Click here for DISCLAIMER Document starts on next page UNITED STATES ENVIRONMENTAL PROTECTION AGENCY WASHINGTON, D. C. 20460 FEB 22 1994 OFFICE OF WATER EPA 823 B 94 001 MEMORANDUM SUBJECT: Use of the Water Effect Ratio Standards FROM : Tudor T. Davies, Director Office of Science and TO: Water Management Division Directors, Regions I  X State Water Quality Standards Program Directors PURPOSE There are two purposes for this memorandum.

The first is to transmit the Determination and Use of Water Effect ratios for Metals. EPA committed to developing this guidance to support implementation of federal standards for those States included in the National Toxics Rule.

The second is to provide policy guidance on whether a State's application of a water effect ratio is a site specific criterion adjustment subject to EPA review and approval/ disapproval.

BACKGROUND In the early 1980' s, members of the regulated community expressed concern that EPA's laboratory derived water quality criteria might not accurately reflect site specific conditions because of the effects of water chemistry and the ability of species to adapt over time. In response to these concerns, EPA created three procedures to derive site specific criteria. These procedures were published in the Water Quality Standards Handbook, 1983. 2 Site specific criteria are allowed by regulation and are subject to EPA review and approval. The Federal water quality standards regulation at section 131.11( b)( l) 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. ham 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)( 8). Am 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 mite specific criteria is applicable to criteria expressed as either total recoverable metal or am 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 met 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. 3 Each of the three options for deriving a final water effect ratio presented in this interim guidance meets the l ciedific and technical acceptability test for deriving mite specific criteria.

Option 3 is the simplest, least restrictive and generally thm least expensive approach for situations where l imulatmd downstream water appropriately represents a " site." It is a fully acceptable approach for deriving the water° ffect ratio although it will generally provide a lower water effect ratio than the other 2 optionm. The other 2 option8 may be more costly and time consuming if more than 3 sample period8 and water effect ratio measurements are made, but are more accurate, and ray yield a larger, but more scientifically defensible site l pmcific criterion.

Site specific criteria, propmrly determined, will fully protect existing uses. The waterbody or segment thereof to which the site specific criteria apply must be clearly defined. A mite can be defined by the State and can be any mite, nail or large, including a watershod or basin. However, the site 8pocific criteria must protect the mite am a whole. It is likely to be more cost effective to derive any mite specific criteria for am large an area am pommible or appropriate. Xt is whamited thmt mite specific criteria are ambient water quality criteria applicable to a site. They mre not intended to be direct modifications to National Pollutant Discharge Elimination Systera ( NPDES) penit liritm. In romt cm888 the naitea will be synonymous with a State's standards. " megm8nt" in its watu quality By defining l itmm on a larger l cmle, titiple dischargers can collaborate on water effect ratio tuting and attain appropriate mite specific crituia at a rmduced cost.

More attention ham beon given to water effect ratios recently because of the numerous discussions and mmetingm on the entire question of metals policy and because WERm were specifically applied in the National Toxic8 Rule. In comments on the proposed National Toxic8 Rule, the public gum8tfonmd whether the EPA promulgation l hould be based solely on thm total recoverable form of a metal. For the reasons set forth in the final preamble, EPA chose to promulgate the crit8ria bm8ed on the total recwerabl8 form with a provi8ion for the application of a water effect ratio. In addition, this approach wmm chomen because of the unigue difficulties of attempting to authoritm site specific criteria modifications for nationally prorulgatmd criteria.

EPA now recommends the use of dissolved metal8 for Statms revising their water quality standards. Dimsolvmd crit8ria may also be modified by a site specific adjustment. 4 While the regulatory application of the water effect ratio applied oily to the 10 juri8dictio~ included in the final National Toxicm Rule:` for aquatic life retmls criteria, we underlltood thmt othu. Statu would F interested in applying m to their l dopted water quality stmndard8. The guidance upon which to bm8e the judg# nt of the acceptability of'the water effect ratio applied by the State i8 contained in the attached provide8 l dditionh information on the recmlculation for site 8pecific crituia radificaeioru. procmdurm A central qu88tion concerning UERm i8 whether their use by a Stat8 reeult8 in 8 mite 8pecific criterion 8ubject to EPA review and approval under Section 303( c) of the Clemn Water Act?

Derivation of 8 vmter effect ratio by a State is a 8ite specific criterion adjwtmnt l ubjwt to EPA review and approval/ dbapproval under Section 303( c). There are two optionm by Which thi8 review Can be l Ccarpli8hed.

Option 1: A State my derive and 8 t each individual watu effect ratio dewination to EPA for review and 8pprovml. Thi8 would be accoBpli8hed through the normal review and revision process umed by a State.

Option 2: A State can Bnd its water quality standards'to provide a formal prowdura which include8 derivation of water effect ratios, appropriate definition of situ, and enforceable monitoring provisions to amsure that demignatti usei l ra protected. Both thi8 ProCedure and aa reSUltin criteria would be 8ubject to full public participation requir nts. Public review of a mite 8pecific criterion could be accarplimhed in conjunction with the public review required for permit i8suance. EPA vould review and approve/ disapprove this protocol am a rmvimed 8tandard once. For public information, we recommend thmt once a year the State publi8h a list of sit specific criteria.

An exception to thi8 policy appliem to the waterm of the jurisdiction8 included in the National Toxic8 Rule. The EPA m is not required for the jurisdiction8 included in the National Toxic8 Rule where EPA established the procedure for the State for applicmtion to the. criteria promulgated. The National Toxic% Rule wmm m forpal rulamking process with notice and cment by which EPA pre authorized the use of a correctly applied water effect ratio. That 8 m process has not yet taken place in States not included in the National Toxic8 Rule. 5 However, the National Toxics Rule does not affect State authority to establish scientifically defensible procedures to determine Federally authorized WERm, to certify those WERm in NPDES permit proceedings, or to deny their application based on the State's risk management analysis.

As. demcribed in Section 131.36( b)( iii) of the water quality standards regulation ( the official regulatory reference to the National Toxics Rule), the water effect ratio is a site specific calculation. As indicated on page 60866 of the preamble to the National Toxic8 Rule, the rule was constructed a8 a raWtable presumption. The water affect ratio is assigned a value of 1.0 until a different water effect ratio is derived from l uitable 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 Toxic% Rule and apply another value of the water effect ratio in order to establish a site 8pecific criterion. The site specific criterion is then used to develop appropriate NPDES permit liaitm. The rule thus provides a State with the flexibility to derive an appropriate site specific criterion for specific waterbodiu.

As a point of emphamis, although a water effect ratio affects permit ltiitm 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.

This interim guidance explains and clarifies the use of site specific criteria. It is issued as interim guidance because it will be included am part of the process unduwaf for review and pOSSibl8 revision of the national aquatic life criteria development methodology guidelines. Am 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 am Wfinal. W EPA is interested in and encourages the submittal of high quality datamets that can be used to provide inmights 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). 6 mara is attached to this memorandum a simplified fAow diagram and an inplementation procedure. Thaw arm intmded to aid a user by placing the water affect ratio procadura in tha contmct of procmding from at 8it8 8pocific criterion to a permit limit. Following those attachment8 is tha guidance itself.

cc: Robert Fuciawp8, Ow l4aehaG. Prothro, o( I William Diamond, SASD Hargarat Stasikowski, HECD Hike Cook, OWEC Cynthia Dougherty, OWE Lw Schroer, OCC Swan kpow, OGC Courtney Riordan, ORD ORD ( Duluth and Narragansett Laboratories) ESD Diroctorm, Region8 I  VIII, X ESDBranch, Region IX Wat8r Quality Standards Coordinator8,  ion8 I  x WER Implementation Site Definition Study Plan Development Sampling Design Effluent Considerations Receiving Water Considerations Lab Procedures Testing Organisms Site Specific Criteria Permit Limits Monitoring Requirements WATER EFFECT RATIO IMPLEMENTATION PRELIMINARY ANALYSIS 41 PLAN FORMULATION  Site definition l How my discharges must be accounted for? Tributuies?

sea page 17.

l What is tha watarbody typa? ( i. e., stream, tidal rivar, bay, etc.). Sea page 44 and Appendix A.

l How can tha8e considerations best ba combined to defina the ralavant geographic % ite"? Sea Appendix A e page 02.

 Plan Davalopmant for Regulatory Agency Review l Is wm wthod 1 or 2 appropriate? ( e. g., Is daoign flaw a meaningful concept or ara other considerations paramount?). Saa page 6.

l Define the effluent & racaiving water 666pla location8 l Dascribe tha temporal 8ampla collection protocol8 propmad. See page 48.

l Cm simulated 8ita watar procadura ba done, or is down6treamr mampling required? See Appendix A.

l De6criba the testing protocol8  test opecias, te8t type, ta8t length, etc. Sea page 45, 50; Appendix I.

l Describa the chemical testing proposed. See Appendix C.

l Describa other details of 8tudy  flow measuramant, QA/ QC, numbar of 8ampling pariods proposed, to whom the results are expactad to apply, schedule, etc.

SANPXZNG DESIGN PORSTREAMS  Discuss the quantification of the design streamflow ( a g., 7410)  USGS gage directly, by extrapolation from USGS gage, or 3  Effluents l measure flows to datarmina average for sampling day l collect 24 hour composite using % leanW equipment and appropriata procadura6; avoid the use of the plant's daily composite sample as a ohortcut.

 Streams l measure flow ( use currant mater or read from gage if available) to determine dilution with affluent: and to check if within accaptabla range for use of the data ( i. e., design flow to 10 times the design flow).

l collect 24 hour composite of upstream water. LABORATORY PROCEDURES ( NOTE: mesa are described in detail in interim guidance).

 Select appropriate primmy C secondary tests  Determine appropriate cmcWER and/ or cccWER  tifOXBChUiS~ U8ingCl. p~, WithMthod8 that have adaquata SaMitivity to maasura low concbntration8, and u8a appropriate QA/ QC  Calculate final water affect ratio ( FWER) for sita. Sea page 36.

IMPLEHENTATION  Assign PWER8 and the site 8pecific criteria for oath metal to each di8ChUgU ( if more than 0~).

 parform a waste load allocation and total maximm daily load ( if appropriate) 80 that math discharger is provided l pumitlimit.

 l 8tabli8h monitoring condition for periodic evaluation of irmtroam biology ( raconandad)

 l 8tabli8h a permit condition for pariodic tasting of WER to verify site spacific criterion ( NTR recommendation)

2 United States Office of Water February 1994 Environmental Protection Office of Science & Technology EPA 823 B 94 001 Agency ( Mail Code 4305)

EPA Interim Guidance on Determination and Use of Water Effect Ratios for Metals 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 NOTICES This document has been reviewed by the Environmental Research Laboratories, Duluth, MN and Narragansett, RX ( 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.

ii 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 do comment 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).

iii UNITED STATES ENVIRONMENTAL PROTECTION AGENCY WASHINGTON, D. C. 20460 FEB 22 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 method by which site specific criteria could be developed.

The interim guidance provided in this document supersedes all guidance concerning water effect ratios and the Indicator Species Procedure given in Chapter 4 of the Water Quality Standards Handbook issued by EPA in 1983 and in Guidelines for Deriving Numerical Aquatic Site Specific Water Quality Criteria by Modifying National Criteria, 1984. Appendix B also supersedes the guidance in these earlier documents for the Recalculation Procedure for performing site specific criteria modifications.

This interim guidance fulfills a commitment made in the final rule to establish numeric criteria for priority toxic pollutants ( 57 FR 60848, December 22, 1992, also known as the " National Toxics Rule"). This guidance also is applicable to pollutants other than metals with appropriate modifications, principally to chemical analyses.

Except for the jurisdictions subject to the aquatic life criteria in the national toxics rule, water affect ratios are site specific criteria subject to review and approval by the appropriate EPA Regional Administrator. Site specific criteria are new or revised criteria subject to the normal EPA review requirements established in Clean Water Act § 303( c). For the States in the National Toxics Rule, EPA has established that site specific water affect ratios may be applied to the criteria promulgated in the rule to establish Site specific criteria. The water affect ratio portion of these criteria would still be subject to State review before the development of total maximum daily loads, waste load allocations or translation into NPDES permit limits. EPA would only review these water affect ratios during its oversight review of these State programs or review of State issued permits.

IV 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 wing option 3 with the use of simulated downstream water for a situation where this simulation appropriately represents a " site", is a fully acceptable approach for deriving a water effect ratio for use in determining a site specific criterion, although it will generally provide a lower water effect ratio than the other 2 options.

As indicated in the introduction to this guidance, the determination of a water effect ratio may require substantial resources. A discharger should consider cost effective, preliminary measures described in this guidance ( e. g., use of " clean" sampling and chemical analytical techniques or in non NTR States, a recalculated criterion) to determine if an indicator species site specific criterion is really needed. It may be that an appropriate site specific criterion is actually being attained. In many instances, use of these other measures may eliminate the need for deriving final water effect ratios. The methods described in this interim guidance should be sufficient to develop site specific criteria that resolve concerns of dischargers when there appears to be no instream toxicity from a metal but, where ( a) a discharge appears to exceed existing or proposed water quality based permit limits, or ( b) an instream concentration appears to exceed an existing or proposed water quality criterion.

This guidance describes different methods for determining water effect ratios. Method 1 has 3 options each of which may only require 3 sampling periods. However options 1 and 2 may be expanded and require a much greater effort. While this position statement has discussed the simplest, least expensive option for method 1 ( the single discharge to a stream) to illustrate that site specific criteria are feasible even when only small dischargers are affected, water effect ratios may be calculated using any of the other options described in the guidance if the State/ discharger believe that there is reason to expect that a more accurate site specific criterion will result from the increased cost and complexity inherent in conducting the V additional to& s and analyzing the results. Situation8 where this could be the ca80 include, for rxample, whore 8ea8onal l ffect8 in receiving water quality or in di8Chargm quality need t0 be a88e88ed.

In addition, EPA will cormider Other 8cientifiWlly dUfUn8ibla approach88 in developing final Water° ffUt ratio8 a8 authorized in 40 CFR 131.11. Howevu, EPA 8trongly recomund8 that beforo a State/ di8Charger impleDent8 aXby approach othu than one described in thi8 interim guidance, di8cu88ion8 be held with appropriate EPA regional offices and Office of Rm8earch and Development' 8 8cienti8t8 befor actual testing begins. The8e di8cu88ion8 would be to enare that time and ruource8 ar8 not wa8ted on 8cientifiWlly and technically unaocoptable approachu. It remain8 tpA's re8ponsibility to Bake final deCi8ion8 on tha 8cientific and technical validity of altenativo approaches to dovaloping 8itW8pOCifiC WatU @ ldllity UitUia.

EPA i8 fully cognizant of the continuing d& ate bew88n vhat coxmtitutu guidance and what i8 a regulatory reguirnmt. Developing 8itw8puific critrria i8 a Stat8 regulatory option. U8ing the methodology correctly a8 de8cribed in thi8 guidance a88Urm8 the State that EPA will accept the re8ult. Othu approaches are possible and logically 8hould be discussed with EPA prior to implmaentation.

The Offiw of Science and Technology beliovu that thi8 interim guidance advance8 the 8cionw of determining 8ite 8pMifiC Criteria and prOVidm8 policy gui& nW that Stat88 and EPA Wn u8e in thi8 complex aru. It refloct8 the 8cientific adVanCa8 in the past 10 yoar8 and the mcperienc8 gained from dealing with these i88ue8 in rul world situatiom. Thi8 guidance will help improve implementation of water quality 8tandard8 and be the basis for future progreW.

Office of Science And Technology Office of Water Vi CONTENTS Notices......................

Foreword . . . . . . . . . . . . . . . . . . . . .

Office of Science and Technology Position Statement Appendices . . . .

Figures . . . . . .

Acknowledgments . .

Executive Summary .

Abbreviations . . .

Glossary . . . . .

Preface . . . . . .

Introduction . . .

Method 1 . . . . . A. B. C. D. E. F. G. H. I. J. Experimental Design . Background Information and Initial Decisions Selecting Primary and Secondary Tests . . . . Acquiring and Acclimating Test Organisms . . Collecting and Handling Upstream Water and Effluent . Laboratory Dilution Water. . . . . . . . Conducting Tests Chemical and Other Measurements . . . . Calculating and Interpreting the Results Reporting the Results . . . . . . . . . . . .

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. xiv xvi . 1 17 17 44 45 47 48 49 50 55 57 62 65 76 vii APPENDICES Page A.

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J. Comparison of WERs Determined Using Upstream and Downstream Water . . . . . . . . . . . . . . . . . . . . 79 The Recalculation Procedure . . . . . . . . . . . . . . . 90 Guidance Concerning the Use of " Clean Techniques" and QA/ QC when Measuring Trace Metals . . . . . . . . . . . . 98 Relationships between WERs and the Chemistry and Toxicology of Metals . . . . . . . . . . . . . . . . . . 109 U. S. EPA Aquatic Life Criteria Documents for Metals . . . 134 Considerations Concerning Multiple Metal, Multiple Discharge, and Special Flowing Water Situations . . . . . 135 Additivity and the Two Components of a WER Determined Using Downstream Water . . . . . . . . . . . . . . . . . 139 Special Considerations Concerning the Determination of WERs with Saltwater Species . . . . . . . . . . . . . 145 Suggested Toxicity Tests for Determining WERs for Metals . . . . . . . . . . . . . . . . . . . . . . . 147 Recommended Salts of Metals . . . . . . . . . . . . . . . 153 viii FIGURES Page 1.

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D6. Four Ways to Derive a Permit Limit . . . . . . . . . . . 16 Calculating an Adjusted Geometric Mean . . . . . . . . . 71 An Example Derivation of a FWER . . . . . . . . . . . . . 72 Reducing the Impact of Experimental Variation . . . . . . 73 Calculating an LC50 ( or EC50) by Interpolation . . . . . 74 Calculating a Time Weighted Average . . . . . . . . . . . 75 An Example of the Deletion Process Using Three Phyla . . 97 A Scheme for Classifying Forms of Metal in Water . . . . 111 An Example of the Empirical Extrapolation Process . . . . 125 The Internal Consistency of the Two Approaches . . . . . 126 The Application of the Two Approaches . . . . . . . . . . 128 A Generalized Complexation Curve . . . . . . . . . . . . 131 A Generalized Precipitation Curve . . . . . . . . . . . . 132 ix ACKNOWLEDGEMENTS This document was written by:

Charles E. Stephan, U. S. EPA, ORD, Environmental Research Laboratory, Duluth, MN.

William H. Peltier, U. S. EPA, Region IV, Environmental Services Division, Athens, GA.

David J. Hansen, U. S. EPA, ORD, Environmental Research Laboratory, Narragansett, RI.

Charles G. Delos, U. S. EPA, Office of Water, Health and Ecological Criteria Division, Washington, DC.

Guy 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 Ballantine, 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.

x EXECUTIVE SUMMARY A variety of physical and chemical characteristics of both the water and the metal can influence the toxicity of a metal to aquatic organisms in a surface water. When a site specific aquatic life criterion is derived for a metal, an adjustment procedure based on the toxicological determination of a water effect ratio ( WER) may be used to account for a difference between the toxicity of the metal in laboratory dilution water and its toxicity in the water at the site. If there is a difference in toxicity and it is not taken into account, the aquatic life criterion for the body of water will be more or less protective than intended by EPA's Guidelines for Deriving Numerical National Water Quality Criteria for the Protection of Aquatic Organisms and Their Uses. After a WER is determined for a site, a site specific aquatic life criterion can be calculated by multiplying an appropriate national, state, or recalculated criterion by the WER. Most WERs are expected to be equal to or greater than 1.0, but some might be less than 1.0. because most aquatic life criteria consist of two numbers, i. e., a Criterion maximum Concentration ( CMC) and a Criterion Continuous Concentration ( CCC), either a cmcWER or a cccWER or both might be needed for a site. The cmcWER and the cccWER cannot be assumed to be equal, but it is not always necessary to determine both.

In order to determine a WER, side by side toxicity tests are performed to measure the toxicity of the metal in two dilution waters. One of the waters has to be a water that would be acceptable for use in laboratory toxicity tests conducted for the derivation of national water quality criteria for aquatic life. In most situations, the second dilution water will be a simulated downstream water that is prepared by mixing upstream water and effluent in an appropriate ratio; in other situations, the second dilution water will be a sample of the actual site water to which the site specific criterion is to apply. The WER is calculated by dividing the endpoint obtained in the site water by the endpoint obtained in the laboratory dilution water. A WER should be determined using a toxicity test whose endpoint is close to, but not lower than, the CMC and/ or CCC that is to be adjusted.

A total recoverable WER can be determined if the metal in both of the side by side toxicity tests is analyzed using the total recoverable measurement, and a dissolved WER can be determined if the metal is analyzed in both tests using the dissolved measurement. Thus four WERs can be determined: Total recoverable cmcWER. Total recoverable cccWER. Dissolved cmcWER. Dissolved cccWER. A total recoverable WER is used to calculate a total recoverable site specific criterion from a total recoverable national, state,

xi or recalculated aquatic life criterion, whereas a dissolved WER is used to calculate a dissolved site specific criterion from a dissolved criterion. WERs are determined individually for each metal at each site; WERs cannot be extrapolated from one metal to another, one effluent to another, or one site water to another.

Because determining a WER requires substantial resources, the desirability of obtaining a WER should be carefully evaluated: 1. Determine whether use of " clean techniques" for collecting, handling, storing, preparing, and analyzing samples will eliminate the reason for considering determination of a WER, because existing data concerning concentrations of metals in effluents and surface waters might be erroneously high. 2. Evaluate the potential for reducing the discharge of the metal. 3. Investigate possible constraints on the permit limits, such as antibacksliding and antidegradation requirements and human health and wildlife criteria. 4. Consider use of the Recalculation Procedure. 5. Evaluate the cost effectiveness of determining a WER. If the determination of a WER is desirable, a detailed workplan for should be submitted to the appropriate regulatory authority ( and possibly to the Water Management Division of the EPA Regional Office) for comment. After the workplan is completed, the initial phase should be implemented, the data should be evaluated, and the workplan should be revised if appropriate.

Two methods are wed to determine WERs. Method 1, which is used to determine cccWERs that apply near plumes and to determine all cmcWRRs, uses data concerning three or shore 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 ABBREVIATIONS ACR: Acute Chronic Ratio CCC : Criterion Continuous Concentration CMC: Criterion Maximum Concentration CRM: Certified Reference Material FAV : Final Acute Value FCV : Final Chronic Value FW: Freshwater FWER: Final Water Effect Ratio GMAV: Genus Mean Acute Value HCME: Highest Concentration of the Metal in the Effluent MDR: Minimum Data Requirement NTR: National Toxics Rule QA/ QC: Quality Assurance/ Quality Control SMAV: Species Mean Acute Value SW: Saltwater TDS: Total Dissolved Solids TIE: Toxicity Identification Evaluation TMDL: Total Maximum Daily Load 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 GLOSSARY Acute chronic ratio  an appropriate measure of the acute toxicity of a material divided by an appropriate measure of the chronic toxicity of the same material under the same conditions.

Appropriate regulatory authority  Usually the State water pollution control agency, even for States under the National Toxics Rule; if, however, a State were to waive its section 401 authority, the Water Management Division of the EPA Regional Office would become the appropriate regulatory authority.

Clean techniques  a set of procedures designed to prevent contamination of samples so that concentrations of trace metals can be measured accurately and precisely.

Critical species  a species that is commercially or recreationally important at the site, a species that exists at the site and is listed as threatened or endangered under section 4 of the Endangered Species Act, or a species for which there is evidence that the loss of the species from the site is likely to cause an unacceptable impact on a commercially or recreationally important species, a threatened or endangered species, the abundances of a variety of other species, or the structure or function of the community.

Design flow  the flow used for steady state wasteload allocation modeling.

Dissolved metal  defined here as " metal that passes through either a 0.45 µ m or a 0.40 µ m 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 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) teats 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 PREFACE Several issues need consideration when guidance such as this is written:

1. Degrees of importance: Procedures and methods are series of instructions, than others. but some of the instructions are more important Some instructions are so important that, if they are not followed, the results will be questionable or unacceptable; other instructions are less important, but definitely desirable. Possibly the best way to express various degrees of importance is the approach described in several ASTM Standards, such as in section 3.6 of Standard E729 ( ASTM 1993a), which is modified here to apply to WERs: The words " must", " should", " may", " can", and " might" have specific meanings in this document. " Must" is used to express an instruction that is to be followed, unless a site specific consideration requires a deviation, and is used only in connection with instructions that directly relate to the validity Of toxicity tests, WERs, FWERS, and the Recalculation Procedure. " Should" is used to state instructions that are recommended and are to be followed if reasonably possible. Deviation from one " should" will not invalidate a WER, but deviation from several probably will. Terms such as " is desirable", " is often desirable", and " might be desirable" are used in connection with less important instructions. " May" is used to mean " is ( are) allowed to", " can" is used to mean " is ( are) able to", and " might" is used to mean " could possibly". Thus the classic distinction between " may" and " can" is preserved, and " might" is not used as a synonym for either " may" or " can" This does not eliminate all problems concerning the degree of importance, however. For example, a small deviation from a " must" might not invalidate a WER, whereas a large deviation would. ( Each " must" and " must not" is in bold print for convenience, not for emphasis, in this document.)

2. Educational and explanatory material: Many people have asked for much detail in this document to ensure that as many WERs as possible are determined in an acceptable manner. In addition, some people want justifications for each detail. Much of the detail that is desired by sow 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 3.

4. Separation of " science", best professional judgment" and " regulatory decisions": These can never be completely"

5.

6. 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 CM then select one based on cost effectiveness, personal preference, details of the particular situation, and perceived advantages and disadvantages.

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.

Best professional judgement : 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".

Personal bias Bias can never be eliminated, and some decisions are at the fine line between " bias" and " best xvii 7. professional judgment". The possibility of bias CM 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.

xviii 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 form 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 diluting water because, for example, the site water has a lower pH and/ or hardness than most laboratory waters. Thus although a national aquatic life criterion is intended to be lower than necessary for most sites, a national criterion might not adequately protect the aquatic life at some sites.

Because a national aquatic life criterion might be more or less protective than intended for the aquatic life in most bodies of water, the U. S. EPA provided guidance ( U. S. EPA 1983a, 1984) concerning three procedures that may be used to derive a site specific criterion: 1. The Recalculation Procedure is intended to take into account relevant differences between the sensitivities of the aquatic organisms in the national dataset and the sensitivities of organisms that occur at the site. 2. The Indicator Species Procedure provides for the use of a water effect ratio ( WER) that is intended to take into account relevant differences between the toxicity of the metal in laboratory dilution water and in site water. 3. The Resident Species Procedure is intended to take into account both kinds of differences simultaneously. A site specific criterion is intended to come closer than the national criterion to providing the intended level of protection to the aquatic life at the site, usually by taking into account the biological and/ or chemical conditions ( i. e., the species composition and/ or water quality characteristics) at the site. The fact that the U. S. EPA has made these procedures available should not be interpreted as implying that the agency advocates that states derive site specific criteria before setting state 1 standards. Also, derivation of a site specific criterion does not change the intended level of protection of the aquatic life at the site. Because a WER is expected to appropriately take into account ( a) the site specific toxicity of the metal, and ( b) synergism, antagonism, and additivity with other constituents of the site water, using a WER is more likely to provide the intended level of protection than not using a WER.

Although guidance concerning site specific criteria has been available since 1983 ( U. S. EPA 1983a, 1984), interest has increased in recent years as states have devoted more attention to chemical specific water quality criteria for aquatic life. In addition, interest in water effect ratios ( WERs) increased when the " Interim Guidance" concerning metals ( U. S. EPA 1992) made a fundamental change in the way that WERs are experimentally determined ( see Appendix A), because the change is expected to substantially increase the magnitude of many WERs. Interest was further focused on WERs when they were integrated into some of the aquatic life criteria for metals that were promulgated by the National Toxics Rule ( 57 FR 60848, December 22, 1992). The newest guidance issued by the U. S. EPA ( Prothro 1993) concerning aquatic life criteria for metals affected the determination and use of WERs only insofar as it affected the use of total recoverable and dissolved criteria.

The early guidance concerning WERs ( U. S. EPA 1983a, 1984) contained few details and needs revision, especially to take into account newer guidance concerning metals ( U. S. EPA 1992; Prothro 1993). The guidance presented herein supersedes all guidance concerning WERs and the Indicator Species Procedure given in Chapter 4 of the Water Quality Standards Handbook ( U. S. EPA 1983a) and in U. S. EPA ( 1984). All guidance presented in U. S. EPA ( 1992) is superseded by that presented by Prothro ( 1993) and by this document. Metals are specifically addressed herein because of the National Toxics Rule NTR) and because of current interest in aquatic life criteria for metals; although most of this guidance also applies to other pollutants, same 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 2 composition and water quality characteristics in the Louether. the Additional reasons for addressing both 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 WRRs may also be used with site specific aquatic life criteria that are derived for metals using the Recalculation Procedure described in Appendix B. m obtained For exatrlple, 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 bassd 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, chromiuxn(~~~), chromiumWI), copper, lead, nickel, silver, and zinc and with the Criterion Maximum Concsntration ( CMC) for mercury. WERs may also be used with saltwater criteria for selenium.

The concept of a WER is rather simple: YBvo 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 ret omnended 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.

3 Because a national Criterion is intended toqrotect aquatic life in alsmst all bodies of water and because a WRR is intended to account for a difference between the toxicity of a metal in a laboratory dilution water Md its toxicity in a site water, dischargers who want higher permit limits than those derived on the basis of M existing aquatic life criterion will probably consider determining a WRR. 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, sane will be about 1.0 Md some will be less than 1.0. b. The detemaination of a WER requires substantial resources. c. There are other things that a discharger CM do that might be mOre cost effective than detenaining 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) M instresm concentration appears to axceed M existing or proposed aquatic life criterion. Such situations result from xneasurernent of the concentration of a metal in M effluent or a surface water. It would therefore seem reasonable to ensue that such measurements were not subject to contamination. Usually it is much easier to verify chemical measurements by using l clean techniques' for collecting, handling, storing, preparing, and analyzing soles, than to determine a WER. Clean techniques and some related QA/ QC considerations are dis~ ssed 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 Mtibacksliding and Mtidegradation requirements Md human health and wildlife criteria. 3. Obtain assistance from M aquatic toxicologist who understands the basics of WERs ( see Appendix D), the U. S. EPA's national aquatic life guidelines ( U. S. EPA 19851, the guidance presented by Prothro ( 19931, 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 Md 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 M 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 detemination of a WER. Comparative toxicity tests provide the most useful data, but chemical analysis of the downstream water might be helpful 4 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 ( Tot) 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 soms metals in water, at the present only toxicity tests CM 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 Managemsnt Division of the EPA Regional Office) for coscnent. 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 wlexnented; then the data obtained should bs 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.

Most aquatic life criteria contain both a a4C and a CCC, and it is usually possible to determine both a cmcWRR anda cccWER. The two WERs cannot be assumsd to be equal because the magnitude of a WER will probably depsnd on the sensitivity of the toxicity test ussd and on the percsnt effluent in the site water ( see Appendix D), both of which CM depend on which WER is to be determined. In some cases, it is expected that a larger WER CM be applied to the CCC than to the CMC, and so it would be environmentally conservative to apply ancWERs 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 CXC, a site specifik CCC, and new permit limits. If these new pexmit 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 bs applied to the CCC than to the CRC and so it might not be environmentally conservative to apply a cmcWER to a CCC ( see section A. 4 of Method 1).

5 Some of the guidance contained herein specifically applies to situations in which the pemit 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 smdeling, the guidance will have to be modified, but it is unclear at present what modifications are most appropriate. For wle, 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.

Wo methods are used to determine WERs. Method 1 will probably be used to determine all c! mcWERs and most CccWERs because it CM 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 swle obtained in the ismediate vicinity of a plums is unknown, simulated downstream water is used so that the percent effluent in the sample is known. For exas@ le, if a ssmgle that was supposed to represent a cwlete mix situation was accidently taken in the plume upstream of comnplete mix, the sanlple would probably have a higher percent effluent and a higher WI& R than a sas@ e taken downstream of coa@ ete 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 cmlete mix but outside the plume, overprotection would probably result.

Method 1 will probably be used to determine all cmcWERs and mo8t cccWERs in flowing fresh waters, such as rivers and streams. Method 1 i8 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 mane state and/ or federal agencies refer to as l effluent dependentm, % abitat creating., or ' effluent dominated'. Method 1 is also used to determine cmcWERs in such. large sites as oceans and large lakes, rebendrb, and estuaries ( see Appendix F).

Method 2 is used to determine WEF4s that apply outside the area of plumes in large bodies of water. Such = will be cccWERs and will be determined using sas@ es 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 detenained 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.

6 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 MC) only one discharge has permit limits for that metal. ( This is the single metal single discharge situation.) b. A more complex situation is when a WER is to be determined for only one metal, but more than one discharge has permit limits for that metal. ( This is the single metal multiple discharge situation.) c. An even mOre complex situation is when WEFU 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.) WRRs need to be determined for each metal at each site because extrapolation of a WRR from one metal to another, one effluent to another, or one surface water to another is too uncertain.

Roth methods work well in multiple metal situations, but special tests or additional tests will be necessary to show that the resulting combination of site specific criteria will not be too toxic. Method 2 is better suited to multiple discharge situations than is Method 1. Appendix F provides additional guidance concerning multiple metal and multiple discharge situations, but it does not discuss allocation of waste loads, which is performed when a wasteload allocation ( WLA) or a total maximum daily load ( TMDL) is developed ( U. S. EPA 1991a).

A total recoverable WER CM 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 WRR 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.

Roth Method 1 and Method 2 CM 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 7 test solutions. Both total recoverable and dissolved measurements are to bs 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 wasuremsnts also increases the alternatives available for use of the results. For vie, a dissolved WER that is not useful with a total recoverable criterion might be useful in the future if a dissolved criterion becomss available. Also, as explained in Appendix D, except for experimsntal variation, use of a total recoverable WER with a total recoverable criterion should produce the same total recoverable permit limits as use of 8 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 WEF& will be larger and smre variable than dissolved WERs.

Traditionally, for practical reasons, the requirssmnts concerning such aspects as acclimation of test organisms to test tsnperature and dilution water have not been as stringsnt 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 isportant for a WER as long as the organisms and conditions are idsntical in the two tests. Because WERs are used to adjust aquatic life criteria that are derivsd fran results of laboratory tests, the tests conducted in laboratory dilution water for the detexmination 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 bstween national criteria and site specific criteria, and so it is isportant to c are at least some results obtained in the laboratory dilution water with results obtained in at least one other laboratory.

Three important principles for staking 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 msed 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. The xmst 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 WRRs use a site water in which the concentration of metal probably is not xiegligible. Recause 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, WERa are based on too few data to 88surne that ncaainal 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 amasured. Further, measuring the concentrations might lead to an interpretation of the results that allows a substantially better use of the WE& S.

The appropriate regulatory authority might recoxmen d that one or more conditions be met when a WER is determined in order to reduce the mssibilitv of havina to determine a new WER later: 1.

2.

3. Requireahnts that  me in the existing permit concerning VET testing, Toxicity Identification Evaluation ( TIE), and/ or Toxicity Reduction Rvaluation ( TRR) ( U. S. EPA 1991a). Implewntation of pollution prevention efforts, such as pretreatment, waste minimization, and source reduction. A dewmstration that applicable technology based requirements are being met.

Even if all re caamsnded 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 desired to determine a WRR 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,

9 alkalinity, and concsntrationa of TSS and TOC should be within 10 8 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 QllcwER even if a CCCWER cannot be determined; a cmcWER might not be useful, however, if the petit limits are controlled by the CCC; in such a case, it would probably not be acceptable to assume that the cxncWER is an environmentally consenmtive estimate of the cc. If the WER is detenninsd using downstream water and the toxicity is due to the efflusnt, tests at lower concentrations of the efflusnt might give an indication of the amount of remsdiation needed.

Besides requiring that the WER be valid, the appropriate regulatory authority might consider inposing other conditions for the approval of a site specific criterion based on the WER: 1. Periodic reevaluation of the WER. a. WERs detenninsd in upstream water take into account constituents contributed by point and nonpoint sources and natural runoff; thus a WER should be reevaluated whenever newly implementsd controls or other changes substantially affect such factors as hardness, alkalinity, @ I, 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 WEZs should be reevaluated whenever newly imglewnted control& or other changes might substantially iqmct the effluent, i. e., might inpact the forms and concentrations of the metal, hardness, alkalinity, pH, suspended solids, organic car& n, or other toxic materials. A special concern is the possibility of a shift fraa discharge of nontoxic metal to discharge of toxic mstal such that the concentration of the metal does not increase; analytical chemistry might not detect the change but toxicity tests would. Evsn if no changes are known to have occurred, WBRs should be reevaluated periodically. ( The NTR recmmmAl that NPDES permits include periodic dete nninations of WERs in the monitoring requiremsnts.) With aduance planning, it should usually bs 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 m& al 10 ( e. g., hardness, alkalinity, pH, Tot, 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, Tot, and/ or TSS decrease below the values that existed when the WERs were determined. 3. Periodic reevaluation of the environmental fate of the metal in the effluent ( see Appendix A). 4. WET testing. 5. Instream bioassessments.

Decisions concerning the possible imposition of such conditions should take into account: a. The ratio of the new and old criteria. The greater the increase in the criterion, the more concern there should be about ( 1) the fate of any nontoxic metal that contributes to the WER and ( 2) changes in water quality that might occur within the site. The imposition of one or more conditions should be considered if the WER is used to raise the criterion by, for axample, 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 F'AV to the CMC, and the range of sensitivities of species in the criteria do cument for the metal ( see Appendix E). b. The size of the site. c. The size of the discharge. d. The rate of downstream dilution. 8. Whether the CMC or the CCC controls the pennit 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 bs larger and any decrease in the discharge of such Tot 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 tims in the future.

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 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 sams or a different body of water that is outside the site; nothing contained herein changes this policy in anyway. If no increase in the existing discharge is allowed, the only use of a WgR will be to determine whether an existing discharge needs tobe reduced. Thus a major use of W might be where technology based controls allow concentrations in surface waters to exceed national, state, or recalculated aquatic life criteria. In this case, it might only bs necessary to determine that the WKR is greater than a particular value; it might not be necessary to quantify them. When possible, it might be desirable to show that the wxinnsa WER is grsater than the WER that will be used in order to demonstrate that a margin of safety exists, but again it might not bs necessary to quantify the smximum 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 fran ambient monitoring should be interpreted by comparison with ambient criteria. ( This is not a problem in jurisdictions subject to the NTR because the NTP & fines the smbient criterion as WER x the EPA criterion'.) If a WER is used to adjust pennit limits without adjusting the criterion, the permit limits would allow the criterion to be exceeded. Thw the WRR shouldbs ussd to calculate a site specific criterion, which should then bs used to calculate pennit limits. In sane states, site specific criteria can only bs 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 bs adopted prior to NPDES permit issuance. Moreouer, the EPA Regional Administrator has authority to approve or disapprove all new and revised site specific criteria and to review NPDES permits to verify cwliance with the applicable water quality criteria. I Other aspects of the use of WERs in connection with permit limits, WLAs, and IMDLs are outside the scope of this document. The Technical Support Documen t ( U. S. EPA 1991s) and Prothro ( 1993) provide more information concerning Mlaaentation procedures. Nothing contained herein should be interpreted as changing the three part approach that EPA uses to protect aquatic life: ( 1) numeric chsmical specific water quality criteria for individual pollutants, ( 2) whole effluent toxicity WET) testing, and ( 3) instreaxn bioassessmsnts.

12 Even though there are similarities between WET testing and the determination of WERs, there are important differences. For example, WEFU CM 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 whsn 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 bioassessmsnts are likely to be recoxnended.

A major problem with the determination and use of aquatic life criteria for metals is that no analytical measurement or combination of xneasuremsnts has yet been shown to explain the toxicity of a metal to aquatic plants, invertebrates, amphibians, and fishes over the relevant range of Condition8 in surface waters ( see Appendix 0). 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 mle, the concentration of free metal ion could possibly bs a useful basis for expressing water quality criteria for metals if it could be feasible and could bs used in a way that does not result in widespread underprotection of aquatic life. Some available data, however, might contradict the idea that ths 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 mstals to aquatic life, it i8 al80 not easy to find a qualitative relationship that CM be used to provide adequate protection for the aquatic life in almost all bodies of water without providing as much overprotection for mans 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 measuremsnts 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 use of WERs to address the general conservatism, but expects that sosm WERs will bs less than l. O. 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 WBRs is not a smle solution to the existing general conservatism. It is likely that a pemanent solution will have to be'based on an adsquate quantitative sxplanation of how metals and a& tic 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 axcassive overprotection for too many bodies of water. Work needs to continue on the penaanent solution and, just in case, on inproved alternative approaches.

Use of WRRs to derive site specific criteria is intended to allow a reduction or elimination of the gsneral overprotection 8SSoCi8ted with application of a national Criterion to individual bodies of water, but a major problem is that a WER will rarely bs constant over time, location, and depth in a body of water due to plumsa, mixing, and resuspension. It is possible that dissolved concentration and WERs will be less variable than total recoverable ones. It might also be possible to reduce the impact of the heterogeneity if n 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 bs used in a conservative manner. Because of vwiation bstween bodies of water, national criteria are derived to bs environsmntally conservative for most bodies of water, whereas the WER procsdure, which is intended to reduce the general conServatiSm of national criteria, has to be conservative because of variation among WKF& within a body of ' water.

The consenratism introduced by variation among WEFU is due not to the concetpt 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 CM be measured, but toxicity cannot be smdelled adsquately using available chemical msasurexnsnts. Similarly, the currsnt way that WEE& are used dspends on a linkage betwssn 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 perfonaed when WERs are determined using Method 2, a different way of using the WERs could avoid some of the problans introduced by the dependence on analytical chemistry. * The l sanrple specific WER approach9 could consist of sampling a body of water at a number of locations, determining the WER for each sample, and 14 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 WEB obtained for that sample. Except for experimental variation, when the quotient for a smle is less than 1, the concentration of metal in that sample, is acceptable; when the quotient for a sample is greater than 1, the concentration of metal in that sample is too high. As a check, both the total recoverable measurement and the dissolved measuremsnt should be used because they should provide the sams answer if everything is done correctly and accurately. This approach CM also be used whenever Method 1 is used; although Method 1 is used with simulated downstream water, the sample specific WER approach CM 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 WEF& within a body of water is not a problesn. 3. It eliminates problsms 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 plumsa. 5. It works squally well in single discharge and multiple discharge situations. 6. It automatically accounts for synergism, antagonism, and additivity between toxicant8. This way of using WERs is equivalsnt to expressing the national criterion for a pollutant in terms of toxicity tests whose endpoints equal the CMC` and the CCC; if the site water causes less adverse effect than is defined to bs the sndpoint, 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 framewo rk 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 guotisnt 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 Pigum 1: Four Way8 to Dmrivo 8 Permit Limit 16 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 WEB, decisions concerning experimental design should be given careful attention and need to answer the following questions: 1. Should WERs be determined using upstream water, actual downstream water, and/ or simulated downstream water? 2. Should WERs be determined when the stream flow is equal to, higher than, and/ or lower than the design flow? 3. Which toxicity tests should be used? 4. Should a cmcWER or a cccWER or both be determined? 5. How should a FWER be derived? 6. For metals whose criteria are hardness dependent, at what hardness should WERs be determined? The answers to these questions should be based on the reason that WERs are determined, but the decisions should also take into account some practical considerations.

17 1. Should WERs be determined using upstream water, actual downstream water, and/ or simulated downstream water?

a. Upstream water provides the least complicated way of determining and using WERs because plumes, mixing zones, and effluent variability do not have to be taken into account. Use of upstream water provides the least useful WERs because it does not take into account the presence of the effluent, which is the source of the metal. It is easy to assume that upstream water will give smaller WERs than downstream water, but in some cases downstream water might give smaller WERS ( see Appendix G). Regardless of whether upstream water gives smaller or larger WERs, a WER should be determined using the water to which the site specific criterion is to apply ( see Appendix Al.

b. Actual downstream water might seem to be the most pertinent water to use when WERs are determined, but whether this is true depends on what use is to be made of the WERs. WERs determined using actual downstream water CM be quantitatively interpreted using the sample specific WER approach described at the end of the Introduction. If, however, it is desired to understand the quantitative implications of a WER for an effluent of concern, use of actual downstream water is problematic because the concentration of effluent in the water CM 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 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 p cross section. of the stream. Regardless of where it is decided conceptually that a 8ample 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 i8 not known exactly what the sample actually repre8ent8, there is no way to know how reproducible the 8 18 i8. 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 sdmple specific WER approach is used to interpret the results.

c.  ated dwtrem would 8eem to be the most unnatural of the three kinds of water, but it offer8 several important advantage8 because effluent and upstream water are mixed at a known ratio. Thi8 i8 important because the magnitude of the WRR will often depend on the concentration of effluentSin the downstream water. Mixtures can be prepared to 8imulate the ratio of effluent and upstream water that uci8ts at the edge of the acute mixing zone, at the edge of the chronic mixing zone, at coanplete mix, or at any other point of interest. If desired, a 8ample of effluent can be mixed with a smle on upstream water in different ratios to simulate different point8 in a stream. Also, the ratio. used can be one that rimulates condition8 at design flow or at any other flow.

The sample specific WRR approach can be u8ed with both actual and simulated downstream water. Additional quantitative use8 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, Llimulated downstream water can be used to determine the variation in the WER that is due to variation in the effluent. 0 It also allows comparison of two or more effluents and determination of the interactions of two or mOre effluents. Additivity of WRRs can be studied using simulated downstream water ( see Appendix G); 8tudie8 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 make8 it possible to conduct controlled studies of changes in WERs due to aging and changes in pH.

19 2. In Method 1, therefore, WERs are detexmined using simulated downstream water that is prepared by mixing 8anples of effluent and upstream water in an appropriate ratio.  8t inlportaIltly, Method 1 Can be USad t0 determine a WER that applies in the vicinity of a plum and can be quantitatively extrapolated to the effluent.

Should m be determined when the 8trem flow is equal to, higher than, and/ or lower than the design flow?

WBRS are used in the derivation of 8ite 8pecific criteria when it i8 de8iredthatpennit limit8 be basedon a crituion that takes into account the characteristics of the watu and/ or the metal at the site. In mO8t cases, permit limit8 are calculated using 8teady state models and are ba8d on a design flOWa It is therefore important that WERE he adequately protective under design flow condition, which might be  acted to require that some set8 of 8~ aph8 of l ffluent and uprrtream water he obtained when the actual stream flaw is clo8e to the derrign flow. Collecting 8amgle8 when the stream flow is clo8e to the de8ign flow will limit a WRR determination to the low flaw 8ea8on ( e. g., from mid July to mid October in 8ome placea) and to year8 in which the flow ia sufficiently low.

It 58 al80  rtMt, hmever, that WRR8 that are applied at de8ign flow provide adequate protection at higher flow8. Generalization8 concerning the impact of higher fhw8 on WRR8 are difficult becau8e 8uch flow8 might ( a) reduce hardnesrr, alkalinity, and pH, ( b) increase or decrea8e the concentrations of ' 1Dc and TSS, ( c) reSU8pend toxic and/ Or ! mHx% iC metal from the sediment, and ( d) wa8h additional pollutant8 into the water. Acidic mmwmelt, for example, might lower the WRR both by diluting thm WER Md by reducing the hardness, alkalinity, and @ I; if substantial labile metal is pre8ent, the WRR might be lowered more than the concentration of the metal, po88ibly r88ulting in increased toxicity at flows higher than d88ign flow. Ssnples taken at higher flows might give  11~: WgRrr because the concentration of the effluent is more dilute; however, total recoverable WERs might be larger if the sample is taken just after M event that greatly increase8 the concentration of TSS and/ or TOC because thi8 might increase both ( 1) the concentration of nontoxic particulate metal in the water and ( 2) the capacity of the water to sorb and detoxify m& al.

WEREI are not of concern when the stream flow is lower than the de8ign 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 3. Which toxicity tests should be used?

a.

b.

C.

d.

e.

f. 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 teat 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 sams sensitivities, but the data also support the generalization that less sensitive toxicity tests usually give smaller WERs than more sensitive test8 ( see Appendix D). when the CCC is lower than the CMC, it is likely that a larger WE? I will result from tests that are sansitive at the CCC than frcm tests that are sensitive at the CMC. 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 X50, although the difference might not be measurable in most cases and the LC25 is likely to be nmre variable than the X50. Selecting the percent effect to bs used to define the endpoint might take into accotmt ( 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 LCSO with l day old fathead SliIlIlOWS~, the same percent effect aut be usad 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 uncWER and a CCCWER could be obtained with the same species. The primarv toxicity test used in the determination of a WER should have an andpoint in laboratory dilution water that is close to, but pot low= than, the CMC and/ or CCC to which the WEFI is to be applied. Because the endpoint of the prinmy 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 I as the endpoint of the primary test becomes closer to the CMC and/ or CCC ( see Appendix D). The WER obtained with the primary test should be confirmed with a # ecow test that uses a specie8 that is taxonomically different frost the species usad in the primary test. 1) The endpoint of the secondary test may be hiohcr Jowey than the CMC, the CCC, oi the endpoint of t% primary test.

21 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 aast mot be, wed as the primary and secondary tests, even if one enmint is lethal and the other is sublethal. g. 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 labor& tory dilution water equals the CncorcK. If such a test is not used, the maximum possible WKR probably will not be obtained. h. No rationale exists to support the idea that different species or tests with the dams sensitivity will produce different WBR8. Because the mode of action might differ fraa species to species and/ or from effect to effect, it is easy to speculate that in mans cases the magnitude of aWERwilldependto soms extentonthe 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. i. If the. tests are otherwise acceptable, both cmcWEBs and cccWERs msy be determined using acute and/ or chronic tests and using lethal and/ or sublethal endpoints. The wrtant consideration is the sensitivity of the test, not the duration, species, life stage, or adverse effect used. j. There is no reason to use species that occur at the 8ite; they may be used in the detexmination 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 sufficiant numbars for conducting toxicity tests over the testing period. 3) Additional QA tests will probably be needed ( see section C. 3. b) because data are not likely to be available from other laboratories for comparison with the results in laboratory dilution water. k. Because a WER is a ratio of results obtained with the sasm 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 b+ low the CMC and/ or CCC that is to be adjusted cannot be used as a primary test.

22 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 do cment for the metal of concern ( see Appendix El, 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 datemined. Different tests with similar sensitivities are expected to give similar WERs, except for experimental variation. The purpose of the secondary test is to provide information concerning this assumption and the validity of the WER.

4. Should a ascWER or a CCCWER or both be detexmined?

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 axamDle, a body of water that is protected for put and take fishing might have only a CMC, yhereas 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 CXC 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 pennit limits. A better way to answer the question would bs to also determine how much the controlling value would have to be changed for the other value to becoms 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 ( ssCCC1. 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 ancWER and a CCCWER, although both can be determined only if toxicity tests with appropriate sensitivities are available.

23 Because the phased approach can always be used, it is only, inportant. to decide whether to use a different approach when its use might be cost effectivea De& ding whether to use a different approach and selecting which one to use is cwlex 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 saam site water, but two WERs. could be determined if the mcWER and the CCCWER would be determined using different site waters, e. g., waters that contain different concentrations of the effluent. b. If the CM2 is higher than the CCC, is there a toxicity te8t 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 anda cccWER canbe determined. If the CE is higher than the CCC but no toxicity test has M endpoint in laboratory dilution water between the C24C and the CCC, two WERs cannot be determined if they would be detemained using the sams 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 pexmit limits? I& is cmlex, but reasonably clear, how to make a decision when a steady state model was used, but it is not clear how a decision should bemade when a dynamicmodelwas 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 msthod 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 CK design flow and a CCC design flow. 1 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 ba different from the percent effluent in the site water used in the determination of the CCCWER; thus the two WEFU should ba determined using two different site waters. This does not impact WEFT detemined using upstream water.

24 e. Is there M acute mixing zone? Is there a chronic mixing zone? 1. When WEEU are determined using upstream water, the Dresence or absence of mixing zones has no impact; the cxncWEZ4 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 M 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 axistence 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 froan a cwlete mix situation. ( If the biologically based method is wed, 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 cxncWEF! 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 oftan depend substantially on the percent effluent in the water ( see Appendix D). f. In what situations would it be environmentally consemative to determine one WER and use it to adjust both the cmcWER and the cccWER? Because ( 1) the CMC is never lower than the CCC and ( 2) a more sensitive test will generally give a WER closer to 1, it will be environmentally conservative to use a cmcWER to adjust a CCC when there are no contradicting considerations. In this case, a cmcWER CM be determined and used to adjust both the CMC and the CCC. Because water quality CM 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 Mother set of ixr@ lications when the two WERs are to be determined using different site waters, e. g., when the site waters contain different concentrations of effluent.

25 When WERS are determined using B water, the same site water is used in the determination of both the ancm and the CCCWER. Whenever the two WERs are determined in the sam site water, any difference in the magnitude of the cmcWER and the cccW'ER will probably be due to the sensitivities of the toxicity tests used. Therefore : a.

b.

c.

d.

e.

f.

Q* If more sensitive toxicity tests generally give larger WERS than less sensitive tests, the maximum CccWER ( a cccWER determined with a terrt whose endpoint equals the CCC). will usually be larger than the maximum cmcm because the CCC is never higher than the CMC. Because the CCC is nwer higher than the CMC, the maximum cmcWER will usually be smaller than the m ccc# ER and it will be environmentally conservative to use the cmcWER to adjust the CCC. A CCCWER CM be detemined 8apar8tely from a CmcwER only if there is a tOXiCity test with M endpoint in laboratory dilution water that i8 between the CMC and the CCC. If no such test exists or CM be dwised, only a cmcWER can be determined, but it CM be usad to adjust both the CE and the CCC. Unless the experimental variation is increased, use of a CCCWER, insteadof a awWRR,. to adjust the CCC till usually * rove the accuracy of the resulting site specific CCC. Thus a cccWER may be determined and used whenever desired, if a toxicity t88t has M endpoint in laboratory dilution water between the CMC and the CCC. A CCCWER cannot ba used to adjust a CMC if the cccWR3 was detexmined using M endpoint that was lower than the CE in laboratory dilution water because it will probably reduce the level of protection. Even if there is a toxicity test that has M endpoint in laboratory dilution water that is between the CMC and the CCC, it is not necessary to decide initially whether to determine a CSCWER and/ Or a CCCWER. When upstream water is used, it is always allowable to determine a QllcwER and use it to derive a site specific CMC and a site specific CCC and then decide whether to determine a ccc. 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 detemine a cccWER in addition to a cmcWEX. Further, if it is decided to determine a CCCW'ER, the sams two tests used in the detexmination of the QllcwER could then be used in the determination of the cccWER, with a rwersal 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 When WERs are determined using m water, the magnitude of each WER will probably depend on the concentration of effluent in the downstream water used ( see Appendix 9) l 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 detexmined using the same site water. If there is a chronic mixing zone, cmcWERs should be detemined in the site water that exists at the edge of the chronic mixing zone, whereas cccWERs should be determined for the cmlete mix situation ( see Appendix A). Thus the percent effluent will be higher in the site water wed in the determination of the cmcWER than in the site water used in the determination of the cccwH1. 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  and& CccWERcan 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 detemination of the cmcm will probably have a higher percent effluent than the site water used in the determination of the cccWEFt, 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 ba 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 wortant whether there is a toxicity test whose endpoint in laboratory dilution water is between the CMC and the CCC: a. If no toxicity test has an endpoint in laboratory dilution water between the CMC and the CCC, the two WERs have to be determined with the same test, in which case the ancWEF4 will probably be larger because the 27 percent effluent in the site water will be higher. Because of the difference in percent effluent in the site waters that should be wed in the determinations of the two WER8, use of the QllcwER to adjust the CCC would not be enviroxnnentally corrrrervative, but use of the cccWER to adjust the CWC would be environmentally coa8823ntive. Although both WER8 could be determined, it would also be acceptable to determine only the CCCWKR and u8e 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 detenained using different toxicity tests. An environwntally conservative alternative to deterxnining two WERs would be to determine a hybrid WKR by using ( 1) a toxicity test whose endpoint is above the Ct4C ( i. e., a toxicity test that ir appropriate for the determination of a CmcwER) and ( 2) 8ite water for the complete mix situation ( i. e., 8ite water appropriate for the determination of cccwm . Ituouldbe eavir onmentally conservative to use thir hybrid WER to adjust the CbC and it would be awiroaswntally consenmtive to use this hybrid WER to adjut the CCC. Although both WERs could be determined, it would also be acceptable to determine only the  rid WE! R and use it to adjust both the CMC andtheCCC ( This hybrid WER described here in paragraph b is the same as the CCCWER described in paragraph a above in which no toocicity test had M endpoint in laboratory dilution water between the CMC andtheCCC. 1 5. EowrhouldaPWERbe derived?

Because of  imental variation and variation in the ccSpo8ition of 8urface waters and effluent8, a single deterskination of a WER does not provide sufficient information to justify adjustment of a criterion. After a 8ufficient nulaber of WER8 have been determined in M acceptable manner, a Final Water Effect Ratio ( FWER) ial derived from the WERs, and the FWER is then used to calculate the 8ite specific criterion. If both a site specific CMC and a 8ite 8pecific CCC are to be derived, both a cmcFWER and a cccFWER have to be derived, unless M enviroxnnentally conservative estimate is used in place of the ancFWER and/ or the ccc.

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 20 variation. When a WE3 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 colm( position of the effluent, and ( d) variability in the percept effluent in the downstream water. Variability and the possibility of mistakes and rare events make it necessary to try to comprotnise betwaen ( 1) providing a high probability of adequate protection ahd ( 2) placing too much reliance 0" the smallest experimentally determined WER, which might reflect experimental variation, a mistake, or a rare went rather than a meaningful difference in the WER.

Various ways CM be employed to address variability: a. Replication CM be used to reduce the is@ act of soma 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 saxrgding times depending on such characteristics of the discharge as the average retention time, in case M unusual WEF4 is obtained with the first sample used. c. Because of the possibility of mistakes and rare wants, samples of effluent and upstream water should. be large enough that portions can be stored for later testing or analyses if M unusual WER is obtained. d. It might be possible to reduce the iqmct of the variability in the percent effluent in the downstream water by establishing a relationship between the WER and the percent effluent. Confounding of the sources CM be a problem when more than one source contributes substantial variability.

When pennit limits are calculated using a steady state model, the limits are based on a design flow, e. g., the 7410. It is usually assuwd that a concentration of metal in M effluent that does not cause unacceptable effects. at the design flow will not cause unacceptable effects at higher flows because the smtal is diluted By the increased flow of the upstream water. Dacreased protection might occur, however, if M 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 sufficiantly conservative that M increase in toxicity till not ba great enough to oven& elm the combination of dilution and the assumed conservatism, wen though it is likely that the national criterion is not overprotective of all bodies 29 of water. When WEIR? 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 ixqmrtant to ( 1) determine SOWB WERs thal: correspond to higher flows or ( 2) provide some coMenmtism. If the concentration of effluent in the downstream water decreases as flow increases, wets determined at higher flows are likely to be smaller than WERs deters& n ed at design flow but the concentration of nrstal will also be lower. If the concentration of TSS increases at high flcmm, however, both the WER and the concentration of as& al might increase. If they are determine& in an appropriate manner, WERs detexmined at flows higher than the design flow CM be used in two ways: a. As eavi ronmntally conservative e8timates of wER8 determined at design flow. b. m asaeas whether WERs determined at design flow will provide adequate protection at higher flows.

In order to appropriately take into account seaso~ l and flow effects and their interactions, both ways of using high flow WBRs require that the downstream water used in the detgmaination of theWKRbe similar to that which actually exist8 during the time of concern. In addition, high flow WERs can ba usad in the second way only if the composition of the downstream watar is known. To satisfy the requiresmnts that ( a) the downstream water used in the determination of a WgR be similar to the actual water and ( b) the cmposition 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 swles at the ratio of the flows of the effluent and the upstream water that existed when the s lea were obtained.

For the first way of using high flow WERs, they are used directly a8 eavi ronmentally consenmtive estimates of the design flow WER. For the second way of using high flow WER8, each ir u8ed to calculate the highest concentration of amtal 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 mcperimentally determined WER. This highest concentration of m&& l in the 8ffluMt UEME) CM be calculated as:

where: ccc = the national, state, or recalculated CCC ( or C24C) that is to be adjusted.

30 aFLOW P 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 existgd 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 doivnstream water. In order to calculate a HCME from M 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 CM 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:

The suffix ' dfm indicates that the values used for these quantities in the calculation of the hwER are those that exist at design flow conditions. The additional datum needed in order to calculate the hWER is the concentration of metal in upstream water at design flow conditions; if this is assumed to be zero, the hwER will be environmentally conservative. If a WER is determined when uFLOW equals the design flow, hWE3 = WER.

The two ways of using WERs determined at flows higher than design flow CM be illustrated using the following examples. These exsmplea were formulated using the concept of additivity of WERs ( see Appendix G). A WEB determined in downstream water consists of two components, one due to the effluent ( the eWER) and one due to the upstream water ( the uWER).. If the eWER and uWER are strictly additive, when WERs are determined at various upstream flows, the downstream WERs can be calculated from the composition of the downstream water ( the % effluent and the % upstream water) and the two WERs ( the eWER and the uWEX) using the equation:

31 wgp I 1% effhent) ( eWgR1 l (% umtrum utter) ( um8)

100 In the emsplea below, it is assumed that: a. A site specific CCC is being derived. ;. 2 M& zl,; cc is 2 ug/ L.

d: The eWZR and uk are COllStMt and strictly additive. 8: The flaw of the effluent ( eFLCW) is always 10 cfs. f. The design flaw of the upstream titer ( uFT,, Wdf) is 40 cfr. Therefore:

ar (( 2 w/ L) mm ( 10 Cf8 + wuml  [( ucmc) ( uaq 10 w/ L .

B m ( ffam ( 10 cm + ( ucmnma ( 40 Cf8) ( 2 w/ L) ( 10 Cf8 + 40 cf8) l In the first  18, the USER is assumed to be 5 and SO the upstream site specific CCC ( wsCCC)  ( CCC) ( uWW) = ( 2 ug/ L)( S)  10 UQ/ L. whichamaM uCONC is assumed to be 0.4 ug/ L,

water is that the assimilative capacity of the upstream 9.6 ug/ L.

40 20.0 80.0 12.000 118.4 12.00 63 13.7 06.3 9.795 140.5 14.21 90 10.0 90.0 8.500 166.4 16.80 190 5.0 95.0 6.750 262.4 26.40 490 2.0 98.0 5.700 550.4 55.20 990 1.0 99.0 5.350 1030.4 103.20 1990 0.5 99.5 5.175 1990.4 199.20 . At wte Mxx hWER As the flaw 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 & d hWERs increase due to the increased dilution of the effluent. The ixkrease in hWER at higher flows will not allw any use of the assimilative capacity of the upstream water because the allowed concentration of metal in the effluent is controlled by the lwest hWER, which is the design flaw hwER in this  le. Any WER determined at a higher flaw CM be used as M MVirOIImMt& lly conservative estimate of the design flaw WER, and the hWERs show that the WER of 12 provides adequate protection at all flows. Whti uFLCW equals the design flow of 40 cfs, WER = hwER.

32 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. . At Cmlete Mix hWER Eizxz zrEff. m LEE *,

10 40 20.0 80.0 8.800 80.00 8.800 10 63 13.7 86.3 6.343 80.00 8.800 ix 190 90 10.0 90.0 95.0 4.900 2.950 80.00 8.800 10 490 2'*: 98.0 1.780 80.00 8.800 10 990 1: o 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 c arable WERs in the first example because the uWER dropped from 5 to 1; the limiting value of the F 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 0: ussCCC, i. e., when the assimilative capacity is zero. As in the first example, any WER determined at a flow higher than design flow CM be used as M 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 infonaation is available. When uFLOW equals the design flow of 40 cfs, WER = hWER.