United States         Office of Science and Technology
Environmental Protection Agency Health and Ecological Criteria OiV.     September 1993
Office of Water &       Washington. O.C. 2046O
Office of Research and
Development                  .    '
Sediment Quality Criteria
for the Protection of
Benthic Organisms:
DIELDR1N

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                                CONTENTS


                                                              PAGE

 Foreword	
 Acknowledgments   	       	'	"    .»
 Tables  .	".".'.'.'.".'.".*.'.'.".".	    m
 Figures	        	    V-
 Introduction	    ™
 Partitioning	.'*.......!!.....	    21
 Toxicity of Dieldrin: Water Exposures  . . . . .	    31
 Toxicity of Dieldrin (Actual and Predicted): Sediment Exposures	    4-\
 Criteria Derivation for Dieldrin		" "     5 J
 Criteria Statement	gli
References	

Appendix A: Summary of Acute ValueYfor Dieldrin for Freshwater and Saltwater"
           Species  	            A _i
Appendix B: Summary of Data from Sediment Spiking Experiments with
           Dieldrin" " ' "

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                                      FOREWORD
        Under the Clean Water Act (CWA) the U.S. Environmental Protection Agency (U S
        and the States develop programs for protecting the chemical, physical, and biological
  integrity of the nation's waters.   Section 304(a)(l) directs the Administrator to develop and
  publish "criteria" reflecting the latest scientific knowledge on: (1) the kind and extent of effects
  on human health and welfare, including effects on plankton, fish, shellfish, and wildlife which
  may be expected from the presence of pollutants in any body of water, including ground water
  (2) the concentration and dispersal of pollutants, or their byproducts, through biological, physical
  and chemical processes, and (3) the effects of pollutants  on biological community diversity
 productivity, and stability.  Section 304(a)(2) directs the Administrator to develop and publish
 information on, among other things, the factors necessary for the protection and propagation of
 shellfish, fish, and wildlife for classes and categories of receiving waters.

        To meet this objective, U.S. EPA has periodically issued ambient water quality criteria
 (WQC) guidance beginning with the publication of "Water  QHiality Criteria 1972" (NAS/NAE
 1973).  All criteria guidance through late  1986 was summarized in an U.S. EPA document
 entitled "Quality Criteria for Water, 1986" (U.S. EPA, 1987). Additional WQC documents that
 update criteria for selected chemicals and provide new criteria for other pollutants have also been
 published. In addition to the development of WQC and to continue to comply with the mandate
 of the CWA, U.S. EPA has conducted efforts to develop and publish sediment quality criteria
 (SQC) for some of the 65 toxic pollutants or toxic pollutant categories. Section 104 of the CWA
 authorizes the administrator to  conduct and promote research  into the causes, effects  extent
 prevention, reduction and elimination of pollution, and to publish relevant information.' Section
 104(n)(l) in particular provides for study of the effects of pollution, including sedimentation in
 estuanes, on aquatic life, wildlife, and recreation. U.S. EPA's efforts with respect to sediment
 criteria are also authorized under CWA Section 304(a).

       Toxic contaminants in bottom sediments of the nations's lakes, rivers, wetlands and
 coastal waters create the potential for continued environmental degradation even where water
 column contaminantlevels meet established WQC. In addition, contaminated sediments can lead
 to water quality impacts, even when direct discharges to the  receiving water have ceased  EPA
 intends SQC be used to assess  the extent of sediment contamination, to aid in implementing
 measures to limit or prevent additional contamination, and to  identify and implement appropriate
 remediation activities when needed.

       The criteria presented in this document are the U.S. EPA's best recommendation of the
 concentrations of a  substance that may be present in sediment while still protecting benthic
 organisms from the  effects of that substance.  These criteria are applicable to  a variety of
freshwater and marine sediments because  they  are based  on the biologically available
concentration of the substance in sediments.  These criteria do  not protect against additive
synergistic or antagonistic effects of contaminants or bioaccumulative effects to aquatic life'
wildlife or human health.                                                             '
                                          11

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       The  criteria derivation methods outlined in this document are proposed to provide
protection of benthic organisms from-biological impacts from chemicals present in sediments.
Guidelines and guidance are being developed by U.S. EPA to assist in the application of criteria
presented in this document, in the development  of sediment quality standards, and in other
water-related programs of this Agency.

       These criteria are being issued in support of U.S. EPA'S  regulations  and policy
initiatives.  This document is Agency guidance only. It does not establish or affect legal rights
or obligations. It does not establish a binding norm and is not finally determinative of the issues
addressed.   Agency decisions in any particular case will be made by applying the law ana
regulations on the basis of the specific facts.
                                         111

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 Principal Author

       David J. Hansen

 Coauthors

       Walter J. Berry


       Dominic M. Di Toro


      PaulR. Paquin


      Laurie D. De Rosa


       Frank E. Stahcil, Jr.

       Robert A. Hoke


       Christopher S. Zarba

Technical and Clerical Support

      Heinz P. Kollig

      Glen B. Thursby


      Maria R.  Paruta


      Dinalyn Spears

      Deborah Robson
                               ACKNOWLEDGEMENTS
  U.S. EPA, Environmental Research Laboratory,
  Narragansett, RT
 Science Applications International Corporation,
 Narragansett, RI

 Manhattan College, Bronx, NY
 HydroQual, Inc., Mahwah, NJ

 HydroQual, Inc.,
 Mahwah, NJ

 HydroQual, Inc.,
 Mahwah, NJ
          •,
 U.S. EPA, Environmental Research Laboratory, Athens, GA

 Science Applications International Corporation,
 Hackensack, NJ

 U.S. EPA Headquarters, Office of Water, Washington, DC
U.S. EPA, Environmental Research Laboratory, Athens, GA

Science Applications International Corporation,
Narragansett, RI

NCSC Senior Environmental Employment Program,
Narragansett, RI

Computer Science Corporation, Narragansett, RI

Science Applications International Corporation
Narragansett, RI
                                        IV

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      Persons who have made significant contributions to the development of the approach and
supporting-science used.in the derivation of sediment criteria for nonionic organic contaminants
are as follows:
      Herbert E. Allen

      Gerald T. Ankley


      Christina E.  Cowan

      Dominic M.  Di Tore


      David J. Hansen


      PaulR. Paquin

      Spyros P. Pavlou

      Richard C. Swartz


      Nelson A. Thomas


      Christopher S. Zarba
University of Delaware, Newark, DE

U.S. EPA, Environmental Research Laboratory,
Duluth, MN

Battelle, Richland, WA

HydroQual, Inc., Mahwah, NJ;
Manhattan College, Bronx, NY

U.S. EPA,  Environmental Research Laboratory,
Narragansett, RI

HydroQual, Inc., Mahwah, NJ

Ebasco Environmental, Bellevue, WA

U.S. EPA,  Environmental Research Laboratory,
Newport, OR

U.S. EPA,  Environmental Research Laboratory,
Duluth,  MN

U.S. EPA Headquarters, Office of Water, Washington, DC

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  Table 2-1.

  Table 2-2.



  Table 2-3.



  Table 3-1.



 Table 3-2.




 Table 3-3.
 Table 4-1.


 Table 4-2.

         V

 Table 5-1.


 Table 5-2.
 Dieldrin measured and estimated log10KoW values..


 Summary  of  log10KoW  values  for dieldrin  measured by the U.S  EPA
 Environmental Research Laboratory, Athens, GA.


 Summary of KQC values for dieldrin derived from literature sorption isotherm
 data.


 Chronic sensitivity of freshwater and saltwater organisms to dieldrin
 Test specific data.


 Summary of freshwater and saltwater acute land chronic values, acute-chronic
 ratios, and derivation of final acute values, final acute-chronic ratios, and final
 chronic values for dieldrin.


 Results of approximate randomization test for the equality  of freshwater and
 saltwater FAV distributions for dieldrin and approximate randomization test for
 the equality of benthic and combined benthic and water  column (WQC) FAV
 distributions.


 Summary of tests with dieldrin-spiked sediment.


 Water-only and sediment LC50s used to test the applicability of the equilibrium
 partitioning theory for dieldrin.


 Sediment quality criteria for dieldrin.


Analysis of variance for derivation of sediment quality criteria confidence
limits for dieldrin.
Table 5-3.    Sediment quality criteria confidence limits for dieldrin.
Appendix A. - Summary of acute values for dieldrin for freshwater and saltwater species.


Appendix B. -Summary of data from sediment spiking  experiments with dieldrin. Data
              from these experiments  were used to calculate KOC values (Figure 2-2) and to
              compare mortalities of test organisms with pore water toxic units (Figure 4-1)
              and predicted sediment toxic units (Figure 4-2).
                                         VI

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                                       FIGURES
                          • J
             ,       t
 Figure 1-1.   Chemical structure and physical-chemical properties of dieldrin.

 Figure 2-1    Observed versus predicted (equation 2-4) partition coefficients  for  nonionic
              organic chemicals (dieldrin datum is highlighted).

 Figure 2-2.   Organic carbon-normalized sorption isotherm for dieldrin (top) and probability
              plot of KOC (bottom) from sediment toxicity tests conducted by Hoke and Ankley
              (1991). The line in the top panel represents the relationship predicted with a log
                 of 5.25, that is 0.^=1^ • Cd
 Figure 3-1.   Genus mean acute values from water-only acute toxicity tests using freshwater
              species vs. percentage rank of their sensitivity. Symbols  representing benthic
              species are solid, those representing water column species are open.  Asterisks
              indicate greater than values. A =  adult,  J = juvenile, X = unspecified life
              stage.

 Figure 3-2.   Genus mean acute values from water-only acute toxicity tests using saltwater
              species vs. percentage rank of their sensitivity. Symbols  representing benthic
              species are solid, those representing water column species  are open.  Asterisks
              indicate greater than values.  A = adult, J = juvenile.

 Figure 3-3.   Probability distribution of FAV difference statistics to compare water-only data
              from freshwater vs. saltwater (upper panel) and benthic vs. WQC (lower panel)
              data.

 Figure 4-1.   Percent mortality of  amphipods in  sediments spiked with acenaphthene or
             phenanthrene (Swartz,  1991), endrin (Nebeker et aL, 1989; Schuytema et al.,
              1989), or fluoranthene (Swartz et al.s 1990), and midge in sediments spiked with
             dieldrin (Hoke, 1992) or kepone (Adams et al. , 1985) relative to pore water toxic
             units. Pore water toxic units are ratios of concentrations of chemicals measured
             in individual treatments divided by the water-only LC50 value from water-only
             tests.  (See Appendix B in this  SQC document, Appendix  B in the endrin,
             acenaphthene,  fluoranthene and phenanthrene SQC documents,  and original
             references for raw data.)

Figure 4-2.   Percent mortality of  amphipods in  sediments spiked  with  acenaphthene or
             phenanthrene (Swartz,  1991), dieldrin (Hoke and Ankley, 1991), endrin (Nebeker
             et al. , 1989; Schuytema et al. , 1989) or fluoranthene (Swartz et al. , 1990; DeWitt
           - et al.,  1992) and midge, in dieldrin spiked sediments (Hoke,  1992) relative to
             "predicted sediment toxic units. " Predicted sediment toxic units are the ratios of
             measured  treatment concentrations  for each  chemical  in  sediments  Otg/goc)
             divided by the predicted LC50 Otg/goc) in sediments (Koc x Water-Only LC50
                                         Vll

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              G*g/L) • IKg^/l.OOO&J. (See Appendix B in this document and Appendix B in
           •   theace,naphthene, endrin, fluoranthene, andphenanthrene SQC documents for raw
              data).

  Figure 5-1.   Comparison between SQC concentrations and 95% confidence intervals  effect
              concentrations from benthic organisms expoajd to dieldrin-spiked sediments and
              sediment concentrations predicted to be chronically safe in fresh water sediments
              Concentrations predicted to be chronically safe (Predicted Genus Mean Chronic
              Values, PGMCV) are derived from the Genus Mean Acute Values (GMAV) from
              *™™?y 9,™°U/ lethaHty ***• Acute Chronic "" 
                                      VUl

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       This report has been reviewed by the Health and Ecological Criteria Division, Office oi
Science and Technology, U.S. Environmental Protection Agency, and approved for publication.
Mention  of trade  names or  commercial products  does not constitute  endorsement  or
recommendation for use.
                             AVAILABILITY NOTTCE
      This document is available to the public through the National Technical Information
Service (NTIS), 5285 Port Royal Road, Springfield, VA 22161. NTIS Accession Number
XXXX-XXXXXX.
                                       IX

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                                     SECTION 1
                                   INTRODUCTION
 1.1 GENERAL INFORMATION:
       Under the Clean Water Act (CWA) the U.S. Environmental Protection Agency (U.S.
 EPA) is responsible for protecting the chemical, physical and biological integrity of the nation's
 waters. In keeping with this responsibility, U.S. EPA published ambient water quality criteria
 (WQC) in 1980 for 64 of the 65 toxic pollutants or pollutant categories designated as toxic in
 the CWA. Additional water quality documents that update criteria for selected consent decree
 chemicals and new criteria have been published  since 1980.  These WQC  are numerical
 concentration limits that are the U.S. EPA's best estimate of concentrations protective of human
 health and the presence and uses of aquatic life.  While these WQC play an important role in
 assuring a healthy aquatic environment, they alone are not sufficient to ensure the protection of
 environmental or human health.
      Toxic pollutants in bottom sediments of the nation's lakes, rivers, wetlands, estuaries and
 marine coastal waters create the potential for continued environmental degradation even where
 water-column  concentrations  comply with established WQC.   In  addition,  contaminated
 sediments  can be a significant pollutant source that may cause water quality degradation  to
persist, even when other pollutant sources are stopped.  The absence of defensible sediment
quality criteria (SQC) makes it difficult to accurately assess the extent of the ecological risks  of
contaminated sediments and to identify, prioritize and implement appropriate clean up activities
                                        1-1

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    and source controls.  As a result of the need for a procedure to assist regulatory agencies in
              ***••*•            •                  ,
                       '   •   '
    making decisions concerning contaminated sediment problems,  a U.S. EPA Office of Science



    and Technology, Health and Ecological Criteria Division  (OST/HEC) research team was



    established to review alternative approaches (Chapman, 1987). All of the approaches reviewed



    had both strengths and weaknesses and no single approach was found to be applicable for SQC



    derivation in all situations (U.S. EPA, 1989a). The equilibrium partitioning (EqP) approach was



    selected for nonionic organic chemicals because it presented the greatest promise for generating



    defensible national numerical chemical-specific SQC applicable across a broad range of sediment



    types. The three principal observations that underlie the EqP method of establishing SQC are:



          1.     The concentrations of nonionic organic chemicals  in sediments, expressed on an

                                              •,

                 organic carbon basis, and in pore waters correlate to observed biological effects



                 on sediment dwelling organisms across a range of sediments.



          2.     Partitioning models can relate  sediment concentrations for nonionic organic



                 chemicals on an organic carbon basis to freely dissolved chemical concentrations


                 in pore water.



          3.     The distribution of sensitivities  of benthic  and  water column organisms  to



                 chemicals are similar, thus, the currently established  WQC final chronic values



                 (FCV) can be used to define the acceptable effects concentration of a chemical


                freely-dissolved in pore water.



          The EqP approach, therefore, assumes that: (1) the partitioning of the chemical between



   sediment organic carbon and interstitial water is at equilibrium; (2) the concentration in either



.   phase can be predicted using appropriate partition coefficients  and the measured concentration
                                            1-2

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 in the other phase; (3) organisms receive equivalent exposure from water-only exposures or from
                         .'.''**
 any equilibrated phase: either from pore water via respiration, from sediment via ingestion or

 other sediment-integument exchange or from a mixture of both exposure routes; (4) for nonionic

 chemicals, effect concentrations in sediments on an organic carbon basis can be predicted using

 the organic carbon partition coefficient (Koc) and effects concentrations in water, (5) the FCV

 concentration is an appropriate effects concentration for freely-dissolved chemical in interstitial

 water; and (6) the SQC (pg/goc) derived as the product of the K^ and FCV is protective of

 benthic organisms.  SQC concentrations  presented in  this document are expressed as pg

 chemical/g sediment organic carbon and not on an interstitial water basis because: (1) pore water

 is difficult to adequately sample; and (2) significant amounts of the dissolved chemical may be
                                           if
 associated with dissolved organic carbon; thus, total chemical concentrations in interstitial water

 may overestimate exposure.

        The data that support the EqP  approach  for deriving SQC for nonionic  organic

 chemicals are reviewed by Di Toro et al. (1991) and U.S. EPA, (1993a). Data supporting these

 observations for dieldrin are presented in this document.

      SQC generated using the EqP method are suitable for use in providing guidance to

regulatory agencies because they are:

      1.  numerical values;

      2.  chemical specific;

      3.  applicable to most sediments;

      4.  predictive of biological effects;  and

      5.  protective of benthic organisms.
                                        1-3

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   As is the case with WQC, the SQC reflect the use of available scientific data to: (1) assess the
                            i
   likelihood of significant environmental effects to benthic organisms from chemicals in sediments;

   and (2) to derive regulatory requirements which will protect against these effects.

         It should be emphasized that these criteria are intended to protect benthic organisms from

  the effects of chemicals associated with sediments.  SQC are intended to apply to sediments

  permanently inundated with water, intertidal sediment and to sediments inundated periodicaUy

  for durations sufficient to permit development of benthic assemblages. They do .not apply to

  occasionally inundated soils containing terrestrial organisms. These criteria do not address the

  question of possible contamination of upper trophic level organisms or the synergistic, additive

  or antagonistic effects of multiple chemicals.  SQC addressing these issues may result in values

  lower or higher than those presented in this document. The SQC presented in this document

 represent the U.S. EPA's best recommendation at this time of the concentration of a chemical

 in sediment that will not adversely affect most benthic organisms. SQC  values may be adjusted

 to account for future data.


       SQC values may also need to be adjusted because of site specific  consideration. In spill

 situations,  where chemical equiUbrium between water and sediments has not yet been reached,

 sediment chemical concentrations less than SQC may pose risks  to benthic organisms. This is

 because  for spills, disequilibrium concentrations in interstitial  and overlying water may be

proportionally higher relative to sediment concentrations. Research has  shown that the source

or "quality" of TOG in the sediment does not effect chemical binding (DeWitt et al., 1992).

However, the physical form of the chemical in the sediment may  have an effect.  At some sites

concentrations in excess of the SQC may not pose risks to benthic  organisms, because the



                                         1-4

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  compound may be a component of a paniculate, such as coal or soot, or exceed solubility such
                   '',-.-•

  as undissolved oil or chemical. In these situations, the national SQC would be overly protective


  of benthic organisms and should not be used unless modified using the procedures outlined in


  the "Guidelines for Deriving Site-specific Sediment Quality Criteria for the Protection of Benthic


  Organisms" (US EPA, 1993b). The SQC may be underprotective where the toxicity of other


  chemicals are additive with the SQC  chemical or species of unusual sensitivity occur at the site.


        This document presents the  theoretical basis and the supporting data relevant to the


 derivation of the SQC for dieldrin. An understanding of the "Guidelines for Deriving Numerical


 National Water Quality Criteria for the Protection of Aquatic Organisms and Their Uses"


 (Stephan et aL, 1985), response to public comment (U.S. EPA, 1985) and "Technical Basis for


 Deriving Sediment Quality Criteria for Nonionic Organic Contaminants for the Protection of


 Benthic Organisms by Using Equilibrium Partitioning" (U.S. EPA 1993a) is necessary in order


 to understand  the following text, tables and calculations. Guidance for the acceptable use of


 SQC values is  contained in "Guide for .the Use and Application of Sediment Quality Criteria for

 Nonionic Organic Contaminants" (U.S. EPA,  1993c).




 1.2  GENERAL INFORMATION: DIELDRIN


    Dieldrin is the common name of a persistent, non-systemic organochlorine insecticide used


for control of public health insect pests, termites and locusts.  It is formulated for use as an


emulsifiable concentrate, wettable and dustable powder and granular product. Other than direct


usage of dieldrin, another  source  of dieldrin in the environment stems from the quick


transformation  of aldrin, also an organochlorine pesticide, to dieldrin. Both dieldrin and aldrin
                                        1-5

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  usage peaked in the mid-1960s and declined until the early 1970s.  Dieldriri and aldrin have been
  restricted from registration and production in the United States since 1974 (U.S. EPA, 1980a).
       Dieldrin is a cyclic hydrocarbon having a chlorine substituted methanobridge (Figure 1-1).
  It is structurally similar to endrin, its endo-endo stereoisomer, and has similar physico-chemical
  chlorine properties,  except that it is more difficult to degrade in the environment (Wang, 1988).
  Dieldrin is a colorless crystalline solid at room temperature, having a melting point of about
  176°C and specific  gravity of 1.75 at 20°C.  It also has a vapor pressure of 0.4 mEa and a
  solubility of 0.19 mg/L at 20°C (Hartley and Kidd, 1987).
     Dieldrin is considered to be toxic to aquatic organisms, bees and mammals (Hartley and
 Kidd, 1987).  The acute toxicity of dieldrin ranges from 0.5 to 740 ug/L for freshwater and 0.7
 to  > 100  /tg/L for saltwater  organisms  (Appendix A).    Differences  between  dieldrin
 concentrations causing acute lethality and chronic toxicity in specie.? acutely sensitive to this
 insecticide are small; acute-chronic ratios range from 2.417 to 12.82 for three species (Table 3-
 3).  Dieldrin bioconcentrates in aquatic animals from 400 to 68,000 times the concentration in
 water (U.S. EPA, 1980a).  The WQC for dieldrin (U.S. EPA,  1980a) is derived using a Final
Residue  Value calculated using bioconcentration data  and the FDA action level to protect
marketability of fish  and shellfish;  therefore, the WQC is not  "effects based". The SQC for
dieldrin is effects based.  It is calculated from the Final Chronic Value (FCV) derived in section
3.
1.3 OVERVIEW OF DOCUMENT:
      Section 1 provides a brief review of the EqP methodology, and a summary of the
                                         1-6

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                             JA,     V
                            ci
                 MOLECULAR FORMULA            C12H,C160
                 MOLECULAR WEIGHT               380.93
                 DENSITY            ,,             1.75 g/cc (20-0
                 MELTING POINT                    176°C
                 PHYSICAL FORM             Colorless crystal
                 VAPOR PRESSURE                  0.40 mPa (20eC)
    CAS NUMBER: 60-57-1
    TSL NUMBER:  IO 15750
 COMMON NAME: Dieldrin (also dieldrine and ndieldrin)
   ^TRADE NAME: Endrex (Shell); Hexadrin
CHEMICAL NAME: l,2,3,4,10,10shexachloro-lR,4S,4aS,5R,6R,7S,8SR,8aR-
                 octahydro-6 J-epoxy-l,4:5,8-dimethanoaphthalene (IUPAC)
     FIGURE 1-1. Chemical structure and physical-chemical properties of dieldrin.
                                   1-7

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 physical-chemical properties and aquatic toxicity of dieldrin. Section 2 reviews a variety of

 methods and'data useful in deriving partition coefficients for dieldrin and includes the KQC

 recommended for use in the derivation of the dieldrin SQC. Section 3 reviews aquatic toxicity

 data contained in the dieldrin WQC document (U.S. EPA, 1980a) and new data that were used

 to derive the FCV used in this document to derive the SQC concentration.  In addition, the

 comparative sensitivity of benthic and water column species is examined as the justification for

 the use of the FCV for dieldrin in the derivation of the SQC.  Section 4 reviews data on the

 toxicity of dieldrin in sediments, the need for organic carbon normalization of dieldrin sediment

 concentrations and the accuracy of the EqP prediction of sediment toxicity using KQC and an

 effect concentration in water.  Data from Sections 2, 3 and 4 are used in Section 5 as the basis
                                         »f
 for the derivation of the SQC for dieldrin and its uncertainty.  The SQC for dieldrin is then

compared to STORET (U.S. EPA, 1989b) and National Status and Trends (NOAA, 1991) data

on dieldrin's environmental occurrence in sediments.  Section 6 concludes with the criteria

statement for dieldrin.  The references used in this document are listed in Section 7.
                                       1-8

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                                     SECTION 2



                                  PARTITIONING
 2.1 DESCRIPTION OF THE EQUILIBRIUM PARTITIONING METHODOLOGY:


       Sediment quality criteria (SQC) are the numerical concentrations of individual chemicals


 which are intended to be predictive of biological effects, protective of the presence of benthic


 organisms and applicable to the range of natural sediments from lakes,  streams, estuaries and


 near coastal marine waters. As a consequence, they can be used in much the same way as water


 quality criteria (WQC); ie., the concentration of,a chemical which is protective of the intended


 use such as aquatic life protection. For nonionic organic chemicals, SQC are expressed as ;tg


 chemical/g organic carbon and apply to sediments having Ss 0.2% organic carbon by dry


 weight. A brief overview follows of the concepts which underlie the equilibrium partitioning


 (EqP) methodology for deriving SQC. The methodology is discussed in detail in the "Technical


 Basis for Deriving Sediment  Quality Criteria  for Nonionic  Organic Contaminants for the


 Protection of Benthic Organisms by Using  Equilibrium Partitioning"  (U.S. EPA, 1993a),


 hereafter referred to as the SQC Technical Basis Document.


       Bioavailability of a chemical at a particular sediment concentration often differs from one


 sediment type to another. Therefore, a method is necessary for determining a SQC based on the


bioavailable  chemical  fraction  in  a sediment.    For  nonionic organic  chemicals,  the


concentration-response relationship for the biological effect of concern can most often be
                               lj :

correlated  with the interstitial water (i.e., pore water) concentration  0»g chemical/liter pore
                                       2-1

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  1991).  From a purely practical point of view, this correlation suggests that if it were possible
  to  measure the  pore water chemical concentration,  or predict it from  the  total sediment
  concentration and the relevant  sediment properties, then that concentration could be used to
  quantify the exposure concentration for an organism.  Thus, knowledge of the partitioning of
  chemicals between the solid and liquid phases in a sediment is a necessary component for
  establishing SQC.  It is for this reason that the  methodology described below is called the
  equilibrium partitioning (EqP) method.
       It is shown in the SQC Technical Basis Document (U.S. EPA, 1993a) that the final acute
 values (FAVs) in the WQC documents are appropriate for benthic organisms for a wide range
 of chemicals. (The data showing this for dieldrin are presented in Section 3). Thus, a SQC can
 be established using the final chronic value (FCV) derived using the WQC Guidelines (Stephan
 et aL, 1985) as the acceptable effect concentration in pore or overlying water (see Section 5),
 and the partition coefficient can be used to relate the pore water concentration to the sediment
 concentration via  the partitioning equation. This acceptable concentration in  sediment is the
 SQC.
       The calculation is as follows: Let FCV fog/L) be the acceptable concentration in water
for the chemical  of interest; then compute the SQC using the partition coefficient, (Kp)
O-^giediment)}  between sediment and water:
             SQC = Kp  FCV                                                (2-l)
This is the fundamental equation used to generate the SQC.  Its utility depends upon the
existence of a methodology for quantifying the partition  coefficient, Kp.
      Organic carbon appears to be the dominant sorption phase for nonionic organic chemicals

                                        2-2

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  in naturally occurring sediments and thus controls the bioavailabffity of these compounds in
  sediments. Evidence for this can be found in numerous toxicity tests, bioaccumulation studies
  and chemical analyses of pore water and sediments (Di Toro et aL, 1991).  The evidence for
  dieldrin is discussed in this section and section 4. The organic carbon binding of a chemical in
  sediment is a function of that chemical's organic  carbon partition coefficient (Koc) and the
  weight fraction of organic carbon in the sediment (f^.  The relationship is as follows:
              KP  = foe KOC                                                  (2-2)
       It follows that:
              SQCOC = KOCFCV                                              p_3)

 where SQCoc is the sediment quality criterion on a sediment organic carbon basis.
       Koc is not usually  measured directly  (although it  can be done, see section  2.3).
 Fortunately, KOC is closely related to the octanol-water partition coefficient (KoW) (equation 2-5)
 which has been measured for many compounds, and can be measured very accurately. The next
 section reviews the available information on the KoWfor dieldrin.

 2.2 DETERMINATION OF KOW FOR DIELDRIN:
    Several approaches have been used to determine KOW for the derivation of SQC, as discussed
in the SQC Technical Basis Document. At the U.S. EPA, Environmental Research Laboratory
at Athens, GA (ERL,A) three methods were selected for measurement and two for estimation
of KOW.  The measurement methods were shake-centrifugation (SC), generator column (GCol),
and slow-stir-flask (SSF), and the  estimation  methods were  SPARC  (SPARC Performs
                                        2-3

-------
  Automated Reasoning in Chemistry;  Karickhoff et al., 1989) and CLOGP (Chou and Jure,
  1979). Data were also extracted from the literature. The SC method is a standard procedure in
  the Organization for Economic Cooperation and Development (OECD) guidelines for testing
  chemicals, therefore, it has regulatory precedence.
     In the examination of the literature data primary references were found listing measured
 KQWS for dieldrin ranging from 4.09 to 6.2 (Table 2-1). Primary references were found in the
 literature for estimated log10KoW ranging from 3.54 to 5.40 (Table 2-1).  The range of reported
 values for dieldrin is significantly greater than the range of values for some other compounds,
 and we were not able to determine from studying the primary articles that any value was more
 likely to be accurate than any other.
                                          •.
 TABLE 2-1.  DIELDRIN MEASURED AND ESTIMATED LOG,0KoW VALUES .
METHOD
Measured
Measured
Measured
Measured
Measured
Estimated
Estimated
LOG10KoW
4.09
4.54
4.65
5.40
6.2
3.54
5.40
REFERENCE
Ellington and Stancil, 1988
Brooke, et al., 1986
De Kock and Lord, 1987
De Bruijn et al., 1989
Briggs, 1981
Mabey etal., 1982
SPARC'
•SPARC is from SPARC Performs Automated Reasoning in Chemistry, (Karickhoff et al.,
1989).
  A KOW value for SPARC is also included in Table 2-1.  SPARC is a computer expert system
under development at ERL,A, and the University of Georgia, at Athens. For more information
on SPARC see U.S. EPA (1993a).  The SPARC estimated log10KoW value for dieldrin is 5.40.
                                        2-4

-------
 We had little confidence in the available measured or estimated values for K
-------
       Laboratory studies to characterize adsorption are generally conducted using particle
                  »   •   '       "'      '                   i   '•'
 suspensions.  The high concentrations of solids and turbulent conditions necessary to keep the

 mixture in suspension make data interpretation difficult as a inesult of a particle interaction effect.

 This effect suppresses the partition coefficient relative to that observed for undisturbed sediments

 (Di Toro, 1985; Mackay and Powers, 1987).

 TABLE 2-2.  SUMMARY OF LOG10Kow VALUES FOR DIELDRIN MEASURED BY THE
 U.S. EPA, ENVIRONMENTAL RESEARCH LABORATORY, ATHENS, GA.
SHAKE
CENTRIFUGATION
(SC)
5.04
5.00
5.04
5.03
5.04
4.88
4.99

5.01"
GENERATOR
COLUMN
(GCol)
4.89
4.88
5.1S
5.15
5.26
5.38
5.67'
5.04
5.16b
SLOW STIR
FLASK
(SSF)
5.33
5.43
5.38
5.33
5.43
5.08
5.28

5.34b
      "Value considered outlier and omitted from mean Computation.
      bLog10 of mean measured values.
      Based on analysis of an extensive body of experimental  data for a wide range of

compound types and experimental conditions, the particle interaction model (Di Toro, 1985)

yields the following relationship for estimating KP:
                                                                          (2-4)
1 +
                         KQC / V x
                                       2-6

-------
 where m is the particle concentration in the suspension (kg/L), and u* = 1.4, an empirical
 constant.   The K^ is given by:
                      = 0.00028 + 0.983 log10KoW                              (2-5)
       Figure 2-1 compares observed partition coefficient data for the reversible component with
 calculated values estimated with the particle interaction model (Equation 2-4 and Equation 2-5)
 for a wide range of compounds (Di Toro, 1985). The observed partition coefficient for dieldrin
 using adsorption data (Sharom et aL, 1980) is highlighted on this plot.  The observed log10Kp
 of 1.68 reflects significant particle interaction effects. The observed partition coefficient is more
 than an order of magnitude lower than the value expected in the absence of particle effects ( i.e. ,
 log10Kp = 3.32 from the fooKoc = 2100 L/kg).' KQC was computed from equation 2-5.
       Several sorption isotherm experiments with particle suspensions that provide an additional
 way to compute KQC were found in a comprehensive literature search for partitioning information
 for dieldrin (Table 2-3).  TheKoc values derived from these data are lower than KQC values from
 laboratory measurements of KOW.  The lower KQC can be explained from the particle interaction
 effects. Partitioning in a quiescent setting would result in less desorption and higher KOC. These
 data are presented as examples of particle interaction effects only as 100 percent reversibility
 is assumed in the absence of desorption studies and actual K^ can not be computed.
       In the absence of particle effects,  KQC is related to KQW via Equation 2-5. For logioKow
 = 5.34 (ERL,A, mean measured value), this expression results in an estimate of log10Koc =
5.25.
                                        2-7

-------
          O)

         \

         a

         o
         TJ
         d>

         a>
         €0
         JO
         O
                        Partition Coefficient

                      Reversible Component
                    Predicted  log 10 Kp  (L/kg)
Figure 2-1
Ob^ived versus predicted (equation 2-4) partition coefficients for nonionic
organic chemicals (dieldrin datum is highlighted).
                              2-8

-------
   2.3.2. KQC FROM SEDIMENT TOXTCITY TESTS:
      Measurements of K^ are available from sediment toxicity tests using dieldrin (Hoke and
   Ankley, 1992). These tests were with a sediment having an average organic carbon content of
   1.75 percent (Table 4-1; Appendix B). Dieldrin concentrations were measured in sediments and
   unfiltered pore waters providing the data necessary to calculate the partition coefficient for an
   undisturbed  bedded sediment.  Since it is likely that organic carbon complexing in pore water
   is significant for dieldrin, organic  carbon concentrations were also measured in pore water.
  Figure 2-2 is a plot of the organic carbon-normalized sorption isotherm for dieldrin, where the
  sediment dieldrin concentration frg/g^ is plotted versus calculated free (dissolved) pore water
  concentration (pg/L). Using pore water organic carbon concentrations (DOC), and assuming
  KBOC equal to K^, the calculated tree pore water dieldrin concentration CD 0*g/L) is presented
  in Figure 2-2 is given by:
                   l+mpoeKpoc                                             (2'6>
 where Croro is the measured total pore water concentration  and m^ is the measured DOC
 concentration (U.S. EPA, 1993a).
       The data used to make this plot are included in Appendix B.  The line of unity slope
 corresponding to the log^ = 5.25 derived from SSF is compared to the data. The data from
 the sediment toxicity test fall on the line of unity slope for logl^oc  = 5.25.
       A probability plot of the observed experimental log10Koc  values is shown in the lower
,panel of Figure 2-2. The lo&oKoc values are approximately normally distributed with a mean
 of Log10 KOC = 5.32 and a standard error of the mean of 0.109. This value is in agreement
                                         2-9

-------
                              i-Hiiii—i i  i IIIHI—r-rn
                                         LEGEND
                                      - Hoke ond Ankley, 1992
         0.1 — »  i i HUH — '
001
                     0.1
                                   i  t i mi
                                                  i tun   § i  i
                                                   100
                                                  1000
      CALCULATED PORE WATER  CONCENTRATION  
-------
  with log10Koc = 5.25, which was computed from the SSF determined (Section 2.2) dieldrin

            5.34 (Equation 2-5).


               TABLE 2-3.  SUMMARY OF KOC VALUES FOR DIELDRIN
            DERIVED FROM LITERATURE SORPTION ISOTHERM DATA.
Observed
LogioKoc (SD)
4.20 (0.14)
4.14 (0.15)
4.10
n
4
3
1
Solids (SD)
(g/L)
5.0
16.4 (4.6)
100.0
References
Eye, 1968
Betsffl, 1990
Briggs, 1981
 2.4 SUMMARY OF DERIVATION OF KOC FOR DIELDRIN:

       The KOC selected to calculate the sediment quality criteria for dieldrin is based on the

 regression of log^ to log10KoW (Equation 2-5), using the dieldrin log10KoW of 5.34 recently

 measured by ERL,A. This approach, rather than the use of the K^ from toxicity tests  was

 adopted because the regression equation is based on the most robust dataset available that spans

a broad range of chemicals and particle types, thus encompassing a wide range of K<,w and f^.

The regression equation yields a log^ = 5.25.  This value is in agreement with the log

of 5.32 measured in the sediment toxicity tests.
                                     2-11

-------

-------
                                     SECTIONS

                   TOXICITY OF DIELDRIN: WATER EXPOSURES

 3.1 TOXICTrY OF DffiLDRIN IN WATER: DERIVATION OF DffiLDRIN WATER
    QUALITY CRITERIA:


       The equilibrium partitioning (EqP) method for derivation of sediment quality criteria

 (SQC) uses the dieldrin water quality criterion (WQC) Final Chronic Value (FCV) and partition

 coefficients (Koc) to estimate the maximum concentrations of nonionic organic  chemicals in

 sediments, expressed on an organic carbon basis, that will not cause adverse effects to benthic

 organisms. For this document, life stages of species classed as benthic are either species that

 five in the sediment (infauna) or on the sediment surface (epibenthic) and obtain their food from

 either the sediment or water column (U.S. EPA, 1989c). In this section (1)  the FCV from the

 dieldrin WQC document (U.S. EPA, 1980a) is revised using new aquatic toxicity test data; and

 (2) the use of this FCV is justified as the effects concentration for SQC derivation.

 3.2 ACUTE TOXICITY - WATER EXPOSURES:

      One hundred  and forty five standard toxicity tests with dieldrin have been conducted on

25 freshwater species from 19 genera (Appendix A). Eighty six of these tests are from one

study with the guppy, Poecilla reticulata (Chadwick and Kiigemagi, 1968).  Overall genus mean

acute values (GMAVs) range from 0.5 to 740 /tg/L.  Fishes, damselflys, isopods, glass shrimp,

stoneflies, and mayflies were most sensitive; GMAVs for these taxa range from 0.5 to 24 pg/L.

Seventeen tests on thirteen benthic species from twelve genera are contained in this database

(Figure 3-1; Appendix A). Benthic organisms were among both the most sensitive, and most
                                       3-1

-------
         1000
          100
       i
       ui

       I
       HI


       1
      IU
      S
      CO


      i
      o
10
          0.1
                     A Arthropods

                     D Other Invertebrates

                     O.Fishes
                                                                         Orconectes (A)'

                                                              Gammarus (A,X)
                                                                Simocephalus (J,X)
                                                                                  Daphnia (A,J,X)
Ephemerella (X)
                 Micropterus
                                  _'Asellus (X)
                               fQnoprhynchus (J.X)
                      Pteronarcys (J, naiads)
                                                               [ Acroneuria (naiads)

                                                            *alaemonetes (X)
                                                             *-
       Ischnura (J)
    gpomis(J)

arassius (J,X)
                        Claassenfa (J)


                  Pleronarcella (J)
                  _L
                                               J_
                                  J_
                            20             40            (50            80


                           PERCENTAGE RANK OF FRESHWATER GENERA
                                                                           100
Figure 3-1.   Genus mean acute values from water-only acute toxicity tests using freshwater

              species vs. percentage rank of their sensitivity. Symbols representing benthic

              species are solid, those representing water column species are open.  Asterisks

              indicate greater than values.  A = adult,  J  = juvenile,  X = unspecified life

              stage.
                                           3-2

-------
 resistant, freshwater species to dieldrin; GMAVs range from 0.5 to 740 pg/L. Of the epibenthic



 species  tested,  channel catfish,  stoneflies, mayflies, damselflies, and  isopods  were most



 sensitive;  GMAVs  range from  0.5 to 12 /tg/L.   infaunal species tested include  only the



 stoneflies, Pteronarcvs califomica (LC50 = 4.416 ftgfL) and Pteronarcella badia (LC50 = 0.5



 jtg/L). The final acute value (FAV) derived from the overall GMAVs (Stephan et al. 1985) for



 freshwater organisms is 0.3595 /xg/L (Table 3-2).




       Thirty two acute tests  have been conducted on 23 saltwater species from 21 genera



 (Appendix A).  Overall GMAVs range from 0.70 to > 100 jtg/L.  Sensitivities of saltwater



 organisms were similar to those of freshwater organisms. Fishes and crustaceans were the most



 sensitive. Within this database there are results from 23 tests on benthic life-stages of 16 species



 from 14 genera (Figure 3-2; Appendix A).  Benthic organisms were among both the most



 sensitive, and most resistant, saltwater genera to dieldrin.  The most sensitive benthic species



 is the pink shrimp, Peneaus duorarum. with a flow-through 96 hour LC50 of 0.70 pg/L based



 on measured  concentrations. The American eel, Anquilla rostrata. has a similar sensitivity to



 dieldrin with  a 96 hr LC50 of 0.9 jtg/L. other benthic species for which there are data appear



 less  sensitive; GMAVs range from 4.5 to  >  100 jtg/L.  The FAV derived from the overall



 GMAVs (Stephan et al., 1985) for saltwater organisms is 0.6594 jtg/L (Table 3-2), less than the



 acute value for the economically important P. duorarum.








 3.3 CHRONIC TOXZCITY - WATER EXPOSURES:




       Chronic toxicity  tests have been conducted with dieldrin using two freshwater fish:



rainbow trout, Oncorhynchus myldss. and the guppy,  P. reticulata. and a saltwater mysid,
                                         3-3

-------
         1000
          100
      i
      (II
      I
      I
      CO


      til
      o
           10
                      A Arthropods

                      D Other Invertebrates

                      O Fishes
                                                              Ophyryotrocha *(A)
                                                    Poecilia (J)
                                          Cyprinodon (A)
                                                                       Sphaeroides (A)
                        Fundulus (J)

                                    ~ Crangon (A)
                               Thalassoma (A)
                       .   'MenidiafJ)
                       Mysidopsis (fl)
                   'Micrometrus(A) '
                                                 Morone (J)       s^ Crassostrea (A)
                                                             ^BMugit(A)
                                                             Palaemonetes (A}
                                                       Pagurus(A)
                                                   Gasterosteus (J)
                                             Palaemon (A)
                 Cymatogaster (J)


            'Oncorhynchus (J)
                      'Anguilla (J)

                   Penaeus (A)
         0.1
                           20            40            60            80

                          PERCENTAGE RANK OF SALTWATER GENERA
                                                                       100
Figure 3-2.
Genus mean acute values ftom water-only acute toxicity tests using saltwater
species vs. percentage rank of their sensitivity. Symbok represen
species are solid, those representing water column species a£ open
indicate greater than values. A - adult, J = juvenile
                                          3-4

-------
 Mysidopsis £*> (Table 3-1)..  Both Q. mvkiss and the M- bahia have benthic life stages.

 (Chronic toxicity tests using Q. myldss and P. reticulata fail to meet the test requirement of

 measured concentration for use in  deriving WQC.   Recently, an early life-stage test was

 successfully completed using rainbow trout, Q. mvkiss (Brooke, 1993). The acute-chronic ratio

 ACR, from this test (11.39) was almost identical to the value of 12.82 from unmeasured tests

 with this fish (Table 3-1; 3-2).  This new value will be added to this document following public

 comment. Time did not permit its inclusion in this draft.

       Dieldrin concentrations were not measured in freshwater tests. However, the nominal

 and measured concentrations in the salt water M- bahia chronic test differed by less than 20%

 at all concentrations.  One life x;ycle test has been conducted with Q. mvkiss  (Chadwick and
                                          >^
 Shumway 1969). There was a 97% reduction in survival and a 36% reduction in growth of the

 survivors  in 0.39 jtg/L relative to control fish; all fish died at 1.2 jtg/L. Q. mvkiss were not

 significantly affected at concentrations of 0.012 to 0.12 pg/L. No progeny were tested. The

 other freshwater chronic  test was a three-generation  study using the  guppy, P.  reticulata

 (Roelofs, 1971). Because exposure concentrations were increased from the test with the first

generation to the tests  with the  next two generations, and because there  was no effect at any

concentration in the first test, only results from the second two tests are  reported here (Table

3-2).  There was no effect on P. reticulata survival at dieldrin concentrations from 0.2 to 1.0

/tg/L. Mean brood size was reduced by 32% at 2.5 jtg/L.

      Saltwater M- bahia exposed to dieldrin in a life-cycle test were affected at concentrations

similar to those affecting the two freshwater fish mentioned above. M. bahia exposed to 1.1 and

1.6 /tg/L (U.S. EPA, 1980b) had a 35% and 58% reduction in survival, respectively, relative



                                        3-5

-------
3-6

-------
             r-
             in
             A
      01
             in
             A
                    «n
                    10
                    in
                    A
e

H


I
0
a
«
            M
    w
            e
            o
                   e
                   o
Mysid
i
ete
Pol
Ooh
                 3-7

-------
  to control_M.bjhja.  There were no significant effects at 0.10 to 0.49 jtg/L.  No effects were
                           j   '  '   *
  observed on reproduction at any concentration tested and progeny response was not recorded.

  One  life-cycle and one partial  life-cycle test were  conducted with the polychaete worm,

  .Ophryptrogha diadema (Hooftman and Vink, 1980; Tables 3-1 and 3-2). The observed nominal

  no effect concentration was of 0.1 ^g/L (below limit of analytical detection) for the life-cycle

  test initiated with larvae and 1.2 jtg/L (based on measured concentrations) for the partial life-

  cycle test initiated with adults.  For the life-cycle test with larvae there were 40, 37 81 and 99 %

 decreases in reproductive potential, (combined effect  on number of egg masses and embryo

 survival), relative to carrier control worms at 0.3,1.5,3.1 and 13 jtg/L, respectively. Embryo

 survival was reduced by 35, 16, 61 and 71 % at dieldrin concentrations of 0.3,1.5, 3.1 and 13

 pg/L, respectively. At 13 jig/L dieldrin survival was reduced to 34% relative to the controls.

 In the Q.diadema partial life-cycle test, reproductive potential was reduced by 57, 92, 97 and

 100% relative to the carrier control in concentrations of 2.6, 8, 23 and 72 pgfL. Sixty-three

 percent of adults in 72 /tg/L died. Reductions in egg survival were 39, 70, 62 and 100%

 relative to controls in concentrations of 2.6, 8, 23 and 72 jigfL, respectively.  The chronic

 sensitivity of this species appears similar to that of the other species tested chronically but acute

 sensitivity is low: 96 hr LC50 > 100 jtg/L for adults and larvae.

       The difference between  acute and chronic sensitivity to dieldrin for acutely sensitive

 species is approximately  an order-ofcmagnitude or less (Table 3-2).  The acute-chronic ratio

 (ACR) for acutely insensitive polychaetes was > 56.63 in one test and > 577.4 in a second.

The available ACRs for acutely sensitive species are 2.417 forP. reticulata, 6.129 for M. bahia

and 12.82 for.O. myMss.  The Final Acute-Chronic Ratio (ACR),  the geometric mean of these
                                         3-8

-------
 three values, is 5.748.                    -                 .

        Hie FCVs (Table 3-2), are used as the effect concentrations for calculating the SQC for

 benthic species.  The FCV for freshwater organisms of 0.0625 pg/L is the quotient of the FAV

 of 0.3595 ftg/L and the final ACR of 5.748. Similarly, the FCV for saltwater organisms of

 0.1147 ftg/L is the quotient of the FAV of 0.6594 jtg/L and the final ACR of 5.748.



 3.4 APPLICABILITY OF THE WATER QUALITY CRITERION AS THE EFFECTS
   CONCENTRATION FOR DERIVATION OF THE DffiLDRIN SEDIMENT
    QUALITY CRITERION:

       The use of the FCV (the chronic effects-based WQC concentration) as the effects

 concentration for calculation of the EqP-based SQC assumes that benthic (infaunal and

 epibenthic) species, taken as a group, have sensitivities similar to all benthic and water column

 species  tested to derive the WQC concentration.  Data supporting the reasonableness  of this

 assumption over all chemicals for which there are published or draft WQC documents are

 presented in Di Toro et al. (1991), and the SQC Technical Basis Document (U.S. EPA, 1993a).

 The conclusion of similarity of sensitivity is supported by comparisons between (1) acute values

 for the  most sensitive benthic species and acute values for the most sensitive water column

 species for all chemicals; (2) acute values for all benthic species and acute values for all species

 in the WQC documents across all chemicals after standardizing the LC50 values; (3) FAVs

 calculated for benthic species alone and FAVs calculated for all species in the WQC documents;

and (4)  individual chemical comparisons of benthic species vs. all species.  Only in this last

comparison are dieldrin-specific comparisons of the sensitivity of benthic and all (benthic and

water-column) species conducted. The following paragraphs examine the data on the similarity
                                        3-9

-------
   of sensitivity of benthic and all species for dieldrin.



          For dieldrin,'benthic species account for 12 out of 19 genera tested in fteshwater, and



   14 out of 21 genera tested in saltwater (Figures 3-1, 3-2),,  An initial test of the difference



  between the freshwater and saltwater FAVs for all species (water column and benthic) exposed



  to dieldrin was performed using the Approximate Randomization method (Noreen, 1989). The



  Approximate Randomization method tests the significance level of a test statistic when compared



  to a distribution of statistics generated from many random subsamples. The test statistic in this



  case is the difference between the freshwater FAV, computed from the freshwater (combined



  water column and benthic) species LC50 values, and the saltwater FAV, computed from the



  saltwater (combined water column and benthic) species LC50 values (Table 3-1).  In the



 Approximate Randomization method, the freshwater LC50 values and the saltwater LC50 values



 are combined into one data set. The data set is shuffled, them separated back so that randomly



 generated "freshwater" and "saltwater" FAVs can be computed. The LC50 values are separated



 back such that the number of LC50 values used to calculate the sample FAVs are the same as



 the number used to calculate the original FAVs.  These too FAVs are subtracted and the



 difference used as the sample statistic. This is done many  times so that the sample statistics



 make up a distribution that is representative of the population of FAV differences (Figure 3-3).



 Hie test statistic is compared to this distribution to determine it's level of significance. The null



 hypothesis is that the LC50 values that comprise the saltwater and freshwater data bases are not



 different.  If this is true, the difference between the actual freshwater and saltwater FAVs should



be common to the majority of randomly generated FAV differences.  For dieldrin, the test-



statistic falls at the 31 percentile of the generated FAV differences. Since the probability is less






                                       3-10

-------
  than 95%, the hypothesis of no significant difference in sensitivity for freshwater and saltwater
  species is accepted (Table 3-3).
        Since freshwater and saltwater species showed similar sensitivity, a test of difference in
  sensitivity for benthic and all (benthic and water column species combined, hereafter referred
 to as  "WQC") organisms combining freshwater and saltwater species using the Approximate
 Randomization method was  performed.  The test statistic in this case is the difference between
 the WQC FAV, computed from the WQC LC50 values, and the benthic FAV, computed from
 the benthic organism LC50 values.  This is slightly different then the previous test for saltwater
 and freshwater species. The difference is that saltwater and freshwater species in the first test
 represent two separate groups.  In this test the benthic organisms are a subset of the WQC
 organisms set. In the Approximate Randomization  method for this test, the number of data
 points coinciding with the number of benthic organisms are selected from the WQC data set.
 A "benthic" FAV is computed. The original WQC FAV and the "benthic" FAV  are then used
 to compute the difference statistic.  This is done many times and the distribution that results is
 representative of the population of FAV difference statistics.   The test statistic is compared to
 this distribution to  determine its level of significance. The probability distribution of the
 computed FAV differences are shown in the bottom panel of Figure 3-3. The test statistic for
 this analysis  falls at the 72  percentile and the hypothesis of no difference  in  sensitivity is
accepted (Table 3-3).  This analysis suggests that the FCV for dieldrin based on data from all
tested species is an appropriate effects concentration for benthic organisms.
                                        3-11

-------
  TABtE 3-3. RESULTS OF APPROXIMATE RANDOMIZATION TEST FOR
  THE EQUALITY OF THE FRESHWATER AND SALTWATER FAV
  DISTRIBUTIONS FOR DIELDRIN AND APPROXIMATE
  RANDOMIZATION TEST FOR THE EQUALITY OF BENTHIC AND
  COMBINED BENTHIC AND WATER COLUMN (WQC) FAV
                   DISTRIBUTIONS.

 Compar-
 ison     Habitat or Water Type1  AR Statistic"   Probability6

 Fresh.    Fresh (19)   Salt (21)      loiioS31
 vs Salt

 Bentbic   Bentbic (26) WQC (40)      0.090       72
 vs Water
 Column +
 Bentbic (WQC)
•Values in parentheses are the number of LC50 values used in the comparison.
 AR statistic = FAV difference between original compared groups.
"Probability that the theoretical AR statistic <. tiiiat the observed AR statistic given
 that the samples came from the same population.
                       3-12

-------
  6


  4


  8


  2


  1


  0


 -1


 -2



 -3
                                DIELDRIN
                                ~i—i—i—i—i—i—r
1111111  i i  "lull	1—i—r—i—i—i—i	inin i i  i

FRESHWATER VS SALTWATER
 TT



  O
O  H
                   tf*9
                0CD
               » Himn
            0.1
                  10  20
                                      60
                           80  90
                                                         89
                                                      99.9
 U
 O

 III
 cc^^
 Wsj
 4



 3



 2



 1



 0



 -1



•2


-3
             -   BENTHIC VS WQC
                         i  i  i  i  i—r
                                                 mil I
             -  o

             -o
              ' "iimi  i i u i
           0.1
                                                  ...
         10  20
                                     60
                                  80  90
                                                         99
                                             99.9
Figure 3-3.
                    PROBABILITY


 Robabffity distribution of FAV difference statistics to compare water-only data

 from freshwater vs. saltwater (upper panel) and benthic vs. WQC (lower panel)
                                3-13

-------

-------
                                       SECTION 4




   TOXIOTY OF DIELDRIN (ACTUAL AND PREDICTED): SEDIMENT EXPOSURE



4.1 TOXICITY OF DIELDRIN IN SEDIMENTS:

    The toxicity of dieldrin spiked into clean sediments has been tested with two saltwater

species (a polychaete and the sand shrimp) and two freshwater species (an amphipod and a
                                        •  *,
midge) (Table 4-1). Therefore, generalizations of dieldrin's toxicity across species or sediments

are limited. The endpoint reported in these studies was mortality.  Details about exposure

methodology are provided because, unlike aquatic toxicity tests, sediment testing methodologies

have not been standardized.  Data are available from many experiments using both field and

laboratory sediments contaminated with mixtures of dieldrin and other compounds.  Data from

these studies have not been included here because it is not possible to determine the contribution

of dieldrin to the observed toxicity.

      The only saltwater experiments that tested dieldrin-spiked sediments were conducted by

McLeese et al. (1982) and McLeese and Metcalfe (1980). These began with clean sediments

that were added to dieldrin-coated beakers just prior to the addition of test organisms. This is

in marked contrast to tests with freshwater sediments that were spiked with dieldrin days  or

weeks prior to test initiation.  As a result, the dieldrin concentrations in the sediment and
                                        4-1

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 overlying water varied greatly over the course of these saltwater experiments and exposure
                    ,.•'•*'
 conditions are uncertain.  In addition, transfer of test organisms to freshly prepared beakers

 every 48 hours further complicates interpretation of results of McLeese et al. (1982) because

 exposure conditions change several times during the course of the test.  McLeese et al. (1982)

 tested the effects of dieldrin on the polychaete worm, Nereis virens. in sediment with 2% TOC

 (17% sand and 83% silt and clay) in 12 day toxicity tests. No worms died in 13 /*g/g dry wt

 sediment, the highest concentration tested.  McLeese and Metcalfe (1980) tested the effects of

 dieldrin in sand with a TOC content of 0.28 % on the sand shrimp, Crangon septemspinosa.  The

 4 day LC50 was 0.0041 ug/g dry wt. sediment, and 1.46 ug/g^. Concentrations of dieldrin in

 water overlying  the sediment were 10 times the LC50 in water.  The authors conclude that

 sediment-associated dieldrin contributed little towards the toxicity observed.

    The effects of dieldrin-spiked sediments from three fresh-water sites in Minnesota on the

 fresh-water amphipod, Hvalella azteca have been studied by Hoke and Ankley (1991). The total

 organic carbon (TOC) concentrations in the three sediments were 1.7%, 2.9%, and 8.7%. The

 sediments were rolled in dieldrin-coated jars at 4°C for 23 days. Mortality of H. azteca in these

 flow-through tests was related to sediment exposure because dieldrin concentrations in overlying

 water were generally below detection limits.  There was no dose-response relationship observed

 in the results from the definitive test with one of the sediments (Airport Pond), or in the results

 from further testing with this sediment using H- azteca (Hoke and Ankley, 1992; Hoke 1992).

For this reason only the data from the range finder test with this sediment  are used in the

analysis of the toxicity data (sections 4.1, 4.2, 4.3), and in Figures 4-1 and 4-2. The ten-day

LCSO's increased with increasing TOC when dieldrin concentration was expressed on a dry
                                         4-3

-------
  weight basis,  but  increased  only  slightly  with increasing  organic  carbon when dieldrin
            '.<•''
  concentration was expressed on an organic carbon basis (Table 4-1).  LCSO's normalized to dry

  weight differed by a factor of 21.2 (18.2 to 386 jtg/g)over a 5..0 fold range of TOC.  In contrast,

  the organic carbon normalized LCSO's differed by a factor of 3.4 (1,073 to 3,682 Mg/goc)-

        The effects of dieldrin-spiked sediments from two freshwater sites in Minnesota on the

 fresh water midge, Chironomus tentans. have been studied by Hoke (1992). The TOC contents

 in the two sediments were 1.5 and 2.0%.  The sediments weie rolled in dieldrin coated jars at

 4°C for one month, stored at 4°C for two months, and then rolled at 4°C for an additional

 month. LCSOs normalized to dry weight differed by a factor of 2.89 (0.53 to 1.53 ng/g dry wt).

 LC50s normalized to organic carbon differed by a factor of 2.22 (35.33 to 78.46).  It is not
                                             v
 surprising that organic carbon normalization had little effect, given the small range of TOC (1.5

 to 2.0%).


    Overall,  the need for organic normalization of the concentration of nonionic organic

 chemicals in sediments is presented in the Technical Basis Document (U.S.EPA, 1993a).  The

 need for organic carbon normalization for dieldrin is supported by the dieldrin-spiked toxicity

 tests described above. Although it is important to demonstrate that organic carbon normalization

 is necessary if SQC are to be  developed using the EqP approach, it is fundamentally more

 important to demonstrate that KQC and water only effects concentrations can be used to predict

the effects concentration for dieldrin and other nonionic organic chemicals on an organic carbon

basis for a range of sediments. Evidence supporting this pirediction for dieldrin and other

nonionic organic chemicals follows in section 4.3.
                                         4-4

-------
 4.2    CORRELATION BETWEEN  ORGANISM  RESPONSE AND  PORE  WATER
        CONCENTRATION:

     One corollary of the EqP theory is that freely dissolved pore-water LCSQs for a given

 organism should be constant across sediments of varying organic carbon content (U.S.EPA,

 1993a). Appropriate pore-water values are available from two studies (Table 4-2). Data from

 tests with water column species were not considered in this analysis. Hoke and Ankley (1991)

 found 10-day LC50 values for H. azteca based on pore-water concentrations differed by a factor

 of 8.0 (57.6 to 458 jig/L) for three sediments containing from 1.7 to 8.7% TOC. Therefore,

 pore water normalized LC50 values provide only a slight improvement over LC50s for dieldrin

 expressed on a dry weight basis which varied by a factor of 21.2 (18.2 to 386 /tg/L).  Hoke

 (1992)  found  10-day  LC50  values  for  the,.£. tentans  based  on  predicted pore water

 concentrations (the sediment concentration multiplied by the KQC) differed by a factor of 2.17

 (0.23 to 0.50).  This variability is slightly less than that shown when dry wt (factor of 2.89) is

 used, but similar to that shown when organic carbon (factor  of 2.22) normalization is used.

 Partitioning to dissolved organic carbon was proposed to explain the lack of similarity of LC50

 values based on total pore water dieldrin concentrations.

      A more detailed evaluation of the degree to which the response of benthic organisms can

be predicted from toxic units of substances in pore water can be made utilizing results from

toxicity  tests with sediments  spiked with  other substances, including acenaphthene and

phenanthrene (Swartz,  1991), dieldrin (Hoke 1992), endrin (Nebeker et al., 1989; Schuytema

et al., 1989), fluoranthene (Swartz et al., 1990; DeWitt et al., 1992), or kepone (Adams et al.,

 1985) (Figure 4-1; Appendix B). The data included in this analysis come from tests conducted

at EPA laboratories or from tests which utilized designs at least as rigorous as those conducted


                                        4-5

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  at the EPA  laboratories.  Tests with acenaphthene and  phenanthrene used two  saltwater
                    ,.-•'"
  amphipods (Leptocheirus plumulosus and Eohaustorius estuarius't and marine sediments.  Tests


  with  fluoranthene used a saltwater amphipod (Rhepoxvnius abronius^ and marine sediments.


  Freshwater sediments spiked with  endrin were tested using the amphipod IL azteca:  while


 kepone-spiked sediments were tested using the midge, £,  tentans.  Figure 4-1 presents the


 percentage mortalities of the benthic species tested in individual treatments for each chemical


 versus "pore water toxic units"(PWTU) for all sediments tested. PWTUs are the concentration


 of the chemical in pore  water fctg/L) divided by the water only LC50 (pg/L).  Theoretically,


 50% mortality should occur at one interstitial water toxic unit. At concentrations below one


 PWTU there should be less  than 50% mortality, and at concentrations above one PWTU there


 should be greater than 50% mortality. Figure 4-1 shows that at concentrations below one PWTU


 mortality was generally low, and increased sharply at approximately one PWTU. Therefore this


 comparison supports the  concept that interstitial water concentrations can be used to predict the


 response of an organism to  a chemical that is  not sediment specific.   This  pore  water


 normalization was not used  to derive SQC in this document because of the complexation of


 nonionic organic chemicals with pore water DOC (Section 2) and the difficulties of adequately

 sampling pore waters.



 4.3 TESTS OF THE EQUILIBRIUM PARTITIONING PREDICTION OF SEDIMENT
    TOXICITY:


       SQC derived using the EqP approach utilize partition coefficients and FCVs from WQC


documents to derive the SQC concentration for protection of benthic organisms. The partition


coefficient (KQC) is used to normalize sediment concentrations and predict biologically available
                                        4-7

-------
 concentrations across sediment types. The data required to test the organic carbon normalization

                    •  •    '       '         '                        '
 for dieldrin in sediments are available for 2 benthic species. Data from tests with water column


 species were not included in this analysis. Testing of this component of SQG derivation requires


 three elements: (1) a water-only effect concentration, such as a 10-day LC50 value in j*g/L; (2)


 an identical sediment effect concentration on an organic caitbon basis, such as a 10-day LC50


 value in jig/god and (3) a partition coefficient for the chemical, KQC in L/Kgoc. This section


 presents evidence that the observed effect concentration  iin sediments (2) can be predicted


 utilizing the water effect concentration  (1) and the partition coefficient (3).


    Predicted ten-day LC50 values from dieldrin-spiked sediment tests with IL azteca (Hoke and


 Ankley, 1991) wese calculated (Table 4-2) using the logw KOC value of 5.25 from Section 2 of
                                           >f
 this document and the water-only LC50 value (7.3 ug/L).  Batios of actual to predicted LCSO's


 for dieldrin averaged 1.26 (range 0.827 to 2.83) in tests with three sediments (Table 4-2).


 Similarly, predicted 10-day LC50 values for dieldrin-spiked isedimenf tests with C. tentans were


 calculated using the logw KQC of  5.25 and a 10-day water only LC50 value of 0.29 jig/L.


 Ratios of predicted to actual LCSOs for dieldrin averaged 1.02 (range 0.69 to 1.52) in tests with


 two sediments (Table 4-2). The overall mean for both species was 1.16.


      A more detailed evaluation of the accuracy and precision of the EqP prediction of the


 response of benthic organisms can be made using the  results of toxicity tests with amphipods


 exposed to sediments spiked with acenaphthene, phenanthrene, dieldrin, endrin, or fluoranthene.


The data included in this analysis came from tests conducted at EPA laboratories or from tests


which utilized designs at least as  rigorous as those conducted at the EPA laboratories.  Data


from the kepone experiments are not included because a measured K^ for kepone obtained using
                                          4-8

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   the slow stir flask method is not available. Swartz (1991) exposed the saltwater amphipods £,


  SSfo^rius and'L. plumulpsus to acenaphthene in three marines sediments having organic carbon


  contents ranging from 0.82 to 4.2% and to phenanthrene fin three marine sediments having


  organic carbon contents ranging from 0.82 to 3.6%.  Swartz et al. (1990) exposed the saltwater
                                                               j

  amphipod E, abronius to fluoranthene in three marine sediments having 0.18, 0.31 and 0.48%


  organic carbon.  Hoke and Ankley (1991) exposed the ampMpod g, azteca to three dieldrin-


  spiked freshwater sediments having 1.7, 2.9 and 8.7% organic carbon and Hoke (1992) exposed


 the midge £. tentans to two freshwater dieldrin-spiked sediments having 2.0 and 1.5 % organic


 carbon. Nebeker et al. (1989) and Schuytema et al. (1989) exposed H-  azteca to three endrin-


 spiked sediments having 3.0, 6.1 and 11.2% organic carbon. Figure 4-2presents thepercentage


 mortalities of amphipods in individual treatments of each chemical versus "predicted sediment


 toxic units" (PSTU) for each sediment treatment. PSTUs are the concentration of the chemical


 in sediments fcg/goc) divided by the predicted LC50 Otg/goc) in sediments (the product of KOC


 and the 10-day water-only LC50).  In this normalization, 50% mortality should occur at one


 PSTU. At concentrations below  one PSTU mortality was generally low, and increased sharply


 at one PSTU. The means of the LC50s for these tests calculated on a PSTU basis were 1.90,


 for acenaphthene, 1.16 for dieldrin,  0.44 for endrin, 0.80 for fluoranthene, and 1.22 for


phenanthrene.  The  mean value  for the five chemicals is 0.99.  This illustrates that the EqP


method can account for the effects of different sediment properties and properly predict the


effects concentration in sediments using the effects concentration from water only exposures.
                                        4-10

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                  ','.••••"  SECTIONS

                     CRITERIA DERIVATION FOR DffiLDRIN

 5.1 CRITERIA DERIVATION:
       The water quality criteria (WQC) Final Chronic Value (FCV), without an averaging
 period or return frequency (See section 3), is used to calculate of sediment quality criteria (SQC)
 because it is probable that the concentration of contaminants in sediments are relatively stable
 over time, thus exposure to sedentary benthic species should be chronic and relatively constant.
 This is in contrast to the situation in the water column, where a rapid change in exposure and
 exposures of limited durations can occur due to fluctuations in effluent concentrations, dilutions
 in receiving waters or the free-swimming or planktonic nature of water column organisms.  For
 some particular uses of the SQC it may be appropriate to use the area! extent and  vertical
 stratification of contamination of a sediment at a site in much the same way that averaging
periods or mixing zones are used with WQC.
      The  FCV is the value that should protect 95% of the tested species included in the
calculation of the WQC from chronic effects of the substance. The FCV is the quotient of the
Final Acute Value (FAV), and the final Acute Chronic Ratio (ACR) for the substance.  The
FAV is an estimate of the acute LC50 or EC50 concentration of the substance corresponding to
a cumulative probability of 0.05 for the genera from eight or more families for which acceptable
acute tests have been conducted on the substance.  The ACR is the mean ratio of acute to
Chronic toxicity for three or more species exposed to the substance that meets minimum database
                                       5-1

-------
 requirements.  For more information on the calculation of ACRs, FAVs, and FCVs see the

 National Water Quality  Criteria Guidelines (Stephan et al., 1985).  The FCV used in this

 document differs from the FCV in the dieldrin WQC document (U.S. EPA, 1980) because it

 incorporates recent data not included in that document, and omits some data which does not meet

 the data requirements established in the WQC Guidelines (Stephan et al., 1985).

       The equilibrium partitioning (EqP) method for calculating SQC is based on the following

 procedure.  If FCV (pg/L) is the chronic concentration from the WQC  for the chemical of

 interest, then the SQC 0*g/g sediment),  is computed using the partition coefficient, KP (L/g

 sediment), between sediment and pore water:

             SQC = KP FCV                                                 (5_!)

       Since organic carbon is the predominant sorption phase for nonionic organic chemicals

 in  naturally occurring sediments,  (salinity, grainsize and  other sediment parameters have

 inconsequential roles  in  sorption,  see sections 2.1 and 4.3)  the organic carbon partition

 coefficient, (Koc) can be substituted for KP. Therefore, on a siediment organic carbon basis, the

              , is:
             SQCOC = KOCFCV                                              (5_2)

Since (K^) is presumably independent of sediment type for non-ionic organic chemicals, so also
                                                  \
is SQCoc. Table 5-1 contains the calculation of the dieldrin SQC.

      The organic carbon normalized SQC is applicable to sediments with an organic carbon
                                                /
fraction of foe ^  0.2%.  For sediments with f^, < 0.2%, organic carbon normalization and

SQC may not apply.
                                        5-2

-------
             TABLE 5-1. SEDIMENT QUALITY CRITERIA FOR DffiLDRIN
Type of
Water Body
Fresh Water
Salt Water
(L/kg) W
5.34
5.34
(L/kg)
5.25
5.25
FCV
0*g/L)
0.0625
0.1147
SQCoc
0*g/goc)
11*
20"
            = (105-25 L/kgocWlO-* kgoc/gocKO.0625 ug dieldrin/L) = 11 M



    "SQCoc = (105-25 L/kgocWIO-* kgoc/goc)»(0.1147 jig dieldrin/L) - 20 pg di




        Since organic carbon is the factor controlling the bioavailability of nonionic organic



 compounds in sediments, SQC have been developed on an organic carbon basis, not on a dry



 weight basis.   When the chemical concentrations in sediments are reported as dry weight



 concentration  and organic  carbon  data are available,  it is best to  convert the sediment



 concentration to pg chemical/gram organic carbon. These concentrations can then be directly



 compared to the SQC  value.  This facilitates comparisons  between the  SQC and field



 concentrations relative to identification of hot spots and  the degree to which sediment



 concentrations do or do not exceed SQC values.  The conversion from dry weight to organic



 carbon normalized concentration can be done using the following formula:



       Ug Chemical/goc = A*g Chemical/gDRywT  -~- (% TOC  -^ 100)



                       = ftg Chemical/gDRYWT • 100 -5- %  TOC



       For example, afreshwater sediment with a concentration of 0.1 (tg chemical/gDRYWT and



0.5% TOC has an organic carbon-normalized concentration of 20 /tg/goc (0.1 ng/gDKrw[ • 100



- 0.5  = 20 Mg/goc) which exceeds the freshwater SQC of 11 ^g/g^.  Another freshwater



sediment with the same concentration of dieldrin (0.1 Mg/gDRY w) but a TOC concentration of




                                       5-3

-------
  5.0% would have an organic carbon normalized concentration of 2.0 /ig/goc (0.1 /«g/gDRYWT *
                     i  •    •
  100 -s- 5.0 = 2.0 /*g/goc), which is below the SQC for dieldrin.


        In situations where TOC values for particular sediments are not available, a range of

  TOC values may be used in a "worst case" or "best case" analysis.  In this case, the organic

  carbon-normalized SQC values (SQCoc) may be "converted"  to dry weight-normalized SQC

 values (SQCDRY WT.).  This "conversion" must be done for each level of TOC of interest:


              SQCDRywT  = SQCoc Gtg/goc) • (% TOC * 100)

 where SQCDRYWT is the dry weight normalized SQC value. For example, the SQC value for

 freshwater sediments with 1% organic carbon is 0.11 /xg/g:


              SQCDRYWT. = 11 jtg/goc • 1% TOC -5- 100 == 0.11 /*g/gDRYWT
                                            i
 This method is used in the analysis of the STORET data in section 5.4.

 5.2 UNCERTAINTY ANALYSIS:


    Some of the uncertainty in the calculation of the dieldrirt SQC can be estimated from the

 degree to which the EqP model, which is the basis for the criteria, can rationalize the available

 sediment toxicity data. The EqP model asserts that (1) the bioavailability of nonionic organic

 chemicals from  sediments  is equal on an organic carbon basis, and  (2) that the  effects

 concentration in sediment Otg/goc) can be estimated from thepiraduct of the effects concentration

from water only exposures  0*g/L) and  the partition coefficient KQC (L/kg). The uncertainty

associated with the SQC can be obtained from a quantitative estimate of the degree to which the

available data support these assertions.


    The data used in the uncertainty analysis are from the water-only and sediment toxicity tests

that have been conducted to fulfill the minimum database requirements for the development of
                                         5-4

-------
  SQC (See Section 4.3 and Technical Basis Document, U.S. EPA, 1993a). These freshwater and
  saltwater tests span a range of chemicals and organisms; they include both water-only and
  sediment exposures  and they are replicated within each chemical-organism-exposure media
  treatment. These data were analyzed using an analysis of variance (ANOVA) to estimate the
  uncertainty (i.e. the variance) associated  with varying the exposure media and that associated
  with experimental error.  If the EqP model were perfect, then there would be only experimental
  error.  Therefore, the uncertainty associated with the use of EqP is the variance associated with
  varying exposure  media.
     The data used in the uncertainty analysis are illustrated in Figure 4-2. The data for dieldrin
 are summarized in Appendix B. LCSOs for sediment and water-only tests were computed from
                                            i^
 these data. The EqP model can be used to normalize the data in order to put it on a common
 basis. The LCSOs from water-only exposures (LC50*; ftg/L) are related to the organic carbon-
 normalized LCSOs from sediment exposures (LC50S(OC; ^g/g^ via the partitioning equation:
             LC50S>OC =  oc                                            (5.3)

The EqP model asserts that the toxicity of sediments expressed on an organic carbon basis equals
the toxicity in water tests multiplied by the KOC. Therefore, both LC50SfOC and KOC»LC50W are
estimates of the true LCSOoc for each chemical-organism pair.  In this analysis, the uncertainty
of KOC is not treated separately.   Any error associated with K^ will be reflected in the
uncertainty attributed to varying the exposure media.
    In order to perform an analysis of variance, a model of the random variations is required.
                                        5-5

-------
 As discussed above, experiments that seek to validate equation 5-3 are subject to various sources
                    ,.'''"'       '                      '
 of random variations.  A number of chemicals and organisms! have been tested. Each chemical -

  organism pair was tested in water-only exposures and in different sediments. Let a represent

 the random variation due to this source. Also, each experiment is replicated.  Let € represent

 the random variation due to this source. If the model were perfect, there would be no random

 variations other than that due to experimental error which is reflected in the replications. Hence

 a represents the uncertainty due to the approximations inherent in the model and  € represents

 the experimental error.  Let (trj2 and (rresponding to a water-only or

 sediment exposure; ^ are the population of ln(LC50) for chemical-organism pair i.  The error

 structure  is assumed to be lognormal which corresponds to assuming that the  errors are

proportional to the means, e.g. 20%, rather than absolute quantities, e.g. 1 jtg/L. The statistical

problem is to estimate /iis (oj2, and (
-------
         Table 5-2: ANALYSIS OF VARIANCE FOR DERIVATION OF
         SEDIMENT QUALITY CRITERIA CONFIDENCE LIMITS FOR
                                   DIELDRIN.
Source of Uncertainty
Exposure media
Replication
Sediment Quality Criteria
Parameter Value
  0.2%. For sediments with r^, < 0.2%,  organic carbon normalization and

SQC do not apply.
                                      5-7

-------
                  TABLE 5-3. SEDIMENT QUALITY CRITERIA
                      •  CONFIDENCE LIMITS FOR DIELDRIN
Sediment Quality Criteria
95% Confidence Limits fW/«r__^
Type of
Water Body
Fresh Water
Salt Water
p j^\^i
*3>^^*Oi"^
A*g/goc
11
20
Lower
5.2
9.5
Upper
24
44
• 5.3   COMPARISON OF DIELDRIN SQC AND UNCERTAINTY CONCENTRATIONS TO
       SEDIMENT CONCENTRATIONS THAT ARE TOXIC  OR PREDICTED TO BE
       CHRONICALLY ACCEPTABLE.

       Insight into the magnitude of protection afforded  to  benthic species  by SQC

 concentrations and 95% confidence intervals can be inferred using effect concentrations from

 toxicity tests with benthic species exposed to sediments spiked with dieldrin and sediment

 concentrations predicted to be chronically safe to organisms tested jn water-only exposures

 (Figures 5-1 and 5-2).  Effect concentrations in sediments can be predicted from water-only

 toxicity data and KOC values  (See Section  4).  Chronically acceptable concentrations are

 extrapolated from genus mean acute value (GMAV) from water-only, 96-hour lethality tests

 using acute-chronic ratios (ACR). Therefore, it  may be reasonable  to  combine these two

 predictive procedures to estimate, for dieldrin, chronically acceptable sediment concentrations

 (Predicted Genus Mean Chronic Value, PGMCV)) from GMAVs (Appendix A), ACRs (Table

 3-2) and the KOC (Table 5-1):

             PGMCV = (GMAV -^ ACR)* K^                              (5.7)


                                       5-8

-------
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: A
a
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Arthropod
Other invertebrate
Fishes A A
Log^Koc-525
ACR.5.75 A A
SEDIMENTTESTS: lOdLCSO T
* C. tentans • 56.9ua/goc 1
range 2 tests -35.3 to 78.5»ig/goc ©
® H.,azteea » IddOfig/ggg -^
range 3 tests- 1070 to 3680 Q A A
A O 0
~ A • O A
. o u ~ "
I

A*





-; — i —
1

uppen24ng/goc


- .-----..-„..„«....„„„„„ lower. S^g/goc


— ' 	 -1 — — «- 	 • ' « • ,
                       20         40         60         80        100

                   PERCENTAGE RANK OF FRESHWATER GENERA
Figure 5-1.
Comparison between SQC concentrations and 95% confidence intervals, effect
concentrations from benthic organisms exposed to dieldrin-spiked sediments and
sediment concentrations predicted to be chronically safe in fresh water sediments
SZ^^**0!?15° * °hronicaUy «* (Predicted Genus Mean Chronic
Values, PGMCV) are denved from the Genus Mean Acute Values (GMAV) from

S?V - £SfJ8?1??£?• ACUtC Chn)niC ""^ (ACR) ^ ** values.
PGMCy = (GMAV + ACR)Koc. Symbols for PGMCVs are A for arthropods
O for fishes and D for other invertebrates. Solid symbols are benthic genenr
open symbols water column genera. Arrows indicate greater than values  Error
bars around sediment LC50 values indicate observed range of LC50s
                                      5-9

-------
            10'
      o

     I
     ill

     I
     O
     X
     o
     CO
     z
     HI
     o
     o
     01
     Q
     UJ
     CC
     Q_
10'
           102
10
         Water-only tests;  (96HR LC50 •*• ACR)

         A Arthropod
         D Other Invertebrate
         O Fishes
           AC R- 5.75
           upper:44fig/goc

           • SQC:

            lower: Q.
                        20
                        40
60
80
100
                   PERCENTAGE RANK OF SALTWATER GENERA
Figure 5-2.   Comparison between SQC concentrations and 95% confidence intervals, effect
             concentrations from benthic organisms exposed to dieldrin-spiked sediments and
             sediment concentrations predicted to be chronically safe in salt water sediments.
             Concentrations predicted to be chronically safe (Predicted Genus Mean Chronic
             Values, PGMCV) are derived from the Genus Mean Acute Values (GMAV) from
             water-only 96-hour lethality tests, Acute Chronic Ratios (ACR) and K— values
             PGMCV = (GMAV •*• ACR)Koc. Symbols for PGMCVs are A for arthropods,
             O for fishes and D for other invertebrates.  Solid symbols are benthic genera;
             open symbols water column genera. Arrows indicate greater than values. Error
             bars around sediment LC50 values indicate observed range  of LC50s.
                                      5-10

-------
         In Figures 5-1 and 5-2 each PGMCV for fishes, arthropods or other invertebrates tested
  in water is plotted against the percentage rank of its sensitivity.  Results from toxicity tests with
  benthic organisms exposed to sediments spiked with dieldrin  (Table 4-1) are placed  in the
  PGMCV rank appropriate to the test-specific effect concentration.  (For example, the 10-day
  LC50 for H. izteca, (1,640 ^g/goc) is placed between the PGMCV of 742 fig/g^ for the
  stonefly, Acroneuria,  and the PGMCV of 6,605 pg/g^ for the cladoceran, SJr
  Therefore, LC50  or other effect  concentrations  are intermingled  in  this  figure with
  concentrations predicted to be  chronically  safe.  Care should be taken by  the reader in
  interpreting these data with dissimilar endpoints. The following discussion of SQC, organism
  sensitivities and PGMCVs is not intended to provide accurate predictions of the responses of taxa
  or communities of benthic organisms relative to specific concentrations of dieldrin in sediments
 in the field. It is, however/intended to guide scientists and managers through the complexity
 of available data relative to potential risks to benthic taxa posed by sediments contaminated with
 dieldrin.
       The freshwater SQC for dieldrin (11 ^g/g^ is less than any of the PGMCVs or LC50
 values from spiked sediment toxicity tests.  The PGMCVs for 17 of 19 freshwater genera are
 greater than the upper 95% confidence interval of the SQC (23 pg/g^. The PGMCVs for the
 stonefly Bejonarcella (15 ^g/goc) and Claassenia (18 Mg/goc) are below the SQC upper 95%
 confidence interval.   This illustrates why the slope of the species sensitivity distribution is
 important. It also suggests that if the extrapolation from water only acute lethality tests to
 chronically acceptable sediment concentrations is accurate, these or similarly sensitive genera
may be chronically impacted by sediment concentrations marginally above the SQC and possibly
                                       5-11

-------
  less than the 95% upper confidence interval. For dieldrin, PGMGVs range over three orders

                          »   •  .   - •            •                 •

  of magnitude 'from the most sensitive to the most tolerant genus. A sediment concentration 20



  times the SQC would include the GMCVs of one-half of the 12 benthic genera tested including



  stoneflies, mayflies, isopods and  catfish.  Tolerant benthiic genera such as the amphipod



  •
-------
   assessment of the concentrations of dieldrin in the sediments of the nation's water bodies. Log
   probability plots of dieldrin concentrations on a dry weight basis in sediments are shown in
   Figure 5-3. Dieldrin is found at varying concentrations in sediments from rivers, lakes and near
   coastal water bodies in theTJnited States.  This is due to its widespread use and quantity applied
   during the 1960s and early 1970s.  It was restricted from register and production in the
   United States in 1974.  Median concentrations are generally at or near detection limits in most
   water bodies for data after 1986. There is significant variability with dieldrin concentrations in
  sediments ranging over nine orders of magnitude within the country.
        lie SQC for dieldrin can be compared to existing concentrations of dieMrta in sediments
  of natural water systems in the United States as contained in the STORET database (U.S. EPA,
  1989b).  These data are generaUy reported on a dry weight basis, rather than an organic carbon
  normalized basis.  Therefore, SQC values corresponding to sediment organic carbon levels of
  1 to 10SS are compared to dieldrin's distribution in sediments as examples only. For fresh
 water sediments, SQC values are 0.11 Mg/g dry weigh, in sediments having 1 * organic carbon
 and 1.1 «fe dry weight in sediments having 10% organic carbon; for marine sediments SQC
 are 0.20 nfe dry weigh, and 2.0 Mg/g, dry weight respectively.  Figure 5-3 presents the
 comparisons of these SQC to probability distributions of observed sediment dieldrin levels for
 steams and lakes (fresh water systems, shown on the upper panels) and estuaries (marine
 systems, lowerpanel). For both streams (n = 3075) and lakes (n ^ 457), both the SQC of 0.11
 n/g dry weight for 1%  organic carbon fresh water sediments and the SQC of 1.1 ,,g/g dry
weight for 10% organic carbon fresh water sediments are exceeded by less than 1 % of the data.
m estuaries, the dab (n=160) indicate tha, neither criteria, 0.20 ug/g dry weight for sediments
                                         5-13

-------
     Uj5»
     Wfc

                             •"TIT]	1	1—i—i—i—r-
                              TOTAL SAMPLES:  3075
                              MEASURED  SAMPLES:  590
                                                  J_
                                                  80  90
                                           Jin 11.
                                                     99
                                                                     99.9
10

10
                  'TOTAL SAMPLES:' <57'  '
                  MEASURED SAMPLES:  124
                                                       p"""  '  """"
                                                  1    '••[•« •  • •  •
                                                 80   90
                                                               99
                                                          99.9
            10'
            10'
| ESTUARY
"""1 	 1 	 1 — i — | — i — r — | — inn
TOTAL SAMPLES: 160
MEASURED SAMPLES: 3

                 ' •	'  	in
                              10  20
                              50


                                    EIO  90


                                                              99 .   99.9
Figure 5-3.
                      PROBABILITY





on each figure represents the SQC vlfwhe^roc - iST 27    ? ?"
line represents the SQC when TOC ™%         ~     ' ^lower da*ed
                        5-14

-------
   having 1 % organic carbon or 2.0 Mg/g dry weight for sediments having 10% organic carbon are
   exceeded by the post '1986 samples.  Concentrations of dieldrin in sediments from estuaries are
   two order of magnitude below the SQC value for 1 % organic carbon sediments and three orders
.   of magnitude below the SQC value for sediments with TOCs of 10 %.
         Hie dieldrin distribution in Figure 5-3 includes data from some samples in which the
  dieldrin concentration was below the detection limit.  These data are indicated on the plot as
  "less than"  symbols «), and plotted at the reported  detection limits.  Because these values
  represent upper bounds and not measured values the percentage of samples in which the SQC
  values are actually exceeded may be less than the percentage reported.
        A second database developed as part of the National Status and Trends Program (NOAA,
  1991)  is  also available for assessing  contairdnant levels in  marine sediments that  are
  representative  of areas away from sources  of contamination.  The probability distribution for
  these data,  which can be directly expressed on  an organic carbon basis, is compared to the
  saltwater SQC for dieldrin (20 ^g/g^ on Figure 5-4. Data presented are from sediments with
 0.20 to 31.9  percent organic  carbon.   The median organic carbon normalized dieldrin
 concentration (0.08 Atg/goc) is 2 orders of magnitude below the SQC of 20 Mgoc. Noneofthese
 samples (n=408) exceeded the criteria. Hence, these results are consistent with the preceding
 comparison of the marine SQC to STORE! data.
       Regional  differences in  dieldrin concentrations may affect the above  conclusions
concerning expected criteria exceedences.  This analysis also does not consider other factors
such as the type of samples collected (i.e., whether samples were from surficial grab samples
or vertical core profiles), or the relative frequencies and intensities of sampling in different study
                                         5-15

-------
     i iimni i  mini i  nunii

          0
     i     o
uiiii 111 uiiii 11
mm i  nun 111  •mm , pi,,,,,
       (OO 0/On) 1N3WIQ38
                       -8 §
                                             «2  U
                                             p "S ex
                                             o § eo


                                            ill
                                            3M J
                                            •8
                                            •-3

                                            •8
                                              S
                                            o o\
                                            S ^
                                            8 2
                    t
                    «n
                    5-16

-------
   areas.  It is presented as  an aid in assessing the range of reported dieldrin sediment
   concentrations and the extent to which they may exceed the SQC.

  5.5 UMTTATIONS TO THE APPLICABILITY OF SEDIMENT QUALITY CRITERIA:
        Rarely, if ever,  are  contaminants found alone  in naturally occurring  sediments.
  Obviously, the fact that the concentration of a particular contaminant does not exceed the SQC
  does not mean that other chemicals, for which there are no SQC available, are not present in
  concentrations sufficient to cause harmful effects.  Furthermore, even if SQC were available for
  all of the contaminants in a particular sediment, there might be additive or synergistic effects
  that the criteria do not address.  In this sense the SQC represent "best case" criteria.
        It is theoretically possible that antagonistic reactions between chemicals could reduce the
 toxiciry of a given chemical  such  that it might not cause unacceptable effects  on benthic
 organisms at concentrations above the SQC when it occurs with the antagonistic chemical.
 However, antagonism has rarely been demonstrated. What should be much more common are
 instances where toxic  effects occur at concentrations below the SQC because of the additivity
 of toxicity of many common contaminants (Alabaster and Lloyd, 1982), e.g. heavy metals and
 PAHs, and instances where other toxic compounds for which no SQC exist occur along with
 SQC chemicals.
       Care must be used  in application of EqP-based SQC  in disequilibrium conditions.  In
some  instances site-specific SQC may be required to address this condition. EqP-based SQC
assume that nonionic organic chemicals are in equilibrium with the sediment and IW and are
associated with sediment primarily through adsorption into sediment organic carbon.  In order
                                        5-17

-------
 for these assumptions to be valid, the chemical must be dissolved in IW and partitioned into
 sediment organic carbon.  The chemical must, therefore, be associated with the sediment for a
 sufficient length of time for equilibrium to be reached.  In sediments where particles of
 undissolved dieldrin occur, disequilibrium exists and criteria are over protective.   In liquid
 chemical spill situations disequilibrium concentrations in interstitial and overlying water may be
 proportionately  higher relative to sediment concentrations.   In this case criteria  may   be
 underprotective.
       In very  dynamic  areas, with highly erosional or depositional bedded  sediments,
 equilibrium may not be attained with contaminants.  However, even high KOW nonionic organic
 compounds come to equilibrium in clean sediment in  a period of days,  weeks or months.
 Equilibrium times are shorter for mixtures of two  sediments, each previously at equilibrium.
 This is particularly relevant in tidal situations where  large volumes of sediments are eroded and
 deposited, yet near equilibrium conditions may predominate over large areas. Except for spills
and particulate chemical, near equilibrium is  the rule and disequilibrium is uncommon.  In
instances where it is suspected that EqP does  not apply for  a particular sediment because of
disequilibrium discussed above, site-specific methodologies may be applied (U.S. EPA, 1993b).
                                        5-18

-------
                                      SECTION 6

                                CRITERIA STATEMENT



        The procedures described in the "Technical Basis for Deriving National Sediment Quality

 Criteria for Nonionic Organic Contaminants for the Protection of Benthic Organisms by Using

 Equilibrium  Partitioning"  (U.S. EPA, 1993a)  indicate that benthic organisms  should be

 acceptably protected in freshwater sediments containing .<. 11 /tg dieldrin/g organic carbon and

 saltwater sediments containing  <. 20 /tg dieldrin/g organic carbon, except possibly where a

 locally important species is very sensitive or sediment organic carbon is  < 0.2%.
                                           •„
        Confidence limits of 5.2 to 24 jtg/gbc for freshwater sediments and 9.5 to 44 jtg/goc for

 saltwater sediments are provided as an estimate of the uncertainty associated with the degree to

 which the observed concentration in sediment fag/goc), which may be toxic, can be predicted

 using the organic carbon partition coefficient (K^ and the water-only effects concentration.

 Confidence limits do  not incorporate uncertainty associated with water quality criteria.  An

 understanding of the theoretical basis of the equilibrium partitioning methodology, uncertainty,

 the partitioning and toxicity of dieldrin, and sound judgement are required in the regulatory use

 of SQC and their confidence limits.

       These concentrations represent the U.S. EPA's best judgement at this time of the levels

of dieldrin in sediments that would be protective of benthic species. It is the philosophy of the

Agency and the EPA Science Advisory Board that the use of sediment  quality criteria (SQCs)

as stand-alone, pass-fail criteria is not recommended for all applications and should frequently


                                         6-1

-------
 trigger additional studies at sites under investigation. The upper confidence limit should be
 interpreted as a concentration above which impacts on benthic species should be expected.
 Conversely, the lower confidence limit should be interpreted as a concentration below which
impacts on benthic species should be unlikely.
                                       6-2

-------
                                      SECTION 7
                                    REFERENCES

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 Adema, D.M.M.  1978. Dahpnia magna as a test animal in acute and chronic toxicity tests
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 Betsffl, J.D.  1990.  The  sorption of hydrophobie organic  compounds in the presence  of
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Brooke, L.T.  1993.  Conducting toxicity tests with freshwater organisms exposed to dieldrin
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Butler, P.A.  1963. Commercial Fisheries Investigations,  tt: Pesticide and Wildlife Studies:


                                        7-1

-------
        A Review of Fish and Wildlife Service Investigations During 1961 and 1962. U.S  Fish
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                                           *

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-------
        Toxicology and Chemistry 10:(12)1541-1583.
                    »•''':"'       :-;   -
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 Henderson, C., et al.  1959. Relative toxicity of ten chlorinated huydrocarbon insecticides to
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 Hoke,  R. and G. Ankley, 1991.   Results of dieldrin sediment  spiking study conducted  in
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Hoke,  R. and G.  Ankley, 1992.   Results  of re-test  of Airport  Pond  dieldrin-spiked
       sediments.  Memorandum to D. Hansen, D. Di Toro. January 27.  8pp
                                        7-3

-------
  Hoqftman, R.N. and G. J. Vink, 1980. The determination of toxic effects of pollutants with the
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                                       7-4

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 Nebeker, A.V., G.S. Schuytema, W.L. Griffis, J.A. Barbitta, and L.A. Carey. 1989. Effect
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                                        7-5

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       June 26, 1991.  160 pp.
                 ',.>'•"'
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 Taizwell, C.M. and C. Henderson. 1957. Toxicity of dieldrin to fish. Trans. Am. Fish. Soc
       86:245.

 U.S. Environmental Protection Agency. 1980a. Ambient water quality criteria aldrin/dieldrin.
       Office of Water Regulations and Standards, Criteria and Standards Division. U S  EPA
       Washington,  D.C. EPA 440/5-80-019.

 U.S. Environmental Protection Agency.  1980b.  Unpublished laboratory data.  Environment
       Research Laboratory, Gulf Breeze, Florida.  3pp.

 U.S. Environmental Protection Agency. 1985.  Appendix B - Response to public comments on
       "Guidelines for deriving numerical national water quality criteria for the protection of
       aquatic organisms and their uses." July 19, 1985. Fed. Regist. 50:30793-30796.

 U.S. Environmental Protection Agency. 1987.  Quality Criteria for Water, 1986.  EPA 440/5-
       86-001. May 1, 1987. U.S. Government Printing Office No.  955-002-000008.  406pp.

 U.S. Environmental Protection Agency. 1989a. Sediment classification methods compendium.
       Watershed Protection Division, U.S. EPA. 280 pp.

U.S. Environmental Protection Agency. 1989b.  "Handbook: Water Quality Control Information
       System, STORET," Washington, D.C., 20406.

U.S. Environmental Protection Agency. 1989c. Briefing Report to the EPA Science Advisory
       Board  on the Equilibrium Partitioning Approach to Generating Sediment Quality
       Criteria.  Office of Water Regulations and Standards, Criteria and Standards Division
       132 pp.

U.S. Environmental Protection Agency. 1993a.  Technical Basis for Deriving Sediment Quality
       Criteria for Nonionic Organic Contaminants for the Protection of Benthic Organisms by
       Using Equilibrium Partitioning.  (In Review).

U.S. Environmental Protection Agency.   1993b.    Guidelines  for Deriving Site-Specific


                                       7-6

-------
       Sediment Quality Criteria for the Protection of Benthic Organisms.  (In Review).



           ?rt?1^0?C^n^Cy- 1993C- GuidefortheUseandAppHcationof Sediment
           ity Cntena for Nonionic Organic Chemicals. (In Review).
Wang, Y.S., 1988. The contamination and bioconcentration of aldrin, dieldrin and endrin in

      lower lakes at Rocky Mountain Arsenal.  Ph.D. Thesis, Colorado State
                                      7-7

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