 0.5 

67641 Acetone 0.1 

67663 Chloroform 0.01 

67721 Hexachloroethane 0.001 

70304 Hexachlorophene 0.0003 

71363 n-Butenol 0.1 

** 

** 

150 

0.13 

** 

1.3 

16 

0.0057 

0.29 

51 

** 

0.014 

Estimated RAC (mg/l) RAC (mg/l) 

BCF RL: 10E-6 RL: 10E-6 

3% Lipid 0.0065 kg/day 0.020 kg/day 

40000 

10960 

5.2 

0.40 

29.72 

0.0001 0.00004 

4 

0.0002 

0.003 

1 

0.00006 

0.0009 

87.6 

0.4 8 3 

3804 0.000002 0.0000007 

146.8 0.00002 0.000007 

416 0.8 0.3 

0.77 3000 900 

0.4 6 2 

32.04 0.00002 0.000007 

0.58 2 0.5 

0.84 2 0.7 

0.4 0.1 0.3 

0.4 0.0005 0.0002 

12.2 90 30 

4.92 

0.4 

0.4 

5.56 

700 

40000 

0.71 

9000 

14000 

3000 

4000 

20 6 

0.001 0.0004 

0.00008 0.00002 

2000 500 

Consumption Consumption 

I-1 
Page No. 2 

09/20/90 

CAS Chemi ccl RfD 41* 

m&w Name m/kg/day /ng/kg/day 

USE DNLY FOR SCREENING Values from IRIS 9/l/90 

Consult IRIS for Ufxlate and Uhenever Possible, Site Specific 

lipid, Conscnption and BioaccumLation factors Should be 

Used in Application of RAC in Regulatory Action. 

71432 Benzene 

71556 l,l,l-Trichloroethane 

72208 Endrin 

72435 Bethoxychlor 

72548 p, p’ -ODD 

72559 p,p’-DDE 

74839 Brcmomethane 

74906 Nydrogen cyanide 

75058 Acetonitrile 

75070 Acetaldehyde 

75092 Dichloromethane 

75150 Carbon disulfide 

75252 Brmofon 

75274 Bromodichtoromthane 

75354 l,l-Dichloroethylem 

75694 lrichloromonofluoromethane 

75718 Dichlorodifluoromethane 

75876 ChloraL 

75990 Dalapon, sodium salt 

76131 CFC-113 

76448 NeptachLor 

77474 RexachIorocycLopentadiene 

77781 Dimethyl sulfate 

78002 Tetraethyl lead 

78488 Merphos oxide 

78591 lsophorone 

78831 lsobutyl alcohol 

78864 2-Chlorobutane 

78933 Methyl ethyl ketone 

l * 

0.09 

0.0003 

0.005 

l * 

l * 

0.0014 

0.02 

0.006 

l * 

0.06 

0.1 

0.02 

0.02 

0.009 

0.3 

0.2 

0.002 

0.03 

30 

0.0005 

0.007 

l * 

0.029 

l * 

l * 

0.24 

0.34 

l * 

0.0075 

0.0079 

0.6 

4.5 

l * 

l * 

0.0000001 

0.00003 

0.2 0.0041 

0.3 

l * l * 

0.05 l * 

Estimated RAC (me/l) RAC C&VI, 

BCF RL: lOE-6 RL: lOE-6 

3% Lipid 0.0065 kg/day 0.020 kg/&y 

Consurpt ion Consrnpt ion 

7.84 0.05 0.01 

14.52 70 20 

32.04 0.1 0.03 

2564 0.02 0.007 

12840 O-O00004 0.000001 

coDoo O.ODOOOW o.oODOOo2 

1.13 10 4 

0.4 

0.4 

1.54 

200 50 

0.9 0.3 

11.92 

7.16 

7.44 

13.36 

6 

3.23 

2.31 

63.2 

692 

744 

0.4 

0.1 0.04 

30 10 

0.002 0.0008 

200 80 

400 100 

7 2 

100 50 

5000 2000 

0.000004 0.000001 

0.1 0.03 

9 

0.56 

15.5 

0.4 

0.3 0.09 

6000 2000 

1000 400 

l-2 
Page No. 3 

09/28/90 

CAS Chemical RfD ql* 

~rName ng/kg/day /ng/kg/day 

USE ONLY FOR SCREENING Values from IRIS 9/l/90 

Consult IRIS for Update and bhenever Possible, Site Specific 

lipid, Consmption and Bioaccuwlation Factors Should be 

Used in Application of RAC in Regulatory Action. 

79DD5 1,1,2-Trichloroethane 0.004 

79016 Trichloroethylene l * 

79061 Acrylamide o.DDO2 

79107 Acrylic acid 0.08 

79221 Methyl chlorocarbonate l * 

79345 1,1,2,2-Tetrachloroethane l * 

8OD57 Bishpenol A. 0.05 

81812 Uarfarin o.ODD3 

82688 Pentachloronitrobenzene 0.003 

837% Rotemne 0.004 

84662 Diethyl phthalate 0.8 

8472D Ethylphthalyl l thylglycolate 3 

64742 Dibutyt p)lthalate 0.1 

855007 Diquat 0.0022 

85449 Phthalic hydride 2 

85687 Butyl bentyl phthalate 0.2 

85701 Butylphthalyl butylglycolate 1 

86306 N-Nitrosodiphenylain l * 

07683 HexachIorobutadiene 0.002 

87821 Nexabromobenzene 0.002 

87865 Pentachlorophenol 0.03 

88062 2,4,6-Trichlorophenol l * 

08857 Dinoseb 0.001 

91941 3,3g-dichlorobenridine l * 

92524 l,l-Biphenyl 0.05 

92875 Benridine 0.002 

93652 HCPP 0.003 

93721 2.4.5-TP 0.008 

93765 2,4,5-Trichlorophenoxyacetic acid 0.01 

0.057 

0.011 

4.5 

0.20 

l * 

t* 

l * 

0.0049 

0.078 

c. 

0.45 

230 

l t 

Estimated RAC (awl) RAC (ng/L) 

BCF RL: lOE-6 RL: lDE-6 

3% Lipid 0.0065 kg/day 0.020 kg/day 

Consuption Consmption 

6.64 0.03 

9.84 0.1 

0.4 D.OD6 

0.4 zoo0 

19.52 0.003 

166.8 3 

4.76 0.7 

1160 0.03 

91.2 0.5 

17.2 SW 

7.92 4OOD 

808 1 

1120 

372.4 

50 

397.2 

4560 

1568 

106 

239.2 

117 

243.2 

2.8 

48 

143.2 

81.6 

2 0.6 

30 9 

0.04 0.01 

0.0004 O.OODl 

0.005 0.002 

0.2 0.07 

0.05 0.01 

D.DW2 o.ww7 

2 0.7 

0.00002 D.WOOO5 

0.7 0.2 

0.6 0.2 

1 0.4 

l-3 
Page No. 4 

W/28/90 

USE DNLY FOR SCREENING Values from IRIS 9/l/90 

Consult IRIS for Update and Uhenever Possible, Site Specific 

Lipid, Consumption and Bioaccmulatim Factors Should be 

Used in Application of RAC in Regulatory Action. 

CAS Chemical RfD ql' 

nmtn?rNmw Wkg/&y /mg/kg/&,y 

94746 HCPA 0.0005 27.32 

94757 2,4-Dichloro@wnoxyacetic acid 0.01 2o.w 

94815 BCPB 0.01 50.4 

94826 4-(2,4-Dichlorophenoxytmtyric acid 0.008 38.56 

45487 o-Cresol 0.05 l * 7.6 

95498 o-Chlorotoluene 0.02 88 

95501 1,2-Dichlorobenzene 0.09 l * 103 

95578 2-Chlorophenol 0.005 8.8 

95658 3,4-Dimethylphenol 0.001 24.72 

95943 1,2,4,5-Tetrachlorobenrene 0.0003 1404 

95954 2,4,5-Trichlorophenol 0.1 176.4 

96184 1,2,3-Trichloropropane 0.006 5.84 

98011 Furfural 0.003 0.54 

98077 Benzotrichloride l ,t l * 283 

98828 cume 0.04 138 

98862 Acetophenone 0.1 2.82 

98953 Nitrobenzene 0.0005 4.92 

W354 1,3,5-Trinitrobenrene O.WDOS 1.93 

99650 n-Dinitrobenrene 0.0001 3.08 

100414 Ethylbenrene 0.1 l * 66.8 

loo425 styrene 0.2 29.24 

100647 Benryl chloride ++ 0.17 21.5 

100527 Bentaldehyde 0.1 2.42 

101213 Chlorprophan (CIPC) 0.2 96.4 

101553 p-Bromodiphenyl ether l * tt 2179 

101611 44~Hethylenebis(NN~dimethyl)aniline .* 0.046 888 

103231 Di-(2-ethylhexylHtdipnte 0.7 40000 

103333 Ambenzene l * 0.11 175.2 

1056D2 Caprolactam 0.5 0.4 

Estimsted RAC (mgfl) RAC (tag/l) 

BCF RL: lDE-6 RL: lOE-6 

3% Lipid 0.0065 kg/day 0.020 kg/day 

Conslnption Consumption 

0.2 0.06 

5 2 

2 0.7 

2 0.7 

70 20 

2 0.8 

9 3 

6 2 

0.4 0.1 

0.002 0.0007 

6 2 

10 4 

60 20 

3 1 

400 loo 

1 0.4 

0.3 0.09 

0.4 0.1 

20 5 

80 20 

0.003 D.Wl 

SW 100 

20 7 

0.0003 O.DWW 

0.2 0.06 

0.0006 0.0002 

10000 4000 

1-4 
Page No. 5 

W/28/90 

CAS Chemical RfD v 

nu&erNam mgfkgfday fmgfkgfday 

USE DNLY FOR SCREENING Values from IRIS 9/l/90 

Consult IRIS for Update and Whenever Possible, Site Specific 

Lipid, Consumption and Bioaccmlation Factors Should be 

Used in Application of RAC in Regulatory Action. 

106376 1,4-Dibranobenzem 

106445 p-Cresol 

106478 p-Chloroaniline 

106898 Epichlorohydrin 

106934 1,2-Dibromoethane 

106990 1,3-Butadiene 

107028 Acrolein 

107051 Ally1 chloride 

107062 1,2-Dichloroethane 

107131 Acrylonitrile 

107186 Ally1 alcohol 

107211 Ethylene glycol 

107302 Chloromethyl methyl ether 

108101 Methyl isobutyl ketone 

108316 Maleic anhydride 

lD8394 m-Cresol 

108452 n-Phenylenediamine 

108883 Toluene 

108907 Chlorobenrene 

108918 Cyclohexylmine 

108941 Cyclohexanone 

108952Phenol 

109693 1-Chlorobutane 

11WW Furan 

110543 n-Hexane 

110861 Pyridine 

111444 Bis(chloroethyl)ether 

114261 Baygon 

115297 Endosulfan 

0.01 

0.05 

0.004 

0.002 

l * 

*t 

l * 

l + 

l * 

l * 

0.005 

2 

** 

0.05 

0.1 

0.05 

0.006 

0.2 

0.02 

0.2 

5 

0.6 

l t 

0.001 

.* 

0.001 

t* 

0.004 

0.00005 

.* 

0.0099 

85 

t* 

l * 

910 

0.54 

l * 

l * 

t. 

l * 

l * 

*. 

1.1 

Estimated RAC (mgfl) RAC (Ml) 

BCF RL: 10E-6 RL: lOE-6 

3% Lipid 0.0065 kg/day 0.020 kg/day 

Consumption Consumption 

181.2 

7.6 

5.32 

0.4 

3.76 

5.08 

0.4 

2.20 

2.26 

0.4 

0.4 

0.4 

0.4 

1.38 

0.6 0.2 

70 20 

8 2 

3 0.9 

0.00003 0.00001 

0.000005 0.000002 

0.05 0.02 

100 40 

60000 20000 

400 100 

7.6 

0.4 

25.52 

28 

1.92 

0.69 

2.33 

15.5 

1.75 

179 

0.53 

0.98 

2.81 

70 20 

200 50 

80 30 

8 3 

1000 400 

80000 30000 

3000 900 

6 

20 

0.01 

20 

2 

7 

0.003 

5 

t-5 
Page No. 6 

W/28/90 

CAS Chemical RfD 41. 

rrrkrla w/kg/&y /ng/kg/&y 

USE DNLY FOR SCREENING Values from IRIS 9/l/90 

Consult IRIS for Update and Uhenever Possible, Site Specific 

Lipid, Ccmsrnption and Bioaccumlation Factors Should be 

Used in Application of RAC in Regulatory Action. 

115322 Dicofol 

116063 Aldicarb 

117817 BisCZ-ethylhexyl)phthalate 

118741 Hexachloro&ntene 

118967 2,4,6-Trinitrotoluem 

120127 Anthracene 

120616 Dilcthyl terephthalate 

12D821 1,2,4-Trichlorobenrene 

12D832 2,4-Dichloro@wnol 

121142 2,4/2,6-Dinitrotoluene mixture 

121697 N-N-Dimethylaniline 

121755 Malathion 

121824 RDX 

122349 Simstine 

122394 Diphenylmine 

122429 Propham 

122667 1,2-Diphenylhydrazine 

123331 claleic hydraride 

123911 1,4-Dioxane 

124403 Dinethylamine 

124481 Dibromochloromethane 

126987 Methacrylonitrile 

127184 Tetrachloroethylene 

129000 Pyrefw 

131113 Dimethyl phthalate 

1318% 4,6-Dinitro-o-cyclohexyl phenol 

133062 Captan 

133073 Folpet 

133904 Chloranrbm 

l * 

l * 

0.02 

0.0008 

o.ooo5 

0.3 

0.1 

l * 

0.003 

l * 

0.002 

0.02 

0.003 

0.002 

0.025 

0.02 

l * 

0.5 

l * 

l * 

0.02 

0.0001 

0.01 

0.03 

l * 

0.002 

0.013 

0.1 

0.015 

0.44 

l * 

0.014 

l * 

l * 

0.68 

0.11 

l * 

0.80 

0.011 

l * 

l * 

l * 

0.0035 

t-6 

Estimated RAC (mgfl) RN. (no/l) 

BCF RL: lOE-6 RL: lOE-6 

3% Lipid 0.0065 kg/day 0.020 kg/day 

%8D 

1.21 

4oooo 

188W 

2.27 

550 

11.56 

383.6 

42 

5.96 

11.16 

3.83 

15.36 

115.2 

18 

35.36 

0.4 

0.4 

0.4 

9.24 

0.42 

38.72 

1280 

2.51 

800 

8.24 

Consrnption Consuption 

o.oODDD3 O.OllWDO8 

0.00002 O.DODD% 

o.olm O.DOOl 

2 0.8 

6 2 

loo 30 

0.8 0.3 

0.002 D.ODO9 

2 0.6 

60 20 

1 0.5 

2 0.8 

10 4 

0.0004 O.DoOl 

1OODO 4OOD 

3 0.8 

20 8 

3 0.8 

3 0.9 

0.3 0.08 

0.03 

20 

0.009 

6 
Page No. 7 

W/28/90 

USE ONLY FOR SCREENING Values from IRIS 9/l/90 

Consult IRIS for Update and Whenever Possible, Site Specific 

Lipid, Consumption and BioaccuwLation Factors Should be 

Used in Application of RAC in Regulatory Action. 

CAS Chemical RfD v 

nuber Name WWday /Wkg/day 

Estimated RAC (tag/L) RAC tmg/\) 

BCF RL: lOE-6 RL: lOE-6 

3% Lipid 0.0065 kg/day 0.020 kg/day 

Consumption Consurpt i on 

137268 Thiram 0.005 

139402 Propazina 0.02 

141662 Bidrin 0.0001 

141786 Ethyl acetate 0.9 

143339 Sodim cyanide 0.04 

145733 Endothall 0.02 

148185 Sodim diethyldithiocarbanate 0.03 

150505 lkrphos 0.00003 

151508 Potassim cyanide 0.05 

156605 tram-1.2~Dichloroethylene 0.02 

2D6440 f luoranthene 0.04 

298DDD Methyl parathion 0.00025 

298044 Oisulfoton 0.00004 

300765 Waled 0.002 

302012 HydrazitWHydrazine sulfate l * 

3lWDD2 ALdrin 0.00003 

319846 alpha-Herachlorocyclohexen l * 

319857 beta-tfexschlorocycloherane l * 

319868 deLta-HexachlorocyclohexMw l * 

330541 Diuron 0.002 

330552 Linuron 0.002 

4601% Cyanogen 0.04 

SD4245 4-Aminopyridine tt 

SD6616 Potassius silver cyanide 0.2 

506649 Silver cyanide 0.1 

506683 Cyanogen bromide 0.09 

506774 Chlorine cyanide 0.05 

507200 t-Butylchloride l * 

510156 Chlorobenzilate 0.02 

l * 

3.0 

17 

6.3 

1.8 

t* 

l * 

l * 

l * 

47.2 5 1 

0.4 3 0.9 

0.54 20000 6000 

0.4 600 200 

2.5 90 30 

128D 0.3 0.1 

25.48 0.1 0.03 

60 0.007 0.002 

65.9 0.3 0.1 

1638 0.0000004 0.0000001 

146.8 0.00001 O.OODOO4 

146.8 0.00004 O.DOOOl 

146.8 

23.56 0.9 0.3 

51.6 0.4 0.1 

0.4 

12.2 

434 0.5 0.2 

1-7 
Page No. 8 

09/28/90 

USE ONLY FDR SCREENING Values from IRIS 9/l/90 

Consult IRIS for Update and Whenever Possible, Site Specific 

Lipid, Consumption and Bioaccurulation factors Should be 

Used in Application of RAC in Regulatory Action. 

CAS Chemical RfD 91. 

mm&r Naae w/kg/day /mg/kg/dw 

Estimated RAC (@t/l) RAC OWL) 

BCF RL: lOE-6 RL: lOE-6 

3% Lipid 0.0065 kg/&y 0.020 kg/day 

Conslnpt i on Consunpt ion 

541731 1,3-Dichlorobenzene l * et 

542621 Bariua cyanide 0.07 

542756 1,3-Oichloropropene 0.0003 l + 

542881 Bis(chloromethyl)ethcr l * 220 

544923 Copper cyanide 0.005 

556887 Nitroguanidine 0.1 

556887 Nitroguanidine 0.1 l * 

557211 Zinc cyanide 0.05 

563122 Ethion 0.0005 

563688 Thallim acetate 0.00009 l * 

576261 2,6-Dimethytphenol 0.0006 

592018 Calcius Cyanide 0.04 

598776 1,1,2-Trichloropropane 0.005 

608731 tech-HexachIorocyclohcxane 0.003 1.8 

608935 Pentachtorobenzene 0.0008 

615543 1,2,4-Tribromobenrene 0.005 

621647 I-Nitrosodi-N-propyLmnine l * 7.0 

630104 Selenourea 0.005 

630206 1,1,1,2-TetrachLoroethane 0.03 

709988 PropaniL 0.005 

732116 Phosmet 0.02 

759944 S-Ethyl dipropyLthiocarbamate 0.025 

765344 GlycidyaMehyck 0.0004 

834128 Amtryn 0.009 

886500 Terbutryn 0.001 

924163 N-Nitroso-di-n-butylamine +* 5.4 

930552 N-Nitrosopyrrolidine l * 2.1 

944229 fonofos 0.002 

950378 Methidathion 0.001 

103 

2.94 1 0.4 

1.05 0.00005 0.00002 

191 0.03 0.009 

24.72 0.3 0.08 

17.4 3 1 

146.8 0.00004 0.00001 

5120 0.002 0.0005 

524 0.1 0.03 

1.85 0.0009 0.0003 

39.72 8 3 

105.2 0.5 0.2 

15.2 10 4 

68.4 4 1 

40.4 2 0.8 

83.6 0.1 0.04 

12.68 0.0002 o.oooo5 

0.4 0.01 0.004 

193.6 0.1 0.04 

l-8 
Page No. 9 

W/28/W 

US Chrical 

nnkrName 

USE DNLY FOR SCREENING Values from IRIS 9/l/90 

ConsuLt IRIS for Update and Mwnever Possible, Site Specific 

Lipid, Consrnption and Bioaccuulation Factors Should be 

Used in Application of RAC in Regulatory Action. 

RfD ql* Estimatad RAC EmgIl) RAC tmp/L) 

Wkg/W /ng/kg/W BCf RL: loE-6 RL: lOE-6 

3% Lipid 0.0065 kg/&y 0.020 kg/&y 

Consrnption Consumption 

957517 Diphenamid 0.03 

961115 Tetrachlorovinphos 0.03 

1024573 HeptachLor epoxide 0.000013 

1071836 Glyphosate 0.1 

1116547 N-Nitrosodiethanolmine l * 

11631% DecabrcuaodiphenyL ether 0.01 

1314325 lhallic oxide l * 

1314621 Vanadium pentoxide 0.009 

1314847 Zinc phosphide o.clOo3 

1330207 Xylanes 2 

1332214 Asbestos l * 

1336363 Polychlorinated biphenyls l c 

1445756 DiisopropyL methyl phosphonate 0.08 

1563662 Carbofuran 0.005 

1582098 Triflurelin 0.0075 

1596845 Alar 0.15 

1610180 Prometon 0.015 

1646884 ALdicarb sulfone 0.0003 

1689845 BromoxyniI 0.02 

1689992 Bromoxynil octanoate 0.02 

1861321 Dacthal 0.5 

1861401 Benefin 0.3 

1897456 Chlorothalonil 0.015 

1910425 Paraquat 0.0045 

1912249 Atrazine 0.005 

1918009 Dim&a 0.03 

1918021 Picloram 0.07 

1918167 Propachlor 0.013 

1929777 Vernam 0.001 

9.1 

l * 

2.8 

+* 

l t 

l * 

t* 

7.7 

o.w?? 

t* 

10.72 30 10 

64.8 5 2 

13.72 0.00009 o.owo3 

0.4 

40000 

0.4 2000 700 

14.32 4 1 

1784 0.0008 0.0003 

0.4 4000 1000 

40.8 4 1 

0.4 8 3 

36.36 6 2 

5800 0.04 0.01 

327.6 20 5 

1784 2 0.6 

178.4 0.9 0.3 

26.96 2 0.6 

14.08 20 7 

11.56 70 20 

16.72 9 3 

179.2 0.06 0.02 

0.01 0.003 

0.003 O.DDW 

l-9 
Page No. 10 

W/28/90 

CAS Chmaical 

nt&arName 

USE tMLy FOR SCREENING Values from IRIS 9/l/90 

Consult IRIS for Llpdate and Uhemver Possible, Site Specific 

Lipid, Consumption and Bioacctmutation Factors Should be 

Used in Application of RAC in Regulatory Action. 

RfD ql* Estimated RAC (mg/L) RAC (ne/l) 

Wkg/dsy /ng/kg/day BCF RL: lOE-6 RL: lOE-6 

3% Lipid 0.0065 kg/day 0.020 kg/day 

Conslnption Conslnpticm 

1929824 Nitrapyrin O.Wl5 

2088415 Butylate 0.1 

2050477 p,pl-Dibromodiphenyl ether l * 

2104645 EPN O.WWl 

2164172 fluometuron 0.013 

2212671 Hoalinate 0.002 

2303175 Triallate 0.013 

2310170 PhosaIone 0.0025 

2312350 Propargite 0.02 

2385855 Mi rex 0.000002 

2425861 Captafol 0.002 

2439103 Dodine 0.004 

2691410 Dctahydro-1,3,5,7-tetranitro-1.3,s. 0.05 

2921882 ChLorpyrifoa 0.003 

3337711 Asulm 0.05 

3689245 Tetraethyldithiopyrophosphate o.wDs 

52346a4 Carboxin 0.1 

5902512 ferbacil 0.013 

6108107 epsilon-HexachLorocycLohexana l * 

6533739 ThaLLius carbonate 0.00008 

72871% Prometryn 0.004 

7439921 Lead and canpovds l * 

7439965 Manganese 0.1 

7439976 Mercury, (inorganic) l * 

7440020 Nickel, soluble salts 0.02 

7440144 Radiun 226 and 228 l * 

7440144 Rdius 228 (and 226) l * 

7440224 S i I ver 0.003 

7440360 Antimony 0.0004 

l * 

l . 

l * 

l * 

l * 

c* 

l * 

2-M.S/pCi/L 

79.2 

292 

11882 

648 

10.08 

35.28 

3a8 

0.2 0.07 

4 1 

0.0002 o.oooo5 

10 5 

0.6 0.2 

0.4 0.1 

752 o.oow3 O.WWW 

800 

0.4 

3.37 

5.4 

146.8 

0.04 0.01 

1000 SW 

300 loo 

30 8 

71.2 0.6 0.2 

I-10 
Page No. 11 

W/28/90 

USE ONLY FUR SCREENING Values frun IRIS P/l/M) 

Consult IRIS for Update and Whenever Possible, Site Specific 

Lipid, Consumption and Bioaccmulation Factors Should be 

Used in Application of RAC in Regulatory Action. 

CAS Chamical 

tnm&rNam 

7440382 Arsenic, inorganic 

7440393 Bariu 

7440417 Berylliu 

7440428 Boron (Boron and Borates only) 

7440439 ca&iua 

7448473 ChromitmW) 

7440508 Copper 

7440611 Uraniua, natural 

7446186 ThalliuaCl) sulfate 

7723140 Wtitc phosphorus 

7773868 Amostim sulffmate 

7782414 fluorine (sol&le fluoride) 

7703008 Selenious acid 

7783864 Hydrogen sulfida 

7791120 Thatliu chloride 

7883512 Phosphina 

8001352 Toxapfme 

8tM1589 Creosote 

8007452 Coke oven emissions 

aD65483 Demetal 

10102439 Nitric oxida 

10102440 Nitrogen dioxida 

10102451 lhallim nitrate 

10265926 Itethaidophos 

10453868 Resmethrin 

10595956 N-Nitroso-N-methylethylainc 

12035722 Nickel s&sulfide 

12039520 Thallium selenite 

12122677 Zineb 

RfO 91* Estimated RAC (mg/l) RAC (mg/l) 

mg/WW /Wkg/cJw BCF RL: W-6 RL: 10E-6 

3X Lipid 0.0065 kg/day 0.020 kg/day 

Conslnption Consumption 

l * l c 

0.07 

0.005 4.3 

0.09 

l * l * 

O-DOS l * 

*. l * 

.* l * 

o.woo8 l * 

o.ww2 l * 

0.25 

0.06 

0.003 

0.003 

o.oDow l * 

o.ooo3 

l * 1.1 414 o.oooo2 O.DWW8 

l * l * 

l * 

o.ww4 

0.1 

1 

o-woo9 l * 

O.WW5 0.4 1 0.4 

0.03 11900 0.03 0.009 

** 22 0.4 0.001 0.0004 

t. .t 

o.ww9 l * 

0.05 

I-11 
Page No. 12 

w/28/90 

CAS Chemical 

nu&erNw 

USE DULY FOR SCREENING Values from IRIS 9/l/90 

Consult IRIS for Update and Whenever Possible, Site Specific 

Lipid, Consusptim and Bioaccwlation Factors Should be 

Used in Application of RAC in Regulatory Action. 

12427382 Maneb 0.005 

13463393 Nickel carbonyl l * 

13593038 Quinalphos 0.0005 

13684634 Phemediphan 0.25 

14797558 Nitrate l * 

14797650 Nitrite 0.1 

14059677 Radon 222 l * 

15299997 Napropamide 0.1 

15972608 Alachlor 0.01 

16065831 Chromiu<lIl) 1 

16672810 Ethephon 0.005 

16752775 Hethoaryl 0.025 

17884352 Benaayl 0.05 

19044883 Oryzalin .* 

19488743 Hexachlorodibenro-p-dioxin mixture l * 

19666309 Oxadiaron 0.005 

28859738 Aluinm Phosphide 0.0004 

21087649 Uetribuzin 0.025 

21725462 Cyanazine 0.002 

22224926 Fenmiphos 0.00025 

22967926 Methyl mercury 0.0003 

23135220 oxaryl 0.025 

23564858 Thiophanate-methyl 0.08 

23958585 Pronamide 0.075 

24307264 Wepiquat chloride 0.03 

2505789D Bentazon 0.0025 

25329355 Pentachlorocyclopentadiene l * 

26628228 Sodius aride 0.004 

27314132 Norflurazon 0.04 

RfD ql* 

WWdey /Irg/kg/day 

l * 

1X-b/pCi/L 

62.00 

l * 

Estimated RAC (mg/l) RAC O&I/L) 

BCF RL: lDE-6 RL: lDE-6 

3% Lipid 0.0065 kg/day 0.020 kg/day 

Consumption Consumption 

42.8 0.1 0.04 

124 20 7 

156 7 2 

226 0.5 0.2 

31.1 

3.64 6 

21.5 0.1 

2.92 300 

33.92 20 

2 

0.04 

100 

8 

l-12 
Page No. 13 

W/28/90 

CAS Chemical 

nu&erNme 

USE DNLY FOR SCREENING Values from IRIS 9/l/90 

Consult IRIS for Update and Uhenever Possible, Site Specific 

Lipid, Consumption and Bioaccuwlstion Factors Should be 

Used in Application of RAC in ReguLatory Action. 

RfD ql* 

w/kg/dey /mg/kg/day 

Estimated RAC (mg/l) RAC (ng/l) 

BCF RL: lOE-6 RL: lOE-6 

3% Lipid 0.0065 kg/day 0.020 kg/day 

Consrnption Consumption 

28249776 Thiobencarb 

29232937 Pirimiphos-methyl 

30560191 Acephate 

32534819 Pentabromodiphmyl ether 

32536520 Dctabromodiphenyl ether 

33089611 Amitraz 

33820530 lsopropalin 

34014181 Tebuthiurcm 

35367385 Difl&enruron 

35554440 lnmralil 

36483600 Hexabrcwnodiphenyl ether 

36734197 Iprodione (Rovral) 

39148248 Fosetyl-al 

39515418 Dani tot 

39638329 Bis(Z-chloroisopropyl) ether 

LD888479 letrabromodiphenyl ether 

40487421 Pendimethalin (Proul) 

41851507 Chlorocyclopentadiene 

42874033 Oxyfluorfen 

43121433 Bayleton 

43222486 Di f enmquat 

49690940 Tribromodiphenyl ether 

50471448 Vinclozolin 

51218452 Metolachlor 

51235042 Hexazinone 

51630581 Pydrin 

52315078 Cypermethrin 

52645531 Permethrin 

55285148 Carbosulfan 

0.01 

0.01 

l * 

0.002 

0.003 

0.0025 

0.015 

0.07 

0.02 

0.013 

l * 

0.04 

3 

0.0005 

0.04 

l * 

0.04 

tt 

0.003 

0.03 

0.08 

l * 

0.025 

0.1 

0.033 

0.025 

0.01 

0.05 

0.01 

71.6 2 0.5 

66.7 2 0.5 

0.0087 0.4 3 1 

** 

*t 

5520 0.03 0.01 

0.4 2000 600 

44.8 5 2 

19.84 7 2 

l * 

5170 0.001 0.0003 

21.9 20 6 

l * 

3416 0.1 0.04 

l * 

l * 

11040 0.003 0.001 

39.28 8 3 

252.4 4 1 

348.4 1 0.3 

35200 0.008 0.002 

10320 0.01 0.003 

40000 0.01 0.004 

l-13 
Page No. 14 

W/28/90 

CAS Chemical 

nnkrName 

USE ONLY FDR SCREENING Values from IRIS 9/1/W 

Consult IRIS for Update and Uhenever Possible, Site Specific 

Lipid, Cmslmption and Bioaccutulation Factors Should be 

Used in Application of RAC in Regulatory Action. 

RfD ql* Estimsted RAC (W/l) RAC fmg/l) 

m/kg/day /W/kg/day BCF RL: lOE-6 RL: lOE-6 

3% Lipid 0.0065 kg/&y 0.020 kg/day 

Consumption ConsLnption 

55290647 Dimethipin 

57837191 Metalaxyl 

58138082 Tridiphane 

59756604 fluridone 

60207901 Propiconazole 

60568050 furmcyclox 

62476599 Sodiw acifluorfm 

63936561 Nonabromdiphmyl ether 

64902723 Chlorsulfuron 

65195553 Avermectin Bl 

662152'18 Cyromarine 

66332965 flutolanil 

66841256 Tralathrin 

67485294 Amdro 

677470% Prochlorar 

68085858 Cyhalothrin/Karate 

683593?5 Baythroid 

69409945 Flwalinate 

69806402 Haloxyfop-methyl 

72128020 fcmssafen 

72128020 Fomesafen 

74051802 Sethoxydim 

74115245 Apollo 

74223646 Ally 

76578148 Assure 

76738620 Paclobutrarol 

77182822 Glufosirmte-ammniu 

77323843 Trichlorocyclopentadiane 

77323854 Tetrachlorocyclopentadiemt 

0.02 l * 

0.06 

0.003 

0.08 

0.013 

l * 0.030 

0.013 

l t l * 

0.05 

0.004 

0.0075 

0.06 

0.0075 

0.0003 

0.009 0.15 

0.005 

0.025 

0.01 

0.00005 

.+ 0.19 

l * 0.19 

0.w 

0.013 

0.25 

0.009 

0.013 

0.0004 

l * l * 

l . l * 

24.1 

1856 

19.0 

0.54 

40000 

10700 

40000 

30 9 

0.5 0.2 

30 9 

200 SO 

0.002 0.0007 

0.005 0.002 

0.007 0.002 

0.003 o.ww 

l-14 
Page No. 15 

w/tam 

CAS Chemical 

ru&erName 

USE ONLY FOR SCREENING Values from IRIS g/l/W 

Consult IRIS for Update and Whenever Possible, Site Specific 

Lipid, Consumption and Bioacclnulation Factors Should be 

Used in Application of RAC in Regulatory Action. 

77501634 Lactofen 0.002 

78587050 Savey 0.025 

fO277273 Harmony 0.013 

81335377 lmazsquin o-25 

81335725 Pursuit 0.25 l * 

82558507 lsoxaben 0.05 

82657043 Bipbenthrin 0.015 

83055w6LondaX 0.2 

a550ww NuStar 0.0007 

88671890 Systhane 0.025 

90992324 Chlorinuron-ethyl 0.02 

101200480 Express 0.008 

RfD 41* Estimatad RAC (tag/l) RAC OWL) 

WkgfW /ng/kg/dey BCF RL: lOE-6 RL: lOE-6 

3X Lipid 0.0065 kg/day 0.020 kg/day 

Consumption Cmsurption 

40000 0.004 0.001 

l-15 
United States 

Environmental Protection 

Agency 

Office Of Water 

EPA Technical Support Document 

For Water Quality-based 

Toxics Control 

RESPONSIVENESS SUMMARY 
RESPONSIVENESS SUMMARY 

General 

1. The TSD needs to be peer reviewed. 

EPA Response: We feel EPA has provided ample opportunity for 

review of the TSD. The Williamsburg Workgroup met in 

December, 1988, and was composed of a diverse group of people 

from EPA Regions, States, environmental groups, trade 

associations, academia, private industry, and municipalities. 

This group was formed to determine what changes were needed 

in the document. Based on the Workgroup's inputs, EPA 

produced a first draft TSD in November, 1990. This draft was 

sent out to Workgroup members, and anyone else upon request, 

for review and comment. EPA considered each comment and made 

changes to the document. A second revised draft TSD was 

noticed in the Federal Resister on May 11, 1990. Over 2500 

copies of the new draft were sent out. Between the two 

drafts, we received 120 comments from a diverse group of 

people. EPA considered each individual comment and made 

changes to the document where necessary. 

2. The TSD doesn't apply to CSO discharges. 

EPA Response: We acknowledge that the TSD was written with 

continuous discharges in mind. However, there is no reason 

why the general concepts and some of the recommendations of 

the TSD cannot be extended to CSO or other rainfall related 

discharges. The dynamic model applications in the TSD have 

been used over the last 10 years to address water quality 

problems related to CSOs. EPA believes that this model can 

also be applied to address toxic problems with CSOs. 

Chanter 1: Approaches to Water Quality-based Toxics Control 

Overview: Added more information and clarification to support 

EPA's position with regard to the major issues addressed in the 

comments. 

1. The cause and effect relationship between effluent toxicity 

and instream impacts is not adequately documented, and other 

factors that may cause instream impacts have not been 

addressed. EPA's use of the CETTP studies was criticized. 

EPA Response: The revised draft TSD has been changed to 

include more documentation to support our position on the 

cause and effect relationship between effluent toxicity and 

instream impacts. Supporting information on a study conducted on the Trinity River in Texas was added as well as more 

1 
2. 

3. 

detailed explanation of the CETTP studies. EPA evaluated the 

results of the CETTP studies, the North Carolina studies, and 

the Trinity River study and found that if toxicity is present 

after considering dilution, instream impact will also be 

present. EPA's finding is now clearly stated in the TSD and 

referenced. The TSD now includes a discussion of Parkhurst's 

major criticisms of the CETTP studies and why EPA feels those 

criticisms are unfounded. The TSD also acknowledges that 

biological, physical, and chemical factors of the community 

can influence the actual effects that effluent toxicity may 

cause in the receiving water. 

Toxicity test method precision is too variable to be used in 

NPDES permits. 

EPA Response: The revised draft includes all available 

precision data for both acute and chronic toxicity tests; this 

includes intralaboratory and all available interlaboratory 

test results. Raw precision data is presented in table form 

in Appendix A and discussed within the chapter. EPA 

evaluations, as well as published literature reviews which 

include estimates of whole effluent precision data are 

presented. EPA is comfortable with the conclusion that whole 

effluent toxicity tests are no more variable than chemical 

analytical methods and therefore stands behind the 

recommendation that toxicity test methods be used in NPDES 

permits. 

The same data contained in the TSD has been used by EPA in 

proposing adoption of the toxicity methods into EPA's regulations at 40 CFR 136. The toxicity methods were proposed 

for adoption on December 4, 1989. Comments were received and 

will be answered upon notice of EPA's decision regarding these 

methods. The decision is expected in April 1991. 

The biologicalcriteria/bioassessment approach is not yet part 

of EPA's water quality regulations and should not yet be used 

in the regulatory process. For the best assessment of sources 

and causes, EPA should use a g9weight-of-evidence approach." 

EPA Response: As previously stated in the TSD (per section 

131,11(b)(2) of the Water Quality Standards Regulation), 

biological criteria can supplement existing chemical-specific 

criteria and provide an alternative to chemical-specific 

criteria where such criteria cannot be established. To 

acknowledge the current evolving status of the use of biocriteria, EPA's statement in the TSD has been revised to read: 

"To better protect the biological integrity of aquatic 

communities, EPA recommends that States besin to develop and 

implement biological criteria in their water quality standards." 

2 
EPA does not agree with use of the llweight-of-evidence'M 

approach in this context because biosurveys are too complex 

to override the other two methods (i.e., chemical-specific 

and whole effluent toxicity). The TSD now explains 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 (Chapter 3). The results of one 

assessment technique should not be used to contradict or 

overrule the results of the other(s). 

However, EPA recognizes that there are instances when the 

whole effluent, chemical specific, and biological criteria 

approaches will give disparate results. The TSD was revised 

to recommend that permitting authorities use a more complex 

way to assess excursions beyond standards and establishing 

permit limits to provide assurance that simplifying assumptions are not the cause of the apparent discrepancy. The TSD 

also now includes examples of where the whole effluent 

toxicity test protocols may conflict with critical environmental parameters to lead to an apparent disparity between 

the whole effluent toxicity and chemical specific approaches. 

4. The problem of false positives (i.e., instream impact predicted by toxicity test where none exists) is not addressed. 

EPA Response: The revised draft TSD now addresses the problem 

of toxicity test interferences caused by environmental 

parameters and explains that 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 carefully look 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 (TIE) may be necessary to 

3 
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 toxicity 

tests conducted. 

5. Current standard algal toxicity test methods lack the ability 

to provide useful data on the ecological impact of a discharge 

and should not be recommended as a test species. EPA should 

recommend three species representing two different phyla be 

used for toxicity testing. EPA's recommendation to use 

surrogate (rather than resident) species in toxicity testing 

as being more protective seems contrary to the site specific 

nature of the permitting process. 

EPA Response: To address concerns with the algal test and to 

allow more flexibility in testing, the recommendation in the 

TSD has been changed to "fish, invertebrate, and plant." EPA 

will not modify its recommendation to include only 2 phyla. 

EPA's objective in requiring 3 species from 3 different phyla 

is to be predictively protective. 

The TSD's discussion of the use of surrogate species was 

expanded to explain that to use a resident organism, a 

facility would have to develop a protocol to culture the 

organism and to assess intra- and inter-laboratory 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 their health cannot be assured. 

6. More documentation is needed for the statement that the ttIC25 

is approximately the analogue of an NOEC.@@ Before EPA makes 

such a broad ranging recommendation, there must be sufficient 

data to establish an overall relationship. 

EPA Response: The language in the revised draft has been 

clarified to better explain what an IC25 is and how it is 

calculated. EPA believes that there is sufficient data to 

support the statement that the "IC25 is approximately the 

analogue of an NOEC." The data in Appendix A is presented so 

that both hypothesis testing and IC25 calculations of an NOEC 

can be compared. A statistical analysis using minimum 

significant differences is graphically presented in Figure l1. Figure l-l shows that an NOEC calculated using the IC25 

is comparable to an NOEC calculated using hypothesis testing, 

and that the relationship is statistically sound. 

4 
7. The data in the TSD shows that whole effluent toxicity is not 

additive. 

EPA Response: The cited article in the TSD and the TIE data 

were reviewed. These data support EPA's position that acute 

toxicity is additive. However, the data do not support 

additivity of chronic toxicity. Therefore, the TSD was 

changed to reflect this. 

8. There are insufficient chemical-specific field data to support 

that exceedances of th8 criteria cause in&ream impacts. 

EPA Response: The field studies referenced in Chapter 1 for 

chemical specific criteria investigations were conducted over 

twenty years ago. The field investigators dosed a stream with 

toxicants to measure the response. This approach is not 

possible today because it would violate the States water 

quality standards. Since that time, EPA has developed a 

method of using laboratory toxicity data on specific chemicals 

to derive data. 

ChaDter 2: Water Oualitv Criteria and Standards 

Overview: Revised introduction to summarize key regulatory 

requirements; reorganized into clearly defined aquatic life and 

human health discussions; added more information on what should be 

considered when allowing mixing zones; added more in-depth 

discussion of criteria for human health protection. 

1. Water quality criteria are not reliable due to data gaps or 

errors in derivation. Site-specific criteria should not be 

limited to being more stringent than the national criteria. 

EPA Response: Changing the general procedure for deriving WQ 

criteria is not within the scope of the TSD. In the near 

future EPA expects to re-examine the general procedures for 

deriving aquatic life criteria, and will request public 

comment thereon. While EPA requests public comment on all WQ 

criteria documents before publishing them in final form, EPA 

accepts comment on criteria at any time and can correct errors 

through criteria summary documents that it distributes from 

time to time. Finally, there is no Agency policy, set forth 

in the TSD or elsewhere, that prevents state-wide or sitespecific criteria from being less stringent than the national 

criteria. 

2. Provide more clarity on how to prohibit lethality within the 

mixing zones? 

EPA Response: The TSD has been amended to clarify the goals 

of EPA's recommendations on mixing zones. It now states that 

mixing zone conditions should not be lethal to organisms 

5 
passing through it. Survival of organisms that might wish to 

reside permanently in a mixing zone is not assured by past or 

current policy. 

Because the chronic criteria may be exceeded at the end of 

the pipe there is a potential for lethality to sensitive 

organisms that attempt to reside permanently in the mixing 

zone. Part of the intent of the recommendations of the 1985 

TSD and 1991 TSD is to protect the survival of organisms 

passing through the mixing zone. In all cases, exposures to 

concentrations above the CCC and CMC cannot be correctly 

interpreted without accounting for the duration of exposure. 

3. Th8 TSD is flawed because it assumes that mixing zones exist. 

The Great Lakes Water Quality Agreement pr8ClUd8S us8 of flow 

augmentation as a substitute for adequate treatment. 

EPA Response: The TSD recommendations in no way authorize 

mixing zones where otherwise prohibited. In addition, the 

recommendations on mixing zones do not advocate management of 

reservoirs for flow augmentation. 

4. TSD should at a minimum be against mixing zones for persistent and bioacccumulative toxicants, and the burden of 

justifying mixing zones for non-persistent and non-bioacccumulative toxicants should be on the discharger. The 

Great Lakes Water Quality Agreement calls for zero discharge 

of pollutants and elimination of persistent and bioaccumulative toxicants. The Clean Water Act also has a goal of zero 

discharge of pollutants. 

EPA Response: The TSD continues to note that EPA regulations 

allow mixing zones at the discretion of the State. The TSD 

also discusses options that should be considered when 

determining whether to allow mixing zones for aquatic life and 

human health protection. For protection of aquatic life, a 

mixing zone may be permitted as long as its size is 

sufficiently limited that it does not significantly impair 

the integrity of the water body as a whole, and it does not 

cause lethality to organisms passing through the mixing zone. 

For protection of human health, mixing zones should be 

restricted such that they do not encroach on areas often used 

by the public for fishing, and particularly where stationary 

species such as shellfish are harvested; mixing zones may also 

be restricted to compensate for uncertainties in the 

protectiveness of the water quality criteria or uncertainties 

in the assimilative capacity (TMDL) of the water body. 

Bioaccumulative pollutant problems are not fundamentally 

caused by mixing zones. Bioaccumulation is generally a 

system-wide problem that occurs when the appropriate TMDL for 

a water body as a whole is exceeded. Consequently, EPA does 

not consider mixing zone restrictions to be the best mechanism 

6 
for addressing such problems. However, eliminating mixing 

zones can be used as a device to further reduce the loading 

below the TMDL, although the results are not as predictable 

as a direct reduction of the TMDL would be. Chapter 2 has 

been now been modified to state the factors that should be 

considered in judging whether a mixing zone causes significant 

health or ecological risks. 

It should be noted that the TSD deals with WQ based effluent 

limits (i.e., those needed to protect aquatic life and human 

health), implemented primarily under the specific requirements of Sections 301, 303, and 304 of the CWA. Under this 

framework, zero discharge of pollutants is generally required 

where the water quality standard or the Total Maximum Daily 

Load is set at zero. 

5. The TSD should allow mixing zones for all toxicants including bioaccumulative pollutants. Furthermore, fate processes 

(such as sedimentation or decay) that occur within mixing 

zones should be taken into consideration. 

EPA Response: The TSD now sets forth specific conditions 

under which denial of mixing zones would be appropriate. EPA 

regulations also allow (while neither encouraging nor 

discouraging) States to use mixing zones. The recommendations of the TSD implicitly discourage consideration of fate 

processes such as sedimentation or decay. Mixing zones are 

an allowance for variations in concentration due to incomplete mixing, over small spatial scales, usually too small 

for fate processes to significantly reduce concentrations. 

6. Limiting all dischargers to 0.3 TUa is akin to a technology 

based approach such as a 30 mg/L BOD or suspended solids 

limit. The TSD should consider the resulting instream water 

quality. 

EPA Response: The mixing zone discussion has been modified 

to provide different alternatives for assuring that instream 

goals and standards are met. Not exceeding 0.3 TUa at the 

end of the pipe is one of the recommended ways to assure 

survival of organisms passing through the mixing zone. 

Nevertheless, the discussion has been modified to deemphasize the technological requirements and emphasize the 

attainment of instream goals. 

7. EPA should provided more information on the tests it used to 

arrive at an LC5O/LC1 ratio of 0.3. According to the data 

presented in the TSD, the 0.3 ratio is overly conservative in 

most cases. 

EPA Response: The magnitude of the acute WET criterion is 

based on data collected from a number of facilities in EPA 

7 
Region 4. These data show that 90% of the facilities 

exhibited an LC1 which was no less than 0.3 times the LC50. 

This is different from the 0.5 factor used to derive a 

chemical specific acute criterion. 

8. The acute toxicity criterion of 0.3 TUa is below detection. 

How would it be implemented? 

EPA Response: The implementation of this criterion would be 

identical to that used for specific chemicals. This implementation is expressed in Chapter 5. 

9. The l-hour averaging period for the acute criteria (CMC) is 

overly restrictive, does not correspond to the 48-96 hour 

toxicity tests, and cannot be modeled with existing EPA WQ 

models. Elsewhere in the document, EPA indicates that 24 

hours is an appropriate averaging period for modeling 

purposes. Both concentration and exposure time are important, since for many toxicants the 96-hour LC50 is dramatically higher than the l-hour LC50. The 1-hour averaging 

period is technically unsupported by the limited information 

presented. EPA appears to have edited the data presented in 

Appendix D to include only those data that support the l-hour 

averaging period. EPA should present all available data on 

the effect of exposure duration on toxicity. 

EPA Response: EPA agrees that both concentration and exposure 

time are important. The TSD has now been modified to note 

that the l-hour averaging period is based on ammonia, a fastacting toxicant. As the l-hour averaging period was intended 

to be protective even for the fastest acting toxicants, it may 

be overly conservative for many pollutants. Consequently, the 

TSD recommends allowance for site-specific (or chemicalspecific) modification of the averaging periods. Alternative 

averaging periods can be developed from data on the time 

course of mortality in acute toxicity tests. 

EPA expects that for many pollutants, such site-specific or 

state-wide alternative averaging periods, if developed using 

adequate data, may be greater than the period recommended for 

national criteria. Furthermore, EPA recognizes that a 24- 

hour acute averaging period may be appropriate in some 

modeling contexts, where concentrations do not change rapidly 

over short time periods. 

While Appendix D of the TSD presents some examples of 

pollutants for which a short averaging period, on the order 

of hours, might be appropriate, the TSD was not intended as 

the mechanism by which EPA would develop a rationale for the 

acute averaging period. The averaging periods were set forth 

in 1985 in the *'Guidelines for Deriving...National Criteria". 

In the near future EPA intends to review and perhaps modify 

8 
10. 

11. 

the Guidelines, and may at that time consider the feasibility 

of different acute averaging periods for different chemicals. 

The rationale for the I-day averaging period is weak. Chronic 

toxicity tests for most species require much longer than four 

days. The field studies presented in Chapter 1 indicate that 

longer term excursions are needed to produce ecological 

effects. EPA inappropriately picked examples of pollutants 

with low acute-chronic ratios to justify the r-day averaging 

period. A do-day averaging period (per the freshwater ammonia 

oriteria document) should be acceptable, particularly in cases 

where concentrations do not change rapidly. 

EPA Response: The TSD has been modified to note that EPA 

selected the 4-day chronic averaging period based on the 

shortest period that chronic effects may be observed for 

certain chemicals. 

As the 4-day period was selected for provide adequate 

protection in all cases, EPA recognizes that longer averaging periods may be appropriate for many pollutants, and 

recommends using site-specific or state-wide pollutantspecific alternative averaging periods, where scientifically 

supported. EPA believes that selection of an appropriate 

chronic averaging period is technically difficult, with less 

applicable data than is available for selecting the acute 

averaging period. EPA agrees that the acute-chronic ratio is 

a confounding influence in interpreting the duration needed 

to produce a chronic effect endpoint. Where an appropriate 

acute criterion is in force, the chronic averaging period need 

not be shortened simply because the acute-chronic ratio is 

low. 

The TSD does not supersede the freshwater ammonia criteria 

document, and thus EPA still recommends an averaging period 

of as long as 30 days for ammonia, where concentrations do 

not vary excessively. 

EPA's recommended once in three year return frequency for 

criteria excursions is overly conservative. Appendix D 

presents time periods needed for ecological recovery from 

severe or catastrophic stresses, not slight stresses caused 

by marginal criteria excursions. The frequency of significant criteria excursions, comparable to those that caused 

the measurable ecological impacts set forth in Chapter 1, 

would be much less than for marginal criteria excursions. 

EPA should present data on the ecological differences between 

sites with different excursion frequencies. EPA should 

develop guidance on how to establish site-specific allowable 

frequencies. 

9 
EPA Response: EPA has used Appendix D to set forth information on time periods needed for ecological recovery from 

severe or catastrophic stresses. EPA's recommended 3-year 

return interval was set forth in the 1985 "Guidelines for 

Deriving Numeric National Criteria...", and a review or 

revision of this recommended frequency was not within the 

scope of the TSD. Nevertheless, EPA intends to address the 

excursion frequency during the upcoming revision of the 

Guidelines. 

EPA expects that criteria exceedances can cause adverse 

effects and that the magnitude of the effect will depend on 

many things including the magnitude and duration of the 

exceedance. EPA believes that all adverse effects are not 

necessarily unacceptable, but that pollution should not be 

allowed to subject aquatic communities to long-term or regular 

short-term adverse effects. All dramatic adverse effects are 

certainly unacceptable. 

EPA believes that the 3-year return interval can be justified 

by the Appendix D data if one makes the assumption that the 

type of ecological impact shown in Appendix D could be caused 

by fairly small criteria excursions. The concentrations 

causing the Appendix D impacts were in fact not known. EPA 

recognizes that the chemical and ecological field data 

summarized in Chapter 1 suggest that successive excursions 

well above the criteria would be needed to cause severe 

impacts. EPA also recognizes that the probability of large 

excursions can be calculated to be extremely small compared 

to the probability of marginal excursions. 

EPA does not have information to allow direct comparisons of 

ecological quality versus criteria excursion frequency, except 

possibly as could be inferred from the field data shown in 

Chapter 1. EPA does not intend at this time to set forth 

guidance on developing site-specific allowable frequencies. 

Nevertheless, in general, EPA recommends that ecosystems not 

spend a substantial portion of time in a state of recovery 

from pollution stresses, and that pollution stresses not 

significantly increase the total stress experienced by 

organisms in the ecosystem. If the criteria are set appropriately, a marginal excursion might be expected to have 

little or no measurable impact, and little or no time period 

needed for recovery. The probability of a marginal criteria 

excursion nevertheless has a calculable relationship with the 

probabilities of severe criteria excursions. Consequently, 

a scientifically justified site-specific or state-wide 

frequency could be developed by considering (a) the 

probability (estimated by simulation or by statistical 

calculation) of a range of excursions of differing severity, 

coupled with (b) the estimated ecological recovery period for 

10 
the corresponding different degrees of impact. Based on the 

total period of recovery from a full range of possible events, 

compared with the sum of return intervals for such events, the 

allowable frequency for the marginal criteri'a excursion could 

be established. 

12. The once in three year excursion frequency does not take into 

consideration the likelihood of apparent excursions caused by 

the inherent variability of the analytical tests. 

EPA Response: The allowable frequency for criteria excursions should refer to true excursions of the criteria, not to 

spurious excursions caused by analytical variability or error. 

In evaluating data on chemical concentrations or toxicity 

units, it is desirable to subtract the analytical error log 

variance from the observed log variance in order to arrive at 

the true log variance contributing to criteria excursions. 

13. The IRIS (Integrated Risk Information System) data base should 

be peer reviewed before it is used to update and generate 

RACS. The TSD should not imply that a State can update its 

standards simply by inserting the latest XRIS information into 

the equation used to calculate the criteria. 

EPA Response: EPA's IRIS data base reflects the latest 

information about the Agency's.health assessments for specific 

chemicals. While the material in the data base is internally 

reviewed, the Agency does not plan to have it undergo external 

peer review. Use of IRIS information for developing state 

water quality standards or discharge permits in no way 

relieves the State of applicable requirements for public 

notice and comment. 

14. The fish consumption rates used to derive residue-based 

criteria and RACs are unrealistically high for many waters. 

EPA Response: EPA recommends using site-specific fish 

consumption rates whenever such information can be obtained. 

15. The harmonic mean is not an appropriate design flow. The 

arithmetic mean flow should be used. 

EPA Response: For carcinogens it is appropriate to determine 

the long-term arithmetic mean exposure concentration. Because 

flow is not normally distributed, using the arithmetic mean 

flow for design purposes will underestimate the mean 

concentration. 

Using the downstream harmonic mean flow will result in closely 

estimating the mean concentration, providing that the 

streamflow is not dominated by the effluent flow, and provided 

that the effluent input is not correlated to the streamflow. 
16. Discussion of sediment criteria and biological criteria is 

premature. The TSD should not advocate that states and 

regions implement regulatory controls based on such criteria. 

EPA Response: The biological criteria and sediment criteria 

sections have been modified. EPA has undertaken development 

of biological criteria and sediment criteria with the intent 

that they would, after development, have regulatory applications. 

EPA does not .intend to imply that these approaches can or 

should necessarily be used at this time to implement controls. Nevertheless, EPA believes that the states, the 

regulated community, and the interested general public need 

to know how EPA is proceeding with these criteria, and what 

the future regulatory implications may be. EPA is not 

advocating that sediment criteria, by themselves, be used to 

establish remediation target levels. EPA also recognizes that 

sediment criteria cannot be used for setting discharge limits 

without first developing a scientifically sound basis for 

predicting the effect of effluent quality on sediment quality. 

EPA is not suggesting that sediment quality concerns would 

necessarily be more limiting on dischargers than water quality 

concerns. 

17. Aquatic life protection as measured by whole effluent toxicity 

and chemical specific criteria are not applicable to waters 

without aquatic life designated uses. 

EPA Response: The TSD explains that numeric water quality 

criteria are developed by States to protect the designated 

uses within the water quality standards. However, the TSD 

also reiterated EPA's position, as expressed in the June 2, 

1989, Federal Resister preamble on the 304(l) promulgation 

was added, that the narrative criteria apply to all waters to 

prohibit acute toxicity. 

18. The food chain multiplier factors in the RAC calculation 

should be deleted because it is contrary to measured levels 

and BCF estimations. 

EPA Response: The differences between bioconcentration and 

bioaccumulation have been recognized in the scientific 

literature for a number of years. Data published by Thomann 

in Environmental Science and Technolosv (June 1989) show that 

bioaccumulation can be over 100 times higher than 

bioconcentration. Published critiques on EPA's dioxin 

criterion have also expressed that bioaccumulation is more 

important than bioconcentration for pollutants with log water 

octanol partition coefficients greater than 6. EPA believes 

12 
that consideration of bioaccumulation through use of the food 

chain multiplier is consistent with the existing knowledge of 

bioconcentration factors. 

Chauter 3: Effluent Characterization 

Overview: Simplified the chapter organization; cited regulatory 

requirements (40 CFR 122.44(d)); revised the discussion on effluent 

bioconcentration evaluation to conform with the new draft document. 

1. Determining the need for an effluent limit in the absence of 

effluent monitoring data does not address the required factors 

of 40 CFR 5 lzz.rr(d)(l)(ii). 

EPA Response: EPA maintains the position that regulatory 

authorities may make a finding of reasonable potential even 

where effluent monitoring data is not available. EPA 

disagrees with the commenter on the issue of whether the 

required factors can be addressed in the absence of effluent 

monitoring data. Box 3-l of the TSD was revised to illustrate 

how the recommended procedure for making reasonable potential 

determinations in the absence of effluent monitoring data will 

address each of the 4 required factors. 

2. Including bioconcentration and bioaccuntulation recommendations in the TSD is premature. Approaches to this problem 

require much more peer review, input and development. 

EPA Response: EPA agrees with this comment. The draft 

bioconcentration guidance that was referenced in the draft 

chapter 3 has not yet been released for public comment and 

does require additional peer review. The majority of the 

chapter 3 discussion on the specifics of this guidance was 

removed; the remaining sections specifically state that the 

procedures in the draft guidance should not be used by 

regulatory authorities until the guidance is finalized by EPA. 

3. In determining reasonable potential, the cumulative effluent 

discharge to a receiving water should be considered instead 

of single discharges. 

EPA Response: EPA agrees with this comment. The draft of 

chapter 3 did not draw a clear enough distinction between the 

terms @*causelV and l'contributes to" in the context of the 

reasonable potential determination. Chapter 3 now states that 

where multiple discharges collectively are causing or show the 

reasonable potential to cause or contribute to an excursion 

of water quality standards, limits must be developed for each 

discharger to protect against such collective excursions. 

This is underscored by adding the exact regulatory language 

of 122.44(d). Finally, the document now has recommendations 

on the use of toxicity testing in multiple discharge 

situations. 

13 
4. Guidance is needed on how to demonstrate that crhemicalspecific limits alone are sufficient to achieve applicable 

water quality standards, thus obviating the need for a WET 

limit. 

EPA Response: EPA agrees with this comment. Regulations at 

122.44(d)(l)(v) provide that WET limits are not necessary 

where the permitting authority demonstrates that chemicalspecific limits are sufficient to protect water quality 

standards. The draft of chapter 3 did not clearly recognize 

this point and provided no guidance on how to make this 

demonstration. Chapter 3 now reiterates this regulatory 

provision with a new section entitled VJsing a Chemicalspecific Limit to Control Toxicity.11 This new section 

recommends that the discharger conduct a TIE to identify 

causative toxicants. Where the causative toxicants are 

controlled by chemical-specific limits, the permitting 

authority may make the determination that WET limits are not 

necessary. 

5. The multiple conservative assumptions in the effluent 

characterization methodology (the effluent is most toxic to 

the most sensitive life stage at the time of lowest stream 

flow and peak design flow) are overkill. 

EPA Response: EPA does not agree that multiple conservative 

assumptions amount to overkill. This comment implies that 

EPA recommends establishing effluent limits to protect against 

toxic impacts that are never projected to occur. In truth, 

EPA only recommends establishing effluent limits where toxic 

impacts are projected to occur. Estimates of toxic impact 

should be made assuming that the effluent is most toxic to the 

most sensitive species or lifestage at the time of lowest 

available dilution because these are conditions that can be 

expected to occur. For the most part chapter 3 remains 

unchanged as a result of this comment. However, EPA has added 

a short discussion that suggests that the regulatory authority 

may choose to assess reasonable potential using a stochastic 

dilution model which incorporates both ambient dilution and 

effluent variability to project toxic impact. 

6. EPA should not recommend a 3 species minimum. The algae test 

and the marine tests are not sufficiently studied. 

EPA Response: Chapter 3 continues to recommend as a minimum 

that 3 species be tested quarterly for a minimum of 1 year 

where toxicity tests are used to make decisions regarding the 

need for WET limits. Experience indicates that algal tests 

can be a highly sensitive test species for some pollutants. 

Furthermore, using a plant adds another trophic level to the 

test regimen. EPA rejects the assertion that the algal tests 

are not sufficiently studied. For both freshwater and marine 

14 
waters, the use of 3 species is more protective than 2 species 

since a wider range of species sensitivity can be measured. 

7. EPA should not recommend that ambient toxicity tests be 

conducted at worst case low flow conditions. 

EPA Response: EPA disagrees with this comment. Chapter 3 

continues to recommend that ambient toxicity testing be 

conducted during appropriate low flow or worst case design 

periods. In order for the results of ambient toxicity tests 

to form the basis for decisions about whether toxicity 

controls are needed, the test must reflect the conditions that 

such controls would be designed to protect. If a regulatory 

authority's policy is to protect at the 7410 flow, than the 

ambient tests must be conducted at flows that are very near 

the 7410. Otherwise, the regulatory authority will learn very 

little from the ambient test about whether toxicity limits are 

necessary for a particular discharge. 

8. Reasonable potential determinations should not be based on 

whole effluent toxicity data alone. Toxicity data and 

instream survey data should be used together in a weight of 

evidence approach. 

EPA Response: We disagree. As discussed in the response to 

comment no. 2 in Chapter 1, EPA considers that water quality 

standards apply independently of each other. Whole effluent 

toxicity measures a different biological endpoint than do 

instream survey data. 

9. One piece of effluent data projecting an excursion above a 

water quality standard is insufficient to justify setting an 

effluent limit. 

EPA Response: EPA disagrees. EPA's position is that where 

even one data point shows that an excursion of a state water 

quality standard is projected, the Director may determine that 

permit limits for whole effluent toxicity or for specific 

chemicals are necessary. In making such a determination, 

NPDES regulations requires that the Director also account for 

existing controls on point and nonpoint sources, the 

variability of the pollutant parameter in the effluent, the 

sensitivity of the species to toxicity testing (for whole 

effluent), and where appropriate the dilution of the effluent 

in the receiving water. In addition, the Director should 

consider all other available information pertaining to the 

discharger to assist in making an informed judgement. 

15 
Chapter 4: Exnosure Assessment and Wasteload Allocation 

Overview: Clarified terminology for mixing zone and design flow 

conditions. 

1. Clarify or reevaluate Agency's position with respect to key 

mixing zone issues: including how to establish geographical 

boundaries of mixing zones and point of application of 

criterion for persistent bioaccumulative pollutants, and how 

to prevent lethality in the mixing zone. 

EPA Response: The revised TSD states as it did in previous 

drafts that the mixing zone size should be minimized and the 

dimensions should be based on the site-specific conditions. 

As previously stated, site-specific evaluations should also 

be conducted by the permitting agency to determine whether to 

allow a mixing zone for discharge of bioaccumulative 

pollutants. The TSD was revised to expand the three 

approaches for preventing lethality in the mixing zone to four 

approaches. The new approach includes submission of actual 

data to show that a drifting organism would not be exposed to 

l-hour average concentrations exceeding the CMC. In addition, 

clarification was provided on the approach not requiring the 

use of a high velocity discharge to show that the CMC is met. 

Approaches for preventing lethality in the mixing zone conform 

with the position taken in Chapter 2. 

2. Clarify or reevaluate Agency's position on appropriate 

critical flows for toxicants which have potential human health 

impacts and/or aquatic life impacts. 

EPA Response: It was determined that the harmonic mean flow 

was appropriate for evaluation of human health impacts that 

are of concern due to long-term exposures (e.g., cancer). 

The TSD was revised to clarify the use of harmonic mean flow 

and its appropriateness for use in water quality modeling to 

evaluate human health impacts (e.g., a step-by-step calculation procedure was added). Additional clarification was 

provided for the use of hydrologically-based 7410 and lQl0 

flows for the evaluation of "worst case" scenarios for 

determining potential aquatic life impacts. 

3. EPA should present u valid modeling approaches and determine the best water quality model for each water quality 

condition. 

EPA Response: The models described in the text were those 

EPA considers to be comprehensive to evaluate most water 

quality conditions. Because of all the potential sitespecific scenarios, it is beyond the scope of the document to 

present al-J of the valid approaches that could be used for 

16 
water quality modeling. The permitting authority needs to 

evaluate the data available for the site and select the water 

quality model that is best for calculating the receiving water 

concentrations and the TMDLs. 

4. specific caveats should be added to the model descriptions. 

EPA Response: It is beyond the scope of the document to 

provide all of the caveats that are applicable for each of 

the models. Although the commenters presented specific 

caveats for a model, they were not incorporated since it was 

beyond the scope of revising the document to determine all 

the cases that the caveat could apply. 

5. Commenters want additional explanation or justification for 

the assumptions and applications for the equations presented 

in the TSD. 

EPA Response: The TSD already contained a sufficient explanation or referenced the documents which contained the 

rationale. 

6. The water quality model "DYNTOX1# and the software program RRD 

FLOW are not available. 

EPA Response: DYNTOX is not currently available but should 

be accessible by early 1991. HHD FLOW is not available, but 

DFLOW is available, and the TSD has been changed to reference 

this software package. 

7. Guidance on regulating nonpoint sources should be included in 

the TSD. 

EPA Response: It is beyond the scope of the document to 

include information on regulating non-point sources. 

8. Guidance should be given in the discussion on design flow for 

persistent pollutants. 

EPA Response: Persistent pollutants should be assessed in 

the same way as bioaccumulative pollutants. 

9. The human health section does not provide direction regarding the percentage of fish that are taken from a given area. 

EPA Response: The percentage of fish should be determined 

based on site-conditions. The WLA criteria are conservative 

estimates based on heavy consumption of fish or a potentially 

large contaminated area. 

17 
10. The TSDs discussion on mixing Zone8 contribution to human 

health is overemphasized. 

EPA Response: The TSDs discussion on mixing zones has been 

revised and does not overemphasize the contributions to human 

health. 

11. EPA should revise its discussion of modeling WLAs for human 

health toxicants because it is misleading since it suggests 

that the method should be used whether the applicable 

criterion is a drinking water standard intended to prevent 

acute effects on humans or criterion established for Deriving Numerical National Water Quality Criteria. 

EPA Response: The TSD does not discuss acute human health 

effects; therefore, it was determined that the discussion is 

not misleading. 

Chapter 5: Permit Reffuirements 

Overview: Added clarifications and examples to support the 

existing text; gave equal weight to developing limits based on a 

dynamic and steady state wasteload allocations; added discussion 

on metals, average and maximum permit limits, single dilution 

tests, variability, and mass-based limits. 

1. The two value, steady-state model approach to permit limit 

derivation is overly conservative with too many built-in 

safety factors. As a result the limits derived are too 

stringent. 

EPA Response: EPA has revised Chapter 5 of the TSD to 

emphasize the development of permit limits that are as exact 

as possible to attain and maitain water quality standards. 

Chapter 5 now recommends the use of the statistical limit 

derivation procedure which provides for two options. The 

first option uses a steady state model approach for development of the wasteload allocation (WLA) and long term average 

(LTA). Since this approach relies on critical condition 

assumptions regarding effluent characteristics and receiving 

water characteristics, this approach may derive limits that 

are more restrictive than the second option. The effluent 

characteristics of importance are pollutant concentrations, 

pollutant concentration variation, and effluent flow. The 

receiving water characteristics of importance are pollutant 

concentrations and receiving water flow. 

Where a discharger or permitting authority believes that the 

steady state model approach results in overly restrictive 

permit limits, under the revised TSD the discharger or 

permitting authority has the alternative of using a dynamic 

18 
model approach. Under the dynamic model approach worst case 

assumptions are minimized and more accurate receiving water 

concentrations of a pollutant can be calculated. In general, 

dynamic models account for the daily variations of and 

relationships between flow, effluent and environmental 

conditions and therefore directly determine the probability 

that a water quality standards exceedance will occur. Because 

of this, dynamic models can be used to develop wasteload 

allocations which more exactly maintain the water quality 

standards at the return frequency requirements of the 

standards. The WLA is first developed by iteratively running 

the dynamic model with successively lower LTAs until the model 

shows compliance with the water quality standards. With this 

approach now a recommended option in the TSD, this major issue 

is resolved. A disadvantage of using dynamic model outputs 

to develop permit limits is the lack of necessary data for 

effluent variability and receiving water flows. 

2. The discussion of below detection levels is confusing. EPA 

should not be setting water quality-based limits below 

detection levels, especially since variability of a test 

method is greater the closer the results are to the detection 

level. 

EPA Response: The discussion of below detection level limits 

in the draft TSD was unclear. Since the time of the draft 

TSD, EPA issued its guidance on setting permit limits below 

the detection level for dioxin. This guidance uses a minimum 

level (ML) to ascertain compliance with limits set below 

detection levels. The TSD was revised to follow the dioxin 

guidance (May 21', 1990 Memorandum from LaJuana Wilcher, 

"Strategy for the Regulation of Discharges of PHDDs and PHDFs 

from Pulp and Paper Mills to Waters of the United StateslV). 

Specific values for the minimum level are found in the 

description of methods 1624 and 1625 for some organic 

compounds. (See the appendix to 40 CFR 136.) 

3. EPA's criteria recommend that limits be derived for toxicant 

in the soluble form or biologically available form, yet many 

permit limits are being written and compliance being based 

upon the total recoverable form of metals. The TSD should 

address this issue and give guidance on developing limits only 

for bioavailable forms of pollutants. 

EPA Response: The TSD has been revised to cite the regulatory requirements at 40 CFR 122.45(c) and to provide three 

options for use where a state has not developed a method for 

determining total recoverable permit limits based on a 

dissolved or acid soluble water quality criterion. The three 

options are 1) assume complete availability of the total 

recoverable metali 2) use the method in the EPA wasteload 

allocation guidance manual for toxics in rivers to relate the 

19 
two, and 3) use site specific data to develop a relationship. 

4. EPA's regulations do not allow permit limits based on one day 

maximum concentrations for POTBS. 

EPA Response: The NPDES regulation at 40 CFR 122.45(d) 

require the use of a 7-day average unless impractical. The 

discussion on the expression of permit limits now states that 

EPA considers the 7-day average limit for POTWs to be 

impractical for the purposes of controlling the discharge of 

toxics. The reason for this statement is that control of the 

7-day average in lieu of control of the l-day maximum will 

allow for unmeasured short-term excursions of an acute water 

quality standard. 

5. TREs should not be required as a permit condition to respond 

to a violation of a whole effluent toxicity limit. The proper 

response by the permitting authority is through use of 

enforcement mechanisms. 

EPA Response: The discussion on TREs was changed to not 

recommend that the TRE or accelerated monitoring be included 

in the permit to respond to permit limit violations but rather 

be part of the enforcement response. This was done to help 

distinguish between monitoring only provisions and enforcement 

of permit limits. This should reduce confusion about the 

difference between enforcement on a single event violation of 

a limit and the need for multiple violations before a TRE is 

warranted. 

6. Permit limits should be increased to consider analytical 

variability. EPA should consider using the approach it 

proposed in the Amelia River study. 

EPA Response: EPA disagrees with the proposal. A discussion 

was added to explain how EPA considers analytical variability 

in developing permit limits. The discussion states that since 

this variability is an intrinsic part of all data collection 

(effluent monitoring, wasteload allocation development, and 

water quality standards development), and that the variability 

can go both ways (higher or lower), that EPA does not separate 

it out from all other variability factors. EPA's Amelia River 

study is not final; the approach cited may not be included in 

the final report. 

Although difficult, it may be possible to determine what 

proportion of the observed variability can be attributed to 

sampling error, and what proportion can be attributed to the 

method of measurement. Regardless, the TSD makes use of a 

coefficient of variation that includes both sources of 

variability. This is not unreasonable since sampling for 

monitoring purposes also results in the inclusion of these 

20 
two sources of variability. 

concentrations 

Any consideration of upstream 

variability. 

in the WLA will also include analytical 

There is no "true" calculation in the process 

of developing permit limits because there is analytical 

uncertainty throughout the process. Instead, LTAs should be 

calculated from the WLAs using the same CV that is also used 

to calculate permit limits from the most limiting LTA. It is 

unimportant exactly what CV is used because the most 

restrictive LTA itself is used only for determining the 

desired treatment performance level. However, this procedure 

assures (99 or 95% confident) that the permit limits will be 

less than or equal to the more limiting LTA. 

7. Since permit limits were derived based on 95th and 99th 

percentile probabilities, that occasional 9XC88danC8S of 

permit limits should be allowed on the same basis. 

EPA Response: In statistics, the selection of an acceptable 

probability level reflects the level of confidence that is 

desired of the results. As such, an acceptable level must be 

defined prior to performing any statistical procedure. As 

stated in the TSD, the probability basis of 0.99 for the daily 

maximum limit, and 0.95 for the average monthly limit have 

been used historically in connection with development of the 

effluent guideline limitations and have been well accepted 

upheld in legal challenges to the guidelines. These values 

are tied to monitoring frequencies that are required for each 

limit. There is no mixing of two probability bases since they 

are distinct and separate limits. The goal in establishing 

these levels is to allow the regulatory agency to distinguish 

between adequately operated wastewater treatment plants with 

normal variability from poorly operated treatment plants. 

8. Permit limits should vary with flow conditions in the 

receiving waters since exposure is baS8d on dilution with 

receiving water flow. 

EPA R8SpOnS8: EPA partially agrees but only to the extent 

that the limits are seasonally based. The seasonal approach 

has been used by permitting authorities for setting permit 

limits to protect against excursions of dissolved oxygen and 

ammonia standards. However, seasonal limits are different 

than limits which vary daily based on river flow. EPA is not 

convinced that a daily variable approach would be universally 

practicalgivenwastewatertreatmentresponse and performance; 

for this reason EPA has not included procedures for this 

approach. In addition, the discharger has the option of using 

dynamic modeling to develop permit limits. Since dynamic 

modeling considers all receiving water flows, this option 

would provide the discharger a less restrictive permit limit 

than would be obtained by using steady state modeling. 

21 
9. If limits W8r8 derived that W8r8 Overly Stringent due to the 

worst case assumption of the Steady stat8 model, and if in 

the future the permittee conducted dynamic modeling which 

r8SUlt8d in less stringent limits , permittees would be bound 

to the previous limits due to the anti-backsliding requirements. 

EPA R8SpOnS8: There is no absolute prohibition. EPA's 

September 1989 guidance document on antibacksliding contains 

EPA's interpretation of the Clean Water Act provisions in 

§402(0) and §303(d)(4). This guidance also does not contain 

an absolute prohibition. It is EPA's position that the CWA's 

anti-backsliding provision and EPA's existing regulations do 

not uniformly prohibit the incorporation into a permit of less 

stringent limits, standards, or conditions. In certain 

situations (i.e., under Sections 402(o) and 303(d)(4) of the 

CWA) , 

less stringent limits or conditions may be permissible. 

Section 402(o)(l) provides that backsliding from water 

quality-based limits is prohibited except in compliance with 

Section 303(d)(4). Section 303(d)(4)(A) only allows establishment of less stringent limits in a permit for discharge 

into a non-attainment water only if two conditions are met: 

1) the existing permit limit must have been based on a TMDL 

or other WLA established under Section 303, and 2) attainment 

of water quality standards must be assured. Section 

303(d)(4)(B) allows establishment of less stringent limits in 

a permit for discharge into an attained water only where 

relaxation is consistent with a State's antidegradation 

policy. 

Section 402(o)(2) also outlines exceptions to the general 

prohibition against backsliding from water quality-based 

permit limitations. Under Section 402(o)(3), backsliding may 

be allowed: 1) where there have been material and substantial 

alterations or additions to the facility; 2) where good cause 

exists due to events beyond the permittee's control and for 

which there is no reasonably available remedy; 3) where the 

permittee has installed and properly operated and maintained 

required treatment facilities; and 4) where new information 

justifies backsliding from water quality-based permit 

limitations and other Section 301(b)(l)(C) limitations. 

10. EPA does not have guidance on how to conduct a chronic TRE 

and therefore limits should not b8 derived based upon chronic 

endpoints. 

EPA R8SpOnS8: EPA is aware of the need for guidance on 

conducting TREs for chronic toxicity. EPA's Duluth laboratory is near completion of a draft guidance document. The 

guidance document will be widely available when finished. 

Regardless, the lack of a finished guidance document is not 

22 
a valid reason to allow for excursions above a narrative 

standard as measured by chronic whole effluent toxicity. 

Some dischargers have been able to comply with chronic 

toxicity limits and identify and control sources of chronic 

toxicity. EPA experience has shown that portions of the 

published TRE procedures for solving incidences of acute 

toxicity (EPA/600/2-88/070, EPA/600/2-88,'062, EPA/600/3- 

88/034) can be used for resolving incidences of chronic 

toxicity. 

11. EPA should allow adequate time for facilities to come into 

compliance with water quality-based permit limits. 

EPA Response: EPA is aware that facilities may need time to 

comply with newly established effluent limits. This has been 

accomplished in the past in NPDES permits by allowing 

compliance schedules within the permit. Regulatory agencies 

use of compliance schedules for water quality-based effluent 

limits are governed by recent decisions regarding the StarKist Caribe ruling by EPA's Chief Judicial Officer (CJO). On 

March 8, 1989, in review of the evidentiary hearing request 

by Star-Kist Caribe, the CJO ruled that compliance schedules 

for water quality-based effluent limits may not be included 

in NPDES permits unless explicitly authorized by the State in 

its water quality standards or implementation regulations. 

The ruling was based on an interpretation of section 

301(b)(l)(C) of the Clean Water Act. Later, on Septermber 4, 

1990, the CJO granted a stay of the ruling to allow EPA and 

States to use compliance schedules for water quality-based 

limits where such schedules are consistent with State policy. 

In any case, the allowance for compliance schedules is a State 

decision which may ultimately need to be expressed in water 

quality standards. 

12. Effluent limits should b8 set within the ability of treatment 

t8ChnOlOgy. 

EPA R8SpOnS8: EPA is aware that there may be a number of 

water quality-based permit limits for toxics which may be 

presently unachievable with existing wastewater treatment 

technology. However, the NPDES regulations at 40 CFR 

122.44(d) require that effluent limits more stringent than 

those established based on a treatment technology basis must 

be set to achieve water quality standards. 

13. The limit derivation prOC8dUr8S for human health should us8 

the same Statistical prOC8dUreS as used for deriving limits 

for aquatic life. Permit limits should be derived from the 

harmonic mean effluent concentration. 

EPA Response: Since compliance with permit limitations is by 

regulation determined on a daily and monthly basis, it is 

necessary to set permit limitations expressed in these 

23 
14. 

contexts that meet a given WLA every month. The statistical 

procedures for permit limit derivation in the TSD are designed 

to accomplish this for aquatic life protection where the use 

of shorter term averaging periods is consistent with two 

number aquatic life criteria. 

HOWeVer, if the TSD procedures were directly used for setting 

permit limits on bioconcentratable pollutants, both maximum 

daily and average monthly permit limits could exceed the 

wasteload allocation necessary to meet the criterion. These 

two permit limits would assure that the long term average 

effluent discharge would comply with the human health derived 

WLA onlv if the assessment of the effluent variability was 

precise. With bioconcentratable pollutants where exposure 

duration ranges up to 70 years, EPA believes that effluent 

variability cannot be reliably estimated from existing data 

for exposure periods a year. If the effluent variability was 

over-estimated when establishing the permit limits, then a 

facility could be discharging in compliance with the permit 

limits but would be exceeding the wasteload allocation for 

human health protection. This approach is clearly 

unacceptable. 

This problem does not arise when using the TSD statistical 

procedure for setting permit limits for protecting against 

aquatic toxicity. In this case, the monthly average and daily 

maximum permit limits are more closely related to the four day 

average and one hour maximum used as exposure periods for the 

criteria. Any imprecision in assessing effluent variability 

would therefore not have as great an effect on the permit 

limits. 

Effluents may not always demonstrate a log-normal distribution. The TSD should present procedures for using Other 

distributions. 

EPA R8SpOnS8: EPA believes, after reviewing the database used 

to establish effluent guidelines, that the log-normal 

distribution best characterizes effluents. EPA's analysis of 

these data are provided in Appendix E. The general 

characteristics of the lognormal distribution (it is only 

positive and is skewed towards extreme high values) make it 

an appropriate distribution for dealing with effluent 

concentrations. According to Gilbert (1987), the lognormal 

distribution is the only available 2-parametric distribution 

that can routinely be applied to environmental data. Since 

the Agency is not providing derivation procedures for 

alternate data distributions, it is not necessary to test for 

lognormality of effluent data. Permitting authorities can 

develop their own methods using other probabilistic 

distributions. 

24 
Chapter 6: Enforcement 

oV8rVi8W: No major changes made in recommendations; added 

clarifications and explanations. 

1. One effluent test failure should not equal a violation; 

provisions should b8 mad8 to delay punitive enforcement action 

Wh8r8 legitimate efforts are being made; permittees should not 

be held in violation of their toxicity limit While conducting 

a TRE. 

EPA R8SpOnS8: The current draft still emphasizes the 

principle that any single exceedance is a violation and is 

subject to a full range of enforcement responses. However, 

the draft has been caveated with discussion on EPA's guidance 

which outlines a systematic review of all violations to 

determine the appropriate level of enforcement. EPA's 

enforcement guidance is included in an appendix. 

25 
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