r/EPA
United States
Environmental Protection
Agency
Office of Water
(4305)
EPA-823-B-94-005a
August 1994
Water Quality Standards
Handbook:
Second Edition
"... to restore and maintain the chemical,
physical, and biological integrity of the Nation's
waters."
Contains Update #1
August 1994
Section 101 (a) of the Clean Water Act
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WATER QUALITY STANDARDS
HANDBOOK
SECOND EDITION
Water Quality Standards Branch
Office of Science and Technology
U.S. Environmental Protection Agency
Washington, DC 20460
September 1993
Contains update #1
August 1994
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Table of Contents
FOREWORD
Dear Colleague:
The following document entitled Water Quality Standards Handbook - Second Edition provides guidance
issued in support of the Water Quality Standards Regulation (40 CFR 131, as amended). This Handbook includes
the operative provisions of the first volume of the Handbook issued in 1983 and incorporates subsequent guidance
issued since 1983. The 1993 Handbook contains only final guidance previously issued by EPAit contains no
new guidance.
Since the 1983 Handbook has not been updated hi ten years, we hope that this edition will prove valuable
by pulling together current program guidance and providing a coherent document as a foundation for State and
Tribal water quality standards programs. The Handbook also presents some of the evolving program concepts
designed to reduce human and ecological risks, such as endangered species protection; criteria to protect wildlife,
wetlands, and sediment quality; biological criteria to better define desired biological communities in aquatic
ecosystems; and nutrient criteria.
This Handbook is intended to serve as a "living document," subject to future revisions as the water quality
standards program moves forward, and to reflect the needs and experiences of EPA and the States. To this end,
the Handbook is published in a loose leaf format designed to be placed in three ring binders. This copy of the
Handbook includes updated material for 1994 (see Appendix X), and EPA anticipates publishing additional
changes periodically and providing them to Handbook recipients. To ensure that you will receive these updates,
please copy the reader response card in Appendix W and mail it to the address on the reverse.
The Handbook also contains a listing, by title and date, of the guidance issued since the Handbook was
first published in 1983 that is incorporated in the Second Edition. Copies of these documents are available upon
request.
The Water Quality Standards Handbook - Second Edition provides guidance on the national water quality
standards program. EPA regional offices and States may have additional guidance that provides more detail on
selected topics of regional interest. For information on regional or State guidance, contact the appropriate
regional water quality standards coordinator listed in Appendix U.
EPA invites participation from interested parties in the water quality standards program, and appreciates
questions on this guidance as well as suggestions and comments for improvement. Questions or comments may
be directed to the EPA regional water quality standards coordinators or to:
David Sabock, Chief
U.S. Environmental Protection Agency
Water Quality Standards Branch (4305)
401 M Street, S.W.
Washington, D.C. 20460
Telephone (202) 475-7315
Betsy Southerland, Acting Director
Standards and Applied Science Division
(8/15/94) ' ill
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Water Quality Standards Handbook - Second Edition
Note to the Reader
The Water Quality Standards Handbook, first issued in 1983, is a compilation of EPA's
guidance on the water quality standards program and provides direction for States in reviewing,
revising and implementing water quality standards. The Water Quality Standards Handbook -
Second Edition retains all the guidance in the 1983 Handbook unless such guidance was specifically
revised in subsequent years. An annotated list of the major guidance and policy documents on the
water quality standards program issued since 1983 is included in the Introduction and material added
to the Second Edition by periodic updates since 1993 is summarized in Appendix X. Material in the
Handbook contains only guidance previously issued by EPA; it contains no new guidance.
The guidance contained in each of the documents listed in the Introduction is either:
1) incorporated in its entirety, or summarized, in the text of the appropriate section of this
Handbook, or 2) attached as an appendix (see Table of Contents). If there is uncertainty or
perceived inconsistency on any of the guidance incorporated into this Handbook, the reader is
directed to review the original guidance documents or call the Water Quality Standards Branch at
(202) 260-1315. Copies of all original guidance documents not attached as appendices may be
obtained from the source listed for each document in the Reference section of this Handbook.
Limited free copies of this Handbook may be obtained from:
Office of Water Resource Center, RC-4100
U. S. Environmental Protection Agency
401 M Street, S.W.
Washington, DC 20460
Telephone: (202) 260-7786 (voice mail publication request line)
Copies may also be obtained from:
Education Resource Information Center/Clearinghouse for Science, Mathematics and Environmental
Education (ERIC)
1929 Kenny Road
Columbus, OH 43210-1080 (Telephone: 614-292-6717)
(VISA, Mastercard and purchase order numbers from schools and businesses accepted)
U.S. Department of Commerce
National Technical Information Service (NTIS)
5285 Port Royal Road
Springfield, VA 22161 (Telephone: 1-800-553-6847)
(American Express, VISA and Mastercard accepted)
Robert S. Shippen
Editor
IV (8/15/94)
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Table of Contents
TABLE OF CONTENTS
Foreword iii
Note to the Reader iv
Table of Contents v
Glossary GLOSS-1
Introduction INT-1
History of the Water Quality Standards Program INT-1
Handbook Changes Since 1983 INT-5
Overview of the Water Quality Standards Program INT-8
The Role of WQS in the Water Quality Management Program INT-13
Future Program Directions INT-14
Chapter 1 - General Provisions (40 CFR 131 - Subpart A)
1.1 Scope - 40 CFR 131.1 1-1
1.2 Purpose - 40 CFR 131.2 1-1
1.3 Definitions - 40 CFR 131.3 1-1
1.4 State Authority - 40 CFR 131.4 1-2
1.5 EPA Authority - 40 CFR 131.5 . 1-3
1.6 Requirements for Water Quality Standards Submission - 40 CFR 131.6 ...... 1-4
1.7 Dispute Resolution Mechanism - 40 CFR 131.7 1-4
1.8 Requirements for Indian Tribes To Qualify for the WQS Program - 40 CFR
131.8 1-9
1.9 Adoption of Standards for Indian Reservation Waters 1-18
Endnotes 1-21
Chapter 2 - Designation of Uses (40 CFR 131.10)
2.1 Use Classification - 40 CFR 131.10(a) 2-1
2.2 Consider Downstream Uses - 40 CFR 131.10(b) 2-4
2.3 Use Subcategories - 40 CFR 131.10(c) 2-5
2.4 Attainability of Uses - 40 CFR 131.10(d) 2-5
2.5 Public Hearing for Changing Uses - 40 CFR 131.10(e) 2-6
2.6 Seasonal Uses - 40 CFR 131.10(f) 2-6
2.7 Removal of Designated Uses - 40 CFR 131.10(g) and (h) . 2-6
2.8 Revising Uses to Reflect Actual Attainment - 40 CFR 131.10(i) 2-8
2.9 Use Attainability Analyses - 40 CFR 131.10(j) and (k) 2-9
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Water Quality Standards Handbook - Second Edition
Chapter 3 - Water Quality Criteria (40 CFR 131.11)
3.1 EPA Section 304(a) Guidance 3-1
3.2 Relationship of Section 304(a) Criteria to State Designated Uses 3-10
3.3 State Criteria Requirements 3-12
3.4 Criteria for Toxicants 3-13
3.5 Forms of Criteria 3-23
3.6 Policy on Aquatic Life Metals Criteria 3-34
3.7 Site-Specific Aquatic Life Criteria 3-38
Endnotes 3-45
Chapter 4 - Antidegradation (40 CFR 131.12)
4.1 History of Antidegradation 4-1
4.2 Summary of the Antidegradation Policy 4-1
4.3 State Antidegradation Requirements 4-2
4.4 Protection of Existing Uses - 40 CFR 131.12(a)(l) 4-3
4.5 Protection of Water Quality in High-Quality Waters - 40 CFR 131.12(a)(2) .... 4-6
4.6 Applicability of Water Quality Standards to Nonpoint Sources Versus Enforceability
of Controls 4-9
4.7 Outstanding National Resource Waters (ONRW) - 40 CFR 131.12(a)(3) 4-10
4.8 Antidegradation Application and Implementation 4-10
Chapter 5 - General Policies (40 CFR 131.13)
5.1 Mixing Zones 5-1
5.2 Critical Low-Flows 5-9
5.3 Variances From Water Quality Standards 5-11
Chapter 6 - Procedures for Review and Revision of Water Quality Standards
(40 CFR 131 - Subpart C)
6.1 State Review and Revision 6-1
6.2 EPA Review and Approval 6-8
6.3 EPA Promulgation 6-13
Chapter 7 - The Water Quality-based Approach to Pollution Control
7.1 Determine Protection Level 7-2
7.2 Conduct Water Quality Assessment 7-3
7,3 Establish Priorities 7-5
7.4 Evaluate Water Quality Standards for Targeted Waters 7-6
7.5 Define and Allocate Control Responsibilities 7-7
7.6 Establish Source Controls 7-8
7.7 Monitor and Enforce Compliance 7-12
7.8 Measure Progress 7-13
References REF-1
VI (8/15/94)
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Table of Contents
Appendices:
A - Water Quality Standards Regulation - 40 CFR 131.
B - Chronological Summary of Federal Water Quality Standards Promulgation Actions.
C - Biological Criteria: National Program Guidance for Surface Waters, April 1990.
D - National Guidance: Water Quality Standards for Wetlands, July 1990.
E - An Approach for Evaluating Numeric Water Quality Criteria for Wetlands Protection,
July 1991.
F - Coordination Between the Environmental Protection Agency, Fish and Wildlife Service
and National Marine Fisheries Service Regarding Development of Water Quality
Criteria and Water Quality Standards Under the Clean Water Act, July 1992.
G - Questions and Answers on: Antidegradation, August 1985.
H - Derivation of the 1985 Aquatic Life Criteria.
I - List of EPA Water Quality Criteria Documents.
J - Attachments to Office of Water Policy and Technical Guidance on Interpretation and
Implementation of Aquatic Life Metals Criteria, October 1993.
K - Procedures for the Initiation of Narrative Biological Criteria, October 1992.
L - Interim Guidance on Determination and Use of Water-Effect Ratios for Metals,
February 1994.
M - Reserved.
N - IRIS [Integrated Risk Information System] Background Paper.
O- Reserved.
P - List of 126 Section 307(a) Priority Toxic Pollutants.
Q - Wetlands and 401 Certification: Opportunities and Guidelines for States and Eligible
Indian Tribes - April 1989.
R - Policy on the Use of Biological Assessments and Criteria in the Water Quality
Program, May 1991.
S - Reserved.
T - Use Attainability Analysis Case Studies.
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Water Quality Standards Handbook - Second Edition
U - List of EPA Regional Water Quality Standards Coordinators.
V - Water Quality Standards Program Document Request Forms.
W - Update Request Form for Water Quality Standards Handbook - Second Edition.
X - Summary of Updates
Vlll (8/15/94)
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Glossary
"Completely mixed condition" is defined as no measurable difference in the concentration of a
pollutant exists across a transect of the water body (e.g., does not vary by 5%) (USEPA,
1991a.)
"Criteria" are elements of State water quality standards, expressed as constituent concentrations, levels,
or narrative statements, representing a quality of water that supports a particular use. When
criteria are met, water quality will generally protect the designated use (40 CFR 131.3.)
"Criteria 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 (USEPA, 1991a.)
"Criteria 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 (USEPA, 1991a.)
"Critical life stage" is the period of time hi 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
(USEPA, 1991a.)
"Critical species" is a species that is commercially or recreationally important at the site, a species that
exists at the site and is listed as threatened or endangered under section 4 of the Endangered
Species Act, or a species for which there is evidence that the loss of the species from the site
is likely to cause an unacceptable impact on a commercially or recreationally important species,
a threatened or endangered species, the abundances of a variety of other species, or the structure
or function of the community (USEPA, 1994a.)
"Design flow" is the flow used for steady-state waste load allocation modeling (USEPA, 1991a.)
"Designated uses" are those uses specified in water quality standards for each water body or segment
whether or not they are being attained (40 CFR 131.3.)
"Discharge length scale" is the square root of the cross-sectional area of any discharge outlet (USEPA,
1991a.)
"Diversity" is the number and abundance of biological taxa hi a specified location (USEPA, 1991a.)
"Effective 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 (USEPA, 1991a.)
"Existing uses" are those uses actually attained in the water body on or after November 28, 1975,
whether or not they are included hi the water quality standards (40 CFR 131.3.)
(8/15/94) GLOSS-3
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Water Quality Standards Handbook - Second Edition
"Federal Indian Reservation," "Indian Reservation," or "Reservation" is defined as all land within
the limits of any Indian reservation under the jurisdiction of the United States Government,
notwithstanding the issuance of any patent, and including rights-of-way running through the
reservation (40 CFR 131.3.)
"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 (USEPA, 1991a.)
"Frequency" is how often criteria can be exceeded without unacceptably affecting the community
(USEPA, 1991a.)
"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 (USEPA, 199la.)
"Indian Tribe" or "Tribe" describes any Indian Tribe, band, group, or community recognized by the
Secretary of the Interior and exercising governmental authority over a Federal Indian reservation
(40 CFR 131.3.)
"Inhibition concentration" (1C) is a point estimate of the toxicant concentration that would cause a
given percent reduction (e.g., IC25) in a non-lethal biological measurement of the test
organisms, such as reproduction or growth (USEPA, 199la.)
"Lethal concentration" is the point estimate of the toxicant concentration that would be lethal to a
given percentage of the test organisms during a specified period (USEPA, 199la.)
"Lipophilic" is a high affinity for lipids (fats) (USEPA, 1991a.)
"Load allocations" (LA) the portion of a receiving water TMDL that is attributed either to one of its
existing or future nonpoint sources of pollution or to natural background sources (USEPA,
1991a.)
"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 hi chronic or subchronic
human epidemiology studies or animal exposure (USEPA, 1991a.)
"Magnitude" is how much of a pollutant (or pollutant parameter such as toxicity), expressed as a
concentration or toxic unit is allowable (USEPA, 1991a.)
"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 (USEPA, 1991a.)
"Mixing zone" is an area where an effluent discharge undergoes initial dilution and is extended to cover
the secondary mixing in the ambient water body. A mixing zone is an allocated impact zone
where water quality criteria can be exceeded as long as acutely toxic conditions are prevented
(USEPA, 1991a.)
GLOSS-4 (8/15/94)
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Glossary
"Navigable waters" refer to the waters of the United States, including the territorial seas (33 USC
1362.)
"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 (USEPA, 1991a.)
"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 (USEPA, 1991a.)
"Nonthreshold effects" are associated with exposure to chemicals that have no safe exposure levels.
(i.e., cancer) (USEPA, 1991a.)
"Persistent pollutant" is not subject to decay, degradation, transformation, volatilization, hydrolysis,
or photolysis (USEPA, 1991a.)
"Pollution" is defined as the man-made or man-induced alteration of the chemical, physical, biological
and radiological integrity of water (33 USC 1362.)
"Priority pollutants" are those pollutants listed by the Administrator under section 307(a) of the Act
(USEPA, 1991a.)
"Reference ambient concentration" (RAC) is the concentration of a chemical in water which will not
cause adverse impacts to human health; RAC is expressed in units of mg/1 (USEPA, 1991a.)
"Reference conditions" describe the characteristics of water body segments least impaired by human
activities. As such, reference conditions can be used to describe attainable biological or habitat
conditions for water body segments with common watershed/catchment characteristics within
defined geographical regions.
"Reference tissue concentration" (RTC) is the concentration of a chemical in edible fish or shellfish
tissue which will not cause adverse impacts to human health when ingested. RTC is expressed
in units of mg/kg (USEPA, 1991a.)
"Reference dose" (RfD) is an estimate of the daily exposure to human population that is likely to be
without appreciable risk of deleterious effect during a lifetime; derived from NOAEL or LOAEL
(USEPA, 1991a.)
"Section 304(a) criteria" are developed by EPA under authority of section 304(a) of the Act based on
the latest scientific information on the relationship that the effect of a constituent concentration
has on particular aquatic species and/or human health. This information is issued periodically
to the States as guidance for use in developing criteria (40 CFR 131.3.)
"Site-specific aquatic life criterion" is a water quality criterion for aquatic life that has been derived
to be specifically appropriate to the water quality characteristics and/or species composition at
a particular location (USEPA, 1994a.)
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Water Quality Standards Handbook - Second Edition
"States" include: the 50 States, the District of Columbia, Guam, the Commonwealth of Puerto Rico,
Virgin Islands, American Samoa, the Trust Territory of the Pacific Islands, and the
Commonwealth of the Northern Mariana Islands, and Indian Tribes that EPA determines qualify
for treatment as States for the purposes of water quality standards (40 CFR 131.3.)
"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 (USEPA, 1991a.)
"STORET" is EPA's computerized water quality database that includes physical, chemical, and
biological data measured hi water bodies throughout the United States (USEPA, 1991a.)
"Sublethal" refers to a stimulus below the level that causes death (USEPA, 1991a.)
"Synergism" is the characteristic property of a mixture of toxicants that exhibits a greater-than-additive
total toxic effect (USEPA, 1991a.)
"Threshold effects" result from chemicals that have a safe level (i.e., acute, subacute, or chronic
human health effects) (USEPA, 1991a.)
"Total maximum daily load" (TMDL) is the sum of the individual waste load allocations (WLAs) and
load allocations (LAs); 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 (USEPA, 1991a.)
"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 (USEPA, 1991a.)
"Toxic pollutant" refers to those pollutants, or combination of pollutants, including disease-causing
agents, which after discharge and upon exposure, ingestion, inhalation, or assimilation into any
organism, either directly from the environment or indirectly by ingestion through food chains,
will, or on the basis of information available to the administrator, cause death, disease,
behavioral abnormalities, cancer, genetic mutations, physiological malfunctions (including
malfunctions in reproduction) or physical deformations, hi such organisms or their offspring (33
USC section 1362.)
"Toxic units" (TUs) are a measure of toxicity hi an effluent as determined by the acute toxicity units
(TUa) or chronic toxicity units (TUc) measured (USEPA, 1991a.)
"Toxic unit acute" (TUa) 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/LC50) (USEPA, 1991a.)
"Toxic unit chronic" (TUc) 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., 100/NOEC)
(USEPA, 1991a.)
GLOSS-6 (8/15/94)
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Glossary
"Use attainability analysis" (UAA) is a structured scientific assessment of the factors affecting the
attainment of the use which may include physical, chemical, biological, and economic factors
as described in section 131.10(g) (40 CFR 131.3.)
"Waste load allocation" (WLA) is the portion of a receiving water's TMDL that is allocated to one
of its existing or future point sources of pollution (USEPA, 1991a.)
"Waters of the United States" refer to:
(1) all waters which are currently used, were used in the past, or may be susceptible to use
in interstate or foreign commerce, including all waters which are subject to the ebb and
flow of the tide;
(2) all interstate waters, including interstate wetlands;
(3) all other waters such as intrastate lakes, rivers, streams (including intermittent streams),
mudflats, sandflats, wetlands, sloughs, prairie potholes, wet meadows, playa lakes, or
natural ponds the use or degradation of which would affect or could affect interstate or
foreign commerce, including any such waters:
(i) which are or could be used by interstate or foreign travelers for recreational or
other purposes;
(ii) from which fish or shellfish are or could be taken and sold in interstate or foreign
commerce; or
(iii) which are or could be used for industrial purposes by industries in interstate
commerce.
(4) all impoundments of waters otherwise defined as waters of the United States under this
definition;
(5) tributaries of waters in paragraphs (1) through (4) of this definition;
(6) the territorial sea; and
(7) wetlands adjacent to waters (other than waters that are themselves wetlands) identified
in paragraphs (1) through (6) of this definition. "Wetlands" are defined as those areas
that are inundated or saturated by surface or groundwater at a frequency and duration
sufficient to support, and that under normal circumstances do support, a prevalence of
vegetation typically adapted for life in saturated soil conditions. Wetlands generally
include swamps, marshes, bogs, and similar areas.
Waste treatment systems, including treatment ponds or lagoons designed to meet the
requirements of the Act (other than cooling ponds as defined in 40 CFR 423.11(m) which also
meet the criteria for this definition) are not waters of the United States. (40 CFR 232.2.)
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Water Quality Standards Handbook - Second Edition
"Water-effect ratio" (WER) is an appropriate measure of the toxicity of a material obtained in
a site water divided by the same measure of the toxicity of the same material obtained
simultaneously in a laboratory dilution water (USEPA, 1994a.)
"Water quality assessment" is an evaluation of the condition of a water body using biological surveys,
chemical-specific analyses of pollutants hi water bodies, and toxicity tests (USEPA, 1991a.)
"Water quality limited segment" refers to any segment where it is known that water quality does not
meet applicable water quality standards and/or is not expected to meet applicable water quality
standards even after application of technology-based effluent limitations required by sections
301(b)(l)(A) and (B) and 306 of the Act (40 CFR 131.3.)
"Water quality standards" (WQS) are provisions of State or Federal law which consist of a designated
use or uses for the waters of the United States, water quality criteria for such waters based upon
such uses. Water quality standards are to protect public health or welfare, enhance the quality
of the water and serve the purposes of the Act (40 CFR 131.3.)
"Whole-effluent toxicity" is the total toxic effect of an effluent measured directly with a toxicity test
(USEPA, 1991a.)
GLOSS-8 (8/15/94)
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Introduction
Section 401 certification and FERC licenses
(USEPA, 199 Ih), clarifies the range of
water quality standards elements that States
need to apply when making CWA section
401 certification decisions. Section 401 of
the CWA is discussed in section 7.6.3.
Technical Support Document for Water Quality-
based Toxics Control, (USEPA, 199la),
provides technical guidance for assessing and
regulating the discharge of toxic substances
to the waters of the United States.
Policy on the Use of Biological Assessments and
Criteria in the Water Quality Program
(USEPA, 1991i), provides the basis for
EPA's policy that biological surveys shall be
fully integrated with toxicity and chemical-
specific assessment methods in State water
quality programs. Further discussion of this
policy is contained in section 3.3.
Numeric Water Quality Criteria for Wetlands
(Appendix E), evaluates EPA's numeric
aquatic life criteria to determine how they
can be applied to wetlands. Wetland aquatic
life criteria are discussed in section 3.5.6.
Endangered Species Act Joint Guidance
(Appendix F), establishes a procedure by
which EPA, the U.S. Fish and Wildlife
Service, and the National Marine Fisheries
Service will consult on the development of
water quality criteria and standards.
Office of Water Policy and Technical Guidance on
Interpretation and Implementation of Aquatic
Life Metals Criteria (USEPA, 1993f),
transmits Office of Water (OW) policy and
guidance on the interpretation and
implementation of aquatic life criteria for the
management of metals. Section 3.6
discusses EPA's policy on aquatic life metals
criteria.
Interpretation of Federal Antidegradation
Regulatory Requirement (USEPA, 1994a),
provides guidance on the interpretation of
the antidegradation policy in 40 CFR
131.12(a)(2) as it relates to nonpoint
sources. Antidegradation and nonpoint
sources are discussed in Section 4.6.
Interim Guidance on Determination and Use of
Water-Effect Ratios for Metals (Appendix
L), provides interim guidance concerning the
experimental determination of water-effect
ratios (WERs) for metals and supersedes all
guidance concerning water-effect ratios and
the Indicator Species Procedure in USEPA,
1983a and in USEPA, 1984f. It also
supersedes the guidance in these earlier
documents for the Recalculation Procedure
for performing site-specific aquatic life
criteria modifications. Site-specific aquatic
life criteria are discussed in Section 3.7.
The guidance contained in each of the above
documents is either incorporated into the text of
the appropriate section of this Handbook or
attached as appendices (see Table of Contents).
The reader is directed to the original guidance
documents for the explicit guidance on the topics
discussed. Copies of all original guidance
documents not attached as appendices may be
obtained from the source listed for each document
in the Reference section of this Handbook.
The Water Quality Standards Handbook - Second
Edition is reorganized from the 1983 Handbook.
An overview to Water Quality Standards and
Water Quality Management programs has been
added, and chapters 1 through 6 are organized to
parallel the provisions of the Water Quality
Standards Regulation. Chapter 7 briefly
introduces the role of water quality standards in
the water quality-based approach to pollution
control.
The Water Quality Standards Handbook - Second
Edition retains all the guidance in the 1983
Handbook unless such guidance was specifically
revised in subsequent years.
(8/15/94)
INT-7
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Water Quality Standards Handbook - Second Edition
OVERVIEW OF THE WATER QUALITY STANDARDS PROGRAM
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 preventing degradation of water
quality through antidegradation provisions. 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.
"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:
* include provisions for restoring and
maintaining chemical, physical, and
biological integrity of State waters;
wherever attainable, achieve a level of water
quality that provides for the protection and
propagation of fish, shellfish, and wildlife,
and recreation in and on the water
("fishable/swimmable"); and
* consider the use and value of State waters
for public water supplies, propagation of fish
and wildlife, recreation, agriculture and
industrial purposes, and navigation.
Section 303(c) of the Clean Water Act provides
the statutory basis for the water quality standards
program. The regulatory requirements governing
the program, the Water Quality Standards
Regulation, are published at 40 CFR 131. The
Regulation is divided into four subparts (A
through D), which are summarized below.
General Provisions (40 CFR 131 - Subpart A)
Subpart A includes the scope (section 131.1) and
purpose (section 131.2) of the Regulation,
definitions of terms used in the Regulation
(section 131.3), State (section 131.4) and EPA
(section 131.5) authority for water quality
standards, and the minimum requirements for a
State water quality standards submission (section
131.6).
On December 12, 1991, the EPA promulgated
amendments to Subpart A of the Water Quality
Standards Regulation in response to the CWA
section 518 requirements (see 56 F.R. 64875).
The Amendments:
establish a mechanism to resolve
unreasonable consequences that may result
from an Indian Tribe and a State adopting
differing water quality standards on common
bodies of water (section 131.7); and
add procedures by which an Indian Tribe can
qualify for the section 303 water quality
standards and section 401 certification
programs of the Clean Water Act (section
131.8).
The sections of Subpart A are discussed in chapter
1.
Establishment of Water Quality Standards -
(Subpart B)
Subpart B contains regulatory requirements that
must be included in State water quality standards:
designated uses (section 131.10), criteria that
protect the designated uses (section 131.11), and
an antidegradation policy that protects existing
uses and high water quality (section 131.12).
Subpart B also provides for State discretionary
policies, such as mixing zones and water quality
standards variances (section 131.13).
Each of these sections is summarized below and
discussed in detail in chapters 2 through 5
respectively.
Designation of Uses
The Water Quality Standards Regulation requires
that States specify appropriate water uses to be
INT-8
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CHAPTER 2
DESIGNATION OF USES
(40 CFR 131.10)
Table of Contents
2.1 Use Classification - 40 CFR 131.10(a) 2-1
2.1.1 Public Water Supplies 2-1
2.1.2 Protection and Propagation of Fish, Shellfish, and Wildlife 2-1
2.1.3 Recreation 2-2
2.1.4 Agriculture and Industry 2-3
2.1.5 Navigation 2-4
2.1.6 Other Uses 2-4
2.2 Consider Downstream Uses - 40 CFR 131.10(b) 2-4
2.3 Use Subcategories - 40 CFR 131.10(c) 2-5
2.4 Attainability of Uses - 40 CFR 131.10(d) 2-5
2.5 Public Hearing for Changing Uses - 40 CFR 131.10(e) 2-6
2.6 Seasonal Uses - 40 CFR 131.10(f) 2-6
2.7 Removal of Designated Uses - 40 CFR 131.10(g) and (h) 2-6
2.7.1 Step 1 - Is the Use Existing? 2-6
2.7.2 Step 2 - Is the Use Specified in Section 101(a)(2)? 2-8
2.7.3 Step 3 - Is the Use Attainable? 2-8
2.7.4 Step 4 - Is a Factor from 131.10(g) Met? 2-8
2.7.5 Step 5 - Provide Public Notice 2-8
2.8 Revising Uses to Reflect Actual Attainment - 40 CFR 131.10(1) 2-8
2.9 Use Attainability Analyses - 40 CFR 131.100) and (k) 2-9
2.9.1 Water Body Survey and Assessment - Purpose and Application 2-9
2.9.2 Physical Factors 2-10
2.9.3 Chemical Evaluations 2-12
2.9.4 Biological Evaluations 2-12
2.9.5 Approaches to Conducting the Physical, Chemical, and Biological
Evaluations 2-15
2.9.6 Estuarine Systems 2-18
2.9.7 Lake Systems 2-23
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Chapter 2 - Designation of Uses
federally permitted or licensed activities that may
result in a discharge to waters of the United
States. The decision to grant or to deny
certification, or to grant a conditional certification
is based on a State's determination regarding
whether the proposed activity will comply with
applicable water quality standards and other
provisions. Thus, States may deny certification
and prohibit EPA from issuing an NPDES permit
that would violate water quality standards.
Section 401 also allows a State to participate in
extraterritorial actions that will affect that State's
waters if a federally issued permit is involved.
In addition to the above sources for solutions,
when the problem arises between a State and an
Indian Tribe qualified for treatment as a State for
water quality standards, the dispute resolution
mechanism could be invoked (see section 1.7, of
this Handbook).
Use Subcategories - 40 CFR 131.10(c)
States are required to designate uses considering,
at a minimum, those uses listed in section 303(c)
of the Clean Water Act (i.e., public water
supplies, propagation of fish and wildlife,
recreation, agriculture and industrial purposes,
and navigation). However, flexibility inherent in
the State process for designating uses allows the
development of subcategories of uses within the
Act's general categories to refine and clarify
specific use classes. 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. Determination of non-attainment in
waters with broad use categories may be difficult
and open to alternative interpretations. If a
determination of non-attainment is in dispute,
regulatory actions will be difficult to accomplish
(USEPA, 1990a).
The State selects the level of specificity it desires
for identifying designated uses and subcategories
of uses (such as whether to treat recreation as a
single use or to define a subcategory for
secondary recreation). However, the State must
be at least as specific as the uses listed in sections
101(a) and 303(c) of the Clean Water Act.
Subcategories of aquatic life uses may be on the
basis of attainable habitat (e.g., coldwater versus
warmwater habitat); innate differences in
community structure and function (e.g., high
versus low species richness or productivity); or
fundamental differences in important community
components (e.g., warmwater fish communities
dominated by bass versus catfish). Special uses
may also be designated to protect particularly
unique, sensitive, or valuable aquatic species,
communities, or habitats.
Data collected from biosurveys as part of a
developing biocriteria program may assist States
in refining aquatic life use classes by revealing
consistent differences among aquatic communities
inhabiting different waters of the same designated
use. Measurable biological attributes could then
be used to divide one class into two or more
subcategories (USEPA, 1990a).
If States adopt subcategories that do not require
criteria sufficient to fully protect the goal uses in
section 101(a)(2) of the Act (see section 2.1,
above), a use attainability analysis pursuant to 40
CFR 131.10(j) must be conducted for waters to
which these subcategories are assigned. Before
adopting subcategories of uses, States must
provide notice and opportunity for a public
hearing because these actions are changes to the
standards.
Attainability of Uses - 40 CFR
When designating uses, States may wish to
designate only the uses that are attainable.
However, if the State does not designate the uses
specified in section 101(a)(2) of the Act, the State
must perform a use attainability analysis under
section 131.10(j) of the regulation. States are
encouraged to designate uses that the State
believes can be attained in the future.
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Water Quality Standards Handbook - Second Edition
"Attainable uses" are, at a minimum, the uses
(based on the State's system of water use
classification) that can be achieved 1) when
effluent limits under sections 301(b)(l)(A) and (B)
and section 306 of the Act are imposed on point
source dischargers and 2) when cost-effective and
reasonable best management practices are imposed
on nonpoint source dischargers.
Public Hearing for Changing Uses - 40
CFR
The Water Quality Standards Regulation requires
States to provide opportunity for public hearing
before adding or removing a use or establishing
subcategories of a use. As mentioned in section
2.2 above, the State should consider
extraterritorial effects of such changes.
Seasonal Uses - 40 CFR 131.10(f)
In some areas of the country, uses are practical
only for limited seasons. EPA recognizes
seasonal uses in the Water Quality Standards
Regulation. States may specify the seasonal uses
and criteria protective of that use as well as the
time frame for the "... season, so long as the
criteria do not prevent the attainment of any more
restrictive uses attainable in other seasons."
For example, in many northern areas, body
contact recreation is possible only a few months
out of the year. Several States have adopted
primary contact recreational uses, and the
associated microbiological criteria, for only those
months when primary contact recreation actually
occurs, and have relied on less stringent
secondary contact recreation criteria to protect for
incidental exposure in the "non-swimming"
season.
Seasonal uses that may require more stringent
criteria are uses that protect sensitive organisms
or life stages during a specific season such as the
early life stages of fish and/or fish migration
(e.g., EPA's Ambient Water Quality Criteria for
Dissolved Oxygen (see Appendix I) recommends
more stringent dissolved oxygen criteria for the
early life stages of both coldwater and warmwater
fish).
Removal of Designated Uses - 40 CFR
131.10(g) and (h)
Figure 2-1 shows how and when designated uses
may be removed.
2.7.1 Step 1 - Is the Use Existing?
Once a use has been designated for a particular
water body or segment, the water body or water
body segment cannot be reclassified for a
different use except under specific conditions. If
a designated use is an existing use (as defined in
40 CFR 131.3) for a particular water body, the
existing use cannot be removed unless a use
requiring more stringent criteria is added (see
2-6
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Chapter 3 - Water Quality Criteria
CHAPTERS
WATER QUALITY CRITERIA
(40 CFR 131.11)
Table of Contents
3.1 EPA Section 304(a) Guidance 3-1
3.1.1 State Use of EPA Criteria Documents 3-1
3.1.2 Criteria for Aquatic Life Protection 3-2
3.1.3 Criteria for Human Health Protection 3-3
3.2 Relationship of Section 304(a) Criteria to State Designated Uses 3-10
3.2.1 Recreation 3-10
3.2.2 Aquatic Life 3-11
3.2.3 Agricultural and Industrial Uses 3-11
3.2.4 Public Water Supply 3-11
3.3 State Criteria Requirements 3-12
3.4 Criteria for Toxicants 3-13
3.4.1 Priority Toxic Pollutant Criteria 3-13
3.4.2 Criteria for Nonconventional Pollutants 3-23
3.5 Forms of Criteria 3-23
3.5.1 Numeric Criteria 3-24
3.5.2 Narrative Criteria 3-24
3.5.3 Biological Criteria 3-26
3.5.4 Sediment Criteria 3-28
3.5.5 Wildlife Criteria 3-31
3.5.6 Numeric Criteria for Wetlands 3-33
3.6 Policy on Aquatic Life Criteria for Metals 3-34
3.6.1 Background 3-34
3.6.2 Expression of Aquatic Life Criteria 3-34
3.6.3 Total Maximum Daily Loads (TMDLs) and National Pollutant Discharge
Elimination System (NPDES) Permits 3-36
3.6.4 Guidance on Monitoring 3-37
3.7 Site-Specific Aquatic Life Criteria 3-38
3.7.1 History of Site-Specific Criteria Guidance 3-38
3.7.2 Preparing to Calculate Site-Specific Criteria 3-40
3.7.3 Definition of a Site . 3-41
3.7.4 The Recalculation Procedure 3-41
3.7.5 The Water-Effect Ratio (WER) Procedure 3-43
3.7.6 The Resident Species Procedure 3-44
Endnotes 3-45
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Chapter 3 - Water Quality Criteria
CHAPTER 3
WATER QUALITY CRITERIA
The term "water quality criteria" has two different
definitions under the Clean Water Act (CWA).
Under section 304(a), EPA publishes water
quality criteria that consist of scientific
information regarding concentrations of specific
chemicals or levels of parameters in water that
protect aquatic life and human health (see section
3.1 of this Handbook). The States may use these
contents as the basis for developing enforceable
water quality standards. Water quality criteria are
also elements of State water quality standards
adopted under section 303 (c) of the CWA (see
sections 3.2 through 3.6 of this Handbook).
States are required to adopt water quality criteria
that will protect the designated use(s) of a water
body. These criteria must be based on sound
scientific rationale and must contain sufficient
parameters or constituents to protect the
designated use.
EPA Section 304(a) Guidance
EPA and a predecessor agency have produced a
series of scientific water quality criteria guidance
documents. Early Federal efforts were the
"Green Book" (FWPCA, 1968) and the "Red
Book" (USEPA, 1976). EPA also sponsored a
contract effort that resulted in the "Blue Book"
(NAS/NAE, 1973). These early efforts were
premised on the use of literature reviews and the
collective scientific judgment of Agency and
advisory panels. However, when faced with the
need to develop criteria for human health as well
as aquatic life, the Agency determined that new
procedures were necessary. Continued reliance
solely on existing scientific literature was deemed
inadequate because essential information was not
available for many pollutants. EPA scientists
developed formal methodologies for establishing
scientifically defensible criteria. These were
subjected to review by the Agency's Science
Advisory Board of outside experts and the public.
This effort culminated on November 28, 1980,
when the Agency published criteria development
guidelines for aquatic life and for human health,
along with criteria for 64 toxic pollutants
(USEPA, 1980a,b). Since that initial publication,
the aquatic life methodology was amended
(Appendix H), and additional criteria were
proposed for public comment and finalized as
Agency criteria guidance. EPA summarized the
available criteria information in the "Gold Book"
(USEPA, 1986a), which is updated from time to
time. However, the individual criteria documents
(see Appendix I), as updated, are the official
guidance documents.
EPA's criteria documents provide a
comprehensive lexicological evaluation of each
chemical. For toxic pollutants, the documents
tabulate the relevant acute and chronic toxicity
information for aquatic life and derive the criteria
maximum concentrations (acute criteria) and
criteria continuous concentrations (chronic
criteria) that the Agency recommends to protect
aquatic life resources. The methodologies for
these processes are described in Appendices H
and J and outlined in sections 3.1.2 and 3.1.3 of
this Handbook.
3.1.1
State Use of EPA Criteria Documents
EPA's water quality criteria documents are
available to assist States in:
adopting water quality standards that include
appropriate numeric water quality criteria;
interpreting existing water quality standards
that include narrative "no toxics in toxic
amounts" criteria;
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Water Quality Standards Handbook - Second Edition
* making listing decisions under section 304(1)
of the CWA;
* writing water quality-based NPDES permits
and individual control strategies; and
providing certification under section 401 of
the CWA for any Federal permit or license
(e.g., EPA-issued NPDES permits, CWA
section 404 permits, or Federal Energy
Regulatory Commission licenses).
In these situations, States have primary authority
to determine the appropriate level to protect
human health or welfare (in accordance with
section 303(c)(2) of the CWA) for each water
body. However, under the Clean Water Act,
EPA must also review and approve State water
quality standards; section 304(1) listing decisions
and draft and final State-issued individual control
strategies; and in States where EPA writes
NPDES permits, EPA must develop appropriate
water quality-based permit limitations. The States
and EPA therefore have a strong interest in
assuring that the decisions are legally defensible,
are based on the best information available, and
are subject to full and meaningful public comment
and participation. It is very important that each
decision be supported by an adequate record.
Such a record is critical to meaningful comment,
EPA's review of the State's decision, and any
subsequent administrative or judicial review.
Any human health criterion for a toxicant is based
on at least three interrelated considerations:
* cancer potency or systemic toxicity,
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,
when different from EPA's recommendation, must
be clear and in the public record. If a State relies
on EPA's section 304(a) criteria document (or
other EPA documents), the State may reference
and rely on the data in these documents and need
not create duplicative or new material for
inclusion in their records. However, where site-
specific issues arise or the State decides to adopt
an approach to any one of these three factors that
differs from the approach in EPA's criteria
document, the State must explain its reasons in a
manner sufficient for a reviewer to determine that
the approach chosen is based on sound scientific
rationale (40 CFR 131.11(b)).
3.1.2 Criteria for Aquatic Life Protection
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 Appendix H 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 plants and
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 other water
quality characteristics, such as pH, temperature,
or hardness. Separate criteria are developed for
fresh and salt waters.
The Water Quality Standards Regulation allows
States to develop numerical criteria or modify
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Chapter 3 - Water Quality Criteria
EPA's recommended criteria to account for
site-specific or other scientifically defensible
factors. Guidance on modifying national criteria
is found in sections 3.6 and 3.7. 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 EPA guidelines (Appendix H).
Magnitude for Aquatic Life Criteria
Water quality criteria for aquatic life contain two
expressions of allowable magnitude: a criterion
maximum concentration (CMC) to protect against
acute (short-term) effects; and a criterion
continuous concentration (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, or reproduction.
Where appropriate, the calculated criteria may be
lowered to be protective ofcomercially or
recreationally important species.
Duration for Aquatic Life Criteria
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 an
upper limit on the average 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.
For chronic criteria, EPA recommends an
averaging period of 4 days. That is, the 4-day
average exposure should not exceed the CCC.
Frequency for Aquatic Life Criteria
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 an average frequency for
excursions of both acute and chronic criteria not
to exceed once in 3 years. 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 (Appendix H). EPA selected
the 3-year average frequency of criteria
exceedence with the intent of providing for
ecological recovery from a variety of severe
stresses. This return interval is roughly
equivalent to a 7Q10 design flow condition.
Because of the nature of the ecological recovery
studies available, the severity of criteria
excursions could not be rigorously related to the
resulting ecological impacts. Nevertheless, EPA
derives its criteria intending that a single marginal
criteria excursion (i.e., a slight excursion over a
1-hour period for acute or over a 4-day period for
chronic) would 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.
3.1.3 Criteria for Human Health Protection
This section reviews EPA's procedures used to
develop assessments of human health effects in
developing water quality criteria and reference
ambient concentrations. A more complete human
health effects discussion is included in the
Guidelines and Methodology Used in the
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Water Quality Standards Handbook - Second Edition
Preparation of Health Effects Assessment Chapters
of the Consent Decree Water Documents
(Appendix J). The procedures contained in this
document are used in the development and
updating of EPA water quality criteria and may be
used in updating State criteria and in developing
State criteria for those pollutants lacking EPA
human health criteria. The procedures may also
be applied as site-specific interpretations of
narrative standards and as a basis for permit limits
under 40 CFR 122.44 (d)(l)(vi).
Magnitude and Duration
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 establish that the
duration for human health criteria for carcinogens
should 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 owing to a
wide variety of endpoints: some developmental
(and thus age-specific and perhaps gender-
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.
Human Exposure Considerations
A complete human exposure evaluation for toxic
pollutants of concern for bioaccumulation would
encompass not only estimates of exposures due to
fish consumption but also exposure from
background concentrations and other exposure
routes, The more important of these include
recreational and occupational contact, dietary
intake from other than fish, intake from air
inhalation, and drinking water consumption. For
section 304(a) criteria development, EPA typically
considers only exposures to a pollutant that occur
through the ingestion of water and contaminated
fish and shellfish. This is the exposure default
assumption, although the human health guidelines
provide for considering other sources where data
are available (see 45 F.R. 79354). Thus the
criteria are based on an assessment of risks
related to the surface water exposure route only
(57 F.R. 60862-3).
The consumption of contaminated fish tissue is of
serious concern because the presence of even
extremely low ambient concentrations of
bioaccumulative 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 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 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.
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Chapter 3 - Water Quality Criteria
Fish Consumption Values
EPA's human health criteria have assumed a
human body weight of 70 kg and the consumption
of 6.5 g of fish and shellfish per day. Based on
data collected in 1973-74, 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. The mean lipid content
of fish and shellfish tissue consumed in this study
was estimated to be 3.0 percent (USEPA, 1980c).
Currently, four levels of fish and shellfish
consumption are provided in EPA guidance
(USEPA, 199 la):
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. This consumption level is based
on the average of both consumers and
nonconsumers of.
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. This average
consumption level also includes both
consumers and nonconsumers of.
165 g/day to represent consumption of fish
and shellfish from marine, estuarine, and
freshwaters by the 99.9th percentile of the
U.S. population consuming the most fish or
seafood.
180 g/day to represent a "reasonable worst
case" based on the assumption that some
individuals would consume fishand shellfish
at a rate equal to the combined consumption
of red meat, poultry, fish, and shellfish in
the United States.
EPA is currently 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
and shellfish in their diets (maximally
exposed individuals).
Many States use EPA's 6.5 g/day consumption
value. However, some States use the above-
mentioned 20 g/day value and, for saltwaters,
37 g/day. In general, EPA recommends that the
consumption values used in deriving criteria from
the formulas in this chapter reflect the most
current, relevant, and/or site-specific information
available.
Bioaccumulation Considerations
The ratio of the contaminant concentrations in fish
tissue versus that in 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). The bioaccumulation
factor (BAF) is defined similarly to the BCF
except that it includes contaminant uptake from
both water and food. Under laboratory
conditions, measurements of tissue/water
partitioning are generally considered to involve
uptake from water only. On the other hand, both
processes are likely to apply in the field since the
entire food chain is exposed.
The BAF/BCF ratio ranges from 1 to 100, with
the highest ratios applying to organisms in higher
trophic levels, and to chemicals with logarithm of
the octanol-water partitioning coefficient (log P)
close to 6.5.
Bioaccumulation considerations are integrated into
the criteria equations by using food chain
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Water Quality Standards Handbook - Second Edition
multipliers (FMs) in conjunction with the BCF.
The bioaccumulation and bioconcentration factors
for a chemical are related as follows:
BAF = FM x BCF
By incorporating the FM and BCF terms into the
criteria equations, bioaccumulation can be
addressed.
In Table 3-1, FM values derived from the work
of Thomann (1987, 1989) are listed according to
log P value and trophic level of the organism.
For chemicals with log P values greater than
about 7, there is additional 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. Trophic level 4
organisms are typically the most desirable species
for sport fishing and, therefore, FMs for trophic
level 4 should generally be used in the equations
for calculating criteria. In those very rare
situations where only lower trophic level
organisms are found, e.g., possibly oyster beds,
an FM for a lower trophic level might be
considered.
Measured BAFs (especially for those chemicals
with log P values above 6.5) reported in the
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1.0
1.0
1.0
1.1
I.I
1.1
1.1
1.2
1 3
& *ff
1.4
1.5
1.8
2.1
21-
.5
3.0
3.7
4.6
5.9
7.5
9.8
13
17
21
25
29
34
39
45
45*
' I v ^ 5 '
"4
"l.O
Y.b
1.0
1.0
1-0
1-0
1.1
1.1
1.1
l.i
1,2
Y3
s*s
1,4
1,6 '
2.6
2.6
3/%
.2
4.3
5.8
8.0
11
16
23
33
47
67
75
84
92
98
100
100*
experimentally measured BAFs in calculating the
criterion, the (FM x BCF) term is replaced by the
BAF in the equations in the following section.
Relatively few BAFs have been measured
accjuately and reported, and their application to
sites other than the specific ecosystem where they
were developed is problematic and subject to
uncertainty. The option is also 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.
* TheS6 recommended FMs are conservative
Hvfs for. log 3> values greater than 6.5 may range from
the values given to as low as 0.1 for contaminants with
very low bioavailability.
Table 3-1. Estimated Food Chain
Multipliers (FMs)
Updating Human Health Criteria Using
IRIS
EPA recommends that States use the most current
risk information in the process of updating human
3-6
(8/15/94)
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Chapter 3 - Water Quality Criteria
health criteria. The Integrated Risk Information
System (IRIS) (Barns and Dourson, 1988;
Appendix N) is an electronic data base of the
USEPA that provides chemical-specific risk
information on the relationship between chemical
exposure and estimated human health effects. Risk
assessment information contained in 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 Agency-wide
consensus. Risk assessment information and
values are updated on a monthly basis and are
approved for Agency-wide use. 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.
IRIS contains two types of quantitative risks
values: the oral Reference Dose (RfD) 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 qi*) represents the upper
bound cancer-causing potential resulting from
lifetime exposure to a substance. The RfD or the
oral carcinogenic potency estimate is used in the
derivation of EPA human health criteria.
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 IRIS contains the Agency's most
recent quantitative risk assessment values, current
IRIS values should be used by States in updating
or developing new human health criteria. This
means that the 1980 human health criteria should
be updated with the latest IRIS values. The
procedure for deriving an updated human health
water quality criterion would require inserting the
current Rfd or carcinogenic potency estimate on
IRIS into the equations in Exhibit 3.1 or 3.2, as
appropriate.
ERA'S
water quality
criterion
available
Evaluate other
sources of data,
e.g., FDA action
levels, MCLs, risk
assessment, fish
consumption
advisory levels
Figure 3-1. Procedure for determining an
updated criterion using IRIS
data.
Figure 3-1 shows the procedure for determining
an updated criterion using IRIS data. If a
chemical has both carcinogenic and non-
carcinogenic effects, i.e., both a cancer potency
estimate and a RfD, both criteria should be
calculated. The most stringent criterion applies.
Calculating Criteria for Non-carcinogens
The RfD is an estimate of the daily exposure to
the human population that is likely to be without
appreciable risk of causing 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 studies. (Note: "LOAEL"
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Water Quality Standards Handbook - Second Edition
and "NOAEL" refer to animal and human
toxicology and are therefore distinct from the
aquatic toxicity terms "no-observed-effect
concentration" (NOEC) and "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 a
LOAEL (USEPA, 1988a). An additional
uncertainty factor may be applied to account for
significant weakness or gaps in the database.
The RfD is a threshold below which systemic
toxic 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 IRIS are values that
represent EPA's consensus (and have uncertainty
spanning perhaps an order of magnitude). This
means an RfD of 1.0 mg/kg/day could range from
0.3 to 3.0 mg/kg/day.
For noncarcinogenic effects, an updated criterion
can be derived using the equation in Exhibit 3-1.
If the receiving water body 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.
Calculating Criteria for Carcinogens
Any human health criterion for a carcinogen is
based on at least three interrelated considerations:
cancer potency, exposure, and risk
characterization. When developing State criteria,
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
through the consumption of contaminated fish
and/or other aquatic life would require a criterion
of zero. The zero level is based upon the
assumption of non-threshold effects (i.e., no safe
level exists below which any increase in exposure
does not result in an increased risk of cancer) for
carcinogens. However, because a publicly
acceptable policy for safety does not require the
absence of all risk, a numerical estimate of
pollutant concentration (in jitg/1) which
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 population of
specified size due to an increase in exposure
(e.g., 10"6 risk level = 1 additional cancer in a
population of 1 million). 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 million people or 10"
6) 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 10'5 to 10"7 as examples.
The cancer potency estimate, or slope factor
(formerly known as the q{*), 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 which 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 criteria for
carcinogens, the oral cancer potency estimates or
slope factors from IRIS are used.
It is important to note that cancer potency factors
may overestimate or underestimate the actual risk.
Such potency estimates are subject to great
uncertainty because of two primary factors:
3-8
(8/15/94)
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Chapter 3 - Water Quality Criteria
C (rag/I) = (RID x WT> - 0>T -UN) x WT
WI + IFC x L x FM x BCF]
where:
WT
DT
IN
WI
PC
L
FM
BCF
updated water quality criterion (mg/1)
oral reference dose (nag 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 I/day)
daily fish consumption (kg fish/day)
ratio of Hpid fractioii offish tissue consumed to 5%
food chain multiplier (from Table 3-1)
bioconcentration factor (mg toxicant/kg fish divided by mg toxicant/L
water) for fish with 3 % lipid content
Exhibit 3-1. Equation for Deriving Human Health Criteria Based on Noncarcinogenic Effects
adequacy of the cancer data base (i.e.,
human vs. animal data); and
limited information regarding the mechanism
of cancer causation.
Risk levels of 10'5, 10'6, and 10'7 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
10~6 may yield a higher overall risk level if
multiple carcinogenic chemicals are present.
For carcinogenic effects, the criterion can be
determined by using the equation in Exhibit 3-2.
If the receiving water body is not designated as a
drinking water source, the factor WI can be
deleted.
Deriving Quantitative Risk Assessments in
the Absence of IRIS Values
The RfDs or cancer potency estimates comprise
the existing dose-response factors for developing
criteria. 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-response factors can be
developed based upon extrapolation of acute
and/or chronic animal data to concentrations of
exposure protective of fish consumption by
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3-9
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Water Quality Standards Handbook - Second Edition
C(mg/l) =
(RLxWT)
where:
C
RL
WT
qi*
wi
FC
L
FM
BCF
r, i
q,* [WI + FC x L x (FM x BCF)}
updated water quality criterion (mg/1)
risk level (10"*) where x is usually in the range of 4 to 6
' !
* A .
weight of an average human adult (70 kg)
^ ,"',
carcinogenic potency factor (kg day/mg)
average human adult water intake (2 I/day)
daily fish consumption (kg fish/day)
ratio of lipid fraction of fish tissue consumed to 3% assumed by EPA
!
food chain multiplier (from Table 3-1)
" E s
bioconcentration factor (mg toxicant/kg fish divided by mg toxicant/L
water) for fish with 3% lipid content
Exhibit 3-2. Equation for Deriving Human Health Criteria Based on Carcinogenic Effects
humans.
Relationship of Section 304(a) Criteria
to State Designated Uses
The section 304(a)(l) criteria published by EPA
from time to time can be used to support the
designated uses found in State standards. The
following sections briefly discuss the relationship
between certain criteria and individual use
classifications. Additional information on this
subject also can be found in the "Green Book"
(FWPCA, 1968); the "Blue Book" (NAS/NAE,
1973); the "Red Book" USEPA, 1976); the EPA
Water Quality Criteria Documents (see Appendix
I); the"Gold Book" (USEPA, 1986a); and future
EPA section 304(a)(l) water quality criteria
publications.
Where a water body is designated for more than
one use, criteria necessary to protect the most
sensitive use must be applied. The following four
sections discuss the major types of use categories.
3.2.1 Recreation
Recreational uses of water include activities such
as swimming, wading, boating, and fishing.
Often insufficient data exist on the human health
effects of physical and chemical pollutants,
including most toxics, to make a determination of
criteria for recreational uses. However, as a
general guideline, recreational waters that contain
chemicals in concentrations toxic or otherwise
harmful to man if ingested, or irritating to the
skin or mucous membranes of the human body
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Chapter 3 - Water Quality Criteria
upon brief immersion, should be avoided. The
section 304(a)(l) human health effects criteria
based on direct human drinking water intake and
fish consumption might provide useful guidance in
these circumstances. Also, section 304(a)(l)
criteria based on human health effects may be
used to support this designated use where fishing
is included in the State definition of "recreation."
In this latter situation, only the portion of the
criterion based on fish consumption should be
used. Section 304(a)(l) criteria to protect
recreational uses are also available for certain
physical, microbiological, and narrative "free
from" aesthetic criteria.
Research regarding bacteriological indicators has
resulted in EPA recommending that States use
Escherichia coli or enterococci as indicators of
recreational water quality (USEPA, 1986b) rather
than fecal coliform because of the better
correlation with gastroenteritis in swimmers.
The "Green Book" and "Blue Book" provide
additional information on protecting recreational
uses such as pH criteria to prevent eye irritation
and microbiological criteria based on aesthetic
considerations.
3.2.2 Aquatic Life
The section 304(a)(l) criteria for aquatic life
should be used directly to support this designated
use. If subcategories of this use are adopted
(e.g., to differentiate between cold water and
warm water fisheries), then appropriate criteria
should be set to reflect the varying needs of such
subcategories.
3.2.3 Agricultural and Industrial Uses
The "Green Book" (FWPCA, 1968) and "Blue
Book" (NAS/NAE, 1973) provide some
information on protecting agricultural and
industrial uses. Section 304(a)(l) criteria for
protecting these uses have not been specifically
developed for numerous parameters pertaining to
these uses, including most toxics.
Where criteria have not been specifically
developed for these uses, the criteria developed
for human health and aquatic life are usually
sufficiently stringent to protect these uses. States
may also establish criteria specifically designed to
protect these uses.
3.2.4 Public Water Supply
The drinking water exposure component of the
section 304(a)(l) criteria based on human health
effects can apply directly to this use classification.
The criteria also may be appropriately modified
depending upon whether the specific water supply
system falls within the auspices of the Safe
Drinking Water Act's (SDWA) regulatory control
and the type and level of treatment imposed upon
the supply before delivery to the consumer. The
SDWA controls the presence of contaminants in
finished ("at-the-tap") drinking water.
A brief description of relevant sections of the
SDWA is necessary to explain how the Act will
work in conjunction with section 304(a)(l) criteria
in protecting human health from the effects of
toxics due to consumption of water. Pursuant to
section 1412 of the SDWA, EPA has promulgated
"National Primary Drinking Water Standards" for
certain radionuclide, microbiological, organic, and
inorganic substances. These standards establish
maximum contaminant levels (MCLs), which
specify the maximum permissible level of a
contaminant in water that may be delivered to a
user of a public water system now defined as
serving a minimum of 25 people. MCLs are
established based on consideration of a range of
factors including not only the health effects of the
contaminants but also treatment capability,
monitoring availability, and costs. Under section
1401(l)(D)(i) of the SDWA, EPA is also allowed
to establish the minimum quality criteria for water
that may be taken into a public water supply
system.
Section 304(a)(l) criteria provide estimates of
pollutant concentrations protective of human
health, but do not consider treatment technology,
costs, and other feasibility factors. The section
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Water Quality Standards Handbook - Second Edition
304(a)(l) criteria also include fish
bioaccumulation and consumption factors in
addition to direct human drinking water intake.
These numbers were not developed to serve as
"at-the-tap" drinking water standards, and they
have no regulatory significance under the SDWA.
Drinking water standards are established based on
considerations, including technological and
economic feasibility, not relevant to section
304(a)(l) criteria. Section 304(a)(l) criteria are
more analogous to the maximum contaminant
level goals (MCLGs) (previously known as
RMCLs) under section 1412(b)(l)(B) of the
SDWA in which, based upon a report from the
National Academy of Sciences, the Administrator
should set target levels for contaminants in
drinking water at which "no known or anticipated
adverse effects occur and which allow an adequate
margin of safety." MCLGs do not take treatment,
cost, and other feasibility factors into
consideration. Section 304(a)(l) criteria are, in
concept, related to the health-based goals specified
in the MCLGs.
MCLs of the SDWA, where they exist, control
toxic chemicals in finished drinking water.
However, because of variations in treatment,
ambient water criteria may be used by the States
as a supplement to SDWA regulations. When
setting water quality criteria for public water
supplies, States have the option of applying
MCLs, section 304(a)(l) human health effects
criteria, modified section 304(a)(l) criteria, or
controls more stringent than these three to protect
against the effects of contaminants by ingestion
from drinking water.
For treated drinking water supplies serving 25
people or greater, States must control
contaminants down to levels at least as stringent
as MCLs (where they exist for the pollutants of
concern) in the finished drinking water.
However, States also have the options to control
toxics in the ambient water by choosing section
304(a)(l) criteria, adjusted section 304(a)(l)
criteria resulting from the reduction of the direct
drinking water exposure component in the criteria
calculation to the extent that the treatment process
reduces the level of pollutants, or a more stringent
contaminant level than the former three options.
State Criteria Requirements
Section 131.11(a)(l) of the Regulation requires
States to adopt water quality criteria to protect the
designated use(s). The State criteria must be
based on sound scientific rationale and must
contain sufficient parameters or constituents to
protect the designated use(s). For waters with
multiple use designations, the criteria must
support the most sensitive use.
In section 131.11, States are encouraged to adopt
both numeric and narrative criteria. Aquatic life
criteria should protect against both short-term
(acute) and long-term (chronic) effects. Numeric
criteria are particularly 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 may also 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 where there are no numeric criteria in the
State standards. Narrative criteria also can be
used where toxicity cannot be traced to a
particular pollutant.
Section 131.11(a)(2) requires States to develop
implementation procedures which explain how the
State will ensure that narrative toxics criteria are
met.
To more fully protect aquatic habitats, it is EPA's
policy that States fully integrate chemical-specific,
whole-effluent, and biological assessment
approaches in State water quality programs (see
Appendix R). Specifically, each of these three
methods can provide a valid assessment of non-
attainment of designated aquatic life uses but can
rarely demonstrate use attainment separately.
Therefore, EPA supports a policy of independent
application of these three water quality assessment
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Chapter 3 - Water Quality Criteria
approaches. Independent application means that
the validity of the results of any one of the
approaches does not depend on confirmation by
one or both of the other methods. This policy is
based on the unique attributes, limitations, and
program applications of each of the three
approaches. Each method alone can provide valid
and independently sufficient evidence of non-
attainment of water quality standards, irrespective
of any evidence, or lack thereof, derived from the
other two approaches. The failure of one method
to confirm impacts identified by another method
does not negate the results of the initial
assessment.
It is also EPA's policy that States should
designate aquatic life uses that appropriately
address biological integrity and adopt biological
criteria necessary to protect those uses (see
section 3.5.3 and Appendices C, K, and R).
Criteria for Toxicants
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
(Appendix A) and the Water Quality Act of 1987
which amended the Clean Water Act (Public Law
100-4) include more specific requirements for the
particular toxicants listed pursuant to CWA
section 307(a). For regulatory purposes, EPA has
translated the 65 compounds and families of
compounds listed pursuant to section 307(a) into
126 more specific substances, which EPA refers
to as "priority toxic pollutants." The 126 priority
toxic pollutants are listed in the WQS regulation
and in Appendix P of this Handbook. 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)(l) criteria guidance (see
section 3.4.1);
requirements applicable to priority toxic
pollutants that have not been the subject of
CWA section 304(a)(l) criteria guidance (see
section 3.4.1); and
requirements applicable to all other toxicants
(e.g., non-conventional pollutants like
ammonia and chlorine) (see section 3.4.2).
3.4.1 Priority Toxic Pollutant Criteria
The criteria requirements applicable to priority
toxic pollutants (i.e., the first two categories
above) are specified in CWA section 303 (c) (2) (B).
Section 303(c)(2)(B), as added by the Water
Quality Act of 1987, provides that:
Whenever a State reviews water quality
standards pursuant to paragraph (1) of
this subsection, or revises or adopts
new standards pursuant to this
paragraph, such State shall adopt
criteria for all toxic pollutants listed,
pursuant to section 307(a)(l) of this Act
for which criteria have been published
under section 304(a), the discharge or
presence of which in the affected
waters could reasonably be expected to
interfere with those designated uses
adopted by the State, as necessary to
support such designated uses. Such
criteria shall be specific numerical
criteria for such toxic pollutants.
Where such numerical criteria are not
available, whenever a State reviews
water quality standards pursuant to
paragraph (1), or revises or adopts new
standards pursuant to this paragraph,
such State shall adopt criteria based on
biological monitoring or assessment
methods consistent with information
published pursuant to section 304(a)(8).
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
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3-13
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Water Quality Standards Handbook - Second Edition
methods or previously adopted
numerical criteria.
EPA, in devising guidance for section
303(c)(2)(B), attempted to provide States with the
maximum flexibility that complied with the
express statutory language but also with the
overriding congressional objective: prompt
adoption and implementation of numeric toxics
criteria. EPA believed that flexibility was
important so that each State could comply with
section 303(c)(2)(B) and to the extent possible,
accommodate its existing water quality standards
regulatory approach.
General Requirements
To carry out the requirements of section
303(c)(2)(B), whenever a State revises its water
quality standards, it must review all available
information and data to first determine whether
the discharge or the presence of a toxic pollutant
5s interfering with or is likely to interfere with the
attainment of the designated uses of any water
body segment.
If the data indicate that it is reasonable to expect
the toxic pollutant to interfere with the use, or it
actually is interfering with the use, then the State
must adopt a numeric limit for the specific
pollutant. If a State is unsure whether a toxic
pollutant is interfering with, or is likely to
interfere with, the designated use and therefore is
unsure that control of the pollutant is necessary to
support the designated use, the State should
undertake to develop sufficient information upon
which to make such a determination. Presence of
facilities that manufacture or use the section
307(a) toxic pollutants or other information
indicating that such pollutants are discharged or
will be discharged strongly suggests that such
pollutants could be interfering with attaining
designated uses. If a State expects the pollutant
not to interfere with the designated use, then
section 303(1)(2)(B) does not require a numeric
standard for that pollutant.
Section 303(c)(2)(B) addresses only pollutants
listed as "toxic" pursuant to section 307(a) of the
Act, which are codified at 40 CFR 131.36(b).
The section 307(a) list contains 65 compounds and
families of compounds, which potentially include
thousands of specific compounds. The Agency
has interpreted that list to include 126 "priority"
toxic pollutants for regulatory purposes.
Reference in this guidance to toxic pollutants or
section 307(a) toxic pollutants refers to the 126
priority toxic pollutants unless otherwise noted.
Both the list of priority toxic pollutants and
recommended criteria levels are subject to change.
The national criteria recommendations published
by EPA under section 304(a) (see section 3.1,
above) of the Act include values for both acute
and chronic aquatic life protection; only chronic
criteria recommendations have been established to
//////////////////////////// ///,///»
'//////////////////,,,...
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Chapter 3 - Water Quality Criteria
protect human health. To comply with the
statute, a State needs to adopt aquatic life and
human health criteria where necessary to support
the appropriate designated uses. Criteria for the
protection of human health are needed for water
bodies designated for public water supply. When
fish ingestion is considered an important activity,
then the human health-related water quality
criteria recommendation developed under section
304(a) of the CWA should be used; that is, the
portion of the criteria recommendation based on
fish consumption. For those pollutants designated
as carcinogens, the recommendation for a human
health criterion is generally more stringent than
the aquatic life criterion for the same pollutant.
In contrast, the aquatic life criteria
recommendations for noncarcinogens are
generally more stringent than the human health
recommendations. When a State adopts a human
health criterion for a carcinogen, the State needs
to select a risk level. EPA has estimated risk
levels of 10'5, 10'6, and lO'7 in its criteria
documents under one set of exposure assumptions.
However, the State is not limited to choosing
among the risk levels published in the section
304(a) criteria documents, nor is the State limited
to the base case exposure assumptions; it must
choose the risk level for its conditions and explain
its rationale.
EPA generally regulates pollutants treated as
carcinogens in the range of 10"6 to 10"4 to protect
average exposed individuals and more highly
exposed populations. However, if a State selects
a criterion that represents an upper bound risk
level less protective than 1 in 100,000 (e.g., 10"5),
the State needs to have substantial support in the
record for this level. This support focuses on two
distinct issues. First, the record must include
documentation that the decision maker considered
the public interest of the State in selecting the risk
level, including documentation of public
participation in the decision making process as
required by the Water Quality Standards
Regulation at 40 CFR 131.20(b). Second, the
record must include an analysis showing that the
risk level selected, when combined with other risk
assessment variables, is a balanced and reasonable
estimate of actual risk posed, based on the best
and most representative information available.
The importance of the estimated actual risk
increases as the degree of conservatism in the
selected risk level diminishes. EPA carefully
evaluates all assumptions used by a State if the
State chose to alter any one of the standard EPA
assumption values (57 F.R. 60864, December 22,
1993).
EPA does not intend to propose changes to the
current requirements regarding the bases on which
a State can adopt numeric criteria (40 CFR
131.11(b)(l)). Under EPA's regulation, in
addition to basing numeric criteria on EPA's
section 304(a) criteria documents, States may also
base numeric criteria on site-specific
determinations or other scientifically defensible
methods.
EPA expects each State to comply with the new
statutory requirements in any section 303(c) water
quality standards review initiated after enactment
of the Water Quality Act of 1987. The structure
of section 303(c) is to require States to review
their water quality standards at least once each 3
year period. Section 303(c)(2)(B) instructs States
to include reviews for toxics criteria whenever
they initiate a triennial review. Therefore, even
if a State has complied with section 303(c)(2)(B),
the State must review its standards each triennium
to ensure that section 303(c)(2)(B) requirements
continue to be met, considering that EPA may
have published additional section 304(a) criteria
documents and that the State will have new
information on existing water quality and on
pollution sources.
It should be noted that nothing in the Act or in the
Water Quality Standards Regulation restricts the
right of a State to adopt numeric criteria for any
pollutant not listed pursuant to section 307(a)(l),
and that such criteria may be expressed as
concentration limits for an individual pollutant or
for a toxicity parameter itself as measured by
whole-effluent toxicity testing. However, neither
numeric toxic criteria nor whole-effluent toxicity
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should be used as a surrogate for, or to supersede
the other.
State Options
States may meet the requirements of CWA section
303(c)(2)(B) by choosing one of three
scientifically and technically sound options (or
some combination thereof):
(1) Adopt statewide numeric criteria in State
water quality standards for all section 307(a)
toxic pollutants for which EPA has
developed criteria guidance, regardless of
whether the pollutants are known to be
present;
(2) Adopt specific numeric criteria in State
water quality standards for section 307(a)
toxic pollutants as necessary to support
designated uses where such pollutants are
discharged or are present in the affected
waters and could reasonably be expected to
interfere with designated uses;
(3) Adopt a "translator procedure" to be applied
to a narrative water quality standard
provision that prohibits toxicity in receiving
waters. Such a procedure is to be used by
the State in calculating derived numeric
criteria, which shall be used for all purposes
under section 303(c) of the CWA. At a
minimum, such criteria need to be developed
for section 307(a) toxic pollutants, as
necessary to support designated uses, where
these pollutants are discharged or present in
the affected waters and could reasonably be
expected to interfere with designated uses.
Option 1 is consistent with State authority to
establish water quality standards. Option 2 most
directly reflects the CWA requirements and is the
option recommended by EPA. Option 3, while
meeting the requirements of the CWA, is best
suited to supplement numeric criteria from option
1 or 2. The three options are discussed in more
detail below.
OPTION 1
Adopt statewide numeric criteria in State water
quality standards for all section 307(a) toxic
pollutants for which EPA has developed criteria
guidance, regardless of whether the pollutants
are known to be present.
Pro:
simple, straightforward implementation
ensures that States will satisfy statute
makes maximum uses of EPA
recommendations
gets specific numbers into State water quality
standards fast, at first
Con:
some priority toxic pollutants may not be
discharged in State
may cause unnecessary monitoring by States
might result in "paper standards"
Option 1 is within a State's legal authority under
the CWA to adopt broad water quality standards.
This option is the most comprehensive approach
to satisfy the statutory requirements because it
would include all of the priority toxic pollutants
for which EPA has prepared section 304(a)
criteria guidance for either or both aquatic life
protection and human health protection. In
addition to a simple adoption of EPA's section
304(a) guidance as standards, a State must select
a risk level for those toxic pollutants which are
carcinogens (i.e., that cause or may cause cancer
in humans).
Many States find this option attractive because it
ensures comprehensive coverage of the priority
toxic pollutants with scientifically defensible
criteria without the need to conduct a resource-
intensive evaluation of the particular segments and
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pollutants requiring criteria. This option also
would not be more costly to dischargers than
other options because permit limits would be
based only on the regulation of the particular
toxic pollutants in their discharges and not on the
total listing in the water quality standards. Thus,
actual permit limits should be the same under any
of the options.
The State may also exercise its authority to use
one or more of the techniques for adjusting water
quality standards:
establish or revise designated stream uses
based on use attainability analyses (see
section 2.9);
develop site-specific criteria; or
allow short-term variances (see section 5.3)
when appropriate.
All three of these techniques may apply to
standards developed under any of the three
options discussed in this guidance. It is likely
that States electing to use option 1 will rely more
on variances because the other two options are
implemented with more site-specific data being
available. It should be noted, however, that
permits issued pursuant to such water quality
variances still must comply with any applicable
antidegradation and antibacksliding requirements.
OPTION 2
Adopt specific numeric criteria in State water
quality standards for section 307(a) toxic
pollutants as necessary to support designated
uses where such pollutants are discharged or
are present in the affected waters and could
reasonably be expected to interfere with
designated uses.
Pro:
directly reflects statutory requirement
standards based on demonstrated need to
control problem pollutants
State can use EPA's section 304(a) national
criteria recommendations or other
scientifically acceptable alternative, including
site-specific criteria
State can consider current or potential toxic
pollutant problems
State can go beyond section 307(a) toxics
list, as desired
Con:
may be difficult and time consuming to
determine if, and which, pollutants are
interfering with the designated use
adoption of standards can require lengthy
debates on correct criteria limit to be
included in standards
successful State toxic control programs based
on narrative criteria may be halted or slowed
as the State applies its limited resources to
developing numeric standards
difficult to update criteria once adopted as
part of standards
to be absolutely technically defensible, may
need site-specific criteria in many situations,
leading to a large workload for regulatory
agency
EPA recommends that a State use this option to
meet the statutory requirement. It directly reflects
all the Act's requirements and is flexible,
resulting in adoption of numeric water quality
standards as needed. To assure that the State is
capable of dealing with new problems as they
arise, EPA also recommends that States adopt a
translator procedure the same as, or similar to,
that described in option 3, but applicable to all
chemicals causing toxicity and not just priority
pollutants as is the case for option 3.
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Beginning in 1988, EPA provided States with
candidate lists of priority toxic pollutants and
water bodies in support of CWA section 304(1)
implementation. These lists were developed
because States were required to evaluate existing
and readily available water-related data to comply
with section 304(1), 40 CFR 130.10(d). A similar
"strawman" analysis of priority pollutants
potentially requiring adoption of numeric criteria
under section 303(c)(2)(B) was furnished to most
States in September or October of 1990 for their
use in ongoing and subsequent triennial reviews.
The primary differences between the "strawman"
analysis and the section 304(1) candidate lists were
that the "strawman" analysis (1) organized the
results by chemical rather than by water body, (2)
included data for certain STORET monitoring
stations that were not used in constructing the
candidate lists, (3) included data from the Toxics
Release Inventory database, and (4) did not
include a number of data sources used in
preparing the candidate lists (e.g., those, such as
fish kill information, that did not provide
chemical-specific information).
EPA intends for States, at a minimum, to use the
information gathered in support of section 304(1)
requirements as a starting point for identifying (1)
water segments that will need new and/or revised
water quality standards for section 307(a) toxic
pollutants, and (2) which priority toxic pollutants
require adoption of numeric criteria. In the
longer term, EPA expects similar determinations
to occur during each triennial review of water
quality standards as required by section 303(c).
In identifying the need for numeric criteria, EPA
is encouraging States to use information and data
such as:
presence or potential construction of
facilities that manufacture or use priority
toxic pollutants;
ambient water monitoring data, including
those for sediment and aquatic life (e.g., fish
tissue data);
NPDES permit applications and permittee
self-monitoring reports;
effluent guideline development documents,
many of which contain section 307(a)(l)
priority pollutant scans;
pesticide and herbicide application
information and other records of pesticide or
herbicide inventories;
public water supply source monitoring data
noting pollutants with Maximum
Contaminant Levels (MCLs); and
any other relevant information on toxic
pollutants collected by Federal, State,
interstate agencies, academic groups, or
scientific organizations.
States are also expected to take into account
newer information as it became available, such as
information in annual reports from the Toxic
Chemical Release Inventory requirements of the
Emergency Planning and Community Right-To-
Know Act of 1986 (Title III, Public Law 99-499).
Where the State's review indicates a reasonable
expectation of a problem from the discharge or
presence of toxic pollutants, the State should
identify the pollutant(s) and the relevant
segment(s). In making these determinations,
States should use their own EPA-approved criteria
or existing EPA water quality criteria for
purposes of segment identification. After the
review, the State may use other means to establish
the final criterion as it revises its standards.
As with option 1, a State using option 2 must
follow all its legal and administrative
requirements for adoption of water quality
standards. Since the resulting numeric criteria are
part of a State's water quality standards, they are
required to be submitted by the State to EPA for
review and either approval or disapproval.
EPA believes this option offers the State optimum
flexibility. For section 307(a) toxic pollutants
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adversely affecting designated uses, numeric
criteria are available for permitting purposes. For
other situations, the State has the option of
defining site-specific criteria.
OPTION 3
Adopt a procedure to be applied to the
narrative water quality standard provision that
prohibits toxicity in receiving waters. Such a
procedure would be used by a State hi
calculating derived numeric criteria to be used
for all purposes of water quality criteria under
section 303(c) of the CWA. At a minimum
such criteria need to be derived for section
307(a) toxic pollutants where the discharge or
presence of such pollutants hi the affected
waters could reasonably be expected to
interfere with designated uses, as necessary to
support such designated uses.
Pro:
allows a State flexibility to control priority
toxic pollutants
reduces time and cost required to adopt
specific numeric criteria as water quality
standards regulations
allows immediate use of latest scientific
information available at the time a State
needs to develop derived numeric criteria
revisions and additions to derived numeric
criteria can be made without need to revise
State law
State can deal more easily with a situation
where it did not establish water quality
standards for the section 307(a) toxic
pollutants during the most recent triennial
review
State can address problems from non-section
307(a) toxic pollutants
Con:
EPA is currently on notice that a derived
numeric criterion may invite legal challenge
once the necessary procedures are adopted to
enhance legal defensibility (e.g., appropriate
scientific methods and public participation
and review), actual savings in time and costs
may be less than expected
public participation in development of
derived numeric criteria may be limited
when such criteria are not addressed in a
hearing on water quality standards
EPA believes that adoption of a narrative standard
along with a translator mechanism as part of a
State's water quality standard satisfies the
substantive requirements of the statute. These
criteria are subject to all the State's legal and
administrative requirements for adoption of
standards plus review and either approval or
disapproval by EPA, and result in the
development of derived numeric criteria for
specific section 307(a) toxic pollutants. They are
also subject to an opportunity for public
participation. Nevertheless, EPA believes the
most appropriate use of option 3 is as a
supplement to either option 1 or 2. Thus, a State
would have formally adopted numeric criteria for
toxic pollutants that occur frequently; that have
general applicability statewide for inclusion in
NPDES permits, total maximum daily loads, and
waste load allocations; and that also would have
a sound and predictable method to develop
additional numeric criteria as needed. This
combination of options provides a complete
regulatory scheme.
Although the approach in option 3 is similar to
that currently allowed in the Water Quality
Standards Regulation (40 CFR 131.11(a)(2)), this
guidance discusses several administrative and
scientific requirements that EPA believes are
necessary to comply with section 303(c)(2)(B).
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(1) The Option 3 Procedure Must Be Used To
Calculate Derived Numeric Water Quality
Criteria
States must adopt a specific procedure to be
applied to a narrative water quality criterion. To
satisfy section 303(c)(2)(B), this procedure shall
be used by the State in calculating derived
numeric criteria, which shall be used for all
purposes under section 303(c) of the CWA. Such
criteria need to be developed for section 307(a)
toxic pollutants as necessary to support designated
uses, where these pollutants are discharged or are
present in the affected waters and could
reasonably be expected to interfere with the
designated uses.
To assure protection from short-term exposures,
the State procedure should ensure development of
derived numeric water quality criteria based on
valid acute aquatic toxicity tests that are lethal to
half the affected organisms (LC50) for the species
representative of or similar to those found in the
State. In addition, the State procedure should
ensure development of derived numeric water
quality criteria for protection from chronic
exposure by using an appropriate safety factor
applicable to this acute limit. If there are
saltwater components to the State's aquatic
resources, the State should establish appropriate
derived numeric criteria for saltwater in addition
to those for freshwater.
The State's documentation of the tests should
include a detailed discussion of its quality control
and quality assurance procedures. The State
should also include a description (or reference
existing technical agreements with EPA) of the
procedure it will use to calculate derived acute
and chronic numeric criteria from the test data,
and how these derived criteria will be used as the
basis for deriving appropriate TMDLs, WLAs,
and NPDES permit limits.
As discussed above, the procedure for calculating
derived numeric criteria needs to protect aquatic
life from both acute and chronic exposure to
specific chemicals. Chronic aquatic life criteria
are to be met at the edge of the mixing zone.
The acute criteria are to be met (1) at the end-of-
pipe if mixing is not rapid and complete and a
high rate diffuser is not present; or (2) after
mixing if mixing is rapid and complete or a high
rate diffuser is present. (See EPA's Technical
Support Document for Water Quality-based Toxics
Control, USEPA 199la.)
EPA has not established a national policy
specifying the point of application in the receiving
water to be used with human health criteria.
However, EPA has approved State standards that
apply human health criteria for fish consumption
at the mixing zone boundary and/or apply the
criteria for drinking water consumption, at a
minimum, at the point of use. EPA has also
proposed more stringent requirements for the
application of human health criteria for highly
bioaccumulative pollutants in the Water Quality
guidance for the Great Lakes System (50 F.R.
20931, 21035, April 16, 1993) including
elimination of mixing zones.
In addition, the State should also include an
indication of potential bioconcentration or
bioaccumulation by providing for:
« laboratory tests that measure the steady-state
bioconcentration rate achieved by a
susceptible organism; and/or
field data in which ambient concentrations
and tissue loads are measured to give an
appropriate factor.
In developing a procedure to be used in
calculating derived numeric criteria for the
protection of aquatic life, the State should
consider the potential impact that bioconcentration
has on aquatic and terrestrial food chains.
The State should also use the derived
bioconcentration factor and food chain multiplier
to calculate chronically protective numeric criteria
for humans that consume aquatic organisms. In
calculating this derived numeric criterion, the
State should indicate data requirements to be met
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Chapter 3 - Water Quality Criteria
when dealing with either threshold (toxic) or non-
threshold (carcinogenic) compounds. The State
should describe the species and the minimum
number of tests, which may generally be met by
a single mammalian chronic test if it is of good
quality and if the weight of evidence indicates that
the results are reasonable. The State should
provide the method to calculate a derived numeric
criterion from the appropriate test result.
Both the threshold and non-threshold criteria for
protecting human health should contain exposure
assumptions, and the State procedure should be
used to calculate derived numeric criteria that
address the consumption of water, consumption of
fish, and combined consumption of both water
and fish. The State should provide the
assumptions regarding the amount of fish and the
quantity of water consumed per person per day,
as well as the rationale used to select the
assumptions. It needs to include the number of
tests, the species necessary to establish a dose-
response relationship, and the procedure to be
used to calculate the derived numeric criteria.
For non-threshold contaminants, the State should
specify the model used to extrapolate to low dose
and the risk level. It should also address
incidental exposure from other water sources
(e.g., swimming). When calculating derived
numeric criteria for multiple exposure to
pollutants, the State should consider additive
effects, especially for carcinogenic substances,
and should factor in the contribution to the daily
intake of toxicants from other sources (e.g., food,
air) when data are available.
(2) The State Must Demonstrate That the
Procedure Results hi Derived Numeric
Criteria Are Protective
The State needs to demonstrate that its procedures
for developing criteria, including translator
methods, yield fully protective criteria for human
health and for aquatic life. EPA's review process
will proceed according to EPA's regulation of 40
CFR 131.11, which requires that criteria be based
on sound scientific rationale and be protective of
all designated uses. EPA will use the expertise
and experience it has gained in developing section
304(a) criteria for toxic pollutants by application
of its own translator method (USEPA, 1980b;
USEPA, 1985b).
Once EPA has approved the State's procedure,
the Agency's review of derived numeric criteria,
for example, for pollutants other than section
307(a) toxic pollutants resulting from the State's
procedure, will focus on the adequacy of the data
base rather than the calculation method. EPA
also encourages States to apply such a procedure
to calculate derived numeric criteria to be used as
the basis for deriving permit limitations for
nonconventional pollutants that also cause
toxicity.
(3) The State Must Provide Full Opportunity
for Public Participation in Adoption of the
Procedure
The Water Quality Standards Regulation requires
States to hold public hearings to review and revise
water quality standards in accordance with
provisions of State law and EPA's Public
Participation Regulation (40 CFR 25). Where a
State plans to adopt a procedure to be applied to
the narrative criterion, it must provide full
opportunity for public participation in the
development and adoption of the procedure as part
of the State's water quality standards. .
While it is not necessary for the State to adopt
each derived numeric criterion into its water
quality standards and submit it to EPA for review
and approval, EPA is very concerned that all
affected parties have adequate opportunity to
participate in the development of a derived
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numeric criterion even though it is not being
adopted directly as a water quality standard.
A State can satisfy the need to provide an
opportunity for public participation in the
development of derived numeric criteria in several
ways, including:
a specific hearing on the derived numeric
criterion;
the opportunity for a public hearing on an
NPDES permits as long as public notice is
given that a criterion for a toxic pollutant as
part of the permit issuance is being
contemplated; or
a hearing coincidental with any other hearing
as long as it is made clear that development
of a specific criterion is also being
undertaken.
For example, as States develop their lists and
individual control strategies (ICSs) under section
304(1), they may seek full public participation.
NPDES regulations also specify public
participation requirements related to State permit
issuance. Finally, States have public participation
requirements associated with Water Quality
Management Plan updates. States may take
advantage of any of these public participation
requirements to fulfill the requirement for public
review of any resulting derived numeric criteria.
In such cases, the State must give prior notice that
development of such criteria is under
consideration.
(4) The Procedure Must Be Formally Adopted
and Mandatory
Where a State elects to supplement its narrative
criterion with an accompanying implementing
procedure, it must formally adopt such a
procedure as a part of its water quality standards.
The procedure must be used by the State to
calculate derived numeric criteria that will be used
as the basis for all standards' purposes, including
the following: developing TMDLs, WLAs, and
limits in NPDES permits; determining whether
water use designations are being met; and
identifying potential nonpoint source pollution
problems.
(5) The Procedure Must Be Approved by EPA
as Part of the State's Water Quality
Standards Regulation
To be consistent with the requirements of the Act,
the State's procedure to be applied to the narrative
criterion must be submitted to EPA for review
and approval, and will become a part of the
State's water quality standards. (See 40 CFR
131.21 for further discussion.) This requirement
may be satisfied by a reference in the standards to
the procedure, which may be contained in another
document, which has legal effect and is binding
on the State, and all the requirements for public
review, State implementation, and EPA review
and approval are satisfied.
Criteria Based on Biological Monitoring
For priority toxic pollutants for which EPA has
not issued section 304(a)(l) 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:
whole-effluent toxicity control methods;
biological criteria methods; or
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 is also consistent with
the intent of Congress to expand State standards
programs beyond chemical-specific approaches.
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States should also consider developing protocols
to derive and adopt numeric criteria for priority
toxic pollutants (or other pollutants) where EPA
has not issued section 304(a) criteria guidance.
The State should consider available laboratory
toxicity test data that may be sufficient to support
derivation of chemical-specific criteria. Existing
data need not be as comprehensive as that
required to meet EPA's 1985 guidelines in order
for a State to use its own protocols to derive
criteria. EPA has described such protocols in the
proposed Water Quality Guidance for the Great
Lakes System (58 F.R. 20892, at 21016, April 16,
1993.) This is particularly important where other
components of a State's narrative criterion
implementation procedure (e.g., WET 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).
Biologically based monitoring or assessment
methods serve as the basis for control where no
specific numeric criteria exist or where calculation
or application of pollutant-by-pollutant criteria
appears infeasible. Also, these methods may
serve as a supplemental measurement of
attainment of water quality standards in addition
to numeric and narrative criteria. The
requirement for both numeric criteria and
biologically based methods demonstrates that
section 303(c)(2)(B) contemplates that States
develop a comprehensive toxics control program
regardless of the status of EPA's section 304(a)
criteria.
The whole-effluent toxicity (WET) testing
procedure is the principal biological monitoring
guidance developed by EPA to date. The purpose
of the WET procedure is to control point source
dischargers of toxic pollutants. The procedure is
particularly useful for monitoring and controlling
the toxicity of complex effluents that may not be
well controlled through chemical-specific numeric
criteria. As such, biologically based effluent
testing procedures are a necessary component of
a State's toxics control program under section
303(c)(2)(B) and a principal means for
implementing a State's narrative "free from
toxics" standard.
Guidance documents EPA considers to serve the
purpose of section 304(a)(8) include the Technical
Support Document for Water Quality-based Toxics
Control (USEPA, 1991a; Guidelines for Deriving
National Water Quality Criteria for the Protection
of Aquatic Organisms and Their Uses (Appendix
H); Guidelines and Methodology Used in the
Preparation of Health Effect Assessment Chapters
of the Consent Decree Water Criteria Documents
(Appendix J); Methods for Measuring Acute
Toxicity of Effluents to Freshwater and Marine
Organisms (USEPA, 1991d); Short-Term Methods
for Estimating the Chronic Toxicity of Effluents
and Receiving Waters to Freshwater Organisms
(USEPA, 199le); and Short-Term Methods for
Estimating the Chronic Toxicity of Effluents and
Receiving Waters to Marine and Estuarine
Organisms (USEPA, 1991f).
3.4.2 Criteria for Nonconventional Pollutants
Criteria requirements applicable to toxicants that
are not priority toxic pollutants (e.g., ammonia
and chlorine), 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 (discussed in sections 3.5.1 and 3.5.2,
below) may be applied to meet these
requirements.
Forms of Criteria
States are required to adopt water quality criteria,
based on sound scientific rationale, that contain
sufficient parameters or constituents to protect the
designated use. EPA believes that an effective
State water quality standards program should
include both parameter-specific (e.g., ambient
numeric criteria) and narrative approaches.
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3.5.1 Numeric Criteria
Numeric criteria are required where 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 be adequately 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 that are not priority toxic
pollutants 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).
3.5.2 Narrative Criteria
To supplement numeric criteria for toxicants, all
States have also adopted narrative criteria for
toxicants. Such narrative criteria are statements
that describe the desired water quality goal, such
as the following:
All waters, including those within
mixing zones, shall be free from
substances attributable to wastewater
discharges or other pollutant sources
that:
(1) Settle to form objectional
deposits;
(2) Float as debris, scum, oil, or
other matter forming nuisances;
(3) Produce objectionable color, odor,
taste, or turbidity;
(4) Cause injury to, or are toxic to,
or produce adverse physiological
responses in humans, animals, or
plants; or
(5) Produce undesirable or nuisance
aquatic life (54 F.R. 28627, July
6, 1989).
EPA considers that the narrative criteria apply to
all designated uses at all flows and are necessary
to meet the statutory requirements of section
303(c)(2)(A) of the CWA.
Narrative toxic criteria (No. 4, above) 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 toxic criteria are cited
as a basis for establishing whole-effluent toxicity
controls in EPA permitting regulations at 40 CFR
122.44(d)(l)(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 (Exhibit 3-3) should
address all mechanisms to be 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 chemical-specific numeric criteria (e.g.,
procedures for establishing chemical-specific
permit limits under the NPDES permitting
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State implementation procedures for narrative toxics criteria should describe the following:
Specific, scientifically defensible methods by which the State will implement its narrative
toxics standard for all toxicants, including:
- methods for chemical-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 toxicity criteria and/or
controls; and
- methods for developing and implementing biological criteria.
* How these methods will be integrated in the State's toxics control program (i.e., how the
State will proceed when the specified methods produce conflicting or inconsistent results).
Application criteria and information needed to apply numerical criteria, for example:
- methods the State will use to identify those pollutants to be regulated in a specific
discharge;
- an incremental cancer risk level for carcinogens;
- methods for identifying compliance thresholds in permits where calculated limits are
below detection;
- methods for selecting appropriate hardness, pH, 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; and
- other methods and information needed to apply standards on a case-by-case basis.
Exhibit 3-3. Components of a State Implementation Procedure for Narrative Toxics Criteria
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program). Implementation procedures must also
address State programs to control whole-effluent
toxicity (WET) 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 section 3.5.3 and
Appendices C, K, and R).
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 (WET) controls is clear. EPA regulations
at 40 CFR 122.44(d)(l)(v) require NPDES
permits to contain WET 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 is also required by
EPA regulations at 40 CFR 122.44(d)(l). 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 131,
EPA has regulations at 40 CFR 122.44 that cover
the National Surface Water Toxics Control
Program. These regulations are intrinsically
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)(l)(vi) provides the permitting authority
with several options for establishing effluent limits
when a State does not have a chemical-specific
numeric criterion for a pollutant present in an
effluent at a concentration that causes or
contributes to a violation of the State's narrative
criteria.
3.5.3 Biological Criteria
The Clean Water Act of 1972 directs EPA to
develop programs that will evaluate, restore, and
maintain the chemical, physical, and biological
integrity of the Nation's waters. In response to
this directive, States and EPA have implemented
chemically based water quality programs that
address significant water pollution problems.
However, over the past 20 years, it has become
apparent that these programs alone cannot identify
and address all surface water pollution problems.
To help create a more comprehensive program,
EPA is setting a priority for the development of
biological criteria as part of State water quality
standards. This effort will help States and EPA
(1) achieve the biological integrity objective of the
CWA set forth in section 101, and (2) comply
with the statutory requirements under sections 303
and 304 of the Act (see Appendices C and K).
Regulatory Bases for Biocriteria
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 Clean
Water Act. Section 101 (a) of the CWA gives the
general goal of 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(a) 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(c), which requires EPA to
develop criteria based on biological assessment
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methods when numerical criteria are not
established.
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:
the effects of pollutants on aquatic
community components (". . . plankton,
fish, shellfish, wildlife, plant life . . .") and
community attributes (". . . biological
community diversity, productivity, and
stability . . .") in any body of water; and
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 . . . ."
Once biocriteria are formally adopted into State
standards, biocriteria and aquatic life use
designations serve as direct, legal endpoints for
determining aquatic life use attainment/non-
attainment. CWA section 303(c)(2)(B) provides
that when numeric criteria are not available,
States shall adopt criteria for toxics based on
biological monitoring or assessment methods;
biocriteria can be used to meet this requirement.
Development and Implementation of
Biocriteria
Biocriteria are numerical values or narrative
expressions that describe the expected reference
biological integrity of aquatic communities
inhabiting waters of a designated aquatic life use.
In the most desirable scenario, these would be
waters that are either in pristine condition or
minimally impaired. However, in some areas
these conditions no longer exist and may not be
attainable. In these situations, the reference
biological communities represent the best
attainable conditions. In either case, the reference
conditions then become the basis for developing
biocriteria for major surface water types (streams,
rivers, lakes, wetlands, estuaries, or marine
waters).
Biological criteria support designated aquatic life
use classifications for application in State
standards (see chapter 2). 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.,
coldwater versus warmwater stream systems as
represented by distinctive trout or bass fish
communities, respectively). Special uses may
also 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. Biological criteria can be used
to verify improvement in water quality in
response to regulatory and other improvement
efforts and to detect new or continuing
degradation of waters. Biological criteria also
provide a framework for developing improved
best management practices and management
measures for nonpoint source impacts. Numeric
biological criteria can provide effective
monitoring criteria for more definitive evaluation
of the health of an aquatic ecosystem.
The assessment of the biological integrity of a
water body should include measures of the
structure and function of the aquatic community
within a specified habitat. Expert knowledge of
the system is required for the selection of
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appropriate biological components and
measurement indices. The development and
implementation of biological criteria requires:
* selection of surface waters to use in
developing reference conditions for each
designated use;
measurement of the structure and function of
aquatic communities in reference surface
waters to establish biological criteria;
* measurement of the physical habitat and
other environmental characteristics of the
water resource; and
* establishment of a protocol to compare the
biological criteria to biota in comparable test
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 and the supporting
biological monitoring and assessment techniques.
3.5.4 Sediment 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 sections 104 and
304(a)(l) and (2) of the CWA as follows:
section 104(n)(l) authorizes the
Administrator to establish national programs
that study the effects of pollution, including
sedimentation, in estuaries on aquatic life;
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;
section 304(a)(2) directs the Administrator to
develop and publish information on, among
other issues, "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
could also be incorporated into those regulations.
EPA's current sediment criteria development
effort, as described below, focuses on criteria for
the protection of aquatic life. EPA anticipates
potential future expansion of this effort to include
sediment criteria for the protection of human
health.
Chemical Approach to Sediment Criteria
Development
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 predictions to be made regarding the
level of contaminant concentration that benthic
and other organisms may be exposed to. Chronic
water quality criteria, or possibly other
toxicological endpoints, can then be used to
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predict potential biological effects. In addition to
the development of sediment criteria, EPA is also
working to develop a standardized sediment
toxicity test that could be used with or
independently of sediment criteria to assess
chronic effects in fresh and marine waters.
Equilibrium Partitioning (EqP) Sediment
Quality Criteria (SQC) are the U.S.
Environmental Protection Agency'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.
Application of Sediment 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. To make the most informed
decisions, EPA believes that a comprehensive
approach using SQC and biological test methods
is preferred.
Sediment criteria will be particularly valuable in
site-monitoring applications where sediment
contaminant concentrations are gradually
approaching a criterion over time or as a
preventive tool to ensure that point and nonpoint
sources of contamination are controlled and that
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 Applications
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
sediment quality criteria would be in a tiered
approach. In such an application, when
contaminants in sediments exceed the sediment
quality criteria 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
criterion has been violated. (In locations where
contamination significantly exceeds a criterion, no
additional testing would be required. Where
sediment contaminant levels are close to a
criterion, additional testing might be necessary.)
Contaminants in a sediment at concentrations less
than the sediment criterion would not be of
concern. However, in some cases the sediment
could not be considered safe because it might
contain other contaminants above safe levels for
which no sediment criteria exist. In addition, the
synergistic, antagonistic, or additive effects of
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several contaminants in the sediments may be of
concern.
Additional testing in other tiers of an evaluation
approach, such astoxicity tests, 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
include the following:
* Establish permit limits for point sources to
ensure that uncontaminated sediments remain
uncontaminated or sediments already
contaminated have an opportunity to cleanse
themselves. Of course, this would occur
only after criteria and the means to tie point
sources to sediment contamination are
developed.
* Establish target levels for nonpoint sources
of sediment contamination.
* For remediation activities, SQC would be
valuable in identifying:
- need for remediation,
- spatial extent of remediation area,
- benefits derived from remediation
activities,
- responsible parties,
- impacts of depositing contaminated
sediments in water environments, and
- success of remediation activities.
In tiered testing sediment evaluation processes,
sediment criteria and biological testing procedures
work very well together.
Sediment Criteria Status
Science Advisory Board Review
The Science Advisory Board has completed a
second review of the EqP approach to deriving
sediment quality criteria for non-ionic
contaminants. The November 1992 report
(USEPA, 1992c) endorses the EqP approach to
deriving criteria as ". . . sufficiently valid to be
used in the regulatory process if the uncertainty
associated with the method is considered,
described, and incorporated," and that "EPA
should ... establish criteria on the basis of
present knowledge within the bounds of
uncertainty. ..."
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."
Sediment Criteria Documents and
Application Guidance
EPA efforts at producing sediment criteria
documents are being directed first toward
phenanthrene, fluoranthene, dieldrin,
acenaphthene, and endrin. Efforts are also being
directed towards producing a guidance document
on the derivation and interpretation of sediment
quality criteria. The criteria documents were
announced in the Federal Register in January
1994; the public comment period ended June
1994. Final documents and implementation
guidance should be available in early 1996.
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Methodology for Developing Sediment
Criteria for Metal Contaminants
EPA is proceeding to develop a methodology for
calculating sediment criteria for benthic toxicity to
metal contaminants, with key work focused on
identifying and understanding the role of acid
volatile sulfides (AVS), and other binding factors,
in controlling the bioavailability of metal
contaminants. A variety of field and laboratory
verification studies are under way to add
additional support to the methodology. Standard
AVS sampling and analytical procedures are
under development. Presentation of the metals
methodology to the SAB for review is anticipated
for Fall 1994.
Biological Approach to Sediment Criteria
Development
Under the Contaminated Sediment Management
Strategy, EPA programs have committed to using
consistent biological methods to determine if
sediments are contaminated. In the water
program, these biological methods will be used as
a complement to the sediment-chemical criteria
under development. The biological methods
consist of both toxicity and bioaccumulation tests.
Freshwater and saltwater benthic species, selected
to represent the sensitive range of species'
responses to toxicity, are used in toxicity tests to
measure sediment toxicity. Insensitive freshwater
and saltwater benthic species that form the base of
the food chain are used in toxicity tests to
measure the bioaccumulation potential of
sediment. In FY 1994, acute toxicity tests and
bioaccumulation tests selected by all the Agency
programs should be standardized and available for
use. Training for States and EPA Regions on
these methods is expected to begin in FY1995.
In the next few years, research will be conducted
to develop standardized chronic toxicity tests for
sediment as well as toxicity identification
evaluation (TIE) methods. The TIE approach will
be used to identify the specific chemicals in a
sediment causing acute or chronic toxicity in the
test organisms. Under the Contaminated
Sediment Management Strategy, EPA's programs
have also agreed to incorporate these chronic
toxicity and TIE methods into their sediment
testing when they are available.
3.5.5 Wildlife Criteria
Terrestrial and avian species are useful as
sentinels for the health of the ecosystem as a
whole. In many cases, damage to wildlife
indicates that the ecosystem itself is damaged.
Many wildlife species that are heavily dependent
on the aquatic food web reflect the health of
aquatic systems. In the case of toxic chemicals,
terminal predators such as otter, mink, gulls,
terns, eagles, ospreys, and turtles are useful as
integrative indicators of the status or health of the
ecosystem.
Statutory and Regulatory Authority
Section 101(a)(2) of the CWA sets, as an interim
goal of,
. . . wherever attainable . . . water
quality which provides for the
protection and propagation of fish,
shellfish, and wildlife . . . (emphasis
added).
Section 304(a)(l) of the Act also requires EPA to:
. . . develop and publish . . . criteria for
water quality accurately reflecting . . . the
kind and extent of all identifiable effects on
health and welfare including . . . wildlife.
The Water Quality Standards Regulation reflect
the statutory goals and requirements by requiring
States to adopt, where attainable, the CWA
section 101(a)(2) goal uses of protection and
propagation of fish, shellfish, and wildlife (40
CFR 131.10), and to adopt water quality criteria
sufficient to protect the designated use (40 CFR
131.11).
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Wildlife Protection in Current Aquatic
Criteria
Current water quality criteria methodology is
designed to protect fish, benthic invertebrates, and
zooplankton; however, there is a provision in the
current aquatic life criteria guidelines (Appendix
H) that is intended to protect wildlife that
consume aquatic organisms from the
bioaccumulative potential of a compound. The
final residue value can be based on either the
FDA Action Level or a wildlife feeding study.
However, if maximum permissible tissue
concentration is not available from a wildlife
feeding study, a final residue value cannot be
derived and the criteria quantification procedure
continues without further consideration of wildlife
impacts. Historically, wildlife have been
considered only after detrimental effects on
wildlife populations have been observed in the
environment (this occurred with relationship to
DDT, selenium, and PCBs).
Wildlife Criteria Development
EPA's national wildlife criteria effort began
following release of a 1987 Government
Accounting Office study entitled Wildlife
Management - National Refuge Contamination Is
Difficult To Confirm and Clean Up (GAO, 1987).
After waterfowl deformities observed at Kesterson
Wildlife Refuge were linked to selenium
contamination in the water, Congress requested
this study and recommended that "the
Administrator of EPA, in close coordination with
the Secretary of the Interior, develop water
quality criteria for protecting wildlife and their
refuge habitat."
In November of 1988, EPA's Environmental
Research Laboratory in Corvallis sponsored a
workshop entitled Water Quality Criteria To
Protect Wildlife Resources, (USEPA, 1989g)
which was co-chaired by EPA and the Fish and
Wildlife Service (FWS). The workshop brought
together 26 professionals from a variety of
institutions, including EPA, FWS, State
governments, academia, and consultants who had
expertise in wildlife toxicity, aquatic toxicity,
ecology, environmental risk assessment, and
conservation. Efforts at he workshop focused on
evaluating the need for, and developing a strategy
for production of wildlife criteria. Two
recommendations came out of that workshop:
(1) The process by which ambient
water quality criteria are
established should be modified to
consider effects on wildlife; and
(2) chemicals should be prioritized
based on their potential to
adversely impact wildlife species.
Based on the workshop recommendations,
screening level wildlife criteria (SLWC) were
calculated for priority pollutants and chemicals of
concern submitted by the FWS to gauge the extent
of the problem by:
(1) evaluating whether existing water
quality criteria for aquatic life are
protective of wildlife, and
(2) prioritizing chemicals for their potential
to adversely impact wildlife species.
There were 82 chemicals for which EPA had the
necessary toxicity information as well as ambient
water quality criteria, advisories, or lowest-
observed-adverse-effect levels (LOAELs) to
compare with the SLWC values. As would be
expected, the majority of chemicals had SLWC
larger than existing water quality criteria,
advisories, or LOAELs for aquatic life.
However, the screen identified classes of
compounds for which current ambient water
quality criteria may not be adequately protective
of wildlife: chlorinated alkanes, benzenes,
phenols, metals, DDT, and dioxins. Many of
these compounds are produced in very large
amounts and have a variety of uses (e.g.,
solvents, flame retardants, organic syntheses of
fungicides and herbicides, and manufacture of
plastics and textiles. The manufacture and use of
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these materials produce waste byproduct). Also,
5 of the 21 are among the top 25 pollutants
identified at Superfund sites in 1985 (3 metals, 2
organics).
Following this initial effort, EPA held a national
meeting in April 19921 to constructively discuss
and evaluate proposed methodologies for deriving
wildlife criteria to build consensus among the
scientific community as to the most defensible
scientifically approach(es) to be pursued by EPA
in developing useful and effective wildlife criteria.
The conclusions of this national meeting were as
follows:
wildlife criteria should have a tissue-residue
component when appropriate;
peer-review of wildlife criteria and data sets
should be used in their derivation;
wildlife criteria should incorporate methods
to establish site-specific wildlife criteria;
additional amphibian and reptile toxicity data
are needed;
further development of inter-species
toxicological sensitivity factors are needed;
and
criteria methods should measure biomarkers
in conjunction with other studies.
On April 16, 1993, EPA proposed wildlife
criteria in the Water Quality Guidance for the
Great Lakes System (58 F.R. 20802). The
proposed wildlife criteria are based on the current
EPA noncancer human health criteria approach.
In this proposal, in addition to requesting
comments on the proposed Great Lakes criteria
and methods, EPA also requested comments on
possible modifications of the proposed Great
Lakes approach for consideration in the
development of national wildlife criteria.
3.5.6 Numeric Criteria for Wetlands
Extension of the EPA national 304(a) numeric
aquatic life criteria to wetlands is recommended
as part of a program to develop standards and
criteria for wetlands. Appendices D and E
provide an overview of the need for standards and
criteria for wetlands. The 304(a) numeric aquatic
life criteria are designed to be protective of
aquatic life for surface waters and are generally
applicable to most wetland types. Appendix E
provides a possible approach, based on the site-
specific guidelines, for detecting wetland types
that might not be protected by direct application
of national 304(a) criteria. The evaluation can be
simple and inexpensive for those wetland types
for which sufficient water chemistry and species
assemblage data are available, but will be less
useful for wetland types for which these data are
not readily available. In Appendix E, the site-
specific approach is described and recommended
for wetlands for which modification of the 304(a)
numeric criteria are considered necessary. The
results of this type of evaluation, combined with
information on local or regional environmental
threats, can be used to prioritize wetland types
(and individual criteria) for further site-specific
evaluations and/or additional data collection.
Close coordination among regulatory agencies,
wetland scientists, and criteria experts will be
required.
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Policy on Aquatic Life Criteria for
Metals
It is the policy of the Office of Water that the use
of dissolved metal to set and measure compliance
with water quality standards is the recommended
approach, because dissolved metal more closely
approximates the bioavailable fraction of metal in
the water column than does total recoverable
metal. This conclusion regarding metals
bioavailability is supported by a majority of the
scientific community within and outside EPA.
One reason is that a primary mechanism for water
column toxicity is adsorption at the gill surface
which requires metals to be in the dissolved form.
Until the scientific uncertainties are better
resolved, a range of different risk management
decisions can be justified by a State. EPA
recommends that State water quality standards be
based on dissolved metala conversion factor
must be used in order to express the EPA criteria
articulated as total recoverable as dissolved. (See
the paragraph below for technical details on
developing dissolved criteria.) EPA will also
approve a State risk management decision to adopt
standards based on total recoverable metal, if
those standards are otherwise approvable as a
matter of law. (Office of Water Policy and
Technical Guidance on Interpretation and
Implementation of Aquatic Life Metals Criteria
USEPA, 1993f)
3.6.1 Background
The implementation of metals criteria is complex
due to the site-specific nature of metals toxicity.
This issue covers a number of areas including the
expression of aquatic life criteria; total maximum
daily loads (TMDLs), permits, effluent
monitoring, and compliance; and ambient
monitoring. The following Sections, based on the
policy memorandum referenced above, provide
additional guidance in each of these areas.
Included in this Handbook as Appendix J are
three guidance documents issued along with the
Office of Water policy memorandum with
additional technical details. They are: Guidance
Document on Expression of Aquatic Life Criteria
as Dissolved Criteria (Attachment #2), Guidance
Document on Dynamic Modeling and Translators
(Attachment #3), and Guidance Document on
Monitoring (Attachment #4). These will be
supplemented as additional information becomes
available.
Since metals toxicity is significantly affected by
site-specific factors, it presents a number of
programmatic challenges. Factors that must be
considered in the management of metals in the
aquatic environment include: toxicity specific to
effluent chemistry; toxicity specific to ambient
water chemistry; different patterns of toxicity for
different metals; evolution of the state of the
science of metals toxicity, fate, and transport;
resource limitations for monitoring, analysis,
implementation, and research functions; concerns
regarding some of the analytical data currently on
record due to possible sampling and analytical
contamination; and lack of standardized protocols
for clean and ultraclean metals analysis. The
States have the key role in the risk management
process of balancing these factors in the
management of water programs. The site-specific
nature of this issue could be perceived as
requiring a permit-by-permit approach to
implementation. However, EPA believes that this
guidance can be effectively implemented on a
broader level, across any waters with roughly the
same physical and chemical characteristics, and
recommends that States work with the EPA with
that perspective in mind.
3.6.2 Expression of Aquatic Life Criteria
Dissolved vs. Total Recoverable Metal
A major issue is whether, and how, to use
dissolved metal concentrations ("dissolved metal")
or total recoverable metal concentrations ("total
recoverable metal") in setting State water quality
standards. In the past, States have used both
approaches when applying the same EPA Section
304(a) criteria guidance. Some older criteria
documents may have facilitated these different
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approaches to interpretation of the criteria because
the documents were somewhat equivocal with
regards to analytical methods. The May 1992
interim guidance continued the policy that either
approach was acceptable.
The position that the dissolved metals approach is
more accurate has been questioned because it
neglects the possible toxicity of paniculate metal.
It is true that some studies have indicated that
particulate metals appear to contribute to the
toxicity of metals, perhaps because of factors such
as desorption of metals at the gill surface, but
these same studies indicate the toxicity of
particulate metal is substantially less than that of
dissolved metal.
Furthermore, any error incurred from
excluding the contribution of particulate metal will
generally be compensated by other factors which
make criteria conservative. For example, metals
in toxicity tests are added as simple salts to
relatively clean water. Due to the likely presence
of a significant concentration of metals binding
agents in many discharges and ambient waters,
metals in -toxicity tests would generally be
expected to be more bioavailable than metals in
discharges or in ambient waters.
If total recoverable metal is used for the
purpose of specifying water quality standards, the
lower bioavailability of particulate metal and
lower bioavailability of sorbed metals as they are
discharged may result in an overly conservative
water quality standard. The use of dissolved
metal in water quality standards gives a more
accurate result in the water column. However,
total recoverable measurements in ambient water
have value, in that exceedences of criteria on a
total recoverable basis are an indication that metal
loadings could be a stress to the ecosystem,
particularly in locations other than the water
column (e.g., in the sediments).
The reasons for the potential consideration of total
recoverable measurements include risk
management considerations not covered by
evaluation of water column toxicity alone. The
ambient water quality criteria are neither designed
nor intended to protect sediments, or to prevent
effects in the food webs containing sediment
dwelling organisms. A risk manager, however,
may consider sediments and food chain effects
and may decide to take a conservative approach
for metals, considering that metals are very
persistent chemicals. This conservative approach
could include the use of total recoverable metal in
water quality standards. However, since
consideration of sediment impacts is not
incorporated into the criteria methodology, the
degree of conservatism inherent in the total
recoverable approach is unknown. The
uncertainty of metal impacts in sediments stem
from the lack of sediment criteria and an
imprecise understanding of the fate and transport
of metals. EPA will continue to pursue research
and other activities to close these knowledge gaps.
Dissolved Criteria
In the toxicity tests used to develop EPA metals
criteria for aquatic life, some fraction of the metal
is dissolved while some fraction is bound to
particulate matter. The present criteria were
developed using total recoverable metal
measurements or measures expected to give
equivalent results in toxicity tests, and are
articulated as total recoverable. Therefore, in
order to express the EPA criteria as dissolved, a
total recoverable to dissolved conversion factor
must be used. Attachment #2 in Appendix J
provides guidance for calculating EPA dissolved
criteria from the published total recoverable
criteria. The data expressed as percentage metal
dissolved are presented as recommended values
and ranges. However, the choice within ranges is
a State risk management decision. EPA has
recently supplemented the data for copper and is
proceeding to further supplement the data for
copper and other metals. As testing is completed,
EPA will make this information available and this
is expected to reduce the magnitude of the ranges
for some of the conversion factors provided.
EPA also strongly encourages the application of
dissolved criteria across a watershed or
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waterbody, as technically sound and the best use
of resources.
Site-Specific Criteria Modifications
While the above methods will correct some site-
specific factors affecting metals toxicity, further
refinements are possible. EPA has issued
guidance for three site-specific criteria
development methodologies: recalculation
procedure, water-effect ratio (WER) procedure
(called the indicator species procedure in previous
guidance) and resident species procedure. (See
Section 3.7 of this Chapter.)
In the National Toxics Rule (57 PR 60848,
December 22, 1992), EPA recommended the
WER as an optional method for site-specific
criteria development for certain metals. EPA
committed in the NTR preamble to provide
additional guidance on determining the WERs.
The Interim Guidance on the Determination and
Use of Water-Effect Ratios for Metals was issued
by EPA on February 22, 1994 and is intended to
fulfill that commitment. This interim guidance
supersedes all guidance concerning water-effect
ratios and the recalculation procedure previously
issued by EPA. This guidance is included as
Appendix L to this Handbook.
In order to meet current needs, but allow for
changes suggested by protocol users, EPA issued
the guidance as "interim." EPA will accept
WERs developed using this guidance, as well as
by using other scientifically defensible protocols.
3.6.3 Total Maximum Daily Loads (TMDLs)
and National Pollutant Discharge
Elimination System (NPDES) Permits
Dynamic Water Quality Modeling
Although not specifically part of the reassessment
of water quality criteria for metals, dynamic or
probabilistic models are another useful tool for
implementing water quality criteria, especially for
those criteria protecting" aquatic life. These
models provide another way to incorporate site-
specific data. The Technical Support Document
for Water Quality-based Toxics Control (TSD)
(USEPA, 199la) describes dynamic, as well as
static (steady-state) models. Dynamic models
make the best use of the specified magnitude,
duration, and frequency of water quality criteria
and, therefore, provide a more accurate
representation of the probability that a water
quality standard exceedence will occur. In
contrast, steady-state models frequently apply a
number of simplifying, worst case assumptions
which makes them less complex but also less
accurate than dynamic models.
Dynamic models have received increased attention
over the last few years as a result of the
widespread belief that steady-state modeling is
over-conservative due to environmentally
conservative dilution assumptions. This belief has
led to the misconception that dynamic models will
always lead to less stringent regulatory controls
(e.g., NPDES effluent limits) than steady-state
models, which is not true in every application of
dynamic models. EPA considers dynamic models
to be a more accurate approach to implementing
water quality criteria and continues to recommend
their use. Dynamic modeling does require a
commitment of resources to develop appropriate
data. (See Appendix J, Attachment #3 and the
USEPA, 1991a for details on the use of dynamic
models.)
Dissolved-Total Metal Translators
Expressing ambient water quality criteria for
metals as the dissolved form of a metal poses a
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need to be able to translate from dissolved metal
to total recoverable metal for TMDLs and
NPDES permits. TMDLs for metals must be able
to calculate: (1) dissolved metal in order to
ascertain attainment of water quality standards,
and (2) total recoverable metal in order to achieve
mass balance necessary for permitting purposes.
EPA's NPDES regulations require that limits of
metals in permits be stated as total recoverable in
most cases (see 40 CFR §122.45(c)) except when
an effluent guideline specifies the limitation in
another form of the metal, the approved analytical
methods measure only dissolved metal, or the
permit writer expresses a metals limit in another
form (e.g., dissolved, valent specific, or total)
when required to carry out provisions of the
Clean Water Act. This is because the chemical
conditions in ambient waters frequently differ
substantially from those in the effluent, and there
is no assurance that effluent paniculate metal
would not dissolve after discharge. The NPDES
rule does not require that State water quality
standards be expressed as total recoverable;
rather, the rule requires permit writers to translate
between different metal forms in the calculation of
the permit limit so that a total recoverable limit
can be established. Both the TMDL and NPDES
uses of water quality criteria require the ability to
translate between dissolved metal and total
recoverable metal. Appendix J, Attachment #3
provides guidance on this translation.
3.6.4 Guidance on Monitoring
Use of Clean Sampling and Analytical
Techniques
In assessing waterbodies to determine the potential
for toxicity problems due to metals, the quality of
the data used is an important issue. Metals data
are used to determine attainment status for water
quality standards, discern trends in water quality,
estimate background loads for TMDLs, calibrate
fate and transport models, estimate effluent
concentrations (including effluent variability),
assess permit compliance, and conduct research.
The quality of trace level metal data, especially
below 1 ppb, may be compromised due to
contamination of samples during collection,
preparation, storage, and analysis. Depending on
the level of metal present, the use of "clean" and
"ultraclean" techniques for sampling and analysis
may be critical to accurate data for
implementation of aquatic life criteria for metals.
The significance of the sampling and analysis
contamination problem increases as the ambient
and effluent metal concentration decreases and,
therefore, problems are more likely in ambient
measurements. "Clean" techniques refer to those
requirements (or practices for sample collection
and handling) necessary to produce reliable
analytical data in the part per billion (ppb) range.
"Ultraclean" techniques refer to those
requirements or practices necessary to produce
reliable analytical data in the part per trillion (ppt)
range. Because typical concentrations of metals
in surface waters and effluents vary from one
metal to another, the effect of contamination on
the quality of metals monitoring data varies
appreciably.
EPA plans to develop protocols on the use of
clean and ultra-clean techniques and is
coordinating with the United States Geological
Survey (USGS) on this project, because USGS has
been doing work on these techniques for some
time, especially the sampling procedures. Draft
protocols for clean techniques were presented at
the Norfolk, VA analytical methods conference in
the Spring of 1994 and final protocols are
expected to be available in early 1995. The
development of comparable protocols for ultra-
clean techniques is underway and are expected to
be available in late 1995. In developing these
protocols, we will consider the costs of these
techniques and will give guidance as to the
situations where their use is necessary. Appendix
L, pp. 98-108 provide some general guidance on
the use of clean analytical techniques. We
recommend that this guidance be used by States
and Regions as an interim step, while the clean
and ultra-clean protocols are being developed.
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Use of Historical Data
The concerns about metals sampling and analysis
discussed above raise corresponding concerns
about the validity of historical data. Data on
effluent and ambient metal concentrations are
collected by a variety of organizations including
Federal agencies (e.g., EPA, USGS), State
pollution control agencies and health departments,
local government agencies, municipalities,
industrial dischargers, researchers, and others.
The data are collected for a variety of purposes as
discussed above.
Concern about the reliability of the sample
collection and analysis procedures is greatest
where they have been used to monitor very low
level metal concentrations. Specifically, studies
have shown data sets with contamination problems
during sample collection and laboratory analysis,
that have resulted in inaccurate measurements.
For example, in developing a TMDL for New
York Harbor, some historical ambient data
showed extensive metals problems in the harbor,
while other historical ambient data showed only
limited metals problems. Careful resampling and
analysis in 1992/1993 showed the latter view was
correct. The key to producing accurate data is
appropriate quality assurance (QA) and quality
control (QC) procedures. EPA believes that most
historical data for metals, collected and analyzed
with appropriate QA and QC at levels of 1 ppb or
higher, are reliable. The data used in
development of EPA criteria are also considered
reliable, both because they meet the above test
and because the toxicity test solutions are created
by adding known amounts of metals.
With respect to effluent monitoring reported by an
NPDES permittee, the permittee is responsible for
collecting and reporting quality data on a
Discharge Monitoring Report (DMR). Permitting
authorities should continue to consider the
information reported to be true, accurate, and
complete as certified by the permittee. Where the
permittee becomes aware of new information
specific to the effluent discharge that questions the
quality of previously submitted DMR data, the
permittee must promptly submit that information
to the permitting authority. The permitting
authority will consider all information submitted
by the permittee in determining appropriate
enforcement responses to monitoring/reporting
and effluent violations. (See Appendix J,
Attachment #4 for additional details.)
Site-Specific Aquatic Life Criteria
The purpose of this section is to provide guidance
for the development of site-specific water quality
criteria which reflect local environmental
conditions. Site-specific criteria are allowed by
regulation and are subject to EPA review and
approval. The Federal water quality standards
regulation at section 131.11(b)(l)(ii) provides
States with the opportunity to adopt water quality
criteria that are ".. .modified to reflect site-specific
conditions." Site-specific criteria, as with all
water quality criteria, must be based on a sound
scientific rationale in order to protect the
designated use. Existing guidance and practice
are that EPA will approve site-specific criteria
developed using appropriate procedures.
A site-specific criterion is intended to come closer
than the national criterion to providing the
intended level of protection to the aquatic life at
the site, usually by taking into account the
biological and/or chemical conditions (i.e., the
species composition and/or water quality
characteristics) at the site. The fact that the U.S.
EPA has made these procedures available should
not be interpreted as implying that the agency
advocates that states derive site-specific criteria
before setting state standards. Also, derivation of
a site-specific criterion does not change the
intended level of protection of the aquatic life at
the site.
3.7.1 History of Site-Specific Criteria
Guidance
National water quality criteria for aquatic life may
be under- or over-protective if:
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(1) the species at the site are more or less
sensitive than those included in the national
criteria data set (e.g., the national criteria
data set contains data for trout, salmon,
penaeid shrimp, and other aquatic species
that have been shown to be especially
sensitive to some materials), or
(2) physical and/or chemical characteristics of
the site alter the biological availability
and/or toxicity of the chemical (e.g.,
alkalinity, hardness, pH, suspended solids
and salinity influence the concentration(s) of
the toxic form(s) of some heavy metals,
ammonia and other chemicals).
Therefore, it is appropriate that site-specific
procedures address each of these conditions
separately as well as the combination of the two.
In the early 1980's, EPA recognized that
laboratory-derived water quality criteria might not
accurately reflect site-specific conditions and, in
response, created three procedures to derive site-
specific criteria. This Handbook contains the
details of these procedures, referenced below.
1. The Recalculation Procedure is intended to
take into account relevant differences
between the sensitivities of the aquatic
organisms in the national dataset and the
sensitivities of organisms that occur at the
site (see Appendix L, pp. 90-97).
2. The Water-Effect Ratio Procedure (called the
Indicator Species Procedure in USEPA,
1983a; 1984f ) provided for the use of a
water-effect ratio (WER) that is intended to
take into account relevant differences
between the toxicities of the chemical in
laboratory dilution water and in site water
(see Appendix L).
3. The Resident Species Procedure intended to
take into account both kinds of differences
simultaneously (see Section 3.7.6).
These procedures were first published in the 1983
Water Quality Standards Handbook (USEPA,
1983a) and expanded upon in the Guidelines for
Deriving Numerical Aquatic Site-Specific Water
Quality Criteria by Modifying National Criteria
(USEPA, 1984f). Interest has increased in recent
years as states have devoted more attention to
chemical-specific water quality criteria for aquatic
life. In addition, interest in water-effect ratios
increased when they were integrated into some of
the aquatic life criteria for metals that were
promulgated for several states in the National
Toxics Rule (57 FR 60848, December 22, 1992).
The Office of Water Policy and Technical
Guidance on Interpretation and Implementation of
Aquatic Life Criteria for Metals (USEPA, 1993f)
(see Section 3.6 of this Handbook) provided
further guidance on site-specific criteria for metals
by recommending the use of dissolved metals for
setting and measuring compliance with water
quality standards.
The early guidance concerning WERs (USEPA,
1983a; 1984f) contained few details and needed
revision, especially to take into account newer
guidance concerning metals. To meet this need,
EPA issued Interim Guidance on the
Determination and Use of Water-Effect Ratios for
Metals in 1994 (Appendix L). Metals are
specifically addressed in Appendix L because of
the National Toxics Rule and because of current
interest in aquatic life criteria for metals; although
most of this guidance also applies to other
pollutants, some obviously applies only to metals.
Appendix L supersedes all guidance concerning
water-effect ratios and the Indicator Species
Procedure given in Chapter 4 of the Water
Quality Standards Handbook (USEPA, 1983a) and
in Guidelines for Deriving Numerical Aquatic Site-
Specific Water Quality Criteria by Modifying
National Criteria (USEPA, 1984f). Appendix L
(p. 90-98) also supersedes the guidance in these
earlier documents for the Recalculation Procedure
for performing site-specific criteria modifications.
The Resident Species Procedure remains
essentially unchanged since 1983 (except for
changes in the averaging periods to conform to
the 1985 aquatic life criteria guidelines (USEPA,
1985b) and is presented in Section 3.7.6, below.
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The previous guidance concerning site-specific
procedures did not allow the Recalculation
Procedure and the WER procedure to be used
together in the derivation of a site-specific aquatic
life criterion; the only way to take into account
both species composition and water quality
characteristics in the determination of a site-
specific criterion was to use the Resident Species
Procedure. A specific change contained Appendix
L is that, except in jurisdictions that are subject to
the National Toxics Rule, the Recalculation
Procedure and the WER Procedure may now be
used together provided that the recalculation
procedure is performed first. Both the
Recalculation Procedure and the WER Procedure
are based directly on the guidelines for deriving
national aquatic life criteria (USEPA 1985 ) and,
when the two are used together, use of the
Recalculation Procedure must be performed first
because the Recalculation Procedure has specific
implications concerning the determination of the
WER.
3.7.2 Preparing to
Criteria
Calculate Site-Specific
Adopting site-specific criteria in water quality
standards is a State optionnot a requirement.
Moreover, EPA is not advocating that States use
site-specific criteria development procedures for
setting all aquatic life criteria as opposed to using
the National Section 304(a) criteria
recommendations. Site-specific criteria are not
needed in all situations. When a State considers
the possibility of developing site-specific criteria,
it is essential to involve the appropriate EPA
Regional office at the start of the project.
This early planning is also essential if it appears
that data generation and testing may be conducted
by a party other than the State or EPA. The State
and EPA need to apply the procedures judiciously
and must consider the complexity of the problem
and the extent of knowledge available concerning
the fate and effect of the pollutant under
consideration. If site-specific criteria are
developed without early EPA involvement in the
planning and design of the task, the State may
expect EPA to take additional time to closely
scrutinize the results before granting any approval
to the formally adopted standards.
The following sequence of decisions need to be
made before any of the procedures are initiated:
* verify that site-specific criteria are actually
needed (e.g., that the use of clean sampling
and/or analytical techniques, especially for
metals, do not result in attainment of
standards.)
+ Define the site boundaries.
4 Determine from the national criterion
document and other sources if physical
and/or chemical characteristics are known to
affect the biological availability and/or
toxicity of a material of interest.
4 If data in the national criterion document
and/or from other sources indicate that the
range of sensitivity of the selected resident
species to the material of interest is different
from the range for the species in the national
criterion document, and variation in physical
and/or chemical characteristics of the site
water is not expected to be a factor, use the
Recalculation Procedure (Section 3.7.4).
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4 If data in the national criterion document
and/or from other sources indicate that
physical and/or chemical characteristics of
the site water may affect the biological
availability and/or toxicity of the material of
interest, and the selected resident species
range of sensitivity is similar to that for the
species in the national criterion document,
use the Water-Effect Ratio Procedure
(Section 3.7.5).
4 If data in the national criterion document
and/or from other sources indicated that
physical and/or chemical characteristics of
the site water may affect the biological
availability and/or toxicity of the material of
interest, and the selected resident species
range of sensitivity is different from that for
the species in the national criterion
document, and if both these differences are
to be taken into account, use the
Recalculation Procedure in conjunction with
the Water-Effect Ratio Procedure or use the
Resident Species Procedure (Section 3.7.6).
3.7.3 Definition of a Site
Since the rationales for site-specific criteria are
usually based on potential differences in species
sensitivity, physical and chemical characteristics
of the water, or a combination of the two, the
concept of site must be consistent with this
rationale.
In the general context of site-specific criteria, a
"site" may be a state, region, watershed,
waterbody, or segment of a waterbody. The site-
specific criterion is to be derived to provide
adequate protection for the entire site, however
the site is defined.
If water quality effects on toxicity are not a
consideration, the site can be as large as a
generally consistent biogeographic zone permits.
For example, large portions of the Chesapeake
Bay, Lake Michigan, or the Ohio River may be
considered as one site if their respective aquatic
communities do not vary substantially. However,
when a site-specific criterion is derived using the
Recalculation Procedure, all species that "occur at
the site" need to be taken into account when
deciding what species, if any, are to be deleted
from the dataset. Unique populations or less
sensitive uses within sites may justify a
designation as a distinct site.
If the species of a site are lexicologically
comparable to those in the national criteria data
set for a material of interest, and physical and/or
chemical water characteristics are the only factors
supporting modification of the national criteria,
then the site can be defined on the basis of
expected changes in the material's biological
availability and/or toxicity due to physical and
chemical variability of the site water. However,
when a site-specific criterion is derived using a
WER, the WER is to be adequately protective of
the entire site. If, for example, a site-specific
criterion is being derived for an estuary, WERs
could be determined using samples of the surface
water obtained from various sampling stations,
which, to avoid confusion, should not be called
"sites". If all the WERs were sufficiently similar,
one site-specific criterion could be derived to
apply to the whole estuary. If the WERs were
sufficiently different, either the lowest WER could
be used to derive a site-specific criterion for the
whole estuary, or the data might indicate that the
estuary should be divided into two or more sites,
each with its own criterion.
3.7.4 The Recalculation Procedure
The Recalculation Procedure is intended to cause
a site-specific criterion to appropriately differ
from a national aquatic life criterion if justified by
demonstrated pertinent toxicological differences
between the aquatic species that occur at the site
and those that were used in the derivation of the
national criterion. There are at least three reasons
why such differences might exist between the two
sets of species.
4 First, the national dataset contains aquatic
species that are sensitive to many pollutants,
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but these and comparably sensitive species
might not occur at the site.
4 Second, a species that is critical at the site
might be sensitive to the pollutant and
require a lower criterion. (A critical species
is a species that is commercially or
recreationally important at the site, a species
that exists at the site and is listed as
threatened or endangered under section 4 of
the Endangered Species Act, or a species for
which there is evidence that the loss of the
species from the site is likely to cause an
unacceptable impact on a commercially or
recreationally important species, a threatened
or endangered species, the abundances of a
variety of other species, or the structure or
function of the community.)
+ Third, the species that occur at the site
might represent a narrower mix of species
than those in the national dataset due to a
limited range of natural environmental
conditions.
The procedure presented in Appendix L, pp. 90-
98 is structured so that corrections and additions
can be made to the national dataset without the
deletion process being used to take into account
taxa that do not occur at the site; in effect, this
procedure makes it possible to update the national
aquatic life criterion. All corrections and
additions that have been approved by EPA are
required, whereas use of the deletion process is
optional. The deletion process may not be used
to remove species from the criterion calculation
that are not currently present at a site due to
degraded conditions.
The Recalculation Procedure is more likely to
result in lowering a criterion if the net result of
addition and deletion is to decrease the number of
genera in the dataset, whereas the procedure is
more likely to result in raising a criterion if the
net result of addition and deletion is to increase
the number of genera in the dataset.
For the lipid soluble chemicals whose national
Final Residue Values are based on Food and Drug
Administration (FDA) action levels, adjustments
in those values based on the percent lipid content
of resident aquatic species is appropriate for the
derivation of site-specific Final Residue Values.
For lipid-soluble materials, the national Final
Residue Value is based on an average 11 percent
lipid content for edible portions for the freshwater
chinook salmon and lake trout and an average of
10 percent lipids for the edible portion for
saltwater Atlantic herring. Resident species of
concern may have higher (e.g., Lake Superior
siscowet, a race of lake trout) or lower (e.g.,
many sport fish) percent lipid content than used
for the national Final Residue Value.
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For some lipid-soluble materials such as
polychlorinated biphenyls (PCB) and DDT, the
national Final Residue Value is based on wildlife
consumers of fish and aquatic invertebrate species
rather than an FDA action level because the
former provides a more stringent residue level.
See the National Guidelines (USEPA, 1985b) for
details.
For the lipid-soluble materials whose national
Final Residue Values are based on wildlife
effects, the limiting wildlife species (mink for
PCB and brown pelican for DDT) are considered
acceptable surrogates for resident avian and
mammalian species (e.g., herons, gulls, terns,
otter, etc.) Conservatism is appropriate for those
two chemicals, and no less restrictive modification
of the national Final Residue Value is appropriate.
The site-specific Final Residue Value would be
the same as the national value.
3.7.5 The Water-Effect Ratio (WER)
Procedure
The guidance on the Water-Effect Ratio Procedure
presented in Appendix L is intended to produce
WERs that may be used to derive site-specific
aquatic life criteria from most national and state
aquatic life criteria that were derived from
laboratory toxicity data.
As indicated in Appendix L, the
determination of a water-effect ratio may require
substantial resources. A discharger should
consider cost-effective, preliminary measures
described in this Appendix L (e.g., use of "clean"
sampling and chemical analytical techniques
especially for metals, or in non-NTR States, a
recalculated criterion) to determine if an indicator
species site-specific criterion is really needed. In
many instances, use of these other measures may
eliminate the need for deriving water-effect ratios.
The methods described in the 1994 interim
guidance (Appendix L) should be sufficient to
develop site-specific criteria that resolve concerns
of dischargers when there appears to be no
instream toxicity but, where (a) a discharge
appears to exceed existing or proposed water
quality-based permit limits, or (b) an instream
concentration appears to exceed an existing or
proposed water quality criterion.
WERs obtained using the methods described in
Appendix L should only be used to adjust aquatic
life criteria that were derived using laboratory
toxicity tests. WERs determined using the
methods described herein cannot be used to adjust
the residue-based mercury Criterion Continuous
Concentration (CCC) or the field-based selenium
freshwater criterion.
Except in jurisdictions that are subject to the
NTR, the WERs may also be used with site-
specific aquatic life criteria that are derived using
the Recalculation Procedure described in
Appendix L (p.90).
Water-Effect Ratios in the Derivation of
Site-Specific Criteria
A central question concerning WERs is whether
their use by a State results in a site-specific
criterion subject to EPA review and approval
under Section 303 (c) of the Clean Water Act?
Derivation of a water-effect ratio by a State is a
site-specific criterion adjustment subject to EPA
review and approval/disapproval under Section
303(c). There are two options by which this
review can be accomplished.
Option 1:
A State may derive and submit each individual
water-effect ratio determination to EPA for review
and approval. This would be accomplished
through the normal review and revision process
used by a State.
Option 2:
A State can amend its water quality standards to
provide a formal procedure which includes
derivation of water-effect ratios, appropriate
definition of sites, and enforceable monitoring
provisions to assure that designated uses are
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protected. Both this procedure and the resulting
criteria would be subject to full public
participation requirements. EPA would review
and approve/disapprove this protocol as a revised
standard as part of the State's triennial
review/revision. After adoption of the procedure,
public review of a site-specific criterion could be
accomplished in conjunction with the public
review required for permit issuance. For public
information, EPA recommends that once a year
the State publish a list of site-specific criteria.
An exception to this policy applies to the waters
of the jurisdictions included in the National
Toxics Rule. The EPA review is not required for
the jurisdictions included in the National Toxics
Rule where EPA established the procedure for the
State for application to the criteria promulgated.
The National Toxics Rule was a formal
rulemaking process (with notice and comment) in
which EPA pre-authorized the use of a correctly
applied water-effect ratio. That same process has
not yet taken place in States not included in the
National Toxics Rule.
However, the National Toxics Rule does not
affect State authority to establish scientifically
defensible procedures to determine Federally
authorized WERs, to certify those WERs in
NPDES permit proceedings, or to deny their
application based on the State's risk management
analysis.
As described in Section 131.36(b)(iii) of the water
quality standards regulation (the official regulatory
reference to the National Toxics Rule), the water-
effect ratio is a site-specific calculation. As
indicated on page 60866 of the preamble to the
National Toxics Rule, the rule was constructed as
a rebuttable presumption. The water-effect ratio is
assigned a value of 1.0 until a different water-
effect ratio is derived from suitable tests
representative of conditions in the affected
waterbody. It is the responsibility of the State to
determine whether to rebut the assumed value of
1.0 in the National Toxics Rule and apply another
value of the water-effect ratio in order to establish
a site-specific criterion. The site-specific criterion
is then used to develop appropriate NPDES permit
limits. The rule thus provides a State with the
flexibility to derive an appropriate site-specific
criterion for specific waterbodies.
As a point of emphasis, although a water-effect
ratio affects permit limits for individual
dischargers, it is the State in all cases that
determines if derivation of a site-specific criterion
based on the water-effect ratio is allowed and it is
the State that ensures that the calculations and
data analysis are done completely and correctly.
3.7.6 The Resident Species Procedure
The resident Species Procedure for the derivation
of a site-specific criterion accounts for differences
in resident species sensitivity and differences in
biological availability and/or toxicity of a material
due to variability in physical and chemical
characteristics of a site water. Derivation of the
site-specific criterion maximum concentration
(CMC) and site-s;pecific criterion continuous
concentration (CCC) are accomplished after the
complete acute toxicity minimum data set
requirements have been met by conducting tests
with resident species in site water. Chronic tests
may also be necessary. This procedure is
designed to compensate concurrently for any real
differences between the sensitivity range of
species represented in the national data set and for
site water which may markedly affect the
biological availability and/or toxicity of the
material of interest.
Certain families of organisms have been specified
in the National Guidelines acute toxicity minimum
data set (e.g., Salmonidae in fresh water and
Penaeidae or Mysidae in salt water); if this or any
other requirement cannot be met because the
family or other group (e.g., insect or benthic
crustacean) in fresh water is not represented by
resident species, select a substitute(s) from a
sensitive family represented by one or more
resident species and meet the 8 family minimum
data set requirement. If all the families at the site
have been tested and the minimum data set
requirements have not been met, use the most
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Chapter 3 - Water Quality Criteria
sensitive resident family mean acute value as the
site-specific Final Acute Value.
To derive the criterion maximum concentration
divide the site-specific Final Acute Value by two.
The site-specific Final Chronic Value can be
obtained as described in the Appendix L. The
lower of the site-specific Final Chronic Value (as
described in the recalculation procedure -
Appendix L, p. 90) and the recalculated site-
specific Final Residue Value becomes the site-
specific criterion continuous concentration unless
plant or other data (including data obtained from
the site-specific tests) indicates a lower value is
appropriate. If a problem is identified, judgment
should be used in establishing the site-specific
criterion.
The frequency of testing (e.g., the need for
seasonal testing) will be related to the variability
of the physical and chemical characteristics of site
water as it is expected to affect the biological
availability and/or toxicity of the material of
interest. As the variability increases, the
frequency of testing will increase. Many of the
limitations discussed for the previous two
procedures would also apply to this procedure.
Endnotes
1. Proceedings in production.
Contact: Ecological Risk Assessment Branch (4304)
U.S. Environmental Protection Agency
401 M Street, S.W.
Washington, DC 20460
Telephone (202) 260-1940
(8/15/94) 3-45
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Chapter 4 - Antidegradation
CHAPTER 4
ANTIDEGRADATION
(40 CFR 131.12)
Table of Contents
4.1 History of Antidegradation 4-1
4.2 Summary of the Antidegradation Policy 4-1
4.3 State Antidegradation Requirements 4-2
4.4 Protection of Existing Uses - 40 CFR 131.12(a)(l) 4-3
4.4.1 Recreational Uses 4-4
4.4.2 Aquatic Life/Wildlife Uses 4-5
4.4.3 Existing Uses and Physical Modifications 4-5
4.4.4 Existing Uses and Mixing Zones 4-6
4.5 Protection of Water Quality in High-Quality Waters - 40 CFR 131.12(a)(2) 4-6
4.6 Applicability of Water Quality Standards to Nonpoint Sources Versus Enforceability
of Controls 4-9
4.7 Outstanding National Resource Waters (ONRW) - 40 CFR 131.12(a)(3) ....... 4-10
4.8 Antidegradation Application and Implementation 4-10
4.8.1 Antidegradation, Load Allocation, Waste Load Allocation, Total Maximum
Daily Load, and Permits 4-12
4.8.2 Antidegradation and the Public Participation Process 4-13
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Chapter 4 - Antidegradation
High-quality waters are those whose quality
exceeds that necessary to protect the section
101(a)(2) goals of the Act, regardless of use
designation. All parameters do not need to be
better quality than the State's ambient criteria for
the water to be deemed a "high-quality water."
EPA believes that it is best to apply
antidegradation on a parameter-by-parameter
basis. Otherwise, there is potential for a large
number of waters not to receive antidegradation
protection, which is important to attaining the
goals of the Clean Water Act to restore and
maintain the integrity of the Nation's waters.
However, if a State has an official interpretation
that differs from this interpretation, EPA will
evaluate the State interpretation for conformance
with the statutory and regulatory intent of the
antidegradation policy. EPA has accepted
approaches that do not use a strict pollutant-by-
pollutant basis (USEPA, 1989c).
In "high-quality waters," under 131.12(a)(2),
before any lowering of water quality occurs, there
must be an antidegradation review consisting of:
a finding that it is necessary to accommodate
important economical or social development
in the area in which the waters are located
(this phrase is intended to convey a general
concept regarding what level of social and
economic development could be used to
justify a change in high-quality waters);
full satisfaction of all intergovernmental
coordination and public participation
provisions (the intent here is to ensure that
no activity that will cause water quality to
decline in existing high-quality waters is
undertaken without adequate public review
and intergovernmental coordination); and
assurance that the highest statutory and
regulatory requirements for point sources,
including new source performance standards,
and best management practices for nonpoint
source pollutant controls are achieved (this
requirement ensures that the limited
provision for lowering water quality of high-
quality waters down to "fishable/swimmable"
levels will not be used to undercut the Clean
Water Act requirements for point source and
nonpoint source pollution control;
furthermore, by ensuring compliance with
such statutory and regulatory controls, there
is less chance that a lowering of water
quality will be sought to accommodate new
economic and social development).
In addition, water quality may not be lowered to
less than the level necessary to fully protect the
"fishable/swimmable" uses and other existing
uses. This provision is intended to provide relief
only in a few extraordinary circumstances where
the economic and social need for the activity
clearly outweighs the benefit of maintaining water
quality above that required for
"fishable/swimmable" water, and both cannot be
achieved. The burden of demonstration on the
individual proposing such activity will be very
high. In any case, moreover, the existing use
must be maintained and the activity shall not
preclude the maintenance of a
"fishable/swimmable" level of water quality
protection.
The antidegradation review requirements of this
provision of the antidegradation policy are
triggered by any action that would result in the
lowering of water quality in a high-quality water.
Such activities as new discharges or expansion of
existing facilities would presumably lower water
quality and would not be permissible unless the
State conducts a review consistent with the
previous paragraph. In addition, no permit may
be issued, without an antidegradation review, to
a discharger to high-quality waters with effluent
limits greater than actual current loadings if such
loadings will cause a lowering of water quality
(USEPA, 1989c).
Antidegradation is not a "no growth" rule and was
never designed or intended to be such. It is a
policy that allows public decisions to be made on
important environmental actions. Where the State
intends to provide for development, it may decide
under this section, after satisfying the
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Water Quality Standards Handbook - Second Edition
requirements for intergovernmental coordination
and public participation, that some lowering of
water quality in "high-quality waters" is necessary
to accommodate important economic or social
development. Any such lower water quality must
protect existing uses fully, and the State must
assure that the highest statutory and regulatory
requirement for all new and existing point sources
and all cost-effective and reasonable BMPs for
nonpoint source control are being achieved on the
water body.
Section 131.12(a)(2) does not REQUIRE a State
to establish BMPs for nonpoint sources where
such BMP requirements do not exist. We
interpret Section 131.12(a)(2) as REQUIRING
States to adopt an antidegradation policy that
includes a provision that will assure that all cost-
effective and reasonable BMPs established under
State authority are implemented for nonpoint
sources before the State authorizes degradation of
high quality waters by point sources (see USEPA,
1994a.)
Section 131.12(a)(2) does not mandate that States
establish controls on nonpoint sources. The Act
leaves it to the States to determine what, if any,
controls on nonpoint sources are needed to
provide for attainment of State water quality
standards (See CWA Section 319.) States may
adopt enforceable requirements, or voluntary
programs to address nonpoint source pollution.
Section 40 CFR 131.12(a)(2) does not require that
States adopt or implement best management
practices for nonpoint sources prior to allowing
point source degradation of a high quality water.
However, States that have adopted nonpoint
source controls must assure that such controls are
properly implemented before authorization is
granted to allow point source degradation of water
quality.
The rationale behind the antidegradation
regulatory statement regarding achievement of
statutory requirements for point sources and all
cost effective and reasonable BMPs for nonpoint
sources is to assure that, in high quality waters,
where there are existing point or nonpoint source
control compliance problems, proposed new or
expanded point sources are not allowed to
contribute additional pollutants that could result in
degradation. Where such compliance problems
exist, it would be inconsistent with the philosophy
of the antidegradation policy to authorize the
discharge of additional pollutants in the absence of
adequate assurance that any existing compliance
problems will be resolved.
EPA's regulation also requires maintenance of
high quality waters except where the State finds
that degradation is "necessary to accommodate
important economic and social development in the
area in which the waters are located." (40 CFR
Part 131.12(a) (Emphasis added)). We believe
this phrase should be interpreted to prohibit point
source degradation as unnecessary to
accommodate important economic and social
development if it could be partially or completely
prevented through implementation of existing
State-required BMPs.
EPA believes that its antidegradation policy
should be interpreted on a pollutant-by-pollutant
and waterbody-by-waterbody basis. For example,
degradation of a high quality waterbody by a
proposed new BOD source prior to
implementation of required BMPs on the same
waterbody that are related to BOD loading should
not be allowed. However, degradation by the
new point source of BOD should not be barred
solely on the basis that BMPs unrelated to BOD
loadings, or which relate to other waterbodies,
have not been implemented.
We recommend that States explain in their
antidegradation polices or procedures how, and to
what extent, the State will require implementation
of otherwise non-enforceable (voluntary) BMPs
before allowing point source degradation of high
quality waters. EPA understands this
recommendation exceeds the Federal requirements
discussed in this guidance. For example,
nonpoint source management plans being
developed under section 319 of the Clean Water
Act are likely to identify potential problems and
certain voluntary means to correct those
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Chapter 4 - Antidegradation
problems. The State should consider how these
provisions will be implemented in conjunction
with the water quality standards program.
Applicability of Water Quality
Standards to Nonpoint Sources
Versus Enforceability of Controls
The requirement in Section 131.21(a)(2) to
implement existing nonpoint source controls
before allowing degradation of a high quality
water, is a subset of the broader issue of the
applicability of water quality standards versus the
enforceability of controls designed to implement
standards. A discussion of the broader issue is
included here with the intent of further clarifying
the nonpoint source antidegradation question. In
the following discussion, the central message is
that water quality standards apply broadly and it
is inappropriate to exempt whole classes of
activities from standards and thereby invalidate
that broader, intended purpose of adopted State
water quality standards.
Water quality standards serve the dual function of
establishing water quality goals for a specific
waterbody and providing the basis for regulatory
controls. Water quality standards apply to both
point and nonpoint sources. There is a direct
Federal implementation mechanism to regulate
point sources of pollution but no parallel Federal
regulatory process for nonpoint sources. Under
State law, however, States can and do adopt
mandatory nonpoint source controls.
State water quality standards play the central role
in a State's water quality management program,
which identifies the overall mechanism States use
to integrate the various Clean Water Act water
quality control elements into a coherent
management framework. This includes, for
example: (1) setting and revising water quality
standards for all surface waterbodies, (2)
monitoring water quality to provide information
upon which water quality-based decisions will be
made, progress evaluated, and success measured,
(3) preparing a water quality inventory report
under section 305 (b) which documents the status
of the States's water quality, (4) developing a
water quality management plan which lists the
standards, and prescribes the regulatory and
construction activities necessary to meet the
standards, (5) calculating total maximum daily
loads and wasteload allocations for point sources
of pollution and load allocations for nonpoint
sources of pollution in the implementation of
standards, (6) implementing the section 319
management plan which outlines the State's
control strategy for nonpoint sources of pollution,
and (7) developing permits under Section 402.
Water quality standards describe the desired
condition of the aquatic environment, and, as
such, reflect any activity that affects water
quality. Water quality standards have broad
application and use in evaluating potential impacts
of water quality from a broad range of causes and
sources and are not limited to evaluation of effects
caused by the discharge of pollutants from point
sources. In this regard, States should have in
place methods by which the State can determine
whether or not their standards have been achieved
(including uses, criteria, and implementation of an
antidegradation policy). Evaluating attainment of
standards is basic to successful application of a
State's water quality standards program. In the
broad application of standards, these evaluations
are not limited to those activities which are
directly controlled through a mandatory process.
Rather, these evaluations are an important
component of a State's water quality management
program regardless of whether or not an
enforcement procedure is in place for the activity
under review.
Water quality standards are implemented through
State or EPA-issued water quality-based permits
and through State nonpoint source control
programs. Water quality standards are
implemented through enforceable NPDES permits
for point sources and through the installation and
maintenance of BMPs for nonpoint sources.
Water quality standards usually are not considered
self-enforcing except where they are established as
enforceable under State law. Application of water
quality standards in the overall context of a water
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quality management program, however, is not
limited to activities for which there are
enforceable implementation mechanisms.
In simple terms, applicability and enforceability
are two distinctly separate functions in the water
quality standards program. Water quality
standards are applicable to all waters and in all
situations, regardless of activity or source of
degradation. Implementation of those standards
may not be possible in all circumstances; in such
cases, the use attainability analysis may be
employed. In describing the desired condition of
the environment, standards establish a benchmark
against which all activities which might affect that
desired condition are, at a minimum, evaluated.
Standards serve as the basis for water quality
monitoring and there is value in identifying the
source and cause of a exceedance even if, at
present, those sources of impact are not regulated
otherwise controlled.
It is acceptable for a State to specify particular
classes of activities for which no control
requirements have been established in State law.
It is not acceptable, however, to specify that
standards do not apply to particular classes of
activities (e.g. for purposes of monitoring and
assessment). To do so would abrogate one of the
primary functions of water quality standards.
Outstanding National Resource
Waters (ONRW) - 40 CFR
Outstanding National Resource Waters (ONRWs)
are provided the highest level of protection under
the antidegradation policy. The policy provides
for protection of water quality in high-quality
waters that constitute an ONRW by prohibiting
the lowering of water quality. ONRWs are often
regarded as highest quality waters of the United
States: That is clearly the thrust of 131.12(a)(3).
However, ONRW designation also offers special
protection for waters of "exceptional ecological
significance." These are water bodies 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 or whose
characteristics cannot be adequately described by
these parameters (such as wetlands).
The regulation requires water quality to be
maintained and protected in ONRWs. EPA
interprets this provision to mean no new or
increased discharges to ONRWs and no new or
increased discharge to tributaries to ONRWs that
would result in lower water quality in the
ONRWs. The only exception to this prohibition,
as discussed in the preamble to the Water Quality
Standards Regulation (48 F.R. 51402), permits
States to allow some limited activities that result
in temporary and short-term changes in the water
quality of ONRW. Such activities must not
permanently degrade water quality or result in
water quality lower than that necessary to protect
the existing uses in the ONRW. It is difficult to
give an exact definition of "temporary" and
"short-term" because of the variety of activities
that might be considered. However, in rather
broad terms, EPA's view of temporary is weeks
and months, not years. The intent of EPA's
provision clearly is to limit water quality
degradation to the shortest possible time. If a
construction activity is involved, for example,
temporary is defined as the length of time
necessary to construct the facility and make it
operational. During any period of time when,
after opportunity for public participation in the
decision, the State allows temporary degradation,
all practical means of minimizing such
degradation shall be implemented. Examples of
situations in which flexibility is appropriate are
listed in Exhibit 4-1.
Antidegradation
Implementation
Application and
Any one or a combination of several activities
may trigger the antidegradation policy analysis.
Such activities include a scheduled water quality
standards review, the establishment of new or
revised load allocations, waste load allocations,
total maximum daily loads, issuance of NPDES
permits, and the demonstration of need for
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Chapter 4 - Antide%radation
Example 3, A national park wishes to replace a defective septic tank-drainfietd
'. system in a campground. The campground is located immediately
adjacent to a smalt stream with the ONRW use
Under the regulation, the construction eould occur if best management practices were
scrupulously followed to minimize any disturbance of Water quality Or aquatic habitat.
Example 2 Same situation except the campground is Served by a smalt sewage
treatment plant already discharging to the ONRW. It is desired to
enlarge the treatment system and provide higher levels of treatment.
Under the regulation, this water-quality-enhancing action would he permitted if there Was
only temporary increase in. sediment and, perhaps, in organic loading, which Would occur
during the actual construction phase.
Example 3 A National forest with a mature, second growth of trees which are
suitable for harvesting, with associated road repair and
re-stabilization. Streams in the area are designated as ONRW and
support trout fishing.
The regulation intends that best management practices for timber harvesting be followed
, and might include preventive measures more stringent than for similar logging in less
environmentally Sensitive areas. Of course, if the lands Were being considered for
designation as wilderness areas, or other similar designations, EPA's regulation should not
be construed as encouraging or condoning timbering operations. The regulation allows
only temporary and short-term water quality degradation while maintaining existing uses
or new uses consistent with the purpose of the management of the ONRW area.
Other examples of these types of activities include maintenance and/or repair of existing boat ramps or boat
docks, restoration of existing sea walls, repair of existing stormwater pipes, and replacement or repair of
existing bridges.
Exhibit 4-1. Examples of Allowable Temporary Lowering of Water Quality in
Outstanding National Resource Waters
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Water Quality Standards Handbook - Second Edition
advanced treatment or request by private or public
agencies or individuals for a special study of the
water body.
Nonpoint source activities are not exempt from
the provisions of the antidegradation policy. The
language of section 131.12 (a)(2) of the
regulation: "Further, the State shall assure that
there shall be achieved the highest statutory and
regulatory requirements for all new and existing
point sources and all cost-effective and reasonable
best management practices for nonpoint source
control ..." reflects statutory provisions of the
Clean Water Act. While it is true that the Act
does not establish a federally enforceable program
for nonpoint sources, it clearly intends that the
BMPs developed and approved under sections
205Q), 208, 303(e), and 319 be aggressively
implemented by the States.
4.8.1 Antidegradation, Load Allocation,
Waste Load Allocation, Total Maximum
Daily Load, and Permits
In developing or revising a load allocation (LA),
waste load allocation (WLA), or total maximum
daily load (TMDL) to reflect new information or
to provide for seasonal variation, the
antidegradation policy, as an integral part of the
State water quality standards, must be applied as
discussed in this section.
The TMDL/WLA/LA process distributes the
allowable pollutant loadings to a water body. Such
allocations also consider the contribution to
pollutant loadings from nonpoint sources. This
process must reflect applicable State water quality
standards including the antidegradation policy.
No waste load allocation can be developed or
NPDES permit issued that would result in
standards being violated. With respect to
antidegradation, that means existing uses must be
protected, water quality may not be lowered in
ONRWs, and in the case of waters whose quality
exceeds that necessary for the section 101(a)(2)
goals of the Act, an activity cannot result in a
lowering of water quality unless the applicable
public participation, intergovernmental review,
and baseline control requirements of the
antidegradation policy have been met. Once the
LA, WLA, or TMDL revision is completed, the
resulting permits must incorporate discharge
limitations based on this revision.
When a pollutant discharge ceases for any reason,
the waste load allocations for the other
dischargers in the area may be adjusted to reflect
the additional loading available consistent with the
antidegradation policy under two circumstances:
In "high-quality waters" where after the full
satisfaction of all public participation and
intergovernmental review requirements, such
adjustments are considered necessary to
accommodate important economic or social
development, and the "threshold" level
requirements (required point and nonpoint
source controls) are met.
In less than "high-quality waters," when the
expected improvement in water quality (from
the ceased discharge) would not cause a
better use to be achieved.
The adjusted loads still must meet water quality
standards, and the new waste load allocations
must be at least as stringent as technology-based
limitations. Of course, all applicable
requirements of the section 402 NPDES permit
regulations would have to be satisfied before a
permittee could increase its discharge.
If a permit is being renewed, reissued or modified
to include less stringent limitations based on the
revised LA/WLA/TMDL, the same
antidegradation analysis applied during the
LA/WLA/TMDL stage would apply during the
permitting stage. It would be reasonable to allow
the showing made during the LA/WLA/TMDL
stage to satisfy the antidegradation showing at the
permit stage. Any restrictions to less stringent
limits based on antibacksliding would also apply.
If a State issues an NPDES permit that violates
the required antidegradation policy, it would be
subject to a discretionary EPA veto under section
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Chapter 4 - Antidegradation
402(d) or to a citizen challenge. In addition to
actions on permits, any waste load allocations and
total maximum daily loads violating the
antidegradation policy are subject to EPA
disapproval and EPA promulgation of a new
waste load allocation/total maximum daily load
under section 303(d) of the Act. If a significant
pattern of violation was evident, EPA could
constrain the award of grants or possibly revoke
any Federal permitting capability that had been
delegated to the State. Where EPA issues an
NPDES permit, EPA will, consistent with its
NPDES regulations, add any additional or more
stringent effluent limitations required to ensure
compliance with the State antidegradation policy
incorporated into the State water quality
standards. If a State fails to require compliance
with its antidegradation policy through section 401
certification related to permits issued by other
Federal agencies (e.g., a Corps of Engineers
section 404 permit), EPA could comment
unfavorably upon permit issuance. The public, of
course, could bring pressure upon the permit
issuing agency.
For example applications of antidegradation in the
WLA and permitting process, see Exhibit 4-2.
4.8.2 Antidegradation and
Participation Process
the Public
Antidegradation, as with other water quality
standards activities, requires public participation
and intergovernmental coordination to be an
effective tool in the water quality management
process. 40 CFR 131.12(a)(2) contains explicit
requirements for public participation and
intergovernmental coordination when determining
whether to allow lower water quality in high-
quality waters. Nothing in either the water
quality standards or the waste load allocation
regulations requires the same degree of public
participation or intergovernmental coordination for
such non-high-quality waters as is required for
high-quality waters. However public participation
would still be provided in connection with the
issuance of a NPDES permit or amendment of a
208 plan. Also, if the action that causes
reconsideration of the existing waste loads (such
as dischargers withdrawing from the area) will
result in an improvement in water quality that
makes a better use attainable, even if not up to the
"fishable/swimmable" goal, then the water quality
standards must be upgraded and full public review
is required for any action affecting changes in
standards. Although not specifically required by
the standards regulation between the triennial
reviews, we recommend that the State conduct a
use attainability analysis to determine if water
quality improvement will result in attaining higher
uses than currently designated in situations where
significant changes in waste loads are expected.
The antidegradation public participation
requirement may be satisfied in several ways.
The State may hold a public hearing or hearings.
The State may also satisfy the requirement by
providing public notice and the opportunity for the
public to request a hearing. Activities that may
affect several water bodies in a river basin or sub-
basin may be considered in a single hearing. To
ease the resource burden on both the State and
public, standards issues may be combined with
hearings on environmental impact statements,
water management plans, or permits. However,
if this is done, the public must be clearly
informed that possible changes in water quality
standards are being considered along with other
activities. It is inconsistent with the water quality
standards regulation to "back-door" changes in
standards through actions on EIS's, waste load
allocations, plans, or permits.
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Water Quality Standards Handbook - Second Edition
Example 1
Several facilities on a stream segment discharge phosphoruS"COntaining wastes.
Ambient phosphorus concentrations meet the designated class B (nan*
fishable/swimmable) standards, but barely, Three dischargers achieve
elimination by developing land treatment systems. As a result, actual water
quality improves (i.e., phosphorus levels decline) but not quite to the level
needed to meet class A (ftshabte/swimmable) standards. Can the remaining
dischargers now be allowed to increase their phosphorus discharge without an
antidegradation analysis with the result that water quality declines (phosphorus
levels increase) to previous levels? ^
Nothing in the water quality standards regulation explicitly prohibits this* Of course* changes in their
NPDES permit limits may be subject to non-water quality constraints, suck as BPT, BAT, or the
NPDES antibacksliding provisions4 which may restrict the increased loads".
Example 2
Suppose, in the above situation, water quality improves to the point that actual
water quality now meets class A requirements. Is the answer different?
Yes. The standards must be upgraded (see section 2.8).
Example 3
As an alternative case, suppose phosphorus loadings go do wn and water quality
improves because of a change in farming practices (e,g., initiation of a
successful nonpoint source program.) Are the above answers the same?
Yes. Whether the improvement results from a change in point or nonpoint source activity js immaterial
to how any aspect of the standards regulation operates. Section 131.1
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Chapter 7 - The Water Quality-Based Approach to Pollution Control
CHAPTER?
THE WATER QUALITY-BASED
APPROACH TO
POLLUTION CONTROL
Table of Contents
7.1 Determine Protection Level 7-2
7.2 Conduct Water Quality Assessment 7-3
7.2.1 Monitor Water Quality 7-3
7.2.2 Identify Impaired (Water Quality-Limited) Waters 7-3
7.3 Establish Priorities 7-5
7.4 Evaluate Water Quality Standards for Targeted Waters 7-6
7.5 Define and Allocate Control Responsibilities 7-7
7.6 Establish Source Controls 7-8
7.6.1 Point Source Control - the NPDES Process 7-9
7.6.2 Nonpoint Source Controls 7-lQ
7.6.3 CWA Section 401 Certification 7-10
7.7 Monitor and Enforce Compliance 7-12
7.8 Measure Progress 7_13
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Chapter 7 - The Water Quality-Based Approach to Pollution Control
7.51 Define and Allocate Control
^ Responsibilities
For a water quality-limited water that still
requires a TMDL, a State must establish a TMDL
that quantifies pollutant sources, and a margin of
safety, and allocates allowable loads to the
contributing point and nonpoint source discharges
so that the water quality standards are attained.
The development of TMDLs should be
accomplished by setting priorities, considering the
geographic area impacted by the pollution
problem, and in some cases where there are
uncertainties from lack of adequate data, using a
phased approach to establishing control measures
based on the TMDL.
Many water pollution concerns are areawide
phenomena caused by multiple dischargers,
multiple pollutants (with potential synergistic and
additive effects), or nonpoint sources.
Atmospheric deposition and ground water
discharge may also result in significant pollutant
loadings to surface waters. As a result, EPA
recommends that States develop TMDLs on a
watershed basis to efficiently and effectively
manage the quality of surface waters. . -
The TMDL process is a rational method for
weighing the competing pollution concerns and
developing an integrated pollution reduction
strategy for point and nonpoint sources. The
TMDL process allows States to take a holistic
view of their water quality problems from the
perspective of instream conditions. Although
States may define a water body to correspond
with their current programs, it is expected that
States will consider the extent of pollution
problems and sources when defining the
geographic area for developing TMDLs. In
general, the geographical approach for TMDL
development supports sound environmental
management and efficient use of limited water
quality program resources. In cases where
TMDLs are developed on watershed levels, States
should consider organizing permitting cycles so
that all permits in a given watershed expire at the
same time.
Mathematical modeling is a valuable tool for
assessment of all types of water pollution
problems. Dissolved oxygen depletion and
nutrient enrichment from point sources are the
traditional modeling problems of the past. They
continue to be problems and are joined by such
new challenges as nonpoint source loadings, urban
stormwater runoff, toxics, and pollutants
involving sediment and bioaccumulative pathways.
These new pollutants and pathways require the
use of new models.
All models are simplifications of reality that
express our scientific understanding of the
important processes. Where we don't fully
understand the process(es), or cannot collect the
data that would be required to set parameters in a
model that would simulate the processes), we
make simplifying assumptions. All of these
iin'l^
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Water Quality Standards Handbook - Second Edition
simplifications increase the uncertainty of our
ability to predict responses of already highly-
variable systems. While the use of conservative
assumptions does reduce the possibility of
underestimating pollutants effects on the
waterbody, the use of conservative assumptions
does not reduce the uncertainty. Calibration of a
model to given waterbody does more to reduce
uncertainty surrounding the system's response to
reduced pollutant loadings. Sensitivity analyses
can further this process.
For TMDLs involving both traditional and
nontraditional problems, the margins of safety can
be increased and additional monitoring required to
verify attainment of water quality standards, and
provide data needed to recalculate the TMDL if
necessary (the phased approach).
EPA regulations provide that load allocations for
nonpoint sources and natural background "are best
estimates of the loading which may range from
reasonably accurate estimates to gross allotments
. . ." (40 CFR 130.2(g)). A phased approach to
developing TMDLs may be appropriate where
nonpoint sources are involved and where estimates
are based on limited information. Under the
phased approach, TMDL includes monitoring
requirements and a schedule for reassessing
TMDL allocations to ensure attainment of water
quality standards. Uncertainties that cannot be
quantified may also exist for certain pollutants
discharged primarily by point sources. In such
situations a large margin of safety and follow-up
monitoring are appropriate.
By pursuing the phased approach where
applicable, a State can move forward to
implement water quality-based control measures
and adopt an explicit schedule for implementation
and assessment. States can also use the phased
approach to address a greater number of water
bodies including threatened waters or watersheds
that would otherwise not be managed. Specific
requirements relating to the phased approach are
discussed in Guidance for Water Quality-based
Decisions: Vie TMDL Process (USEPA 199 Ic).
Establish Source Controls
Once a TMDL has been established for a water
body (or watershed) and the appropriate source
loads developed, implementation of control
actions should proceed. The State or EPA is
responsible for implementation, the first step
being to update the water quality management
plan. Next, point and nonpoint source controls
should be implemented to meet waste load
allocations and load allocations, respectively.
Various pollution allocation schemes (i.e.,
determination of allowable loading from different
pollution sources in the same water body) can be
employed by States to optimize alternative point
and nonpoint source management strategies.
The NPDES permitting process is used to limit
effluent from point sources. Section 7.6.1
provides a more complete description of the
NPDES process and how it fits into the water
quality-based approach to permitting.
Construction decisions regarding publicly owned
treatment works (POTWs), including advanced
treatment facilities, must also be based on the
more stringent of technology-based or water
quality-based limitations. These decisions should
be coordinated so that the facility plan for the
discharge is consistent with the limitations in the
permit.
In the case of nonpoint sources, both State and
local laws may authorize the implementation of
nonpoint source controls such as the installation of
best management practices (BMPs) or other
management measures. CWA section 319 and
Coastal Zone Act Reauthorization Amendments of
1990 (CZARA) section 6217 State management
programs may also be utilized to implement
nonpoint source control measures and practices to
ensure improved water quality. Many BMPs may
be implemented through section 319 programs
even where State regulatory programs do not
exist. In such cases, a State needs to document
the coordination that may be necessary among
State and local agencies, landowners, operators,
and managers and then evaluate BMP
7-8
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Chapter 7 - The Water Quality-Based Approach to Pollution Control
implementation, maintenance, and overall
effectiveness to ensure that load allocations are
achieved. Section 7.6.2 discusses some of the
programs associated with implementation of
nonpoint source control measures.
States may also grant, condition, or deny
"certification" for a federally permitted or
licensed activity that may result in a discharge to
the waters of the United States, if it is the State
where the discharge will originate. The State
decision is based on a State's determination of
whether the proposed activity will comply with
the requirements of certain sections of the Clean
Water Act, including water quality standards
under section 303. Section 7.6.3 of this
Handbook contains further discussion of section
401 certification.
7.6.1 Point Source Control - the NPDES
Process
Both technology-based and water quality-based
controls are implemented through the National
Pollutant Discharge Elimination System (NPDES)
permitting process. Permit limits based on
TMDLs are called water quality-based limits.
Waste load allocations establish the level of
effluent quality necessary to protect water quality
in the receiving water and to ensure attainment of
water quality standards. Once allowable loadings
have been developed through WLAs for specific
pollution sources, limits are incorporated into
NPDES permits. It is important to ensure that the
WLA accounts for the fact that effluent quality
is often highly variable. The WLA and permit
limit should be calculated to prevent water quality
standards impairment at all times. The reader is
referred to the Technical Support Document for
Water Quality-based Toxics Control (USEPA,
199 la) for additional information on deriving
permit limits.
As a result of the 1987 Amendments to the Act,
Individual Control Strategies (ICSs) were
established under section 304(1)(1) for certain
point source discharges of priority toxic
pollutants. ICSs consist of NPDES permit limits
and schedules for achieving such limits, along
with documentation showing that the control
measures selected are appropriate and adequate
(e.g., fact sheets including information on how
water quality-based limits were developed, such
as total maximum daily loads and waste load
allocations). Point sources with approved ICSs
are to be in compliance with those ICSs as soon
as possible or in no case later than 3 years from
the establishment of the ICS (typically by 1992 or
1993).
When establishing WLAs for point sources in a
watershed, the TMDL record should show that, in
the case of any credit for future nonpoint source
reductions (1) there is reasonable assurance that
nonpoint source controls will be implemented and
maintained, or (2) that nonpoint source reductions
are demonstrated through an effective monitoring
program. Assurances may include the application
or utilization of local ordinances, grant
conditions, or other enforcement authorities. For
example, it may be appropriate to provide that a
permit may be reopened when a WLA requiring
more stringent limits is necessary because
attainment of a nonpoint source load allocation
was not demonstrated.
Some compliance implementation time may, in
certain situations, be necessary and appropriate
for permittees to meet new permit limits based on
new standards. Under the Administrator's April
16, 1990 decision in an NPDES appeal (Star-Kist
Caribe Inc.. NPDES Appeal No. 88-5), the
Administrator stated that the only basis in which
a permittee may delay compliance after July 1,
1977 (for a post July 1977 standard), is pursuant
to a schedule of compliance established in the
permit which is authorized by the State in the
water quality standard itself or in other State
implementing regulations. Standards are made
applicable to individual dischargers through
NPDES permits which reflects the applicable
Federal or State water quality standards. When a
permit is issued, a schedule of compliance for
water quality-based limitations may be included,
as necessary.
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Water Quality Standards Handbook - Second Edition
7.6.2 Nonpoint Source Controls
In addition to permits for point sources, nonpoint
sources controls such as management measures or
best management practices (BMPs) are also to be
implemented so that surface water quality
objectives are met. To fully address water bodies
impaired or threatened by nonpoint source
pollution, States should implement their nonpoint
source management programs and ensure adoption
of control measures or practices by all
contributors of nonpoint source pollution to the
targeted watersheds.
Best management practices are the primary
mechanism in section 319 of the CWA to enable
achievement of water quality standards. Section
319 requires each State, in addition to developing
the assessment reports discussed in section 7.2.1
of this Handbook, to adopt NPS management
programs to control NPS pollution.
Sections 208(b)(2)(F) through (K) of the CWA
also require States to set forth procedures and
methods including land use requirements, to
control to the extent feasible nonpoint sources of
pollution reports.
Section 6217 of the Coastal Zone Reauthorization
Amendments of 1990 (CZARA) requires that
States with federally approved coastal zone
management programs develop Coastal Nonpoint
Pollution Control Programs to be approved by
EPA and NOAA. EPA and NOAA have issued
Coastal Nonpoint Pollution Control Program;
Program Development and Approval Guidance
(NOAA/EPA, 1993), which describes the
program development and approval process and
requirements. State programs are to employ an
initial technology-based approach generally
throughout the coastal management area, to be
followed by a more stringent water quality-based
approach to address known water quality
problems. The Management Measures generally
implemented throughout the coastal management
area are described in Guidance Specifying
Management Measures for Sources of Nonpoint
Pollution in Coastal Waters (USEPA, 1993b).
7.6.3 CWA Section 401 Certification
States may grant, condition, or deny
"certification" for a federally permitted or
licensed activity that may result in a discharge to
the waters of the United States, if it is the State
where the discharge will originate. The language
of section 401(a)(l) is very broad with respect to
the activities it covers:
[A]ny activity, including, but not
limited to, the construction or operation
of facilities, which may result in any
discharge . . .
requires water quality certification.
EPA has identified five Federal permits and/or
licenses that authorize activities that may result in
a discharge to the waters: permits for point
source discharge under section 402 and discharge
of dredged and fill material under section 404 of
the Clean Water Act; permits for activities in
navigable waters that may affect navigation under
sections 9 and 10 of the Rivers and Harbors Act
(RHA); and licenses required for hydroelectric
projects issued under the Federal Power Act.
There are likely other Federal permits and
licenses, such as permits for activities on public
lands, and Nuclear Regulatory Commission
licenses, which may result in a discharge and thus
require 401 certification. Each State should work
with EPA and the Federal agencies active in its
State to determine whether 401 certification is in
fact applicable.
7-10
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Chapter 7 - The Water Quality-Based Approach to Pollution Control
Congress intended for the States to use the water
quality certification process to ensure that no
Federal license or permits would be issued that
would violate State standards or become a source
of pollution in the future. Also, because the
States' certification of a construction permit or
license also operates as certification for an
operating permit (except in certain instances
specified in section 401 (a) (3)), it is imperative for
a State review to consider all potential water
quality impacts of the project, both direct and
indirect, over the life of the project.
In addition, when an activity requiring 401
certification in one State (i.e. the State in which
the discharge originates) will have an impact on
the water quality of another State, the statute
provides that after receiving notice of application
from a Federal permitting or licensing agency,
EPA will notify any States whose water quality
'may be affected. Such States have the right to
submit their objections and request a hearing.
EPA may also submit its evaluation and
recommendations. If the use of conditions cannot
ensure compliance with the affected State's water
quality requirements, the Federal permitting or
licensing agency shall not issue such permit or
license.
The decision to grant, condition, or deny
certification is based on a State's determination
from data submitted by an applicant (and any
other information available to the State) whether
the proposed activity will comply with the
requirements of certain sections of the Clean
Water Act enumerated in section 401(a)(l).
These requirements address effluent limitations
for conventional and nonconventional pollutants,
water quality standards, new source performance
standards, and toxic pollutants (sections 301, 302,
303, 306, and 307). Also included are
requirements of State law or regulation more
stringent than those sections or their Federal
implementing regulations.
States adopt surface water quality standards
pursuant to section 303 of the Clean Water Act
and have broad authority to base those standards
on the waters' use and value for "... public
water supplies, propagation of fish and wildlife,
recreational purposes, and . . . other purposes"
(33 U.S.C. section 1313 (c)(2)(A)). All permits
must include effluent limitations at least as
stringent as needed to maintain established
beneficial uses and to attain the quality of water
designated by States for their waters. Thus, the
States' water quality standards are a critical
concern of the 401 certification process.
If a State grants water quality certification to an
applicant for a Federal license or permit, it is in
effect saying that the proposed activity will
comply with State water quality standards (and the
other CWA and State law provisions enumerated
above). The State may thus deny certification
because the applicant has not demonstrated that
the project will comply with those requirements.
Or it may place whatever limitations or conditions
on the certification it determines are necessary to
ensure compliance with those provisions, and with
any other "appropriate" requirements of State law.
If a State denies certification, the Federal
permitting or licensing agency is prohibited from
issuing a permit or license. While the procedure
varies from State to State, a State's decision to
grant or deny certification is ordinarily subject to
an administrative appeal, with review in the State
courts designated for appeals of agency decisions.
Court review is typically limited to the question of
whether the State agency's decision is supported
by the record and is not arbitrary or capricious.
The courts generally presume regularity in agency
procedures and defer to agency expertise in their
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Water Quality Standards Handbook - Second Edition
review. (If the applicant is a Federal agency,
however, at least one Federal court has ruled that
the State's certification decision may be reviewed
by the Federal courts.)
States may also waive water quality certification,
either affirmatively or involuntarily. Under
section 401(a)(l), if the State fails to act on a
certification request "within a reasonable time
(which shall not exceed one year)" after the
receipt of an application, it forfeits its authority to
grant conditionally or to deny certification.
The most important regulatory tools for the
implementation of 401 certification are the States'
water quality standards regulations and their 401
certification implementing regulations and
guidelines. Most Tribes do not yet have water
quality standards, and developing them would be
a first step prior to having the authority to
conduct water quality certification. Also, many
States have not adopted regulations implementing
their authority to grant, deny, and condition water
quality certification. Wetland and 401
Certification: Opportunities and Guidelines for
States and Eligible Indian Tribes (USEPA, 1989a)
discusses specific approaches, and elements of
water quality standards and 401 certification
regulations that EPA views as effective to
implement the States' water quality certification
authority.
Monitor and Enforce Compliance
As noted throughout the previous sections,
monitoring is a crucial element of water
quality-based decision making. Monitoring
provides data for assessing compliance with water
quality-based controls and for evaluating whether
the TMDL and control actions that are based on
the TMDL protect water quality standards.
With point sources, dischargers are required to
provide reports on compliance with NPDES
permit limits. Their discharge monitoring reports
(DMR) provide a key source of effluent quality
data. In some instances, dischargers may also be
required in the permit to assess the impact of their
discharge on the receiving water. A monitoring
requirement can be put into the permit as a
special condition as long as the information is
collected for purposes of writing a permit limit.
States should also ensure that effective monitoring
programs are in place for evaluating nonpoint
source control measures. EPA recognizes
monitoring as a high-priority activity in a State's
nonpoint source management program (55 F.R.
35262, August 28, 1990). To facilitate the
implementation and evaluation of NPS controls,
States should consult current guidance (USEPA,
199 Ig); (USEPA, 1993b). States are also
encouraged to use innovative monitoring programs
(e.g., rapid bioassessments (USEPA, 1989e), and
volunteer monitoring (USEPA, 1990b) to provide
for adequate point and nonpoint source monitoring
coverage.
Dischargers are monitored to determine whether
or not they are meeting their permit conditions
and to ensure that expected water quality
improvements are achieved. If a State has not
been delegated authority for the NPDES permit
program, compliance reviews of all permittees in
that State are the responsibility of EPA. EPA
retains oversight responsibility for State
compliance programs in NPDES-delegated States.
NPDES permits also contain self-monitoring
requirements that are the responsibility of the
individual discharger. Data obtained through self-
monitoring are reported to the appropriate
regulatory agency.
Based on a review of data, EPA or a State
regulatory agency determines whether or not a
NPDES permittee has complied with the
requirements of the NPDES permit. If a facility
has been identified as having apparent violations,
EPA or the State will review the facility's
compliance history. This review focuses on the
magnitude, frequency, and duration of violations.
A determination of the appropriate enforcement
response is then made. EPA and States are
authorized to bring civil or criminal action against
facilities that violate their NPDES permits. State
7-12
(8/15/94)
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Chapter 7 - The Water Quality-Based Approach to Pollution Control
nonpoint source programs are enforced under
State law and to the extent provided by State law.
Once control measures have been implemented,
the impaired waters should be assessed to
determine if water quality standards have been
attained or are no longer threatened. The
monitoring program used to gather the data for
this assessment should be designed based on the
specific pollution problems or sources. For
example, it is difficult to ensure, a priori, that
implementing nonpoint source controls will
achieve expected load reductions due to
inadequate selection of practices or measures,
inadequate design or implementation, or lack of
full participation by all contributing nonpoint
sources (USEPA, 1987e). As a result, long-term
monitoring efforts must be consistent over time to
develop a data base adequate for analysis of
control actions.
Measure Progress
If the water body achieves the applicable State
water quality standards, the water body may be
removed from the 303 (d) list of waters still
needing TMDLs. If the water quality standards
are not met, the TMDL and allocations of load
and waste loads must be modified. This
modification should be based on the additional
data and information gathered as required by the
phased approach for developing a TMDL, where
appropriate; as part of routine monitoring
activities; and when assessing the water body for
water quality standards attainment.
(8/15/94) 7-13
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References
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Standards, Washington DC. EPA 440/5-86-001. USGPO #955-002-00000-8. (Source #7.)
REF-4 (8/15/94)
-------
References
. 1986b. Ambient Water Quality Criteria for Bacteria. Office of Water Regulations and
Standards, Washington DC. EPA 440/5-84-002. PB 86-158045. (Source #2.)
. 1986c. Technical Guidance Manual for Performing Waste Load Allocations, Book 6,
Design Conditions. Office of Water Regulations and Standards, Washington, DC. EPA 440/4-
87-002. (Source #10.)
. 1986d. Technical Guidance Manual for Performing Waste Load Allocations, Book VI,
Design Conditions: Chapter 1 - Stream Design Flow for Steady-State Modeling. Office of
Water Regulations and Standards, Washington, DC. EPA 440/4-87-004. (Source #10.)
. 1986e. Answers to Questions on Nonpoint Sources and WQS. (Memorandum from
Assistant Administrator for Water to Water Division Director, Region X; March 7.)
Washington, DC. (Source #11.)
. 1986f. Determination of "Existing Uses" for Purposes of Water Quality Standards
Implementation. (Memorandum from Director, Criteria and Standards Division to Water
Management Division Directors, Region I - X, WQS Coordinators, Region I - X; April 7.)
Washington, DC. (Source #11.)
. 1986. Technical Guidance Manual for Performing Waste Load Allocations. Book IV
Lakes, Reservoirs, and Impoundments. Chapter 3 Toxic Substances. Office of Water
Regulations and Standards, Washington, DC. EPA 440/4-87-002. (Source #10.)
. 1987d. Nonpoint Source Controls and Water Quality Standards. (Memorandum from
Chief, Nonpoint Source Branch to Regional Water Quality Branch Chiefs; August 19.)
Washington, DC. (Source #11.)
. 1987e. Setting Priorities: The Key to Nonpoint Source Control. Office of Water
Regulations and Standards. Washington, DC. (Source #8.)
. 1988a. Short-term Methods for Estimating the Chronic Toxicity of Effluents and Receiving
Waters to Marine and Estuarine Organisms. Office of Research and Development, Cincinnati,
OH. EPA 600/4-87-028.
. 1988d. State Clean Water Strategies; Meeting the Challenges for the Future. Office of
Water. Washington, DC. (Source #5.)
. 1988e. Guidance for State Implementation of Water Quality Standards for CWA Section
303(c)(2)(B). Office of Water. Washington, DC. (Source #10.)
. 1989a. Wetlands and 401 Certification: Opportunities for States and Eligible Indian
Tribes. Office of Wetlands Protection, Washington, DC. (Source #12.)
. 1989b. Exposure Factors Handbook. Office of Health and Environmental Assessment,
Washington, DC. EPA 600/8-89-043. (Source #9.)
(8/15/94) REF-5
-------
Water Quality Standards Handbook - Second Edition
. 1989c. Application of Antidegradation Policy to the Niagara River. (Memorandum from
Director, Office of Water Regulations and Standards to Director, Water Management Division,
Region II; August 4.) Washington, DC. (Source #11.)
. 1989d. Selecting Priority Nonpoint Source Projects: You Better Shop Around. Office of
Water; and Office of Policy, Planning and Evaluation. Washington, DC. EPA 506/2-89-003.
(Source #13.)
. 1989e. Rapid Bioassessment Protocols for Use in Streams and Rivers. Assessment and
Watershed Protection Division. Washington, DC. EPA 444/4-89-001. (Source #14.)
_. 1989f. EPA Designation of Outstanding National Resource Waters. (Memorandum from
Acting Director, Criteria and Standards Division to Regional Water Management Division
Directors; May 25.) Washington, DC. (Source #11.)
. 1989g. Guidance for the Use of Conditional Approvals for State WQS. (Memorandum
from Director, Office of Water Regulations and Standards to Water Division Directors,
Regions I - X; June 20.) Washington, DC. (Source #11.)
. 1989h. Designation of Recreation Uses. (Memorandum from Director, Criteria and
Standards Division to Director, Water Management Division, Region IV; September 7.)
Washington, DC. (Source #11.)
. 1989L Water Quality Criteria to Protect Wildlife Resources. Environmental Research
Laboratory. Corvallis, OR. EPA 600/3-89-067. NTIS #PB 89-220016. (Source #2.)
. 1989J. Assessing Human Health Risks from Chemically Contaminated Fish and Shellfish:
a Guidance Manual. Office of Water Regulations and Standards. Washington, DC. EPA
503/8-89-002. (Source #10.)
. 1990a. Biological Criteria, National Program Guidance for Surface Waters. Office of
Water Regulations and Standards, Washington, DC. EPA 440/5-90-004. (Source #10)
. 1990b. Volunteer Water Monitoring: A Guide for State Managers. Office of Water.
Washington, DC. EPA 440/4-90-010. (Source #14.)
. 1990c. The Lake and Reservoir Restoration Guidance Manual, Second Edition. Office of
Water. Nonpoint Source Branch. Washington, DC. EPA 440/4-90-006. (Source #14.)
. 199 la. Technical Support Document for Water Quality-based Toxics Control. Office of
Water, Washington, DC. EPA 505/2-90-001. NTIS #PB 91-127415. (Source #2.)
. 1991b. Methods for the Determination of Metals in Environmental Samples.
Environmental Monitoring Systems Laboratory, Cincinnati, OH 45268. EPA 600/4-91-010.
(Source #9.)
. 199 Ic. Guidance for Water Quality-based Decisions: The TMDL Process. Office of
Water, Washington, DC. EPA 440/4-91-001 (Source #14.)
REF-6 (8/15/94)
-------
References
. 199 Id. Methods for Measuring the Acute Toxicity of Effluents to Aquatic Organisms. 4th.
ed. Office of Research and Development, Cincinnati, OH. EPA 600/4-90-027. (Source #9.)
. 1991e. Short-term Methods for Estimating the Chronic Toxicity of Effluents and Receiving
Waters to Freshwater Organisms. 3d. ed. Office of Research and Development, Cincinnati,
OH. EPA 600/4-91-002. (Source #9.)
. 199 If. Short-term Methods for Estimating the Chronic Toxicity of Effluents and Receiving
Waters to Marine and Estuarine Organisms. 2d. ed. Office of Research and Development,
Cincinnati, OH. EPA 600/4-91-003. (Source #9.)
. 199 Ig. Watershed Monitoring and Reporting Requirements for Section 319 National
Monitoring Program Projects. Assessment and Watershed Protection Division. Washington
DC. (Source #8.)
. 199Ih. Section 401 Certification and FERC Licenses. (Memorandum from Assistant
Administrator, Office of Water to Secretary, Federal Energy Regulatory Commission; January
18.) Washington, DC. (Source #11.)
. 199li. Policy on the Use of Biological Assessments and Criteria in the Water Quality
Program. (Memorandum from Director, Office of Science and Technology to Water
Management Division Directors, Regions I - X; June 19.) (Source #4.)
. 1992b. Interim Guidance on Interpretation and Implementation of Aquatic Life Criteria
for Metals. 57 F.R. 24041. Office of Science and Technology. Washington, DC. (Source
#4.)
. 1993a. Guidelines for Preparation of the 1994 State Water Quality Assessments 305(b)
Reports. Office of Wetlands, Oceans and Watersheds. Washington, DC. (Source #14.)
. 1993b. Guidance Specifying Management Measures for Sources ofNonpoint Pollution in
Coastal Waters. Office of Water. Washington, DC. 840-B-92-002. (Source #8.)
. 1993c. Geographic Targeting: Selected State Examples. Office of Water. Washington,
DC. EPA 841-B-93-001. (Source #14.)
: 1993d. Final Guidance on the Award and Management ofNonpoint Source Program
Implementation Grants Under Section 319(h) of the Clean Water Act for Fiscal Year 1994 and
Future Years. Office of Water. Washington, DC. (Source #8.)
. 1993e. Guidance for Assessing Chemical Contaminant Data for Use in Fish Advisories;
Volume 1 - Fish Sampling and Analysis (in preparation). Office of Water. Washington, DC.
EPA 823-R-93-002. (Source #9.)
. 1993f. Office of Water Policy and Technical Guidance on Interpretation and
Implementation of Aquatic Life Metals Criteria. Office of Water. Washignton, DC.
(Source #10.)
(8/15/94) REF-7
-------
Water Quality Standards Handbook - Second Edition
_. 1994a. Interpretation of Federal Antidegradation Regulatory Requirement. Office of
Science and Technology. Washington, DC. (Source 11.)
. 1994b. Interim Guidance on Determination and Use of Water-Effect Ratios for Metals.
Office of Water. Washington, DC. EPA-823-B-94-001. (Source #10.)
Vernberg, W.B. 1983. Responses to Estuarine Stress. In: Ecosystems of the World: Estuaries and
Enclosed Seas. B.H. Ketchum, ed. Elsevier Scientific Publishing Company, New York, pp.
43-63.
Versar. 1984. Draft Assessment of International Mixing Zone Policies. Avoidance/Attraction
Characteristics, and Available Prediction Techniques. USEPA, Office of Water Regulations
and Standards and USEPA Office of Pesticides and Toxic Substances, Washington, DC.
Windom, H.L., J.T. Byrd, R.G. Smith, and F. Huan. 1991. Inadequacy of NASQAN Data for
Assessing Metals Trends in the Nation's Rivers. Environ. Sci. Technol. 25, 1137.
SOURCES OF DOCUMENTS
(1) Seth Ausubel
U.S. Environmental Protection Agency
Region 2
26 Federal Plaza
New York, NY 10278
Ph: (212) 264-6779
(2) National Technical Information Center
(NTIS)
5285 Front Royal Road
Springfield, VA 22161
Ph: (703)487-4650
(3) U. S. Environmental Protection Agency
Region 1
Water Quality Standards Coordinator
Water Division
JFK Federal Building
One Congress Street
Boston, MA 02203
Ph: (617) 565-3533
(4) U. S. Environmental Protection Agency
Health and Ecological Criteria Division
401 M Street, S.W. (4304)
Washington, DC 20460
Ph: (202)260-5389
(See Appendix V)
(5) U. S. Environmental Protection Agency
Office of Water
401 M Street, S.W. (4301)
Washington, DC 20460
Ph: (202)260-5700
REF-8
(8/15/94)
-------
References
(6) "U.S. General Accounting Office
Post Office Box 6015
Gaithersburg, MD 20877
Telephone: 202-512-6000
(First copy free)
(7) U.S. Government Printing Office
Superintendent of Documents
North Capitol Street H Streets, NW
Washington, DC 20401
Ph: (202) 783-3238
(8) U. S. Environmental Protection Agency
Nonpoint Source Control Branch
401 M Street, S.W. (4305F)
Washington, DC 20460
Ph: (202) 260-7100
(9) U.S. Environmental Protection Agency
Center for Environmental Research
Office of Research and Development
Room G72
26 West Martin Luther King Drive
Cincinnati, OH 45268
Ph: (513) 569-7562
(10) U. S. Environmental Protection Agency
Office of Water Resource Center
RC-4100
401 M Street, S.W.
Washington, DC 20460
Ph: (202) 260-7786 (voice mail
publication request line)
(See Appendix V)
(11) U. S. Environmental Protection Agency
Standards and Applied Science Division
401 M Street, S.W. (4305)
Washington, DC 20460
Fax: (202) 260-9830
Ph: (202)260-7301
(See Appendix V)
(12) U. S. Environmental Protection Agency
Wetlands Division
401 M Street, S.W. (4502F)
Washington, DC 20460
Ph: (202)260-7719
(13) EPIC
U. S. Environmental Protection Agency
11029 Kenwood Road
Building 5
Cincinnati, OH 45242
Fax: (513) 569-7186
Ph: (513)569-7980
(14) U. S. Environmental Protection Agency
Assessment and Watershed Protection
Division
401 M Street, S.W. (4503F)
Washington, DC 20460
Ph: (202) 260-7166
(8/15/94)
REF-9
-------
-------
vvEPA
United States
Environmental Protection
Agency
Office of Water
(4305)
EPA-823-B-94-005b
August 1994
Water Quality Standards
Handbook: Second Edition
Appendixes
Contains update #1
August 1994
"... to restore and maintain the chemical,
physical, and biological integrity of the Nation's
waters."
Section 101 (a) of the Clean Water Act
-------
-------
APPENDIX J
Attachments to Office of Water Policy and
Technical Guidance on Interpretation and
Implementation of Aquatic Life Metals Criteria
WATER QUALITY STANDARDS HANDBOOK
SECOND EDITION
-------
-------
ATTACHMENT #2
GUIDANCE DOCUMENT
ON DISSOLVED CRITERIA
Expression of Aquatic Life Criteria
October 1993
-------
-------
10-1-93
Percent Dissolved in Aquatic Toxicity Tests on Metals
The attached table contains all the data that were found
concerning the percent of the total recoverable metal that was
dissolved in aquatic toxicity tests. This table is intended to
contain the available data that are relevant to the conversion of
EPA's aquatic life criteria for metals from a total recoverable
basis to a dissolved basis. (A factor of 1.0 is used to convert
aquatic life criteria for metals that are expressed on the basis
of the acid-soluble measurement to criteria expressed on the
basis of the total recoverable measurement.) Reports by Grunwald
(1992) and Brungs et al. (1992) provided references to many of
the documents in vhich pertinent data were found. Each document
was obtained and examined to determine whether it contained
useful data.
"Dissolved" is defined as metal that passes through a 0.45-^m
membrane filter. If otherwise acceptable, data that were
obtained using 0.3-^m glass fiber filters and 0.1-^m, membrane
filters were used, and are identified in the table; these data
did not seem to be outliers.
Data were used only if the metal was in a dissolved inorganic
form when it was added to the dilution water. In addition, data
were used only if they were generated in water that would have
been acceptable for use as a dilution water in tests used in the
derivation of water quality criteria for aquatic life; in
particular, the pH had to be between 6.5 and 9.0, and the
concentrations of total organic carbon (TOC) and total suspended
solids (TSS) had to be below 5 mg/L. Thus most data generated
using river water would not be used.
Some data were not used for other reasons. Data presented by
farroll et al. (1979) for cadmium were not used because 9 of the
36 values were above 150%. Data presented by Davies et al.
(1976) for lead and Holcombe and Andrew (1978) for zinc were not
used because "dissolved" was defined on the basis of
polarography, rather than filtration.
Beyond this, the data were not reviewed for quality. Horowitz et
al. (1992) reported that a number of aspects of the filtration
procedure might affect the results. In addition, there might be
concern about use of "clean techniques" and adequate QA/QC.
Each line in the table is intended to represent a separate piece
of information. All of the data in the table were determined in
fresh water, because no saltwater data were found. Data are
becoming available for copper in salt water from the New York
-------
Harbor study; based on the first set of testss, Hansen (1993)
suggested that the average percent of the copper that is
dissolved in sensitive saltwater tests is in the range of 76 to
82 percent.
A thorough investigation of the percent of total recoverable
metal that is dissolved in toxicity tests might attempt to
determine if the percentage is affected by test technique
(static, renewal, flow-through), feeding (were the test animals
fed and, if so, what food and how much), water quality
characteristics (hardness, alkalinity, pH, salinity), test
organisms (species, loading), etc.
The attached table also gives the freshwater criteria
concentrations (CMC and CCC) because percentages for total
recoverable concentrations much (e.g., more than a factor of 3)
above or below the CMC and CCC are likely to be less relevant.
When a criterion is expressed as a hardness equation, the range
given extends from a hardness of 50 mg/L to a, hardness of 200
mg/L.
The following is a summary of the available information for each
metal:
Arsenic fill)
The data available indicate that the percent dissolved is about
100, but all the available data are for concentrations that are
much higher than the CMC and CCC.
Cadmium
Schuytema et al. (1984) reported that "there were no real
differences" between measurements of total and dissolved cadmium
at concentrations of 10 to 80 ug/L (pH - 6.7 to 7.8, hardness =
25 mg/L, and alkalinity = 33 mg/L); total and dissolved
concentrations were said to be "virtually equivalent".
The CMC and CCC are close together and only range from 0.66 to
8.6 ug/L. The only available data that are known to be in the
range of the CMC and CCC were determined with a glass fiber
filter. The percentages that are probably most relevant are 75,
92, 89, 78, and 80.
Chromium fill)
The percent dissolved decreased as the total recoverable
concentration increased, even though the highest concentrations
reduced the pH substantially. The percentages that are probably
-------
most relevant to the CMC are 50-75, whereas the percentages that
are probably most relevant to the CCC are 86 and 61.
Chromium(VI)
The data available indicate that the percent dissolved is about
100, but all the available data are for concentrations that are
much higher than the CMC and CCC.
Copper
Howarth and Sprague (1978) reported that the total and dissolved
concentrations of copper were "little different" except when the
total copper concentration was above 500 ug/L at hardness = 360
mg/L and pH = 8 or 9. Chakoumakos et al. (1979) found that the
percent dissolved depended more on alkalinity than on hardness,
pH, or the total recoverable concentration of copper.
Chapman (1993) and Lazorchak (1987) both found that the addition
of daphnid food affected the percent dissolved very little, even
though Chapman used yeast-trout chow-alfalfa whereas Lazorchak
used algae in most tests, but yeast-trout chow-alfalfa in some
tests. Chapman (1993) found a low percent dissolved with and
without food, whereas Lazorchak (1987) found a high percent
dissolved with and without food. All of Lazorchak's values were
in high hardness water; Chapman's one value in high hardness
water was much higher than his other values.
Chapman (1993) and Lazorchak (1987) both compared the effect of
food on the total recoverable LC50 with the effect of food on the
dissolved LC50. Both authors found that food raised both the
dissolved LC50 and the total recoverable LC50 in about the same
proportion, indicating that food did not raise the total
recoverable LC50 by sorbing metal onto food particles; possibly
the food raised both LCSOs by (a) decreasing the toxicity of
dissolved metal, (b) forming nontoxic dissolved complexes with
_he metal, or (c) reducing uptake.
The CMC and CCC are close together and only range from 6.5 to 34
ug/L. The percentages that are probably most relevant are 74,
95, 95, 73, 57, 53, 52, 64, and 91.
Lead
The data presented in Spehar et al. (1978) were from Holcombe et
al. (1976). Both Chapman (1993) and Holcombe et al. (1976) found
that the percent dissolved increased as the total recoverable
concentration increased. It would seem reasonable to expect more
precipitate at higher total recoverable concentrations and
-------
therefore a lower percent dissolved at higher concentrations.
The increase in percent dissolved with increasing concentration
might be due to a lowering of the pH as more metal is added if
the stock solution was acidic.
The percentages that are probably most relevant to the CMC are 9,
18, 25, 10, 62, 68, 71, 75, 81, and 95, whereas the percentages
that are probably most relevant to the CCC are 9 and 10.
Mercury
The only percentage that is available is 73, but it is for a
concentration that is much higher than the CMC.
Nickel
The percentages that are probably most relevant to the CMC are
88, 93, 92, and 100, whereas the only percentage that is probably
relevant to the CCC is 76.
Selenium
No data are available.
There is a CMC, but not a CCC. The percentage dissolved seems to
be greatly reduced by the food used to feed daphnids, but not by
the food used to feed fathead minnows, ;he percentages that are
probably most relevant to the CMC are 4. 79 ', 79, 73, 91, 90, and
93.
Zinc
The CMC and CCC are close together and only range from 59 to 210
ug/L. The percentages that are probably most relevant are 31,
77, 77, 99, 94, 100, 103, and 96.
-------
Recommended Values (%)A and Ranges of Measured Percent Dissolved
Considered Most Relevant in Fresh Water
Metal CMC CCC
Recommended Recommended
Value f%) (Range %) Value (%) (Range
Arsenic(III)
Cadmium
Chromium (III)
Chromium (VI)
Copper
Lead
Mercury
Nickel
Selenium
Silver
Zinc
95
85
85
95
85
50
35
85
NAE
85
85
100-1048
75-92
50-75
100B
52-95
9-95
73B
88-100
NAC
41-93
31-103
95
85
85
95
85
25
NAE
85
NAE
YYD
85
100-1048
75-92
61-86
100B
52-95
9-10
NAE
76
NAC
YYD
31-103
A The recommended values are based on current knowledge and are
subject to change as more data becomes available.
B All available data are for concentrations that are much higher
than the CMC.
c NA = No data are available.
D YY = A CCC is not available, and therefore cannot be adjusted.
E NA = Bioaccumulative chemical and not appropriate to adjust to
percent dissolved.
-------
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References
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Development. Part I. Fathead Minnows (Pimephales promelas) and
Goldfish (Carassius auratus) as Standard Fish in Bioassays and
Their Reaction to Potential Reference Toxicants. EPA-600/3-76-
061a. National Technical Information Service, Springfield, VA.
Page 24.
Benoit, D.A. 1975. Chronic Effects of Copper on Survival,
Growth, and Reproduction of the Bluegill (Lepomis macrochirus).
Trans. Am. Fish. Soc. 104:353-358.
Brungs, W.A., T.S. Holderman, and M.T. Southerland. 1992.
Synopsis of Water-Effect Ratios for Heavy Metals as Derived for
Site-Specific Water Quality Criteria.
Call, D.J., L.T. Brooke, and D.D. Vaishnav. 1982. Aquatic
Pollutant Hazard Assessments and Development of a Hazard
Prediction Technology by Quantitative Structure-Activity
Relationships. Fourth Quarterly Report. University of
Wisconsin-Superior, Superior, WI.
Carlson, A.R., H. Nelson, and D. Hammermeister. 1986a.
Development and Validation of Site-Specific Water Quality
Criteria for Copper. Environ. Toxicol. Chem. 5:997-1012.
Carlson, A.R., H. Nelson, and D. Hammermeister. 1986b.
Evaluation of Site-Specific Criteria for Copper and Zinc: An
Integration of Metal Addition Toxicity, Effluent and Receiving
Water Toxicity, and Ecological Survey Data. EPA/600/S3-86-026.
National Technical Information Service, Springfield, VA.
Carroll, J.J., S.J. Ellis, and W.S. Oliver. 1979. Influences of
Hardness Constituents on the Acute Toxicity of Cadmium to Brook
Trout (Salvelinus fontinalis).
Chakoumakos, C., R.C. Russo, and R.V. Thurston. 1979. Toxicity
of Copper to Cutthroat Trout (Salmo clarki) under Different
Conditions of Alkalinity, pH, and Hardness. Environ. Sci.
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Chapman, G.A. 1993. Memorandum to C. Stephan. June 4.
Davies, P.H.y J.P. Goettl, Jr., J.R. Sinley, and N.F. Smith.
1976. Acute and Chronic Toxicity of Lead to Rainbow Trout Salmo
gairdneri, in Hard and Soft Water. Water Res. 10:199-206.
Finlayson, B.J., and K.M Verrue. 1982. Toxicities of Copper,
Zinc, and Cadmium Mixtures to Juvenile Chinook Salmon. Trans.
Am. Fish. Soc. 111:645-650.
13
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Geckler, J.R., W.B. Horning, T.M. Neiheisel, Q.H. Pickering, E.L.
Robinson, and C.E. Stephan. 1976. Validity of Laboratory Tests
for Predicting Copper Toxicity in Streams. EPA-600/3-76-116.
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118.
Grunwald, D. 1992. Metal Toxicity Evaluation: Review, Results,
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Hammermeister, D., C. Northcott, L. Brooke, and D. Call. 1983.
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JRB Associates. 1983. Demonstration of the Site-specific
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Krawczyk, A.K. Ratcliff, and J.H. Gakstatter. 1984. Toxicity of
Cadmium in Water and Sediment Slurries to Daphnia magna.
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of Water Quality Criteria-Based Metal Mixtures on Three Aquatic
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Sprague, J.B. 1964. Lethal Concentration of Copper and Zinc for
Young Atlantic Salmon. J. Fish. Res. Bd. Canada 21:17-9926.
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15
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ATTACHMENT #3
GUIDANCE DOCUMENT
ON DYNAMIC MODELING AND TRANSLATORS
August 1993
Total Maximum Daily Loads (TMDLs) and Permits
o Dynamic Water Quality Modeling
Although not specifically part of the reassessment of water quality criteria for metals,
dynamic or probabilistic models are another useful tool for implementing water quality
criteria, especially those for protecting aquatic life. Dynamic models make best use of the
specified magnitude, duration, and frequency of water quality criteria and thereby provide a
more accurate calculation of discharge impacts on ambient water quality. In contrast, steady-
state modeling is based on various simplifying assumptions which makes it less complex and
less accurate than dynamic modeling. Building on accepted practices in water resource
engineering, ten years ago OW devised methods allowing the use of probability distributions
in place of worst-case conditions. The description of these models and their advantages and
disadvantages is found in the 1991 Technical Support Document for Water Quality-based
Toxic Control (TSD).
Dynamic models have received increased attention in the last few years as a result of
the perception that static modeling is over-conservative due to environmentally conservative
dilution assumptions. This has led to the misconception that dynamic models will always
justify less stringent regulatory controls (e.g. NPDES effluent limits) than static models. In
effluent dominated waters where the upstream concentrations are relatively constant,
however, a dynamic model will calculate a more stringent wasteload allocation than will a
steady state model. The reason is that the critical low flow required by many State water
quality standards in effluent dominated streams occurs more frequently than once every three
years. When other environmental factors (e.g. upstream pollutant concentrations) do not
vary appreciably, then the overall return frequency of the steady state model may be greater
than once in three years. A dynamic modeling approach, on the other hand, would be more
stringent, allowing only a once in three year return frequency. As a result, EPA considers
dynamic models to be a more accurate rather than a less stringent approach to implementing
water quality criteria.
The 1991 TSD provides recommendations on the use of steady state and dynamic
water quality models. The reliability of any modeling technique greatly depends on the
accuracy of the data used in the analysis. Therefore, the selection of a model also depends
upon the data. EPA recommends that steady state wasteload allocation analyses generally be
used where few or no whole effluent toxicity or specific chemical measurements are
available, or where daily receiving water flow records are not available. Also, if staff
resources are insufficient to use and defend the use of dynamic models, then steady state
models may be necessary. If adequate receiving water flow and effluent concentration data
are available to estimate frequency distributions, EPA recommends that one of the dynamic
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wasteload allocation modeling techniques be used to derive wasteload allocations which will
more exactly maintain water quality standards. The minimum data required for input into
dynamic models include at least 30 years of river flow data and one year of effluent and
ambient pollutant concentrations.
o Dissolved-Total Metal Translators
When water quality criteria are expressed as the dissolved form of a metal, there is a
need to translate TMDLs and NPDES permits to and from the dissolved form of a metal to
the total recoverable form. TMDLs for toxic metals must be able to calculate 1) the
dissolved metal concentration in order to ascertain attainment of water quality standards and
2) the total recoverable metal concentration in order to achieve mass balance. In meeting
these requirements, TMDLs consider metals to be conservative pollutants and quantified as
total recoverable to preserve conservation of mass. The TMDL calculates the dissolved or
ionic species of the metals based on factors such as total suspended solids (TSS) and ambient
pH. (These assumptions ignore the complicating factors of metals interactions with other
rnetals.) In addition, this approach assumes that ambient factors influencing metal
partitioning remain constant with distance down the river. This assumption probably is valid
under the low flow conditions typically used as design flows for permitting of metals (e.g.,
7Q10, 4B3, etc) because erosion, resuspension, and wet weather loadings are unlikely to be
significant and river chemistry is generally stable. In steady-state dilution modeling, metals
releases may be assumed to remain fairly constant (concentrations exhibit low variability)
with time.
EPA's NPDES regulations require that metals limits in permits be stated as total
recoverable in most cases (see 40 CFR §122.45(c)). Exceptions occur when an effluent
guideline specifies the limitation in another form of the metal or the approved analytical
methods measure only the dissolved form. Also, the permit writer may express a metals
limit hi another form (e.g., dissolved, valent, or total) when required, in highly unusual
cases, to carry out the provisions of the CWA.
The preamble to the September 1984 National Pollutant Discharge Elimination System
Permit Regulations states that the total recoverable method measures dissolved metals plus
that portion of solid metals that can easily dissolve under ambient conditions (see 49 Federal
Register 38028, September 26, 1984). This method is intended to measure metals hi the
effluent that are or may easily become environmentally active, while not measuring metals
that are expected to settle out and remain inert.
The preamble cites, as an example, effluent from an electroplating facility that adds
lime and uses clarifiers. This effluent will be a combination of solids not removed by the
clarifiers and residual dissolved metals. When the effluent from the clarifiers, usually with a
high pH level, mixes with receiving water having significantly lower pH level, these solids
instantly dissolve. Measuring dissolved metals in the effluent, in this case, would
underestimate the impact on the receiving water. Measuring with the total metals method, on
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the other hand, would measure metals that would be expected to disperse or settle out and
remain inert or be covered over. Thus, measuring total recoverable metals in the effluent
best approximates the amount of metal likely to produce water quality impacts.
However, the NPDES rule does not require in any way that State water quality
standards be in the total recoverable form; rather, the rule requires permit writers to consider
the translation between differing metal forms in the calculation of the permit limit so that a
total recoverable limit can be established. Therefore, both the TMDL and NPDES uses of
water quality criteria require the ability to translate from the dissolved form and the total
recoverable form.
Many toxic substances, including metals, have a tendency to leave the dissolved phase
and attach to suspended solids. The partitioning of toxics between solid and dissolved phases
can be determined as a function of a pollutant-specific partition coefficient and the
concentration of solids. This function is expressed by a linear partitioning equation:
C= ££
I+KJ-TSS-W-
where,
C = dissolved phase metal concentration,
CTf = total metal concentration,
TSS = total suspended solids concentration, and
Kd = partition coefficient.
A key assumption of the linear partitioning equation is that the sorption reaction
reaches dynamic equilibrium at the point of application of the criteria; that is, after allowing
for initial mixing the partitioning of the pollutant between the adsorbed and dissolved forms
can be used at any location to predict the fraction of pollutant in each respective phase.
Successful application of the linear partitioning equation relies on the selection of the
partition coefficient. The use of a partition coefficient to represent the degree to which
toxics adsorb to solids is most readily applied to organic pollutants; partition coefficients for
metals are more difficult to define. Metals typically exhibit more complex speciation and
complexation reactions than organics and the degree of partitioning can vary greatly
depending upon site-specific water chemistry. Estimated partition coefficients can be
determined for a number of metals, but waterbody or site-specific observations of dissolved
and adsorbed concentrations are preferred.
EPA suggests three approaches for instances where a water quality criterion for a
metal is expressed in the dissolved form in a State's water quality standards:
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1. Using clean analytical techniques and field sampling procedures with appropriate
QA/QC, collect receiving water samples and determine site specific values of Kd for
each metal. Use these Kd values to "translate" between total recoverable and
dissolved metals in receiving water. This approach is more difficult to apply because
it relies upon the availability of good quality measurements of ambient metal
concentrations. This approach provides an accurate assessment of the dissolved metal
fraction providing sufficient samples are collected. EPA's initial recommendation is
that at least four pairs of total recoverable and dissolved ambient metal measurements
be made during low flow conditions or 20 pairs over all flow conditions. EPA
suggests that the average of data collected during low flow or the 95th percentile
highest dissolved fraction for all flows be used. The low flow average provides a
representative picture of conditions during the rare low flow events. The 95th
percentile highest dissolved fraction for all flows provides a critical condition
approach analogous to the approach used to identify low flows and other critical
environmental conditions.
2. Calculate the total recoverable concentration for the purpose of setting the permit
limit. Use a value of 1 unless the permittee has collected data (see #1 above) to show
that a different ratio should be used. The value of 1 is conservative and will not err
on the side of violating standards. This approach is very simple to apply because it
places the entire burden of data collection and analysis solely upon permitted
facilities. In terms of technical merit, it has the same characteristics of the previous
approach. However, permitting authorities may be faced with difficulties in
negotiating with facilities on the amount of data necessary to determine the ratio and
the necessary quality control methods to assure that the ambient data are reliable.
3. Use the historical data on total suspended solids (TSS) in receiving waterbodies at
appropriate design flows and Kd values presented in the Technical Guidance Manual
for Performing Waste Load Allocations. Book II. Streams and Rivers. EPA-440/4-
84-020 (1984) to "translate" between (total recoverable) permits limits and dissolved
metals in receiving water. This approach is fairly simple to apply. However, these
Kd values are suspect due to possible quality assurance problems with the data used to
develop the values. EPA's initial analysis of this approach and these values in one
site indicates that these Kd values generally over-estimate the dissolved fraction of
metals in ambient waters (see Figures following). Therefore, although this approach
may not provide an accurate estimate of the dissolved fraction, the bias in the estimate
is likely to be a conservative one.
EPA suggests that regulatory authorities use approaches #1 and #2 where States
express their water quality standards in the dissolved form. In those States where the
standards are in the total recoverable or acid soluble form, EPA recommends that no
translation be used until the time that the State changes the standards to the dissolved form.
Approach #3 may be used as an interim measure until the data are collected to implement
approach #1.
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ATTACHMENT #4
GUIDANCE DOCUMENT
ON CLEAN ANALYTICAL TECHNIQUES AND MONITORING
October 1993
Guidance on Monitoring
o Use of Clean Sampling and Analytical Techniques
Pages 98-108 of the WER guidance document (Appendix L of the Water Quality
Standards Handbook-Second Edition) provides some general guidance on the use of clean
techniques. The Office of Water recommends that this guidance be used by States and
Regions as an interim step while the Office of Water prepares more detailed guidance.
o Use of Historical DMR Data
With respect to effluent or ambient monitoring data reported by an NPDES permittee
on a Discharge Monitoring Report (DMR), the certification requirements place the burden on
the permittee for collecting and reporting quality data. The certification regulation at 40
CFR 122.22(d) requires permittees, when submitting information, to state: "I certify under
penalty of law that this document and all attachments were prepared under my direction or
supervision in accordance with a system designed to assure that qualified personnel properly
gather and evaluate the information submitted. Based on my inquiry of the person or persons
who manage the system, or those persons directly responsible for gathering the information,
the information submitted is, to the best of my knowledge and belief, true, accurate, and
complete. I am aware that there are significant penalties for submitting false information,
including the possibility of fine and imprisonment for knowing violations."
Permitting authorities should continue to consider the information reported in DMRs
to be true, accurate, and complete as certified by the permittee. Under 40 CFR 122.41(1)(8),
however, as soon as the permittee becomes aware of new information specific to the effluent
discharge that calls into question the accuracy of the DMR data, the permittee must submit
such information to the permitting authority. Examples of such information include a new
finding that the reagents used hi the laboratory analysis are contaminated with trace levels of
metals, or a new study that the sampling equipment imparts trace metal contamination. This
information must be specific to the discharge and based on actual measurements rather than
extrapolations from reports from other facilities. Where a permittee submits information
supporting the contention that the previous data are questionable and the permitting authority
agrees with the findings of the information, EPA expects that permitting authorities will
consider such information in determining appropriate enforcement responses.
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18
In addition to submitting the information described above, the permittee also must
develop procedures to assure the collection and analysis of quality data that are true,
accurate, and complete. For example, the permittee may submit a revised quality assurance
plan that describes the specific procedures to be undertaken to reduce or eliminate trace
metal contamination.
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APPENDIX L
Interim Guidance on Determination and
Use of Water-Effect Ratios for Metals
WATER QUALITY STANDARDS HANDBOOK
SECOND EDITION
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FEE 22 1994
EPA-823-B-94-001
MEMORANDUM
SUBJECT: Use of the Water-Effect Ratio in Water Quality
Standards
FROM: Tudor T. Davies, Director
Office of Science and Technology
TO: Water Management Division Directors, Regions I - X
State Water Quality Standards Program Directors
PURPOSE
There are two purposes for this memorandum.
The first is to transmit the Interim Guidance on the
Determination and Use of Water-Effect Ratios for Metals. EPA
committed to developing this guidance to support implementation
of federal standards for those States included in the National
Toxics Rule.
The second is to provide policy guidance on whether a
State's application of a water-effect ratio is a site-specific
criterion adjustment subject to EPA review and
approval/disapproval.
BACKGROUND
In the early 1980's, members of the regulated community
expressed concern that EPA's laboratory-derived water quality
criteria might not accurately reflect site-specific conditions
because of the effects of water chemistry and the ability of
species to adapt over time. In response to these concerns, EPA
created three procedures to derive site-specific criteria. These
procedures were published in the Water Quality Standards
Handbook, 1983.
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Site-specific criteria are allowed by regulation and are
subject to EPA review and approval. The Federal water quality
standards regulation at section 131.11(b)(1) provides States with
the opportunity to adopt water quality criterisi that are
"...modified to reflect site-specific conditions." Under section
131.5 (a) (2), EPA reviews standards to determine: "whether a State
has adopted criteria to protect the designated water uses."
On December 22, 1992, EPA promulgated the National Toxics
Rule which established Federal water quality standards for 14
States which had not met the requirements of Clean Water Act
Section 303(c)(2)(B). As part of that rule, EPA gave the States
discretion to adjust the aquatic life criteria for metals to
reflect site-specific conditions through use of a water-effect
ratio. A water-effect ratio is a means to account for a
difference between the toxicity of the metal in. laboratory
dilution water and its toxicity in the water at. the site.
In promulgating the National Toxics Rule, EPA committed to
issuing updated guidance on the derivation of water-effect
ratios. The guidance reflects new information since the
previous guidance and is more comprehensive in order to provide
greater clarity and increased understanding. This new guidance
should help standardize procedures for deriving water-effect
ratios and make results more comparable and defensible.
Recently, an issue arose concerning the most appropriate
form of metals upon which to base water quality standards. On
October 1, 1993, EPA issued guidance on this issue which
indicated that measuring the dissolved form of metal is the
recommended approach. This new policy however, is prospective
and does not affect the criteria in the National Toxics Rule.
Dissolved metals criteria are not generally numerically equal to
total recoverable criteria and the October 1, 1.993 guidance
contains recommendations for correction factors for fresh water
criteria. The determination of site-specific criteria is
applicable to criteria expressed as either total recoverable
metal or as dissolved metal.
DISCUSSION
Existing guidance and practice are that EPA will approve
site- specific criteria developed using appropriate procedures.
That policy continues for the options set forth in the interim
guidance transmitted today, regardless of whether the resulting
criterion is equal to or more or less stringent than the EPA
national 304(a) guidance. This interim guidance supersedes all
guidance concerning water-effect ratios previously issued by the
Agency.
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Each of the three options for deriving a final water-effect
ratio presented in this interim guidance meets the scientific and
technical acceptability test for deriving site-specific criteria.
Option 3 is the simplest, least restrictive and generally the
least expensive approach for situations where simulated
downstream water appropriately represents a "site." It is a
fully acceptable approach for deriving the water-effect ratio
although it will generally provide a lower water-effect ratio
than the other 2 options. The other 2 options may be more costly
and time consuming if more than 3 sample periods and water-effect
ratio measurements are made, but are more accurate, and may yield
a larger, but more scientifically defensible site specific
criterion.
Site-specific criteria, properly determined, will fully
protect existing uses. The waterbody or segment thereof to which
the site-specific criteria apply must be clearly defined. A site
can be defined by the State and can be any size, small or large,
including a watershed or basin. However, the site-specific
criteria must protect the site as a whole. It is likely to be
more cost-effective to derive any site-specific criteria for as
large an area as possible or appropriate. It is emphasized that
site-specific criteria are ambient water quality criteria
applicable to a site. They are not intended to be direct
modifications to National Pollutant Discharge Elimination System
(NPDES) permit limits. In most cases the "site" will be
synonymous with a State's "segment" in its water quality
standards. By defining sites on a larger scale, multiple
dischargers can collaborate on water-effect ratio testing and
attain appropriate site-specific criteria at a reduced cost.
More attention has been given to water-effect ratios
recently because of the numerous discussions and meetings on the
entire question of metals policy and because WERs were
specifically applied in the National Toxics Rule. In comments on
the proposed National Toxics Rule, the public questioned whether
the EPA promulgation should be based solely on the total
recoverable form of a metal. For the reasons set forth in the
final preamble, EPA chose to promulgate the criteria based on the
total recoverable form with a provision for the application of a
water-effect ratio. In addition, this approach was chosen
because of the unique difficulties of attempting to authorize
site-specific criteria modifications for nationally promulgated
criteria.
EPA now recommends the use of dissolved metals for States
revising their water quality standards. Dissolved criteria may
also be modified by a site-specific adjustment.
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While the regulatory application of the water-effect ratio
applied only to the 10 jurisdictions included in the final
National Toxics Rule for aquatic life metals criteria, we
understood that other States would be interested in applying WERs
to their adopted water quality standards. The guidance upon
which to base the judgment of the acceptability of the water-
effect ratio applied by the State is contained in the attached
Interim Guidance on The Determination and Use of Water-Effect
Ratios for Metals. It should be noted that this guidance also
provides additional information on the recalculation procedure
for site-specific criteria modifications.
(WER) in non-National Toxics
Status of the Water-effect Ratio
Rule States
A central question concerning WERs is whether their use by a
State results in a site-specific criterion subject to EPA review
and approval under Section 303 (c) of the Clean Water Act?
Derivation of a water-effect ratio by a State is a site-
specific criterion adjustment subject to EPA review and
approval/disapproval under Section 303(c). There are two options
by which this review can be accomplished.
Option 1: A State may derive and submit each individual
water-effect ratio determination to EPA for review and
approval. This would be accomplished through the normal
review and revision process used by a State.
Option 2: A State can amend its water quality standards to
provide a formal procedure which includes derivation of
water-effect ratios, appropriate definition of sites, and
enforceable monitoring provisions to assure that designated
uses are protected. Both this procedure and the resulting
criteria would be subject to full public participation
requirements. Public review of a site-specific criterion
could be accomplished in conjunction with the public review
required for permit issuance. EPA would review and
approve/disapprove this protocol as a revised standard once.
For public information, we recommend that once a year the
State publish a list of site-specific criteria.
An exception to this policy applies to the waters of the
jurisdictions included in the National Toxics Rule. The EPA
review is not required for the jurisdictions included in the
National Toxics Rule where EPA established the procedure for the
State for application to the criteria promulgated. The National
Toxics Rule was a formal rulemaking process with notice and
comment by which EPA pre-authorized the use of a correctly
applied water-effect ratio. That same process has not yet taken
place in States not included in the National Toxics Rule.
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However, the National Toxics Rule does not affect State authority
to establish scientifically defensible procedures to determine
Federally authorized WERs, to certify those WERs in NPDES permit
proceedings, or to deny their application based on the State's
risk management analysis.
As described in Section 131.36(b) (iii) of the water quality
standards regulation (the official regulatory reference to the_
National Toxics Rule), the water-effect ratio is a site-specific
calculation. As indicated on page 60866 of the preamble to the
National Toxics Rule, the rule was constructed as a rebuttable
presumption. The water-effect ratio is assigned a value of 1.0
until a different water-effect ratio is derived from suitable
tests representative of conditions in the affected waterbody. It
is the responsibility of the State to determine whether to rebut
the assumed value of 1.0 in the National Toxics Rule and apply
another value of the water-effect ratio in order to establish a
site-specific criterion. The site-specific criterion is then
used to develop appropriate NPDES permit limits. The rule thus
provides a State with the flexibility to derive an appropriate
site-specific criterion for specific waterbodies.
As a point of emphasis, although a water-effect ratio
affects permit limits for individual dischargers, it is the State
in all cases that determines if derivation of a site-specific
criterion based on the water-effect ratio is allowed and it is
the State that ensures that the calculations and data analysis
are done completely and correctly.
CONCLUSION
This interim guidance explains and clarifies the use of
site-specific criteria. It is issued as interim guidance because
it will be included as part of the process underway for review
and possible revision of the national aquatic life criteria
development methodology guidelines. As part of that review, this
interim guidance is subject to amendment based on comments,
especially those from the users of the guidance. At the end of
the guidelines revision process the guidance will be issued as
"final."
EPA is interested in and encourages the submittal of high
quality datasets that can be used to provide insights into the
use of these guidelines and procedures. Such data and technical
comments should be submitted to Charles E. Stephan at EPA's
Environmental Research Laboratory at Duluth, MN. A complete
address, telephone number and fax number for Mr. Stephan are
included in the guidance itself. Other questions or comments
should be directed to the Standards and Applied Science Division
(mail code 4305, telephone 202-260-1315).
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There is attached to this memorandum a simplified flow
diagram and an implementation procedure. These are intended to
aid a user by placing the water-effect ratio procedure in the
context of proceeding from at site-specific criterion to a permit
limit. Following these attachments is the guidance itself.
Attachments
cc: Robert Perciasepe, OW
Martha G. Prothro, OW
William Diamond, SASD
Margaret Stasikowski, HECD
Mike Cook, OWEC
Cynthia Dougherty, OWEC
Lee Schroer, OGC
Susan Lepow, OGC
Courtney Riordan, ORD
ORD (Duluth and Narragansett Laboratories)
ESD Directors, Regions I - VIII, X
BSD Branch, Region IX
Water Quality Standards Coordinators, Regions I - X
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WATER-EFFECT RATIO IMPLEMENTATION
PRELIMINARY ANALYSIS & PLAN FORMULATION
- Site definition
How many discharges must be accounted for? Tributaries?
See page 17.
What is the waterbody type? (i.e., stream, tidal river,
bay, etc.). See page 44 and Appendix A.
How can these considerations best be combined to define
the relevant geographic "site"? See Appendix A @ page
82.
- Plan Development for Regulatory Agency Review
Is WER method 1 or 2 appropriate? (e.g., Is design flow
a meaningful concept or are other considerations
paramount?). See page 6.
Define the effluent & receiving water sample locations
Describe the temporal sample collection protocols
proposed. See page 48.
Can simulated site water procedure be done, or is
downstream sampling required? See Appendix A.
Describe the testing protocols - test species, test
type, test length, etc. See page 45, 50; Appendix I.
Describe the chemical testing proposed. See Appendix C.
Describe other details of study - flow measurement,
QA/QC, number of sampling periods proposed, to whom the
results are expected to apply, schedule, etc.
SAMPLING DESIGN FOR STREAMS
- Discuss the quantification of the design streamflow (e.g.,
7Q10) - USGS gage directly, by extrapolation from USGS
gage, or ?
- Effluents
measure flows to determine average for sampling day
collect 24 hour composite using "clean" equipment and
appropriate procedures; avoid the use of the plant's
daily composite sample as a shortcut.
- Streams
measure flow (use current meter or read from gage if
available) to determine dilution with effluent; and to
check if within acceptable range for use of the data
(i.e., design flow to 10 times the design flow).
collect 24 hour composite of upstream water.
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LABORATORY PROCEDURES (NOTE: These are described in detail in
interim guidance).
- Select appropriate primary & secondary tests
- Determine appropriate cmcWER and/or cccWER
- Perform chemistry using clean procedures, with methods
that have adequate sensitivity to measure: low
concentrations, and use appropriate QA/QC
- Calculate final water-effect ratio (FWER) for site.
See page 36.
IMPLEMENTATION
- Assign FWERs and the site specific criteria for each metal
to each discharger (if more than one).
- perform a waste load allocation and total maximum daily
load (if appropriate) so that each discharger is provided
a permit limit.
- establish monitoring condition for periodic evaluation of
instream biology (recommended)
- establish a permit condition for periodic testing of WER
to verify site-specific criterion (NTR recommendation)
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Interim Guidance on
Determination and Use of
Water-Effect Ratios for Metals
February 1994
U.S. Environmental Protection Agency
Office of Water
Office of Science and Technology
Washington, D.C.
Office of Research and Development
Environmental Research Laboratories
Duluth, Minnesota
Narragansett, Rhode Island
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NOTICES
This document has been reviewed by the Environmental Research
Laboratories, Duluth, MN and Narragansett, RI (Office of Research
and Development) and the Office of Science and Technology (Office
of Water), U.S. Environmental Protection Agency, and approved for
publication.
Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.
11
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FOREWORD
This document provides interim guidance concerning the
experimental determination of water-effect ratios (WERs) for
metals; some aspects of the use of WERs are also addressed. It
is issued in support of EPA regulations and policy initiatives
involving the application of water quality criteria and standards
for metals. This document is agency guidance only. It does not
establish or affect legal rights or obligations. It does not
establish a binding norm or prohibit alternatives not included in
the document. It is not finally determinative of the issues
addressed. Agency decisions in any particular case will be made
by applying the law and regulations on the basis of specific
facts when regulations are promulgated or permits are issued.
This document is expected to be revised periodically to reflect
advances in this rapidly evolving area. Comments, especially
those accompanied by supporting data, are welcomed and should be
sent to: Charles E. Stephan, U.S. EPA, 6201 Congdon Boulevard,
Duluth MN 55804 (TEL: 218-720-5510; FAX: 218-720-5539).
ill
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FEE 22 1994
OFFICE OF SCIENCE AND TECHNOLOGY POSITION STATEMENT
Section 131.11(b)(ii) of the water quality standards
regulation (40 CFR Part 131) provides the regulatory mechanism
for a State to develop site-specific criteria for use in water
quality standards. Adopting site-specific criteria in water
quality standards is a State option--not a requirement. The
Environmental Protection Agency (EPA) in 1983 provided guidance
on scientifically acceptable methods by which site-specific
criteria could be developed.
The interim guidance provided in this document supersedes all
guidance concerning water-effect ratios and the Indicator Species
Procedure given in Chapter 4 of the Water Quality Standards
Handbook issued by EPA in 1983 and in Guidelines for Deriving
Numerical Aquatic Site-Specific Water Quality Criteria by
Modifying National Criteria, 1984. Appendix B also supersedes
the guidance in these earlier documents for the Recalculation
Procedure for performing site-specific criteria modifications.
This interim guidance fulfills a commitment made in the final
rule to establish numeric criteria for priority toxic pollutants
(57 FR 60848, December 22, 1992, also known as the "National
Toxics Rule"). This guidance also is applicable to pollutants
other than metals with appropriate modifications, principally to
chemical analyses.
Except for the jurisdictions subject to the aquatic life
criteria in the national toxics rule, water-effect ratios are
site-specific criteria subject to review and approval by the
appropriate EPA Regional Administrator. Site-specific criteria
are new or revised criteria subject to the normal EPA review
requirements established in Clean Water Act § 303 (c) . For the
States in the National Toxics Rule, EPA has established that
site-specific water-effect ratios may be applied to the criteria
promulgated in the rule to establish site-specific criteria. The
water-effect ratio portion of these criteria would still be
subject to State review before the development of total maximum
daily loads, waste load allocations or translation into NPDES
permit limits. EPA would only review these water-effect ratios
during its oversight review of these State programs or review of
State-issued permits.
IV
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Each of the three options for deriving a final water-effect
ratio presented on page 36 of this interim guidance meets the
scientific and technical acceptability test for deriving site-
specific criteria specified in the water quality standards
regulation (40 CFR 131.11(a)). Option 3 is the simplest, least
restrictive and generally the least expensive approach for
situations where simulated downstream water appropriately
represents a "site." Option 3 requires experimental
determination of three water-effect ratios with the primary test
species that are determined during any season (as long as the
downstream flow is between 2 and 10 times design flow
conditions.) The final WER is generally (but not always) the
lowest experimentally determined WER. Deriving a final water-
effect ratio using option 3.with the use of simulated downstream
water for a situation where this simulation appropriately
represents a "site", is a fully acceptable approach for deriving
a water-effect ratio for use in determining a site-specific
criterion, although it will generally provide a lower water-
effect ratio than the other 2 options.
As indicated in the introduction to this guidance, the
determination of a water-effect ratio may require substantial
resources. A discharger should consider cost-effective,
preliminary measures described in this guidance (e.g., use of
"clean" sampling and chemical analytical techniques or in non-NTR
States, a recalculated criterion) to determine if an indicator .
species site-specific criterion is really needed. It may be that
an appropriate site-specific criterion is actually being
attained. In many instances, use of these other measures may
eliminate the need for deriving final water-effect ratios. The
methods described in this interim guidance should be sufficient
to develop site-specific criteria that resolve concerns of
dischargers when there appears to be no instream toxicity from a
metal but, where (a) a discharge appears to exceed existing or
proposed water quality-based permit limits, or (b) an instream
concentration appears to exceed an existing or proposed water
quality criterion.
This guidance describes 2 different methods for determining
water-effect ratios. Method 1 has 3 options each of which may
only require 3 sampling periods. However options 1 and 2 may be
expanded and require a much greater effort. While this position
statement has discussed the simplest, least expensive option for
method 1 (the single discharge to a stream) to illustrate that
site specific criteria are feasible even when only small
dischargers are affected, water-effect ratios may be calculated
using any of the other options described in the guidance if the
State/discharger believe that there is reason to expect that a
more accurate site-specific criterion will result from the
increased cost and complexity inherent in conducting the
v
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additional tests and analyzing the results. Situations where
this could be the case include, for example, where seasonal
effects in receiving water quality or in discharge quality need
to be assessed.
In addition, EPA will consider other scientifically defensible
approaches in developing final water-effect ratios as authorized
in 40 CFR 131.11. However, EPA strongly recommends that before a
State/discharger implements any approach other than one described
in this interim guidance, discussions be held with appropriate
EPA regional offices and Office of Research and Development's
scientists before actual testing begins. These discussions would
be to ensure that time and resources are not wasted on
scientifically and technically unacceptable approaches. It
remains EPA's responsibility to make final decisions on the
scientific and technical validity of alternative approaches to
developing site-specific water quality criteria.
EPA is fully cognizant of the continuing debate between what
constitutes guidance and what is a regulatory requirement.
Developing site-specific criteria is a State regulatory option.
Using the methodology correctly as described in this guidance
assures the State that EPA will accept the result. Other
approaches are possible and logically should be discussed with
EPA prior to implementation.
The Office of Science and Technology believes that this
interim guidance advances the science of determining site-
specific criteria and provides policy guidance that States and
EPA can use in this complex area. It reflects the scientific
advances in the past 10 years and the experience gained from
dealing with these issues in real world situations. This
guidance will help improve implementation of water quality
standards and be the basis for future progress.
Tudor T. Davies, Director
Office of Science And Technology
Office of Water
VI
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CONTENTS
Page
Notices ii
Foreword iii
Office of Science and Technology Position Statement iv
Appendices viii
Figures ix
Acknowledgments x
Executive Summary xi
Abbreviations xiii
Glossary xiv
Preface xvi
Introduction 1
Method 1 17
A. Experimental Design 17
B. Background Information and Initial Decisions 44
C. Selecting Primary and Secondary Tests 45
D. Acquiring and Acclimating Test Organisms 47
E. Collecting and Handling Upstream Water and Effluent . . 48
F. Laboratory Dilution Water 49
G. Conducting Tests . 50
H. Chemical and Other Measurements 55
I. Calculating and Interpreting the Results 57
J. Reporting the Results 62
Method 2 65
References 76
VI1
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A.
B.
C.
D.
E.
F.
G.
H.
I.
APPENDICES
Comparison of WERs Determined Using Upstream and
Downstream Water
Pacre
The Recalculation Procedure
79
90
Guidance Concerning the Use of "Clean Techniques" and
QA/QC when Measuring Trace Metals 98
Relationships between WERs and the Chemistry and
Toxicology of Metals 109
U.S. EPA Aquatic Life Criteria Documents for Metals .
134
Considerations Concerning Multiple-Metal, Multiple-
Discharge, and Special Flowing-Water Situations 135
Additivity and the Two Components of a WER Determined
Using Downstream Water 139
Special Considerations Concerning the Determination
of WERs with Saltwater Species 145
Suggested Toxicity Tests for Determining WERs
for Metals
147
J. Recommended Salts of Metals ..... 153
vxn
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FIGURES
Page
1. Four Ways to Derive a Permit Limit 16
2. Calculating an Adjusted Geometric Mean 71
3 . An Example Derivation of a FWER 72
4. Reducing the Impact of Experimental Variation 73
5. Calculating an LC50 (or EC50) by Interpolation 74
6. Calculating a Time-Weighted Average 75
Bl. An Example of the Deletion Process Using Three Phyla . . 97
Dl. A Scheme for Classifying Forms of Metal in Water .... Ill
D2. An Example of the Empirical Extrapolation Process .... 125
D3. The Internal Consistency of the Two Approaches 126
D4. The Application of the Two Approaches 128
D5. A Generalized Complexation Curve 131
D6. A Generalized Precipitation Curve 132
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ACKNOWLEDGMENTS
This document was written by:
Charles E. Stephan, U.S. EPA, ORD, Environmental Research
Laboratory, Duluth, MN.
William H. Peltier, U.S. EPA, Region IV, Environmental
Services Division, Athens, GA.
David J. Hansen, U.S. EPA, ORD, Environmental Research
Laboratory, Narragansett, RI.
Charles G. Delos, U.S. EPA, Office of Water, Health
and Ecological Criteria Division, Washington, DC.
Gary A. Chapman, U.S. EPA, ORD, Environmental Research
Laboratory (Narragansett), Pacific Ecosystems Branch,
Newport, OR.
The authors thank all the people who participated in the open
discussion of the experimental determination of water-effect
ratios on Tuesday evening, January 26, 1993 in Annapolis, MD.
Special thanks go to Herb Allen, Bill Beckwith, Ken Bruland, Lee
Dunbar, Russ Erickson, and Carlton Hunt for their technical input
on this project, although none of them necessarily agree with
everything in this document. Comments by Kent Ballentine, Karen
Sourdine, Mark Hicks, Suzanne Lussier, Nelson Thomas, Bob Spehar,
Fritz Wagener, Robb Wood, and Phil Woods on various drafts, or
portions of drafts, were also very helpful, as were discussions
with several other individuals.
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EXECUTIVE SUMMARY
A variety of physical and chemical characteristics of both the
water and the metal can influence the toxicity of a metal to
aquatic organisms in a surface water. When a site-specific
aquatic life criterion is derived for a metal, an adjustment
procedure based on the toxicological determination of a water-
effect ratio (WER) may be used to account for a difference
between the toxicity of the metal in laboratory dilution water
and its toxicity in the water at the site. If there is a
difference in toxicity and it is not taken into account, the
aquatic life criterion for the body of water will be more or less
protective than intended by EPA's Guidelines for Deriving
Numerical National Water Quality Criteria for the Protection of
Aquatic Organisms and Their Uses. After a WER is determined for
a site, a site-specific aquatic life criterion can be calculated
by multiplying an appropriate national, state, or recalculated
criterion by the WER. Most WERs are expected to be equal to or
greater than 1.0, but some might be less than 1.0. Because most
aquatic life criteria consist of two numbers, i.e., a Criterion
Maximum Concentration (CMC) and a Criterion Continuous
Concentration (CCC), either a cmcWER or a cccWER or both might be
needed for a site. The cmcWER and the cccWER cannot be assumed
to be equal, but it is not always necessary to determine both.
In order to determine a WER, side-by-side toxicity tests are
performed to measure the toxicity of the metal in two dilution
waters. One of the waters has to be a water that- would be
acceptable for use in laboratory toxicity tests conducted for the
derivation of national water quality criteria for aquatic life.
In most situations, the second dilution water will be a simulated
downstream water that is prepared by mixing upstream water and
effluent in an appropriate ratio; in other situations, the second
dilution water will be a sample of the actual site water to which
the site-specific criterion is to apply. The WER is calculated
by dividing the endpoint obtained in the site water by the
endpoint obtained in the laboratory dilution water. A WER should
be determined using a toxicity test whose endpoint is close to,
but not lower than, the CMC and/or CCC that is to be adjusted.
A total recoverable WER can be determined if the metal in both of
the side-by-side toxicity tests is analyzed using the total
recoverable measurement, and a dissolved WER can be determined if
the metal is analyzed in both tests using the dissolved
measurement. Thus four WERs can be determined:
Total recoverable cmcWER.
Total recoverable cccWER.
Dissolved cmcWER.
Dissolved cccWER.
A total recoverable WER is used to calculate a total recoverable
site-specific criterion from a total recoverable national, state,
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or recalculated aquatic life criterion, whereas a dissolved WER
is used to calculate a dissolved site-specific criterion from a
dissolved criterion. WERs are determined individually for each
metal at each site; WERs cannot be extrapolated from one metal to
another, one effluent to another, or one site water to another.
Because determining a WER requires substantial resources, the
desirability of obtaining a WER should be carefully evaluated:
1. Determine whether use of "clean techniques" for collecting,
handling, storing, preparing, and analyzing samples will
eliminate the reason for considering determination of a WER,
because existing data concerning concentrations of metals in
effluents and surface waters might be erroneously high.
2. Evaluate the potential for reducing the discharge of the
metal.
3. Investigate possible constraints on the permit limits, such as
antibacksliding and antidegradation requirements and human
health and wildlife criteria.
4. Consider use of the Recalculation Procedure.
5. Evaluate the cost-effectiveness of determining a WER.
If the determination of a WER is desirable, a detailed workplan
for should be submitted to the appropriate regulatory authority
(and possibly to the Water Management Division of the EPA
Regional Office) for comment. After the workplan is completed,
the initial phase should be implemented, the data should be
evaluated, and the workplan should be revised if appropriate.
Two methods are used to determine WERs. Method 1, which is used
to determine cccWERs that apply near plumes and to determine all
cmcWERs, uses data concerning three or more distinctly separate
sampling events. It is best if the sampling events occur during
both low-flow and higher-flow periods. When sampling does not
occur during both low and higher flows, the site-specific
criterion is derived in a more conservative manner due to greater
uncertainty. For each sampling event, a WER is determined using
a selected toxicity test; for at least one of the sampling
events, a confirmatory WER is determined using a different test.
Method 2, which is used to determine a cccWER for a large body of
water outside the vicinities of plumes, requires substantial
site-specific planning and more resources than Method 1. WERs
are determined using samples of actual site water obtained at
various times, locations, and depths to identify the range of
WERs in the body of water. The WERs are used to determine how
many site-specific CCCs should be derived for the body of water
and what the one or more CCCs should be.
The guidance contained herein replaces previous agency guidance
concerning (a) the determination of WERs for use in the
derivation of site-specific aquatic life criteria for metals and
(b) the Recalculation Procedure. This guidance is designed to
apply to metals, but the principles apply to most pollutants.
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ABBREVIATIONS
ACR: Acute-Chronic Ratio
CCC: Criterion Continuous Concentration
CMC: Criterion Maximum Concentration
CRM: Certified Reference Material
FAV: Final Acute Value
FCV: Final Chronic Value
FW: Freshwater
FWER: Final Water-Effect Ratio
GMAV: Genus Mean Acute Value
HCME: Highest Concentration of the Metal in the Effluent
MDR: Minimum Data Requirement
NTR: National Toxics Rule
QA/QC: Quality Assurance/Quality Control
SMAV: Species Mean Acute Value
SW: Saltwater
TDS: Total Dissolved Solids
TIE: Toxicity Identification Evaluation
TMDL: Total Maximum Daily Load
TOG: Total Organic Carbon
TRE: Toxicity Reduction Evaluation
TSD: Technical Support Document
TSS: Total Suspended Solids
WER: Water-Effect Ratio
WET: Whole Effluent Toxicity
WLA: Wasteload Allocation
Xlll
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GLOSSARY
Acute-chronic ratio - an appropriate measure of the acute
toxicity of a material divided by an appropriate
measure of the chronic toxicity of the same material
under the same conditions.
Appropriate regulatory authority - Usually the State water
pollution control agency, even for States under the National
Toxics Rule; if, however, a State were to waive its section
401 authority, the Water Management Division of the EPA
Regional Office would become the appropriate regulatory
authority.
Clean techniques - a set of procedures designed to prevent
contamination of samples so that concentrations of
trace metals can be measured accurately and precisely.
Critical species - a species that is commercially or
recreationally important at the site, a species that exists
at the site and is listed as threatened or endangered under
section 4 of the Endangered Species Act, or a species for
which there is evidence that the loss of the species from
the site is likely to cause an unacceptable impact on a
commercially or recreationally important species, a
threatened or endangered species, the abundances of a
variety of other species, or the structure or function of
the community.
Design flow - the flow used for steady-state wasteload
allocation modeling.
Dissolved metal - defined here as "metal that passes through
either a 0.45-/*m or a 0.40-^tm membrane filter".
Endpoint - the concentration of test material that is expected to
cause a specified amount of adverse effect.
Final Water-Effect Ratio - the WER that is used in the
calculation of a site-specific aquatic life criterion.
Flow-through test - a test in which test solutions flow into
the test chambers either intermittently (every few
minutes) or continuously and the excess flows out.
Labile metal - metal that is in water and will readily
convert from one form to another when in a
nonequilibrium condition.
Particulate metal - metal that is measured by the total
recoverable method but not by the dissolved method.
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Primary test - the toxicity test used in the determination
of a Final Water-Effect Ratio (FWER); the specification
of the test includes the test species, the life stage
of the species, the duration of the test, and the
adverse effect on which the endpoint is based.
Refractory metal - metal that is in water and will not
readily convert from one form to another when in a
nonequilibrium condition, i.e., metal that is in water
and is not labile.
Renewal test - a test in which either the test solution in a
test chamber is renewed at least once during the test
or the test organisms are transferred into a new test
solution of the same composition at least once during
the test.
Secondary test - a toxicity test that is usually conducted
along with the primary test only once to test the
assumptions that, within experimental variation, (a)
similar WERs will be obtained using tests that have
similar sensitivities to the test material, and (b)
tests that are less sensitive to the test material will
usually give WERs that are closer to 1.
Simulated downstream water - a site water prepared by mixing
effluent and upstream water in a known ratio.
Site-specific aquatic life criterion - a water quality
criterion for aquatic life that has been derived to be
specifically appropriate to the water quality
characteristics and/or species composition at a
particular location.
Site water - upstream water, actual downstream water, or
simulated downstream water in which a toxicity test is
conducted side-by-side with the same toxicity test in a
laboratory dilution water to determine a WER.
Static test - a test in which the solution and organisms
that are in a test chamber at the beginning of the test
remain in the chamber until the end of the test.
Total recoverable metal - metal that is in aqueous solution
after the sample is appropriately acidified and
digested and insoluble material is separated.
Water-effect ratio - an appropriate measure of the toxicity
of a material obtained in a site water divided by the
same measure of the toxicity of the same material
obtained simultaneously in a laboratory dilution water.
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PREFACE
Several issues need consideration when guidance such as this is
written:
1. Degrees of importance: Procedures and methods are series of
instructions, but some of the instructions are more important
than others. Some instructions are so important that, if they
are not followed, the results will be questionable or
unacceptable; other instructions are less important, but
definitely desirable. Possibly the best way to express
various degrees of importance is the approach described in
several ASTM Standards, such as in section 3.6 of Standard
E729 (ASTM 1993a), which is modified here to apply to WERs:
The words "must", "should", "may", "can", and "might" have
specific meanings in this document. "Must" is used to
express an instruction that is to be followed, unless a
site-specific consideration requires a deviation, and is
used only in connection with instructions that directly
relate to the validity of toxicity tests, WERs, FWERs, and
the Recalculation Procedure. "Should" is used to state
instructions that are recommended and are to be followed if
reasonably possible. Deviation from one "should" will not
invalidate a WER, but deviation from several probably will.
Terms such as "is desirable", "is often desirable", and
"might be desirable" are used in connection with less
important instructions. "May" is used to mean "is (are)
allowed to", "can" is used to mean "is (are) able to", and
"might" is used to mean "could possibly". Thus the classic
distinction between "may" and "can" is preserved, and
"might" is not used as a synonym for either "may" or "can".
This does not eliminate all problems concerning the degree of
importance, however. For example, a small deviation from a
"must" might not invalidate a WER, whereas a large deviation
would. (Each "must" and "must not" is in bold print for
convenience, not for emphasis, in this document.)
2. Educational and explanatory material; Many people have asked
for much detail in this document to ensure that as many WERs
as possible are determined in an acceptable manner. In
addition, some people want justifications for each detail.
Much of the detail that is desired by some people is based on
"best professional judgment", which is rarely considered an
acceptable justification by people who disagree with a
specified detail. Even if details are taken from an EPA
method or an ASTM standard, they were often included in those
documents on the basis of best professional judgment. In
contrast, some people want detailed methodology presented
without explanatory material. It was decided to include as
much detail as is feasible, and to provide rationale and
explanation for major items.
xv i
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3. Alternatives: When more than one alternative is both
scientifically sound and appropriately protective, it seems
reasonable to present the alternatives rather than presenting
the one that is considered best. The reader can then select
one based on cost-effectiveness, personal preference, details
of the particular situation, and perceived advantages and
disadvantages.
4. Separation of "science", "best professional judgment" and
"regulatory decisions": These can never be completely
separated in this kind of document; for example, if data are
analyzed for a statistically significant difference, the
selection of alpha is an important decision, but a rationale
for its selection is rarely presented, probably because the
selection is not a scientific decision. In this document, an
attempt has been made to focus on good science, best
professional judgment, and presentation of the rationale; when
possible, these are separated from "regulatory decisions"
concerning margin of safety, level of protection, beneficial
use, regulatory convenience, and the goal of zero discharge.
Some "regulatory decisions" relating to implementation,
however, should be integrated with, not separated from,
"science" because the two ought to be carefully considered
together wherever science has implications for implementation.
5. Best professional judgment: Much of the guidance contained
herein is qualitative rather than quantitative, and much
judgment will usually be required to derive a site-specific
water quality criterion for aquatic life. In addition,
although this version of the guidance for determining and
using WERs attempts to cover all major questions that have
arisen during use of the previous version and during
preparation of this version, it undoubtedly does not cover all
situations, questions, and extenuating circumstances that
might arise in the future. All necessary decisions should be
based on both a thorough knowledge of aquatic toxicology and
an understanding of this guidance; each decision should be
consistent with the spirit of this guidance, which is to make
best use of "good science" to derive the most appropriate
site-specific criteria. This guidance should be modified
whenever sound scientific evidence indicates that a site-
specific criterion produced using this guidance will probably
substantially underprotect or overprotect the aquatic life at
the site of concern. Derivation of site-specific criteria for
aquatic life is a complex process and requires knowledge in
many areas of aquatic toxicology; any deviation from this
guidance should be carefully considered to ensure that it is
consistent with other parts of this guidance and with "good
science".
6. Personal bias: Bias can never be eliminated, and some
decisions are at the fine line between "bias" and "best
xvii
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professional judgment". The possibility of bias can be
eliminated only by adoption of an extreme position such as "no
regulation" or "no discharge". One way to deal with bias is to
have decisions made by a team of knowledgeable people.
7. Teamwork; The determination of a WER should be a cooperative
team effort beginning with the completion of the initial
workplan, interpretation of initial data, revision of the
workplan, etc. The interaction of a variety of knowledgeable,
reasonable people will help obtain the best results for the
expenditure of the fewest resources. Members of the team
should acknowledge their biases so that the team can make best
use of the available information, taking into account its
relevancy to the immediate situation and its quality.
xvi 11
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INTRODUCTION
National aquatic life criteria for metals are intended to protect
the aquatic life in almost all surface waters of the United
States (U.S. EPA 1985). This level of protection is accomplished
in two ways. First, the national dataset is required to contain
aquatic species that have been found to be sensitive to a variety
of pollutants. Second, the dilution water and the metal salt
used in the toxicity tests are required to have physical and
chemical characteristics that ensure that the metal is at least
as toxic in the tests as it is in nearly all surface waters. For
example, the dilution water is to be low in suspended solids and
in organic carbon, and some forms of metal (e.g., insoluble metal
and metal bound by organic complexing agents) cannot be used as
the test material. (The term "metal" is used herein to include
both "metals" and "metalloids".)
Alternatively, a national aquatic life criterion might not
adequately protect the aquatic life at some sites. An untested
species that is important at a site might be more sensitive than
any of the tested species. Also, the metal might be more toxic
in site water than in laboratory dilution water because, for
example, the site water has a lower pH and/or hardness than most
laboratory waters. Thus although a national aquatic life
criterion is intended to be lower than necessary for most sites,
a national criterion might not adequately protect the aquatic
life at some sites.
Because a national aquatic life criterion might be more or less
protective than intended for the aquatic life in most bodies of
water, the U.S. EPA provided guidance (U.S. EPA 1983a,1984)
concerning three procedures that may be used to derive a site-
specific criterion:
1. The Recalculation Procedure is intended to take into account
relevant' differences between the sensitivities of the aquatic
organisms in the national dataset and the sensitivities of
organisms that occur at the site.
2. The Indicator Species Procedure provides for the use of a
water-effect ratio (WER) that is intended to take into account
relevant differences between the toxicity of the metal in
laboratory dilution water and in site water.
3. The Resident Species Procedure is intended to take into
account both kinds of differences simultaneously.
A site-specific criterion is intended to come closer than the
national criterion to providing the intended level of protection
to the aquatic life at the site, usually by taking into account
the biological and/or chemical conditions (i.e., the species
composition and/or water quality characteristics) at the site.
The fact that the U.S. EPA has made these procedures available
should not be interpreted as implying that the agency advocates
that states derive site-specific criteria before setting state
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standards. Also, derivation of a site-specific criterion does
not change the intended level of protection of the aquatic life
at the site. Because a WER is expected to appropriately take
into account (a) the site-specific toxicity of the metal, and (b)
synergism, antagonism, and additivity with other constituents of
the site water, using a WER is more likely to provide the
intended level of protection than not using a WER.
Although guidance concerning site-specific criteria has been
available since 1983 (U.S. EPA 1983a,1984), interest has
increased in recent years as states have devoted more attention
to chemical-specific water quality criteria for aquatic life. In
addition, interest in water-effect ratios (WERs) increased when
the "Interim Guidance" concerning metals (U.S. EPA 1992) made a
fundamental change in the way that WERs are experimentally
determined (see Appendix A), because the change is expected to
substantially increase the magnitude of many WERs. Interest was
further focused on WERs when they were integrated into some of
the aquatic life criteria for metals that were promulgated by the
National Toxics Rule (57 FR 60848, December 22, 1992). The
newest guidance issued by the U.S. EPA (Prothro 1993) concerning
aquatic life criteria for metals affected the determination and
use of WERs only insofar as it affected the use of total
recoverable and dissolved criteria.
The early guidance concerning WERs (U.S. EPA 1983a,1984)
contained few details and needs revision, especially to take into
account newer guidance concerning metals (U.S. EPA 1992; Prothro
1993). The guidance presented herein supersedes all guidance
concerning WERs and the Indicator Species Procedure given in
Chapter 4 of the Water Quality Standards Handbook (U.S. EPA
1983a) and in U.S. EPA (1984). All guidance presented in U.S.
EPA (1992) is superseded by that presented by Prothro (1993) and
by this document. Metals are specifically addressed herein
because of the National Toxics Rule (NTR) and because of current
interest in aquatic life criteria for metals; silthough most of
this guidance also applies to other pollutants, some obviously
applies only to metals.
Even though this document was prepared mainly because of the NTR,
the guidance contained herein concerning WERs is likely to have
impact beyond its use with the NTR. Therefore, it is appropriate
to also present new guidance concerning the Recalculation
Procedure (see Appendix B) because the previous guidance (U.S.
EPA 1983a,1984) concerning this procedure also contained few
details and needs revision. The NTR does not allow use of the
Recalculation Procedure in jurisdictions subject to the NTR.
The previous guidance concerning site-specific procedures did not
allow the Recalculation Procedure and the WER procedure to be
used together in the derivation of a site-specific aquatic life
criterion; the only way to take into account both species
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composition and water quality characteristics in the
determination of a site-specific criterion was to use the
Resident Species Procedure. A specific change contained herein
is that, except in jurisdictions that are subject to the NTR, the
Recalculation Procedure and the WER Procedure may now be used
together. Additional reasons for addressing both the
Recalculation Procedure and the WER Procedure in this document
are that both procedures are based directly on the guidelines for
deriving national aquatic life criteria (U.S. EPA 1985) and, when
the two are used together, use of the Recalculation Procedure has
specific implications concerning the determination of the WER.
This guidance is intended to produce WERs that may be used to
derive site-specific aquatic life criteria for metals from most
national and state aquatic life criteria that were derived from
laboratory toxicity data. Except in jurisdictions that are
subject to the NTR, the WERs may also be used with site-specific
aquatic life criteria that are derived for metals using the
Recalculation Procedure described in Appendix B. WERs obtained
using the methods described herein should not be used to adjust
aquatic life criteria that were derived for metals in other ways.
For example, because they are designed to be applied to criteria
derived on the basis of laboratory toxicity tests, WERs
determined using the methods described herein cannot be used to
adjust the residue-based mercury Criterion Continuous
Concentration (CCC) or the field-based selenium freshwater
criterion. For the purposes of the NTR, WERs may be used with
the aquatic life criteria for arsenic, cadmium, chromium(III),
chromium(VI), copper, lead, nickel, silver, and zinc and with the
Criterion Maximum Concentration (CMC) for mercury. WERs may also
be used with saltwater criteria for selenium.
The concept of a WER is rather simple:
Two side-by-side toxicity tests are conducted - one test using
laboratory dilution water and the other using site water. The
endpoint obtained using site water is divided by the endpoint
obtained using laboratory dilution water. The quotient is the
WER, which is multiplied times the national, state, or
recalculated aquatic life criterion to calculate the site-
specific criterion.
Although the concept is simple, the determination and use of WERs
involves many considerations.
The primary purposes of this document are to:
1. Identify steps that should be taken before the determination
of a WER is begun.
2. Describe the methods recommended by the U.S. EPA for the
determination of WERs.
3. Address some issues concerning the use of WERs.
4. Present new guidance concerning the Recalculation Procedure.
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Before Determining a WER
Because a national criterion is intended to protect aquatic life
in almost all bodies of water and because a WER is intended to
account for a difference between the toxicity of a metal in a
laboratory dilution water and its toxicity in a site water,
dischargers who want higher permit limits than those derived on
the basis of an existing aquatic life criterion will probably
consider determining a WER. Use of a WER should be considered
only as a last resort for at least three reasons:
a. Even though some WERs will be substantially greater than 1.0,
some will be about 1.0 and some will be less than 1.0.
b. The determination of a WER requires substantial resources.
c. There are other things that a discharger can do that might be
more cost-effective than determining a WER.
The two situations in which the determination of a WER might
appear attractive to dischargers are when (a) a discharge appears
to exceed existing or proposed water quality-based permit limits,
and (b) an instream concentration appears to exceed an existing
or proposed aquatic life criterion. Such situations result from
measurement of the concentration of a metal in an effluent or a
surface water. It would therefore seem reasonable to ensure that
such measurements were not subject to contamination. Usually it
is much easier to verify chemical measurements by using "clean
techniques" for collecting, handling, storing, preparing, and
analyzing samples, than to determine a WER. Clean techniques and
some related QA/QC considerations are discussed in Appendix C.
In addition to investigating the use of "clean techniques", other
steps that a discharger should take prior to beginning the
experimental determination of a WER include:
1. Evaluate the potential for reducing the discharge of the
metal.
2. Investigate such possible constraints on permit limits as
antibacksliding and antidegradation requirements and human
health and wildlife criteria.
3. Obtain assistance from an aquatic toxicologist who understands
the basics of WERs (see Appendix D), the U.S. EPA's national
aquatic life guidelines (U.S. EPA 1985), the guidance
presented by Prothro (1993), the national criteria document
for the metal(s) of concern (see Appendix E), the procedures
described by the U.S. EPA (1993a,b,c) for acute and chronic
toxicity tests on effluents and surface waters, and the
procedures described by ASTM (1993a,b,c,d,e) for acute and
chronic toxicity tests in laboratory dilution water.
4. Develop an initial definition of the site to which the site-
specific criterion is to apply.
5. Consider use of the Recalculation Procedure (see Appendix B).
6. Evaluate the cost-effectiveness of the determination of a WER.
Comparative toxicity tests provide the most useful data, but
chemical analysis of the downstream water might be helpful
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because the following are often true for some metals:
a. The lower the percent of the total recoverable metal in the
downstream water that is dissolved, the higher the WER.
b. The higher the concentration of total organic carbon (TOG)
and/or total suspended solids (TSS), the higher the WER.
It is also true that the higher the concentration of nontoxic
dissolved metal, the higher the WER. Although some chemical
analyses might provide useful information concerning the
toxicities of some metals in water, at the present only
toxicity tests can accurately reflect the toxicities of
different forms of a metal (see Appendix D).
7. Submit a workplan for the experimental determination of the
WER to the appropriate regulatory authority (and possibly to
the Water Management Division of the EPA Regional Office) for
comment. The workplan should include detailed descriptions of
the site; existing criterion and standard; design flows; site
water; effluent; sampling plan; procedures that will be used
for collecting, handling, and analyzing samples of site water
and effluent; primary and secondary toxicity tests; quality
assurance/quality control (QA/QC) procedures; Standard
Operating Procedures (SOPs) ,- and data interpretation.
After the workplan is completed, the initial phase should be
implemented; then the data obtained should be evaluated, and the
workplan should be revised if appropriate. Developing and
modifying the workplan and analyzing and interpreting the data
should be a cooperative effort by a team of knowledgeable people.
Two Kinds of WERs
Most aquatic life criteria contain both a CMC and a CCC, and it
is usually possible to determine both a cmcWER and a cccWER. The
two WERs cannot be assumed to be equal because the magnitude of a
WER will probably depend on the sensitivity of the toxicity test
used and on the percent effluent in the site water (see Appendix
D), both of which can depend on which WER is to be determined.
In some cases, it is expected that a larger WER can be applied to
the CCC than to the CMC, and so it would be environmentally
conservative to apply cmcWERs to CCCs. In such cases it is
possible to determine a cmcWER and apply it to both the CMC and
the CCC in order to derive a site-specific CMC, a site-specific
CCC, and new permit limits. If these new permit limits are
controlled by the new site-specific CCC, a cccWER could be
determined using a more sensitive test, possibly raising the
site-specific CCC and the permit limits again. A cccWER may, of
course, be determined whenever desired. Unless the experimental
variation is increased, use of a cccWER will usually improve the
accuracy of the resulting site-specific CCC.
In some cases, a larger WER cannot be applied to the CCC than to
the CMC and so it might not be environmentally conservative to
apply a cmcWER to a CCC (see section A.4 of Method 1).
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Steady-state and Dynamic Models
Some of the guidance contained herein specifically applies to
situations in which the permit limits were calculated using
steady-state modeling; in particular, some samples are to be
obtained when the actual stream flow is close to the design flow.
If permit limits were calculated using dynamic modeling, the
guidance will have to be modified, but it is unclear at present
what modifications are most appropriate. For example, it might
be useful to determine whether the magnitude of the WER is
related to the flow of the upstream water and/or the effluent.
Two Methods
Two methods are used to determine WERs. Method 1 will probably
be used to determine all cmcWERs and most cccWERs because it can
be applied to situations that are in the vicinities of plumes.
Because WERs are likely to depend on the concentration of
effluent in the water and because the percent effluent in a water
sample obtained in the immediate vicinity of a plume is unknown,
simulated downstream water is used so that the percent effluent
in the sample is known. For example, if a sample that was
supposed to represent a complete-mix situation was accidently
taken in the plume upstream of complete mix, the sample would
probably have a higher percent effluent and a higher WER than a
sample taken downstream of complete mix; use of the higher WER to
derive a site-specific criterion for the complete-mix situation
would result in underprotection. If the sample were accidently
taken upstream of complete mix but outside the plume,
overprotection would probably result.
Method 1 will probably be used to determine all cmcWERs and most
cccWERs in flowing fresh waters, such as rivers and streams.
Method 1 is intended to apply not only to ordinary rivers and
streams but also to streams that some people might consider
extraordinary, such as streams whose design flows are zero and
streams that some state and/or federal agencies refer to as
11 effluent-dependent", " habit at-creat ing ", or "effluent-
dominated" . Method 1 is also used to determine cmcWERs in such
large sites as oceans and large lakes, reservoirs, and estuaries
(see Appendix F).
Method 2 is used to determine WERs that apply outside the area of
plumes in large bodies of water. Such WERs will be cccWERs and
will be determined using samples of actual site water obtained at
various times, locations, and depths in order to identify the
range of WERs that apply to the body of water. These
experimentally determined WERs are then used to decide how many
site-specific criteria should be derived for the body of water
and what the criterion (or criteria) should be. Method 2
requires substantially more resources than Method 1.
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The complexity of each method increases when the number of metals
and/or the number of discharges is two or more:
a. The simplest situation is when a WER is to be determined for
only one metal and only one discharge has permit limits for
that metal. (This is the single-metal single-discharge
situation.)
b. A more complex situation is when a WER is to be determined for
only one metal, but more than one discharge has permit limits
for that metal. (This is the single-metal multiple-discharge
situation.)
c. An even more complex situation is when WERs are to be
determined for more than one metal, but only one discharge has
permit limits for any of the metals. (This is the multiple-
metal single-discharge situation.)
d. The most complex situation is when WERs are to be determined
for more than one metal and more than one discharge has permit
limits for some or all of the metals. (This is the multiple-
metal multiple-discharge situation.)
WERs need to be determined for each metal at each site because
extrapolation of a WER from one metal to another, one effluent to
another, or one surface water to another is too uncertain.
Both methods work well in multiple-metal situations, but special
tests or additional tests will be-necessary to show that the
resulting combination of site-specific criteria will not be too
toxic. Method 2 is better suited to multiple-discharge
situations than is Method 1. Appendix F provides additional
guidance concerning multiple-metal and multiple-discharge
situations, but it does not discuss allocation of waste loads,
which is performed when a wasteload allocation (WLA) or a total
maximum daily load (TMDL) is developed (U.S. EPA 1991a).
Two Analytical Measurements
A total recoverable WER can be determined if the metal in both of
the side-by-side toxicity tests is analyzed using the total
recoverable measurement; similarly, a dissolved WER can be
determined if the metal in both tests is analyzed using the
dissolved measurement. A total recoverable WER is used to
calculate a total recoverable site-specific criterion from an
aquatic life criterion that is expressed using the total
recoverable measurement, whereas a dissolved WER is used to
calculate a dissolved site-specific criterion from a criterion
that is expressed in terms of the dissolved measurement. Figure
1 illustrates the relationships between total recoverable and
dissolved criteria, WERs, and the Recalculation Procedure.
Both Method 1 and Method 2 can be used to determine a total
recoverable WER and/or a dissolved WER. The only difference in
the experimental, procedure is whether the WER is based on
measurements of total recoverable metal or dissolved metal in the
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test solutions. Both total recoverable and dissolved
measurements are to be performed for all tests to help judge the
quality of the tests, to provide a check on the analytical
chemistry, and to help understand the results; performing both
measurements also increases the alternatives available for use of
the results. For example, a dissolved WER that is not useful
with a total recoverable criterion might be useful in the future
if a dissolved criterion becomes available. Also, as explained
in Appendix D, except for experimental variation, use of a total
recoverable WER with a total recoverable criterion should produce
the same total recoverable permit limits as use of a dissolved
WER with a dissolved criterion; the internal consistency of the
approaches and the data can be evaluated if both total
recoverable and dissolved criteria and WERs are determined. It
is expected that in many situations total recoverable WERs will
be larger and more variable than dissolved WERs.
The Quality of the Toxicitv Tests
Traditionally, for practical reasons, the requirements concerning
such aspects as acclimation of test organisms to test temperature
and dilution water have not been as stringent for toxicity tests
on surface waters and effluents as for tests using laboratory
dilution water. Because a WER is a ratio calculated from the
results of side-by-side tests, it might seem that acclimation is
not important for a WER as long as the organisms and conditions
are identical in the two tests. Because WERs are used to adjust
aquatic life criteria that are derived from results of laboratory
tests, the tests conducted in laboratory dilution water for the
determination of WERs should be conducted in the same way as the
laboratory toxicity tests used in the derivation of aquatic life
criteria. In the WER process, the tests in laboratory dilution
water provide the vital link between national criteria and site-
specific criteria, and so it is important to compare at least
some results obtained in the laboratory dilution water with
results obtained in at least one other laboratory.
Three important principles for making decisions concerning the
methodology for the side-by-side tests are:
1. The tests using laboratory dilution water should be conducted
so that the results would be acceptable for use in the
derivation of national criteria.
2. As much as is feasible, the tests using site water should be
conducted using the same procedures as the tests using the
laboratory dilution water.
3. All tests should follow any special requirements that are
necessary because the results are to be used to calculate a
WER. Some such special requirements are imposed because the
criterion for a rather complex situation is being changed
based on few data, so more assurance is required that the data
are high quality.
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The most important special requirement is that the concentrations
of the metal are to be measured using both the total recoverable
and dissolved methods in all toxicity tests used for the
determination of a WER. This requirement is necessary because
half of the tests conducted for the determination of WERs use a
site water in which the concentration of metal probably is not
negligible. Because it is likely that the concentration of metal
in the laboratory dilution water is negligible, assuming that the
concentration in both waters is negligible and basing WERs on the
amount of metal added would produce an unnecessarily low value
for the WER. In addition, WERs are based on too few data to
assume that nominal concentrations are accurate. Nominal
concentrations obviously cannot be used if a dissolved WER is to
be determined. Measured dissolved concentrations at the
beginning and end of the test are used to judge the acceptability
of the test, and it is certainly reasonable to measure the total
recoverable concentration when the dissolved concentration is
measured. Further, measuring the concentrations might lead to an
interpretation of the results that allows a substantially better
use of the WERs.
Conditions for Determining a WER
The appropriate regulatory authority might recommend that one or
more conditions be met when a WER is determined in order to
reduce the possibility of having to determine a new WER later:
1. Requirements that are in the existing permit concerning WET
testing, Toxicity Identification Evaluation (TIE), and/or
Toxicity Reduction Evaluation (TRE) (U.S. EPA 1991a).
2. Implementation of pollution prevention efforts, such as
pretreatment, waste minimization, and source reduction.
3. A demonstration that applicable technology-based requirements
are being met.
If one or more of these is not satisfied when the WER is
determined and is implemented later, it is likely that a new WER
will have to be determined because of the possibility of a change
in the composition of the effluent.
Even if all recommended conditions are satisfied, determination
of a WER might not be possible if the effluent, upstream water,
and/or downstream water are toxic to the test organisms. In some
such cases, it might be possible to determine a WER, but
remediation of the toxicity is likely to be required anyway. It
is unlikely that a WER determined before remediation would be
considered acceptable for use after remediation. If it is
desired to determine a WER before remediation and the toxicity is
in the upstream water, it might be possible to use a laboratory
dilution water or a water from a clean tributary in place of the
upstream water; if a substitute water is used, its water quality
characteristics should be similar to those of the upstream water
(i.e., the pH should be within 0.2 pH units and the hardness,
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alkalinity, and concentrations of TSS and TOG should be within 10
% or 5 tng/L, whichever is greater, of those in the upstream
water). If the upstream water is chronically toxic, but not
acutely toxic, it might be possible to determine a cmcWER even if
a cccWER cannot be determined; a cmcWER might not be useful,
however, if the permit limits are controlled by the CCC; in such
a case, it would probably not be acceptable to assume that the
cmcWER is an environmentally conservative estimate of the cccWER.
If the WER is determined using downstream water and the toxicity
is due to the effluent, tests at lower concentrations of the
effluent might give an indication of the amount of remediation
needed.
Conditions for Using a WER
Besides requiring that the WER be valid, the appropriate
regulatory authority might consider imposing other conditions for
the approval of a site-specific criterion based on the WER:
1. Periodic reevaluation of the WER.
a. WERs determined in upstream water take into account
constituents contributed by point and nonpoint sources and
natural runoff; thus a WER should be reeveiluated whenever
newly implemented controls or other changes substantially
affect such factors as hardness, alkalinity, pH, suspended
solids, organic carbon, or other toxic materials.
b. Most WERs determined using downstream water are influenced
more by the effluent than the upstream water. Downstream
WERs should be reevaluated whenever newly implemented
controls or other changes might substantially impact the
effluent, i.e., might impact the forms and concentrations
of the metal, hardness, alkalinity, pH, suspended solids,
organic carbon, or other toxic materials. A special
concern is the possibility of a shift from discharge of
nontoxic metal to discharge of toxic metal such that the
concentration of the metal does not increase; analytical
chemistry might not detect the change but toxicity tests
would.
Even if no changes are known to have occurred, WERs should be
reevaluated periodically. (The NTR recommends that NPDES
permits include periodic determinations of WERs in the
monitoring requirements.) With advance planning, it should
usually be possible to perform such reevaluations under
conditions that are at least reasonably similar to those that
control the permit limits (e.g., either design-flow or high-
flow conditions) because there should be a reasonably long
period of time during which the reevaluation can be performed.
Periodic determination of WERs should be designed to answer
questions, not just generate data.
2. Increased chemical monitoring of the upstream water, effluent,
and/or downstream water, as appropriate, for water quality
characteristics that probably affect the toxicity of the metal
10
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(e.g., hardness, alkalinity, pH, TOG, and TSS) to determine
whether conditions change. The conditions at the times the
samples were obtained should be kept on record for reference.
The WER should be reevaluated whenever hardness, alkalinity, pH,
TOG, and/or TSS decrease below the values that existed when the
WERs were determined.
3. Periodic reevaluation of the environmental fate of the metal
in the effluent (see Appendix A).
4. WET testing.
5. Instream bioassessments.
Decisions concerning the possible imposition of such conditions
should take into account:
a. The ratio of the new and old criteria. The greater the
increase in the criterion, the more concern there should be
about (1) the fate of any nontoxic metal that contributes to
the WER and (2) changes in water quality that might occur
within the site. The imposition of one or more conditions
should be considered if the WER is used to raise the criterion
by, for example, a factor of two, and especially if it is
raised by a factor of five or more. The significance of the
magnitude of the ratio can be judged by comparison with the
acute-chronic ratio, the factor of two that is the ratio of
the FAV to the CMC, and the range of sensitivities of species
in the criteria document for the metal (see Appendix E).
b. The size of the site.
c. The size of the discharge.
d. The rate of downstream dilution.
e. Whether the CMC or the CCC controls the permit limits.
When WERs are determined using upstream water, conditions on the
use of a WER are more likely when the water contains an effluent
that increases the WER by adding TOG and/or TSS, because the WER
will be larger and any decrease in the discharge of such TOG
and/or TSS might decrease the WER and result in underprotection.
A WER determined using downstream water is likely to be larger
and quite dependent on the composition of the effluent; there
should be concern about whether a change in the effluent might
result in underprotection at some time in the future.
Implementation Considerations
In some situations a discharger might not want to or might not be
allowed to raise a criterion as much as could be justified by a
WER:
1. The maximum possible increase is not needed and raising the
criterion more than needed might greatly raise the cost if a
greater increase would require more tests and/or increase the
conditions imposed on approval of the site-specific criterion.
2. Such other constraints as antibacksliding or antidegradation
requirements or human health or wildlife criteria might limit
the amount of increase regardless of the magnitude of the WER.
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3. The permit limits might be limited by an aquatic life
criterion that applies outside the site. It is EPA policy
that permit limits cannot be so high that they inadequately
protect a portion of the same or a different body of water
that is outside the site; nothing contained herein changes
this policy in any way.
If no increase in the existing discharge is allowed, the only use
of a WER will be to determine whether an existing discharge needs
to be reduced. Thus a major use of WERs might be where
technology-based controls allow concentrations in surface waters
to exceed national, state, or recalculated aquatic life criteria.
In this case, it might only be necessary to determine that the
WER is greater than a particular value; it might not be necessary
to quantify the WER. When possible, it might be desirable to
show that the maximum WER is greater than the WER that will be
used in order to demonstrate that a margin of safety exists, but
again it might not be necessary to quantify the maximum WER.
In jurisdictions not subject to the NTR, WERs should be used to
derive site-specific criteria, not just to calculate permit
limits, because data obtained from ambient monitoring should be
interpreted by comparison with ambient criteria. (This is not a
problem in jurisdictions subject to the NTR because the NTR
defines the ambient criterion as "WER x the EPA criterion".) If
a WER is used to adjust permit limits without adjusting the
criterion, the permit limits would allow the criterion to be
exceeded. Thus the WER should be used to calculate a site-
specific criterion, which should then be used to calculate permit
limits. In some states, site-specific criteria can only be
adopted as revised criteria in a separate, independent water
quality standards review process. In other states, site-specific
criteria can be developed in conjunction with the NPDES
permitting process, as long as the adoption of a site-specific
criterion satisfies the pertinent water quality standards
procedural requirements (i.e., a public notice and a public
hearing). In either case, site-specific criteria are to be
adopted prior to NPDES permit issuance. Moreover, the EPA
Regional Administrator has authority to approve or disapprove all
new and revised site-specific criteria and to review NPDES
permits to verify compliance with the applicable water quality
criteria.
Other aspects of the use of WERs in connection with permit
limits, WIiAs, and TMDLs are outside the scope of this document.
The Technical Support Document (U.S. EPA 1991a) and Prothro
(1993) provide more information concerning implementation
procedures. Nothing contained herein should be interpreted as
changing the three-part approach that EPA uses to protect aquatic
life: (1) numeric chemical-specific water quality criteria for
individual pollutants, (2) whole effluent toxicity (WET) testing,
and (3) instream bioassessments.
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Even though there are similarities between WET testing and the
determination of WERs, there are important differences. For
example, WERs can be used to derive site-specific criteria for
individual pollutants, but WET testing cannot. The difference
between WET testing and the determination of WERs is less when
the toxicity tests used in the determination of the WER are ones
that are used in WET testing. If a WER is used to make a large
change in a criterion, additional WET testing and/or instream
bioassessments are likely to be recommended.
The Sample-Specific WER Approach
A major problem with the determination and use of aquatic life
criteria for metals is that no analytical measurement or
combination of measurements has yet been shown to explain the
toxicity of a metal to aquatic plants, invertebrates, amphibians,
and fishes over the relevant range of conditions in surface
waters (see Appendix D). It is not just that insufficient data
exist to justify a relationship; rather, existing data possibly
contradict some ideas that could possibly be very useful if true.
For example, the concentration of free metal ion could possibly
be a useful basis for expressing water quality criteria for
metals if it could be feasible and could be used in a way that
does not result in widespread underprotection of aquatic life.
Some available data, however, might contradict the idea that the
toxicity of copper to aquatic organisms is proportional to the
concentration or the activity of the cupric ion. Evaluating the
usefulness of any approach based on metal speciation is difficult
until it is known how many of the species of the metal are toxic,
what the relative toxicities are, whether they are additive (if
more than one is toxic), and the quantitative effects of the
factors that have major impacts on the bioavailability and/or
toxicity of the toxic species. Just as it is not easy to find a
useful quantitative relationship between the analytical chemistry
of metals and the toxicity of metals to aquatic life, it is also
not easy to find a qualitative relationship that can be used to
provide adequate protection for the aquatic life in almost all
bodies of water without providing as much overprotection for some
bodies of water as results from use of the total recoverable and
dissolved measurements.
The U.S. EPA cannot ignore the existence of pollution problems
and delay setting aquatic life criteria until all scientific
issues have been adequately resolved. In light of uncertainty,
the agency needs to derive criteria that are environmentally
conservative in most bodies of water. Because of uncertainty
concerning the relationship between the analytical chemistry and
the toxicity of metals, aquatic life criteria for metals are
expressed in terms of analytical measurements that result in the
criteria providing more protection than necessary for the aquatic
life in most bodies of water. The agency has provided for the
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use of WERs to address the general conservatism., but expects that
some WERs will be less than 1.0 because national, state, and
recalculated criteria are not necessarily environmentally
conservative for all bodies of water.
It has become obvious, however, that the determination and use of
WERs is not a simple solution to the existing general
conservatism. It is likely that a permanent solution will have
to be based on an adequate quantitative explanation of how metals
and aquatic organisms interact. In the meantime, the use of
total recoverable and dissolved measurements to express criteria
and the use of site-specific criteria are intended to provide
adequate protection for almost all bodies of water without
excessive overprotection for too many bodies of water. Work
needs to continue on the permanent solution and, just in case, on
improved alternative approaches.
Use of WERs to derive site-specific criteria is intended to allow
a reduction or elimination of the general overprotection
associated with application of a national criterion to individual
bodies of water, but a major problem is that a WER will rarely be
constant over time, location, and depth in a body of water due to
plumes, mixing, and resuspension. It is possible that dissolved
concentrations and WERs will be less variable than total
recoverable ones. It might also be possible to reduce the impact
of the heterogeneity if WERs are additive across time, location,
and depth (see Appendix G). Regardless of what approaches,
tools, hypotheses, and assumptions are utilized, variation will
exist and WERs will have to be used in a conservative manner.
Because of variation between bodies of water, national criteria
are derived to be environmentally conservative for most bodies of
water, whereas the WER procedure, which is intended to reduce the
general conservatism of national criteria, has to be conservative
because of variation among WERs within a body of water.
The conservatism introduced by variation among WERs is due not to
the concept of WERs, but to the way they are used. The reason
that national criteria are conservative in the first place is the
uncertainty concerning the linkage of analytical chemistry and
toxicity; the toxicity of solutions can be measured, but toxicity
cannot be modelled adequately using available chemical
measurements. Similarly, the current way that WERs are used
depends on a linkage between analytical chemistry and toxicity
because WERs are used to derive site-specific criteria that are
expressed in terms of chemical measurements.
Without changing the amount or kind of toxicity testing that is
performed when WERs are determined using Method 2, a different
way of using the WERs could avoid some of the problems introduced
by the dependence on analytical chemistry. The "sample-specific
WER approach" could consist of sampling a body of water at a
number of locations, determining the WER for each sample, and
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measuring the concentration of the metal in each sample. Then
for each individual sample, a quotient would be calculated by
dividing the concentration of metal in the sample by the product
of the national criterion times the WER obtained for that sample.
Except for experimental variation, when the quotient for a sample
is less than 1, the concentration of metal in that sample is
acceptable; when the quotient for a sample is greater than 1, the
concentration of metal in that sample is too high. As a check,
both the total recoverable measurement and the dissolved
measurement should be used because they should provide the same
answer if everything is done correctly and accurately. This
approach can also be used whenever Method 1 is used; although
Method 1 is used with simulated downstream water, the sample-
specific WER approach can be used with either simulated
downstream water or actual downstream water.
This sample-specific WER approach has several interesting
features:
1. It is not a different way of determining WERs; it is merely a
different way of using the WERs that are determined.
2. Variation among WERs within a body of water is not a problem.
3. It eliminates problems concerning the unknown relationship
between toxicity and analytical chemistry.
4. It works equally well in areas that are in or near plumes and
in areas that are away from plumes.
5. It works equally well in single-discharge and multiple-
discharge situations.
6. It automatically accounts for synergism, antagonism, and
additivity between toxicants.
This way of using WERs is equivalent to expressing the national
criterion for a pollutant in terms of toxicity tests whose
endpoints equal the CMC and the CCC; if the site water causes
less adverse effect than is defined to be the endpoint, the
concentration of that pollutant in the site water does not exceed
the national criterion. This sample-specific WER approach does
not directly fit into the current framework wherein criteria are
derived and then permit limits are calculated from the criteria.
If the sample-specific WER approach were to produce a number of
quotients that are greater than 1, it would seem that the
concentration of metal in the discharge(s) should be reduced
enough that the quotient is not greater than 1. Although this
might sound straightforward, the discharger(s) would find that a
substantial reduction in the discharge of a metal would not
achieve the intended result if the reduction was due to removal
of nontoxic metal. A chemical monitoring approach that cannot
differentiate between toxic and nontoxic metal would not detect
that only nontoxic metal had been removed, but the sample-
specific WER approach would.
15
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Figure 1: Four Ways to Derive a Permit Limit
Total Recoverable Criterion
\/
Recalculation
Procedure
_v
Total
Recoverable
crncWER
and/or cccWER
v
Total Recoverable
Site-specific Criterion
\/
Total Recoverable Permit Limit
Dissolved Criterion = (TR Criterion) (% dissolved in toxicity tests)
\/
\/
Recalculation
Procedure
Dissolved
cmcWER
and/or cccWER
\/
\/
Dissolved Site-
specific Criterion
_v
Net % contribution from the total recoverable metal in the effluent
to the dissolved metal in the downstream water. (This will probably
change if the total recoverable concentration in the effluent changes.)
\/
Total Recoverable Permit Limit
For both the total recoverable and dissolved measurements, derivation of an
optional site-specific criterion is described on the right. If both, the
Recalculation Procedure and the WER procedure are used, the Recalculation
Procedure must be performed first. (The Recalculation Procedure cannot be
used in jurisdictions that are subject to the National Toxics Rule.)
16
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METHOD 1: DETERMINING WERs FOR AREAS IN OR NEAR PLUMES
Method 1 is based on the determination of WERs using simulated
downstream water and so it can be used to determine a WER that
applies in the vicinity of a plume. Use of simulated downstream
water ensures that the concentration of effluent in the site
water is known, which is important because the magnitude of the
WER will often depend on the concentration of effluent in the
downstream water. Knowing the concentration of effluent makes it
possible to quantitatively relate the WER to the effluent.
Method 1 can be used to determine either cmcWERs or cccWERs or
both in single-metal, flowing freshwater situations, including
streams whose design flow is zero and "effluent-dependent"
streams (see Appendix F). As is also explained in Appendix F,
Method 1 is used when cmcWERs are determined for "large sites",
although Method 2 is used when cccWERs are determined for "large
sites". In addition, Appendix F addresses special considerations
regarding multiple-metal and/or multiple-discharge situations.
Neither Method 1 nor Method 2 covers all important methodological
details for conducting the side-by-side toxicity tests that are
necessary in order to determine a WER. Many references are made
to information published by the U.S. EPA (1993a,b,c) concerning
toxicity tests on effluents and surface waters and by ASTM
(1993a,b,c,d,e,f) concerning tests in laboratory dilution water.
Method 1 addresses aspects of toxicity tests that (a) need
special attention when determining WERs and/or (b) are usually
different for tests conducted on effluents and tests conducted in
laboratory dilution water. Appendix H provides additional
information concerning toxicity tests with saltwater species.
A. Experimental Design
Because of the variety of considerations that have important
implications for the determination of a WER, decisions
concerning experimental design should be given careful
attention and need to answer the following questions:
1. Should WERs be determined using upstream water, actual
downstream water, and/or simulated downstream water?
2. Should WERs be determined when the stream flow is equal to,
higher than, and/or lower than the design flow?
3. Which toxicity tests should be used?
4. Should a cmcWER or a cccWER or both be determined?
5. How should a FWER be derived?
6. For metals whose criteria are hardness-dependent, at what
hardness should WERs be determined?
The answers to these questions should be based on the reason
that WERs are determined, but the decisions should also take
into account some practical considerations.
17
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1. Should WERs be determined using upstream water, actual
downstream water, and/or simulated downstream water?
a. Upstream water provides the least complicated way of
determining and using WERs because plumes, mixing
zones, and effluent variability do not have to be taken
into account. Use of upstream water provides the least
useful WERs because it does not take into account the
presence of the effluent, which is the source of the
metal. It is easy to assume that upstream water will
give smaller WERs than downstream water, but in some
cases downstream water might give smaller WERs (see
Appendix G). Regardless of whether upstream water
gives smaller or larger WERs, a WER should be
determined using the water to which the site-specific
criterion is to apply (see Appendix A).
b. Actual downstream water might seem to be the most
pertinent water to use when WERs are determined, but
whether this is true depends on what use is to be made
of the WERs. WERs determined using actual downstream
water can be quantitatively interpreted using the
sample-specific WER approach described at the end of
the Introduction. If, however, it is desired to
understand the quantitative implications of a WER for
an effluent of concern, use of actual downstream water
is problematic because the concentration of effluent in
the water can only be known approximately.
Sampling actual downstream water in areas that are in
or near plumes is especially difficult. The WER
obtained is likely to depend on where the sample is
taken because the WER will probably depend on the
percent effluent in the sample (see Appendix D). The
sample could be taken at the end of the pipe, at the
edge of the acute mixing zone, at the edge of the
chronic mixing zone, or in a completely mixed
situation. If the sample is taken at the edge of a
mixing zone, the composition of the sample will
probably differ from one point to another along the
edge of the mixing zone.
If samples of actual downstream water are to be taken
close to a discharge, the mixing patterns and plumes
should be well known. Dye dispersion studies
(Kilpatrick 1992) are commonly used to determine
isopleths of effluent concentration and complete mix;
dilution models (U.S. EPA 1993d) might also be helpful
when selecting sampling locations. The most useful
samples of actual downstream water are probably those
taken just downstream of the point at. which complete
mix occurs or at the most distant point that is within
18
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the site to which the site-specific criterion is to
apply. When samples are collected from a complete-mix
situation, it might be appropriate to composite samples
taken over a cross section of the stream. Regardless
of where it is decided conceptually that a sample
should be taken, it might be difficult to identify
where the point exists in the stream and how it changes
with flow and over time. In addition, if it is not
known exactly what the sample actually represents,
there is no way to know how reproducible the sample is.
These problems make it difficult to relate WERs
determined in actual downstream water to an effluent of
concern because the concentration of effluent in the
sample is not known; this is not a problem, however, if
the sample-specific WER approach is used to interpret
the results.
Simulated downstream water would seem to be the most
unnatural of the three kinds of water, but it offers
several important advantages because effluent and
upstream water are mixed at a known ratio. This is
important because the magnitude of the WER will often
depend on the concentration of effluent in the
downstream water. Mixtures can be prepared to simulate
the ratio of effluent and upstream water that exists at
the edge of the acute mixing zone, at the edge of the
chronic mixing zone, at complete mix, or at any other
point of interest. If desired, a sample of effluent
can be mixed with a sample on upstream water in
different ratios to simulate different points in a
stream. Also, the ratio used can be one that simulates
conditions at design flow or at any other flow.
The sample-specific WER approach can be used with both
actual and simulated downstream water. Additional
quantitative uses can be made of WERs determined using
simulated downstream water because the percent effluent
in the water is known, which allows quantitative
extrapolations to the effluent. In addition, simulated
downstream water can be used to determine the variation
in the WER that is due to variation in the effluent.
It also allows comparison of two or more effluents and
determination of the interactions of two or more
effluents. Additivity of WERs can be studied using
simulated downstream water (see Appendix G); studies of
toxicity within plumes and studies of whether increased
flow of upstream water can increase toxicity are both
studies of additivity of WERs. Use of simulated
downstream water also makes it possible to conduct
controlled studies of changes in WERs due to aging and
changes in pH.
19
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In Method 1, therefore, WERs are determined using
simulated downstream water that is prepared by mixing
samples of effluent and upstream water in an appropriate
ratio. Most importantly, Method 1 can be used to
determine a WER that applies in the vicinity of a plume
and can be quantitatively extrapolated to the effluent.
2. Should WERs be determined when the stream flow is equal
to, higher than, and/or lower than the design flow?
WERs are used in the derivation of site-specific criteria
when it is desired that permit limits be based on a
criterion that takes into account the characteristics of
the water and/or the metal at the site. In most cases,
permit limits are calculated using steady-state models and
are based on a design flow. It is therefore important
that WERs be adequately protective under design-flow
conditions, which might be expected to require that some
sets of samples of effluent and upstream water be obtained
when the actual stream flow is close to the design flow.
Collecting samples when the stream flow is close to the
design flow will limit a WER determination to the low-flow
season (e.g., from mid-July to mid-October in some places)
and to years in which the flow is sufficiently low.
It is also important, however, that WERs that are applied
at design flow provide adequate protection at higher
flows. Generalizations concerning the impact of higher
flows on WERs are difficult because such flows might (a)
reduce hardness, alkalinity, and pH, (b) increase or
decrease the concentrations of TOG and TSS, (c) resuspend
toxic and/or nontoxic metal from the sediment, and (d)
wash additional pollutants into the water. Acidic
snowmelt, for example, might lower the WER both by
diluting the WER and by reducing the hardness, alkalinity,
and pH; if substantial labile metal is present, the WER
might be lowered more than the concentration of the metal,
possibly resulting in increased toxicity at flows higher
than design flow. Samples taken at higher flows might
give smaller WERs because the concentration of the
effluent is more dilute; however, total recoverable WERs
might be larger if the sample is taken just after an event
that greatly increases the concentration of TSS and/or TOG
because this might increase both (1) the concentration of
nontoxic particulate metal in the water and (2) the
capacity of the water to sorb and detoxify metal.
WERs are not of concern when the stream flow is lower than
the design flow because these are acknowledged times of
reduced protection. Reduced protection might not occur,
however, if the WER is sufficiently high when the flow is
lower than design flow.
20
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3. Which toxicity tests should be used?
a. As explained in Appendix D, the magnitude of an
experimentally determined WER is likely to depend on
the sensitivity of the toxicity test used. This
relationship between the magnitude of the WER and the
sensitivity of the toxicity test is due to the aqueous
chemistry of metals and is not related to the test
organisms or the type of test. The available data
indicate that WERs determined with different tests do
not differ greatly if the tests have about the same
sensitivities, but the data also support the
generalization that less sensitive toxicity tests
usually give smaller WERs than more sensitive tests
(see Appendix D).
b. When the CCC is lower than the CMC, it is likely that a
larger WER will result from tests that are sensitive at
the CCC than from tests that are sensitive at the CMC.
c. The considerations concerning the sensitivities of two
tests should also apply to two endpoints for the same
test. For any lethality test, use of the LC25 is
likely to result in a larger WER than use of the LC50,
although the difference might not be measurable in most
cases and the LC25 is likely to be more variable than
the LC50. Selecting the percent effect to be used to
define the endpoint might take into account (a) whether
the endpoint is above or below the CMC and/or the CCC
and (b) the data obtained when tests are conducted.
Once the percent effect is selected for a particular
test (e.g., a 48-hr LC50 with 1-day-old fathead
minnows)., the same percent effect must be used whenever
that test is used to determine a WER for that effluent.
Similarly, if two different tests with the same species
(e.g., a lethality test and a sublethal test) have
substantially different sensitivities, both a cmcWER
and a cccWER could be obtained with the same species.
d. The primary toxicity test used in the determination of
a WER should have an endpoint in laboratory dilution
water that is close to, but not lower than, the CMC
and/or CCC to which the WER is to be applied.
e. Because the endpoint of the primary test in laboratory
dilution water cannot be lower than the CMC and/or CCC,
the magnitude of the WER is likely to become closer to
1 as the endpoint of the primary test becomes closer to
the CMC and/or-CCC (see Appendix D).
f. The WER obtained, with the primary test should be
confirmed with a secondary test that uses a species
that is taxonomically different from the species used
in the primary test.
1) The endpoint of the secondary test may be higher or
lower than the CMC, the CCC, or the endpoint of the
primary test.
21
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2) Because of the limited number of toxicity tests that
have sensitivities near the CMC or CCC for a metal,
it seems unreasonable to require that the two
species be further apart taxonomically than being in
different orders.
Two different endpoints with the same species must not
be used as the primary and secondary tests, even if one
endpoint is lethal and the other is sublethal.
If more sensitive toxicity tests generally give larger
WERs than less sensitive tests, the maximum value of a
WER will usually be obtained using a toxicity test
whose endpoint in laboratory dilution water equals the
CMC or CMC. If such a test is not used, the maximum
possible WER probably will not be obtained.
No rationale exists to support the idea that different
species or tests with the same sensitivity will produce
different WERs. Because the mode of action might
differ from species to species and/or from effect to
effect, it is easy to speculate that in some cases the
magnitude of a WER will depend to some extent on the
species, life stage, and/or kind of test, but no data
are available to support conclusions concerning the
existence and/or magnitude of any such differences.
If the tests are otherwise acceptable, both cmcWERs and
cccWERs may be determined using acute and/or chronic
tests and using lethal and/or sublethal endpoints. The
important consideration is the sensitivity of the test,
not the duration, species, life stage, or adverse
effect used.
There is no reason to use species that occur at the
site; they may be used in the determination of a WER if
desired, but:
1) It might be difficult to determine which of the
species that occur at the site are sensitive to the
metal and are adaptable to laboratory conditions.
2) Species that occur at the site might be harder to
obtain in sufficient numbers for conducting toxicity
tests over the testing period.
3) Additional QA tests will probably be needed (see
section C.3.b) because data are not likely to be
available from other laboratories for comparison
with the results in laboratory dilution water.
Because a WER is a ratio of results obtained with the
same test in two different dilution waters, toxicity
tests that are used in WET testing, for example, may be
used, even if the national aquatic life guidelines
(U.S. EPA 1985) do not allow use of the test in the
derivation of an aquatic life criterion. Of course, a
test whose endpoint in laboratory dilution water is
below the CMC and/or CCC that is to be adjusted cannot
be used as a primary test.
22
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1. Because there is no rationale that suggest that it
makes any difference whether the test is conducted with
a species that is warmwater or coldwater, a fish or an
invertebrate, or resident or nonresident at the site,
other than the fact that less sensitive tests are
likely to give smaller WERs, such considerations as the
availability of test organisms might be important in
the selection of the test. Information in Appendix I,
a criteria document for the metal of concern (see
Appendix E), or any other pertinent source might be
useful when selecting primary and secondary tests.
m. A test in which the test organisms are not fed might
give a different WER than a test in which the organisms
are fed just because of the presence of the food (see
Appendix D). This might depend on the metal, the type
and amount of food, and whether a total recoverable or
dissolved WER is determined.
Different tests with similar sensitivities are expected to
give similar WERs, except for experimental variation. The
purpose of the secondary test is to provide information
concerning this assumption and the validity of the WER.
4. Should a cmcWER or a cccWER or both be determined?
This question does not have to be answered if the
criterion for the site contains either a CMC or a CCC but
not both. For example, a body of water that is protected
for put-and-take fishing might have only a CMC, whereas a
stream whose design flow is zero might have only a CCC.
When the criterion contains both a CMC and a CCC, the
simplistic way to answer the question is to determine
whether the CMC or the CCC controls the existing permit
limits; which one is controlling depends on (a) the ratio
of the CMC to the CCC, (b) whether the number of mixing
zones is zero, one, or two, and (c) which steady-state or
dynamic model was used in the calculation of the permit
limits. A better way to answer the question would be to
also determine how much the controlling value would have
to be changed for the other value to become controlling;
this might indicate that it would not be cost-effective to
derive, for example, a site-specific CMC (ssCMC) without
also deriving a site-specific CCC (ssCCC). There are also
other possibilities: (1) It might be appropriate to use a
phased approach, i.e., determine either the cmcWER or the
cccWER and then decide whether to determine the other.
(2) It might be appropriate and environmentally
conservative to determine a WER that can be applied to
both the CMC and the CCC. (3) It is always allowable to
determine and use both a cmcWER and a cccWER, although
both can be determined only if toxicity tests with
appropriate sensitivities are available.
23
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Because the phased, approach can always be used, it is only
important to decide whether to use a different approach
when its use might be cost-effective. Deciding whether to
use a different approach and selecting which one to use is
complex because a number of considerations need to be
taken into account:
a. Is the CMC equal to or higher than the CCC?
If the CMC equals the CCC, two WERs cannot be
determined if they would be determined using the
same site water, but two WERs could be determined if
the cmcWER and the cccWER would be determined using
different site waters, e.g., waters that contain
different concentrations of the effluent.
b. If the CMC is higher than the CCC, is there a toxicity
test whose endpoint in laboratory dilution water is
between the CMC and the CCC?
If the CMC is higher than the CCC and there is a
toxicity test whose endpoint in laboratory dilution
water is between the CMC and the CCC, both a cmcWER
and a cccWER can be determined. If the CMC is
higher than the CCC but no toxicity test has an
endpoint in laboratory dilution water between the
CMC and the CCC, two WERs cannot be determined if
they would be determined using the same site water;
two WERs could be determined if they were determined
using different site waters, e.g., waters that
contain different concentrations of the effluent.
c. Was a steady-state or a dynamic model used in the
calculation of the permit limits?
It is complex, but reasonably clear, how to make a
decision when a steady-state model was used, but it
is not clear how a decision should, be made when a
dynamic model was used.
d. If a steady-state model was used, were one or two
design flows used, i.e., was the hydrologically based
steady-state method used or was the biologically based
steady-state method used?
When the hydrologically based method is used, one
design flow is used for both the CMC and the CCC,
whereas when the biologically based method is used,
there is a CMC design flow and a CCC design flow.
When WERs are determined using downstream water, use
of the biologically based method will probably cause
the percent effluent in the site water used in the
determination of the cmcWER to be different from the
percent effluent in the site water used in the
determination of the cccWER; thus the two WERs
should be determined using two different site
waters. This does not impact WERs determined using
upstream water.
24
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e. Is there an acute mixing zone? Is there a chronic
mixing zone?
1. When WERs are determined using upstream water,
the presence or absence of mixing zones has no
impact; the cmcWER and the cccWER will both be
determined using site water that contains zero
percent effluent, i.e., the two WERs will be
determined using the same site water.
2. Even when downstream water is used, whether there
is an acute mixing zone affects the point of
application of the CMC or ssCMC, but it does not
affect the determination of any WER.
3. The existence of a chronic mixing zone has
important implications for the determination of
WERs when downstream water is used (see Appendix
A). When WERs are determined using downstream
water, the cmcWER should be determined using
water at the edge of the chronic mixing zone,
whereas the cccWER should be determined using
water from a complete-mix situation. (If the
biologically based method is used, the two
different design flows should also be taken into
account when determining the percent effluent
that should be in the simulated downstream
water.) Thus the percent effluent in the site
water used in the determination of the cmcWER
will be different from the percent effluent in
the site water used in the determination of the
cccWER; this is important because the magnitude
of a WER will often depend substantially on the
percent effluent in the water (see Appendix D).
f. In what situations would it be environmentally
conservative to determine one WER and use it to adjust
both the cmcWER and the cccWER?
Because (1) the CMC is never lower than the CCC and
(2) a more sensitive test will generally give a WER
closer to 1, it will be environmentally conservative
to use a cmcWER to adjust a CCC when there are no
contradicting considerations. In this case, a
cmcWER can be determined and used to adjust both the
CMC and the CCC. Because water quality can affect
the WER, this approach is necessarily valid only if
the cmcWER and the cccWER are determined in the same
site water. Other situations in which it would be
environmentally conservative to use one WER to
adjust both the CMC and the CCC are described below.
These considerations have one set of implications when
both the cmcWER and cccWER are to be determined using the
same site water, and another set of implications when the
two WERs are to be determined using different site waters,
e.g., when the site waters contain different
concentrations of effluent.
25
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When WERs are determined using upstream water, the same
site water is used in the determination of both the cmcWER
and the cccWER. Whenever the two WERs are determined in
the same site water, any difference in the magnitude of
the cmcWER and the cccWER will probably be due to the
sensitivities of the toxicity tests used. Therefore:
a. If more sensitive toxicity tests generally give larger
WERs than less sensitive tests, the maximum cccWER (a
cccWER determined with a test whose endpoint equals the
CCC) will usually be larger than the maximum cmcWER
because the CCC is never higher than the CMC.
b. Because the CCC is never higher than the CMC, the
maximum cmcWER will usually be smaller than the maximum
cccWER and it will be environmentally conservative to
use the cmcWER to adjust the CCC.
c. A cccWER can be determined separately from a cmcWER
only if there is a toxicity test with an endpoint in
laboratory dilution water that is between the CMC and
the CCC. If no such test exists or can be devised,
only a cmcWER can be determined, but it can be used to
adjust both the CMC and the CCC.
d. Unless the experimental variation is increased, use of
a cccWER, instead of a cmcWER, to adjust the CCC will
usually improve the accuracy of the resulting site-
specific CCC. Thus a cccWER may be determined and used
whenever desired, if a toxicity test has an endpoint in
laboratory dilution water between the CMC and the CCC.
e. A cccWER cannot be used to adjust a CMC if the cccWER
was determined using an endpoint that was lower than
the CMC in laboratory dilution water because it will
probably reduce the level of protection.
f. Even if there is a toxicity test that has an endpoint
in laboratory dilution water that is between the CMC
and the CCC, it is not necessary to decide initially
whether to determine a cmcWER and/or a cccWER. When
upstream water is used, it is always aillowable to
determine a cmcWER and use it to derive a site-specific
CMC and a site-specific CCC and then decide whether to
determine a cccWER.
g. If there is a toxicity test whose endpoint in
laboratory dilution water is between the CCC and the
CMC, and if this test is used as the secondary test in
the determination of the cmcWER, this test will provide
information that should be very useful for deciding
whether to determine a cccWER in addition to a cmcWER.
Further, if it is decided to determine a cccWER, the
same two tests used in the determination of the cmcWER
could then be used in the determination of the cccWER,
with a reversal of their roles as primary and secondary
tests. Alternatively, a cmcWER and a cccWER could be
determined simultaneously if both tests are conducted
on each sample of site water.
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When WERs are determined using downstream water, the
magnitude of each WER will probably depend on the
concentration of effluent in the downstream water used
(see Appendix D). The first important consideration is
whether the design flow is greater than zero, and the
second is whether there is a chronic mixing zone.
a. If the design flow is zero, cmcWERs and/or cccWERs that
are determined for design-flow conditions will both be
determined in 100 percent effluent. Thus this case is
similar to using upstream water in that both WERs are
determined in the same site water. When WERs are
determined for high-flow conditions, it will make a
difference whether a chronic mixing zone needs to be
taken into account, which is the second consideration.
b. If there is no chronic mixing zone, both WERs will be
determined for the complete-mix situation; this case is
similar to using upstream water in that both WERs are
determined using the same site water. If there is a
chronic mixing zone, cmcWERs should be determined in
the site water that exists at the edge of the chronic
mixing zone, whereas cccWERs should be determined for
the complete-mix situation (see Appendix A). Thus the
percent effluent will be higher in the site water used
in the determination of the cmcWER than in the site
water used in the determination of the cccWER. Because
a site water with a higher percent effluent will
probably give a larger WER than a site water with a
lower percent effluent, both a cmcWER and a cccWER can
be determined even if there is no test whose endpoint
in laboratory dilution water is between the CMC and the
CCC. There are opposing considerations, however:
1) The site water used in the determination of the
cmcWER will probably have a higher percent effluent
than the site water used in the determination of the
cccWER, which will tend to cause the cmcWER to be
larger than the cccWER.
2) If there is a toxicity test whose endpoint in
laboratory dilution water is between the CMC and the
CCC, use of a more sensitive test in the
determination of the cccWER will tend to cause the
cccWER to be larger than the cmcWER.
One consequence of these opposing considerations is that
it is not known whether use of the cmcWER to adjust the
CCC would be environmentally conservative; if this
simplification is not known to be conservative, it should
not be used. Thus it is important whether there is a
toxicity test whose endpoint in laboratory dilution water
is between the CMC and the CCC:
a. If no toxicity test has an endpoint in laboratory
dilution water between the CMC and the CCC, the two
WERs have to be determined with the same test, in which
case the cmcWER will probably be larger because the
27
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percent effluent in the site water will be higher.
Because of the difference in percent effluent in the
site waters that should be used in the determinations
of the two WERs, use of the cmcWER to adjust the CCC
would not be environmentally conservative, but use of
the cccWER to adjust the CMC would be environmentally
conservative. Although both WERs could be determined,
it would also be acceptable to determine only the
cccWER and use it to adjust both the CMC and the CCC.
b. If there is a toxicity test whose endpoint in
laboratory dilution water is between the CMC and the
CCC, the two WERs could be determined using different
toxicity tests. An environmentally conservative
alternative to determining two WERs would be to
determine a hybrid WER by using (1) a toxicity test
whose endpoint is above the CMC (i.e., a toxicity test
that is appropriate for the determination of a cmcWER)
and (2) site water for the complete-mix situation
(i.e., site water appropriate for the determination of
cccWER). It would be environmentally conservative to
use this hybrid WER to adjust the CMC and it would be
environmentally conservative to use this hybrid WER to
adjust the CCC. Although both WERs could be
determined, it would also be acceptable to determine
only the hybrid WER and use it to adjust both the CMC
and the CCC. (This hybrid WER described here in
paragraph b is the same as the cccWER described in
paragraph a above in which no toxicity test had an
endpoint in laboratory dilution water between the CMC
and the CCC.)
How should a FWER be derived?
Background
Because of experimental variation and variation in the
composition of surface waters and effluents, a single
determination of a WER does not provide sufficient
information to justify adjustment of a criterion. After a
sufficient number of WERs have been determined in an
acceptable manner, a Final Water-Effect Ratio (FWER) is
derived from the WERs, and the FWER is then used to
calculate the site-specific criterion. If both a site-
specific CMC and a site-specific CCC are to be derived,
both a cmcFWER and a cccFWER have to be derived, unless an
environmentally conservative estimate is used in place of
the cmcFWER and/or the cccFWER.
When a WER is determined using upstream water, the two
major sources of variation in the WER are (a) variability
in the quality of the upstream water, much of which might
be related to season and/or flow, and (b) experimental
28
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variation. When a WER is determined in downstream water,
the four major sources of variation are (a) variability in
the quality of the upstream water, much of which might be
related to season and/or flow, (b) experimental variation,
(c) variability in the composition of the effluent, and
(d) variability in the percent effluent in the downstream
water. Variability and the possibility of mistakes and
rare events make it necessary to try to compromise between
(1) providing a high probability of adequate protection
and (2) placing too much reliance on the smallest
experimentally determined WER, which might reflect
experimental variation, a mistake, or a rare event rather
than a meaningful difference in the WER.
Various ways can be employed to address variability:
a. Replication can be used to reduce the impact of some
sources of variation and to verify the importance of
others.
b. Because variability in the composition of the effluent
might contribute substantially to the variability of
the WER, it might be desirable to obtain and store two
or more samples of the effluent at slightly different
times, with the selection of the sampling times
depending on such characteristics of the discharge as
the average retention time, in case an unusual WER is
obtained with the first sample used.
c. Because of the possibility of mistakes and rare events,
samples of effluent and upstream water should be large
enough that portions can be stored for later testing or
analyses if an unusual WER is obtained.
d. It might be possible to reduce the impact of the
variability in the percent effluent in the downstream
water by establishing a relationship between the WER
and the percent effluent.
Confounding of the sources can be a problem when more than
one source contributes substantial variability.
When permit limits are calculated using a steady-state
model, the limits are based on a design flow, e.g., the
7Q10. It is usually assumed that a concentration of metal
in an effluent that does not cause unacceptable effects at
the design flow will not cause unacceptable effects at
higher flows because the metal is diluted by the increased
flow of the upstream water. Decreased protection might
occur, however, if an increase in flow increases toxicity
more than it dilutes the concentration of metal. When
permit limits are based on a national criterion, it is
often assumed that the criterion is sufficiently
conservative that an increase in toxicity will not be
great enough to overwhelm the combination of dilution and
the assumed conservatism, even though it is likely that
the national criterion is not overprotective of all bodies
29
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of water. When WERs are used to reduce the assumed
conservatism, there is more concern about the possibility
of increased toxicity at flows higher than the design flow
and it is important to (1) determine some WERs that
correspond to higher flows or (2) provide some
conservatism. If the concentration of effluent in the
downstream water decreases as flow increases, WERs
determined at higher flows are likely to be smaller than
WERs determined at design flow but the concentration of
metal will also be lower. If the concentration of TSS
increases at high flows, however, both the WER and the
concentration of metal might increase. If they are
determined in an appropriate manner, WERs determined at
flows higher than the design flow can be used in two ways :
a. As environmentally conservative estimates of WERs
determined at design flow.
b. To assess whether WERs determined at design flow will
provide adequate protection at higher flows .
In order to appropriately take into account seasonal and
flow effects and their interactions, both ways of using
high- flow WERs require that the downstream water used in
the determination of the WER be similar to that which
actually exists during the time of concern. In addition,
high- flow WERs can be used in the second way only if the
composition of the downstream water is known. To satisfy
the requirements that (a) the downstream water used in the
determination of a WER be similar to the actual water and
(b) the composition of the downstream water be known, it
is necessary to obtain samples of effluent and upstream
water at the time of concern and to prepare a simulated
downstream water by mixing the samples at the ratio of the
flows of the effluent and the upstream water that existed
when the samples were obtained.
For the first way of using high- flow WERs, they are used
directly as environmentally conservative estimates of the
design- flow WER. For the second way of using high- flow
WERs, each is used to calculate the highest concentration
of metal that could be in the effluent without causing the
concentration of metal in the downstream water to exceed
the site-specific criterion that would be derived for that
water using the experimentally determined WER. This
highest concentration of metal in the effluent (HCME) can
be calculated as :
r/n- - [ (CCC) (WER) (eFLOW + uFLOW) ] - [ (uCONC) (uFLOW) ]
HCME -
where :
CCC = the national, state, or recalculated CCC (or CMC)
that is to be adjusted.
30
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eFLOW = the flow of the effluent that was the basis of the
preparation of the simulated downstream water.
This should be the flow of the effluent that
existed when the samples were taken.
uFLOW = the flow of the upstream water that was the basis
of the preparation of the simulated downstream
water. This should be the flow of the upstream
water that existed when the samples were taken.
uCONC = the concentration of metal in the sample of
upstream water used in the preparation of the
simulated downstream water.
In order to calculate a HCME from an experimentally
determined WER, the only information needed besides the
flows of the effluent and the upstream water is the
concentration of metal in the upstream water, which should
be measured anyway in conjunction with the determination
of the WER.
When a steady-state model is used to derive permit limits,
the limits on the effluent apply at all flows; thus, each
HCME can be used to calculate the highest WER (hWER) that
could be used to derive a site-specific criterion for the
downstream water at design flow so that there would be
adequate protection at the flow for which the HCME was
determined. The hWER is calculated as:
, = (HCME) (eFLOWdf} + (uCONCdf) (uFLOWdf]
(CCC) (eFLOWdf + uFLOWdf)
The suffix "df" indicates that the values used for these
quantities in the calculation of the hWER are those that
exist at design-flow conditions. The additional datum
needed in order to calculate the hWER is the concentration
of metal in upstream water at design-flow conditions; if
this is assumed to be zero, the hWER will be
environmentally conservative. If a WER is determined when
uFLOW equals the design flow, hWER = WER.
The two ways of using WERs determined at flows higher than
design flow can be illustrated using the following
examples. These examples were formulated using the
concept of additivity of WERs (see Appendix G). A WER
determined in downstream water consists of two components,
one due to the effluent (the eWER) and one due to the
upstream water (the uWER). If the eWER and uWER are
strictly additive, when WERs are determined at various
upstream flows, the downstream WERs can be calculated from
the composition of the downstream water (the % effluent
and the % upstream water) and the two WERs (the eWER and
the uWER) using the equation:
31
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(% effluent) (eWER) + (% upstream water) (uWER)
100
In the examples below, it is assumed that:
a. A site-specific CCC is being derived.
b. The national CCC is 2 ug/L.
c. The eWER is 40.
d. The eWER and uWER are constant and strictly additive.
e. The flow of the effluent (eFLOW) is always 10 cf s .
f . The design flow of the upstream water (uFLOWdf) is 40
cfs.
Therefore :
.. _ [(2 ug/L) (WER) (10 cfs + UFLOW)
HCME --
- [ (uCONC) (uFLOW) ]
10 ug/L
, = (HCME} (10 cfs) + (uCONCdf) (40 cfs}
(2 ug/L) (10 cfs + 40 cfs)
In the first example, the uWER is assumed to be 5 and so
the upstream site-specific CCC (ussCCC) = (CCC)(uWER) =
(2 ug/L)(5) = 10 ug/L. uCONC is assumed to be 0.4 ug/L,
which means that the assimilative capacity of the upstream
water is 9.6 ug/L.
eFLOW
(cfs)
10
10
10
10
10
10
10
uFLOW
(cfs)
40
63
90
190
490
990
1990
At Complete Mix
% Eff. % UPS.
WER
20
13
10
5
2
1
0
.0
.7
.0
.0
.0
.0
.5
80
86
90
95
98
99
99
.0
.3
.0
.0
.0
.0
.5
12
9
8
6
5
5
5
.000
.795
.500
.750
.700
.350
.175
HCME
(ug/L)
118.4
140.5
166.4
262.4
550.4
1030.4
1990.4
hWER
12.00
14.21
16.80
26.40
55.20
103 .20
199.20
As the flow of the upstream water increases, the WER
decreases to a limiting value equal to uWER. Because the
assimilative capacity is greater than zero, the HCMEs and
hWERs increase due to the increased dilution of the
effluent. The increase in hWER at higher flows will not
allow any use of the assimilative capacity of the upstream
water because the allowed concentration of metal in the
effluent is controlled by the lowest hWER, which is the
design-flow hWER in this example. Any WER determined at a
higher flow can be used as an environmentally conservative
estimate of the design-flow WER, and the hWERs show that
the WER of 12 provides adequate protection at all flows.
When uFLOW equals the design flow of 40 cfs, WER = hWER.
32
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In the second example, uWER is assumed to be 1, which
means that ussCCC = 2 ug/L. uCONC is assumed to be 2
ug/L, so that uCONC = ussCCC. The assimilative capacity
of the upstream water is 0 ug/L.
eFLOW
(cfs)
10
10
10
10
10
10
10
uFLOW
(cfs)
40
63
90
190
490
990
1990
At Complete Mix
% Eff. % Ups.
WER
20
13
10
5
2
1
0
.0
.7
.0
.0
.0
.0
.5
80
86
90
95
98
99
99
.0
.3
.0
.0
.0
.0
.5
8
6
4
2
1
1
1
.800
.343
.900
.950
.780
.390
.195
HCME
(ucr/L)
80.00
80.00
80.00
80.00
80.00
80.00
80.00
hWER
8.800
8.800
8.800
8.800
8.800
8.800
8.800
All the WERs in this example are lower than the comparable
WERs in the first example because the uWER dropped from 5
to 1; the limiting value of the WER at very high flow is
1. Also, the HCMEs and hWERs are independent of flow
because the increased dilution does not allow any more
metal to be discharged when uCONC = ussCCC, i.e., when the
assimilative capacity is zero. As in the first example,
any WER determined at a flow higher than design flow can
be used as an environmentally conservative estimate of the
design-flow WER and the hWERs show that the WER of 8.8
determined at design flow will provide adequate protection
at all flows for which information is available. When
uFLOW equals the design flow of 40 cfs, WER = hWER.
In the third example, uWER is assumed to be 2, which means
that ussCCC = 4 ug/L. uCONC is assumed to be 1 ug/L; thus
the assimilative capacity of the upstream water is 3 ug/L.
eFLOW
(cfs)
10
10
10
10
10
10
10
uFLOW
(cfs)
40
63
90
190
490
990
1990
At Complete Mix
% Eff. % UPS.
WER
20
13
10
5
2
1
0
.0
.7
.0
.0
.0
.0
.5
80
86
90
95
98
99
99
.0
.3
.0
.0
.0
.0
.5
9
7
5
3
2
2
2
.600
.206
.800
.900
.760
.380
.190
HCME
(ucr/L)
92.0
98.9
107.0
137.0
227.0
377.0
677.0
hWER
9.60
10.29
11.10
14.10
23.10
38.10
68.10
All the WERs in this example are intermediate between the
comparable WERs in the first two examples because the uWER
is now 2, which is between 1 and 5; the limiting value of
the WER at very high flow is 2. As in the other examples,
any WER determined at a flow higher than design flow can
be used as an environmentally conservative estimate of the
33
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design-flow WER and the hWERs show that the WER of 9.6
determined at design flow will provide adequate protection
at all flows for which information is available. When
uPLOW equals the design flow of 40 cfs, WER = hWER.
If this third example is assumed to be subject to acidic
snowmelt in the spring so that the eWER and uWER are less-
than-additive and result in a WER of 4.8 (rather than 5.8)
at a uFLOW of 90 cfs, the third HCME would be 87 ug/L, and
the third hWER would be 9.1. This hWER is lower than the
design-flow WER of 9.6, so the site-specific criterion
would have to be derived using the WER of 9.1, rather than
the design-flow WER of 9.6, in order to provide the
intended level of protection. If the eWER and uWER were
less-than-additive only to the extent that the third WER
was 5.3, the third HCME would be 97 ug/L and the third
hWER would be 10.1. In this case, dilution by the
increased flow would more than compensate for the WERs
being less-than-additive, so that the design-flow WER of
9.6 would provide adequate protection at a uFLOW of 90
cfs. Auxiliary information might indicate whether an
unusual WER is real or is an accident; for example, if the
hardness, alkalinity, and pH of snowmelt are all low, this
information would support a low WER.
If the eWER and uWER were more-than-additive so that the
third WER was 10, this WER would not be an environmentally
conservative estimate of the design-flow WER. If a WER
determined at a higher flow is to be used as an estimate
of the design-flow WER and there is reason to believe that
the eWER and the uWER might be more-than-additive, a test
for additivity can be performed (see Appendix G).
Calculating HCMEs and hWERs is straightforward if the WERs
are based on the total recoverable measurement. If they
are based on the dissolved measurement, it is necessary to
take into account the percent of the total recoverable
metal in the effluent that becomes dissolved in the
downstream water.
To ensure adequate protection, a group of WERs should
include one or more WERs corresponding to flows near the
design flow, as well as one or more WERs corresponding to
higher flows.
a. Calculation of hWERs from WERs determined at various
flows and seasons identifies the highest WER that can
be used in the derivation of a site-specific criterion
and still provide adequate protection at all flows for
which WERs are available. Use of hWERs eliminates the
need to assume that WERs determined at design flow will
provide adequate protection at higher flows. Because
hWERs are calculated to apply at design flow, they
34
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apply to the flow on which the permit limits are based.
The lowest of the hWERs ensures adequate protection at
all flows, if hWERs are available for a sufficient
range of flows, seasons, and other conditions.
b. Unless additivity is assumed, a WER cannot be
extrapolated from one flow to another and therefore it
is not possible to predict a design-flow WER from a WER
determined at other conditions. The largest WER is
likely to occur at design flow because, of the flows
during which protection is to be provided, the design
flow is the flow at which the highest concentration of
effluent will probably occur in the downstream water.
This largest WER has to be experimentally determined;
it cannot be predicted.
The examples also illustrate that if the concentration of
metal in the upstream water is below the site-specific
criterion for that water, in the limit of infinite
dilution of the effluent with upstream water, there will
be adequate protection. The concern, therefore, is for
intermediate levels of dilution. Even if the assimilative
capacity is zero, as in the second example, there is more
concern at the lower or intermediate flows, when the
effluent load is still a major portion of the total load,
than at higher flows when the effluent load is a minor
contribution.
The Options
To ensure adequate protection over a range of flows, two
types of WERs need to be determined:
Type 1 WERs are determined by obtaining samples of
effluent and upstream water when the downstream
flow is between one and two times higher than
what it would be under design-flow conditions.
Type 2 WERs are determined by obtaining samples of
effluent and upstream water when the downstream
flow is between two and ten times higher than
what it would be under design-flow conditions.
The only difference between the two types of samples is
the downstream flow at the time the samples are taken.
For both types of WERs, the samples should be mixed at the
ratio of the flows that existed when the samples were
taken so that seasonal and flow-related changes in the
water quality characteristics of the upstream water are
properly related to the flow at which they occurred. The
ratio at which the samples are mixed does not have to be
the exact ratio that existed when the samples were taken,
but the ratio has to be known, which is why simulated
downstream water is used. For each Type 1 WER and each
Type 2 WER that is determined, a hWER is calculated.
35
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Ideally, sufficient numbers of both types of WERs would be
available and each WER would be sufficiently precise and
accurate and the Type 1 WERs would be sufficiently similar
that the FWER could be the geometric mean of the Type 1
WERs, unless the FWER had to be lowered because of one or
more hWERs. If an adequate number of one or both types of
WERs is not available, an environmentally conservative WER
or hWER should be used as the FWER.
Three Type 1 and/or Type 2 WERs, which were determined
using acceptable procedures and for which there were at
least three weeks between any two sampling events, must be
available in order for a FWER to be derived. If three or
more are available, the FWER should be derived from the
WERs and hWERs using the lowest numbered option whose
requirements are satisfied:
1. If there are two or more Type 1 WERs:
a. If at least nineteen percent of all of the WERs are
Type 2 WERs, the derivation of the FWER depends on
the properties of the Type 1 WERs:
1) If the range of the Type 1 WERs is not greater
than a factor of 5 and/or the range of the ratios
of the Type 1 WER to the concentration of metal
in the simulated downstream water is not greater
than a factor of 5, the FWER is the lower of (a)
the adjusted geometric mean (see Figure 2) of all
of the Type 1 WERs and (b) the lowest hWER.
2) If the range of the Type 1 WERs is greater than a
factor of 5 and the range of the ratios of the
Type 1 WER to the concentration of metal in the
simulated downstream water is greater than a
factor of 5, the FWER is the lowest of (a) the
lowest Type 1 WER, (b) the lowest hWER, and (c)
the geometric mean of all the Type 1 and Type 2
WERs, unless an analysis of the joint
probabilities of the occurrences of WERs and
metal concentrations indicates that a higher WER
would still provide the level of protection
intended by the criterion. (EPA intends to
provide guidance concerning such an analysis.)
b. If less than nineteen percent of all of the WERs are
Type 2 WERs, the FWER is the lower of (1) the lowest
Type 1 WER and (2) the lowest hWER.
2. If there is one Type 1 WER, the FWER is the lowest of
(a) the Type 1 WER, (b) the lowest hWER, and (c) the
geometric mean of all of the Type 1 and Type 2 WERs.
3. If there are no Type 1 WERs, the FWER is the lower of
(a) the lowest Type 2 WER and (b) the lowest hWER.
If fewer than three WERs are available and a site-specific
criterion is to be derived using a WER or a FWER, the WER
or FWER has to be assumed to be 1. Examples of deriving
FWERs using these options are presented in Figure 3.
36
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The options are designed to ensure that:
a. The options apply equally well to ordinary flowing
waters and to streams whose design flow is zero.
b. The requirements for deriving the FWER as something
other than the lowest WER are not too stringent.
c. The probability is high that the criterion will be
adequately protective at all flows, regardless of the
amount of data that are available.
d. The generation of both types of WERs is encouraged
because environmental conservatism is built in if both
types of WERs are not available in acceptable numbers.
e. The amount of conservatism decreases as the quality and
quantity of the available data increase.
The requirement that three WERs be available is based on a
judgment that fewer WERs will not provide sufficient
information. The requirement that at least nineteen
percent of all of the available WERs be Type 2 WERs is
based on a judgment concerning what constitutes an
adequate mix of the two types of WERs: when there are five
or more WERs, at least one-fifth should be Type 2 WERs.
Because each of these options for deriving a FWER is
expected to provide adequate protection, anyone who
desires to determine a FWER can generate three or more
appropriate WERs and use the option that corresponds to
the WERs that are available. The options that utilize the
least useful WERs are expected to provide adequate
protection because of the way the FWER is derived from the
WERs. It is intended that, on the average, Option la will
result in the highest FWER, and so it is recommended that
data generation should be designed to satisfy the
requirements of this option if possible. For example, if
two Type 1 WERs have been determined, determining a third
Type 1 WER will require use of Option Ib, whereas
determining a Type 2 WER will require use of Option la.
Calculation of the FWER as an adjusted geometric mean
raises three issues:
a. The level of protection would be greater if the lowest
WER, rather than an adjusted mean, were used as the
FWER. Although true, the intended level of protection
is provided by the national aquatic life criterion
derived according to the national guidelines; when
sufficient data are available and it is clear how the
data should be used, there is no reason to add a
substantial margin of safety and thereby change the
intended level of protection. Use of an adjusted
geometric mean is acceptable if sufficient data are
available concerning the WER to demonstrate that the
adjusted geometric mean will provide the intended level
of protection. Use of the lowest of three or more WERs
would be justified, if, for example, the criterion had
37
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been lowered to protect a commercially important
species and a WER determined with that species was
lower than WERs determined with other species.
b. The level of protection would be greater if the
adjustment was to a probability of 0.95 rather than to
a probability of 0.70. As above, the intended level of
protection is provided by the national aquatic life
criterion derived according to the national guidelines.
There is no need to substantially increase the level of
protection when site-specific criteria are derived.
c. It would be easier to use the more common arithmetic
mean, especially because the geometric mean usually
does not provide much more protection than the
arithmetic mean. Although true, use of the geometric
mean rather than the arithmetic mean is justified on
the basis of statistics and mathematics; use of the
geometric mean is also consistent with the intended
level of protection. Use of the arithmetic mean is
appropriate when the values can range, from minus
infinity to plus infinity. The geometric mean (GM) is
equivalent to using the arithmetic mean of the
logarithms of the values. WERs cannot be negative, but
the logarithms of WERs can. The distribution of the
logarithms of WERs is therefore more likely to be
normally distributed than is the distribution of the
WERs. Thus, it is better to use the GM of WERs. In
addition, when dealing with quotients, use of the GM
reduces arguments about the correct way to do some
calculations because the same answer is obtained in
different ways. For example, if WER1 = (Nl)/(Dl) and
WER2 = (N2)/(D2), then the GM of WER1 and WER2 gives
the same value as [(GM of Nl and N2)/(GM of Dl and D2)]
and also equals the square root of
{[(Nl) (N2)]/[(D1) (D2)] }.
Anytime the FWER is derived as the lowest of a series of
experimentally determined WERs and/or hWERs, the magnitude
of the FWER will depend at least in part: on experimental
variation. There are at least three ways that the
influence of experimental variation on the FWER can be
reduced:
a. A WER determined with a primary test can be replicated
and the geometric mean of the replicates used as the
value of the WER for that determination. Then the FWER
would be the lowest of a number of geometric means
rather than the lowest of a number of individual WERs.
To be true replicates, the replicate determinations of
a WER should not be based on the same test in
laboratory dilution water, the same sample of site
water, or the same sample of effluent.
b. If, for example, Option 3 is to be used with three Type
2 WERs and the endpoints of both the primary and
38
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secondary tests in laboratory dilution water are above
the CMC and/or CCC to which the WER is to apply, WERs
can be determined with both the primary and secondary
tests for each of the three sampling times. For each
sampling time, the geometric mean of the WER obtained
with the primary test and the WER obtained with the
secondary test could be calculated; then the lowest of
these three geometric means could be used as the FWER.
The three WERs cannot consist of some WERs determined
with one of the tests and some WERs determined with the
other test; similarly the three WERs cannot consist of
a combination of individual WERs obtained wirth-the '
primary and/or secondary tests and geometric means of
results of primary and secondary tests.
c. As mentioned above, because the variability of the
effluent might contribute substantially to the
variability of the WERs, it might be desirable to
obtain and store more than one sample of the effluent
when a WER is to be determined in case an unusual WER
is obtained with the first sample used.
Examples of the first and second ways of reducing the
impact of experimental variation are presented in Figure
4. The availability of these alternatives does not mean
that they are necessarily cost-effective.
6. For metals whose criteria are hardness-dependent, at what
hardness should WERs be determined?
The issue of hardness bears on such topics as acclimation
of test organisms to the site water, adjustment of the
hardness of the site water, and how an experimentally
determined WER should be used. If all WERs were
determined at design-flow conditions, it might seem that
all WERs should be determined at the design-flow hardness.
Some permit limits, however, are not based on the hardness
that is most likely to occur at design flow; in addition,
conducting all tests at design-flow conditions provides no
information concerning whether adequate protection will be
provided at other flows. Thus, unless the hardnesses of
the upstream water and the effluent are similar and do not
vary with flow, the hardness of the site water will not be
the same for all WER determinations.
Because the toxicity tests should be begun within 36 hours
after the samples1 of effluent and upstream water are
collected, there is little time to acclimate organisms to
a sample-specific hardness. One alternative would be to
acclimate the organisms to a preselected hardness and then
adjust the hardness of the site water, but adjusting the
hardness of the site water might have various effects on
the toxicity of the metal due to competitive binding and
ionic impacts on the test organisms and on the speciation
39
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of the metal; lowering hardness without also diluting the
WER is especially problematic. The least objectionable
approach is to acclimate the organisms to a laboratory
dilution water with a hardness in the range of 50 to 150
mg/L and then use this water as the laboratory dilution
water when the WER is determined. In this way, the test
organisms will be acclimated to the laboratory dilution
water as specified by ASTM (1993a,b,c,d,e).
Test organisms may be acclimated to the site water for a
short time as long as this does not cause the tests to
begin more than 36 hours after the samples were collected.
Regardless of what acclimation procedure is used, the
organisms used for the toxicity test conducted using site
water are unlikely to be acclimated as well as would be
desirable. This is a general problem with toxicity tests
conducted in site water (U.S. EPA 1993a,b,c; ASTM 1993f),
and its impact on the results of tests is unknown.
For the practical reasons given above, an experimentally
determined WER will usually be a ratio of endpoints
determined at two different hardnesses and will thus
include contributions from a variety of differences
between the two waters, including hardness. The
disadvantages of differing hardnesses are that (a) the
test organisms probably will not be adequately acclimated
to site water and (b) additional calculations will be
needed to account for the differing hardnesses; the
advantages are that it allows the generation of data
concerning the adequacy of protection at various flows of
upstream water and it provides a way of overcoming two
problems with the hardness equations: (1) it is not known
how applicable they are to hardnesses outside the range of
25 to 400 mg/L and (2) it is not known how applicable they
are to unusual combinations of hardness, alkalinity, and
pH or to unusual ratios of calcium and magnesium.
The additional calculations that are necessary to account
for the differing hardnesses will also overcome the
shortcomings of the hardness equations. The purpose of
determining a WER is to determine how much metal can be in
a site water without lowering the intended level of
protection. Each experimentally determined WER is
inherently referenced to the hardness of the laboratory
dilution water that was used in the determination of the
WER, but the hardness equation can be used to calculate
adjusted WERs that are referenced to other hardnesses for
the laboratory dilution water. When used to adjust WERs,
a hardness equation for a CMC or CCC can be used to
reference a WER to any hardness for a laboratory dilution
water, whether it is inside or outside the range of 25 to
400 mg/L, because any inappropriateness in the equation
40
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will be automatically compensated for when the adjusted
WER is used in the derivation of a FWER and permit limits.
For example, the hardness equation for the freshwater CMC
for copper gives CMCs of 9.2, 18, and 34 ug/L at
hardnesses of 50, 100, and 200 mg/L, respectively. If
acute toxicity tests with Ceriodaphnia reticulata gave an
EC50 of 18 ug/L using a laboratory dilution water with a
hardness of 100 mg/L and an EC50 of 532.2 ug/L in a site
water, the resulting WER would be 29.57. It can be
assumed that, within experimental variation, ECSOs of 9.2
and 34 ug/L and WERs of 57.85 and 15.65 would have been
obtained if laboratory dilution waters with hardnesses of
50 and 200 mg/L, respectively, had been used, because the
EC50 of 532.2 ug/L obtained in the site water does not
depend on what water is used for the laboratory dilution
water. The WERs of 57.85 and 15.65 can be considered to
be adjusted WERs that were extrapolated from the
experimentally determined WER using the hardness equation
for the copper CMC. If used correctly, the experimentally
determined WER and all of the adjusted WERs will result in
the same permit limits because they are internally
consistent and are all based on the EC50 of 532.2 ug/L
that was obtained in site water.
A hardness equation for copper can be used to adjust the
WER if the hardness of the laboratory dilution water used
in the determination of the WER is in the range of 25 to
400 mg/L (preferably in the range of about 40 to 250 mg/L
because most of the data used to derive the equation are
in this range). However, the hardness equation can be
used to adjust WERs to hardnesses outside the range of 25
to 400 mg/L because the basis of the adjusted WER does not
change the fact that the EC50 obtained in site water was
532.2 ug/L. If the hardness of the site water was 16
mg/L, the hardness equation would predict an EC50 of 3.153
ug/L, which would result in an adjusted WER of 168.8.
This use of the hardness equation outside the range of 25
to 400 mg/L is valid only if the calculated CMC is used
with the corresponding adjusted WER. Similarly, if the
hardness of the site water had been 447 mg/L, the hardness
equation would predict an EC50 of 72.66 ug/L, with a
corresponding adjusted WER of 7.325. If the hardness of
447 mg/L were due to an effluent that contained calcium
chloride and the alkalinity and pH of the site water were
what would usually occur at a hardness of 50 mg/L rather
than 400 mg/L, any inappropriateness in the calculated
EC50 of 72.66 ug/L will be compensated for in the adjusted
WER of 7.325, because the adjusted WER is based on the
EC50 of 532.2 ug/L that was obtained using the site water.
41
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In the above examples it was assumed that at a hardness of
100 tng/L the EC50 for C. reticulata equalled the CMC,
which is a very reasonable simplifying assumption. If,
however, the WER had been determined with the more
resistant Daphnia pulex and ECSOs of 50 ug/L and 750 ug/L
had been obtained using a laboratory dilution water and a
site water, respectively, the CMC given by the hardness
equation could not be used as the predicted EC50. A new
equation would have to be derived by changing the
intercept so that the new equation gives an EC50 of 50
ug/L at a hardness of 100 mg/L; this new equation could
then be used to calculate adjusted ECSOs, which could then
be used to calculate corresponding adjusted WERs:
Hardness EC50 WER
(mcr/L) fucr/L)
16 8.894 84,33
50 26.022 28.82
100 50.000* 15.00*
200 96.073 7.81
447 204.970 3.66
The values marked with an asterisk are the assumed
experimentally determined values; the others were
calculated from these values . At each hardness the
product of the EC50 times the WER equals 750 ug/L because
all of the WERs are based on the same EC50 obtained using
site water. Thus use of the WER allows explication of the
hardness equation for a metal to conditions to which it
otherwise might not be applicable.
HCMEs can then be calculated using either the
experimentally determined WER or an adjusted WER as long
as the WER is applied to the CMC that corresponds to the
hardness on which the WER is based. For example, if the
concentration of copper in the upstream water was 1 ug/L
and the flows of the effluent and upstream water were 9
and 73 cfs, respectively, when the samples were collected,
the HCME calculated from the WER of 15.00 would be:
HCME = (17.73 ug/L) (15) (9 + 73 cfs) - (1 ug/L) (73 cfs) = 2415 u
9
because the CMC is 17.73 ug/L at a hardness of 100 mg/L.
(The value of 17.73 ug/L is used for the CMC instead of 18
ug/L to reduce roundoff error in this example.) If the
hardness of the site water was actually 447 ug/L, the HCME
could also be calculated using the WER of 3.66 and the CMC
of 72.66 ug/L that would be obtained from the CMC hardness
equation:
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HOME = (72-66 "a/*-) (3.66) (9 + 73 cfs) - (1 ug/L) (73 cfs) = 2415 ug/L _
9 C-fs
Either WER can be used in the calculation of the HOME as
long as the CMC and the WER correspond to the same
hardness and therefore to each other, because:
(17.73 ug/L) (15) = (72.66 ug/L) (3.66) .
Although the HCME will be correct as long as the hardness,
CMC, and WER correspond to each other, the WER used in the
derivation of the FWER must be the one that is calculated
using a hardness equation to be compatible with the
hardness of the site water. If the hardness of the site
water was 447 ug/L, the WER used in the derivation of the
FWER has to be 3.66; therefore, the simplest approach is
to calculate the HCME using the WER of 3.66 and the
corresponding CMC of 72.66 ug/L, because these correspond
to the hardness of 447 ug/L, which is the hardness of the
site water.
In contrast, the hWER should be calculated using the CMC
that corresponds to the design hardness. If the design
hardness is 50 mg/L, the corresponding CMC is 9.2 ug/L.
If the design flows of the effluent and the upstream water
are 9 and 20 cfs, respectively, and the concentration of
metal in upstream water at design conditions is 1 ug/L,
the hWER obtained from the WER determined using the site
water with a hardness of 447 mg/L would be:
- (2415 ug/L) (9 cfs) + (1 ug/L) (20 cfs) = _.. 54
(9.2 ug/L) (9 cfs + 20 cfs) 81'b4 '
None of these calculations provides a way of extrapolating
a WER from one site-water hardness to another. The only
extrapolations that are possible are from one hardness of
laboratory dilution water to another; the adjusted WERs
are based on predicted toxicity in laboratory dilution
water, but they are all based on measured toxicity in site
water. If a WER is to apply to the design flow and the
design hardness, one or more toxicity tests have to be
conducted using samples of effluent and upstream water
obtained under design-flow conditions and mixed at the
design-flow ratio to produce the design hardness. A WER
that is specifically appropriate to design conditions
cannot be based on predicted toxicity in site water; it
has to be based on measured toxicity in site water that
corresponds to design-flow conditions. The situation is
more complicated if the design hardness is not the
hardness that is most likely to occur when effluent and
upstream water are mixed at the ratio of the design flows.
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B. Background Information and Initial Decisions
1. Information should be obtained concerning the effluent and
the operating and discharge schedules of the discharger.
2. The spatial extent of the site to which the WER and the
site-specific criterion are intended to apply should be
defined (see Appendix A). Information concerning
tributaries, the plume, and the point of complete mix
should be obtained. Dilution models (U.S. EPA 1993d) and
dye dispersion studies (Kilpatrick 1992) might provide
information that is useful for defining sites for cmcWERs.
3. If the Recalculation Procedure (see Appendix B) is to be
used, it should be performed.
4. Pertinent information concerning the calculation of the
permit limits should be obtained:
a. What are the design flows, i.e., the flow of the
upstream water (e.g., 7Q10) and the flow of the
effluent that are used in the calculation of the permit
limits? (The design flows for the CMC and CCC might be
the same or different.)
b. Is there a CMC (acute) mixing zone and/or a CCC
(chronic) mixing zone?
c. What are the dilution(s) at the edge(s) of the mixing
zone(s)?
d. If the criterion is hardness-dependent, what is the
hardness on which the permit limits are based? Is this
a hardness that is likely to occur under design-flow
conditions?
5. It should be decided whether to determine a cmcWER and/or
a cccWER.
6. The water quality criteria document (see Appendix E) that
serves as the basis of the aquatic life criterion should
be read to identify any chemical or toxicological
properties of the metal that are relevant.
7. If the WER is being determined by or for a discharger, it
will probably be desirable to decide what is the smallest
WER that is desired by the discharger (e.g., the smallest
WER that would not require a reduction in the amount of
metal discharged). This "smallest desired WER" might be
useful when deciding whether to determine a WER. If a WER
is determined, this "smallest desired WER" might be useful
when selecting the range of concentrations to be tested in
the site water.
8. Information should be read concerning health and safety
considerations regarding collection and handling of
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effluent and surface water samples and conducting toxicity
tests (U.S. EPA l-993a; ASTM 1993a) . Information should
also be read concerning safety and handling of the
metallic salt that will be used in the preparation of the
stock solution.
9. The proposed work should be disqussed with the appropriate
regulatory authority (and possibly the Water Management
Division of the EPA Regional Office) before deciding how
to proceed with the development of a detailed workplan.
10. Plans should be made to perform one or more rangefinding
tests in both laboratory dilution water and site water
(see section G.7).
C. Selecting Primary and Secondary Tests
1. For each WER (cmcWER and/or cccWER) to be determined, the
primary and secondary tests should be selected using the
rationale presented in section A.3, the information in
Appendix I, the information in the criteria document for
the metal (see Appendix E), and any other pertinent
information that is available. When a specific test
species is not specified, also select the species.
Because at least three WERs must be determined with the
primary test, but only one must be determined with the
secondary test, selection of the tests might be influenced
by the availability of the species (and the life stage in
some cases) during the planned testing period.
a. The description of a "test" specifies not only the test
species and the duration of the test but also the life
stage of the species and the adverse effect on which
the results are to be based, all of which can have a
major impact on the sensitivity of the test.
b. The endpoint (e.g., LC50, EC50, IC50) of the primary
test in laboratory dilution water should be as close as
possible, but it must not be below, the CMC and/or CCC
to which the WER is to be applied, because for any two
tests, the test that has the lower endpoint is likely
to give the higher WER (see Appendix D).
NOTE: If both the Recalculation Procedure and a WER are
to be used in the derivation of the site-specific
criterion, the Recalculation Procedure must be
completed first because the recalculated CMC
and/or CCC must be used in the selection of the
primary and secondary tests.
c. The endpoint (e.g., LC50, EC50, IC50) of the secondary
test in laboratory dilution water should be as close as
possible, but may be above or below, the CMC and/or CCC
to which the WER is to be applied.
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1) Because few toxicity tests have endpoints close to
the CMC and CCC and because the major use of the
secondary test is confirmation (see section I.7.b),
the endpoint of the secondary test may be below the
CMC or CCC. If the endpoint of the secondary test
in laboratory dilution water is above the CMC and/or
CCC, it might be possible to use the results to
reduce the impact of experimental variation (see
Figure 4). If the endpoint of the primary test in
laboratory dilution water is above the CMC and the
endpoint of the secondary test is between the CMC
and CCC, it should be possible to determine both a
cccWER and a cmcWER using the same two tests.
2) It is often desirable to conduct the secondary test
when the first primary test is conducted in case the
results are surprising; conducting both tests the
first time also makes it possible to interchange the
primary and secondary tests, if desired, without
increasing the number of tests that need to be
conducted. (If results of one or more rangefinding
tests are not available, it might be desirable to
wait and conduct the secondary test when more
information is available concerning the laboratory
dilution water and the site water.)
The primary and secondary tests must be conducted with
species in different taxonomic orders; at least one
species must be an animal and, when feasible, one species
should be a vertebrate and the other should be an
invertebrate. -A plant cannot be used if nutrients and/or
chelators need to be added to either or both dilution
waters in order to determine the WER. It is desirable to
use a test and species for which the rate of success is
known to be high and for which the test organisms are
readily available. (If the WER is to be used with a
recalculated CMC and/or CCC, the species used in the
primary and secondary tests do not have to be on the list
of species that are used to obtain the recalculated CMC
and/or CCC.)
There are advantages to using tests suggested in Appendix
I or other tests of comparable sensitivity for which data
are available from one or more other laboratories.
a. A good indication of the sensitivity of the test is
available. This helps ensure that the endpoint in
laboratory dilution water is close to the CMC and/or
CCC and aids in the selection of concentrations of the
metal to be used in the rangefinding and/or definitive
toxicity tests in laboratory dilution water. Tests
with other species such as species that occur at the
site may be used, but it is sometimes more difficult to
obtain, hold, and test such species.
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b. When a WER is determined and used, the results of the
tests in laboratory dilution water provide the
connection between the data used in the derivation of
the national criterion and the data obtained in site
water, i.e., the results in laboratory dilution water
are a vital link in the derivation and use of a WER.
It is, therefore, important to be able to judge the
quality of the results in laboratory dilution water.
Comparison of results with data from other laboratories
evaluates all aspects of the test methodology
simultaneously, but for the determination of WERs, the
most important aspect is the quality of the laboratory
dilution water because the dilution water is the most
important difference between the two side-by-side tests
from which the WER is calculated. Thus, two tests must
be conducted for which data are available on the metal
of concern in a laboratory dilution water from at least
one other laboratory. If both the primary and
secondary tests are ones for which acceptable data are
available from at least one other laboratory, these are
the only two tests that have to be conducted. If,
however, the primary and/or secondary tests are ones
for which no results are already available for the
metal of concern from another laboratory, the first or
second time a WER is determined at least two additional
tests must be conducted in the laboratory dilution
water in addition to the tests that are conducted for
the determination of WERs (see sections F.5 and 1.5).
1) For the determination of a WER, data are not
required for a reference toxicant with either the
primary test or the secondary test because the above
requirement provides similar data for the metal for
which the WER is actually being determined.
2) See Section 1.5 concerning interpretation of the
results of these tests before additional tests are
conducted.
D. Acquiring and Acclimating Test Organisms
1. The test organisms should be obtained, cultured, held,
acclimated, fed, and handled as recommended by the U.S.
EPA (1993a,b,c) and/or by ASTM (1993a,b,c,d,e). All test
organisms must be acceptably acclimated to a laboratory
dilution water that satisfies the requirements given in
sections F.3 and F.4; an appropriate number of the
organisms may be randomly or impartially removed from the
laboratory dilution water and placed in the site water
when it becomes available in order to acclimate the
organisms to the site water for a while just before the
tests are begun.
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2. The organisms used in a pair of side-by-side tests must be
drawn from the same population and tested under identical
conditions.
E. Collecting and Handling Upstream Water and Effluent
1. Upstream water will usually be mixed with effluent to
prepare simulated downstream water. Upstream water may
also be used as a site water if a WER is to be determined
using upstream water in addition to or instead of
determining a WER using downstream water. The samples of
upstream water must be representative; they must not be
unduly affected by recent runoff events (or other erosion
or resuspension events) that cause higher levels of TSS
than would normally be present, unless there is particular
concern about such conditions.
2. The sample of effluent used in the determination of a WER
must be representative; it must be collected during a
period when the discharger is operating normally.
Selection of the date and time of sampling of the effluent
should take into account the discharge pattern of the
discharger. It might be appropriate to collect effluent
samples during the middle of the week to allow for
reestablishment of steady-state conditions after shutdowns
for weekends and holidays,- alternatively, if end-of-the-
week slug discharges are routine, they should probably be
evaluated. As mentioned above, because the variability of
the effluent might contribute substantially to the
variability of the WERs, it might be desirable to obtain
and store more than one sample of the effluent when WERs
are to be determined in case an unusual WER is obtained
with the first sample used.
3. When samples of site water and effluent are collected for
the determination of the WERs with the primary test, there
must be at least three weeks between one sampling' event
and the next. It is desirable to obtain samples in at
least two different seasons and/or during times of
probable differences in the characteristics of the site
water and/or effluent.
4. Samples of upstream water and effluent must be collected,
transported, handled, and stored as recommended by the
U.S. EPA (1993a). For example, samples of effluent should
usually be composites, but grab samples are acceptable if
the residence time of the effluent is sufficiently long.
A sufficient volume should be obtained so that some can be
stored for additional testing or analyses if an unusual
WER is obtained. Samples must be stored at 0 to 4°C in
the dark with no air space in the sample container.
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5. At the time of collection, the flow of both the upstream
water and the effluent must be either measured or
estimated by means of correlation with a nearby U.S.G.S.
gauge, the pH of both upstream water and effluent must be
measured, and samples of both upstream water and effluent
should be filtered for measurement of dissolved metals.
Hardness, TSS, TOG, and total recoverable and dissolved
metal must be measured in both the effluent and the
upstream water. Any other water quality characteristics,
such as total dissolved solids (TDS) and conductivity,
that are monitored monthly or more often by the permittee
and reported in the Discharge Monitoring Report must also
be measured. These and the other measurements provide
information concerning the representativeness of the
samples and the variability of the upstream water and
effluent.
6. "Chain of custody" procedures (U.S. EPA 1991b) should be
used for all samples of site water and effluent,
especially if the data might be involved in a legal
proceeding.
7. Tests must be begun within 36 hours after the collection
of the samples of the effluent and/or the site water,
except that tests may be begun more than 36 hours after
the collection of the samples if it would require an
inordinate amount of resources to transport the samples to
the laboratory and begin the tests within 36 hours.
8. If acute and/or chronic tests are to be conducted with
daphnids and if the sample of the site water contains
predators, the site water must be filtered through a 37-/zm
sieve or screen to remove predators.
F. Laboratory Dilution Water
1. The laboratory dilution water must satisfy the
requirements given by U.S. EPA (1993a,b,c) or ASTM
(1993a,b,c,d,e). The laboratory dilution water must be a
ground water, surface water, reconstituted water, diluted
mineral water, or dechlorinated tap water that has been
demonstrated to be acceptable to aquatic organisms. If a
surface water is used for acute or chronic tests with
daphnids and if predators are observed in the sample of
the water, it must be filtered through a 37-/zm sieve or
screen to remove the predators. Water prepared by such
treatments as deionization and reverse osmosis must not be
used as the laboratory dilution water unless salts,
mineral water, hypersaline brine, or sea salts are added
as recommended by U.S. EPA (1993a) or ASTM (1993a).
49
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2. The concentrations of both TOC and TSS must be less than 5
mg/L.
3. The hardness of the laboratory dilution water should be
between 50 and 150 mg/L and must be between 40 and 220
mg/L. If the criterion for the metal is hardness-
dependent, the hardness of the laboratory dilution water
must not be above the hardness of the site water, unless
the hardness of the site water is below 50 mg/L.
4. The alkalinity and pH of the laboratory dilution water
must be appropriate for its hardness; values for
alkalinity and pH that are appropriate for some hardnesses
are given by U.S. EPA (1993a) and ASTM (1993a); other
corresponding values should be determined by
interpolation. Alkalinity should be adjusted using sodium
bicarbonate, and pH should be adjusted using aeration,
sodium hydroxide, and/or sulfuric acid.
5. It would seem reasonable that, before any samples of site
water or effluent are collected, the toxicity tests that
are to be conducted in the laboratory dilution water for
comparison with results of the same tests from other
laboratories (see sections C.3.b and 1.5) should be
conducted. These should be performed at the hardness,
alkalinity, and pH specified in sections F.3 and F.4.
G. Conducting Tests
1. There must be no differences between the side-by-side
tests other than the composition of the dilution water,
the concentrations of metal tested, and possibly the water
in which the test organisms are acclimated just prior to
the beginning of the tests.
2. More than one test using site water may be conducted side-
by-side with a test using laboratory dilution water; the
one test in laboratory dilution water will be used in the
calculation of several WERs, which means that it is very
important that that one test be acceptable.
3. Facilities for conducting toxicity tests should be set up
and test chambers should be selected and cleaned as
recommended by the U.S. EPA (1993a,b,c) and/or ASTM
(1993a,b,c,d,e).
4. A stock solution should be prepared using an inorganic
salt that is highly soluble in water.
a. The salt does not have to be one that was used in tests
that were used in the derivation of the national
criterion. Nitrate salts are generally acceptable;
50
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chloride and sulfate salts of many metals are also
acceptable (see Appendix J). It is usually desirable
to avoid use of a hygroscopic salt. The salt used
should meet A.C.S. specifications for reagent-grade, if
such specifications are available; use of a better
grade is usually not worth the extra cost. No salt
should be used until information concerning safety and
handling has been read.
b. The stock solution may be acidified (using metal-free
nitric acid) only as necessary to get the metal into
solution.
c. The same stock solution must be used to add metal to
all tests conducted at one time.
5. For tests suggested in Appendix I, the appendix presents
the recommended duration and whether the static or renewal
technique should be used; additional information is
available in the references cited in the appendix.
Regardless of whether or not or how often test solutions
are renewed when these tests are conducted for other
purposes, the following guidance applies to all tests that
are conducted for the determination of WERs:
a. The renewal technique must be used for tests that last
longer than 48 hr.
b. If the concentration of dissolved metal decreases by
more than 50 % in 48 hours in static or renewal tests,
the test solutions must be renewed every 24 hours.
Similarly, if the concentration of dissolved oxygen
becomes too low, the test solutions must be renewed
every 24 hours. If one test in a pair of tests is a
renewal test, both tests must be renewal tests.
c. When test solutions are to be renewed, the new test
solutions must be prepared from the original unspiked
effluent and water samples that have been stored at 0
to 4°C in the dark with no air space in the sample
container.
d. The static technique may be used for tests that do not
last longer than 48 hours unless the above
specifications require use of the renewal technique.
If a test is used that is not suggested in Appendix I, the
duration and technique recommended for a comparable test
should be used.
6. Recommendations concerning temperature, loading, feeding,
dissolved oxygen, aeration, disturbance, and controls
given by the U.S. EPA (1993a,b,c) and/or ASTM
(1993a,b,c,d,e) must be followed. The procedures that are
used must be used in both of the side-by-side tests.
7. To aid in the selection of the concentrations of metals
that should be used in the test solutions in site water, a
static rangefinding test should be conducted for 8 to 96
51
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hours, using a dilution factor of 10 (or 0.1) or 3.2 (or
0.32) increasing from about a factor of 10 below the value
of the endpoint given in the criteria document for the
metal or in Appendix I of this document for tests with
newly hatched fathead minnows. If the test is not in the
criteria document and no other data are available, a mean
acute value or other data for a taxonomically similar
species should be used as the predicted value. This
rangefinding test will provide information concerning the
concentrations that should be used to bracket the endpoint
in the definitive test and will provide information
concerning whether the control survival will be
acceptable. If dissolved metal is measured in one or more
treatments at the beginning and end of the rangefinding
test, these data will indicate whether the concentration
should be expected to decrease by more than 50 % during
the definitive test. The rangefinding test may be
conducted in either of two ways:
a. It may be conducted using the samples of effluent and
site water that will be used in the definitive test.
In this case, the duration of the rangefinding test
should be as long as possible within the limitation
that the definitive test must begin within 36 hours
after the samples of effluent and/or site water were
collected, except as per section E.7.
b. It may be conducted using one set of samples of
effluent and upstream water with the definitive tests
being conducted using samples obtained at a later date.
In this case the rangefinding test might give better
results because it can last longer, but there is the
possibility that the quality of the effluent and/or
site water might change. Chemical analyses for
hardness and pH might indicate whether any major
changes occurred from one sample to the next.
Rangefinding tests are especially desirable before the
first set of toxicity tests. It might be desirable to
conduct rangefinding tests before each individual
determination of a WER to obtain additional information
concerning the effluent, dilution water, organisms, etc.,
before each set of side-by-side tests are begun.
8. Several considerations are important in the selection of
the dilution factor for definitive tests. Use of
concentrations that are close together will reduce the
uncertainty in the WER but will require more
concentrations to cover a range within which the endpoints
might occur. Because of the resources necessary to
determine a WER, it is important that endpoints in both
dilution waters be obtained whenever a set of side-by-side
tests are conducted. Because static and renewal tests can
be used to determine WERs, it is relatively easy to use
more treatments than would be used in flow-through tests.
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The dilution factor for total recoverable metal must be
between 0.65 and 0.99, and the recommended factor is 0.7.
Although factors between 0.75 and 0.99 may be used, their
use will probably not be cost-effective. ..Because there is
likely to be more uncertainty in the predicted value of
the endpoint in site water, 6 or 7 concentrations are
recommended in the laboratory dilution water, and 8 or 9
in the simulated downstream water, at a dilution factor of
0.7. It might be desirable to use even more treatments in
the first of the WER determinations, because the design of
subsequent tests can be based on the results of the first
tests if the site water, laboratory dilution water, and
test organisms do not change too much. The cost of adding
treatments can be minimized if the concentration of metal
is measured only in samples from treatments that will be
used in the calculation of the endpoint.
9. Each test must contain a dilution-water control. The
number of test organisms intended to be exposed to each
treatment, including the controls, must be at least 20.
It is desirable that the organisms be distributed between
two or more test chambers per treatment. If test
organisms are not randomly assigned to the test chambers,
they must be assigned impartially (U.S. EPA 1993a; ASTM
1993a) between all test chambers for a pair of side-by-
side tests. For example, it is not acceptable to assign
20 organisms to one treatment, and then assign 20
organisms to another treatment, etc. Similarly, it is not
acceptable to assign all the organisms to the test using
one of the dilution waters and then assign organisms to
the test using the other dilution water. The test
chambers should be assigned to location in a totally
random arrangement or in a randomized block design.
10. For the test using site water, one of the following
procedures should be used to prepare the test solutions
for the test chambers and the "chemistry controls" (see
section H.I):
a. Thoroughly mix the sample of the effluent and place the
same known volume of the effluent in each test chamber;
add the necessary amount of metal, which will be
different for each treatment; mix thoroughly; let stand
for 2 to 4 hours; add the necessary amount of upstream
water to each test chamber; mix thoroughly; let stand
for 1 to 3 hours.
b. Add the necessary amount of metal to a large sample of
the effluent and also maintain an unspiked sample of
the effluent; perform serial dilution using a graduated
cylinder and the well-mixed spiked and unspiked samples
of the effluent; let stand for 2 to 4 hours; add the
necessary amount of upstream water to each test
chamber; mix thoroughly; let stand for 1 to 3 hours.
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c. Prepare a large volume of simulated downstream water by
mixing effluent and upstream water in the desired
ratio; place the same known volume of the simulated
downstream water in each test chamber; add the
necessary amount of metal, which will be different for
each treatment; mix thoroughly and let stand for 1 to 3
hours.
d. Prepare a large volume of simulated downstream water by
mixing effluent and upstream water in the desired
ratio; divide it into two portions; prepare a large
volume of the highest test concentration of metal using
one portion of the simulated downstream water; perform
serial dilution using a graduated cylinder and the
well-mixed spiked and unspiked samples of the simulated
downstream water; let stand for 1 to 3 hours.
Procedures "a" and "b" allow the metal to equilibrate
somewhat with the effluent before the solution is diluted
with upstream water.
11. For the test using the laboratory dilution water, either
of the following procedures may be used to prepare the
test solutions for the test chambers and the "chemistry
controls" (see section H.I):
a. Place the same known volume of the laboratory dilution
water in each test chamber; add the necessary amount of
metal, which will be different for each treatment; mix
thoroughly; let stand for 1 to 3 hours.
b. Prepare a large volume of the highest test
concentration in the laboratory dilution water; perform
serial dilution using a graduated cylinder and the
well-mixed spiked and unspiked samples of the
laboratory dilution water; let stand for 1 to 3 hours.
12. The test organisms, which have been acclimated as per
section D.I, must be added to the test chambers for the
site-by-side tests at the same time. The time at which
the test organisms are placed in the test chambers is
defined as the beginning of the tests, which must be
within 36 hours of the collection of the samples, except
as per section E.7.
13. Observe the test organisms and record the effects and
symptoms as specified by the U.S. EPA (1993a,b,c) and/or
ASTM (1993a,b,c,d,e). Especially note whether the
effects, symptoms, and time course of toxicity are the
same in the side-by-side tests.
14. Whenever solutions are renewed, sufficient solution should
be prepared to allow for chemical analyses.
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H. Chemical and Other Measurements
1. To reduce the possibility of contamination of test
solutions before or during tests, thermometers and probes
for measuring pH and dissolved oxygen must not be placed
in test chambers that will provide data concerning effects
on test organisms or data concerning the concentration of
the metal. Thus measurements of pH, dissolved oxygen, and
temperature before or during a test must be performed
either on "chemistry controls" that contain test organisms
and are fed the same as the other test chambers or on
aliquots that are removed from the test chambers. The
other measurements may be performed on the actual test
solutions at the beginning and/or end of the test or the
renewal.
2. Hardness (in fresh water) , or salinity (in salt water), pH,
alkalinity, TSS, and TOG must be measured on the upstream
water, the effluent, the simulated and/or actual
downstream water, and the laboratory dilution water.
Measurement of conductivity and/or total dissolved solids
(TDS) is recommended in fresh water.
3. Dissolved oxygen, pH, and temperature must be measured
during the test at the times specified by the U.S. EPA
(1993a,b,c) and/or ASTM (1993a,b,c,d,e). The measurements
must be performed on the same schedule for both of the
side-by-side tests. Measurements must be performed on
both the chemistry controls and actual test solutions at
the end of the test. .
4. Both total recoverable and dissolved metal must be
measured in the upstream water, the effluent, and
appropriate test solutions for each of the tests.
a. The analytical measurements should be sufficiently
sensitive and precise that variability in analyses will
not greatly increase the variability of the WERs. If
the detection limit of the analytical method that will
be used to determine the metal is greater than one-
tenth of the CCC or CMC that is to be adjusted, the
analytical method should probably be improved or
replaced (see Appendix C). If additional sensitivity
is needed, it is often useful to separate the metal
from the matrix because this will simultaneously
concentrate the metal and remove interferences.
Replicate analyses should be performed if necessary to
reduce the impact of analytical variability.
1) EPA methods (U.-S..EPA 1983b,1991c) should usually be
used for both total recoverable and dissolved
measurements, but in some cases alternate methods
might have to be used in order to achieve the
necessary sensitivity. Approval for use of
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alternate methods is to be requested from the
appropriate regulatory authority.
b. All measurements of metals must be performed using
appropriate QA/QC techniques. Clean techniques for
obtaining, handling, storing, preparing, and analyzing
the samples should be used when necessary to^achieve
blanks that are sufficiently low (see Appendix C).
c. Rather than measuring the metal in all test solutions,
it is often possible to store samples and then analyze
only those that are needed to calculate the results of
the toxicity tests. For dichotomous data (e.g.,
either-or data; data concerning survival), the metal in
the following must be measured:
1) all concentrations in which some, but not all, of
the test organisms were adversely affected.
2) the highest concentration that did not adversely
affect any test organisms.
3) the lowest concentration that adversely affected all
of the test organisms.
4) the controls.
For data that are not dichotomous (i.e., for count and
continuous data), the metal in the controls and in the
treatments that define the concentration-effect curve
must be measured; measurement of the concentrations of
metals in other treatments is desirable.
d. In each treatment in which the concentration of metal
is to be measured, both the total recoverable and
dissolved concentrations must be measured:
1) Samples must be taken for measurement of total
recoverable metal once for a static test, and once
for each renewal for renewal tests; in renewal
tests, the samples are to be taken after the
organisms have been transferred to the new test
solutions. When total recoverable metal is measured
in a test chamber, the whole solution in the chamber
must be mixed before the sample is taken for
analysis; the solution in the test chamber must not
be acidified before the sample is taken. The sample
must be acidified after it is placed in the sample
container.
2) Dissolved metal must be measured at the beginning
and end of each static test; in a renewal test, the
dissolved metal must be measured at the beginning of
the test and just before the solution is renewed the
first time. When dissolved metal is measured in a
test chamber, the whole solution in the test chamber
must be mixed before a sufficient amount is removed
for filtration; the solution in the test chamber
must not be acidified before the sample is taken.
The sample must be filtered within one hour after it
is taken, and the filtrate must be acidified after
filtration.
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5. Replicates, matrix spikes, and other QA/QC checks must be
performed as required by the U.S. EPA (1983a,1991c).
I. Calculating and Interpreting the Results
1. To prevent roundoff error in subsequent calculations, at
least four significant digits must be retained in all
endpoints, WERs, and FWERs. This requirement is not based
on mathematics or statistics and does not reflect the
precision of the value; its purpose is to minimize concern
about the effects of rounding off on a site-specific
criterion. All of these numbers are intermediate values
in the calculation of permit limits and should not be
rounded off as if they were values of ultimate concern.
2. Evaluate the acceptability of each toxicity test
individually.
a. If the procedures used deviated from those specified
above, particularly in terms of acclimation,
randomization, temperature control, measurement of
metal, and/or disease or disease-treatment, the test
should be rejected; if deviations were numerous and/or
substantial, the test must be rejected.
b. Most tests are unacceptable if more than 10 percent of
the organisms in the controls were adversely affected,
but the limit is higher for some tests; for the tests
recommended in Appendix I, the references given should
be consulted.
c. If an LC50 or EC50 is to be calculated:
1) The percent of the organisms that were adversely
affected must have been less than 50 percent, and
should have been less than 37 percent, in at least
one treatment other than the control.
2) In laboratory dilution water the percent of the
organisms that were adversely affected must have
been greater than 50 percent, and should have been
greater than 63 percent, in at least one treatment.
In site water the percent of the organisms that were
adversely affected should have been greater than 63
percent in at least one treatment. (The LC50 or
EC50 may be a "greater than" or "less than" value in
site water, but not in laboratory dilution water.)
3) If there was an inversion in the data (i.e., if a
lower concentration killed or affected a greater
percentage of the organisms than a higher
concentration), it must not have involved more than
two concentrations that killed or affected between
20 and 80 percent of the test organisms.
If an endpoint other than an LC50 or EC50 is used or if
Abbott's formula is used, the above requirements will
have to be modified accordingly.
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d. Determine whether there was anything unusual about the
test results that would make them questionable.
e. If solutions were not renewed every 24 hours, the
concentration of dissolved metal must not have
decreased by more than 50 percent from the beginning to
the end of a static test or from the beginning to the
end of a renewal in a renewal test in test
concentrations that were used in the calculation of the
results of the test.
Determine whether the effects, symptoms, and time course
of toxicity was the same in the side-by-side tests in the
site water and the laboratory dilution water. For
example, did mortality occur in one acute test, but
immobilization in the other? Did most deaths occur before
24 hours in one test, but after 24 hours in the other? In
sublethal tests, was the most sensitive effect the same in
both tests? If the effects, symptoms, arid/or time course
of toxicity were different, it might indicate that the
test is questionable or that additivity, synergism, or
antagonism occurred in site water. Such information might
be particularly useful when comparing tests that produced
unusually low or high WERs with tests that produced
moderate WERs.
Calculate the results of each test:
a. If the data for the most sensitive effect are
dichotomous, the endpoint must be calculated as a LC50,
EC50, LC25, EC25, etc., using methods described by the
U.S. EPA (1993a) or ASTM (1993a). If two or more
treatments affected between 0 and 100 percent in both
tests in a side-by-side pair, probit amalysis must be
used to calculate results of both tests, unless the
probit model is rejected by the goodness of fit test in
one or both of the acute tests. If probit analysis
cannot be used, either because fewer than two
percentages are between 0 and 100 percent or because
the model does not fit the data, computational
interpolation must be used (see Figure 5); graphical
interpolation must not be used.
1} The same endpoint (LC50, EC25, etc.) and the same
computational method must be used for both tests
used in the calculation of a WER.
2) The selection of the percentage used to define the
endpoint might be influenced by the. percent effect
that occurred in the tests and the correspondence
with the CCC and/or CMC.
3) If no treatment killed or affected more than 50
percent of the test organisms and the test was
otherwise acceptable, the LC50 or EC50 should be
reported to be greater than the highest test
concentration.
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4) If no treatment other than the control killed or
affected less than 50 percent of the test organisms
and the test was otherwise acceptable, the LC50 or
EC50 should be reported to be less than the lowest
test concentration.
b. If the data for the most sensitive effect are not
dichotomous, the endpoint must be calculated using a
regression-type method (Hoekstra and Van Ewijk 1993;
Stephan and Rogers 1985) , such as linear interpolation
(U.S. EPA 1993b,c) or a nonlinear regression method
(Barnthouse et al. 1987; Suter et al. 1987; Bruce and
Versteeg 1992). The selection of the percentage used
to define the endpoint might be influenced by the
percent effect that occurred in the tests and the
correspondence with the CCC and/or CMC. The endpoints
in the side-by-side tests must be based on the same
amount of the same adverse effect so that the WER is a
ratio of identical endpoints. The same computational
method must be used for both tests used in the
calculation of the WER.
c. Both total recoverable and dissolved results should be
calculated for each test.
d. Results should be based on the time-weighted average
measured metal concentrations (see Figure 6).
The acceptability of the laboratory dilution water must be
evaluated by comparing results obtained with two sensitive
tests using the laboratory dilution water with results _
that were obtained using a comparable laboratory dilution
water in one or more other laboratories (see sections
C.3.b and F.5).
a. If, after taking into account any known effect of
hardness on toxicity, the new values for the endpoints
of both of the tests are (1) more than a factor of 1.5
higher than the respective means of the values from the
other laboratories or (2) more than a factor of 1.5
lower than the respective means of values from the
other laboratories or (3) lower than the respective
lowest values available from other laboratories or (4)
higher than the respective highest values available
from other laboratories, the new and old data must be
carefully evaluated to determine whether the laboratory
dilution water used in the WER determination was
acceptable. For example, there might have been an
error in the chemical measurements, which might mean
that the results of all tests performed in the WER
determination need to be adjusted and that the WER
would not change. It is also possible that the metal
is more or less toxic in the laboratory dilution water
used in the WER determination. Further, if the new
data were based on measured concentrations but the old
data were based on nominal concentrations, the new data
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should probably be considered to be better than the
old. Evaluation of results of any other toxicity tests
on the same or a different metal using the same
laboratory dilution water might be useful.
b. If, after taking into account any known effect of
hardness on toxicity, the new values for the endpoints
of the two tests are not either both higher or both
lower in comparison than data from other laboratories
(as per section a above) and if both of the new values
are within a factor of 2 of the respective means of the
previously available values or are within the ranges of
the values, the laboratory dilution water used in the
WER determination is acceptable.
c. A control chart approach may be used if sufficient data
are available.
d. If the comparisons do not indicate that the laboratory
dilution water, test method, etc., are acceptable, the
tests probably should be considered unacceptable,
unless other toxicity data are available to indicate
that they are acceptable.
Comparison of results of tests between laboratories
provides a check on all aspects of the test procedure; the
emphasis here is on the quality of the laboratory dilution
water because all other aspects of the side-by-side tests
on which the WER is based must be the same, except
possibly for the concentrations of metal used and the
acclimation just prior to the beginning of the tests.
6. If all the necessary tests and the laboratory dilution
water are acceptable, a WER must be calculated by dividing
the endpoint obtained using site water by the endpoint
obtained using laboratory dilution water.
a. If both a primary test and a secondary test were
conducted using both waters, WERs must be calculated
for both tests.
b. Both total recoverable and dissolved WERs must be
calculated.
c. If the detection limit of the analytical method used to
measure the metal is above the endpoint in laboratory
dilution water, the detection limit must be used as the
endpoint, which will result in a lower WER than would
be obtained if the actual concentration had been
measured. If the detection limit of the analytical
method used is above the endpoint in site water, a WER
cannot be determined.
7. Investigation of the WER.
a. The results of the chemical measurements of hardness,
alkalinity, pH, TSS, TOG, total recoverable metal,
dissolved metal, etc., on the effluent and the upstream
water should be examined and compared with previously
available values for the effluent and upstream water,
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respectively, to determine whether the samples were
representative and to get some indication of the
variability in the composition, especially as it might
affect the toxicity of the metal and the WER, and to
see if the WER correlates with one or more of the
measurements.
b. The WERs obtained with the primary and secondary tests
should be compared to determine whether the WER
obtained with the secondary test confirmed the WER
obtained with the primary test. Equally sensitive
tests are expected to give WERs that are similar (e.g.,
within a factor of 3), whereas a test that is less
sensitive will probably give a smaller WER than a more
sensitive test (see Appendix D). Thus a WER obtained
with a primary test is considered confirmed if either
or both of the following are true:
1) the WERs obtained with the primary and secondary
tests are within a factor of 3.
2) the test, regardless of whether it is the primary or
secondary test, that gives a higher endpoint in the
laboratory dilution water also gives the larger WER.
If the WER obtained with the secondary test does not
confirm the WER obtained with the primary test, the
results should be investigated. In addition, WERs
probably should be determined using both tests the next
time samples are obtained and it would be desirable to
determine a WER using a third test. It is also
important to evaluate what the results imply about the
protectiveness of any proposed site-specific criterion.
c. If the WER is larger than 5, it should be investigated.
1) If the endpoint obtained using the laboratory
dilution water was lower than previously reported
lowest value or was more than a factor of two lower
than an existing Species Mean Acute Value in a
criteria document, additional tests in the
laboratory dilution water are probably desirable.
2) If a total recoverable WER was larger than 5 but the
dissolved WER was not, is the metal one whose WER is
likely to be affected by TSS and/or TOG and was the
concentration of TSS and/or TOG high? Was there a
substantial difference between the total recoverable
and dissolved concentrations of the metal in the
downstream water?
3) If both the total recoverable and dissolved WERs
were larger than 5, is it likely that there is
nontoxic dissolved metal in the downstream water?
d. The adverse effects and the time-course of effects in
the side-by-side tests should be compared. If they are
different, it might indicate that the site-water test
is questionable or that additivity, synergism, or
antagonism occurred in the site water. This might be
especially important if the WER obtained with the
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secondary test did not confirm the WER obtained with
the primary test or if the WER was very large or small.
8. If at least one WER determined with the primary test was
confirmed by a WER that was simultaneously determined with
the secondary test, the cmcFWER and/or the cccFWER should
be derived as described in section A.5.
9. All data generated during the determination of the WER
should be examined to see if there are any implications
for the national or site-specific aquatic life criterion.
a. If there are data for a species for which data were not
previously available or unusual data for a species for
which data were available, the national criterion might
need to be revised.
b. If the primary test gives an LC50 or EC50 in laboratory
dilution water that is the same as the national CMC,
the resulting site-specific CMC should be similar to
the LC50 that was obtained with the primary test using
downstream water. Such relationships might serve as a
check on the applicability of the use of WERs.
c. If data indicate that the site-specific criterion would
not adequately protect a critical species, the site-
specific criterion probably should be lowered.
J. Reporting the Results
A report of the experimental determination of a WER to the
appropriate regulatory authority must include the following:
1. Name(s) of the investigator(s), name and location of the
laboratory, and dates of initiation and termination of the
tests.
2. A description of the laboratory dilution water, including
source, preparation, and any demonstrations that an
aquatic species can survive, grow, and reproduce in it.
3. The name, location, and description of the discharger, a
description of the effluent, and the design flows of the
effluent and the upstream water.
4. A description of each sampling station, date, and time,
with an explanation of why they were selected, and the
flows of the upstream water and the effluent at the time
the samples were collected.
5. The procedures used to obtain, transport, and store the
samples of the upstream water and the effluent.
6. Any pretreatment, such as filtration, of the effluent,
site water, and/or laboratory dilution water.
7. Results of all chemical and physical measurements on
upstream water, effluent, actual and/or simulated
downstream water, and laboratory dilution water, including
hardness (or salinity), alkalinity, pH, and concentrations
of total recoverable metal, dissolved metal, TSS, and TOC.
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8. Description of the experimental design, test chambers,
depth and volume of solution in the chambers, loading and
lighting, and numbers of organisms and chambers per
treatment.
9. Source and grade of the metallic salt, and how the stock
solution was prepared, including any acids or bases used.
10. Source of the test organisms, scientific name and how
verified, age, life stage, means and ranges of weights
and/or lengths, observed diseases, treatments, holding and
acclimation procedures, and food.
11. The average and range of the temperature, pH, hardness (or
salinity), and the concentration of dissolved oxygen (as %
saturation and as mg/L) during acclimation, and the method
used to measure them.
12. The following must be presented for each toxicity test:
a. The average and range of the measured concentrations of
dissolved oxygen, as % saturation and as mg/L.
b. The average and range of the test temperature and the
method used to measure it.
c. The schedule for taking samples of test solutions and
the methods used to obtain, prepare, and store them.
d. A summary table of the total recoverable and dissolved
concentrations of the metal in each treatment,
including all controls, in which they were measured.
e. A summary table of the values of the toxicological
variable(s) for each treatment, including all controls,
in sufficient detail to allow an independent
statistical analysis of the data.
f. The endpoint and the method used to calculate it.
g. Comparisons with other data obtained by conducting the
same test on the same metal using laboratory dilution
water in the same and different laboratories; such data
may be from a criteria document or from another source.
h. Anything unusual about the test, any deviations from
the procedures described above, and any other relevant
information.
13. All differences, other than the dilution water and the
concentrations of metal in the test solutions, between the
side-by-side tests using laboratory dilution water and
site water.
14. Comparison of results obtained with the primary and
secondary tests.
15. The WER and an explanation of its calculation.
A report of the derivation of a FWER must include the
following:
1. A report of the determination of each WER that was
determined for the derivation of the FWER; all WERs
determined with secondary tests must be reported along
with all WERs that were determined with the primary test.
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The design flow of the upstream water and the effluent and
the hardness used in the derivation of the permit limits,
if the criterion for the metal is hardness-dependent.
A summary table must be presented that contains the
following for each WER that was derived:
a. the value of the WER and the two endpoints from which
it was calculated.
b. the hWER calculated from the WER.
c. the test and species that was used.
d. the date the samples of effluent and site water were
collected.
e. the flows of the effluent and upstream water when the
samples were taken.
f. the following information concerning the laboratory
dilution water, effluent, upstream water, and actual
and/or simulated downstream water: hardness (salinity),
alkalinity, pH, and concentrations of total recoverable
metal, dissolved metal, TSS, and TOG.
A detailed explanation of how the FWER was derived from
the WERs that are in the summary table.
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METHOD 2: DETERMINING cccWERs FOR AREAS AWAY FROM PLUMES
Method 2 might be viewed as a simple process wherein samples of
site water are obtained from locations within a large body of
fresh or salt water (e.g., an ocean or a large lake, reservoir,
or estuary), a WER is determined for each sample, and the FWER is
calculated as the geometric mean of some or all of the WERs. In
reality, Method 2 is not likely to produce useful results unless
substantial resources are devoted to planning and conducting the
study. Most sites to which Method 2 is applied will have long
retention times, complex mixing patterns, and a number of
dischargers. Because metals are persistent, the long retention
times mean that the sites are likely to be defined to cover
rather large areas; thus such sites will herein be referred to
generically as "large sites". Despite the differences between
them, all large sites require similar special considerations
regarding the determination of WERs. Because Method 2 is based
on samples of actual surface water (rather than simulated surface
water), no sample should be taken in the vicinity of a plume and
the method should be used to determine cccWERs, not cmcWERs. If
WERs are to be determined for more than one metal, Appendix F
should be read.
Method 2 uses many of the same methodologies as Method 1, such as
those for toxicity tests and chemical analyses. Because the
sampling plan is crucial to Method 2 and the plan has to be based
on site-specific considerations, this description of Method 2
will be more qualitative than the description of Method 1.
Method 2 is based on use of actual surface water samples, but use
of simulated surface water might provide information that is
useful for some purposes:
1. It might be desirable to compare the WERs for two discharges
that contain the same metal. This might be accomplished by
selecting an appropriate dilution water and preparing two
simulated surface waters, one that contains a known
concentration of one effluent and one that contains a known
concentration of the other effluent. The relative magnitude
of the two WERs is likely to be more useful than the absolute
values of the WERs themselves.
2. It might be desirable to determine whether the eWER for a
particular effluent is additive with the WER of the site water
(see Appendix G). This can be studied by determining WERs for
several different known concentrations of the effluent in site
water.
3. An event such as a rain might affect the WER because of a
change in the water quality, but it might also reduce the WER
just by dilution of refractory metal or TSS. A proportional
decrease in the WER and in the concentration of the metal
(such as by dilution of refractory metal) will not result in
underprotection; if, however, dilution decreases the WER
65
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proportionally more than it decreases the concentration of
metal in the downstream water, underprotection is likely to
occur. This is essentially a determination of whether the WER
is additive when the effluent is diluted with rain water (see
Appendix G).
4. An event that increases TSS might increase the total
recoverable concentration of the metal and the total
recoverable WER without having much effect on either the
dissolved concentration or the dissolved WER.
In all four cases, the use of simulated surface water is useful
because it allows for the determination of WERs using known
concentrations of effluent.
An important step in the determination of any WER is to define
the area to be included in the site. The major principle that
should be applied when defining the area is the: same for all
sites: The site should be neither too small nor too large. If
the area selected is too small, permit limits might be
unnecessarily controlled by a criterion for an area outside the
site, whereas too large an area might unnecessarily incorporate
spatial complexities that are not relevant to the discharge(s) of
concern and thereby unnecessarily increase the cost of
determining the WER. Applying this principle is likely to be
more difficult for large sites than for flowing-water sites.
Because WERs for large sites will usually be determined using
actual, rather than simulated, surface water, there are five
major considerations regarding experimental design and data
analysis:
1. Total recoverable WERs at large sites might vary so much
across time, location, and depth that they are not very
useful. An assumption should be developed that an
appropriately defined WER will be much more similar across
time, location, and depth within the site than will a total
recoverable WER. If such an assumption cannot be used, it is
likely that either the FWER will have to be set equal to the
lowest WER and be overprotective for most of the site or
separate site-specific criteria will have to be derived for
two or more sites.
a. One assumption that is likely to be worth testing is that
the dissolved WER varies much less across time, location,
and depth within a site than the total recoverable WER. If
the assumption proves valid, a dissolved WER can be applied
to a dissolved national water quality criterion to derive a
dissolved site-specific water quality criterion that will
apply to the whole site.
b. A second assumption that might be worth testing is that the
WER correlates with a water quality characteristic such as
TSS or TOG across time, location, and depth.
c. Another assumption that might be worth testing is that the
dissolved and/or total recoverable WER is mostly due to
66
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nontoxic metal rather than to a water quality
characteristic that reduces toxicity. If this is true and
if there is variability in the WER, the WER'will correlate
with the concentration of metal in the site water. This is
similar to the first assumption, but this one can allow use
of both total recoverable and dissolved WERs, whereas the
first one only allows use of a dissolved WER.
If WERs are too variable to be useful and no way can be found
to deal with the variability, additional sampling will
probably be required in order to develop a WER and/or a site-
specific water quality criterion that is either (a) spatially
and/or temporally dependent or (b) constant and
environmentally conservative for nearly all conditions.
2. An experimental design should be developed that tests whether
the assumption is of practical value across the range of
conditions that occur at different times, locations, and
depths within the site. Each design has to be formulated
individually to fit the specific site. The design should try
to take into account the times, locations, and depths at which
the extremes of the physical, chemical, and biological
conditions occur within the site, which will require detailed
information concerning the site. In addition, the
experimental design should balance available resources with
the need for adequate sampling.
a. Selection of the number and timing of sampling events
should take into account seasonal, weekly, and daily
considerations. Intensive sampling should occur during the
two most extreme seasons, with confirmatory sampling during
the other two seasons. Selection of the day and time of
sample collection should take into account the discharge
schedules of the major industrial and/or municipal
discharges. For example, it might be appropriate to
collect samples during the middle of the week to allow for
reestablishment of steady-state conditions after shutdowns
for weekends and holidays; alternatively, end-of-the-week
slug discharges are routine in some situations. In coastal
sites, the tidal cycle might be important if facilities
discharge, for example, over a four-hour period beginning
at slack high tide. Because the highest concentration of
effluent in the surface water probably occurs at ebb tide,
determination of WERs using site water samples obtained at
this time might result in inappropriately large WERs that
would result in underprotection at other times; samples
with unusually large WERs might be especially useful for
testing assumptions. The importance of each consideration
should be determined on a case-by-case basis.
b. Selection of the number and locations of stations to be
sampled within a sampling event should consider the site as
a whole and take into account sources of water and
discharges, mixing patterns, and currents (and tides in
coastal areas). If the site has been adequately
67
-------
characterized, an acceptable design can probably be
developed using existing information concerning (1) sources
of the metal and other pollutants and (2) the spatial and
temporal distribution of concentrations of the metal and
water quality factors that might affect the toxicity of the
metal. Samples should not be taken within or near mixing
zones or plumes of dischargers; dilution models (U.S. EPA
1993) and dye dispersion studies (Kilpatrick 1992) can
indicate areas that should definitely be avoided. Maps,
current charts, hydrodynamic models, and water quality
models used to allocate waste loads and derive permit
limits are likely to be helpful when determining when and
where to obtain site-water samples. Available information
might provide an indication of the acceptability of site
water for testing selected species. The larger and more
complex the site, the greater the number of sampling
locations that will be needed.
c. In addition to determining the horizontal location of each
sampling station, the vertical location (i.e., depth) of
the sampling point needs to be selected. Known mixing
regimes, the presence of vertical stratification of TSS
and/or salinity, concentration of metal, effluent plumes,
tolerance of test species, and the need to obtain samples
of site water that span the range of site conditions should
be considered when selecting the depth at which the sample
is to be taken. Some decisions concerning depth cannot be
made until information is obtained at the: time of sampling;
for example, a conductivity meter, salinometer, or
transmissometer might be useful for determining where and
at what depth to collect samples. Turbidity might
correlate with TSS and both might relate to the toxicity of
the metal in site water; salinity can indicate whether the
test organisms and the site water are compatible.
Because each site is unique, specific guidance cannot be given
here concerning either the selection of the appropriate number
and locations of sampling stations within a site or the
frequency of sampling. All available information concerning
the site should be utilized to ensure that the times,
locations, and depths of samples span the range of water
quality characteristics that might affect the toxicity of the
metal:
a. High and low concentrations of TSS.
b. High and low concentrations of effluents.
c. Seasonal effects.
d. The range of tidal conditions in saltwater situations.
The sampling plan should provide the data needed to allow an
evaluation of the usefulness of the assumption(s) that the
experimental design is intended to test. Statisticians should
play a key role in experimental design and data analysis, but
professional judgment that takes into account pertinent
biological, chemical, and toxicological considerations is at
least as important as rigorous statistical analysis when
68
-------
interpreting the data and determining the degree to which the
data correspond to the assumption(s).
3. The details of each sampling design should be formulated with
the aid of people who understand the site and people who have
a working knowledge of WERs. Because of the complexity of
designing a WER study for large sites, the design team should
utilize the combined expertise and experience of individuals
from the appropriate EPA Region, states, municipalities,
dischargers, environmental groups, and others who can
constructively contribute to the design of the study.
Building a team of cooperating aquatic toxicologists, aquatic
chemists, limnologists, oceanographers, water quality
modelers, statisticians, individuals from other key
disciplines, as well as regulators and those regulated, who
have knowledge of the site and the site-specific procedures,
is central to success of the derivation of a WER for a large
site. Rather than submitting the workplan to the appropriate
regulatory authority (and possibly the Water Management
Division of the EPA Regional Office) for comment at the end,
they should be members of the team from the beginning.
4. Data from one sampling event should always be analyzed prior
to the next sampling event with the goal of improving the
sampling design as the study progresses. For example, if the
toxicity of the metal in surface water samples is related to
the concentration of TSS, a water quality characteristic such
as turbidity might be measured at the time of collection of
water samples and used in the selection of the concentrations
to be used in the WER toxicity tests in site water. At a
minimum, the team that interprets the results of one sampling
event and plans the next should include an aquatic
toxicologist, a metals chemist, a statistician, and a modeler
or other user of the data.
5. The final interpretation of the data and the derivation of the
FWER(s) should be performed by a team. Sufficient data are
likely to be available to allow a quantitative estimate of
experimental variation, differences between species, and
seasonal differences. It will be necessary to decide whether
one site-specific criterion can be applied to the whole area
or whether separate site-specific criteria need to be derived
for two or more sites. The interpretation of the data might
produce two or more alternatives that the appropriate
regulatory authority could subject to a cost-benefit analysis.
Other aspects of the determination of a WER for a large site are
likely to be the same as described for Method 1. For example:
a. WERs should be determined using two or more sensitive species;
the suggestions given in Appendix I should be considered when
selecting the tests and species to be used.
69
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b. Chemical analyses of site water, laboratory dilution water,
and test solutions should follow the requirements for the
specific test used and those given in this document.
c. If tests in many surface water samples are compared to one
test in a laboratory dilution water, it is very important that
that one test be acceptable. Use of (1) rangefinding tests,
(2) additional treatments beyond the standard five
concentrations plus controls, and (3) dilutions that are
functions of the known concentration-effect relationships
obtained with the toxicity test and metal of concern will help
ensure that the desired endpoints and WERs can be calculated.
d. Measurements of the concentrations of both total recoverable
and dissolved metal should be targeted to the test
concentrations whose data will be used in the calculation of
the endpoints.
e. Samples of site water and/or effluent should be collected,
handled, and transported so that the tests can begin as soon
as is feasible.
f. If the large site is a saltwater site, the considerations
presented in Appendix H ought to be given attention.
70
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Figure 2: Calculating an Adjusted Geometric Mean
Where n = the number of experimentally determined WERs in a set,
the "adjusted geometric mean" of the set is calculated as
follows:
a. Take the logarithm of each of the WERs. The logarithms can be
to any base, but natural logarithms (base e) are preferred for
reporting purposes.
b. Calculate x = the arithmetic mean of the logarithms.
c. Calculate s = the sample standard deviation of the
logarithms:
n - 1
d. Calculate SE = the standard error of the arithmetic mean:
SE = s/i/n _
e. Calculate A = x- (t0-7) (SE) , where t0 7 is the value of Student's
t statistic for a one-sided probability of 0.70 with 12-1
degrees of freedom. The values of t0-7 for some common
degrees of freedom (df) are:
1 0.727
2 0.617
3 0.584
4 0.569
5 0.559
6 0.553
7 0.549
8 0.546
9 0.543
10 0.542
11 0.540
12 0.539
The values of fc0 7 for more degrees of freedom are available,
for example, on page T-5 of Natrella (1966).
f. Take the antilogarithm of A.
This adjustment of the geometric mean accounts for the fact that
the means of fifty percent of the sets of WERs are expected to be
higher than the actual mean; using the one-sided value of t for
0.70 reduces the percentage to thirty.
71
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Figure 3: An Example Derivation of a FWER
This example assumes that cccWERs were determined monthly using
simulated downstream water that was prepared by mixing upstream
water with effluent at the ratio that existed when the samples
were obtained. Also, the flow of the effluent is always 10 cfs,
and the design flow of the upstream water is 40 cfs. (Therefore,
the downstream flow at design-flow conditions is 50 cfs.) The
concentration of metal in. upstream water at design flow is 0.4
ug/L, and the CCC is 2 ug/L. Each FWER is derived from the WERs
and hWERs that are available through that month.
Month
March
April
May
June
July
Aug.
Sept.
Oct.
Nov.
Dec.
Jan.
Feb.
ePLOW
(cfs)
10
10
10
10
10
10
10
10
10
10
10
10
uFLOW
(cfs)
850
289
300
430
120
85
40
45
150
110
180
244
uCONC
(ucr/L)
0.8
0.6
0.6
0.6
0.4
0.4
0.4
0.4
0.4
0.4
0.6
0.6
WER
HOME
(ucr/L)
hWER
FWER
5
6
5
5
7
10
12
11
7
3
6
6
.2a
.0=
.8C
.7C
.Oc
.5e
.Oe
.Oe
.5°
.5C
.9C
.lc
826
341
341
475
177
196
118
119
234
79
251
295
.4
.5
.6
.8
.2
.1
.4
.2
.0
.6
.4
.2
82
34
34
47
17
19
12
12
23
8
25
29
.80
.31
.32
.74
.88
.77
.00
.08
.56
.12
.30
.68
1
1
1
5
5
6
10
10
10
8
8
8
.Ob
.Ob
.Ob
.7d
.7d
.80f
.693
.883
.88g
.12h
.12h
.12h
Neither Type 1 nor Type 2; the downstream flow (i.e., the sum
of the eFLOW and the uFLOW) is > 500 cfs.
The total number of available Type 1 and Type 2 WERs is less
than 3.
A Type 2 WER; the downstream flow is between 100 and 500 cfs.
No Type 1 WER is available; the FWER is the lower of the
lowest Type 2 WER and the lowest hWER.
A Type 1 WER; the downstream flow is between 50 and 100 cfs.
One Type 1 WER is available; the FWER is the geometric mean of
all Type 1 and Type 2 WERs.
Two or more Type 1 WERs are available and the range is less
than a factor of 5; the FWER is the adjusted geometric mean
(see Figure 2) of the Type 1 WERs, because all the hWERs are
higher.
Two or more Type 1 WERs are available and the range is not
greater than a factor of 5; the FWER is the lowest hWER
because the lowest hWER is lower than the adjusted geometric
mean of the Type 1 WERs.
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Figure 4: Reducing the Impact of Experimental Variation
When the FWER is the lowest of, for example, three WERs, the
impact of experimental variation can be reduced by conducting
additional primary tests. If the endpoint of the secondary test
is above the CMC or CCC to which the FWER is to be applied, the
additional tests can also be conducted with the secondary test.
Month
April
May
June
Lowest
Case 1
(Primary
Test)
4.801
2.552
9.164
2.552
(Primary
Test)
4.801
2.552
9.164
Case 2
(Primary
Test)
3.565
4.190
6.736
Geometric
Mean
4.137
3 .270
7.857
3.270
Month
April
May
June
Lowest
Case 3
(Primary (Second.
Test) Test)
4.801
2.552
9.164
3.163
5.039
7.110
Geo.
Mean
3.897
3.586
8.072
3.586
Case 4
(Primary (Second.
Test) Test)
4.801
2 .552
9.164
3.163
2.944
7.110
Geo.
Mean
3.897
2.741
8.072
2.741
Case 1 uses the individual WERs obtained with the primary test
for the three months, and the FWER is the lowest of the three
WERs. In Case 2, duplicate primary tests were conducted in each
month, so that a geometric mean could be calculated for each
month; the FWER is the lowest of the three geometric means.
In Cases 3 and 4, both a primary test and a secondary test were
conducted each month and the endpoints for both tests in
laboratory dilution water are above the CMC or CCC to which the
FWER is to be applied. In both of these cases, therefore, the
FWER is the lowest of the three geometric means.
The availability of these alternatives does not mean that they
are necessarily cost-effective.
73
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Figure 5: Calculating an LC50 (or EC50) by Interpolation
When fewer than two treatments kill some but not all of the
exposed test organisms, a statistically sound estimate of an LC50
cannot be calculated. Some programs and methods produce LCBOs
when there are fewer than two "partial kills", but such results
are obtained using interpolation, not statistics. If (a) a test
is otherwise acceptable, (b) a sufficient number of organisms are
exposed to each treatment, and (c) the concentrations are
sufficiently close together, a test with zero or one partial kill
can provide all the information that is needed concerning the
LC50. An LC50 calculated by interpolation should probably be
called an "approximate LC50" to acknowledge the lack of a
statistical basis for its calculation, but this does not imply
that such an LC50 provides no useful toxicological information.
If desired, the binomial test can be used to calculate a
statistically sound probability that the true LC50 lies between
two tested concentrations (Stephan 1977) .
Although more complex interpolation methods can be used, they
will not produce a more useful LC50 than the method described
here. Inversions in the data between two test concentrations
should be removed by pooling the mortality data for those two
concentrations and calculating a percent mortality that is then
assigned to both concentrations . Logarithms to a base other than
10 can be used if desired. If PI and P2 are the percentages of
the test organisms that died when exposed to concentrations Cl
and C2, respectively, and if Cl < C2, PI < P2, 0 £ PI s 50,
and 50 s P2 & 100, then:
50 - Pi
P2 - Pi
C = Log- Cl + P(Log C2 - Log Cl)
LC50 = 10C
If PI - 0 and P2 = 100, LC50 = J(C1) (C2) .
If PI - P2 = 50, LC50 = V(C1) (C2) .
If PI » 50, LC50 = Cl.
If P2 = 50, LC50 = C2.
If Cl - 4 mg/L, C2 = 7 'mg/L, PI = 15 %, and P2 == 100 %,
then LC50 = 5.036565 mg/L.
Besides the mathematical requirements given above, the following
toxicological recommendations are given in sections G.8 and 1.2:
a. 0.65 < C1/C2 < 0.99.
b. 0 s PI < 37.
c. 63 < P2 s 100.
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Figure 6: Calculating a Time-Weighted Average
If a sampling plan (e.g., for measuring metal in a treatment in a
toxicity test) is designed so .that a series of values are
obtained over time in such a way that each value contains the
same amount of information (i.e., represents the same amount of
time), then the most meaningful average is the arithmetic
average. In most cases, however, when a series of values is
obtained over time, some values contain more information than
others; in these cases the most meaningful average is a time-
weighted average (TWA). If each value contains the same amount
of information, the arithmetic average will equal the TWA.
A TWA is obtained by multiplying each value by a weight and then
dividing the sum of the products by the sum of the weights. The
simplest approach is to let each weight be the duration of time
that the sample represents. Except for the first and last
samples, the period of time represented by a sample starts
halfway to the previous sample and ends halfway to the next
sample. The period of time represented by the first sample
starts at the beginning of the test, and the period of time
represented by the last sample ends at the end of the test. Thus
for a 96-hr toxicity test, the sum of the weights will be 96 hr.
The following are hypothetical examples of grab samples taken
from 96-hr flow-through tests for two common sampling regimes:
Sampling Cone. Weight Product Time-weighted average
time (hr) (mcr/L) (hr) (hr) (mq/L) (mg/L)
0 12 48 576
96 14 48 672
96 1248 1248/96 = 13.00
0 8 12 96
24 6 24 144
48 7 24 168
72 9 24 216
96 8 12. 96
96 720 720/96 = 7.500
When all the weights are the same, the arithmetic average equals
the TWA. Similarly, if only one sample is taken, both the
arithmetic average and the TWA equal the value of that sample.
The rules are more complex for composite samples and for samples
from renewal tests. In all cases, however, the sampling plan can
be designed so that the TWA equals the arithmetic average.
75
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REFERENCES
ASTM. 1993a. Guide for Conducting Acute Toxicity Tests with
Pishes, Macroinvertebrates, and Amphibians. Standard E729.
American Society for Testing and Materials, Philadelphia, PA.
ASTM. 1993b. Guide for Conducting Static Acute Toxicity Tests
Starting with Embryos of Four Species of Saltwater Bivalve
Molluscs. Standard E724. American Society for Testing and
Materials, Philadelphia, PA.
ASTM. 1993c. Guide for Conducting Renewal Life-Cycle Toxicity
Tests with Daphnia magna. Standard E1193. American Society for
Testing and Materials, Philadelphia, PA.
ASTM. 1993d. Guide for Conducting Early Life-Stage Toxicity
Tests with Fishes. Standard E1241. American Society for Testing
and Materials, Philadelphia, PA.
ASTM. 1993e. Guide for Conducting Three-Brood, Renewal Toxicity
Tests with Ceriodaphnia dubia. Standard E1295. American Society
for Testing and Materials, Philadelphia, PA.
ASTM. 1993f. Guide for Conducting Acute Toxicity Tests on
Aqueous Effluents with Fishes, Macroinvertebrates, and
Amphibians. Standard E1192. American Society for Testing and
Materials, Philadelphia, PA.
Barnthouse, L.W., G.W. Suter, A.E. Rosen, and J.J. Beauchamp.
1987. Estimating Responses of Fish Populations to Toxic
Contaminants. Environ. Toxicol. Chem. 6:811-824.
Bruce, R.D., and D.J. Versteeg. 1992. A Statistical Procedure
for Modeling Continuous Toxicity Data. Environ. Toxicol. Chem.
11:1485-1494.
Hoekstra, J.A., and P.M. Van Ewijk. 1993. Alternatives for the
No-Observed-Effect Level. Environ. Toxicol. Chem. 12:187-194.
Kilpatrick, F.A. 1992. Simulation of Soluble Waste Transport
and Buildup in Surface Waters Using Tracers. Open-File Report
92-457. U.S. Geological Survey, Books and Open-File Reports, Box
25425, Federal Center, Denver, CO 80225.
Natrella, M.G. 1966. Experimental Statistics. National Bureau
of Standards Handbook 91. (Issued August 1, 1963; reprinted
October 1966 with corrections). U.S. Government Printing Office,
Washington, DC.
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Prothro, M.G. 1993. Memorandum titled "Office of Water Policy
and Technical Guidance on Interpretation and Implementation of
Aquatic Life Metals Criteria". October 1.
Stephan, C.E. 1977. Methods for Calculating an LC50. In:
Aquatic Toxicology and Hazard Evaluation. (F.L. Mayer and J.L.
Hamelink, eds.) ASTM STP 634. American Society for Testing and
Materials, Philadelphia, PA. pp. 65-84.
Stephan, C.E., and J.W. Rogers. 1985. Advantages of Using
Regression Analysis to Calculate Results of Chronic Toxicity
Tests. In: Aquatic Toxicology and Hazard Assessment: Eighth
Symposium. (R.C. Bahner and D.J. Hansen, eds.) ASTM STP 891.
American Society for Testing and Materials, Philadelphia, PA.
pp. 328-338.
Suter, G.W., A.E. Rosen, E. Linder, and D.F. Parkhurst. 1987.
Endpoints for Responses of Fish to Chronic Toxic Exposures.
Environ. Toxicol. Chem. 6:793-809.
U.S. EPA. 1983a. Water Quality Standards Handbook. Office of
Water Regulations and Standards, Washington, DC.
U.S. EPA. 1983b. Methods for Chemical Analysis of Water and
Wastes. EPA-600/4-79-020. National Technical Information
Service, Springfield, VA.
U.S. EPA. 1984. Guidelines for Deriving Numerical Aquatic Site-
Specific Water Quality Criteria by Modifying National Criteria.
EPA-600/3-84-099 or PB85-121101. National Technical
Information Service, Springfield, VA.
U.S. EPA. 1985. Guidelines for Deriving Numerical National
Water Quality Criteria for the Protection of Aquatic Organisms
and Their Uses. PB85-227049. National Technical Information
Service, Springfield, VA.
U.S. EPA. 1991a. Technical Support Document for Water Quality-
based Toxics Control. EPA/505/2-90-001 or PB91-127415.
National Technical Information Service, Springfield, VA.
U.S. EPA. 1991b. Manual for the Evaluation of Laboratories
Performing Aquatic Toxicity Tests. EPA/600/4-90/031. National
Technical Information Service, Springfield, VA.
U.S. EPA. 1991c. Methods for the Determination of Metals in
Environmental Samples. EPA-600/4-91-010. National Technical
Information Service, Springfield, VA.
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U.S. EPA. 1992. Interim Guidance on Interpretation and
Implementation of Aquatic Life Criteria for Metals. Office of
Science and Technology, Health and Ecological Criteria Division,
Washington, DC.
U.S. EPA. 1993a. Methods for Measuring the Acute Toxicity of
Effluents and Receiving Waters to Freshwater and Marine
Organisms. Fourth Edition. EPA/600/4-90/027F. National
Technical Information Service, Springfield, VA.
U.S. EPA. 1993b. Short-term Methods for Estimating the Chronic
Toxicity of Effluents and Receiving Waters to Freshwater
Organisms. Third Edition. EPA/600/4-91/002. National Technical
Information Service, Springfield, VA.
U.S. EPA. 1993c. Short-Term Methods for Estimating the Chronic
Toxicity of Effluents and Receiving Waters to Marine and
Estuarine Organisms. Second Edition. EPA/600/4-91/003.
National Technical Information Service, Springfield, VA.
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Appendix A: Comparison of WERs Determined Using Upstream and
Downstream Water
The "Interim Guidance" concerning metals (U.S. EPA 1992) made a
fundamental change in the way WERs should be experimentally
determined because it changed the source of the site water. The
earlier guidance (U.S. EPA 1983,1984) required that upstream
water be used as the site water, whereas the newer guidance (U.S.
EPA 1992) recommended that downstream water be used as the site
water. The change in the source of the site water was merely an
acknowledgement that the WER that applies at a location in a body
of water should, when possible, be determined using the water
that occurs at that location.
Because the change in the source of the dilution water was
expected to result in an increase in the magnitude of many WERs,
interest in and concern about the determination and use of WERs
increased. When upstream water was the required site water, it
was expected that WERs would generally be low and that the
determination and use of WERs could be fairly simple. After
downstream water became the recommended site water, the
determination and use of WERs was examined much more closely. It
was then realized that the determination and use of upstream WERs
was more complex than originally thought. It was also realized
that the use of downstream water greatly increased the complexity
and was likely to increase both the magnitude and the variability
of many WERs. Concern about the fate of discharged metal also
increased because use of downstream water might allow the
discharge of large amounts of metal that has reduced or no.
toxicity at the end of the pipe. The probable increases in the
complexity, magnitude, and variability of WERs and the increased
concern about fate, increased the importance of understanding the
relevant issues as they apply to WERs determined using both
upstream water and downstream water.
A. Characteristics of the Site Water
The idealized concept of an upstream water is a pristine water
that is relatively unaffected by people. In the real world,
however, many upstream waters contain naturally occurring
ligands, one or more effluents, and materials from nonpoint
sources; all of these might impact a WER. If the upstream
water receives an effluent containing TOC and/or TSS that
contributes to the WER, the WER will probably change whenever
the quality or quantity of the TOC and/or TSS changes. In
such a case, the determination and use of the WER in upstream
water will have some of the increased complexity associated
with use of downstream water and some of the concerns
associated with multiple-discharge situations (see Appendix
F). The amount of complexity will depend greatly on the
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number and type of upstream point and nonpoint sources, the
frequency and magnitude of fluctuations, and whether the WER
is being determined above or below the point of complete mix
of the upstream sources.
Downstream water is a mixture of effluent and upstream water,
each of which can contribute to the WER, and so there are two
components to a WER determined in downstream, water: the
effluent component and the upstream component. The existence
of these two components has the following implications:
1. WERs determined using downstream water are likely to be
larger and more variable than WERs determined using
upstream water.
2. The effluent component should be applied only where the
effluent occurs, which has implications concerning
implementation.
3. The magnitude of the effluent component of a WER will
depend on the concentration of effluent in the downstream
water. (A consequence of this is that the effluent
component will be zero where the concentration of effluent
is zero, which is the point of item 2 above.)
4. The magnitude of the effluent component of a WER is likely
to vary as the composition of the effluent varies.
5, Compared to upstream water, many effluents contain higher
concentrations of a wider variety of substances that can
impact the toxicity of metals in a wider variety of ways,
and so the effluent component of a WER can be due to a
variety of chemical effects in addition to such factors as
hardness, alkalinity, pH, and humic acid.
6. Because the effluent component might be due, in whole or in
part, to the discharge of refractory metal (see Appendix
D), the WER cannot be thought of simply as being caused by
the effect of water quality on the toxicity of the metal.
Dealing with downstream WERs is so much simpler if the
effluent WER (eWER) and the upstream WER (uWER) are additive
that it is desirable to understand the concept of additivity
of WERs, its experimental determination, and its use (see
Appendix G).
B. The Implications of Mixing Zones.
When WERs are determined using upstream water, the presence or
absence of mixing zones has no impact; the cmcWER and the
cccWER will both be determined using site water that contains
zero percent of the effluent of concern, i.e., the two WERs
will be determined using the same site water.
When WERs are determined using downstream water, the magnitude
of each WER will probably depend on the concentration of
effluent in the downstream water used (see Appendix D). The
concentration of effluent in the site water will depend on
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where the sample is taken, which will not be the same for the
cmcWER and the cccWER if there are mixing zone(s). Most, if
not all, discharges have a chronic (CCC) mixing zone; many,
but not all, also have an acute (CMC) mixing zone. The CMC
applies at all points except those inside a CMC mixing zone;
thus if there is no CMC mixing zone, the CMC applies at the
end of the pipe. The CCC applies at all points outside the
CCC mixing zone. It is generally assumed that if permit
limits are based on a point in a stream at which both the CMC
and the CCC apply, the CCC will control the permit limits,
although the CMC might control if different averaging periods
are appropriately taken into account. For this discussion, it
will be assumed that the same design flow (e.g., 7Q10) is used
for both the CMC and the CCC.
If the cmcWER is to be appropriate for use inside the chronic
mixing zone, but the cccWER is to be appropriate for use
outside the chronic mixing zone, the concentration of effluent
that is appropriate for use in the determination of the two
WERs will not be the same. Thus even if the same toxicity
test is used in the determination of the cmcWER and the
cccWER, the two WERs will probably be different because the
concentration of effluent will be different in the two site
waters in which the WERs are determined.
If the CMC is only of concern within the CCC mixing zone, the
highest relevant concentration of metal will occur at the edge
of the CMC mixing zone if there is a CMC mixing zone; the
highest concentration will occur at the end of the pipe if
there is no CMC mixing zone. In contrast, within the CCC
mixing zone, the lowest cmcWER will probably occur at the
outer edge of the CCC mixing zone. Thus the greatest level of
protection would be provided if the cmcWER is determined using
water at the outer edge of the CCC mixing zone, and then the
calculated site-specific CMC is applied at the edge of the CMC
mixing zone or at the end of the pipe, depending on whether
there is an acute mixing zone. The cmcWER is likely to be
lowest at the outer edge of the CCC mixing zone because of
dilution of the effluent, but this dilution will also dilute
the metal. If the cmcWER is determined at the outer edge of
the CCC mixing zone but the resulting site-specific CMC is
applied at the end of the pipe or at the edge of the CMC
mixing zone, dilution is allowed to reduce the WER but it is
not allowed to reduce the concentration of the metal. This
approach is environmentally conservative, but it is probably
necessary given current implementation procedures. (The
situation might be more complicated if the uWER is higher than
the eWER or if the two WERs are less-than-additive.)
A comparable situation applies to the CCC. Outside the CCC
mixing zone, the CMC and the CCC both apply, but it is assumed
that the CMC can be ignored because the CCC will be more
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restrictive. The cccWER should probably be determined for the
complete-mix situation, but the site-specific CCC will have to
be met at the edge of the CCC mixing zone. Thus dilution of
the WER from the edge of the CCC mixing zone to the point of
complete mix is taken into account, but dilution of the metal
is not.
If there is neither an acute nor a chronic mixing zone, both
the CMC and the CCC apply at the end of the pipe, but the CCC
should still be determined for the complete-mix situation.
C. Definition of site.
In the general context of site-specific criteria, a "site" may
be a state, region, watershed, waterbody, segment of a
waterbody, category of water (e.g., ephemeral streams), etc.,
but the site-specific criterion is to be derived to provide
adequate protection for the entire site, however the site is
defined. Thus, when a site-specific criterion is derived
using the Recalculation Procedure, all species that "occur at
the site" need to be taken into account when deciding what
species, if any, are to be deleted from the dataset.
Similarly, when a site-specific criterion is derived using a
WER, the WER is to be adequately protective of the entire
site. If, for example, a site-specific criterion is being
derived for an estuary, WERs could be determined using samples
of the surface water obtained from various sampling stations,
which, to avoid confusion, should not be called "sites". If
all the WERs were sufficiently similar, one site-specific
criterion could be derived to apply to the whole estuary. If
the WERs were sufficiently different, either the lowest WER
could be used to derive a site-specific criterion for the
whole estuary, or the data might indicate that the estuary
should be divided into two or more sites, each with its own
criterion.
The major principle that should be applied when defining the
area to be included in the site is very simplistic: The site
should be neither too small nor too large.
1. Small sites are probably appropriate for cmcWERs, but
usually are not appropriate for cccWERs because metals are
persistent, although some oxidation states are not
persistent and some metals are not persistent in the water
column. For cccWERs, the smaller the defined site, the
more likely it is that the permit limits will be controlled
by a criterion for an area that is outside the site, but
which could have been included in the site without
substantially changing the WER or increasing the cost of
determining the WER.
2. Too large an area might unnecessarily increase the cost of
determining the WER. As the size of the site increases,
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the spatial and temporal variability is likely to increase,
which will probably increase the number of water samples in
which WERs will need to be determined before a site-
specific criterion can be derived.
3. Events that import or resuspend TSS and/or TOG are likely
to increase the total recoverable concentration of the
metal and the total recoverable WER while having a much
smaller effect on the dissolved concentration and the
dissolved WER. Where the concentration of dissolved metal
is substantially more constant than the concentration of
total recoverable metal, the site can probably be much
larger for a dissolved criterion than for a total
recoverable criterion. If one criterion is not feasible
for the whole area, it might be possible to divide it into
two or more sites with separate total recoverable or
dissolved criteria or to make the criterion dependent on a
water quality characteristic such as TSS or salinity.
4. Unless the site ends where one body of water meets another,
at the outer edge of the site there will usually be an
instantaneous decrease in the allowed concentration of the
metal in the water column due to the change from one
criterion to another, but there will not be an
instantaneous decrease in the actual concentration of metal
in the water column. The site has to be large enough to
include the transition zone in which the actual
concentration decreases so that the criterion outside the
site is not exceeded.
It is, of course, possible in some situations that relevant
distant conditions (e.g., a lower downstream pH) will
necessitate a low criterion that will control the permit
limits such that it is pointless to determine a WER.
When a WER is determined in upstream water, it is generally
assumed that a downstream effluent will not decrease the WER.
It is therefore assumed that the site can usually cover a
rather large geographic area.
When a site-specific criterion is derived based on WERs
determined using downstream water, the site should not be
defined in the same way that it would be defined if the WER
were determined using upstream water. The eWER should be
allowed to affect the site-specific criterion wherever the
effluent occurs, but it should not be allowed to affect the
criterion in places where the effluent does not occur. In
addition, insofar as the magnitude of the effluent component
at a point in the site depends on the concentration of
effluent, the magnitude of the WER at a particular point will
depend on the concentration of effluent at that point. To the
extent that the eWER and the uWER are additive, the WER and
the concentration of metal in the plume will decrease
proportionally (see Appendix G).
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When WERs are determined using downstream water, the following
considerations should be taken into account when the site is
defined:
1. If a site-specific criterion is derived using a WER that
applies to the complete-mix situation, the upstream edge of
the site to which this criterion applies should be the
point at which complete mix actually occurs. If the site
to which the complete-mix WER is applied starts at the end
of the pipe and extends all the way across the stream,
there will be an area beside the plume that will not be
adequately protected by the site-specific criterion.
2. Upstream of the point of complete mix, it will usually be
protective to apply a site-specific criterion that was
derived using a WER that was determined using upstream
water.
3. The plume might be an area in which the concentration of
metal could exceed a site-specific criterion without
causing toxicity because of simultaneous dilution of the
metal and the eWER. The fact that the plume is much larger
than the mixing zone might not be important if there is no
toxicity within the plume. As long as the concentration of
metal in 100 % effluent does not exceed that allowed by the
additive portion of the eWER, from a toxicological
standpoint neither the size nor the definition of the plume
needs to be of concern because the metal will not cause
toxicity within the plume. If there is no toxicity within
the plume, the area in the plume might be like a
traditional mixing zone in that the concentration of metal
exceeds the site-specific criterion, but it would be
different from a traditional mixing zone in that the level
of protection is not reduced.
Special considerations are likely to be necessary in order to
take into account the eWER when defining a site related to
multiple discharges (see Appendix F).
D. The variability in the experimental determination of a WER.
When a WER is determined using upstream water, the two major
sources of variation in the WER are (a) variability in the
quality of the site water, which might be related to season
and/or flow, and (b) experimental variation. Ordinary day-to-
day variation will account for some of the variability, but
seasonal variation is likely to be more important.
As explained in Appendix D, variability in the concentration
of nontoxic dissolved metal will contribute to the variability
of both total recoverable WERs and dissolved WERs; variability
in the concentration of nontoxic particulate metal will
contribute to the variability in a total recoverable WER, but
not to the variability in a dissolved WER. Thus, dissolved
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WERs are expected to be less variable than total recoverable
WERs, especially where events commonly increase TSS and/or
TOG. In some cases, therefore, appropriate use of analytical
chemistry can greatly increase the usefulness of the
experimental determination of WERs. The concerns regarding
variability are increased if an upstream effluent contributes
to the WER.
When a WER is determined in downstream water, the four major
sources of variability in the WER are (a) variability in the
quality of the upstream water, which might be related to
season and/or flow, (b) experimental variation, (c)
variability in the composition of the effluent, and (d)
variability in the ratio of the flows of the upstream water
and the effluent. The considerations regarding the first two
are the same as for WERs determined using upstream water;
because of the additional sources of variability, WERs
determined using downstream water are likely to be more
variable than WERs determined using upstream water.
It would be desirable if a sufficient number of WERs could be
determined to define the variable factors in the effluent and
in the upstream water that contribute to the variability in
WERs that are determined using downstream water. Not only is
this likely to be very difficult in most cases, but it is also
possible that the WER will be dependent on interactions
between constituents of the effluent and the upstream water,
i.e., the eWER and uWER might be additive, more-than-additive,
or less-than-additive (see Appendix G). When interaction
occurs, in order to completely understand the variability of
WERs determined using downstream water, sufficient tests would
have to be conducted to determine the means and variances of:
a. the effluent component of the WER.
b. the upstream component of the WER.
c. any interaction between the two components.
An interaction might occur, for example, if the toxicity of a
metal is affected by pH, and the pH and/or the buffering
capacity of the effluent and/or the upstream water vary
considerably.
An increase in the variability of WERs decreases the
usefulness of any one WER. Compensation for this decrease in
usefulness can be attempted by determining WERs at more times;
although this will provide more data, it will not necessarily
provide a proportionate increase in understanding. Rather
than determining WERs at more times, a better use of resources
might be to obtain more information concerning a smaller
number of specially selected occasions.
It is likely that some cases will be so complex that achieving
even a reasonable understanding will require unreasonable
resources. In contrast, some WERs determined using the
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methods presented herein might be relatively easy to
understand if appropriate chemical measurements are performed
when WERs are determined.
1. If the variation of the total recoverable WER is
substantially greater than the variation of the comparable
dissolved WER, there is probably a variable and substantial
concentration of particulate nontoxic metal. It might be
advantageous to use a dissolved WER just because it will
have less variability than a total recoverable WER.
2. If the total recoverable and/or dissolved WER correlates
with the total recoverable and/or dissolved concentration
of metal in the site water, it is likely that a substantial
percentage of the metal is nontoxic. In this case the WER
will probably also depend on the concentration of effluent
in the site water and on the concentration of metal in the
effluent.
These approaches are more likely to be useful when WERs are
determined using downstream water, rather than upstream water,
unless both the magnitude of the WER and the concentration of
the metal in the upstream water are elevated by an upstream
effluent and/or events that increase TSS and/or TOG.
Both of these approaches can be applied to WERs that are
determined using actual downstream water, but the second can
probably provide much better information if it is used with
WERs determined using simulated downstream water that is
prepared by mixing a sample of the effluent with a sample of
the upstream water. In this way the composition and
characteristics of both the effluent and the upstream water
can be determined, and the exact ratio in the downstream water
is known.
Use of simulated downstream water is also a way to study the
relation between the WER and the ratio of effluent to upstream
water at one point in time, which is the most direct way to
test for additivity of the eWER and the uWER (see Appendix G).
This can be viewed as a test of the assumption that WERs
determined using downstream water will decrease as the
concentration of effluent decreases. If this assumption is
true, as the flow increases, the concentration of effluent in
the downstream water will decrease and the WER will decrease.
Obtaining such information at one point in time is useful, but
confirmation at one or more other times would be much more
useful.
E. The fate of metal that has reduced or no toxicity.
Metal that has reduced or no toxicity at the end of the pipe
might be more toxic at some time in the future. For example,
metal that is in the water column and is not toxic now might
become more toxic in the water column later or might move into
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the sediment and become toxic. If a WER allows a surface
water to contain as much toxic metal as is acceptable, the WER
would not be adequately protective if metal that was nontoxic
when the WER was determined became toxic in the water column,
unless a compensating change occurred. Studies of the fate of
metals need to address not only the changes that take place,
but also the rates of the changes.
Concern about the fate of discharged metal justifiably raises
concern about the possibility that metals might contaminate
sediments. The possibility of contamination of sediment by
toxic and/or nontoxic metal in the water column was one of the
concerns that led to the establishment of EPA's sediment
quality criteria program, which is developing guidelines and
criteria to protect sediment. A separate .program was
necessary because ambient water quality criteria are not
designed to protect sediment. Insofar as technology-based
controls and water quality criteria reduce the discharge of
metals, they tend to reduce the possibility of contamination
of sediment. Conversely, insofar as WERs allow an increase in
the discharge of metals, they tend to increase the possibility
of contamination of sediment.
When WERs are determined in upstream water, the concern about
the fate of metal with reduced or no toxicity is usually small
because the WERs are usually small. In addition, the factors
that result in upstream WERs being greater than 1.0 usually
are (a) natural organic materials such as humic acids and (b)
water quality characteristics such as hardness, alkalinity,
and pH. It is easy to assume that natural organic materials
will not degrade rapidly, and it is easy to monitor changes in
hardness, alkalinity, and pH. Thus there is usually little
concern about the fate of the metal when WERs are determined
in upstream water, especially if the WER is small. If the WER
is large and possibly due at least in part to an upstream
effluent, there is more concern about the fate of metal that
has reduced or no toxicity.
When WERs are determined in downstream water, effluents are
allowed to contain virtually unlimited amounts of nontoxic
particulate metal and nontoxic dissolved metal. It would seem
prudent to obtain some data concerning whether the nontoxic
metal might become toxic at some time in the future whenever
(1) the concentration of nontoxic metal is large, (2) the
concentration of dissolved metal is below the dissolved
national criterion but the concentration of total recoverable
metal is substantially above the total recoverable national
criterion, or (3) the site-specific criterion is substantially
above the national criterion. It would seem appropriate to:
a. Generate some data concerning whether "fate" (i.e.,
environmental processes).will cause any of the nontoxic
metal to become toxic due to oxidation of organic matter,
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oxidation of sulfides, etc. For example, a WER could be
determined using a sample of actual or simulated downstream
water, the sample aerated for a period of time (e.g., two
weeks), the pH adjusted if necessary, and another WER
determined. If aeration reduced the WER, shorter and
longer periods of aeration could be used to study the rate
of change.
b. Determine the effect of a change in water quality
characteristics on the WER; for example, determine the
effect of lowering the pH on the WER if influent lowers the
pH of the downstream water within the area to which the
site-specific criterion is to apply.
c. Determine a WER in actual downstream water to demonstrate
whether downstream conditions change sufficiently (possibly
due to degradation of organic matter, multiple dischargers,
etc.) to lower the WER more than the concentration of the
metal is lowered.
If environmental processes cause nontoxic metal to become
toxic, it is important to determine whether the time scale
involves days, weeks, or years.
Summary
When WERs are determined using downstream water, the site water
contains effluent and the WER will take into account not only the
constituents of the upstream water, but also the toxic and
nontoxic metal and other constituents of the effluent as they
exist after mixing with upstream water. The determination of the
WER automatically takes into account any additivity, synergism,
or antagonism between the metal and components of the effluent
and/or the upstream water. The effect of calcium, magnesium, and
various heavy metals on competitive binding by such organic
materials as humic acid is also taken into account. Therefore, a
site-specific criterion derived using a WER is likely to be more
appropriate for a site than a national, state, or recalculated
criterion not only because it takes into account the water
quality characteristics of the site water but also because it
takes into account other constituents in the effluent and
upstream water.
Determination of WERs using downstream water causes a general
increase in the complexity, magnitude, and variability of WERs,
and an increase in concern about the fate of metal that has
reduced or no toxicity at the end of the pipe. In addition,
there are some other drawbacks with the use of downstream water
in the determination of a WER:
1. It might serve as a disincentive for some dischargers to
remove any more organic carbon and/or particulate matter than
required, although WERs for some metals will not be related to
the concentration of TOG or TSS.
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2. If conditions change, a WER might decrease in the future.
This is not a problem if the decrease is due to a reduction in
nontoxic metal, but it might be a problem if the decrease is
due to a decrease in TOG or TSS or an increase in competitive
binding.
3. If a WER is determined when the effluent contains refractory
metal but a change in operations results in the discharge of
toxic metal in place of refractory metal, the site-specific
criterion and the permit limits will not provide adequate
protection. In most cases chemical monitoring probably will
not detect such a change, but toxicological monitoring
probably will.
Use of WERs that are determined using downstream water rather
than upstream water increases:
1. The importance of understanding the various issues involved in
the determination and use of WERs.
2. The importance of obtaining data that will provide
understanding rather than obtaining data that will result in
the highest or lowest WER.
3. The appropriateness of site-specific criteria.
4. The resources needed to determine a WER.
5. The resources needed to use a WER.
6. The resources needed to monitor the acceptability of the
downstream water.
A WER determined using upstream water will usually be smaller,
less variable, and simpler to implement than a WER determined
using downstream water. Although in some situations a downstream
WER might be smaller than an upstream WER, the important
consideration is that a WER should be determined using the water
to which it is to apply.
References
U.S. EPA. 1983. Water Quality Standards Handbook. Office of
Water Regulations and Standards, Washington, DC.
U.S. EPA. 1984. Guidelines for Deriving Numerical Aquatic Site-
Specific Water Quality Criteria by Modifying National Criteria.
EPA-600/3-84-099 or PB85-121101. National Technical
Information Service, Springfield, VA.
U.S. EPA. 1992. Interim Guidance on Interpretation and
Implementation of Aquatic Life Criteria for Metals. Office of
Science and Technology, Health and Ecological Criteria Division,
Washington, DC.
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Appendix B: The Recalculation Procedure
NOTE: The National Toxics Rule (NTR) does not allow use of the
Recalculation Procedure in the derivation of a site-
specific criterion. Thus nothing in this appendix applies
to jurisdictions that are subject to the NTR.
The Recalculation Procedure is intended to cause a site-specific
criterion to appropriately differ from a national aquatic life
criterion if justified by demonstrated pertinent toxicological
differences between the aquatic species that occur at the site
and those that were used in the derivation of the national
criterion. There are at least three reasons why such differences
might exist between the two sets of species. First, the national
dataset contains aquatic species that are sensitive to many
pollutants, but these and comparably sensitive species might not
occur at the site. Second, a species that is critical at the
site might be sensitive to the pollutant and require a lower
criterion. (A critical species is a species that is commercially
or recreationally important at the site, a species that exists at
the site and is listed as threatened or endangered under section
4 of the Endangered Species Act, or a species for which there is
evidence that the loss of the species from the site is likely to
cause an unacceptable impact on a commercially or recreationally
important species, a threatened or endangered species, the
abundances of a variety of other species, or the structure or
function of the community.) Third, the species that occur at the
site might represent a narrower mix of species than those in the
national dataset due to a limited range of natural environmental
conditions. The procedure presented here is structured so that
corrections and additions can be made to the national dataset
without the deletion process being used to take into account taxa
that do and do not occur at the site; in effect, this procedure
makes it possible to update the national aquatic life criterion.
The phrase "occur at the site" includes the species, genera,
families, orders, classes, and phyla that:
a. are usually present at the site.
b. are present at the site only seasonally due to migration.
c. are present intermittently because they periodically return to
or extend their ranges into the site.
d. were present at the site in the past, are not currently
present at the site due to degraded conditions, and are
expected to return to the site when conditions improve.
e. are present in nearby bodies of water, are not currently
present at the site due to degraded conditions, and are
expected to be present at the site when conditions improve.
The taxa that "occur at the site" cannot be determined merely by
sampling downstream and/or upstream of the site at one point in
time. "Occur at the site" does not include taxa that were once
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present at the site but cannot exist at the site now due to
permanent physical alteration of the habitat at the site
resulting from dams, etc.
The definition of the "site" can be extremely important when
using theRecalculation Procedure. For example, the number of
taxa that occur at the site will generally decrease as the size
of the site decreases. Also, if the site is defined to be very
small, the permit limit might be controlled by a criterion that
applies outside (e.g., downstream of) the site.
Note: If the variety of aquatic invertebrates, amphibians, and
fishes is so limited that species in fewer than eight
families occur at the site, the general Recalculation
Procedure is not applicable and the following special
version of the Recalculation Procedure must be used:
1. Data must be available for at least one species in
each of the .families that occur at the site.
2. The lowest Species Mean Acute Value that is available
for a species that occurs at the site must be used as
the FAV.
3. The site-specific CMC and CCC must be calculated as
described below in part 2 of step E, which is titled
"Determination of the CMC and/or CCC".
The concept of the Recalculation Procedure is to create a dataset
that is appropriate for deriving a site-specific criterion by
modifying the national dataset in some or all of three ways:
a. Correction of data that are in the national dataset.
b. Addition of data to the national dataset.
c. Deletion of data that are in the national dataset.
All corrections and additions that have been approved by_U.S. EPA
are required, whereas use of the deletion process is optional.
The Recalculation Procedure is more likely to result in lowering
a criterion if the net result of addition and deletion is to
decrease the number of genera in the dataset, whereas the
procedure is more likely to result in raising a criterion if the
net result of addition and deletion is to increase the number of
genera in the dataset.
The Recalculation Procedure consists of the following steps:
A. Corrections are made in the national dataset.
B. Additions are made to the national dataset.
C. The deletion process may be applied if desired.
D. If the new dataset does not satisfy the applicable Minimum
Data Requirements (MDRs), additional pertinent data must be
generated; if the new data are approved by the U.S. EPA, the
Recalculation Procedure must be started again at step B with
the addition of the new data.
E. The new CMC or CCC or both are determined.
F. A report is written.
Each step is discussed in more detail below.
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A. Corrections
1. Only corrections approved by the U.S. EPA may be made.
2. The concept of "correction" includes removal of data that
should not have been in the national dataset in the first
place. The concept of "correction" does not include removal
of a datum from the national dataset just because the quality
of the datum is claimed to be suspect. If additional data are
available for the same species, the U.S. EPA will decide which
data should be used, based on the available guidance (U.S. EPA
1985); also, data based on measured concentrations are usually
preferable to those based on nominal concentrations.
3. Two kinds of corrections are possible:
a. The first includesthose corrections that are known to and
have been approved by the U.S. EPA; a list of these will be
available from the U.S. EPA.
b. The second includes those corrections that are submitted to
the U.S. EPA for approval. If approved, these will be
added to EPA's list of approved corrections.
4. Selective corrections are not allowed. All corrections on
EPA's newest list must be made.
B. Additions
1. Only additions approved by the U.S. EPA may be made.
2. Two kinds of additions are possible:
a. The first includes those additions that are known to and
have been approved by the U.S. EPA; a list of these will be
available from the U.S. EPA.
b. The second includes those additions that are submitted to
the U.S. EPA for approval. If approved, these will be
added to EPA's list of approved additions.
3. Selective additions are not allowed. All additions on EPA's
newest list must be made.
C. The Deletion Process
The basic principles are:
1. Additions and corrections must be made as per steps A and B
above, before the deletion process is performed.
2. Selective deletions are not allowed. If any species is to be
deleted, the deletion process described below must be applied
to all species in the national dataset, after any necessary
corrections and additions have been made to the national
dataset. The deletion process specifies which species must be
deleted and which species must not be deleted. Use of the
deletion process is optional, but no deletions are optional
when the deletion process is used.
3. Comprehensive information must be available concerning what
species occur at the site; a species cannot be deleted based
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on incomplete information concerning the species that do and
do not satisfy the definition of "occur at the site".
4. Data might have to be generated before the deletion process is
begun:
a. Acceptable pertinent toxicological data must be available
for at least one species in each class of aquatic plants,
invertebrates, amphibians, and fish that contains a species
that is a critical species at the site.
b. For each aquatic plant, invertebrate, amphibian, and fish
species that occurs at the site and is listed as threatened
or endangered under section 4 of the Endangered Species
Act, data must be available or be generated for an
acceptable surrogate species. Data for each surrogate
species must be used as if they are data for species that
occur at the site.
If additional data are generated using acceptable procedures
(U.S. EPA 1985) and they are approved by the U.S. EPA, the
Recalculation Procedure must be started again at step B with
the addition of the new data.
5. Data might have to be generated after the deletion process is
completed. Even if one or more species are deleted, there
still are MDRs (see step D below) that must be satisfied. If
the data remaining after deletion do not satisfy the
applicable MDRs, additional toxicity tests must be conducted
using acceptable procedures (U.S. EPA 1985) so that all MDRs
are satisfied. If the new data are approved by the U.S. EPA,
the Recalculation Procedure must be started again at step B
with the addition of new data.
6. Chronic tests do not have to be conducted because the national
Final Acute-Chronic Ratio (FACR) may be used in the derivation
of the site-specific Final Chronic Value (FCV). If acute-
chronic ratios (ACRs) are available or are generated so that
the chronic MDRs are satisfied using only species that occur
at the site, a site-specific FACR may be derived and used in
place of the national FACR. Because a FACR was not used in
the derivation of the freshwater CCC for cadmium, this CCC can
only be modified the same way as a FAV; what is acceptable
will depend on which species are deleted.
If any species are to be deleted, the following deletion process
must be applied:
a. Obtain a copy of the national dataset, i.e., tables 1, 2,
and 3 in the national criteria document (see Appendix E).
b. Make corrections in and/or additions to the national
dataset as described in steps A and B above.
c. Group all the species in the dataset taxonomically by
phylum, class, order, family, genus, and species.
d. Circle each species that satisfies the definition of "occur
at the site" as presented on the first page of this
appendix, and including any data for species that are
surrogates of threatened or endangered species that occur
at the site.
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e. Use the following step-wise process to determine
which of the uncircled species must be deleted and
which must not be deleted:
1. Does the genus occur at the site?
If "No", go to step 2.
If "Yes", are there one or more species in the genus
that occur at the site but are not in the
dataset?
If "No", go to step 2.
If "Yes", retain the uncircled species.*
2. Does the family occur at the site?
If "No", go to step 3.
If "Yes", are there one or more genera in the family
that occur at the site but are not in the
dataset?
If "No", go to step 3.
If "Yes", retain the uncircled species.*
3. Does the order occur at the site?
If "No", go to step 4.
If "Yes", does the dataset contain a circled species
that is in the same order?
If "No", retain the uncircled species.*
If "Yes", delete the uncircled species.*
4. Does the class occur at the site?
If "No", go to step 5.
If "Yes", does the dataset contain a circled species
that is in the same class?
If "No", retain the uncircled species.*
If "Yes", delete the uncircled species.*
5. Does the phylum occur at the site?
If "No", delete the uncircled species.*
If "Yes", does the dataset contain a circled species
that is in the same phylum?
If "No", retain the uncircled species.*
If "Yes", delete the uncircled species.*
* - Continue the deletion process by starting at step 1 for
another uncircled species unless all uncircled species
in. the dataset have been considered.
The species that are circled and those that are retained
constitute the site-specific dataset. (An example of the
deletion process is given in Figure Bl.)
This deletion process is designed to ensure that:
a. Each species that occurs both in the national dataset and
at the site also occurs in the site-specific dataset.
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b. Each species that occurs at the site but does not occur in
the national dataset is represented in the site-specific
dataset by all species in the national dataset that are in
the same genus.
c. Each genus that occurs at the site but does not occur in
the national dataset is represented in the site-specific
dataset by all genera in the national dataset that are in
the same family.
d. Each order, class, and phylum that occurs both in the
national dataset and at the site is represented in the
site-specific dataset by the one or more species in the
national dataset that are most closely related to a species
that occurs at the site.
D. Checking the Minimum Data Requirements
The initial MDRs for the Recalculation Procedure are the same as
those for the derivation of a national criterion. If a specific
requirement cannot be satisfied after deletion because that kind
of species does not occur at the site, a taxonomically similar
species must be substituted in order to meet the eight MDRs:
If no species of the kind required occurs at the site, but a
species in the same order does, the MDR can only be satisfied
by data for a species that occurs at the site and is in that
order; if no species in the order occurs at the site, but a
species in the class does, the MDR can only be satisfied by
data for a species that occurs at the site and is in that
class. If no species in the same class occurs at the site,
but a species in the phylum does, the MDR can only be
satisfied by data for a species that occurs at the site and is
in that phylum. If no species in the same phylum occurs at
the site, any species that occurs at the site and is not used
to satisfy a different MDR can be used to satisfy the MDR. If
additional data are generated using acceptable procedures
(U.S. EPA 1985) and they are approved by the U.S. EPA, the
Recalculation Procedure must be started again at step B with
the addition of the new data.
If fewer than eight families of aquatic invertebrates,
amphibians, and fishes occur at the site, a Species Mean Acute
Value must be available for at least one species in each of the
families and the special version of the Recalculation Procedure
described on the second page of this appendix must be used.
E. Determining the CMC and/or CCC
1. Determining the FAV:
a. If the eight family MDRs are satisfied, the site-specific
FAV must be calculated from Genus Mean Acute Values using
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the procedure described in the national aquatic life
guidelines (U.S. EPA 1985).
b. If fewer than eight families of aquatic invertebrates,
amphibians, and fishes occur at the site, the lowest
Species Mean Acute Value that is available for a species
that occurs at the site must be used as the FAV, as per the
special version of the Recalculation Procedure described on
the second page of this appendix.
2. The site-specific CMC must be calculated by dividing the site-
specific FAV by 2. The site-specific FCV must be calculated
by dividing the site-specific FAV by the national FACR (or by
a site-specific FACR if one is derived). (Because a FACR was
not used to derive the national CCC for cadmium in fresh
water, the site-specific CCC equals the site-specific FCV. )
3. The calculated FAV, CMC, and/or CCC must be lowered, if
necessary, to (1) protect an aquatic plant, invertebrate,
amphibian, or fish species that is a critical species at the
site, and (2) ensure that the criterion is not likely to
jeopardize the continued existence of any endangered or
threatened species listed under section 4 of the Endangered
Species Act or result in the destruction or adverse
modification of such species' critical habitat.
F. Writing the Report
The report of the results of use of the Recalculation Procedure
must include:
1. A list of all species of aquatic invertebrates, amphibians,
and fishes that are known to "occur at the site", along with
the source of the information.
2. A list of all aquatic plant, invertebrate, amphibian, and fish
species that are critical species at the site, including all
species that occur at the site and are listed as threatened or
endangered under section 4 of the Endangered Species Act.
3. A site-specific version of Table 1 from a criteria document
produced by the U.S. EPA after 1984.
4. A site-specific version of Table 3 from a criteria document
produced by the U.S. EPA after 1984.
5. A list of all species that were deleted.
6. The new calculated FAV, CMC, and/or CCC.
7. The lowered FAV, CMC, and/or CCC, if one or more were lowered
to protect a specific species.
Reference
U.S. EPA. 1985. Guidelines for Deriving Numerical National
Water Quality Criteria for the Protection of Aquatic Organisms
and Their Uses. PB85-227049. National Technical Information
Service, Springfield, VA.
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Figure Bl: An Example of the Deletion Process Using Three Phyla
SPECIES THAT ARE IN THE THREE PHYLA AND OCCUR AT THE SITE
Phylum Class Order Family Species
Annelida
Bryozoa
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Hirudin. Rhynchob.
(No species in this
Osteich. Cyprinif.
Osteich. Cyprinif.
Osteich. Cyprinif.
Osteich. Cyprinif.
Osteich. Salmonif.
Osteich.
Osteich.
Percifor.
Percifor.
Amphibia Caudata
Glossiph. Glossip. complanata
phylum occur at the site.)
Cyprinid. Carassius auratus
Cyprinid. Notropis anogenus
Cyprinid. Phoxinus eos
Catostom. Carpiodes carpio
Osmerida. Osmerus mordax
Centrarc. Lepomis cyanellus
Centrarc. Lepomis humilis
Ambystom. Ambystoma gracile
SPECIES THAT ARE IN THE THREE PHYLA AND IN THE NATIONAL DATASET
Phylum Class Order Family Species Code
Tubifex tubifex P
Lophopod. carteri D
Petromyzon marinus D
Carassius auratus S
Notropis hudsonius G
Notropis stramineus G
Phoxinus eos S
Phoxinus oreas D
Tinea tinea D
Ictiobus bubalus F
Oncorhynchus mykiss O
Lepomis cyanellus S
Lepomis macrochirus G
Perca flavescens D
Xenopus laevis C
Annelida
Bryozoa
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Chordata
Oligoch.
Phylact .
Cephala.
Osteich.
Osteich.
Osteich.
Osteich.
Osteich.
Osteich.
Osteich.
Osteich.
Osteich.
Osteich.
Osteich.
Amphibia
Haplotax.
Petromyz .
Cyprinif .
Cyprinif .
Cyprinif.
Cyprinif .
Cyprinif .
Cyprinif .
Cyprinif .
Salmonif .
Percifor.
Percifor.
Percifor.
Anura
Tubifici
Lophopod
Petromyz
Cyprinid
Cyprinid
Cyprinid
Cyprinid
Cyprinid
Cyprinid
Catostom
Salmonid
Centrarc
Centrarc
Percidae
Pipidae
Explanations of Codes:
S = retained because this Species occurs at the site.
G = retained because there is a species in this Genus that
occurs at the site but not in the national dataset.
F = retained because there is a genus in this Family that
occurs at the site but not in the national dataset.
O = retained because this Order occurs at the site and is not
represented by a lower taxon.
C = retained because this Class occurs at the site and is not
represented by a lower taxon.
P = retained because this Phylum occurs at the site and is not
represented by a lower taxon.
D = deleted because this species does not satisfy any of the
requirements for retaining species.
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Appendix C: Guidance Concerning the Use of "Clean Techniques" and
QA/QC when Measuring Trace Metals
Note: This version of this appendix contains more information
than the version that was Appendix B of Prothro (1993).
Recent information (Shiller and Boyle 1987; Windom et al. 1991)
has raised questions concerning the quality of reported
concentrations of trace metals in both fresh and salt (estuarine
and marine) surface waters. A lack of awareness of true ambient
concentrations of metals in fresh and salt surface waters can be
both a cause and a result of the problem. The ranges of
dissolved metals that are typical in surface waters of the United
States away from the immediate influence of discharges (Bruland
1983; Shiller and Boyle 1985,1987; Trefry et al. 1986; Windom et
al. 1991) are:
Metal Salt water Fresh water
(ug/L) (ug/L)
Cadmium 0.01 to 0.2 0.002 to 0.08
Copper 0.1 to 3. 0.4 to 4.
Lead 0.01 to 1. 0.01 to 0.19
Nickel 0.3 to 5. 1. to 2.
Silver 0.005 to 0.2
Zinc 0.1 to 15. 0.03 to 5.
The U.S. EPA (1983,1991) has published analytical methods for
monitoring metals in waters and wastewaters, but these methods
are inadequate for determination of ambient concentrations of
some metals in some surface waters. Accurate and precise
measurement of these low concentrations requires appropriate
attention to seven areas:
1. Use of "clean techniques" during collecting, handling,
storing, preparing, and analyzing samples to avoid
contamination.
2. Use of analytical methods that have sufficiently low detection
limits.
3. Avoidance of interference in the quantification (instrumental
analysis) step.
4. Use of blanks to assess contamination.
5. Use of matrix spikes (sample spikes) and certified reference
materials (CRMs) to assess interference and contamination.
6. Use of replicates to assess precision.
7. Use of certified standards.
In a strict sense, the term "clean techniques" refers to
techniques that reduce contamination and enable the accurate and
precise measurement of trace metals in fresh and salt surface
waters. In a broader sense, the term also refers to related
issues concerning detection limits, quality control, and quality
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assurance. Documenting data quality demonstrates the amount of
confidence that can be placed in the data, whereas increasing the
sensitivity of methods reduces the problem of deciding how to
interpret results that are reported to be below detection limits.
This appendix is written for those analytical laboratories that
want guidance concerning ways to lower detection limits, increase
accuracy, and/or increase precision. The ways to achieve these
goals are to increase the sensitivity of the analytical methods,
decrease contamination, and decrease interference. Ideally,
validation of a procedure for measuring concentrations of metals
in surface water requires demonstration that agreement can be
obtained using completely different procedures beginning with the
sampling step and continuing through the quantification step
(Bruland et al. 1979), but few laboratories have the resources to
compare two different procedures. Laboratories can, however, (a)
use techniques that others have found useful for improving
detection limits, accuracy, and precision, and (b) document data
quality through use of blanks, spikes, CRMs, replicates, and
standards.
Nothing contained or not contained in this appendix adds to or
subtracts from any regulatory requirement set forth in other EPA
documents concerning analyses of metals. A WER can be acceptably
determined without the use of clean techniques as long as the
detection limits, accuracy, and precision are acceptable. No
QA/QC requirements beyond those that apply to measuring metals in
effluents are necessary for the determination of WERs. The word
"must" is not used in this appendix. Some items, however, are
considered so important by analytical chemists who have worked to
increase accuracy and precision and lower detection limits in
trace-metal analysis that "should" is in bold print to draw
attention to the item. Most such items are emphasized because
they have been found to have received inadequate attention in
some laboratories performing trace-metal analyses.
In general, in order to achieve accurate and precise measurement
of a particular concentration, both the detection limit and the
blanks should be less than one-tenth of that concentration.
Therefore, the term "metal-free" can be interpreted to mean that
the total amount of contamination that occurs during sample
collection and processing (e.g., from gloves, sample containers,
labware, sampling apparatus, cleaning solutions, air, reagents,
etc.) is sufficiently low that blanks are less than one-tenth of
the lowest concentration that needs to be measured.
Atmospheric particulates can be a major source of contamination
(Moody 1982; Adeloju and Bond 1985). The term "class-100" refers
to a specification concerning the amount of particulates in air
(Moody 1982); although the specification says nothing about the
composition of the particulates, generic control of particulates
can greatly reduce trace-metal blanks. Except during collection
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of samples, initial cleaning of equipment, and handling of
samples containing high concentrations of metals, all handling of
samples, sample containers, labware, and sampling apparatus
should be performed in a class-100 bench, room, or glove box.
Neither the "ultraclean techniques" that might be necessary when
trace analyses of mercury are performed nor safety in analytical
laboratories is addressed herein. Other documents should be
consulted if one or both of these topics are of concern.
Avoiding contamination by use of "clean techniques"
Measurement of trace metals in surface waters should take into
account the potential for contamination during each step in the
process. Regardless of the specific procedures used for
collection, handling, storage, preparation (digestion,
filtration, and/or extraction), and quantification (instrumental
analysis), the general principles of contamination control should
be applied. Some specific recommendations are:
a. Powder-free (non-talc, class-100) latex, polyethylene, or
polyvinyl chloride (PVC, vinyl) gloves should be worn during
all steps from sample collection to analysis. (Talc seems to
be a particular problem with zinc; gloves made with talc
cannot be decontaminated sufficiently.) Gloves should only
contact surfaces that are metal-free; gloves should be changed
if even suspected of contamination.
b. The acid used to acidify samples for preservation and
digestion and to acidify water for final cleaning of labware,
sampling apparatus, and sample containers should be metal-
free. The quality of the acid used should be better than
reagent-grade. Each lot of acid should be cinalyzed for the
metal(s) of interest before use.
c. The water used to prepare acidic cleaning solutions and to
rinse labware, sample containers, and sampling apparatus may
be prepared by distillation, deionization, or reverse osmosis,
and should be demonstrated to be metal-free.
d. The work area, including bench tops and hoods, should be
cleaned (e.g., washed and wiped dry with lint-free, class-100
wipes) frequently to remove contamination.
e. All handling of samples in the laboratory, including filtering
and analysis, should be performed in a class-100 clean bench
or a glove box fed by particle-free air or nitrogen; ideally
the clean bench or glove box should be located within a class-
100 clean room.
f. Labware, reagents, sampling apparatus, and sample containers
should never be left open to the atmosphere; they should be
stored in a class-100 bench, covered with plastic wrap, stored
in a plastic box, or turned upside down on a clean surface.
Minimizing the time between cleaning and using will help
minimize contamination.
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g. Separate sets of sample containers, labware, and sampling
apparatus should be dedicated for different kinds of samples,
e.g., surface water samples, effluent samples, etc.
h. To avoid contamination of clean rooms, samples that contain
very high concentrations of metals and do not require use of
"clean techniques" should not be brought into clean rooms.
i. Acid-cleaned plastic, such as high-density polyethylene
(HDPE), low-density polyethylene (LDPE), or a fluoroplastic,
should be the only material that ever contacts a sample,
except possibly during digestion for the total recoverable
measurement.
1. Total recoverable samples can be digested in some plastic
containers.
2. HDPE and LDPE might not be acceptable for mercury.
3. Even if acidified, samples and standards containing silver
should be in amber containers.
j. All labware, sample containers, and sampling apparatus should
be acid-cleaned before use or reuse.
1. Sample containers, sampling apparatus, tubing, membrane
filters, filter assemblies, and other labware should be
soaked in acid until metal-free. The amount of cleaning
necessary might depend on the amount of contamination and
the length of time the item will be in contact with
samples. For example, if an acidified sample will be
stored in a sample container for three weeks, ideally the
container should have been soaked in an acidified metal-
free solution for at least three weeks.
2. It might be desirable to perform initial cleaning, for
which reagent-grade acid may be used, before the items are
taken into a clean room. For most metals, items should be
either (a) soaked in 10 percent concentrated nitric acid at
50°C for at least one hour, or (b) soaked in 50 percent
concentrated nitric acid at room temperature for at least
two days; for arsenic and mercury, soaking for up to two
weeks at 50°C in 10 percent concentrated nitric acid might
be required. For plastics that might be damaged by strong
nitric acid, such as polycarbonate and possibly HDPE and
LDPE, soaking in 10 percent concentrated hydrochloric acid,
either in place of or before soaking in a nitric acid
solution, might be desirable.
3. Chromic acid should not be used to clean items that will be
used in analysis of metals.
4. Final soaking and cleaning of sample containers, labware,
and sampling apparatus should be performed in a class-100
clean room using metal-free acid and water. The solution
in an acid bath should be analyzed periodically to
demonstrate that it is metal-free.
k. Labware, sampling apparatus, and sample containers should be
stored appropriately after cleaning:
1. After the labware and sampling apparatus are cleaned, they
may be stored in a clean room in a weak acid bath prepared
using metal-free acid and water. Before use, the items
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should be rinsed at least three times with metal-free water.
After the final rinse, the items should be moved immediately,
with the open end pointed down, to a class-100 clean bench.
Items may be dried on a class-100 clean bench; items should
not be dried in an oven or with laboratory towels. The
sampling apparatus should be assembled in a class-100 clean
room or bench and double-bagged in metal-free polyethylene
zip-type bags for transport to the field; new bags are usually
metal-free.
2. After sample containers are cleaned, they should be filled
with metal-free water that has been acidified to a pH of 2
with metal-free nitric acid (about 0.5 mL per liter) for
storage until use.
1. Labware, sampling apparatus, and sample containers should be
rinsed and not rinsed with sample as necesssiry to prevent high
and low bias of analytical results because acid-cleaned
plastic will sorb some metals from unacidified solutions.
1. Because samples for the dissolved measurement are not
acidified until after filtration, all sampling apparatus,
sample containers, labware, filter holders, membrane
filters, etc., that contact the sample before or during
filtration should be rinsed with a portion of the solution
and then that portion discarded.
2. For the total recoverable measurement, laibware, etc., that
contact the sample only before it is acidified should be
rinsed with sample, whereas items that contact the sample
after it is acidified should not be rinsed. For example,
the sampling apparatus should be rinsed because the sample
will not be acidified until it is in a sample container,
but the sample container should not be rinsed if the sample
will be acidified in the sample container.
3. If the total recoverable and dissolved measurements are to
be performed on the same sample (rather than on two samples
obtained at the same time and place), all the apparatus and
labware, including the sample container, should be rinsed
before the sample is placed in the sample, container; then
an unacidified aliquot should be removed for the total
recoverable measurement (and acidified, digested, etc.) and
an unacidified aliquot should be removed for the dissolved
measurement (and filtered, acidified, etc.) (If a
container is rinsed and filled with sample and an
unacidified aliquot is removed for the dissolved
measurement and then the solution in the container is
acidified before removal of an aliquot for the total
recoverable measurement, the resulting measured total
recoverable concentration might be biased high because the
acidification might desorb metal that had been sorbed onto
the walls of the sample container; the amount of bias will
depend on the relative volumes involved eind on the amount
of sorption and desorption.)
m. Field samples should be collected in a manner that eliminates
the potential for contamination from sampling platforms,
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probes, etc. Exhaust from boats and the direction of wind and
water currents should be taken into account. The people who
collect the samples should be specifically trained on how to
collect field samples. After collection, all handling of
samples in the field that will expose the sample to air should
be performed in a portable class-100 clean bench or glove box.
n. Samples should be acidified (after filtration if dissolved
metal is to be measured) to a pH of less than 2, except that
the pH should be less than 1 for mercury. Acidification
should be done in a clean room or bench, and so it might be
desirable to wait and acidify samples in a laboratory rather
than in the field. If samples are acidified in the field,
metal-free acid can be transported in plastic bottles and
poured into a plastic container from which acid can be removed
and added to samples using plastic pipettes. Alternatively,
plastic automatic dispensers can be used.
o. Such things as probes and thermometers should not be put in
samples that are to be analyzed for metals. In particular, pH
electrodes and mercury-in-glass thermometers should not be
used if mercury is to be measured. If pH is measured, it
should be done on a separate aliquot.
p. Sample handling should be minimized. For example, instead of
pouring a sample into a graduated cylinder to measure the
volume, the sample can be weighed after being ptoured into a
tared container, which is less likely to be subject to error
than weighing the container from which the sample is poured.
(For saltwater samples, the salinity or density should be
taken into account if weight is converted to volume.)
q. Each reagent used should be verified to be metal-free. If
metal-free reagents are not commercially available, removal of
metals will probably be necessary.
r. For the total recoverable measurement, samples should be
digested in a class-100 bench, not in a metallic hood. If
feasible, digestion should be done in the sample container by
acidification and heating.
s. The longer the time between collection and analysis of
samples, the greater the chance of contamination, loss, etc.
t. Samples should be stored in the dark, preferably between 0 and
4°C with no air space in the sample container.
Achieving low detection limits
a. Extraction of the metal from the sample can be extremely
useful if it simultaneously concentrates the metal and
eliminates potential matrix interferences. For example,
ammonium 1-pyrrolidinedithiocarbamate and/or diethylammonium
diethyldithiocarbamate can extract cadmium, copper, lead,
nickel, and zinc (Bruland et al. 1979; Nriagu et al. 1993).
b. The detection limit should be less than ten percent of the
lowest concentration that is to be measured.
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Avoiding interferences
a. Potential interferences should be assessed for the specific
instrumental analysis technique used and for each metal to be
measured.
b. If direct analysis is used, the salt present in high-salinity
saltwater samples is likely to cause interference in most
instrumental techniques.
c. As stated above, extraction of the metal from the sample is
particularly useful because it simultaneously Concentrates the
metal and eliminates potential matrix interferences.
Using blanks to assess contamination
a. A laboratory (procedural, method) blank consists of filling a
sample container with analyzed metal-free water and processing
(filtering, acidifying, etc.) the water through the laboratory
procedure in exactly the same way as a sample. A laboratory
blank should be included in each set of ten or fewer samples
to check for contamination in the laboratory, and should
contain less than ten percent of the lowest concentration that
is to be measured. Separate laboratory blanks should be
processed for the total recoverable and dissolved
measurements, if both measurements are performed.
b. A field (trip) blank consists of filling a sample container
with analyzed metal-free water in the laboratory, taking the
container to the site, processing the water through tubing,
filter, etc., collecting the water in a sample container, and
acidifying the water the same as a field sample. A field
blank should be processed for each sampling trip. Separate
field blanks should be processed for the total recoverable
measurement and for the dissolved measurement, if filtrations
are performed at the site. Field blanks should be processed
in the laboratory the same as laboratory blanks.
Assessing accuracy
a. A calibration curve should be determined for each analytical
run and the calibration should be checked about every tenth
sample. Calibration solutions should be traceable back to a
certified standard from the U.S. EPA or the National Institute
of Science and Technology (NIST).
b. A blind standard or a blind calibration solution should be
included in each group of about twenty samples.
c. At least one of the following should be included in each group
of about twenty samples:
1. A matrix spike (spiked sample; the method of known
additions).
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2. A CRM, if one is available in a matrix that closely
approximates that of the samples. Values obtained for the
CRM should be within the published values.
The concentrations in blind standards and solutions, spikes, and
CRMs should not be more than 5 times the median concentration
expected to be present in the samples.
Assessing precision
a. A sampling replicate should be included with each set of
samples collected at each sampling location.
b. If the volume of the sample is large enough, replicate
analysis of at least one sample should be performed along with
each group of about ten samples.
Special considerations concerning the dissolved measurement
Whereas total recoverable measurements are especially subject to
contamination during digestion, dissolved measurements are
subject to both loss and contamination during filtration.
a. Because acid-cleaned plastic sorbs metal from unacidified
solutions and because samples for the dissolved measurement
are not acidified before filtration, all sampling apparatus,
sample containers, labware, filter holders, and membrane
filters that contact the sample before or during filtration
should be conditioned by rinsing with a portion of the
solution and discarding that portion.
b. Filtrations should be performed using acid-cleaned plastic
filter holders and acid-cleaned membrane filters. Samples
should not be filtered through glass fiber filters, even if
the filters have been cleaned with acid. If positive-pressure
filtration is used, the air or gas should be passed through a
0.2-ptm in-line filter; if vacuum filtration is used, it should
be performed on a class-100 bench.
c. Plastic filter holders should be rinsed and/or dipped between
filtrations, but they do not have to be soaked between
filtrations if all the samples contain about the same
concentrations of metal. It is best to filter samples from
low to high concentrations. A membrane filter should not be
used for more than one filtration. After each filtration, the
membrane filter should be removed and discarded, and the
filter holder should be either rinsed with metal-free water or
dilute acid and dipped in a metal-free acid bath or rinsed at
least twice with metal-free dilute acid; finally, the filter
holder should be rinsed at least twice with metal-free water.
d. For each sample to be filtered, the filter holder and membrane
filter should be conditioned with the sample, i.e., an initial
portion of the sample should be filtered and discarded.
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The accuracy and precision of the dissolved measurement should be
assessed periodically. A large volume of a buffered solution
(such as aerated 0.05 N sodium bicarbonate for analyses in fresh
water and a combination of sodium bicarbonate and sodium chloride
for analyses in salt water) should be spiked so that the
concentration of the metal of interest is in the range of the low
concentrations that are to be measured. Sufficient samples
should be taken alternately for (a) acidification in the same way
as after filtration in the dissolved method and (b) filtration
and acidification using the procedures specified in the dissolved
method until ten samples have been processed in each way. The
concentration of metal in each of the twenty samples should then
be determined using the same analytical procedure. The means of
the two groups of ten measurements should be within 10 percent,
and the coefficient of variation for each group of ten should be
less than 20 percent. Any values deleted as outliers should be
acknowledged.
Reporting results
To indicate the quality of the data, reports of results of
measurements of the concentrations of metals should include a
description of the blanks, spikes, CRMs, replicates, and
standards that were run, the number run, and the results
obtained. All values deleted as outliers should be acknowledged.
Additional information
The items presented above are some of the important aspects of
"clean techniques"; some aspects of quality assurance and quality
control are also presented. This is not a definitive treatment
of these topics; additional information that might be useful is
available in such publications as Patterson and" Settle (1976),
Zief and Mitchell (1976), Bruland et al. (1979), Moody and Beary
(1982), Moody (1982), Bruland (1983), Adeloju arid Bond (1985),
Berman and Yeats (1985), Byrd and Andreae (1986) , Taylor (1987),
Sakamoto-Arnold (1987), Tramontane et al. (1987), Puls and
Barcelona (1989), Windom et al. (1991), U.S. EPA (1992), Horowitz
et al. (1992), and Nriagu et al. (1993).
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References
Adeloju, S.B., and A.M. Bond. 1985. Influence of Laboratory
Environment on the Precision and Accuracy of Trace Element
Analysis. Anal. Chem. 57:1728-1733.
Herman, S.S., and P.A. Yeats. 1985. Sampling of Seawater for
Trace Metals. CRC Reviews in Analytical Chemistry 16:1-14.
Bruland, K.W., R.P. Franks, G.A. Knauer, and J.H. Martin. 1979.
Sampling and Analytical Methods for the Determination of Copper,
Cadmium, Zinc, and Nickel at the Nanogram per Liter Level in Sea
Water. Anal. Chim. Acta 105:233-245.
Bruland, K.W. 1983. Trace Elements in Sea-water. In: Chemical
Oceanography, Vol. 8. (J.P. Riley and R. Chester, eds.)
Academic Press, New York, NY. pp. 157-220.
Byrd, J.T., and M.O. Andreae. 1986. Dissolved and Particulate
Tin in North Atlantic Seawater. Marine Chem. 19:193-200.
Horowitz, A.J., K.A. Elrick, and M.R. Colberg. 1992. The Effect
of Membrane Filtration Artifacts on Dissolved Trace Element
Concentrations. Water Res. 26:753-763.
Moody, J.R. 1982. NBS Clean Laboratories for Trace Element
Analysis. Anal. Chem. 54:1358A-1376A.
Moody, J.R., and E.S. Beary. 1982. Purified Reagents for Trace
Metal Analysis. Talanta 29:1003-1010.
Nriagu, J.O., G. Lawson, H.K.T. Wong, and J.M. Azcue. 1993. A
Protocol for Minimizing Contamination in the Analysis of Trace
Metals in Great Lakes Waters. J. Great Lakes Res. 19:175-182.
Patterson, C.C., and D.M. Settle. 1976. The Reduction in Orders
of Magnitude Errors in Lead Analysis of Biological Materials and
Natural Waters by Evaluating and Controlling the Extent and
Sources of Industrial Lead Contamination Introduced during Sample
Collection and Processing. In: Accuracy in Trace Analysis:
Sampling, Sample Handling, Analysis. (P.O. LaFleur, ed.)
National Bureau of Standards Spec. Publ. 422, U.S. Government
Printing Office, Washington, DC.
Prothro, M.G. 1993. Memorandum titled "Office of Water Policy
and Technical Guidance on Interpretation and Implementation of
Aquatic Life Metals Criteria". October 1.
Puls, R.W.,- and M.J. Barcelona. 1989. Ground Water Sampling for
Metals Analyses. EPA/540/4-89/001. National Technical
Information Service, Springfield, VA.
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Sakamoto-Arnold, C.M., A.K. Hanson, Jr., D.L. Huizenga, and D.R.
Kester. 1987. Spatial and Temporal Variability of Cadmium in
Gulf Stream Warm-core Rings and Associated Waters. J. Mar. Res
45:201-230.
Shiller, A.M., and E. Boyle.
Nature 317:49-52.
1985. Dissolved Zinc in Rivers.
Shiller, A.M., and E.A. Boyle. 1987. Variability of Dissolved
Trace Metals in the Mississippi River. Geochim. Cosmochim. Acta
51:3273-3277.
Taylor, J.K. 1987. Quality Assurance of Chemical Measurements.
Lewis Publishers, Chelsea, MI.
Tramontane, J.M., J.R. Scudlark, and T.M. Church. 1987. A
Method for the Collection, Handling, and Analysis of Trace Metals
in Precipitation. Environ. Sci. Technol. 21:749-753.
Trefry, J.H., T.A. Nelsen, R.P. Trocine, S. Metz., and T.W.
Vetter. 1986. Trace Metal Fluxes through the Mississippi River
Delta System. Rapp. P.-v. Reun. Cons. int. Explor. Mer. 186:277-
288.
U.S. EPA. 1983. Methods for Chemical Analysis of Water and
Wastes. EPA-600/4-79-020. National Technical Information
Service, Springfield, VA. Sections 4.1.1, 4.1.3, and 4.1.4
U.S. EPA. 1991. Methods for the Determination of Metals in
Environmental Samples. EPA-600/4-91-010. National Technical
Information Service, Springfield, VA.
U.S. EPA. 1992. Evaluation of Trace-Metal Levels in Ambient
Waters and Tributaries to New York/New Jersey Harbor for Waste
Load Allocation. Prepared by Battelle Ocean Sciences under
Contract No. 68-C8-0105.
Windom, H.L., J.T. Byrd, R.G. Smith, and F. Huari. 1991.
Inadequacy of NASQAN Data for Assessing Metals Trends in the
Nation's Rivers. Environ. Sci. Technol. 25:1137-1142. (Also see
the comment and response: Environ. Sci. Technol. 25:1940-1941.)
Zief, M., and J.W. Mitchell. 1976. Contamination Control in
Trace Element Analysis. Chemical Analysis Series, Vol. 47.
Wiley, New York, NY.
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Appendix D: Relationships between WERs and the Chemistry and
Toxicology of Metals
The aquatic toxicology of metals is complex in part because the
chemistry of metals in water is complex. Metals usually exist in
surface water in various combinations of particulate and
dissolved forms, some of which are toxic and some of which are
nontoxic. In addition, all toxic forms of a metal are not
necessarily equally toxic, and various water quality
characteristics can affect the relative concentrations and/or
toxicities of some of the forms.
The toxicity of a metal has sometimes been reported to be
proportional to the concentration or activity of a specific
species of the metal. For example, Allen and Hansen (1993)
summarized reports by several investigators that the toxicity of
copper is related to the free cupric ion, but other data do not
support a correlation (Erickson 1993a). For example, Borgmann
(1983) , Chapman and McCrady (1977) , and French and Hunt (1986)
found that toxicity expressed on the basis of cupric ion activity
varied greatly with pH, and Cowan et al. (1986) concluded that at
least one of the copper hydroxide species is toxic. Further,
chloride and sulfate salts of calcium, magnesium, potassium, and
sodium affect the toxicity of the cupric ion (Nelson et al.
1986). Similarly for aluminum, Wilkinson et al. (1993) concluded
that "mortality was best predicted not by the free A13+ activity
but rather as a function of the sum s ( [A13+] + [A1F2+] ) " and that
"no longer can the reduction of Al toxicity in the presence of
organic acids be interpreted simply as a consequence of the
decrease in the free A13+ concentration" .
Until a model has been demonstrated to explain the quantitative
relationship between chemical and toxicological measurements,
aquatic life criteria should be established in an environmentally
conservative manner with provision for site-specific adjustment.
Criteria should be expressed in terms of feasible analytical
measurements that provide the necessary conservatism without
substantially increasing the cost of implementation and site-
specific adjustment. Thus current aquatic life criteria for
metals are expressed in terms of the total recoverable
measurement and/or the dissolved measurement, rather than a
measurement that would be more difficult to perform and would
still require empirical adjustment. The WER is operationally
defined in terms of chemical and toxicological measurements to
allow site-specific adjustments that account for differences
between the toxicity of a metal in laboratory dilution water and
in site water.
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Forms of Metals
Even if the relationship of toxicity to the forms of metals is
not understood well enough to allow setting site-specific water
quality criteria without using empirical adjustments, appropriate
use and interpretation of WERs requires an understanding of how
changes in the relative concentrations of different forms of a
metal might affect toxicity. Because WERs are defined on the
basis of relationships between measurements of toxicity and
measurements of total recoverable and/or dissolved metal, the
toxicologically relevant distinction is between the forms of the
metal that are toxic and nontoxic whereas the chemically relevant
distinction is between the forms that are dissolved and
particulate. "Dissolved metal" is defined here as "metal that
passes through either a 0.45-^m or a 0.40-/an membrane filter" and
"particulate metal" is defined as "total recoverable metal minus
dissolved metal". Metal that is in or on particles that pass
through the filter is operationally defined as "dissolved".
In addition, some species of metal can be converted from one form
to another. Some conversions are the result of reequilibration
in response to changes in water quality characteristics whereas
others are due to such fate processes as oxidation of sulfides
and/or organic matter. Reequilibration usually occurs faster
than fate processes and probably results in any rapid changes
that are due to effluent mixing with receiving water or changes
in pH at a gill surface. To account for rapid changes due to
reequilibration, the terms "labile" and "refractory" will be used
herein to denote metal species that do and do not readily convert
to other species when in a nonequilibrium condition, with
"readily" referring to substantial progression toward equilibrium
in less than about an hour. Although the toxicity and lability
of a form of a metal are not merely yes/no properties, but rather
involve gradations, a simple classification scheme such as this
should be sufficient to establish the principles regarding how
WERs are related to various operationally defined forms of metal
and how this affects the determination and use of WERs.
Figure Dl presents the classification scheme that results from
distinguishing forms of metal based on analytical methodology,
toxicity tests, and lability, as described above. Metal that is
not measured by the total recoverable measurement is assumed to
be sufficiently nontoxic and refractory that it will not be
further considered here. Allowance is made for toxicity due to
particulate metal because some data indicate that particulate
metal might contribute to toxicity and bioaccumulation, although
other data imply that little or no toxicity can be ascribed to
particulate metal (Erickson 1993b). Even if the toxicity of
particulate metal is not negligible in a particular situation, a
dissolved criterion will not be underprotective if the dissolved
criterion was derived using a dissolved WER (see below) or if
there are sufficient compensating factors.
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Figure Dl: A Scheme for Classifying Forms of Metal in Water
Total recoverable metal
Dissolved
Nontoxic
Labile
Refractory
Toxic
Labile
Particulate
Nontoxic
Labile
Refractory
Toxic
Labile
Metal not measured by the total recoverable measurement
Not only can some changes in water quality characteristics shift
the relative concentrations of toxic and nontoxic labile species
of a metal, some changes in water quality can also increase or
decrease the toxicities of the toxic species of a metal and/or
the sensitivities of aquatic organisms. Such changes might be
caused by (a) a change in ionic strength that affects the
activity of toxic species of the metal in water, (b) a
physiological effect whereby an ion affects the permeability of a
membrane and thereby alters both uptake and apparent toxicity,
and (c) toxicological additivity, synergism, or antagonism due to
effects within the organism.
Another possible complication is that a form of metal that is
toxic to one aquatic organism might not be toxic to another.
Although such differences between organisms have not been
demonstrated, the possibility cannot be ruled out.
The Importance of Lability
The only common metal measurement that can be validly
extrapolated from the effluent and the upstream water to the
downstream water merely by taking dilution into account is the
total recoverable measurement. A major reason this measurement
is so useful is because it is the only measurement that obeys the
law of mass balance (i.e., it is the only measurement that is
conservative). Other metal measurements usually do not obey the
law of mass balance because they measure some, but not all, of
the labile species of metals. A measurement of refractory metal
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would be conservative in terms of changes in water quality
characteristics, but not necessarily in regards to fate
processes; such a measurement has not been developed, however.
Permit limits apply to effluents, whereas water quality criteria
apply to surface waters. If permit limits and water quality
criteria are both expressed in terms of total recoverable metal,
extrapolations from effluent to surface water only need to take
dilution into account and can be performed as mass balance
calculations. If either permit limits or water quality criteria
or both are expressed in terms of any other metal measurement,
lability needs to be taken into account, even if both are
expressed in terms of the same measurement.
Extrapolations concerning labile species of metals from effluent
to surface water depend to a large extent on the differences
between the water quality characteristics of the effluent and
those of the surface water. Although equilibrium models of the
speciation of metals can provide insight, the interactions are
too complex to be able to make useful nonempirical extrapolations
from a wide variety of effluents to a wide variety of surface
waters of either (a) the speciation of the metal or (b) a metal
measurement other than total recoverable.
Empirical extrapolations can be performed fairly easily and the
most common case will probably occur when permit limits are based
on the total recoverable measurement but water quality criteria
are based on the dissolved measurement. The empirical
extrapolation is intended to answer the question "What percent of
the total recoverable metal in the effluent becomes dissolved in
the downstream water?" This question can be answered by:
a. Collecting samples of effluent and upstream water.
b. Measuring total recoverable metal and dissolved metal in both
samples.
c. Combining aliquots of the two samples in the. ratio of the
flows when the samples were obtained and mixing for an
appropriate period of time under appropriate conditions.
d. Measuring total recoverable metal and dissolved metal in the
mixture.
An example is presented in Figure D2. This percentage cannot be
extrapolated from one metal to another or from one effluent to
another. The data needed to calculate the percentage will be
obtained each time a WER is determined using simulated downstream
water if both dissolved and total recoverable metal are measured
in the effluent, upstream water, and simulated downstream water.
The interpretation of the percentage is not necessarily as
straightforward as might be assumed. For example, some of the
metal that is dissolved in the upstream water might sorb onto
particulate matter in the effluent, which can be viewed as a
detoxification of the upstream water by the effluent. Regardless
of the interpretation, the described procedure provides a simple
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way of relating the total recoverable concentration in the
effluent to the concentration of concern in the downstream water.
Because this empirical extrapolation can be used with any
analytical measurement that is chosen as the basis for expression
of aquatic life criteria, use of the total recoverable
measurement to express permit limits on effluents does not place
any restrictions on which analytical measurement can be used to
express criteria. Further, even if both criteria and permit
limits are expressed in terms of a measurement such as dissolved
metal, an empirical extrapolation would still be necessary
because dissolved metal is not likely to be conservative from
effluent to downstream water.
Merits of Total Recoverable and Dissolved WERs and Criteria
A WER is operationally defined as the value of an endpoint
obtained with a toxicity test using site water divided by the
value of the same endpoint obtained with the same toxicity test
using a laboratory dilution water. Therefore, just as aquatic
life criteria can be expressed in terms of either the total
recoverable measurement or the dissolved measurement, so can
WERs. A pair of side-by-side toxicity tests can produce both a
total recoverable WER and a dissolved WER if the metal in the
test solutions in both of the tests is measured using both
methods. A total recoverable WER is obtained by dividing
endpoints that were calculated on the basis of total recoverable
metal, whereas a dissolved WER is obtained by dividing endpoints
that were calculated on the basis of dissolved metal. Because of
the way they are determined, a total recoverable WER is used to
calculate a total recoverable site-specific criterion from a
national, state, or recalculated aquatic"life criterion that is
expressed using the total recoverable measurement, whereas a
dissolved WER is used to calculate a dissolved site-specific
criterion from a national, state, or recalculated criterion that
is expressed in terms of the dissolved measurement.
In terms of the classification scheme given in Figure Dl, the
basic relationship between a total recoverable national water
quality criterion and a total recoverable WER is:
A total recoverable criterion treats all the toxic and
nontoxic metal in the site water as if its average
toxicity were the same as the average toxicity of all
the toxic and nontoxic metal in the toxicity tests in
laboratory dilution water on which the criterion is
based.
A total recoverable WER is a measurement of the actual
ratio of the average toxicities of the total
recoverable metal and replaces the assumption that
the ratio is 1.
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Similarly, the basic relationship between a dissolved national
criterion and a dissolved WER is:
A dissolved criterion treats all the toxic and nontoxic
dissolved metal in the site water as if its average
toxicity were the same as the average toxicity of all
the toxic and nontoxic dissolved metal in the
toxicity tests in laboratory dilution water on which
the criterion is based.
A dissolved WER is a measurement of the actual ratio of
the average toxicities of the dissolved metal and
replaces the assumption that the ratio is 1.
In both cases, use of a criterion without a WER involves
measurement of toxicity in laboratory dilution water but only
prediction of toxicity in site water, whereas use of a criterion
with a WER involves measurement of toxicity in both laboratory
dilution water and site water.
When WERs are used to derive site-specific criteria, the total
recoverable and dissolved approaches are inherently consistent.
They are consistent because the toxic effects caused by the metal
in the toxicity tests do not depend on what chemical measurements
are performed; the same number of organisms are killed in the
acute lethality tests regardless of what, if any, measurements of
the concentration of the metal are made. The only difference is
the chemical measurement to which the toxicity is referenced.
Dissolved WERs can be derived from the same pairs of toxicity
tests from which total recoverable WERs are derived, if the metal
in the tests is measured using both the total recoverable and
dissolved measurements. Both approaches start at the same place
(i.e., the amount of toxicity observed in laboratory dilution
water) and end at the same place (i.e., the amount of toxicity
observed in site water). The combination of a total recoverable
criterion and WER accomplish the same thing as the combination of
a dissolved criterion and WER. By extension, whenever a
criterion and a WER based on the same measurement of the metal
are used together, they will end up at the same place. Because
use of a total recoverable criterion with a total recoverable WER
ends up at exactly the same place as use of a dissolved criterion
with a dissolved WER. whenever one WER is determined, both should
be determined to allow (a) a check on the analytical chemistry,
(b) use of the inherent internal consistency to check that the
data are used correctly, and (c) the option of using either
approach in the derivation of permit limits.
An examination of how the two approaches (the total recoverable
approach and the dissolved approach) address the four relevant
forms of metal (toxic and nontoxic particulate metal and toxic
and nontoxic dissolved metal) in laboratory dilution water and in
site water further explains why the two approaches are inherently
consistent. Here, only the way in which the two approaches
address each of the four forms of metal in site water will be
considered:
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a. Toxic dissolved metal:
This form contributes to the toxicity of the site water and
is measured by both chemical measurements. If this is the
only form of metal present, the two WERs will be the same.
b. Nontoxic dissolved metal:
This form does not contribute to the toxicity of the site
water, but it is measured by both chemical measurements.
If this is the only form of metal present, the two WERs
will be the same. (Nontoxic dissolved metal can be the
only form present, however, only if all of the nontoxic
dissolved metal present is refractory. If any labile
nontoxic dissolved metal is present, equilibrium will
require that some toxic dissolved metal also be present.)
c. Toxic particulate metal:
This form contributes to the toxicological measurement in
both approaches; it is measured by the total recoverable
measurement, but not by the dissolved measurement. Even
though it is not measured by the dissolved measurement, its
presence is accounted for in the dissolved approach because
it increases the toxicity of the site water and thereby
decreases the dissolved WER. It is accounted for because
it makes the dissolved metal appear to be more toxic than
it is. Most toxic particulate metal is probably not toxic
when it is particulate; it becomes toxic when it is
dissolved at the gill surface or in the digestive system;
in the surface water, however, it is measured as
particulate metal.
d. Nontoxic particulate metal:
This form does not contribute to the toxicity of the site
water; it is measured by the total recoverable measurement,
but not by the dissolved measurement. Because it is
measured by the total recoverable measurement, but not by
the dissolved measurement, it causes the total recoverable
WER to be higher than the dissolved WER.
In addition to dealing with the four forms of metal similarly,
the WERs used in the two approaches comparably take synergism,
antagonism, and additivity into account. Synergism and
additivity in the site water increase its toxicity and therefore
decrease the WER; in contrast, antagonism in the site water
decreases toxicity and increases the WER.
Each of the four forms of metal is appropriately taken into
account because use of the WERs makes the two approaches
internally consistent. In addition, although experimental
variation will cause the measured WERs to deviate from the actual
WERs, the measured WERs will be internally consistent with the
data from which they were generated. If the percent dissolved is
the same at the test endpoint in the two waters, the two WERs
will be the same. If the percent of the total recoverable metal
that is dissolved in laboratory dilution water is less than 100
percent, changing from the total recoverable measurement to the
dissolved measurement will lower the criterion but it will
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comparably lower the denominator in the WER, thus increasing the
WER. If the percent of the total recoverable metal that is
dissolved in the site water is less than 100 percent, changing
from the total recoverable measurement to the dissolved
measurement will lower the concentration in the site water that
is to be compared with the criterion, but it also lowers the
numerator in the WER, thus lowering the WER. Thus when WERs are
used to adjust criteria, the total recoverable approach and the
dissolved approach result in the same interpretations of
concentrations in the site water (see Figure D3) and in the same
maximum acceptable concentrations in effluents (see Figure D4).
Thus, if WERs are based on toxicity tests whose endpoints equal
the CMC or CCC and if both approaches are used correctly, the two
measurements will produce the same results because each WER is
based on measurements on the site water and then the WER is used
to calculate the site-specific criterion that applies to the site
water when the same chemical measurement is used to express the
site-specific criterion. The equivalency of the two approaches
applies if they are based on the same sample of site water. When
they are applied to multiple samples, the approaches can differ
depending on how the results from replicate samples are used:
a. If an appropriate averaging process is used, the two will be
equivalent.
b. If the lowest value is used, the two approaches will probably
be equivalent only if the lowest dissolved WER and the lowest
total recoverable WER were obtained using the same sample of
site water.
There are several advantages to using a dissolved criterion even
when a dissolved WER is not used. In some situations use of a
dissolved criterion to interpret results of measurements of the
concentration of dissolved metal in site water might demonstrate
that there is no need to determine either a total recoverable WER
or a dissolved WER. This would occur when so much of the total
recoverable metal was nontoxic particulate metal that even though
the total recoverable criterion was exceeded, the corresponding
dissolved criterion was not exceeded. The particulate metal
might come from an effluent, a resuspension event, or runoff that
washed particulates into the body of water. In such a situation
the total recoverable WER would also show that the site-specific
criterion was not exceeded, but there would be no need to
determine a WER if the criterion were expressed on the basis of
the dissolved measurement. If the variation over time in the
concentration of particulate metal is much greater than the
variation in the concentration of dissolved metal, both the total
recoverable concentration and the total recoverable WER are
likely to vary so much over time that a dissolved criterion would
be much more useful than a total recoverable criterion.
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Use of a dissolved criterion without a dissolved WER has three
disadvantages, however:
1. Nontoxic dissolved metal in the site water is treated as if it
is toxic.
2. Any toxicity due to particulate metal in the site water is
ignored.
3. Synergism, antagonism, and additivity in the site water are
not taken into account.
Use of a dissolved criterion with a dissolved WER overcomes all
three problems. For example, if (a) the total recoverable
concentration greatly exceeds the total recoverable criterion,
(b) the dissolved concentration is below the dissolved criterion,
and (c) there is concern about the possibility of toxicity of
particulate metal, the determination of a dissolved WER would
demonstrate whether toxicity due to particulate metal is
measurable.
Similarly, use of a total recoverable criterion without a total
recoverable WER has three comparable disadvantages:
1. Nontoxic dissolved metal in site water is treated as if it is
toxic.
2. Nontoxic particulate metal in site water is treated as if it
is toxic.
3. Synergism, antagonism, and additivity in site water are not
taken into account.
Use of a total recoverable criterion with a total recoverable WER
overcomes all three problems. For example, determination of a
total recoverable WER would prevent nontoxic particulate metal
(as well as nontoxic dissolved metal) in the site water from
being treated as if it is toxic.
Relationships between WERs and the Forms of Metals
Probably the best way to understand what WERs can and cannot do
is to understand the relationships between WERs and the forms of
metals. A WER is calculated by dividing the concentration of a
metal that corresponds to a toxicity endpoint in a site water by
the concentration of the same metal that corresponds to the same
toxicity endpoint in a laboratory dilution water. Therefore,
using the classification scheme given in Figure Dl:
X\- a ~^~ -tv a ~l~ J. Q ~^~ A.Zv ci ~^~ A Ta
WER = £ £ £ s s _
RL + 1?^ + TL + &NL + &TL
The subscripts "S" and "L" denote site water and laboratory
dilution water, respectively, and:
R = the concentration of Refractory metal in a water. (By
definition, all refractory metal is nontoxic metal.)
117
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N ** the concentration of Nontoxic labile metal in a water.
T = the concentration of Toxic labile metal in a water.
AW = the concentration of metal added during a WER determination
that is Nontoxic labile metal after it is added.
AT » the concentration of metal added during a WER determination
that is Toxic labile metal after it is added.
For a total recoverable WER, each of these five concentrations
includes both particulate and dissolved metal, if both are
present; for a dissolved WER only dissolved metal is included.
Because the two side-by-side tests use the same endpoint and are
conducted under identical conditions with comparable test
organisms, Ts + &TS = TL + &TL when the toxic species of the metal
are equally toxic in the two waters. If a difference in water
quality causes one or more of the toxic species of the metal to
be more toxic in one water than the other, or causes a shift in
the ratios of various toxic species, we can define
Ts + Arg
Thus H is a multiplier that accounts for a proportional increase
or decrease in the toxicity of the toxic forms in site water as
compared to their toxicities in laboratory dilution water.
Therefore, the general WER equation is:
WER =
N
H(TL
R
N
(TL
Several things are obvious from this equation:
1. A WER should not be thought of as a simple ratio such as H.
H is the ratio of the toxicities of the toxic species of the
metal, whereas the WER is the ratio of the sum of the toxic
and the nontoxic species of the metal. Only under a very-
specific set of conditions will WER = H. If these conditions
are satisfied and if, in addition, H = l, then WER = 1.
Although it might seem that all of these conditions will
rarely be satisfied, it is not all that rare to find that an
experimentally determined WER is close to 1.
2. When the concentration of metal in laboratory dilution water
is negligible, RL = NL = TL = o and
WER =
Re
N
118
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Even though laboratory dilution water is low in TOG and TSS,
when metals are added to laboratory dilution water in toxicity
tests, ions such as hydroxide, carbonate, and chloride react
with some metals to form some particulate species and some
dissolved species, both of which might be toxic or nontoxic.
The metal species that are nontoxic contribute to &NL, whereas
those that are toxic contribute to ATL . Hydroxide, carbonate,
chloride, TOG, and TSS can increase &NS. Anything that causes
&NS to differ from *NL will cause the WER to differ from 1.
3. Refractory metal and nontoxic labile metal in the site water
above that in the laboratory dilution water will increase the
WER. Therefore, if the WER is determined in downstream water,
rather than in upstream water, the WER will be increased by
refractory metal and nontoxic labile metal in the effluent.
Thus there are three major reasons why WERs might be larger or
smaller than 1:
a. The toxic species of the metal might be more toxic in one
water than in the other, i.e., H* 1.
b. AN might be higher in one water than in the other.
c. R and/or N might be higher in one water than in the other.
The last reason might have great practical importance in some
situations. When a WER is determined in downstream water, if
most of the metal in the effluent is nontoxic, the WER and the
endpoint in site water will correlate with the concentration of
metal in the site water. In addition, they will depend on the
concentration of metal in the effluent and the concentration of
effluent in the site water. This correlation will be best for
refractory metal because its toxicity cannot be affected by water
quality characteristics; even if the effluent and upstream water
are quite different so that the water quality characteristics of
the site water depend on the percent effluent, the toxicity of
the refractory metal will remain constant at zero and the portion
of the WER that is due to refractory metal will be additive.
The Dependence of WERs on the Sensitivity of Toxicitv Tests
It would be desirable if the magnitude of the WER for a site
water were independent of the toxicity test used in the
determination of the WER, so that any convenient toxicity test
could be used. It can be seen from the general WER equation that
the WER will be independent of the toxicity test only if:
H(TT + AT,)
- (2-;. *TJ -- H
which would require that Rs = Ns = ANS = RL = NL = &NL = 0 . (It would
be easy to assume that TL = 0, but it can be misleading in some
situations to make more simplifications than are necessary.)
119
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This is the simplistic concept of a WER that would be
advantageous if it were true, but which is not likely to be true
very often. Any situation in which one or more of the terms is
greater than zero can cause the WER to depend on the sensitivity
of the toxicity test, although the difference in the WERs might
be small.
Two situations that might be common can illustrate how the WER
can depend on the sensitivity of the toxicity test . For these
illustrations, there is no advantage to assuming that H = 1 , so
H will be retained for generality.
1. The simplest situation is when Rs > 0 , i.e., when a
substantial concentration of refractory metal occurs in the
site water. If, for simplification, it is assumed that
Ns = Al\Ts = RL = NL = ANL = 0 , then :
+ Ar£) _ RS +
-- H
(TL + Azy (TL + Azy
The quantity TL + &TL obviously changes as the sensitivity of
the toxicity test changes . When Rs = 0 , then WER = H and the
WER is independent of the sensitivity of the toxicity test .
When Rs > 0 , then the WER will decrease as the sensitivity of
the test decreases because TL + &TL will increase.
2. More complicated situations occur when (Ns + &NS) > o. If, for
simplification, it is assumed that Rs = RL = NL = ANL = 0 , then:
(TL + ATL) (TL + ATL)
a. If (Ns + ANS) > 0 because the site water contains a
substantial concentration of a complexing agent that has an
affinity for the metal and if complexation converts toxic
metal into nontoxic metal, the complexation reaction will
control the toxicity of the solution (Allen 1993) . A
complexation curve can be graphed in several ways, but the
S- shaped curve presented in Figure D5 is most convenient
here. The vertical axis is "% uncomplexed" , which is
assumed to correlate with "% toxic". The "% complexed" is
then the "% nontoxic". The ratio of nontoxic metal to
toxic metal is :
%nontoxic _ %complexed _
= V .
%toxic %uncomplexed
For the complexed nontoxic metal:
concentration of nontoxic metal
V =
concentration of toxic metal
120
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In the site water, the concentration of complexed nontoxic
metal is (N3 + ANS) and the concentration of toxic metal is
(T3 + ATS) , so that:
(Ng + ANS) _ (Ns + ANS)
3 (Ts + ATS) H(TL + ATL)
and
WER =
(TL + ATL/
If the WER is determined using a sensitive toxicity test so
that the % uncomplexed (i.e., the % toxic) is 10 %, then
Vs = (90 %)/(lO %) = 9 , whereas if a less sensitive test is
used so that the % uncomplexed is 50 %, then
Vs = (50 %)/(50 %) = 1. Therefore, if a portion of the WER is
due to a complexing agent in the site water, the magnitude
of the WER can decrease as the sensitivity of the toxicity
test decreases because the % uncomplexed will decrease. In
these situations, the largest WER will be obtained with the
most sensitive toxicity test; progressively smaller WERs
will be obtained with less sensitive toxicity tests. The
magnitude of a WER will depend not only on the sensitivity
of the toxicity test but also on the concentration of the
complexing agent and on its binding constant (complexation
constant, stability constant). In addition, the binding
constants of most complexing agents depend on pH.
If the laboratory dilution water contains a low
concentration of a complexing agent,
NT + ANT
V=L L
TL + ATL
and
+ ATL) + H(TL + ATL) = VgH + H = H(VS
VL(TL + ATL) + (TL + ATL) V
L
The binding constant of the complexing agent in the
laboratory dilution water is probably different from that
of the complexing agent in the site water. Although
changing from a more sensitive test to a less sensitive
test will decrease both Vs and VL, the amount of effect is
not likely to be proportional.
If the change from a more sensitive test to a less
sensitive test were to decrease VL proportionately more
than Vs, the change could result in a larger WER, rather
121
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than a smaller WER, as resulted in the case above when it
was assumed that the laboratory dilution water did not
contain any complexing agent. This is probably most likely
to occur if H = 1 and if Vs < VL, which would mean that
WER < 1. Although this is likely to be a rare situation,
it does demonstrate again the importance of determining
WERs using toxicity tests that have endpoints in laboratory
dilution water that are close to the CMC or CCC to which
the WER is to be applied.
b. If (Ns + &NS) > 0 because the site water contains a
substantial concentration of an ion that will precipitate
the metal of concern and if precipitation converts toxic
metal into nontoxic metal, the precipitation reaction will
control the toxicity of the solution. The "precipitation
curve" given in Figure D6 is analogous to the "complexation
curve" given in Figure D5; in the precipitation curve, the
vertical axis is "% dissolved", which is assumed to
correlate with "% toxic". If the endpoint for a toxicity
test is below the solubility limit of the precipitate,
(Ns + &.NS) = 0, whereas if the endpoint for a toxicity test
is above the solubility limit, (Ns + &NS) > 0 . If WERs are
determined with a series of toxicity tests that have
increasing endpoints that are above the solubility limit,
the WER will reach a maximum value and then decrease. The
magnitude of the WER will depend not only on the
sensitivity of the toxicity test but also on the
concentration of the precipitating agent, the solubility
limit, and the solubility of the precipitate.
Thus, depending on the composition of the site water, a WER
obtained with an insensitive test might be larger, smaller, or
similar to a WER obtained with a sensitive test. Because of the
range of possibilities that exist, the best toxicity test to use
in the experimental determination of a WER is one whose endpoint
in laboratory dilution water is close to the CMC or CCC that is
to be adjusted. This is the rationale that was used in the
selection of the toxicity tests that are suggested in Appendix I.
The available data indicate that a less sensitive toxicity test
usually gives a smaller WER than a more sensitive test (Hansen
1993a). Thus, use of toxicity tests whose endpoints are higher
than the CMC or CCC probably will not result in underprotection;
in contrast, use of tests whose endpoints are substantially below
the CMC or CCC might result in underprotection.
The factors that cause Rs and (Ns + &NS) to be greater than zero
are all external to the test organisms; they are chemical effects
that affect the metal in the water. The magnitude of the WER is
therefore expected to depend on the toxicity test used only in
regard to the sensitivity of the test. If the endpoints for two
122
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different tests occur at the same concentration of the metal, the
magnitude of the WERs obtained with the two tests should be the
same; they should not depend on (a) the duration of the test, (b)
whether the endpoint is based on a lethal or sublethal effect, or
(c) whether the species is a vertebrate or an invertebrate.
Another interesting consequence of the chemistry of complexation
is that the % uncomplexed will increase if the solution is
diluted (Allen and Hansen 1993). The concentration of total
metal will decrease with dilution but the % uncomplexed will
increase. The increase will not offset the decrease and so the
concentration of uncomplexed metal will decrease. Thus the
portion of a WER that is due to complexation will not be strictly
additive (see Appendix G), but the amount of nonadditivity might
be difficult to detect in.toxicity studies of additivity. A
similar effect of dilution will occur for precipitation.
The illustrations presented above were simplified to make it
easier to understand the kinds of effects that can occur. The
illustrations are qualitatively valid and demonstrate the
direction of the effects, but real-world situations will probably
be so much more complicated that the various effects cannot be
dealt with separately.
Other Properties of WERs
1. Because of the variety of factors that can affect WERs, no
rationale exists at present for extrapolating WERs from one
metal to another, from one effluent to another, or from one
surface water to another. Thus WERs should be individually
determined for each metal at each site.
2. The most important information that the determination of a WER
provides is whether simulated and/or actual downstream water
adversely affects test organisms that are sensitive to the
metal. A WER cannot indicate how much metal needs to be
removed from or how much metal can be added to an effluent.
a. If the site water already contains sufficient metal that it
is toxic to the test organisms, a WER cannot be determined
with a sensitive test and so an insensitive test will have
to be used. Even if a WER could be determined with a
sensitive test, the WER cannot indicate how much metal has
to be removed. For example, if a WER indicated that there
was 20 percent too much metal in an effluent, a 30 percent
reduction by the discharger would not reduce toxicity if
only nontoxic metal was removed. The next WER
determination would show that the effluent still contained
too much metal. Removing metal is useful only if the metal
removed is toxic metal. Reducing the total recoverable
concentration does not necessarily reduce toxicity.
123
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b. If the simulated or actual downstream water is not toxic, a
WER can be determined and used to calculate how much
additional metal the effluent could contain and still be
acceptable. Because an unlimited amount of refractory
metal can be added to the effluent without affecting the
organisms, what the WER actually determines is how much
additional toxic metal can be added to the effluent.
The effluent component of nearly all WERs is likely to be due
mostly to either (a) a reduction in toxicity of the metal by
TSS or TOG, or (b) the presence of refractory metal. For both
of these, if the percentage of effluent in the downstream
water decreases, the magnitude of the WER will usually
decrease. If the water quality characteristics of the
effluent and the upstream water are quite different, it is
possible that the interaction will not be additive; this can
affect the portion of the WER that is due to reduced toxicity
caused by sorption and/or binding, but it cannot affect the
portion of the WER that is due to refractory metal.
Test organisms are fed during some toxicity tests, but not
during others; it is not clear whether a WER determined in a
fed test will differ from a WER determined in an unfed test.
Whether there is a difference is likely to depend on the
metal, the type and amount of food, and whether a total
recoverable or dissolved WER is determined. This can be
evaluated by determining two WERs using a test in which the
organisms usually are not fed - one WER with no food added to
the tests and one with food added to the tests. Any effect of
food is probably due to an increase in TOC and/or TSS. If
food increases the concentration of nontoxic metal in both the
laboratory dilution water and the site water, the food will
probably decrease the WER. Because complexes of metals are
usually soluble, complexation is likely to lower both total
recoverable and dissolved WERs; sorption to solids will
probably reduce only total recoverable WERs. The food might
also affect the acute-chronic ratio. Any feeding during a
test should be limited to the minimum necessary.
Ranges of Actual Measured WERs
The acceptable WERs found by Brungs et al. (1992) were total
recoverable WERs that were determined in relatively clean fresh
water. These WERs ranged from about 1 to 15 for both copper and
cadmium, whereas they ranged from about 0.7 to 3 for zinc. The
few WERs that were available for chromium, lead, and nickel
ranged from about 1 to 6. Both the total recoverable and
dissolved WERs for copper in New York harbor range from about 0.4
to 4 with most of the WERs being between 1 and 2 (Hansen 1993b).
124
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Figure D2: An Example of the Empirical Extrapolation Process
Assume the following hypothetical effluent and upstream water:
Effluent:
TE:
DE:
QE-
100 ug/L
10 ug/L
24 cfs
Upstream water:
Ta: 40 ug/L
Dff: 38 ug/L
Qu: 48 Cfs
Downstream water:
60 ug/L
36 ug/L
72 cfs
TD:
(10 % dissolved)
(95 % dissolved)
(60 % dissolved)
where :
T = concentration of total recoverable metal .
D = concentration of dissolved metal.
Q = flow.
The subscripts E, U, and D signify effluent, upstream water, and
downstream water, respectively.
By conservation of flow: QD = QE + Qv .
By conservation of total recoverable metal: TDQD = TEQE + T^J2V .
If P = the percent of the total recoverable metal in the
effluent that becomes dissolved in the downstream water,
P =
For the data given above, the percent of the total recoverable
metal in the effluent that becomes dissolved in the downstream
water is :
(38 ug/L) (48 cfs)} = ~~
p = 100 [(36 ug/L) (72 cfs)
(100 ug/L) (24 cfs)
which is greater than the 10 % dissolved in the effluent and less
than the 60 % dissolved in the downstream water.
125
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Figure D3: The Internal Consistency of the Two Approaches
The internal consistency of the total recoverable and dissolved
approaches can be illustrated by considering the use of WERs to
interpret the total recoverable and dissolved concentrations of a
metal in a site water. For this hypothetical example, it will be
assumed that the national CCCs for the metal are:
200 ug/L as total recoverable metal.
160 ug/L as dissolved metal.
It will also be assumed that the concentrations of the metal in
the site water are:
300 ug/L as total recoverable metal.
120 ug/L as dissolved metal.
The total recoverable concentration in the site water exceeds the
national CCC, but the dissolved concentration does not.
The following results might be obtained if WERs are determined:
In Laboratory Dilution Water
Total recoverable LC50 = 400 ug/L.
% of the total recoverable metal that is dissolved = 80.
(This is based on the ratio of the national CCCs,
which were determined in laboratory dilution water.)
Dissolved LC50 = 320 ug/L.
In Site Water
Total recoverable LC50 = 620 ug/L.
% of the total recoverable metal that is dissolved = 40.
(This is based on the data given above for site water).
Dissolved LC50 =248 ug/L.
WERs
Total recoverable WER = (620 ug/L)/(400 ug/L) =1.55
Dissolved WER = (248 ug/L)/(320 ug/L) = 0.775
Checking the Calculations
Total recoverable WER _ 1.55 _ lab water % dissolved _ 80
Dissolved WER 0.775 site water % dissolved 40
= 2
Site-specific CCCs (ssCCCs)
Total recoverable ssCCC = (200 ug/L) (1.55) = 310 ug/L.
Dissolved ssCCC = (160 ug/L) (0.775) = 124 ug/L.
Both concentrations in site water are below the respective
ssCCCs.
126
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In contrast, the following results might have been obtained when
the WERs were determined:
In Laboratory Dilution Water
Total recoverable LC50 = 400 ug/L.
% of the total recoverable metal that is dissolved = 80.
Dissolved LC50 = 320 ug/L.
In Site Water
Total recoverable LC50 = 580 ug/L.
% of the total recoverable metal that is dissolved = 40.
Dissolved LC50 = 232 ug/L.
WERs
Total recoverable WER = (580 ug/L)/(400 ug/L) = 1.45
Dissolved WER = (232 ug/L)/(320 ug/L) = 0.725
Checking the Calculations
Total recoverable WER _ 1.45 _ lab water % dissolved _ 80
Dissolved WER 0.725 site water % dissolved 40
Site-specific CCCs (ssCCCs)
Total recoverable ssCCC = (200 ug/L) (1.45) = 290 ug/L.
Dissolved ssCCC = (160 ug/L)(0.725) = 116 ug/L.
In this case, both concentrations in site water are above the
respective ssCCCs.
In each case, both approaches resulted in the same conclusion
concerning whether the concentration in site water exceeds the
site-specific criterion.
The two key assumptions are:
1. The ratio of total recoverable metal to dissolved metal in
laboratory dilution water when the WERs are determined equals
the ratio of the national CCCs.
2. The ratio of total recoverable metal to dissolved metal in
site water when the WERs are determined equals the ratio of
the concentrations reported in the site water.
Differences in the ratios that are outside the range of
experimental variation will cause problems for the derivation of
site-specific criteria and, therefore, with the internal
consistency of the two approaches.
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Figure D4: The Application of the Two Approaches
Hypothetical upstream water and effluent will be used to
demonstrate the equivalence of the total recoverable and
dissolved approaches. The upstream water and the effluent will
be assumed to have specific properties in order to allow
calculation of the properties of the downstream water, which will
be assumed to be a 1:1 mixture of the upstream water and
effluent. It will also be assumed that the ratios of the forms
of the metal in the upstream water and in the effluent do not
change when the total recoverable concentration changes.
Upstream water (Flow = 3 cfs)
Total recoverable:
Refractory particulate:
Toxic dissolved:
400 ug/L
200 ug/L
200 ug/L
(50 % dissolved)
Effluent (Flow = 3 cfs)
Total recoverable: 440 ug/L
Refractory particulate: 396 ug/L
Labile nontoxic particulate: 44 ug/L
Toxic dissolved: 0 ug/L (0 % dissolved)
(The labile nontoxic particulate, which is 10 % of the
total recoverable in the effluent, becomes toxic
dissolved in the downstream water.)
Downstream water (Flow = 6 cfs)
Total recoverable:
Refractory particulate:
Toxic dissolved:
420 ug/L
298 ug/L
122 ug/L
(29 % dissolved)
The values for the downstream water are calculated from the
values for the upstream water and the effluent:
Total recoverable: [3(400) + 3(440)1/6 = 420 ug/L
Dissolved: [3(200) + 3(44+0)]/6 = 122 ug/L
Refractory particulate: [3(200) +3(396)]/6 =298 ug/L
Assumed National CCC (nCCC)
Total recoverable = 300 ug/L
Dissolved = 240 ug/L
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Upstream site-specific CCC (ussCCC)
Assume: Dissolved cccWER = 1.2
Dissolved ussCCC = (1.2)(240 ug/L) = 288 ug/L
By calculation: TR ussCCC = (288 ug/L)/(0.5) = 576 ug/L
Total recoverable cccWER = (576 ug/L)/(300 ug/L) = 1.92
nCCC cccWER ussCCC Cone.
Total recoverable: 300 ug/L 1.92 576 ug/L 400 ug/L
Dissolved: 240 ug/L 1.2 288 ug/L 200 ug/L
% dissolved 80 % 50 % 50 %
Neither concentration exceeds its respective ussCCC.
Total recoverable WER _ 1.92 _ lab water % dissolved _ _80^
Dissolved WER 1.2 site water % dissolved ~5Q
Downstream site-specific CCC (dssCCC)
Assume: Dissolved cccWER = 1.8
Dissolved dssCCC = (1.8)(240 ug/L) = 432 ug/L
By calculation: TR dssCCC =
{(432 ug/L-[(200 ug/L)/2])/O.l}+{(400 ug/L)/2} = 3520 ug/L
This calculation determines the amount of dissolved
metal contributed by the effluent, accounts forthe
fact that ten percent of the total recoverable metal
in the effluent becomes dissolved, and adds the total
recoverable metal contributed by the upstream flow.
Total recoverable cccWER = (3520 ug/L)/(300 ug/L) =11.73
nCCC cccWER dssCCC Cone.
Total recoverable: 300 ug/L 11.73 3520 ug/L 420 ug/L
Dissolved: 240 ug/L 1.80 432 ug/L 122 ug/L
% dissolved 80 % 12.27 % 29 %
Neither concentration exceeds its respective dssCCC.
Total recoverable WER _ 11.73 _ lab water % dissolved _ 80
Dissolved WER 1.80 site water % dissolved ~ 12.27
Calculating the Maximum Acceptable Concentration in the Effluent
Because neither the total recoverable concentration nor the
dissolved concentration in the downstream water exceeds its
respective site-specific CCC, the concentration of metal in
the effluent could be increased. Under the assumption that
the ratios of the two forms of the metal in the effluent do
not change when the total recoverable concentration changes,
the maximum acceptable concentration of total recoverable
metal in the effluent can be calculated as follows:
129
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Starting with the total recoverable dssCCC of 3520 ug/L
(6 cfs) (3520 ug/L) - (3 cfs) (400 ug/L) = 6640
3 cfs
Starting with the dissolved dssCCC of 432 ug/L
(6 cfs) (432 ug/L) - (3 cfs) (400 ug/L) (0.5) = 6640 /L
(3 cfs) (0.10)
Checking the Calculations
Total recoverable :
(3 cfs) (6640 ug/L) * (3 cfg) (400 ug/L) = 352Q /L _
6 c.f s
Dissolved:
(3 cfs) (6640 ug/L) (0.10) + (3 cfs) (400 ugr/L) (0.50) = 432 u
6
The value of 0.10 is used because this is the percent of the
total recoverable metal in the effluent that becomes dissolved
in the downstream water.
The values of 3520 ug/L and 432 ug/L equal the downstream
site-specific CCCs derived above.
Another Way to Calculate the Maximum Acceptable Concentration
The maximum acceptable concentration of total recoverable
metal in the effluent can also be calculated from the
dissolved dssCCC of 432 ug/L using a partition coefficient to
convert from the dissolved dssCCC of 432 ug/L to the total
recoverable dssCCC of 3520 ug/L:
[6 Cfsl [ ug/L _ (3 cfs} (40Q
_ °-1227 _ . - = 6640 ug/L .
3 cfs
Note that the value used for the partition coefficient in this
calculation is 0.1227 (the one that applies to the downstream
water when the total recoverable concentration of metal in the
effluent is 6640 ug/L), not 0.29 (the one that applies when
the concentration of metal in the effluent is only 420 ug/L) .
The three ways of calculating the maximum acceptable
concentration give the same result if each is used correctly.
130
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Figure D5: A Generalized Complexation Curve
The curve is for a constant concentration of the complexing
ligand and an increasing concentration of the metal.
100
Q
111
X
111
o
o
LOG OF CONCENTRATION OF METAL
131
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Figure D6: A Generalized Precipitation Curve
The curve is for a constant concentration of the precipitating
ligand and an increasing concentration of the metal.
100,-
a
HI
o
CO
eg
a
LOG OF CONCENTRATION OF METAL
132
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References
Allen, H.E. 1993. Importance of Metal Speciation to Toxicity.
Proceedings of the Water Environment Federation Workshop on
Aquatic Life Criteria for Metals. Anaheim, CA. pp. 5.5-62.
Allen, H.E., and D.J. Hansen. 1993. The Importance of Trace
Metal Speciation to Water Quality Criteria. Paper presented at
Society for Environmental Toxicology and Chemistry. Houston, TX.
November 15.
Borgmann, U. 1983. Metal Speciation and Toxicity of Free Metal
Ions to Aquatic Biota. IN: Aquatic Toxicology. (J.O. Nriagu,
ed.) Wiley, New York, NY.
Brungs, W.A., T.S. Holderman, and M.T. Southerland. 1992.
Synopsis of Water-Effect Ratios for Heavy Metals as Derived for
Site-Specific Water Quality Criteria. U.S. EPA Contract 68-CO-
0070.
Chapman, G.A., and J.K. McCrady. 1977. Copper Toxicity: A
Question of Form. In: Recent Advances in Fish Toxicology. (R.A.
Tubb, ed.) EPA-600/3-77-085 or PB-273 500. National Technical
Information Service, Springfield, VA. pp. 132-151.
Erickson, R. 1993a. Memorandum to C. Stephan. July 14.
Erickson, R. 1993b. Memorandum to C. Stephan. November 12.
French, P., and D.T.E. Hunt. 1986. The Effects of Inorganic
Complexing upon the Toxicity of Copper to Aquatic Organisms
(Principally Fish). IN: Trace Metal Speciation and Toxicity to
Aquatic Organisms - A Review. (D.T.E. Hunt, ed.) Report TR 247.
Water Research Centre, United Kingdom.
Hansen, D.J. 1993a. Memorandum to C.E. Stephan. April 29.
Hansen, D.J. 1993b. Memorandum to C.E. Stephan. October 6.
Nelson, H., D. Benoit, R. Erickson, V. Mattson, and J. Lindberg.
1986. The Effects of Variable Hardness, pH, Alkalinity,
Suspended Clay, and Humics on the Chemical Speciation and Aquatic
Toxicity of Copper. PB86-171444. National Technical Information
Service, Springfield, VA.
Wilkinson, K.J., P.M. Bertsch, C.H. Jagoe, and P.G.C. Campbell.
1993. Surface Complexation of Aluminum on Isolated Fish Gill
Cells. Environ. Sci. Technol. 27:1132-1138.
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Appendix E: U.S. EPA Aquatic Life Criteria Documents for Metals
Metal
EPA Number
NTIS Number
Aluminum
Antimony
Arsenic
Beryllium
Cadmium
Chromium
Copper
Lead
Mercury
Nickel
Selenium
Silver
Thallium
Zinc
EPA
EPA
EPA
EPA
EPA
EPA
EPA
EPA
EPA
EPA
EPA
EPA
EPA
EPA
440/5-86-008
440/5-80-020
440/5-84-033
440/5-80-024
440/5-84-032
440/5-84-029
440/5-84-031
440/5-84-027
440/5-84-026
440/5-86-004
440/5-87-006
440/5-80-071
440/5-80-074
440/5-87-003
PB88-245998
PB81-117319
PB85-227445
PB81-117350
PB85-227031
PB85-227478
PB85-227023
PB85-227437
PB85-227452
PB87-105359
PB88-142237
PB81-117822
PB81-117848
PB87-153581
All are available from:
National Technical Information Service (NTIS)
5285 Port Royal Road
Springfield, VA 22161
TEL: 703-487-4650
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Appendix F: Considerations Concerning Multiple-Metal, Multiple-
Discharge, and Special Flowing-Water Situations
Multiple-Metal Situations
Both Method 1 and Method 2 work well in multiple-metal
situations, although the amount of testing required increases as
the number of metals increases. The major problem is the same
for both methods: even when addition of two or more metals
individually is acceptable, simultaneous addition of the two or
more metals, each at its respective maximum acceptable
concentration, might be unacceptable for at least two reasons:
1. Additivity or synergism might occur between metals.
2. More than one of the metals might be detoxified by the same
complexing agent in the site water. When WERs are determined
individually, each metal can utilize all of the complexing
capacity; when the metals are added together, however, they
cannot simultaneously utilize- all of the complexing capacity.
Thus a discharger might feel that it is cost-effective to try to
justify the lowest site-specific criterion that is acceptable to
the discharger rather than trying to justify the highest site-
specific criterion that the appropriate regulatory authority
might approve.
There are two options for dealing with the possibility of
additivity and synergism between metals:
a. WERs could be developed using a mixture of the metals but it
might be necessary to use several primary toxicity tests
depending on the specific metals that are of interest. Also,
it might not be clear what ratio of the metals should be used
in the mixture.
b. If a WER is determined for each metal individually, one or
more additional toxicity tests must be conducted at the end to
show that the combination of all metals at their proposed new
site-specific criteria is acceptable. Acceptability must be
demonstrated with each toxicity test that was used as a
primary toxicity test in the determination of the WERs for the
individual metals. Thus if a different primary test was used
for each metal, the number of acceptability tests needed would
equal the number of metals. It is possible that a toxicity
test used as the primary test for one metal might be more
sensitive than the CMC (or CCC) for another metal and thus
might not be usable in the combination test unless antagonism
occurs. When a primary test cannot be used, an acceptable
alternative test must be used.
The second option is preferred because it is more definitive; it
provides data for each metal individually and for the mixture.
The first option leaves the possibility that one of the metals is
antagonistic towards another so that the toxicity of the mixture
would increase if the metal causing the antagonism were not
present.
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Multiple-Discharge Situations
Because the National Toxics Rule (NTR) incorporated WERs into the
aquatic life criteria for some metals, it might be envisioned
that more than one criterion could apply to a metal at a site if
different investigators obtained different WERs for the same
metal at the site. In jurisdictions subject to the NTR, as well
as in all other jurisdictions, EPA intends that there should be
nomore than one criterion for a pollutant at a point in a body
of water. Thus whenever a site-specific criterion is to be
derived using a WER at a site at which more than one discharger
has permit limits for the same metal, it is important that all
dischargers work together with the appropriate regulatory
authority to develop a workplan that is designed to derive a
site-specific criterion that adequately protects the entire site.
Method 2 is ideally suited for taking into account more than one
discharger.
Method 1 is straightforward if the dischargers are sufficiently
far downstream of each other that the stream can be divided into
a separate site for each discharger. Method 1 can also be fairly
straightforward if the WERs are additive, but it will be complex
if the WERs are not additive. Deciding whether to use a
simulated downstream water or an actual downstream water can be
difficult in a flowing-water multiple-discharge situation. Use
of actual downstream water can be complicated by the existence of
multiple mixing zones and plumes and by the possibility of
varying discharge schedules; these same problems exist, however,
if effluents from two or more discharges are used to prepare
simulated downstream water. Dealing with a multiple-discharge
situation is much easier if the WERs are additive, and use of
simulated downstream water is the best way to determine whether
the WERs are additive. Taking into account all effluents will
take into account synergism, antagonism, and additivity. If one
of the discharges stops or is modified substantially, however, it
will usually be necessary to determine a new WEIR, except possibly
if the metal being discharged is refractory. Situations
concerning intermittent and batch discharges ne;ed to be handled
on a case-by-case basis.
Special Flowing-Water Situations
Method 1 is intended to apply not only to ordinary rivers and
streams but also to streams that some people might consider
"special", such as streams whose design flows sire zero and
streams that some state and/or federal agencies might refer to as
"effluent-dependent", "habitat-creating", "effluent-dominated",
etc. (Due to differences between agencies, some streams whose
design flows are zero are not considered "effluent-dependent",
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etc., and some "effluent-dependent" streams have design flows
that are greater than zero.) The application of Method 1 to
these kinds of streams has the following implications:
1. If the design flow is zero, at least some WERs ought to be
determined in 100% effluent.
2. If thunderstorms, etc., occasionally dilute the effluent
substantially, at least one WER should be determined in
diluted effluent to assess whether dilution by rainwater might
result in underprotection by decreasing the WER faster than it
decreases the concentration of the metal. This might occur,
for example, if rainfall reduces hardness, alkalinity, and pH
substantially. This might not be a concern if the WER
demonstrates a substantial margin of safety.
3. If the site-specific criterion is substantially higher than
the national criterion, there should be increased concern
about the fate of the metal that has reduced or no toxicity.
Even if the WER demonstrates a substantial margin of safety
(e.g., if the site-specific criterion is three times the
national criterion, but the experimentally determined WER is
11), it might be desirable to study the fate of the metal.
4. If the stream merges with another body of water and a site-
specific criterion is desired for the merged waters, another
WER needs to be determined for the mixture of the waters.
5. Whether WET testing is required is not a WER issue, although
WET testing might be a condition for determining and/or using
a WER.
6. A concern about what species should be present and/or
protected in a stream is a beneficial-use issue, not a WER
issue, although resolution of this issue might affect what
species should be used if a WER is determined. (If the
Recalculation Procedure is used, determining what species
should be present and/or protected is obviously important.)
7. Human health and wildlife criteria and other issues might
restrict an effluent more than an aquatic life criterion.
Although there are no scientific reasons why "effluent-
dependent", etc., streams and streams whose design flows are zero
should be subject to different guidance than other streams, a
regulatory decision (for example, see 40 CFR 131) might require
or allow some or all such streams to be subject to different
guidance. For example, it might be decided on the basis of a use
attainability analysis that one or more constructed streams do
not have to comply with usual aquatic life criteria because it is
decided that the water quality in such streams does not need to
protect sensitive aquatic species. Such a decision might
eliminate any further concern for site-specific aquatic life
criteria and/or for WET testing for such streams. The water
quality might be unacceptable for other reasons, however.
In addition to its use with rivers and streams, Method 1 is also
appropriate for determining cmcWERs that are applicable to near-
field effects of discharges into large bodies of fresh or salt
water, such as an ocean or a large lake, reservoir, or estuary:
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a. The near-field effects of a pipe that extends far into a large
body of fresh or salt water that has a current, such as an
ocean, can probably best be treated the same as a single
discharge into a flowing stream. For example, if a mixing
zone is defined, the concentration of effluent at the edge of
the mixing zone might be used to define how to prepare a
simulated site water. A dye dispersion study (Kilpatrick
1992} might be useful, but a dilution model (U.S. EPA 1993) is
likely to be a more cost-effective way of obtaining
information concerning the amount of dilution at the edge of
the mixing zone.
b. The near-field effects of a single discharge that is near a
shore of a large body of fresh or salt water can also probably
best be treated the same as a single discharge into a flowing
stream, especially if there is a definite plume and a defined
mixing zone. The potential point of impact of near-field
effects will often be an embayment, bayou, or estuary that is
a nursery for fish and invertebrates and/or contains
commercially important shellfish beds. Because of their
importance, these areas should receive special consideration
in the determination and use of a WER, taking into account
sources of water and discharges, mixing patterns, and currents
(and tides in coastal areas). The current and flushing
patterns in estuaries can result in increased pollutant
concentrations in confined embayments and at the terminal up-
gradient portion of the estuary due to poor tidal flushing and
exchange. Dye dispersion studies (Kilpatrick 1992) can be
used to determine the spatial concentration of the effluent in
the receiving water, but dilution models (U.S. EPA 1993) might
not be sufficiently accurate to be useful. Dye studies of
discharges in near-shore tidal areas are especially complex.
Dye injection into the discharge should occur over at least
one, and preferably two or three, complete tidal cycles;
subsequent dispersion patterns should be monitored in the
ambient water on consecutive tidal cycles using an intensive
sampling regime over time, location, and depth. Information
concerning dispersion and the community at risk can be used to
define the appropriate mixing zone(s), which might be used to
define how to prepare simulated site water.
References
Kilpatrick, F.A. 1992. Simulation of Soluble Waste Transport
and Buildup in Surface Waters Using Tracers. Open-File Report
92-457. U.S. Geological Survey, Books and Open-File Reports, Box
25425, Federal Center, Denver, CO 80225.
U.S. EPA. 1993. Dilution Models for Effluent Discharges.
Second Edition. EPA/600/R-93/139. National Technical
Information Service, Springfield, VA.
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Appendix G: Additivity and the Two Components of a WER Determined
Using Downstream Water
The Concept of Additivitv of WERs
In theory, whenever samples of effluent and upstream water are
taken, determination of a WER in 100 % effluent would quantify
the effluent WER (eWER) and determination of a WER in 100 %
upstream water would quantify the upstream WER (uWER);
determination of WERs in known mixtures of the two samples would
demonstrate whether the eWER and the uWER are additive. For
example, if eWER = 40, uWER = 5, and the two WERs are additive, a
mixture of 20 % effluent and 80 % upstream water would give a WER
of 12, except possibly for experimental variation, because:
20(eWER) + 80(uWER) = 20(40) + 80(5) = 800 + 400 = 1200 = 12
100 100 100 100
Strict additivity of an eWER and an uWER will probably be rare
because one or both WERs will probably consist of a portion that
is additive and a portion that is not. The portions of the eWER
and uWER that are due to refractory metal will be strictly
additive, because a change in water quality will not make the
metal more or less toxic. In contrast, metal that is nontoxic
because it is complexed by a complexing agent such as EDTA will
not be strictly additive because the % uncomplexed will decrease
as the solution is diluted; the amount of change in the %
uncomplexed will usually be small and will depend on the
concentration and the binding constant of the complexing agent
(see Appendix D). Whether the nonrefractory portions of the uWER
and eWER are additive will probably also depend on the
differences between the water quality characteristics of the
effluent and the upstream water, because these will determine the
water quality characteristics of the downstream water. If, for
example, 85 % of the eWER and 30 % of the uWER are due to
refractory metal, the WER obtained in the mixture of 20 %
effluent and 80 % upstream water could range from 8 to 12. The
WER of 8 would be obtained if the only portions of the eWER and
uWER that are additive are those due to refractory metal,
because:
20 (0.85) (eWER) + 80 (0.30) (uWER) = 20(0.85) (40) +80(0.30) (5) = 8
100 . 100
The WER could be as high as 12 depending on the percentages of
the other portions of the WERs that are also additive. Even if
the eWER and uWER are not strictly additive, the concept of
additivity of WERs can be useful insofar as the eWER and uWER are
partially additive, i.e., insofar as a portion of at least one of
the WERs is additive. In the example given above, the WER
determined using downstream water that consisted of 20 % effluent
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and 80 % upstream water would be 12 if the eWER and uWER were
strictly additive; the downstream WER would be less than 12 if
the eWER and uWER were partially additive.
The Importance of Additivitv
The major advantage of additivity of WERs can be demonstrated
using the effluent and upstream water that were used above. To
simplify this illustration, the acute-chronic ratio will be
assumed to be large, and the eWER of 40 and the uWER of 5 will be
assumed to be cccWERs that will be assumed to be due to
refractory metal and will therefore be strictly additive. In
addition, the complete-mix downstream water at design-flow
conditions will be assumed to be 20 % effluent and 80 % upstream
water, so that the downstream WER will be 12 as calculated above
for strict additivity.
Because the eWER and the uWER are cccWERs and are strictly
additive, this metal will cause neither acute nor chronic
toxicity in downstream water if (a) the concentration of metal in
the effluent is less than 40 times the CCC and (b) the
concentration of metal in the upstream water is less than 5 times
the CCC. As the effluent is diluted by mixing with upstream
water, both the eWER and the concentration of metal will be
diluted simultaneously; proportional dilution of the metal and
the eWER will prevent the metal from causing acute or chronic
toxicity at any dilution. When the upstream flow equals the
design flow, the WER in the plume will decrease from 40 at the
end of the pipe to 12 at complete mix as the effluent is diluted
by upstream water; because this WER is due to refractory metal,
neither fate processes nor changes in water quality
characteristics will affect the WER. When stream flow is higher
or lower than design flow, the complete-mix WER will be lower or
higher/ respectively, than 12, but toxicity will not occur
because the concentration of metal will also be lower or higher.
If the eWER and the uWER are strictly additive and if the
national CCC is 1 mg/L, the following conclusions are valid when
the concentration of the metal in 100 % effluent is less than 40
mg/L and the concentration of the metal in 100 % upstream water
is less than 5 mg/L:
1. This metal will not cause acute or chronic toxicity in the
upstream water, in 100 % effluent, in the plume, or in
downstream water.
2. There is no need for an acute or a chronic mixing zone where a
lesser degree of protection is provided.
3. If no mixing zone exists, there is no discontinuity at the
edge of a mixing zone where the allowed concentration of metal
decreases instantaneously.
These results also apply to partial additivity as long as the
concentration of metal does not exceed that allowed by the amount
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of additivity that exists. It would be more difficult to take
into account the portions of the eWER and uWER that are not
additive.
The concept of additivity becomes unimportant when the ratios,
concentrations of the metals, or WERs are very different. For
example, if eWER = 40, uWER = 5, and they are additive, a mixture
of 1 % effluent and 99 % upstream water would have a WER of 5.35.
Given the reproducibility of toxicity tests and WERs, it would be
extremely difficult to distinguish a WER of 5 from a WER of 5.35.
In cases of extreme dilution, rather than experimentally
determining a WER, it is probably acceptable to use the limiting
WER of 5 or to calculate a WER if additivity has been
demonstrated.
Traditionally it has been believed that it is environmentally
.conservative to use a WER determined in upstream water (i.e., the
uWER) to derive a site-specific criterion that applies downstream
(i.e., that applies to areas that contain effluent). This belief
is probably based on the assumption that a larger WER would be
obtained in downstream water that contains effluent, but the
belief could also be based on the assumption that the uWER is
additive. It is possible that in some cases neither assumption
is true, which means that using a uWER to derive a downstream
site-specific criterion might result in underprotection. It
seems likely, however, that WERs determined using downstream
water will usually be at least as large as the uWER.
Several kinds of concerns about the use of WERs are actually
concerns about additivity:
1. Do WERs need to be determined at higher flows in addition to
being determined at design flow?
2. Do WERs need to be determined when two bodies of water mix?
3. Do WERs need to be determined for each additional effluent in
a multiple-discharge situation.
In each case, the best use of resources might be to test for
additivity of WERs.
Mixing Zones
In the example presented above, there would be no need for a
regulatory mixing zone with a reduced level of protection if:
1. The eWER is always 40 and the concentration of the metal in
100 % effluent is always less than 40 mg/L.
2. The uWER is always 5 and the concentration of the metal in 100
% upstream water is always less than 5 mg/L.
3. The WERs are strictly additive.
If, however, the concentration exceeded 40 mg/L in 100 %
effluent, but there is some assimilative capacity in the upstream
water, a regulatory mixing zone would be needed if the discharge
were to be allowed to utilize some or all of the assimilative
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capacity. The concept of additivity of WERs can be used to
calculate the maximum allowed concentration of the metal in the
effluent if the eWER and the uWER are strictly additive.
If the concentration of metal in the upstream water never exceeds
0.8 mg/L, the discharger might want to determine how much above
40 mg/L the concentration could be in 100 % effluent. If, for
example, the downstream water at the edge of the chronic mixing
zone under design-flow conditions consists of 70 % effluent and
30 % upstream water, the WER that would apply at the edge of the
mixing zone would be:
IQ(eWER) + 30(uWER) = 70(40) + 30(5) = 2800 + 150 = 2g 5
100 100 100
Therefore, the maximum concentration allowed at this point would
be 29.5 tng/L. If the concentration of the metal in the upstream
water was 0.8 mg/L, the maximum concentration allowed in 100 %
effluent would be 41.8 mg/L because:
70(41.8 mg/L) + 30 (0 .8 mg/L) _ 2926 mg/L + 24 mg/L _ 2g 5 mcr/L
100 100
Because the eWER is 40, if the concentration of the metal in 100
% effluent is 41.8 mg/L, there would be chronic toxicity inside
the chronic mixing zone. If the concentration in 100 % effluent
is greater than 41.8 mg/L, there would be chronic toxicity past
the edge of the chronic mixing zone. Thus even if the eWER and
the uWER are taken into account and they are assumed to be
completely additive, a mixing zone is necessary if the
assimilative capacity of the upstream water is used to allow
discharge of more metal.
If the complete-mix downstream water consists of 20 % effluent
and 80 % upstream water at design flow, the complete-mix WER
would be 12 as calculated above. The complete-mix approach to
determining and using downstream WERs would allow a maximum
concentration of 12 mg/L at the edge of the chronic mixing zone,
whereas the alternative approach resulted in a maximum allowed
concentration of 29.5 mg/L. The complete-mix approach would
allow a maximum concentration of 16.8 mg/L in the effluent
because:
70(16.8 mg/L) + 30(0.8 mg/L) = 1176 mg/L + 24 mg/L = ±2 /L
100 100
In this example, the complete-mix approach limits the
concentration of the metal in the effluent to 16.8 mg/L, even
though it is known that as long as the concentration in 100 %
effluent is less than 40 mg/L, chronic toxicity will not occur
inside or outside the mixing zone. If the WER of 12 is used to
derive a site-specific CCC of 12 mg/L that is applied to a site
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that starts at the edge of the chronic mixing zone and extends
all the way across the stream, there would be overprotection at
the edge of the chronic mixing zone (because the maximum allowed
concentration is 12 mg/L, but a concentration of 29.5 mg/L will
not cause chronic toxicity), whereas there would be
underprotection on the other side of the stream (because the
maximum allowed concentration is 12 mg/L, but concentrations
above 5 mg/L can cause chronic toxicity.)
The Experimental Determination of Additivity
Experimental variation makes it difficult to quantify additivity
without determining a large number of WERs, but the advantages of
demonstrating additivity might be sufficient to make it worth the
effort. It should be possible to decide whether the'eWER and
uWER are strictly additive based on determination of the eWER in
100 % effluent, determination of the uWER in 100 % upstream
water, and determination of WERs in 1:3, 1:1, and 3:1 mixtures of
the effluent and upstream water, i.e., determination of WERs in
100, 75, 50, 25, and 0 % effluent. Validating models of partial
additivity and/or interactions will probably require
determination of more WERs and more sophisticated data analysis
(see, for example, Broderius 1991).
In some cases chemical measurements or manipulations might help
demonstrate that at least some portion of the eWER and/or the
uWER is additive:
1. If the difference between the dissolved WER and the total
recoverable WER is explained by the difference between the
dissolved and total recoverable concentrations, the difference
is probably due to particulate refractory metal.
2. If the WERs in different samples of the effluent correlate
with the concentration of metal in the effluent, all, or
nearly all, of the metal in the effluent is probably nontoxic.
3. A WER that remains constant as the pH is lowered to 6.5 and
raised to 9.0 is probably additive.
The concentration of refractory metal is likely to be low in
upstream water except during events that increase TSS and/or TOG;
the concentration of refractory metal is more likely to be
substantial in effluents. Chemical measurements might help
identify the percentages., of the eWER and the uWER that are due to
refractory metal, but again experimental variation will limit the
usefulness of chemical measurements when concentrations are low.
Summary
The distinction between the two components of a WER determined
using downstream water has the following implications:
1. The magnitude of a WER determined using downstream water will
usually depend on the percent effluent in the sample.
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Insofar as the eWER and uWER are additive, the magnitude of a
downstream WER can be calculated from the eWER, the uWER, and
the ratio of effluent and upstream water in the downstream
water.
The derivation and implementation of site-specific criteria
should ensure that each component is applied only where it
occurs.
a. Underprotection will occur if, for example, any portion of
the eWER is applied to an area of a stream where the
effluent does not occur.
b. Overprotection will occur if, for example, an unnecessarily
small portion of the eWER is applied to an area of a stream
where the effluent occurs.
Even though the concentration of metal might be higher than a
criterion in both a regulatory mixing zone and a plume, a
reduced level of protection is allowed in a mixing zone,
whereas a reduced level of protection is not allowed in the
portion of a plume that is not inside a mixing zone.
Regulatory mixing zones are necessary if, and only if, a
discharger wants to make use of the assimilative capacity of
the upstream water.
It might be cost-effective to quantify the eWER and uWER,
determine the extent of additivity, study variability over
time, and then decide how to regulate the metal in the
effluent.
Reference
Broderius, S.J. 1991. Modeling the Joint Toxicity of
Xenobiotics to Aquatic Organisms: Basic Concepts and Approaches.
In: Aquatic Toxicology and Risk Assessment: Fourteenth Volume.
(M.A. Mayes and M.G. Barren, eds.) ASTM STP 1124. American
Society for Testing and Materials, Philadelphia, PA. pp. 107-
127.
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Appendix H: Special Considerations Concerning the Determination
of WERs with Saltwater Species
1. The test organisms should be compatible with the salinity of
the site water, and the salinity of the laboratory dilution
water should match that of the site water. Low-salinity
stenohaline organisms should not be tested in high-salinity
water, whereas high-salinity stenohaline organisms should not
be tested in low-salinity water; it is not known, however,
whether an incompatibility will affect the WER. If the
community to be protected principally consists of euryhaline
species, the primary and secondary toxicity tests should use
the euryhaline species suggested in Appendix I (or
taxonomically related species) whenever possible, although the
range of tolerance of the organisms should be checked.
a. When Method 1 is used to determine cmcWERs at saltwater
sites, the selection of test organisms is complicated by
the fact that most effluents are freshwater and they are
discharged into salt waters having a wide range of
salinities. Some state water quality standards require a
permittee to meet an LC50 or other toxicity limit at the
end of the pipe using a freshwater species. However, the
intent of the site-specific and national water quality
criteria program is to protect the communities that are at
risk. Therefore, freshwater species should not be used
when WERs are determined for saltwater sites unless such
freshwater species (or closely related species) are in the
community at risk. The addition of a small amount of brine
and the use of salt-tolerant freshwater species is
inappropriate for the same reason. The addition of a large
amount of brine and the use of saltwater species that
require high salinity should also be avoided when salinity
is likely to affect the toxicity of the metal. Salinities
that are acceptable for testing euryhaline species can be
produced by dilution of effluent with sea water and/or
addition of a commercial sea salt or a brine that is
prepared by evaporating site water; small increases in
salinity are acceptable because the effluent will be
diluted with salt water wherever the communities at risk
are exposed in the real world. Only as a last resort
should freshwater species that tolerate low levels of
salinity and are sensitive to metals, such as Daphnia magna
and Hyalella azteca, be used.
b. When Method 2 is used to determine cccWERs at saltwater
sites:
1) If the site water is low-salinity but all the sensitive
test organisms are high-salinity stenohaline organisms,
a commercial sea salt or a brine that is prepared by
evaporating site water may be added in order to increase
the salinity to the minimum level that is acceptable to
the test organisms; it should be determined whether the
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salt or brine reduces the toxicity of the metal and thereby
increases the WER.
2) If the site water is high-salinity, selecting test
organisms should not be difficult because many of the
sensitive test organisms are compatible with high-
salinity water.
2. It is especially important to consider the availability of
test organisms when saltwater species are to be used, because
many of the commonly used saltwater species are not cultured
and are only available seasonally.
3. Many standard published methodologies for tests with saltwater
species recommend filtration of dilution water, effluent,
and/or test solutions through a 37-/*m sieve or screen to
remove predators. Site water should be filtered only if
predators are observed in the sample of the water because
filtration might affect toxicity. Although recommended in
some test methodologies, ultraviolet treatment is often not
needed and generally should be avoided.
4. If a natural salt water is to be used as the laboratory
dilution water, the samples should probably be collected at
slack high tide (± 2 hours). Unless there is stratification,
samples should probably be taken at mid-depth; however, if a
water quality characteristic, such as salinity or TSS, is
important, the vertical and horizontal definition of the point
of sampling might be important. A conductivity meter,
salinometer, and/or transmissometer might be useful for
determining where and at what depth to collect the laboratory
dilution water; any measurement of turbidity will probably
correlate with TSS.
5. The salinity of the laboratory dilution water should be within
± 10 percent or 2 mg/L (whichever is higher) of that of the
site water.
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Appendix I: Suggested Toxicity Tests for Determining WERs for
Metals
Selecting primary and secondary toxicity tests for determining
WERs for metals should take into account the following:
1. WERs determined with more sensitive tests are likely to be
larger than WERs determined with less sensitive tests (see
Appendix D). Criteria are derived to protect sensitive
species and so WERs should be derived to be appropriate for
sensitive species. The appropriate regulatory authority will
probably accept WERs derived with less sensitive tests because
such WERs are likely to provide at least as much protection as
WERs determined with more sensitive tests.
2. The species used in the primary and secondary tests must be in
different orders and should include a vertebrate and an
invertebrate.
3. The test organism (i.e., species and life stage) should be
readily available throughout the testing period.
4. The chances of the test being successful should be high.
5. The relative sensitivities of test organisms vary
substantially from metal to metal.
6. The sensitivity of a species to a metal usually depends on
both the life stage and kind of test used.
7. Water quality characteristics might affect chronic toxicity
differently than they affect acute toxicity (Spehar and
Carlson 1984; Chapman, unpublished; Voyer and McGovern 1991).
8. The endpoint of the primary test in laboratory dilution water
should be as close as possible (but must not be below) the CMC
or CCC to which the WER is to be applied; the endpoint of the
secondary test should be as close as possible (and should not
be below) the CMC or CCC.
9. Designation of tests as acute and chronic has no bearing on
whether they may be used to determine a cmcWER or a cccWER.
The suggested toxicity tests should be considered, but the actual
selection should depend on the specific circumstances that apply
to a particular WER determination.
Regardless of whether test solutions are renewed when tests are
conducted for other purposes, if the concentrations of dissolved
metal and dissolved oxygen remain acceptable when determining
WERs, tests whose duration is not longer than 48 hours may be
static tests, whereas tests whose duration is longer than 48
hours must be renewal tests. If the concentration of dissolved
metal and/or the concentration of dissolved oxygen does not
remain acceptable, the test solutions must be renewed every 24
hours. If one test in a pair of side-by-side tests is a renewal
test, both of the tests must be renewed on the same schedule.
Appendix H should be read if WERs are to be determined with
saltwater species.
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Suggested Tests1 for Determining cmcWERs and cccWERs2.
(Concentrations are to be measured in all tests.)
Metal
Water3
cmcWERs4
cccWERs^
Aluminum FW
Arsenic(III)
Cadmium
Chrom(III) FW
DA
X
CDC
X
FW
SW
FW
SW
DA
BM
DA
MY
GM
CR
SL5 or FM
CR
CDC
MYC
CDC
MYC
FMC
BM
FMC
X
GM
SL or DA
FMC
CDC
Chrom(VI)
Copper
Lead
Mercury
Nickel
Selenium
Silver
Zinc
FW
SW
FW
SW
FW
SW
FW
SW
FW
SW
FW
SW
FW
SW
FW
SW
DA
MY
DA
BM
DA
BM
DA
MY
DA
MY
Y
CR
DA
BM
DA
BM
GM
NE
FM or GM
AR
GM
MYC
GM
BM
FX
BM
Y
MYC
FMC
CR
FM
MY
CDC
MYC
CDC
BMC
CDC
MYC
Y
Y
CDC
MYC
Y
MYC
CDC
MYC
CDC
MYC
GM
NEC
FM
AR
X
X
Y
Y
FMC
BMC
Y
X
FMC
BMC
FMC
BMC
1 The description of a test specifies not only the test species
and the duration of the test but also the life stage of the
species and the adverse effect(s) on which the endpoint is to
be based.
2 Some tests that are sensitive and are used in criteria
documents are not suggested here because the chances of the
test organisms being available and the test being successful
might be low. Such tests may be used if desired.
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FW = Fresh Water; SW = Salt Water.
Two-letter codes are used for acute tests, whereas codes for
chronic tests contain three letters and end in "C". One-
letter codes are used for comments.
In acute tests on cadmium with salmonids, substantial numbers
of fish usually die after 72 hours. Also, the fish are
sensitive to disturbance, and it is sometimes difficult to
determine whether a fish is dead or immobilized.
ACUTE TESTS
AR. A 48-hr EC50 based on mortality and abnormal development from
a static test with embryos and larvae of sea urchins of a
species in the genus Arbacia (ASTM 1993a) or of the species
Strongylocentrotus purpuratus (Chapman 1992).
BM. A 48-hr EC50 based on mortality and abnormal larval
development from a static test with embryos and larvae of a
species in one of four genera (Crassostrea, Mulinia, Mytilus,
Mercenaria) of bivalve molluscs (ASTM 1993b).
CR. A 48-hr EC50 (or LC50 if there is no immobilization) from a
static test with Acartia or larvae of a saltwater crustacean;
if molting does not occur within the first 48 hours, renew at
48 hours and continue the test to 96 hours (ASTM 1993a).
DA. A 48-hr EC50 (or LC50 if there is no immobilization) from a
static test with a species in one of three genera
(Ceriodaphnia, Daphnia, Simocephalus) in the family Daphnidae
(U.S. EPA 1993a; ASTM 1993a).
FM. A 48-hr LC50 from a static test at 25°C with fathead minnow
(Pimephales promelas) larvae that are 1 to 24 hours old (ASTM
1993a; U.S. EPA 1993a). The embryos must be hatched in the
laboratory dilution water, except that organisms to be used
in the site water may be hatched in the site water. The
larvae must not be fed before or during the test and at least
90 percent must survive in laboratory dilution water for at
least six days after hatch.
Note: The following 48-hr LCBOs were obtained at a
hardness of 50 mg/L with fathead minnow larvae that
were 1 to 24 hours old. The metal was measured
using the total recoverable procedure (Peltier
1993) :
Metal LC50 (ucr/L)
Cadmium 13.87
Copper 6.33
Zinc 100.95
149
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FX. A 96-hr LC50 from a renewal test (renew at 48 hours) at 25°C
with fathead minnow (Pimephales promelas) larvae that are 1
to 24 hours old (ASTM 1993a; U.S. EPA 1993a). The embryos
must be hatched in the laboratory dilution water, except that
organisms to be used in the site water may be hatched in the
site water. The larvae must not be fed before or during the
test and at least 90 percent must survive in laboratory
dilution water for at least six days after hatch.
Note: A 96-hr LC50 of 188.14 /KJ/L was obtained at a
hardness of 50 mg/L in a test on nickel with fathead
minnow larvae that were 1 to 24 hours old. The
metal was measured using the total recoverable
procedure (Peltier 1993). A 96-hr LC50 is used for
nickel because substantial mortality occurred after
48 hours in the test on nickel, but not in the tests
on cadmium, copper, and zinc.
GM. A 96-hr EC50 (or LC50 if there is no immobilization) from a
renewal test (renew at 48 hours) with a species in the genus
Gatnmarus (ASTM 1993a) .
MY. A 96-hr EC50 (or LC50 if there is no immobilization) from a
renewal test (renew at 48 hours) with a species in one of two
genera (Mysidopsis, Holmesimysis [nee Ac ant homys i s]) in the
family Mysidae (U.S. EPA 1993a; ASTM 1993a). Feeding is
required during all acute and chronic tests with mysids; for
determining WERs, mysids should be fed four hours before the
renewal at 48 hours and minimally on the non-renewal days.
HE. A 96-hr LC50 from a renewal test (renew at 48 hours) using
juvenile or adult polychaetes in the genus Nereidae (ASTM
1993a).
SL. A 96-hr EC50 (or LC50 if there is no immobilization) from a
renewal test (renew at 48 hours) with a species in one of two
genera (Qncorhvnchus. Salmo) in the family Salmonidae (ASTM
1993a).
CHRONIC TESTS
BMC. A 7-day IC25 from a survival and development renewal test
(renew every 48 hours) with a species of bivalve mollusc,
such as a species in the genus Mulinia. One such test has
been described by Burgess et al. 1992. [Note: When
determining WERs, sediment must not be in the test chamber.]
[Note: This test has not been widely used.]
CDC. A 7-day IC25 based on reduction in survival and/or
reproduction in a renewal test with a species in the genus
Ceriodaphnia in the family Daphnidae (U.S. EPA 1993b). The
150
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test solutions must be renewed every 48 hours. (A 21-day
life-cycle test with Daphnia magna is also acceptable.)
FMC. A 7-day IC25 from a survival and growth renewal test (renew
every 48 hours) with larvae (s 48-hr old) of the fathead
minnow (Pimephales promelas) (U.S. EPA 1993b). When
determining WERs, the fish must be fed four hours before
each renewal and minimally during the non-renewal days.
MYC. A 7-day IC25 based on reduction in survival, growth, and/or
reproduction in a renewal test with a species in one of two
genera (Mysidopsis. Holmesimysis [nee Acanthomysisl) in the
family Mysidae (U.S. EPA 1993c). Mysids must be fed during
all acute and chronic tests; when determining WERs, they
must be fed four hours before each renewal. The test
solutions must be renewed every 24 hours.
NEC. A 20-day IC25 from a survival and growth renewal test (renew
every 48 hours) with a species in the genus Neanthes (Johns
et al. 1991). [Note: When determining WERs, sediment must
not be in the test chamber.] [Note: This test has not been
widely used.]
COMMENTS
X. Another sensitive test cannot be identified at this time, and
so other tests used in the criteria document should be
considered.
Y. Because neither the CCCs for mercury nor the freshwater
criterion for selenium is based on laboratory data concerning
toxicity to aquatic life, they cannot be adjusted using a WER.
REFERENCES
ASTM. 1993a. Guide for Conducting Acute Toxicity Tests with
Fishes, Macroinvertebrates, and Amphibians. Standard E729.
American Society for Testing and Materials, Philadelphia, PA.
ASTM. 1993b. Guide for Conducting Static Acute Toxicity Tests
Starting with Embryos of Four Species of Saltwater Bivalve
Molluscs. Standard E724. American Society for Testing and
Materials, Philadelphia, PA.
Burgess, R., G. Morrison, and S. Rego. 1992. Standard Operating
Procedure for 7-day Static Sublethal Toxicity Tests for Mulinia
lateralis. U.S. EPA, Environmental Research Laboratory,
Narragansett, RI.
151
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Chapman, G.A. 1992. Sea Urchin (Strongvlocentrotus purpuratus)
Fertilization Test Method. U.S. EPA, Newport, OR.
Johns, D.M., R.A. Pastorok, and T.C. Ginn. 1991. A Sublethal
Sediment Toxicity Test using Juvenile Neanthes sp.
(Polychaeta:Nereidae). In: Aquatic Toxicology and Risk
Assessment: Fourteenth Volume. ASTM STP 1124. (M.A. Mayes and
M.G. Barron, eds.) American Society for Testing and Materials,
Philadelphia, PA. pp. 280-293.
Peltier, W.H. 1993. Memorandum to C.E. Stephan. October 19.
Spehar, R.L., and A.R. Carlson. 1984. Derivation of Site-
Specific Water Quality Criteria for Cadmium and the St. Louis
River Basin, Duluth, Minnesota. Environ. Toxicol. Chem. 3:651-
665.
U.S. EPA. 1993a. Methods for Measuring the Acute Toxicity of
Effluents and Receiving Waters to Freshwater and Marine
Organisms. Fourth Edition. EPA/600/4-90/027F. National
Technical Information Service, Springfield, VA.
U.S. EPA. 1993b. Short-term Methods for Estimating the Chronic
Toxicity of Effluents and Receiving Waters to Freshwater
Organisms. Third Edition. EPA/600/4-91/002. National Technical
Information Service, Springfield, VA.
U.S. EPA. 1993c. Short-term Methods for Estimating the Chronic
Toxicity of Effluents and Receiving Waters to Marine and
Estuarine Organisms. Second Edition. EPA/600/4-91/003.
National Technical Information Service, Springfield, VA.
Voyer, R.A., and D.G. McGovern. 1991. Influence of Constant and
Fluctuating Salinity on Responses of Mysidopsis bahia Exposed to
Cadmium in a Life-Cycle Test. Aquatic Toxicol. 19:215-230.
152
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Appendix J: Recommended Salts of Metals
The following salts are recommended for use when determining a
WER for the metal listed. If available, a salt that meets
American Chemical Society (ACS) specifications for reagent-grade
should be used.
Aluminum
*Aluminum chloride 6-hydrate: A1C13»6H2O
Aluminum sulfate 18-hydrate: A12 (SO4) 3«18H2O
Aluminum potassium sulfate 12-hydrate: AlK(SO4) 2«12H2O
Arsenic(III)
*Sodium arsenite: NaAsO2
Arsenic(V)
Sodium arsenate 7-hydrate, dibasic: Na2HAsO4»7H2O
Cadmium
Cadmium chloride 2.5-hydrate: CdCl2»2.5H2O
Cadmium sulfate hydrate: 3CdS04»8H20
Chromium(III)
*Chromic chloride 6-hydrate (Chromium chloride): CrCl3«6H2O
*Chromic nitrate 9-hydrate (Chromium nitrate): Cr(NO3) 3»9H2O
Chromium potassium sulfate 12-hydrate: CrK(SO4) 2«12H2O
Chromium(VI)
Potassium chromate: K2CrO4
Potassium dichromate: K2Cr2O7
*Sodium chromate 4-hydrate: Na2CrO4«4H2O
Sodium dichromate 2-hydrate: Na2Cr2O7«2H2O
Copper
*Cupric chloride 2-hydrate (Copper chloride): CuCl2»2H2O
Cupric nitrate 2.5-hydrate (Copper nitrate) : Cu (NO3) 2«2 . 5H2O
Cupric sulfate 5-hydrate (Copper sulfate): CuSO4»5H2O
Lead
*Lead chloride: PbCl2
Lead nitrate: Pb(NO3)2
Mercury
Mercuric chloride: HgCl2
Mercuric nitrate monohydrate: Hg(NO3)2«H2O
Mercuric sulfate: HgSO4
153
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Nickel
*Nickelous chloride 6-hydrate (Nickel chloride): NiCl2«6H2O
*Nickelous nitrate 6-hydrate (Nickel nitrate): Ni(NO3) 2«6H2O
Nickelous sulfate 6-hydrate (Nickel sulfate): NiSO4«6H2O
Selenium(IV)
*Sodium selenite 5-hydrate: Na2SeO3«5H2O
Selenium(VI)
*Sodium selenate 10-hydrate: Na2SeO4«10H2O
Silver
Silver nitrate: AgNO3
(Even if acidified, standards and samples containing silver
must be in amber containers.)
Zinc chloride: ZnCl2
*Zinc nitrate 6-hydrate: Zn(NO3) 2«6H2O
Zinc sulfate 7-hydrate: ZnSO4»7H2O
*Note: ACS reagent-grade specifications might not be available
for this salt.
No salt should be used until information concerning the safety
and handling of that salt has been read.
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WATER QUALITY STANDARDS
COORDINATORS
Eric Hall, WQS Coordinator
EPA Region 1
Water Division
JFK Federal Building
Boston, MA 02203
617-565-3533
Wayne Jackson, WQS Coordinator
EPA Region 2
Water Division
26 Federal Plaza
New York, NY 10278
212-264-5685
Evelyn MacKnight, WQS Coordinator
EPA Region 3
Water Division
841 Chestnut Street
Philadelphia, PA 19107
215-597-4491
Fritz Wagener, WQS Coordinator
EPA Region 4
Water Division
345 Courtland Street, N.E.
Atlanta, GA 30365
404-347-3555x6633
David Pfeifer, WQS Coordinator
EPA Region 5
Water Division
77 West Jackson Boulevard
Chicago, IL 60604-3507
312-353-9024
Cheryl Overstreet, WQS Coordinator
EPA Region 6
Water Division
1445 Ross Avenue
First Interstate Bank Tower
Dallas, TX 75202
214-655-6643
Larry Shepard, WQS Coordinator
EPA Region 7
Water Complainance Branch
726 Minnesota Avenue
Kansas City, KS 66101
913-551-7441
Bill Wuertherle, WQS Coordinator
EPA Region 8
Water Division
999 18th Street
Denver, CO 80202-2405
303-293-1586
Phil Woods, WQS Coordinator
EPA Region 9
Water Division
75 Hawthorne Street
San Francisco, CA 94105
415-744-1997
Marcia Lagerloef, WQS Coordinator
EPA Region 10
Water Division (WS-139)
1200 Sixth Avenue
Seattle, WA 98101
206-553-0176
-or-
Sally Brough, WQS Coordinator
EPA Region 10
Water Division (WS-139)
1200 Sixth Avenue
Seattle, WA 98101
206-553-1754
(8/15/94)
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TITLE
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Water Quality Standards Regulation, Part II, Environmental Protection Agency, Federal Register,
November 8, 1983
Regulations that govern the development, review, revision and approval of water quality standards
under Section 303 of the Clean Water Act.
Water Quality Standards Handbook, Second Edition, September 1993
Contains guidance issued to date in support of the Water Quality Standards Regulation.
Office of Water Policy and Technical Guidance on Interpretation and Implementation of
Aquatic Life Metals Criteria, EPA 822/F-93-009, October 1993
This memorandum transmits Office of Water policy and guidance on the interpretation and
implementation of aquatic life metals criteria. It covers aquatic life criteria, total maximum daily
loads permits, effluent monitoring, compliance and ambient monitoring.
3.
4.
5.
Water Quality Standards for the 21st Century, 1989
Summary of the proceedings from the first National Conference on water quality standards held in
Dallas, Texas, March 1-3, 1989.
Water Quality Standards for the 21st Century, 1991
Summary of the proceedings from the second National Conference on water quality standards held in
Arlington, Virginia, December 10-12, 1990.
Compilation of Water Quality Standards for Marine Waters, November 1982
Consists of marine water quality standards required by Section 304(a)(6) of the Clean Water Act. The
document identifies marine water quality standards, the specific pollutants associated with such water
quality standards and the particular waters to which such water quality standards apply. The
compilation should not in any way be construed as Agency opinion as to whether the waters listed are
marine waters within the meaning of Section 301(h) of the Clean Water Act or whether discharges to
such waters are qualified for a Section 301(h) modification.
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TITLE
6.
7.
8,
9.
10.
11.
12.
Technical Support Manual: Waterbody Surveys and Assessments for Conducting Use
Attainability Analyses, November 1983
Contains technical guidance to assist States in implementing the revised water quality standards
regulation (48 FR 51400, November 8, 1983). The guidance assists States in answering three key
questions:
a. What are the aquatic protection uses currently being achieved in the \vaterbody?
6. What are the potential uses that can be attained based on the physical, chemical and biological
characteristics of the waterbody?
c. What are the causes of any impairment of the uses?
Technical Support Manual: Waterbody Surveys and Assessments for Conducting Use
Attainability Anises, Volume II: Estuarine Systems
Contains technical guidance to assist States in implementing the revised water quality standards
regulation (48 FR 51400, November 8, 1983). This document addresses the unique characteristics of
estuarine systems and supplements the Technical Support Manual: Waterbodv Summarv and
Assessments for Conducting Use Attainability Analvses (EPA. November 1983).
Technical Support Manual: Waterbody Surveys and Assessments for Conducting Use
Attainability Analyses, Volume III: Lake Systems, November 1984
Contains technical guidance to assist States in implementing the revised water quality standards
regulation (48 FR 51400 November 8, 1983). The document addresses the unique characteristics of
lake systems and supplements two additional guidance documents: Technical Support Manual:
Waterbodv Survev and Assessments for Conducting Use Attainability Analvses EPA. (November 1983)
and Technical Sunnort Manual: Waterbodv Surveys and Assessments for Conducting Use Attainability
Analyses. Vpl 11: Estuarine Systems.
Health Effects Criteria for Marine Recreational Waters, EPA 600/1-80-031, August 1983
Tills report presents health effects quality criteria for marine recreational waters and a
recommendation for a specific criterion. The criteria were among those developed using data collected
from an extensive in-house extramural microbiological research program conducted by the U.S. EPA
over the years 1972-1979.
Health Effects Criteria for Fresh Recreational Waters, EPA 660/1-84-004, August 1984
This report presents health effects criteria for fresh recreational waters and a criterion for the quality
of the bathing water based upon swimming - associated gastrointestinal illness. The criterion was
developed from data obtained during a multi-year freshwater epidemiological-microbiological research
program conducted at bathing beaches near Erie, Pennsylvania and Tulsa, Oklahoma. Three bacterial
indications of fecal pollution were used to measure the water quality: E. Coli, enterococci and fecal
cotiforms.
Introduction to Water Quality Standards, EPA 440/5-88-089, September 1988
A primer on the water quality standards program written in question and answer format. The
publication provides general information about various elements of the water quality standards
program.
Ambient Water Quality Criteria for Bacteria - 1986 EPA 440/5-84-002
Tills document contains bacteriological water quality criteria. The recommended criteria are based on
an estimate of bacterial indicator counts and gastro-intestinal illness rates.
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13. Test Methods for Escherichia Coil and Enterococci; In Water by the Membrane Filter Procedure,
EPA 600/4-85/076, 1985
Contains methods used to measure the bacteriological densities ofE. coli and enterococci in ambient
-waters. A direct relationship between the density of enterococci and E. coli in water and the
occurrence of swimming - associated gastroenteritis has been established through epidemiological
studies of marine and fresh water bathing beaches. These studies have led to the development of
criteria which can be used to establish recreational water standards based on recognized health
effects-water quality relationships.
14. Twenty-Six Water Quality Standards Criteria Summaries, September 1988
These documents contain twenty-six summaries of State/Federal criteria. Twenty-six summaries have
been compiled which contain information extracted from State water quality standards. Titles of the
twenty-six documents are: Acidity-Alkalinity, Antidegradation, Arsenic, Bacteria, Cadmium, Chromium,
Copper, Cyanide, Definitions, Designated Uses, Dissolved Oxygen, Dissolved Solids, General
Provisions, Intermittent Streams, Iron, Lead, Mercury, Mixing Zones, Nitrogen-Ammonia/Nitrate/Nitrite,
Organics, Other Elements, Pesticides, Phosphorus, Temperature, Turbidity, and Zinc.
15. Fifty-Seven State Water Quality Standards Summaries, September 1988
Contains fifty-seven individual summaries of State water quality standards. Included in each summary
is the name of a contact person, use classifications of water bodies, mixing zones, antidegradation
policies and other pertinent information.
16. State Water Quality Standards Summaries, September 1988 (Composite document)
This document contains composite summaries of State water quality standards. The document contains
information about use classifications, antidegradation policies and other information applicable to a
States' water quality standards.
17. Transmittal of Final "Guidance for State Implementation of Water Quality Standards for CWA
Section 303(c)(2)(B)", December 12, 1988
Guidance on State adoption of criteria for priority toxic pollutants. The guidance is designed to help
States comply with the 1987 Amendments to the Clean Water Act which requires States to control
toxics in water quality standards.
18. Chronological Summary of Federal Water Quality Standards Promulgation Actions, January
1993
This document contains the date, type of action and Federal Register citation for State water quality
standards promulgated by EPA. The publication also contains information on Federally promulgated
water quality standards which have been withdrawn and replaced with State approved standards.
19. Status Report: State Compliance with CWA Section 303(c)(2)(b) as of February 4, 1990
Contains information on State efforts to comply with Section 303(c)(2)(B) of the Clean Water Act which
requires adoption of water quality standards for priority pollutants. The report identifies the States
that are compliant as of February 4, 1990, summarizes the status of State actions to adopt priority
pollutants and briefly outlines EPA's plan to federally promulgate standards for noncompliant States.
20. Water Quality Standards for Wetlands: National Guidance, July 1990
Provides guidance for meeting the priority established in the FY 1991 Asencv Operating Guidance to
develop water quality standards for wetlands during the FY 1991-1993 triennium. By the end ofFY
1993, States are required as a minimum to include wetlands in the definition of "State waters,"
establish beneficial uses for wetlands, adopt existing narrative and numeric criteria for wetlands, adopt
narrative biological criteria for wetlands and apply antidegradation policies to -wetlands.
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21. Reference Guide for Water Quality Standards for Indian Tribes, January 1990
Booklet provides an overview of the water quality standards program. Publication is designed
primarily for Indian Tribes that wish to qualify as States for the water quality standards program. The
booklet contains program requirements and a list of reference sources.
22, Developing Criteria to Protect Our Nation's Waters, EPA, September 1990 (Pamphlet)
Pamphlet which briefly describes the water quality standards program and its relationship to water
quality criteria, sediment criteria and biological criteria.
23. Water Quality Standards for the 21st Century, EPA 823-R-92-009, December 1992
Summary of the proceedings from the Third National Conference on Water Quality Standards held in
Las Vegas, Nevada, August 31-September 3, 1992
24. Biological Criteria: National Program Guidance for Surface Waters, EPA-440/5-90-004, April
1990
This document provides guidance for development and implementation of narrative biological criteria.
25. Amendments to the Water Quality Standards Regulation that Pertain to Standards on Indian
Reservations - Final Rule. Environmental Protection Agency, Federal Register, December 12,
1991
This final rule amends the water quality standards regulation by adding: 1) procedures by which an
Indian Tribe may qualify for treatment as a State for purposes of the water quality standards and 401
certification programs and 2) a mechanism to resolve unreasonable consequences that may arise when
an Indian Tribe and a State adopt different water quality standards on a common body of water.
26. Guidance on Water Quality Standards and 401 Certification Programs Administered by Indian
Tribes, December 31, 1991
Tills guidance provides procedures for determining Tribal eligibility and supplements the final rule
"Amendments to the Water Quality Standards Regulation that Pertain to Standards on Indian
Reservations".
27. Water Quality Standards; Establishment of Numeric Criteria for Priority Toxic Pollutants;
State's Compliance - Final Rule, Environmental Protection Agency, Federal Register, December
22, 1992
Tills regulation promulgates for 14 States, the chemical specific, numeric criteria for priority toxic
pollutants necessary to bring all States into compliance with the requirements of Section 303(c)(2)(B)
of the Clean Water Act. Staates determined by EPA to fully comply with Section 303(c)(2)(B)
requirements are not affected by this rule.
28. Interim Guidance on Determinations and Use of Water-Effect Ratios for Metals, EPA 823-B-94-
001, February 1994
This guidance contains specific information on procedures for developing water-effect ratios.
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WATERSHED MODELING SECTION
TITLE
1.
2.
3.
4.
Guidance for Water Quality-based Decisions: The TMDL Process, EPA 440/4-91-001, April 1991
This document defines and clarifies the requirements under Section 303 (d) of the Clean Water Act. Its
purpose is to help State water quality program managers understand the application of total maximum
daily loads (TMDLs) through an integrated, basin-wide approach to controlling point and nonpoint
source pollution. The document describes the steps that are involved in identifying and prioritizing
impaired waters and developing and implementing TMDLs for waters listed under Section 303(d).
Contact: Don Brady (202) 260-5368
Technical Guidance Manual for Performing Waste Load Allocations - Book II Streams and
Rivers - Chapter 1 Biochemical Oxygen Demand/Dissolved Oxygen, EPA 440/4-84-020, September
1983
This chapter presents the underlying technical basis for performing WLA and analysis of BOD/DO
impacts. Mathematical models to calculate water quality impacts are discussed, along with data needs
and data quality.
Contact: Bryan Goodwin (202) 260-1308
Technical Guidance Manual for Performing Waste Load Allocations - Book II Streams and
Rivers - Chapter 2 Nutrient/Eutrophication Impacts, EPA 440/4-84-021, November 1983
This chapter emphasizes the effect of photosynthetic activity stimulated by nutrient discharges on the
DO of a stream or river. It is principally directed at calculating DO concentrations using simplified
estimating techniques.
Contact: Bryan Goodwin (202) 260-1308
Technical Guidance Manual for Performing Waste Load Allocations - Book H Streams and
Rivers - Chapter 3 Toxic Substances, EPA 440/4-84-022, June 1984
This chapter describes mathematical models for predicting toxicant concentrations in rivers. It covers
a range of complexities, from dilution calculations to complex, multi-dimensional, time-varying
computer models. The guidance includes discussion of background information and assumptions for
specifying values.
Contact: Bryan Goodwin (202) 260-1308
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Technical Guidance Manual for Performing Waste Load Allocations - Simplified Analytical
Method for Determining NPDES Effluent Limitations for POTWs Discharging into Low-Flow
Streams
Mis document describes methods primarily intended for "desk top" WLA investigations or screening
studies that use available data for streamflow, effluent flow, and water quality. It is intended for
circumstances where resources for analysis and data acquisition are relatively limited.
Contact: King Boynton (202) 260-7013
6. Technical Guidance Manual for Performing Waste Load Allocations - Book IV Lakes and
Impoundments - Chapter 2 Nutrient/Eutrophication Impacts, EPA 440/4-84-019, August 1983
This chapter discusses lake eutrophication processes and some factors that influence the performance
of WLA analysis and the interpretation of results. Three classes of models are discussed, along with
the application of models and interpretation of resulting calculations. Finally, the document provides
guidance on monitoring programs and simple statistical procedures.
Contact: Bryan Goodwin (202) 260-1308
7. Technical Guidance Manual for Performing Waste Load Allocations - Book IV Lakes, Reservoirs
and Impoundments - Chapter 3 Toxic Substances Impact, EPA 440/4-87-002, December 1986
Tills chapter revi&vs the basic principles of chemical water quality modeling frameworks. The
guidance includes discussion of assumptions and limitations of such modeling frameworks, as well as
the type of information required for model application. Different levels of model complexity are
Illustrated in step-by-step examples.
Contact: Bryan Goodwin (202) 260-1308
8. Technical Guidance Manual for Performing Waste Load Allocations - Book VI Design Conditions
- Chapter I Stream Design Flow for Steady-State Modeling, EPA 440/4-87-004, September 1986
Many state water quality standards (WQS) specify specific design flows. Where such design flows are
not specified in WQS, this document provides a method to assist in establishing a maximum design flow
for the final chronic value (FCV) of any pollutant.
Contact: Bryan Goodwin (202) 260-1308
9. Final Technical Guidance on Supplementary Stream Design Conditions for Steady State
Modeling, December 1988
WQS for many pollutants are written as a function of ambient environmental conditions, such as
temperature, pH or hardness. This document provides guidance on selecting values for these
parameters when performing steady-state WLAs.
Contact: Bryan Goodwin (202) 260-1308
10. Technical Guidance Manual for Performing Waste Load Allocations - Book VH: Permit
Averaging, EPA 440/4-84-023, July 1984
Tills document provides an innovative approach to determining which types of permit limits (daily
maximum, weekly, or monthly averages) should be specified for the steady-state model output, based on
the frequency of acute criteria violations.
Contact: Bryan Goodwin (202) 260-1308 ^
II. Water Quality Assessment: A Screening Procedure for Toxic and Conventional Pollutants in
Surface and Ground Water - Part I - EPA 600/6-8S-022a, September 1985
Tills document provides a range of analyses to be used for water quality assessment. Chapters include
consideration of aquatic fate of toxic organic substances, waste loading calculations, rivers and
streams, impoundments, estuaries, and groundwater.
Contact: Bryan Goodwin (202) 260-1308
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12. Water Quality Assessment: A Screening Procedure for Toxic and Conventional Pollutants in
Surface and Ground Water - Part II - EPA 600/6-85-022b, September 1985
This document provides a range of analyses to be used for water quality assessment. Chapters include
consideration of aquatic fate of toxic organic substances, waste loading calculations, rivers and
streams, impoundments, estuaries, and ground water.
Contact: Bryan Goodwin (202) 260-1308
13. Handbook - Stream Sampling for Waste Load Allocation Applications, EPA 625/6-86/013,
September 1986
This handbook provides guidance in designing stream surveys to support modeling applications for
waste load allocations. It describes the data collection process for model support, and it shows how
models can be used to help design stream surveys. In general, the handbook is intended to educate
field personnel on the relationship between sampling and modeling requirements.
Contact: Bryan Goodwin (202) 260-1308
14. EPA's Review and Approval Procedure for State Submitted TMDLs/WLAs, March 1986
The step-by-step procedure outlined in this guidance addresses the administrative (i.e., non-technical)
aspects of developing TMDLs/WLAs and submitting them to EPA for review and approval. It includes
questions and answers to focus on key issues, pertinent sections of WQM regulations and the CWA,
and examples of correspondence.
Contact: Bryan Goodwin (202) 260-1308
15. Guidance for State Water Monitoring and Wasteload Allocation Programs, EPA 440/4-85-031,
October 1985
This guidance is for use by States and EPA Regions in developing annual section 106 and 2050)
programs. The first part of the document outlines the objectives of the water monitoring program to
conduct assessments and make necessary control decisions. The second part describes the process of
identifying and calculating total maximum daily loads and waste load allocations for point and
nonpoint sources of pollution.
Contact: King Boynton (202) 260-7013
16. Technical Guidance Manual for Performing Waste Load Allocations Book III Estuaries - Part 1 -
Estuaries and Waste Load Allocation Models, EPA 823-R-92-002, May 1990
This document provides technical information and policy guidance for preparing estuarine WLA. It
summarizes the important water quality problems, estuarine characteristics, and the simulation models
available for addressing these problems.
Contact: Bryan Goodwin (202) 260-1308
17. Technical Guidance Manual for Performing Waste Load Allocations Book HI Estuaries - Part 2
Application of Estuarine Waste Load Allocation Models, EPA 823-R-92-003, May 1990
This document provides a guide to monitoring and model calibration and testing, and a case study
tutorial on simulation of WLA problems in simplified estuarine systems.
Contact: Bryan Goodwin (202) 260-1308
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TITLE
18. Technical Guidance Manual for Performing Wasteload Allocations-Book III: Estuaries - Part 3 -
Use of Mixing Zone Models in Estuarine Wasteload Allocations, EPA 823-R-92-004
This technical guidance manual describes the initial mixing wastewater in estuarine and coastal
environments and mixing zone requirements. The important physical processess that govern the
hydrodynamic mixing of aqueous discharges are described, followed by application of available EPA
supported mixing zone models to four case study situations.
Contact: Bryan Goodwin (202) 260-1308
19. Technical Guidance Manual for Performing Wasteload Allocations - Book III - Estuaries - Part 4
- Critical Review of Coastal Embayment and Estuarine Wasteload Allocation Modeling, EPA 823-
R-92-005, August 1992
Tills document summarizes several historical case studies of model use in one freshwater coastal
embayment and a number of estuarine discharge situations.
Contact: Bryan Goodwin (202) 260-1308
20. Technical Support Document for Water Quality-based Toxics Control, EPA 505/2-90-001,
March, 1991
Ttils document discusses assessment approaches, water quality standards, derivation of ambient
criteria, effluent characterization, human health hazard assessment, exposure assessment, permit
requirements, and compliance monitoring. An example is used to illustrate the recommended
procedures.
Contact: King Boynton (202) 260-7013
21. Rates, Constants, and Kinetics Formulations in Surface Water Quality Modeling (Second
Edition), U.S. EPA 600/3-85/040, June 1985
This manual serves as a reference on modeling formulations, constants and rates commonly used in
surface water quality simulations. This manual also provides a range of coefficient values that can be
used to perform sensitivity analyses.
Contact: Bryan Goodwin (202) 260-1308
22. Dynamic Toxics Waste Load Allocation Model (DYNTOX), User's Manual, September 13, 1985
A user's manual which explains how to use the DYNTOX model. It is designed for use in wasteload
allocation of toxic substances.
Contact: Bryan Goodwin (202) 260-1308
23. Windows Front-End to SWMM (Storm Water Management Model), EPA 823-C-94-001, February
1994
A user interface (front-end) to the Storm Water Management Model (SWMM) and supporting
documentation is amiable on diskette. Operating in the Microsoft Windows Environment, this interface
simplifies data entry and model set-up.
Contact: Jerry LaVeck (202) 260-7771
24. Windows Front-End to SWRRBWQ (Simulator for Water Resources in Rural Basins-Water
Quality), EPA 823-C-94-002, February 1994
A user interface (front-end) to the Simulator for Water Resource in Rural Basins-Water Quality model
and supporting documentation is available on diskette. Operating in the Microsoft Windows
em'lronmenl, this interface simplifies data entry and model set-up.
Contact: Jerry LaVeck (202) 260-7771
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ENVIRONMENTAL ASSESSMENT SECTION
TITLE
25. De Minimis Discharges Study: Report to Congress, U.S. EPA 440/4-91-002, November 1991
This report to Congress addresses the requirements of Section 516 by identifying potential de minimis
discharges and recommends effective and appropriate methods of regulating those discharges.
Contact: Rich Healy (202) 260-7812
26. National Study of Chemical Residues in Fish. Volume I, U.S. EPA 823-R-92-008 a, September
1992
This report contains results of a screening study of chemical residues in fish taken from polluted
waters.
Contact: Richard Healy (202) 260-7812
27. National Study of Chemical Residues in Fish. Volume II. U.S. EPA 823-R-92-008 b, September
1992
This report contains results of a screening study of chemical residues in fish taken from polluted
waters.
Contact: Richard Healy (202) 260-7812
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SEDIMENT CONTAMINATION SECTION
TITLE
1. Sediment Classification Methods Compendium, U.S. EPA, EPA 823-R-92-006, September 1992
This compendium is an "encyclopedia" of methods that are used to assess chemically contaminated
sediments. It contains a description of each method, associated advantages and limitations and
existing applications.
Contact: Beverly Baker (202) 260-7037
2. Managing Contaminated Sediments: EPA Decision-Making Processes, Sediment Oversight
Technical Committee, U.S. EPA Report - 506/6-90/002, December, 1990
This document identifies EPA 's current decision-making process (across relevant statutes and
programs) for assessing and managing contaminated sediments. Management activities relating to
contaminated sediments are divided into the following six categories: finding contaminated sediments,
assessment of contaminated sediments, prevention and source controls, remediation, treatment of
removed sediments, and disposal of removed sediments.
Contact: Mike Kravitz (202) 260-7049
3. Contaminated Sediments: Relevant Statutes and EPA Program Activities, Sediment Oversight
Technical Committee, U.S. EPA Report - 506/6-90/003, December, 1990
This document provides information on program office activities relating to contaminated sediment
issues, and the specific statutes under which these activities fall. A table containing major laws or
agreements relevant to sediment quality issues is included.
Contact: Mike Kravitz (202) 260-7049
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9. Special Interest Group (SIG) Forum for Fish Consumption, User's Manual, V.I.O., U.S. EPA
822/8-91/001, February 1992
This user's manual describes various features of the Special Interest Group (SIG) Forum for fish
consumption advisotries, bans and risk management. The manual explains how to access the SIG and
use its data bases, messags, bulletins and other computer files.
Contact: Jeff Bigler (202) 260-1305
10. Consumption Surveys for Fish and Shellfish, A Review and Analysis of Survey Methods, U.S.
EPA-822/R-92-001, February 1992.
This document contains a critical analysis of methods used to determine fish consumption rates of
recreational and subsistence fisherment, groups that have the greates potential for exposure to
contaminants in fish tissues.
Contact: Jeff Bigler (202) 260-1305
11. Proceedings of the U.S. Environmental Protection Agency's National Technical Workshop "PCBs
in Fish Tissue", U.S. EPA/823-R-93-003, September 1993
This documents summarizes the proceedings of the EPA sponsored workshop held on May 10-11, 1993
in Washington, DC.
Contact: Rick Hoffman (202) 260-0642
12. Guidance for Assessing Chemical Contaminant Data for Use in Risk Advisories, Volume 1: Fish
Sampling and Analysis, EPA 823-R-93-002, August 1993
This document provides detailed technical guidance on methods for sampling and analyzing chemical
contaminants in fish and shellfish tissues. It addresses monitoring strategies, selection offish species
and chemical analytes, field and laboratory procedures and data analyses.
Contact: Jeff Bigler (202) 260-1305
13. National Fish Tissue Data Repository User Manual, Version 1.0, EPA 823-B-903-003, November
1993
The U.S. EPA has developed the National Fish Tissue Data Repository (NFTDR) for collection and
storage offish and shellfish contaminants data. The data repository is part of a large EPA data base
system called the Ocean Data Evaluation System (ODES). This manual explains how to access
information from the ODES database.
Contact: Rick Hoffman (202) 260-0642
14. National Fish Tissue Data Repository: Data Entry Guide, Version 1.0, EPA 823-B-93-006,
November 1993
The U.S. EPA has developed the National Fish Tissue Data Repository (NFTDR) for collection and
storage offish and shellfish contaminants data. The data repository is part of a larger EPA data base
system known as the Ocean Data Evaluation System (ODES). This manual assists State and Federal
Agencies in submitting data to the NFTDR
Contact: Rick Hoffman (202) 260-0642
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U.S. EPA
STANDARDS AND APPLIED SCIENCE DIVISION
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APPENDIX X
Summary of Updates
WATER QUALITY STANDARDS HANDBOOK
SECOND EDITION
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