UNITED STATES ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
FEB 221994,
OFFICE OF
WATER
EPA-823-B-94-001
MEMORANDUM
SUBJECT:
FROM:
.TO:
Use of the Mater-Effect Ratio
Standards
Tudor T. Davies, Director
Office of Science and Technology-
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 cf 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.
Printed on Recycled Paper
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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 13l.ll(b)(l) provides states with
the opportunity to adopt water quality criteria 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, 1993..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-
' affect 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. ,
Status of the Water-effect Ratio fWER) in non-National Toxics
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) (ill)- 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 f],ow
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)
BSD Directors, Regions I - VIII, X
BSD Branch, Region IX
Water Quality standards Coordinators, Regions I - X
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CO
CD
CD
DC
LU
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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 Ci
• 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 an 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. .
S ' , . , .- ' .
- 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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United States
Environmental Protection
Agency
Office of Water
Office of Science & Technology
(Mail Code 4305)
February 1994
EPA-823-B-94-001
vvEPA
Interim Guidance
on Determination and Use
of Water-Effect Ratios
for Metals
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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 bejen 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.
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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).
111
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UNITED STATES ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
FEB 2 2 1994
OFFICE OF
WATER
• 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
bv 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 tjiat
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 theses 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
Printed on Recycled Pape
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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 appropropriate 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
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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 regulator/ 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 Of f ice 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
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CONTENTS •
, • Page
Notices . . . . . . . . . . . .../.,. .' . . . . . . ." . . . ii
Foreword . . . . . • • • v •• • • • 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 .......... ^ i ....... 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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~ APPENDICES
Page
A. Comparison of WERs Determined Using. Upstream and
Downstream Water . . . • . . .-.".. • • •. • » v. 79
" B. The Recalculation Procedure 90
C. Guidance Concerning the Use of "Clean Techniques" and
QA/QC when Measuring Trace Metals 98
D. Relationships. between WERs and the Chemistry and
Toxicology of Metals ....... ;....... 109
'* E. U.S. EPA Aquatic Life Criteria Documents for Metals . . ,-. 134
F. Considerations Concerning Multiple-Metal, Multiple-
Discharge, and Special Flowing-Water Situations ..... 135
G. Additivity and the Two Components of a WER Determined
Using Downstream Water . ... . . . ....... 139
H. Special .Considerations Concerning the Determination
of WERs with Saltwater Species 145
I. Suggested Toxicity Tests for Determining WERs
for Metals .. . 147
•
J. Recommended Salts of Metals ......... 153
V111
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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
I ' . - , ' " -, - • . '
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, Du'luth, 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
Gourdine, 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.
. " - .1
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,
XI
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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; VpRs 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 arid 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.
xii •• • •' • '
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ABBREVIATIONS-
ACR: Acute-Chronic.Ratio : • . .
CGC: Criterion Continuous Concentration
i
. 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
TOC: 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
. xiii
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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
1 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-nm or a 0.40-nm 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.
xiv
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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.
xv
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PREFACE
Several issues need cpnsideration 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 .(AS1M 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.
xvi .
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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 •judgment1' 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.
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.
xvixi
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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 sthe 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) the4site-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,19B4)
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; although 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 mav 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 adHust
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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BeforeDetermining 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 (TOO
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 tes.ts 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
i
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
"effluent-dependent", "habitat-creating", 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-dispharge
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 concernincf 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 tcrtal 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.
i
8
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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 TOC should be within 10
% or 5 mg/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 reevaluated 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, TOC, 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, TOC, and/or TS'S 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 TOC and/or TSS, because the WER-
will be larger and any decrease in the discharge of such TOC
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.
. • ' 11 • •• • ' - • '
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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-base.d 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, WLAs, 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
' • • • • ' ' .• • 13
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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
14
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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 tptal 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
endppints 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
I Total Recoverable Criterion
Recalculation
Procedure
Tofcal
Recoverable
^
and/or cccWER
\/
Total Recoverable
Site-specific Criterion
Total Recoverable Permit Limit
Dissolved Criterion » (TR Criterion) (% dinaolved in toricfty tests)!
Recalculation
Procedure
_v
_v
Dissolved
cmcWER
and/or cccWER
Dissolved Site-
Criterion
\/
Net % contribution from the total recoverable metal in the efiluent
to the dissolved metal in the downstream water. (This will probably
change if the total recoverable concentration in the effluent
\/
I 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 consideration's.
17 •
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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 6). 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 time 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 gucintitative 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 6); 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 TOC 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 TOC
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 oust 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 oh 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 tq.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.
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 chroniq
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.l, 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 arid 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 allowcible 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.
26
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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
7 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,
if 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 tes.ts. . 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 6f 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 • '.•••'.'
-------
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 : •
[(CCO (WER) (GFLCW + UFLOW)] - [ (uCONC)' (uFLOtf) ]
- -- -
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:
hWER = (gCME) < OFLOtfdf) •*• ( uCONCdf ) ( uFLOWdf)
(CCC) (eFLOWdf
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
environment ally 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 uWERK 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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HER «= (% effluent) (&WER) + (% upstream water) (uffSR)
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 cfs.
f. The design flow of the upstream water (uFLOWdf) is 40
cfs.
Therefore : .
[(2 ug/L) (ttER) (10 CfB •«• UFLO»).3 - [ ( uOQMQ ( uFLOM 1
10 ug/L ' *
(10 cfs) * (uCONCdf) UP c£g)
(2 UI3/.L) (10 cfs •*• 40 CfB) *
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 uFLOW At. Complete Mix HCME hWER
(cfs) (cfs) % Eff . % UPS. WER (ua/L) _
10 40 20.0 80.0 12.000 118.4 12.00
10
10
10
10.
10
10
63
90
190
490
990
' 1990
13.7
10.0
5.0
2.0
1.0
0.5
86.3
90.0
95.0
98.0
99.0
99.5
9.795
8.500
6.750
5.700
5.350
5.175
140.5
166.4
262.4
550.4
1030.4
1990.4
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
t'he 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 uFLOW At Complete Mix HCME hWER
(cfs) • (cfs) % Eff. % UPS. WER (ua/L)
10 40 20.0. 80.0 8.800 80.00 8.800
10 . 63 13.7 86.3 6.343 80.00 8.800
10 90 10.0 90.0 4.900 ;80.00 8.800
10 190 5.0 95.0 2.950 80.00 8.800
10 .490 2.0 98.0 1.780 ; 80.00 8.800
10 990 1.0 99.0 1.390" 80.00 8.800
10 1990 0.5 99.5 1.195 - 80.00 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 uFLOW At Complete Mix HCME hWER
(cfs) (cfs) % Eff. % UPS. WER . (uo/L)
10 40 20.0 80.0 9.600 92.0 9.60
10 63 13.7 86.3 7.206 98.9 10.29
10 90 10.0 90.0 5.800 .107.0 11.10.
10 ^190 5.0 95.0 3.900 137.0 14.10
10 490 2.0 .98.0 2.760 227.0 23.10
10 990 1.0 99.0 2.380 377.0 38.10
10 .1990 0.5 99.5 2.190 677.0 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
uFLOW 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 6).
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 addltivity 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 bpth 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 6f 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, theTintended 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
1 ' - • " 37 -. " .'.'.'"'.
-------
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 pf 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)V(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)]/[(D1HD2)]}.
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. Ttien 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 with 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 samples 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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f • ' " ' >»,
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
ho*w 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 - ' •' ..-•'.
-------
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 ma/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 mg/L the EC50 for C_. reticulata equalled the CMC,
which is"a very reasonable simplifying assumption. .If,
however, the VJER 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 th€s 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
fmcr/L) (ua/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 application 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:
r
HCME * (17.73 Uff/L) (15) (9 •«• 73 cfs) - (1 ug/L) (73 cfs) m 2415 ^
9 Cfs
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: .
42
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HCME = (72.66 Uff/L) (3.66) (9 +-73 Cfs) - (1 Uff/L) (73 cfs)
Either WER can be used in the calculation of the HCME as
long -as the CMC and the WER correspond to the same
hardness and therefore to each other, because:
(17.73 Uff/L) (15) = (72.66 Uff/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:
hWER - (2415 ug/L) (9 Cfs) * (1 Uff/L) (20 Cfs) - 01 =A
: (9.2 Uff/L) (9 Cfs + 20 Cfs) " 81-54.
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? - ' ' . ,
e. 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, dt
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 WISH. 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
44
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effluent and surface water samples and conducting toxicity
tests (U.S. EPA 1993a; ASTM 1993a) ., Information should
also be read concerning safety and handling of the
metallic salt .that will be used in t^e preparation, of-the
stock solution.
9. The proposed work should be "discussed 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.
1.0. Plans should be made to perform one or more range finding
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, ail of vwhich 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 GCC
to which the WER is to be applied.
•.'..••. 45
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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 oust 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
arid/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.
46 -
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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 fpr
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.
47 " .'••'
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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 analyse's if em unusual
WER is obtained. Samples must be stored at 0 to 4°C in
the dark with no air space in the sample container.
48
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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, TOC, 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 (IDS) 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-jim
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-pn 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 TOG and TSS must be less than 5
mg/L.
3. The hardness pf the laboratory dilution water should be
between 50 arid 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.
i
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.
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 teat
should be used. •
Recommendations concerning temperature, loading, feeding,
dissolved oxygen, aeration, disturbance, and controls
given by the U.S. EPA (1993a,b,c) and/6r 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.
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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8
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 eridpoint 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.
*
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 e,asy to use
more treatments than would be used in flow-through tests.
52
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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.
53 .''.'• ' '
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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 befpre 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 TOC 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 solid.s
(TDS) is recommended in fresh water.
3. Dissolved oxygen, pH, and temperature mast 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
'.55 •• ' •' '
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alternate methods is to be requested from the
appropriate regulatory authority.
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).
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.
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 (il.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.
. ' ' • 57
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if--,
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.
3. 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, and/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. "
4. 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, prbbit analysis 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.
58 . " ' .
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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;
Stephari 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 bo.th 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.S.b and F.5).
a. If, after taking into account any known effect of
hardness on tpxicity, 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
59
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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 jthe 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, TOC, 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,
60
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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. -'
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.
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 dn 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/qr TOC and was the
concentration of TSS and/or TQC 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?
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, syhergism, or
antagonism occurred in the site water. This might be
especially important if the WER obtained with the .
'.-'••: '• 61 • ... '.. „• ' ' •
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8.
secondary, test did not confirm the WER obtained with
the primary test or if the WER was very large or small.
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.
,- i,
J. Reporting the Results-
A report of the experimental determinatior* 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
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
62
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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 tpxicoldgical
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 arid 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 EWER; all WERs
• determined with secondary tests must be reported along
with all WERs that were determined with the primary test,
.••..•' ' • " . ." :•' 63.' . • '- . . . . - • "
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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 TOC.
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 site's,". 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 l.:
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 "'.''•• v / . '
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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 samp 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 sovmuch
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 jtestirig 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 cart 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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nohtoxic 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.
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
infonnatipn 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 durin'g
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 •"..•'• :
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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 pf 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
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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 tp
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 l>e: 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 .•""-•- ' -- ; ' .
-------
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.
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.
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.
Samples of site water and/or effluent should be collected,
handled, and transported so that the tests can begin as soon
as is feasible.
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 pf 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: .
(x - x )2
n - l ' , -
Calculate SE• = the standard error of the arithmetic mean:
SE - s/Jn . __.'.•
Calculate A* x - (fc0.7) (SE) , where t0-7 is the value of Student's
t statistic for a one-sided probability of 0.7D with n - l
degrees of freedom. The values of, t0>7 for some common
degrees of freedom (df) are:. ^
'-•". v ••'•" 1 0.727
2 0.617 .
, 30.584 '
4 0.569
. 5 ' ..." 0.559
6 0.553 •
7 .:. 0.549
•;.\.-;.' .8 _- 0.546 ,
: - \ 9 ."•.'. 0.543 '
10 0.542
.11 0.540
J ' 12 Q-.539 . - ,
" - ' - - .-"•••' • \ •
. The values of t0>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.
eFLOW
(cfs)
10
10
10
10
10 ,
10
10
10
10
10 .
10
1.0
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
5.2a
6.0C
5.8°
5.7C
7.0C
10.5e
12. Oe
11. Oe
7.5C
3.5C
6.9°
6.1C
HCME
(ua/L)
826.4
341.5
341.6
475.8
177.2
196.1
,118.4
119.2'
234.0
79.6
251.4
295.2
hWER
82.80
34.31
34.32
47.74
17.88
19.77
12.00
12.08
23.56
8.12
25.30
29.68
FWER
1.0b
1.0b
1..0b
5.7d
5.7d
6.80f
10.699
10.88g
10.889
8.12h
8.12h
8.12h
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. v
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.
72
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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;tne 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 LCSOs
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 acknpwledge 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 •£ 50
and 50 £ P2 £'100,. then: -
p - 50 ~
P2 - PI
C = Log Cl + P(Log C2 - Log Cl)
LC50 = 10C
If PI = 0 and P2 = 100, LC50
If PI = P2 = 50, LC50 = t/(£D (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 I 2-
a. 0.65 < C1/C2 < 0.99. .
b. 0 £ PI < 37. •
c. 63 < P2 £ 100.
74
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Figure 6: Calculating a Time-Weighted Average
' • - ' " J~ "' '
; , :• •-',';,. 1 .'•• . .*"•:. ' ,
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)1 (hr) (ma/L)' (ma/L)
0 12 48
96 14 48
96 1248 " 1248/96.= 13.00
0 8 12
24 6 24
48 7 24
72 9 24
96 8 12.
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
Fishes, Macroinvertebrates, and Amphibians. Standard E729
American Society for Testing and Materials, Philadelphia, PA. -
ASTM. 1993b. Guide for Conducting Static Acute Toxirity Tests
Starting with Embryos of Four Species of Saltwater Bivalve
Molluscs. Standard E724. American Society for Testina and
Materials, Philadelphia, PA. . H :
ASTM. 1993c. Guide for Conducting Renewal Life-Cycle Toxicity
Tests with Daphnia macma. Standard E1193 . American Society for
Testing and Materials, Philadelphia, PA.
ASTM. 1993d. Guide for Conducting Early Life-Stage Toxicitv
Tests with Fishes. Standard E1241. American Society for Testino
and Materials, Philadelphia, PA.
ASTM. 1993e. Guide for Conducting Three-Brood, Renewal Toxicity
Tests with Cenodaphnia dubia. Standard E1295. American Society
for Testing and Materials, Philadelphia, , PA. .'."..
ASTM. 1993 f. Guide for Conducting Acute Toxicity Tests on
Aqueous Effluents with Fishes, Macroinvertebrates, and
?™thouse< 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.
and ?-J- Versteeg« 1992- A Statistical Procedure
o Contlnuous Toxicity. Data,, Environ. Toxicol., Chem.
11:1485-1494. .
Hoekstra, J.A. , and P.H. Van Ewijk. 1993. Alternatives for the
No-Observed-Ef feet . Level . Environ. Toxicol. Chem, 12:487-194.
Kilpatrick, F.A. 1992. Simulation of Soluble Waste Transport
^ „?£ P in Surface Waters Using Tracers. , Open-File Report
?i7Ji m 5*S'-,Geolog:LCal Survey' Books and Open-File Reports, Box
25425, Federal Center, Denver, CO 80225.
**'G- . 1966- EJtperimental Statistics. National Bureau
of Standards Handbook 91. (Issued August 1, 1963; reprinted
October 1966 with corrections) . U>S. Government Printing Office,
Washington, DC.
76
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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 arid 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. Banner and D.J. Harisen, 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. 1983.a. 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 orPB85-121101. 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 -227 04 9..:. National. Technical Information
Service, Springfield, VA.
U.Si EPA. 1991a. Technical Support pocument for Water Quality-
based Toxics Control. EPA/505/2-90-boi 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. tEPA-60Q/4-91-OiO. National Technical
'Information Service, Springfield, VA.
77
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U.S. EPA. 1992. Interim Guidance oh 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 Metlrods 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.
U.S. EPA. 1993d. Dilution Models for Effluent Discharges.
Second Edition. EPA/600/R-93/139. National Technical
Information Service, Springfield, VA.
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Appendix A: Comparison of WERs Determined Using Upstream and
Downstream Water
** -"' ' ' ' F • *
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 us,e 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 rea-1 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 TQC 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
• ' : . '" . : \ '" 79 * . ' : .....
-------
number and type of upstream point and nonpoint sources the
frS?™V?d' ^gnituje of fluctuations, and whe?he? the 5IR
3 SrSS^ " B*~ *» *** of complete mix
Downstream water is a mixture of -'effluent and upstream water
each of whxch can contribute to the WER, and so there^re twd
components to a WER determined in 'downstream water- the
effluent component and the upstream component . . The existence
? ™se,3two comP°nents has the following implications-
-L. WERS determined using downstream water are likely to be
larger and more variable than WERs determined usino
upstream water. u
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 ) eicj-uent
4. The magnitude of the effluent component of a WER is likelv
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 o? way's,
and so the effluent component of a WER can be due to a
SSJfS ° f^0?1 ef f ects ih addition to such factors as
hardness, alkalinity, pH, and humic acid.
UStot?Le^1USnt comSone2t "^ht be due, in whole or in
Iv r^e dxscharse of refractory metal (see Appendix
25 JJ* ?ann?t be Bought of simply as being Sused by
ffZ**0* Water ^alitY on the toxicity of the metal.
with downstream WERs is so much simpler if the
PS (?WES ' a?d the uPstream WER (SwER) are additive
LS x? desirable to understand the concept of additivity
ppe^dix GK experxmental determination, and its use (see
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
S?f? S^ST °? the effluent of conSem, i.e. , the bwS W^S
will be determined using the same site water.
When WERs are •determined using downstream water, the magnitude
™ will probably depend on the concentration of
f o
ettiuent in the downstream water used (see Appendix D) The
concentration of .effluent in the site water wi?? depend on
80
-------
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
dccWER, 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
• ' -. ''• :•-••••''' ' . .81 •':.••'.. . •'.-.'
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SJ2S? ?' ??S G9cWE\should probably be determined for the
complete-mix situation, but the site-specific CCC will have to
£5 *S£naS the fdge °f the CCC mining zone. Thus dilution of
the WER from the edge of the CCC mixing zone to the- point of
complete mzx is taken into account, but dilution of the metal
3.S HO '""'"
™o *"* !^ithe,r an acute nor a chronic mixing zone, both
the CMC and the CCC apply at the end of the pipe, but the CCC
should sto.ll be determined for the complete-mix situation.
C. Definition of site.
In the general context 'of site-specif ic criteria, a "site" mav
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.
SS1 a£ yT'ra»hen a site-specific criterion is derived using a
WER, the WER is to be adequately protective of the entire
sice, if, for example, a site-specific criterion is being
derived for an estuary, WERs could be determined, using samples
2w £ surface water obtained from various sampling stations,
• H ?i-to av°ld confus:i-on, should not be called "sites". If
• Jll the WERs were sufficiently similar, one site-specific
S Te™=°n COUld 5e.deriv,ed 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 .
data ^^t indicate that the estuary
two- or more sites, each with its own
~ •
^hat ^wuld'be applied when defining the
SoSld hS^??!™38? in th*sit* 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 rnetals are
persistent, although some oxidation states are riot
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 TOC are likely.
to increase the total recoverable concentration of the
metal and the total recoverable
-------
WheneWERs are determined Using downstream water, the followino
defined^10™3 should be taken into account when the site is
If a^site-specific criterion is derived using a WER that
SiX1^6? fc? 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.
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
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
tK^L^S111 thS ?lume- ff ^ere is np toxicity within
tne 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
?^/ y*? the !lte 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 -
°J £°?£OXXC 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
84
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WERs are expected to be less variable than total recoverable
JWERs, especially where events commonly increase1 TSS and/or
TOC. 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 up'stream effluent contributes
to the _ WER. •
When a WER is determined in downstream water, the four major
sourqes 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 inmost 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
: ' ... . . : 85-' . ' '••.'. - -
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methods presented herein might be relatively easy to
S£!r^sdaj! °hemiCal «*««-*. «, performed
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
e£x -Luent . . •
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 concentration7 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
°!!t2rm?-ned us^ng actual downstream water, but the second can
E£2 *S y Pr?vlde 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
ri^?5S;tl?* 2f b°? Jhe effluent and the upstream water .
can be determined, and the exact ratio in the downstream water
xs
Use of simulated, downstream water is 'also a way to study the
ESS"? between Jhe WER and the. ratio of effluent to upstream
££? r. at pnf .??int -in tine,- which is the most direct wa? 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
?™en~a£i°Vf e?fluent ^creases. If this assumption is
true, as the flow increases, the concentration of effluent in
ObLS^ J^ ™a*er Wi^ 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 .
usenuj. . . •
E.
The fate of 'metal that has reduced or no toxicity.
haS reduced or no toxicity at the end of the pipe
^6 ^°X1^ at Some time in the future. For example,
™ £at x? ^ the water col™i and is not toxic now might
become more toxic in the water column later or might move into
86 '.'./;
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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 hot 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 r(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,
, , - !
• . • • ••'••; .... s? ' • •' . • . . ' •
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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 hot 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.
88 .
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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 - ."''••'.'.."' '.-..-.-- ••-.•••'• .'...--•/--• - •-'• ..'.. ..v ,.„'... ...
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.
89
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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, crenera,
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 bodies5 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
90
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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 the Recalculation 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 IKS. 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. v ••"...
Each step is discussed in more detail below.
'"•.-.- .91. "•••".' ' ••'".'•
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A, Corrections . f .
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 frbm the national dataset just because the quality
of the datum is claimed to be suspect. If additional data are
avaxlable -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 includes those 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 no't 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 sit'e; a species cannot be deleted based
92
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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
L . 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 P 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.
"~ . : ' V . ' »' •
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
If
"No", go to step 2.
"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 siter-specific dataset.
94
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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 bv 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 Recruirements
•';'•' • . - .
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 CCG
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
•':•'*''"'• ''' • 95 ••- ' ; ': .. • •
-------
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 the1 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.
FT 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. ...",' •
96
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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 Glass Order Family Species
Annelida Hirudiri. Rhynchob.
Bryozoa {No species in this
Chordata Osteich.. Cyprinif.
Chordata Osteich. Cyprinif.,,
Chordata Osteich. Cyprinif.
Chordata Osteich. Cyprinif.
Chordata Osteich. Salmonif.
Chordata Osteich. Percifor.
Chordata Osteich., Percifor.
Chordata Amphibia Caudata
Glossiph. ... Glossip. complanata
phylum occur at the site.) ,
Cyprinid. Carassius auratus
Cyprinid. Notropis anogenus
Cyprinid. Phpxinus 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
Phoxirius oreas D
Tinea tinea D
Ictiobus bubalus F
Oncorhynchus mykiss O
Lepomis cyanellus S
Lepomis macrochirus G
...Perca flavescens D
Xenopus laeyis 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 .
Cypririif .
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.
97
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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°?
- (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 (estuarin-
and marine surface waters, A lack- of awareness of true ambient
concentrations of metals in fresh and salt surface waters^an £e
both a cause and a result of the problem. The ranges of
dissolved metals that are typical in surface waters of the Unit^r
?QKe%a™a? fro™the immediate influence of discharges m?uXnd
af ? 19^ ale: ** * 1985'1987*' Trefry et al. 1986; wlndom et
Metal
Salt water
Cadmium
Copper
Lead
Nickel
Silver
Zinc
0.01
0.1
0.01
0.3
to
to
to
to
metals
p
some metals
Fresh water
(u.cr/LV
0.002 to 0.08
0.4 to 4. ,
0.01 to 0-.19
1. to: 2.
0.03 to 5.
has published analytical methods for
waters and wastewaters, but these methods
(termination of ambient concentrations of
some surface waters. Accurate and precise
T™ rl^^C' ^i i <•*».» «••«•* wk ^t* A.« J_ __. A_ • _ . •
0.2
3.
1.
5.
0.005 to 0.2
0.1 to 15.
°"Clean
during, collecting, handling,
to avoid
2* J|ed°f ,«* -"ifled. reference ,
6. Use of replicates to assess precision.
Z_! P3?, °? certif ied standards. .
:,J!e?se/Jthe term "clean techniques" refers to
that reduce contamination and enable the accurate .
^..^ 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
98
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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. 891116 items, however, are
"considered so important by analytical, cnemists 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
• . '.-:.-..-. " • ..:' .. • 99 "": •••-' •. •' '"..-' •.
-------
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,
fj*?S£i?nV£nd/°r fraction) and quantification (instrumental
analysis), the general principles of contamination control should
£>e applied. Some specific recommendations are:
Powder-free (non-talc, class-iaO) 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. ««««*
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 analyzed for the
metal (s) of interest before use.
The water used to prepare acidic cleaning. solutions and to
rinse labware, sample containers, and sampling apparatus may
^/^Pa™d^/istillation' deionization" or reverse osmosis,
and should be demonstrated to be metal-free.
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.
a
Srf *Sdi=*g °f^Sa^1^S ±n the ^oratory, 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
100 clean ro °r gl°Ve box should be located within a class -
«£?a£e/ ^sents, sampling apparatus, and sample containers
2?SSS ?^Ver ?S ,L0p.en 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.
100
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Separate sets of sample containers, labware, and sampling
apparatus should be dedicated, for different kinds, of samples;
e.g., surface water samples, effluent samples, «tc.
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.
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.
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.
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
101
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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. , ' • r
1. Labware, sampling apparatus, and sample containers should be
rinsed and not rinsed with sample as necessary to prevent hiah
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, labware, etc., that
contact the sample only before it is acidified should be
rinsed^ith 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
ri <.?0t be aci3ified 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 j.nvt>lved and 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,
102
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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-I'O.O 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 Afield. 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 poured 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.
, , 103
-------
- , . •,_,-
Avoiding interferences
a. Potential interferences should be assessed for the specific
mealSeT analysis technique used and for each me?S to be
b. If direct analysis is used, the salt present in high-salinitv
saltwater samples is likely to cause interference in most
instrumental techniques. «ience in most
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
^^^^^^^"^^^^^^™"*^™""^^""'^^^^~'^~
. (Prob?dura1' method) blank consists of filling a
tain^-?^h analvzfd metal-free water and procesIiSg
,, acidifying, etc.) the water through the laboratnrS
procedure in .exactly the same way as a sample A laboratSrS^
blank should be included in each set of Sn or f ewe? sSp?2
to check for contamination in the laboratory, anl shouS
S S3^ a2 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 .
.
etc coll- rugtu
etc..^ collecting the. water in a sample container
blSSj So2l!hS WatSr the,same as a field SmpL? A SSi
blank should be processed for each sampling trip Senar£i-
field blanks should be processed for the ?otal ?ecovSable
measurement and for the dissolved measurement, if
irtErfaSStSrv'th SitS* Fi?ld. blanks Sh0uid b
in tne laboratory the same as laboratory blanks.
Assessing accuracy
should be determined for each analytical
stsrs-
ol 9 S°» e -l— •*. each g.oup
tdSitions^Pike
-------
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
filter.s 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,. Mf positive-pressure
filtration is used, the air or gas should be passed through a
0.2-nm 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, ^buf they do not rhave 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.
105 '-,":'•
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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 sodiumjchloride
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 techniques11; 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 and 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).
106
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References '•'..
Adeloju, S.B.'v and A.M. Bond. 1985. Influence of Laboratory
Environment on the Precision and Accuracy of Trace Element
Analysis. Anal. Chem. 57:1728-1733.
Berman, 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.D. 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.
. • 107
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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. 1985. - Dissolved Zinc in Rivers
Nature 317:49-52.
Shiller, A.M., and E.A. Boyle. 1987. Variability of Dissolved
Trace Metals in the Mississippi River. Geochim. Cosmochiiru 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 Informal-ion
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. Huan. 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.
108
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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), arid 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 etal. (1993) concluded,
that "mortality was best predicted not by the free' A13+ activity
but rather as a function of the sum Z( [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.'
109
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Forms of Metals
Even if the relationship of toxici'ty 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-fim or a 0.40-pm membrane, filter" and
-particulate metal" is defined as "total recoverable metal minus
dissolved metal". Metal that is in or on particles that pass
tnrough 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 duetto 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
a?vo*ve gradations, a simPle classification scheme such as this
rs,2£uld be sufficient to establish, the principles regarding how
WEKs are related to various operationally defined forms of metal
and how this affects the determination and use of WERs.
i ( . • -
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.
110
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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 met;al 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
-------
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
?r. .?^h 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
112
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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-spepific 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. ,
113
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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
theAtoxic 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
xn 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
(x.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:
114
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a. Toxic dissolved metal: .
.This form contributes to the tdxicity 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 experimentalx
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
.'• . . . us .'••.•'•'.-••
-------
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.
116
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Use of a dissolved criterion without a dissolved ,WER has three
disadvantages, hpwever:
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 i, 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:
The subscripts "S" and "L" denote site water and laboratory
dilution water, respectively, and:
S = the _ concentration of Refractory metal in a water. (By
definition, all refractory metal is nontoxic metal.)
117'
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N 3 the concentration of Nontoxic labile metal in a water.
T = the concentration of Toxic labile metal in a water.
A2\T = 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
_ TB
Thus H is a multiplier that accounts for a proportional increase
or decrease in the toxicity of thfe toxic forms in site water as
compared to their toxicities in laboratory^ dilution water.
Therefore, the- general WER equation is:
NS
RL + NL + *NL + (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, #=i, then WER ~ l.
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 = 0 and
rrrm Rs + Ns
WER = —£ ?~
118
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Even though laboratory dilution water is low in TOC 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 *TL. Hydroxide, carbonate,
chloride,. TOC, and TSS can increase *NS. Anything that, causes
Aflk 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., tf*l.
b. Atf 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 :
which would require that Rs = NS = AWS = RL = NL - AW^ = o . (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
-------
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 ff=l, so.
H will be retained for generality.
1. The simplest situation is when Rs > o, i.e., when a
substantial concentration of refractory metal occurs in the
site water. If, for simplification, it is assumed that
Ns » &NS ** EL « NL = *NL * 0 , then:
.
The quantity TL + &TL Obviously changes as the sensitivity of
the toxicity test changes . When Rs = o , 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 + AWS) > 0. If, for
simplification, it is assumed that Rs « RL '» NL = &NL .= 0, then:
' ur
- a. If (Ns + Aflrs) > 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 199.3) . 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 s % complexed ± v
% toxic %uncoaplexed ' .
For the complexed nontoxic metal: , - '
v _ concentration of nontoxic metal
concentration of toxic metal
120
-------
In the site water, the concentration of complexed nontoxic
metal is (Ns + A#S) and the concentration of toxic metal is
(Ts + ATS) , so that: . •
(Ns.+
(2V+ Aiy H(TL
.„' • VJI(TL + AT,) + H(TL + AT,)
WER = = tyr + J5T =
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,
and ' . .
+ H(TL + ATi) ' : VgH + H H(VS
WER
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 V^, the change could result in a larger WER, rather
121
-------
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 5j 1 and if vs < VL, which would mea.n £hat
WER < l . 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 + jtJfg) > 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 "% dissolved0, which is assumed to
correlate with "% toxic". If the endpoint for a toxicity
test is below the solubility limit of the precipitate,
(Ns + AATS) « 0 , whereas if the endpoint for a toxicity test
is above the solubility limit, (Ns + Aifc) > 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 .,
-------
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) the7'duration of the test, (b)
whether, the endpoint is based on a lethal or sublethal effect, or
ic) 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 dilutipn 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 informat ion ..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
', ,tp 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 ' . .. :' ' '. ; ' ' '
-------
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 pan.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 TOC, 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 nontbxic 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: - '•-.-.
Ts: 100 Ug/L
Ds: 10 ug/L (10 % dissolved)
Q-: - 24 cfs
Upstream water:
40 ug/L
38 ug/L (95 % dissolved)
48 cfs
7* •
J-V
D,,:
Downstream water: , :
TD: 60 ug/L
X>D: . 36 ug/L (60 % dissolved)
O,: 72 cfs
where: .'.''.,'- - -. .'••-.'• ..'"-." '....-. •'_ '.
T = concentration of total recoverable metal.
D = concentration of dissolved metal.
• 0 = flow.. '. • . '•"..- .:- .'..,'...' •. • •'. • • • .'-•-.
The subscripts E, U, and D signify effluent, upstream water, and
downstream water, respectively.
By conservation of flow: QD = QE + Qy .
By conservation of total recoverable metal: TjffD = TJ£E +• T&a .
If P = the percent of the total recoverable metal in the '
effluent .that becomes dissolved in. the downstream water,
For the data given above, the percent of the total recoverable
metal in the effluent that becomes dissolved in the downstream
water is:
P = 1001(36 ug/L) (72 cfs) - (38 ug/L) (48 cfs) ] _-,-«.
(100 ug/L) (24 cfs} ~~~ ~ -- 32%,
\ ' ; -h . ''.'••,,- = • . '
which is greater than the 10 % dissolved in the effluent and less
than the 60 % dissolved in the. downstream water.
125
-------
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.
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 _ 0
Dissolved WER ~ 0.775 ~ site water % dissolved ~ 40
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
-------
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 oag/L) = 1.45
Dissolved WER = (232 ug/L)/(320 ug/L) = 0.725
Checking the Calculations
'v '- ' _, • . ' " , ':• '_ .
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. , ,
, ' ; '; .;; '; ','." ' . >ri :' '•'• 127••' '"• '. .' " - '" " '•". ' • • ': '
-------
Figure D4r 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: 400 ug/L
Refractory particulate: 200 ug/L
Toxic dissolved: 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: 6 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: 420 ug/L
Refractory particulate: 298 ug/L
Toxic dissolved: 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)]/6 = 420 ug/L
Dissolved: [3(200) + 3(44+0)]/6 = 122 ug/L
Refractory particulate: [3(200) -s- 3(396)]/6 = 298 ug/L
Assumed National CCC InCCC)
Total recoverable = 300 ug/L
Dissolved =240 ug/L
128
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Upstream site-specific CCC (ussCCC) ,. ,
t' , ! - ' • - - ' - v : '
Assume: Dissolved cccWER = 1.2 / • ,
Dissolved ussCCG = (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 ~ ~50~ = 1
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])/0.1}+{(400 ug/L)/2} = 3520 ug/L
This calculation determines the amount of dissolved
metal contributed by the effluent, accounts for the
. 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 BO _ K K,
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 ' ' -•'•'.-•'
-------
Starting with the total recoverable dssCCC of 3520 ug/L
(6 cfs) (3520 ug/L) - (3 cfs) (400 ug/L) _
. 3 ~ - •" - -
Starting with the dissolved dssCCC of 432 ug/L
(6 cfs) (4.32 ug/L) - (3 cfs) (400 ug/L) (0.5) --,'
(3 - - ~ -- 664°
Checking the Calculations
Total recoverable :
(3 cfs) (66AO ug/L) + (3 cfs) (400 ug/L) _ ,_0ft
; e cfs ~~~ ^~~ ~ 35ZO
Dissolved:
(3 cfs) (6640 ug/L) (0.10) + (3 cfg) (400 ug/L) (0.50) .^ /T
6 cfs : - - "~ — ~~ = 4'32 uflr/I, .
The value of 0.10 is used because this is the percent of the
total recoverable metal in the effluent that becomes dissolved
xn 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:
16 cf£r] ~ (3
^*.*~.
— '• - '—• = 664°
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.
TOO
O
in
8
8
LOG OF CONCENTRATION OF METAL
131
-------
Figure D6: A Generalized Precipitation Curve
• ; " ' • ' " , I
The curve is for a constant concentration of the precipitatino
ligand and an increasing concentration of the metal.
100
Q
IU
O
CO
CO
5
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. 55-62.
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. (JiO. 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 14i
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, e'd.) Report TR 247
Water Research Centre, United Kingdom.
Hansen, D.J. 1993a. Memorandum to G..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.
133
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Appendix E: U.S. EPA Aquatic Life Criteria Documents for Metals
Metal
Aluminum
Antimony
Arsenic
Beryllium
Cadmium
•
Chromium
Copper
Lead
Mercury
Nickel
Selenium
Silver
Thallium
Zinc
.- EPA Number
EPA 440/5
EPA 440/5
EPA 440/5
EPA 440/5-
EPA 440/5-
EPA 440/5-
EPA 440/5-
EPA 440/5-
EPA 440/5-
EPA 440/5-
EPA 440/5-
EPA 440/5-
EPA 440/5-
EPA 440/5-
-86-008
-80-020
-84-033
-80-024
-84-032
-84-^029
-84-031
-84-027
•84-026
•86-004
87-006
80-071
80-074
87-003
NTIS Numbeir
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
134
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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.
' l - .' " -' ••' 135 ;' ' " "' ', •..= ''•-
-------
Multiple-Discharge Situations
National Toxics Rule (NTR). incorporated WERs into the
life criteria for some metals, it might be envisioned-
that more than one criterion could apply to a metal at a si?e if
*'
has permit
whenever a_site-specific criterion is to be
site at which more than one discharger
for the same metal, it is important that all
IiI—T^ • together with the appropriate recrulat-orv
authority to develop a workplan that is designed to derive a
site-specific criterion that adequately protects the entire site.
discharger? ide?"y ^ted'for taking .into account more than one
Method 1 is straightforward if the dischargers are sufficiently
ar downstream of each other that the stream can be divided into
jer. Method 1 can also be fairly
are additive, but it will be complex
2. Deciding whether to use a
. water or an actual downstream water can be
»^V~T *" " flowin£r-water multiple-discharge situation. Use
actual downstream water can be complicated by the existence of
multiple mixing zones and plumes and by the possibility '
varying discharge schedules, <->iaaa o=™l ~-~iJi___ _-7_^y
if effluents from
.sras
ol f^° a=count synergism, antagonism, -and additivi!? if one
of the discharges stops or is modified substantially, however it
if tha ™l% be necessary to determine a new WER, eicepriossibly
if the metal being discharged is refractory. Situations iJ°sslD-Ly
need to be^andled '
Special Flowing-Water Situations
jntended Jo apply not only to ordinary rivers, and
S^°J° ftreams ^at some people might consider
«™ ««.* as streams whose design flows are zero and
streams that some state and/or federal agencies might refer to as
etc ?Sue SeSdSfc"' "^tat-creating", «effluen?-doBlnIted°,
^oo,:^(?^ differences between agencies, some streams whose
design flows are zero are not considered . »ef fluent-dlpSdent^?
136 .
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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 ho 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-specifid 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:
137 :." '.
-------
* S^ne?r;fieid effe?ts 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 saml as a sinSe
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
?^at?d Slte water. A dye dispersion study (Kilpatrick
1992 might be useful, but a dilution model (U.S. EPA ^3) is
JSSS F bS a more.cost-effective way of obtaining ' S
information concerning the amount of dilution at the edge of
une mixing zone. . . • •
-Hie near-field effects of a single discharge" that is near a
be«tebftL^^%S0dy °f frSSh °r Salt Wat*r 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
!S£ng zone; J^ Potential point of Impac? .of near-?ieJd
effects will often be an embayment, bayou, or estuary that is
a nursery for fish and invertebrates and/or contain?
commercially important shellfish beds. Because of their
i^?SaX?;J?ese areas should receive special consideration
**Sje ^termination and use of a WER, taking into account
?^S ^°f ^ater and disqharges' mining patterns, and currents
(and tides in coastal areas). The current and flushing
So^ ^S J? estuar:Les can result .in increased pollutant
concentrations in confined embayments and at the terminal un-
SSSSJ' P°f,tioS.°f the estuary^ue to poor tidal fining Ld
^ ?g ^ 1_r^e.dlsPersion studies (Kilpatrick 1992) can bl
used to determine the spatial concentration of the effluent in
the receiving^ater, but dilution models (U.S. EpJ J^might
not be sufficiently accurate to be useful. Dye studies of
discharges in near-shore tidal areas are especially complex
Pye injection into the discharge should occur over a?lSaS"
^?^a, P^^rably tw° °r three' c°mplete tidal cycles?
S^ff^ent dispersion patterns should be monitored in the
SSS??^^ °n consecutive^ tidal cycles using an intensive
sampling regime over time, location, and depth Information
de??;?1^ disPersf°n and the community at ?Isk can Sensed to
dSSSS E? appropriate mixing zone(s), which might be used to
define how to prepare simulated site water.
References
S:Lm^ation of Soluble Waste Transport
- Sro cerS US±ng Tracers. Open-File Report
25425 P^S;?;in °giCai Survey' Books and Open-File Reports, Box
-it»4^5, Federal Center, Denver, CO 80225.
?;?* 5P^-^^993- Dilution Models for Effluent Discharges
Second Edition. .EPA/600/R-93/139. National Technical
Information Service, Springfield, VA.
138
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Appendix 6: 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 quantity 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: -
2Q(eWER) * BO(uWER) _ 20(40) + 80(5) _ 800 + 400 1200 _ . _
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)
" ' . 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
-------
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 Additivity .
The major advantage of additivity of WERs can be demonstrated
S?™?*J?e«.£* fluent 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 addit/ve In
addition, the complete-mix downstream water at design-flow
conditions will be assumed to be 20 .% effluent and 80 % upstream
'
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
f£nc^tration 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 b?
?£ lutf^ simultane°usly; proportional dilution of the metal and
S£i2?5 ^ Prevent the metal from causing acute or chronic"
SSSi^fi fc *!£ ^lution. When the upstream flow equals the
design flow/ .the WER in the plume will .decrease from 40 at the
5Sd,°f*-the P13?e t0 P 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
Sa?nS^r^tiCl W^1:L ^ffSCt the WER- ^eri stream fl°w 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 1-he
^So1^ C?° J? 1 ""I711,! the foll°wing conclusions are valid when
metal in 10° % ef f^ent is less than 40
°£ ^ metal >1 water
This metal will not cause acute or chronic toxicity in the
upstream water, in 100 % effluent,, in the plume, or in
downstream water.
There is no need for an acute or a chronic mixing zone where a
lesser degree of protection is provided.
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
3
140
-------
of additivity that exists. It would be more difficult to take
into, account the portions of the eWER and uWER that are not
.additive. . r •
'..'••' ' : . •' '' • • . .
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
• • ' ' , •••'.'-.." 141"":. - • • ': .- • ' ':•
-------
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?
°f in the
™/ water never exceeds
n Jr/r7 ;* discharger might want' to determine how much above
40 mg/L the concentration could 'be in 100 % effluent If £br
example, the downstream water at the edge of the chronic mixino
zone under design-flow conditions consists of }0 % Sf^eS and
zone would be:
TQ(QWER) +30(uNER) _ 70(40) +30(5) 2800+150
100 100 = loo = 29 •5
offSr™/?he IS*****3* concentration allowed at this point would
29.5 mg/L. If the concentration of the metal in the uostream
osre
water was 0.8 mg/L the maximum concentration allowed n 100 I
effluent would be 41.8 mg/L because:
70(41.8 ntcr/L) •»• 30 (0.8 mer/L) _ 2926 mg/L * 24 mall, ,
lOO ~ - " - •JU— =29.5 mg/L .
%flui « r . conce*tration of the metal in 100
* ®f?iuen^ « 41.8 mg/L, there would be chronic toxicity inside
the chronic mixing zone. If the concentration in 100 % effluent
th/SB^o^S11 4J'8 ^g/L^ ^here would be Chronic toxicity pal?
the uraS aL ??vf r
-------
that starts at the edge of the chronic mixing zone and extends
all the way across the stream, there^wpuld 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 Additivitv
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).
• ' " * " ' • NV--.
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 tota^. 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 theieffluent, 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 TOC;
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.
143 • •
-------
Insofar as the eWER and uWER are additive, the magnitude of a
downstream WER can be calculated from the eWER1, 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
OCCU3T£> • *
°CCUr if' f0^ example, any portion of
the eWER is applied to an area of a stream where the
effluent does not occur.
b- ^??rSte?i0n i^J °CTCUr ^f ' for example, an unnecessarily
small portion of the eWER is applied to an area of .a stream
. where the effluent occurs. »i-f«euii
Even though the cpncentration 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
Reference
Broderius, S.J,. 1991. Modeling the Joint Toxicity of
T^°ii°=i°S m° ^*tic Organisms: Basic Concepts and Approaches
/2 V^S C To^xcol°9y and Risk Assessment: Fourteenth Volume
(M.A Mayes and M.G. Barren, eels.. ) ASTM STP 1124. American-
Society for Testing and Materials, Philadelphia, PA. pp. 107-
JL^i / . . .
144
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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
stenohaiine organisms should not be tested in high-salinity
water, whereas high-salinity stenohaiine organisms should not
be tested in low-salinity water; it is not known, however,
t 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 maqna
and Hvalella 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 stenohaiine 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
. "— • , , 145; . . • ' "•'
-------
• f£lt: £.r ^rine Deduces the toxicity of the metal and
thereby increases the WER. mecaj. ana
2) If the site water is high-salinity/selecting test
organxsms should not < be difficult because mSny of
with
2. It is especially important -to consider the availabilitv of
test organxsms when saltwater species are to be used becfii^
many of the commonly used saltwater species art not culSred
and are only available seasonally. cultured
3. Many standard published methodologies for tests with sal tmt- «r-
specxes recommend, filtration of dilution wa^er? effluISt
and/or test solutxons through a 3 7 -pm sieve or screen to
remove predators. Site water should^be fi?terel on!y ±f
predators are observed in the sample of the water because
.,
water gualxty characteristic, such as salinity
or is
win
^'^s^sss^s^.
146
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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. . .
"• ' ' ~ - - f
- . '••:'-' -"- ' ' '• 147 : .-•.'. -' '• ' ; '.
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Suggested Tests1 for Determining cmcWERs and cccWERs2
(Concentrations are to be.measured in all tests.)
Metal
Water3
cmcWERs4
Aluminum
Arsenic (III)
Cadmium
Chrom(Ill)
Chrom(VI)
Copper
Lead
-
Mercury
"*
Nickel
Selenium
Silver
Zinc •
FW
. FW
SW
FW
SW
FW
FW
SW
FW
SW
FW
SW
FW
SW
FW
SW
FW
SW
FW
SW
FW
SW
DA
DA
BM
DA
MY,
GM
DA .
MY
- DA
BM
DA
BM
DA •
MY
DA
MY
Y
CR
DA
BM
DA
BM ,
X
GM
CR
SL5 or FM
CR
SL or DA
GM
NE
FM or GM
AR
GM
MYC
GM
BM
FX
BM
Y
MYC
FMC
CR
FM
MY
u(-:<_:w
CDC
CDC
MYC
CDC
MYC
FMC
CDC
MYC
CDC
BMC
CDC
MYC
Y
Y •
CDC
MYC
Y
MYC
CDC
MYC
CDC "
MYC
Eit\S
X
FMC
BM
FMC
X
CDC
GM
NEC
FM
AR
X
X
Y
Y
I
FMC
BMC
Y
X
FMC
BMC
FMC
BMC
of a test specifies not only the -test species
and the duration of the test but also the life stage of the
adVSrSe effeGt(s) on whi<* the endpoint is to
rtm that are sensitive and are used in criteria
documents are not suggested here because the chances of the
™?Xf°£ga?1SinS oei£g availabl* and .the test being successful
might be low. Such tests may be used if desired!
148
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FW = Fresh Water; SW = Salt Water.
Two-letter codes are used for acute tests, whereas codes for
chronic tests cbntain 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 oh 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
Stronovlocentrotus 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
(Geriodaphnia. 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 LCSOs 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 LG50 (ua/L)
Cadmium 13.87
Copper ; 6.33
Zinc - 100.95
149
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^-?£~hr«.JC52 from a renewal test (renew at 48 hours) at 25°c
with fathead minnow (Pimephales £ronielas ) larvae that are 1
to 24 hours' old (AS1M 1993a; U.S. EPA 1993at The embryos
must be hatched, in the laboratory dilution water/ except ?hat
organisms to be used in the site water may be hatched in the
?£? WaSer; n^1^36 ffiust not be fed bef°re 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 ng/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
• Pro?edure (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. tests
,(°r LG5° if there is no immobilization) from a
48 h°UrS) With a S fr
a
uva te?t (renew at 48 h°"«> using
1993aK Polychaetes in the genus Nereidae (ASTM
CERONIC TESTfi
BMC
I • ' - -•
A 7-day IC25 from a survival and development renewal tecsi-
(renew every 48'hours) with a species of bivalve moilu-sf
•such as a species, in the genus Mulinia. One such test h4s
been described by Burgess et al. 1992. [Note- Wh^n
' *n
test has not been widely used.]
based on reduction in survival and/or
in a renewal test with a species in the genus
in thejfamily Daphnidae (U.S. EPA 1993b). The
150
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MYC
test, solutions must be renewed every 48 hours. (A 21-day
1 life-cycle test with Daphnia maana is also acceptable.)
FMC. A 7-day IC25 from a survival and growth renewal test, (renew
every 48 hours) with larvae (£ 48-hr old) of the fathead
minnow (Pimephales promelas) (UVS. EPA 1993b). When
determining WERs, the fish must be fed four hours before
each renewal and minimally during the non-renewal days.
. A 7-day IC25 based on reduction in survival, growth, and/or
reproduction in a renewal ^est with a species in one of two
genera (Mysidopsis. Holmesimvsis [nee Acanthomvsisl) 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.
rASTM. _ 1993b. Guide for ConductingStatic 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,
Nairragansett, RX.
'• . '• '"....'.-' -151'. . ,
-------
Chapman, G. A. 1992. Sea Urchin (Stronavlocentrotiis
Fertilization Test Method. U.S. EPA, Newport, OR.
Johns, D.M., R.A. Pastorok, and T.G. Ginn. 1991. A Sublethal
Sediment Tbxicity Test using Juvenile Neanthes sp.
(PolychaetarNereidae) . In: Aquatic"' Toxicology and Risk
Assessment: Fourteenth Volume. ASTM STP 1124. (M.A. Mayes and
for
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
oo'5 . .
Hif; EPA* 1993a- Methods for Measuring the Acute Toxicity of
Effluents and Receiving Waters to Freshwater and Marine
Organisms. Fourth Edition. EPA/600/4-90/Q27F. National
Technical Information Service, Springfield, VA.
Short-term Methods for Estimating the Chronic
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, VAJ
Voyer, R.A., and D.G. McGovern. 1991. Influence of "constant and
Fluctuating Salinity on Responses of Mvsidopsis 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 .
Aluirdnum potassium sulf ate 12-hydrate: A1K(SO4)2-12H2O
Arsenic(III) •
*Sodium arsenite: NaAsO2 "
Arsenic(V)
Sodium arsenate 7-hydrate, dibasic: Na2HAsO4 • 7H2O
Cadmium ''-.'', ',
Cadmium chloride 2.5-hydrate: CdCl2.2.5H2Q
Cadmium sulfate hydrate: 3CdSO4-8H2Q .
Chromium(III). .
*Chromic chlpride 6-hydrate (Chromium chloride): CrCl3-6H2O
*Chromic nitrate 9-hydrate (Chromium nitrate): CrJNO3)3-9H2O
Chrpmium potassium sulf ate 12-hydrate :.CrK(SO4) 2.12H2O
' Chromium (VT)
Potassium chromate: K2CrQ4
- Potassium dichromate: K2Cr2O7
*Sodium chromate 4-hydrate: Na2CrO4«4H2O . -
Sodium dichromate 2-hydrate: Na2Cr2O7• 2H2O ',"
Copper ••'.".• , . . " :' , • - ."' ' ''.'"'< '' .'. •: : , '•
*Cupric chloride 2-hydrate (Copper chloride) : C,uCl2.2H2O
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)- NiCl .-SH n
*No.ckelous nitrate 6-hydrate (Nickel nitrate) • Ni (NO > 6H o
Nxckelous sulfate 6-hydrate (Nickel sulfate) i NiSO?'6H2O *
Selenium (TV) .-',,'
*Sodium selenite 5-hydrate: Na2Se03.5H2O
.Selenium ( VI } " .-•....
*Sodium selenate 10-hydrate: Na2SeO4-10H2O
Silver
Silver nitrate: AgNO3
-silver
Zinc ' ;
Zinc chloride : ZnCl2
*Zinc nitrate 6-hydrate: Zn (NO, ) , . 6H,O
Zinc sulfate 7-hydrate: ZnSO4.7H2O •
*Note: ACS reagent -grade ^ specifications might not be available
154
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