R
SEDIMENT CLASSIFICATION METHODS
                 COMPENDIUM
                      Prepared by
              U.S. Environmental Protection Agency
             Sediment Oversight Technical Committee
               EPA Work Assignment Managers

               Beverly Baker and Michael Eravitz
               Office of Science and Technology
                   Washington, DC 20460

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                            ACKNOWLEDGMENTS
This document was prepared by the U.S. Environmental Protection Agency Sediment Oversight
Technical Committee.  The Sediment Oversight Technical Committee, chaired by Dr. Elizabeth
Southerland .of the Office of Science  and Technology, has representation from a  number of
Program Offices in Headquarters and the Regions.

Appreciation is extended to the authors of each chapter contained in this document.  Critical
reviews of portions of the document were provided by the following persons:  G. Allen Burton,
Jr., Tom Chase, Rick Fox, Audrey Massa, George Schupp, and Howard Zar.

Assistance in preparation and production of the Compendium was provided under EPA, Contract
No. 68-C8-0062.

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                                   CONTENTS
Chapter One:

Chapter Two:


Chapter Three:

Chapter Four:

Chapter Five:

Chapter Six:

Chapter Seven:

Chapter Eight:


Chapter Nine:

Chapter Ten:

Chapter Eleven: ,

Chapter Twelve:


Chapter Thirteen:

Chapter Fourteen:
Introduction

Quality Assurance/Quality Control, Sampling, and Analytical
Considerations              •     -,

Bulk Sediment Toxicity Test Approach
                                          * -
Spiked-Sediment Toxicity Test Approach

Intersititial Water Toxicity Identification Evaluation Approach

Equilibrium Partitioning Approach

Tissue Residue Approach
       i"               ,       •'.-'.•
Freshwater Benthic Macroinvertebrate Community Structure and
Function

Marine Benthic Community Structure Assessment

Sediment Quality Triad Approach

Apparent Effects Threshold Approach             ;

A Summary of. the Sediment Assessment Strategy Recommended by
the International Joint Commission

Summary of Sediment-Testing Approach Used for Ocean Disposal

National Status and Trends Program Approach

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           CHAPTER 1
Introduction
1.1 BACKGROUND

    The problem  of contaminated sediments is
widespread in freshwater  and marine  systems
throughout the world.   Contaminated bottom
sediments can have direct  adverse impacts on
bottom fauna.  Contaminated sediments can also
be a long-term source of toxic substances to the
environment and can impact wildlife and humans
through the consumption of  food or  water or
through direct contact:  These impacts may be
present  even though  the overlying water meets
water quality criteria.   As a result, something
more than the traditional  water and  effluent
quality-based control  and monitoring approaches
will be  needed to protect and restore the quality
of  the  Nation's  rivers, lakes, estuaries,  and
embayments.
   ' In recognition of the  significance of the
problem,  the U:S.  Environmental Protection
Agency (EPA) has begun a comprehensive Con-
taminated sediment program. The effort began in
 1985, when EPA  examined the potential national
extent of sediment contamination using existing
sediment monitoring data from the EPA Storage
and Retrieval System (STORET) database (Bolton
et  al.,  1985),  These data were compared to
organic  carbon-normalized  threshold  concen-
trations calculated from existing  water quality
criteria  using the  equilibrium partitioning model.
In  1986, the EPA formed the Sediment Criteria
Technical Advisory Committee to examine possi-
ble approaches .for deriving regulatory criteria for
 sediments.  In 1988, EPA formed two oversight
 committees to take a  comprehensive look at the
 whole range of contaminated sediment issues: the
 Sediment Oversight Steering Committee, which is
" responsible for overall management of the pro-
 gram,  and the  Sediment  Oversight  Technical
 Committee, which is oriented toward technical
 issues  and is  the implementation  arm of the
 Steering  Committee.    These committees  have
 prepared a draft outline describing EPA's Contam-
 inated Sediment  Management Strategy and  have
formed working groups to focus ori specific issues
and approaches to sediment management.  The
committees  are  also sponsoring a  number  of
activities aimed at providing basic information
about contaminated sediment issues.to persons
within the Agency and to the interested public.
This compendium of sediment assessment methods
is one of the committees' products.
    An important initial step in addressing the
contaminated sediments problem is the identi-
fication of scientifically sound methods that can
beused to assess whether and to what extent sedi-
ments are "contaminated" or have the potential for
posing a threat to the environment. The Sediment
Oversight Technical Committee compiled  this
compendium of  sediment  assessment  methods
through the efforts of the committee members and
others who are experienced in the state of the art
in sediment assessment
    Many factors can affect the kinds and magni-
tudes of impacts mat contaminated sediments have
on the environment.  The sediment assessment
tools vary in their suitability and sensitivity for
detecting these different endpoints and effects.  It
is,  therefore,  important to properly  match the
assessment  methods to the site- and program-
specific objectives of the study being conducted.
The suite of assessment methods presented in this
compendium offers a rich repertoire of tools from
which to select the most suitable tests for a given
situation.
    Unfortunately,  there  simply is  no  single
method that will measure all contaminated sedi-
ment impacts at all times and to all  biological
organisms.  This is the result of a number  of
factors, including environmental heterogeneity and
associated sampling problems, variability in the
laboratory exposures, analytical variability, differ-
ing sensitivities of different organisms to different
 types of contaminants, the confounding  effects
 caused by the presence of unmeasured cdntami-.
 nants, the synergistic  and antagonistic effects of
 contaminants,  and the  physical properties  of
 sediments.  While one method will suffice for

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 Sediment Classification Methods Compendium
some circumstances, it is often advisable to use
several complementary methods rather  than a
single one. When several of these approaches are
used together, they can provide additional insights
into the nature and degree of sediment contamina-
tion problems. The use of complementary assess-
ment methods can provide a kind of independent
verification of the degree of sediment contamina-
tion if the conclusions of the different approaches
agree.  If the conclusions differ, that difference
indicates a need for caution in interpreting  the
data since some unusual site-specific circumstanc-
es may be at work. The importance of this type
of verification increases with the significance of
the decisions that must be made using the infor-
mation obtained.  In fact, the actual decision-
making frameworks within which the compendium
methods are used often include this verification in
the concept of tiered testing.
    The assessment methods presented  in  the
compendium are continually being refined and
improved.   Additional methods are also being
developed. As these methods are developed and
verified, they will be incorporated into  future
updates of the compendium.
 the most useful overall measures or predictors of
 ecological impacts currently in use rather than
 procedures that may have  limited application
 outside  of a  particular regulatory framework.
 Nevertheless, many of the methods presented in
 the compendium can be used as part of regulatory
 and/or remedial •actions.
    Guidance on  how. to  use the compendium
 methods in a decision-making framework will be
 provided in forthcoming documents and will likely
 include both chemical and biological methods in
 a tiered hierarchical framework suitable for testing
 various hypotheses and endpoints. Currently such
 a document  has been prepared by the Sediment
 Oversight  Technical   Committee to summarize
 existing EPA  decision-making processes  for
 managing  contaminated  sediments (Managing
 Contaminated Sediments: EPA Decision-Making
 Processes;  USEPA,   1990).   The  information
 provided in  the  compendium  on the  relative
 strengths and weaknesses of the different assess-
 ment methods can provide assistance in selecting
 the appropriate methods.
1.2 OBJECTIVE

    This document is a compendium of scientifi-
cally valid and accepted methods that can be used
to assess sediment quality and predict ecological
impacts.
    Some regulations require the use of certain
types of tests (e.g., the Toxicity  Characteristic
Leaching Procedure under the Resource Conserva-
tion and Recovery Act), criteria (e.g., the limita-
tions in the London Dumping Convention), and
procedures (e.g., risk assessment under the Com-
prehensive Environmental Response, Compensa-
tion, and Liability Act). Additional guidance may
be issued in the future to provide direction when
addressing sediment contamination under particu-
lar regulatory programs including these, or other,
required tests and-approaches.  These other test
procedures will not be presented in this compendi-
um, however, because the intent here is to provide
 1.3 OVERVIEW

     The compendium is organized in the following
 manner.  The remainder of this chapter gives a
 broad overview of the assessment methods in the
 compendium.  The information is presented in
 tabular form to facilitate comparisons between the
 different  methods.   Chapter  2 outlines quality
 assurance/quality control, sampling, and analytical
 considerations  that apply to all of the methods.
 Method-specific  information  is  also provided
 where the procedures differ from the general ones.
    .The remaining chapters give specific informa-
 tion on each of the sediment assessment methods.
 The  information is organized  in a consistent
' manner for each assessment method so the reader
 can readily compare the relative strengths, weak-
 nesses, and applicability of each method in  order
 to  select  the best method(s) for a  specific situa-
 tion. The information provided for each method
 includes the following:
1-2

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                                                                              1—Introduction
   •   How each  method is currently used  or
       could be used;             ,

   •   A detailed description  of the method,
       including types of data, equipment, and
       sampling procedures needed;

   »   The  applicability of the method to the
       protection of wildlife and humans;

   •   The  utility of the method to produce
       numeric sediment quality criteria;

   •   The  method's applicability  to  making
       different types of sediment management
       decisions;

   •   The  method's  advantages,  limitations,
       costs, level of acceptance, and accuracy;

   »   The degree to which the method is actual-
       ly being used  now;

   •  How well it is validated; and

   •  Its potential future uses.

   Extensive references are provided after each
method in case any additional details are required.
The names, addresses, and telephone numbers of
the authors of the descriptions of each method are
provided to facilitate additional follow-up.  Given
the limited level of detail in the compendium, use
of these references is  suggested for actual'imple-
mentation of the methodologies.
   The -i 12 sediment assessment methods  de-
scribed in the  compendium are  summarized in
Table 1-1.  The  assessment methods can  be
categorized in many different ways.  Differentia-
tion could be made between numeric methods and
descriptive methods. Numeric methods are chemi-
cal-specific and can be used to generate numerical
sediment quality criteria (SQC) on a chemical-by-
chemical basis. A potential drawback of descrip-
tive methods is that they are not chemical-specific
and cannot be  used alone to generate numerical
sediment quality criteria for particular chemicals.
On the other hand, descriptive methods  can be
used to directly assess the overall impact of all
chemicals that may be present in a  sediment,
whereas it is difficult to use the chemical-specific
methods to predict the combined effects of several
chemicals.
    Another  differentiation  that is often made
among different sediment assessment methods is
whether they are based on the measurement of the
concentrations of chemicals  of concern or on the
measurement of biological impacts.  For methods
that have ecological validity,  this differentiation
really applies only to the practical implementation
of the methods rather than to their scientific basis
since all ecologically valid methods must ultimate-
ly be based on an  ability to predict or measure
biological effects.. Many of the assessment meth-
ods use both chemical and biological testing or
observation.       '
    Yet another differentiating factor is whether
the method uses interstitial water  (pore water),
elutriate, or bulk sediment (whole,  including the
solids and interstitial water). This difference also
relates primarily to implementation  rather than to
a substantive scientific difference since the chem-
istry of interstitial water and that of the  bulk
sediment are closely linked. Except for contami-
nants that might be transferred directly by inges-
tion,  interstitial  water is the medium, through
which the contaminants in the bulk sediment are
transferred to the affected organisms.          .
    Some  of  the  assessment  methods • (which  ,
would be more  accurately  characterized as ap-
proaches) described in the compendium combine
numeric and descriptive measures.  For example,
the Sediment Quality Triad (Triad) and Apparent
Effects Threshold (AET) approaches employ bulk
sediment toxicity  testing, benthic  community
structure analysis, and concentrations of sediment
contaminants.  The Triad is both descriptive and
.numeric, depending  on  its use.  Typically, the
Triad  approach has  been used in a  descriptive
 manner to identify contaminated sediments. It has
 also been used, however, to generate  criteria, for  .
 several chemical contaminants.  The International
 Joint Commission (UC) approach would be more
 accurately described  as  an assessment strategy
 since  it employs several of  the other sediment
 assessment methods in a tiered, comprehensive
                                                                                             1-3

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Sediment Classification Methods Compendium
                Table 1-1.  Some Characteristics of the Sediment Assessment Methods.
Sediment Method
(Chapter Number)
Bulk Sediment Toxic'rty
P)
Spiked-Sediment
Toxicity
(4)
Interstitial Water Toxicity
(5)
Equilibrium Partitioning
(6)
Tissue Residue
(7)
Freshwater Benthic
Macroinvertebrate
Community Structure
and Function
(8)
Marine Benthic
Community Structure
(9)
Sediment Quality Triad
(10)
Apparent Effects
Threshold
(11)
International Joint
Commission Sediment
Assessment Strategy *
(12)
Sediment-Testing
Approach Used for -
Ocean Disposal
(13)
National Status and
Trends Program
Approach
(14)
Description
Test organisms are exposed to sediments that may contain unknown quantities of potentially toxic
chemicals. At the end of a specified time period, the response of the test organisms is examined
In relation to a specified biological endpoint. •
Dose-response relationships are established by exposing test organisms to sediments that have
been spiked with known amounts of chemicals or mixtures of chemicals. '
The toxic'rty of interstitial water is quantified and identification evaluation procedures are applied to
identify and quantify chemical components responsible for sediment toxicity. The procedures are
implemented in three phases to characterize interstitial water toxic'rty, identify the suspected
toxicant, and confirm toxicant identification. ,
A sediment quality value for a given contaminant is determined by calculating the sediment
concentration of the contaminant that would correspond to an interstitial water concentration
equivalent to the U.S. EPA water quality criterion for the contaminant.
Safe sediment concentrations of specific chemicals are established by determining the sediment
chemical concentration that will result in acceptable tissue residues. Methods to derive unaccept-
able tissue residues are based on chronic water quality criteria and bioconcentration factors,
chronic dose-response experiments or field correlations, and human health risk levels, from the
consumption of freshwater fish or seafood. .
Environmental degradation is measured by evaluating alterations in freshwater benthic community
structure and function.
Environmental degradation is measured by evaluating alterations in marine benthic community
structure. • . - '
Sediment chemical contamination, sediment toxic'rty, and benthic infauna community structure are
measured in the same sediment Correspondence between sediment chemistry, toxicity, and
biological effects is used to determine sediment concentrations that discriminate conditions of
minimal, uncertain, and major biological effects.
An AET is the sediment concentration of a contaminant above which statistically significant
biological effects (e.g., amphipod mortality in bioassays, depressions in the abundance of benthic
infauna) would always be expected. AET values are empirically derived from paired field data for
sediment chemistry and a range of biological effects indicators.
Contaminated sediments are assessed in two stages: (1) an initial assessment that is based on
macrozoobenthic community structure and concentrations of contaminants in sediments and
biological tissues and (2) a detailed assessment that is based on a phased sampling of the
physical, chemical, and biological aspects of the sediment, including laboratory toxicity bioassays.
A tiered testing strategy consisting of physical, chemical, and biological testing to predict benthic
and water column impacts of dredged, sediment disposal.
Three ranges of concentrations are determined for each chemical: the no-effects range, the
possible-effects range, and the probable-effects range. These values are arithmetically deter-
mined from a database consisting of matching chemical and biological data from laboratory
spiked-sediment bioassays, equilibrium-partitioning models, and field studies:
1-4

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                                                                              1—Introduction
procedure. The Sediment-Testing Approach Used
for Ocean Disposal is the  tiered,  comprehen-
sive testing procedure developed by EPA and the
U.S. Army Corps  of Engineers (USAGE)  for
determining the suitability of dredged material for
disposal at designated disposal sites.  The proce-
dure is specified in Evaluation of Dredged Materi-
al Proposed for Ocean Disposal-Testing Manual,
commonly referred to as the 1991 Green Book
(USEPA/USACE, 1991).
    To facilitate the user's selection of the.most
.suitable sediment assessment method, Tables 1-2
through 1-5 highlight the major characteristics of
each method,  information from individual chap-
ters that  is useful  in  management  decisions is,
presented in summary  form and includes method
descriptions and uses,  data and sampling required,
ability  to generate numerical  sediment  quality
criteria, and outlook for future use. More pointed-
ly, the reader will  learn what each method pre-
dicts, what it assumes, how much it will cost, and
why one might choose a particular method over
 another for a specific situation.
    Regardless of which  of  the compendium
 methods one uses, several considerations must be
addressed:   a sampling  program needs  to  be
designed; samples need to be collected, stored,
and analyzed; and quality assurance/quality control
is needed throughout the process to determine the
uncertainty associated  with  the results of the
assessment  Sampling design ahd,QA/QC issues
will be discussed in Chapter 2.
1.4 REFERENCES

Bolton, S.H., RJ. Breteler,  B.W.  Vigon,  JA.
    Scanlon,  and S.L. Clark.   1985.   National
    perspective on sediment quality.    .
USEPA.  1990.  Managing contaminated sedi-
    ments: EPA decision-making processes.  U.S.
    Environmental Protection Agency, Sediment
    Oversight Technical Committee. EPA 506/6-
    90/002. .
USEPA/USACE.  1991. Evaluation of dredged
    material proposed for ocean disposal—Testing
    manual. U.S. Environmental Protection Agen-
    cy and U.S. Army Corps-of Engineers.
                                                                                            1-5

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   Sediment Classification Methods Compendium
Type of Sampl
Required
Reid-collecte
sediments.
Reid-collected sedi
contaminated or
uncontamlnated.
bulk
Reid-colle
sediments.
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          CHAPTER 2
Quality Assurance/Quality Control,
Sampling, and Analytical  Considerations
    The purpose of this chapter is to provide a
brief introduction to some* of the most important
terms and concepts that are integral to the design
of an  adequate program for  sediment  sample
collection, handling, and analysis. This chapter is
intended only  as  a  general guide to sediment
sampling and should not be used as an instruction
manual for  collecting samples.   The  subjects
mentioned will not be dealt with iri an exhaustive
manner.  The reader is referred to the references
cited in this chapter for more complete guidance
on the particular .techniques.
 2.1 ESTABLISHING DATA QUALITY
    OBJECTIVES
        i  ,                      '
    Fundamental to the process of designing a
 study is the establishment of data .quality objec-
 tives (DQOs). The most carefully collected and
 analyzed data are of no use if the data collected
 are insufficient or of the wrong type. To avoid
 either of these and other potentially costly errors,
 EPA has  initiated the use  of the DQO Process.
 The DQO Process is a management tool designed
 to help data users and data collectors design the
 best sampling strategy to reach their objectives
 while minimizing resource requirements.  It is a
 multistep, systematic approach  to data collection
 that enables the manager  to  refine goals and
 objectives and help answer the question, "How
 much data is enough?" As the  steps of the DQO
 Process  are followed, the  decisions made  in
 previous  steps  should be  reviewed to  ensure
 consistency and  cohesiveness.
     The first step in the process is to specify the
 problem and identify limitations of time or re-
 sources on the data-collection effort. This process
 allows one to evaluate his or her current knowl-
 edge base of the problems and identify all avail-
 able resources.  The next step is to identify what
 decisions or activities will be made based on the
data.   The answer to this  question is vital to
ensure the collection of the right type of data.
The decision goals should be as narrow in scope
as possible, and-considerable effort may be re-
quired to define them properly.   ,
    The third step involves identifying all vari-
ables needed to  make a  decisipn.   This  step
focuses on eliminating the potential measurement
or collection of data that may not actually be used
in the decision-making process.  The next step
requires the data collector  to set or define the
boundaries of the study, including the population,
which could consist of people, objects, or media,
and the boundaries on the population, including
space, time, and area.
    Developing a decision rule, or how the data
will be used and summarized, is the next step in
 the process.  This step involves describing how
 the study results will be compiled or calculated
 and defining the decision rule in an "If.:., then ..."
 format.   The statement should incorporate the
 study results as "If the results are this, then the
 action should be this/  For  example, "If PCB
 levels in fish are greater than  2 ppm, then a fish
 consumption advisory will be issued." This step,
 along with the others, helps define the data collec-
 tion effort by identifying the data needed to fulfill
 the decision rule.
    A very important step  in the DQO Process is
 specifying the limits of uncertainty acceptable in
 the data.  These limits can be expressed as accept-
 able false-positive and false-negative error rates
 for the decision.  These error rates .must be based
 on careful consideration of the consequences of
 incorrect conclusions being drawn from the data.
 The definitions of false-positive and false-negative
 errors vary with the decision being defined.  If a
 decision to take regulatory action is being made,
 a possible false-negative error could result in no
 action being taken because incorrect data results
 indicated  there was no problem.  The opposite
-could also occur, where  a false positive error

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Sediment Classification Methods Compendium
results in regulatory action being taken when no
problem exists.  It is essential that the potential
consequences  to  economic, health,  ecological,
political, and  social issues  be considered when
deciding on acceptable false-positive and false-
negative error rates.  This step may involve the
consultation of a qualified statistician.
    Finally, all steps in the DQO Process should
be reviewed to design the most efficient sampling
study.  Considerations including cost, time, de-
fined boundaries, the decision rule, and all other
factors  defined  and specified  during  the DQO
Process should be incorporated.
    One can refer to "Planning Issues for Super-
fund  Site Remediation"  in  Hazardous Material
Control (Ryti and Neptune, 1991) for an excellent
example of applying the DQO Process to an actual
situation.
    Quality assurance and quality  control are
integral  components  of  every aspect  of a pro-
gram's activities.  The collection of reliable data
is contingent on the  use of and adherence to  a
good Quality Assurance Project Plan; the devel-
opment of a sound sampling study is contingent
on  the  use of the DQO Process; and use  and
implementation of the DQO Process is contingent
on a Quality Assurance Program Plan.
2.2 SAMPLING DESIGN
2.2.1  Test, Reference, and Control Sediments

    In sediment  quality evaluations, there is  a
substantial  precedent  for using  comparisons
between sites rather than  comparison of testing
results to an independently set numerical bench-
mark. This is the result of a number of factors
including the standard procedures used in biologi-
cal testing, the paucity of scientifically acceptable
numerical sediment quality criteria or standards,
and  the long-standing "nondegradation"  philoso-
phy   used in  evaluating  the  acceptability  of
dredged  material for open-water disposal.  The
degree of sediment contamination in  a particular
area is often evaluated by comparing the structure
of benthic communities, levels of pollutants, or
bioassay test results in sediments collected from
the area being  investigated with those  in  the
surrounding area. The terms used to describe the
different sediments in the comparisons are  test
sediments,   control  sediments, and  reference
sediments.
    As used in sediment assays and assessments,
a test sediment is sampled from the area  whose
quality is being assessed.  A control sediment is
a  pristine  (or nearly  so) sediment,  free from
localized anthropogenic inputs of pollutants with
contamination present only because of inputs from
the global spread of pollutants (Lee et al., 1989).
A control sediment is fully compatible with the
needs of the organisms  used in  the assay, is
known to not cause toxicity, and is used primarily
to verify the health of the test organisms and the
acceptability of the test conditions (USEPA/USA-
CE, 1991).  The  control sediment  may be artifi-
cially prepared  in  order  to  achieve  sufficient
volumes of a known and consistent quality  for use
in standard testing and for culturing test organisms
(ASTM, 1990).,
    A reference sediment, on the  other hand, is
collected from a location that may  contain low to
moderate levels of pollutants resulting from both
the global inputs and some localized anthropogen-
ic sources, representing the background levels of
pollutants in an  area (Lee et al., 1989).  The
reference sediment is to be as similar as possible
to the test  sediments in grain size, total organic
carbon (TOC), and other  physical  characteristics
(Lee era/., 1989; USEPA/USACE, 1991; ASTM,
1990). The physical environment of the reference
site should also be as similar as possible to that at
the sites where the test sediments will be collect-
ed.  This  is  especially significant for benthic
community structure comparisons, since communi-
ty structure can be very significantly affected by
water depth, physical transport processes such as
waves and currents, sediment grain size, and the
presence of organic debris.
    As used in dredged material assessment, the
results of assays or evaluations on the test sediments
are compared to those obtained  from  reference
sediments to determine whether the test sediments
are contaminated. In contrast, the results of assays
or  evaluations using  the  control  sediments  are
usually compared only to  past results using those
2-2

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                                            2—CM/QC, Sampling, and Analytical Considerations
same control sediments to ensure that the testing
was free of some extraneous factors that may have
affected the reliability of the test. Depending on the
study objectives, however, controls can also be used
as a  benchmark  against which to  compare test
sediments to determine the relative degree of con-
tamination of sediments  collected from  different
sites (ASTM, 1990).
    A clear understanding of the end uses of the
data is essential in the establishment of an appro-
priate sampling program. A cost-effective study
for a qualitative overview of potential contaminat-
ed sediment impacts will differ markedly from one
whose purpose,  is  to  make  statistically-based
numerical comparisons with criteria or indexes, or
to reference sites. .
    Sediment sampling programs are most often
undertaken  to achieve one or more  of the follow-
ing objectives:

     •  To  fulfill a regulatory testing requirement:.

     •  To  determine characteristic ambient lev-
        els;

     •  To monitor trends in contamination levels;
for analysis  together with  some "observation"
samples to supplement the analytical results.
   /Available information about the area to be
sampled and its surroundings should be used in
determining the final sample design. Knowledge
about bottom topography, currents, areas of dredg-
ing and the frequency of dredging, locations of
point  and nonpoint  sources  of contaminants,
distribution of grain sizes, and other factors can
provide the basis for determining which of the
sampling designs to use (e.g., Are there reasons to
expect localized hot spots of contamination?) and
where to place sampling locations  (e.g., Which
parts of the area are likely to be similar enough to
group into the same strata?). Preliminary surveys
of an  area using depth-sounding and sediment-
profiling  equipment can   prove invaluable in
delineating vertical and horizontal distributions of
sediments (DC, 1988).  This information can be
helpful in planning sediment sampling methods
(grab samples or cqre samples) and sample site
selection  (grouping  similar areas  into  strata,
identifying likely locations  of hot spots).
    The methods most often used for selecting the
sample collection sites are haphazard, worst-case,
random,   stratified  random,  and  exhaustive
 (Higgins, 1988).
        To identify hot spots of contamination; or    2.2.1.1  Haphazard
   -' - •   To screen for potential problems.

     These different objectives will lead to differ-
 ent sampling designs. For example, a study for a
 dredging project may have a specific set of guide-
 lines on sampling frequency, sample site selection
 methodology, and other parameters already deter-
 mined by existing specific guidance. The design
 for a study to determine ambient levels will strive
 to obtain uniform, random  coverage of an area
 through the collection of samples from a relatively
 large number of sites. The design for a study to
 track sediment contamination trends will expend
 its resources to sample fewer sites but more often.
 A study to identify hot spots would concentrate
 efforts on fewer sites within zones most likely to
 be contaminated, while an initial screening study
 might take very few, randomly distributed samples
     The haphazard method, whereby one selects
 sampling sites based on whim or ease of imple-
 mentation rather than science or knowledge, really
 reflects the lack of a design. This method has no
 validity and should not be used.

 2.2.1.2 Worst-Case

     The worst-case sampling  design is based on
 knowledge regarding the presence and distribution
 of potential sources of sediment contamination in
 an area.  It is usually considered cost-effective as
 long as the study objectives are being met.  An
 inherent problem with this design is that it results
 in an incomplete characterization of an area and is
 not statistically robust.  However, it can be useful
 as an  initial survey to determine the potential for
 a contamination problem, which  would be fol-
                                                                                               2-3

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 Sediment Classification Methods Compendium
 lowed up with more complete sampling later, if
 needed.   The effectiveness  of this  technique
 depends on the availability  of reliable historical
 information on contamination, sources, bathyme-
 try, currents, and other factors.

 2.2.1.3 Random

     The random sampling design is most useful
 for cases where little is known  about  the likely
 distribution of sediment contamination or sources,
 or when available information indicates a high
 degree of homogeneity in an area. The area to be
 sampled is divided using a grid system.  Samples
 are distributed within the grid randomly, with each
 location having an equal probability  of being
 sampled.  The number of  samples is selected
 statistically  based  on  the  requirements of  the
 survey and the acceptability of false-positive or
 false-negative results.  This design yields statisti-
 cally sound results.

 2.2.1.4 Stratified Random

     The stratified random design  is a variation on
 the previous two designs.  Available information
 is used to identify different zones  that are likely to
 be similar in degree of contamination or' other
 characteristics.  Samples sites  are then  randomly
 selected within the different zones.  This design
 also yields statistically reliable results.

 2.2.1.5 Exhaustive

    In the exhaustive design,  an area is subdivided
 into equal-sized units, each of which is then
 sampled.   This design yields a very  complete
 characterization. However, this design is usually
 very costly because of the large number of sam-
 ples that need to be collected.
                                              j>
 2.2.2  Numbers of Samples

    Statistics can be used to determine the number
 of samples'needed. To use statistics in  this way,
 one needs to decide  what comparisons will  be
 made with the resulting data and what will be the
 desired statistical power of the comparisons (i.e.,
 at what level of confidence will resulting differ-
 ences be tested).  In addition, one needs some
 information  about  the  inherent environmental
 variability in the area (i.e., the likelihood that an
 observed difference is due to ah actual difference
 in  contamination  rather than just  the  natural
 heterogeneity in sediment or benthic population
 characteristics  in the Jirea).   There are many
 different statistical approaches to estimating the
 number of samples required and to interpreting the
 resulting test results. Excellent reviews of statisti-
 cal designs and interpretation are given by Baudo
 (1990) for sediment physical and chemical testing
 and by Downing and Rigler  (1984) for  benthic
 community structure evaluations.
    In  practice,  constraints  on  resources often
 preclude the  use of a purely statistical approach to
 determining  the  number of samples and some
 form of a cost-benefit approach  is often used to
 arrive at a reasonable compromise between statis-
 tical power and the cost of the study. One of the
 major advantages  of  the tiered approaches for
 testing and assessment is the cost savings that
 results  when information is collected relatively
 inexpensively initially and additional resources are
 expended  only when  the information  collected
 thus far is insufficient to make a  decision.
    Guidance on how to select  a  cost-effective
 approach is  usually provided in  very general
 qualitative terms as to the factors that should be
 considered in  arriving at a  decision (USEPA/
 USAGE,  1991; Higgins, 1988;  Plumb,  1981).
 Decisions  are largely  subjective.  However, re-
 searchers  at EPA's  Environmental Research
 Laboratory (ERL)-Narragansett/Newport recently
 developed a four-step procedure to determine the
 optimal cost-effective sampling scheme for marine
 benthic community assessment (USEPA, undated).
 The procedure  begins with  an initial  limited
 sampling using two or more sampling schemes at
 paired sites (test and reference sites). The  "costs"
 in tune and money .are assessed for each sampling
 scheme.   Next, a  statistical  power  analysis is
 conducted to calculate the  number of  replicate
 samples needed to achieve a desired degree of
statistical "power"  for each  sampling  scheme.
Finally, the power-cost efficiencies of the alterna-
tive  sampling schemes are. calculated  and  the
2-4

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                                            2—QA/QC, Sampling, and Analytical Considerations
optimum scheme is selected as the one with the
highest power-cost efficiency.


2.3 QUALITY ASSURANCE/QUALITY
    CONTROL

    Quality assurance  and quality control (QA/
QC) are essential to, the production of environ-
mental monitoring data of known and documented
quality in a cost-effective manner, QA/QC should
be an integral part of the process of study d.esign,
execution,  and data evaluation and interpretation.
    All EPA data-collection programs have imple-
mented Quality Assurance Program Plans designed
and overseen  by their management to ensure the
quality of  all  activities for which their organiza-
tion is responsible.  These programs address all
quality   assurance  issues in  regard^to  policy,
planning, review, and  implementation.  QA Pro-
ject Plans are  a vital part of the QA Program Plan.
A QA Project Plan is a project-specific guidance
compiled to encompass all aspects  of  the sam-
pling/analytical effort.  The preparation of a QA
Project Plan is often met with unnecessary trepida-
tioii.  A QA Project Plan is simply a  written
record of  the plans that must be made and fol-
 lowed in executing a study.  A QA Project Plan
 provides detailed documentation of all facets of
 how and why a particular study will be undertak-
 en. The Plan also describes the alternative actions
 that will be taken in the event that things do not
 gb according  to the original plans. Once all of the
 purposes and procedures of the proposed study are
 recorded in a QA Project Plan, the Plan can be
 improved or modified, if needed, through reviews
 by persons knowledgeable about different aspects
 of the  study (e.g., chemical analysis, sampling
 logistics, navigational positioning, sample preser-
 vation techniques).
     Because the QA Project Plan is a vital tool for
 the data-collection process, it is essential that all
 personnel involved in the project read and under-
 stand the Plan  and that the Plan be available for
 reference throughout the project to ensure proper
 implementation.          .
     QA Project Plans are important for  legal as
 well as scientific reasons.  QA Project Plans are
required  for  all EPA-associated  projects (EPA
Order 53(50.1).  QA Project Plans become part of
contracts that are issued to undertake studies (40
GFR, Part 15).  Furthermore, nonadherence to the
Plan could result hi the data being unusable for
court proceedings or regulatory decisions.
    The QA Project Plan is just as important after
the study is completed and the data are being used
to make an  evaluation or decision.  The Plan
provides the information needed to  assess the .
degree of confidence one can place in the data, as
well as the comparability of the data collected in
a particular study with those from another study.
A common problem that managers and scientists
have with using existing data is not that the old
data are unreliable, but that the data are of un-
known reliability.                            .

2.3.1  QA/QC Terminology

    A number of important concepts and terms
need to be defined to develop an understanding of
what makes up an adequate QA/QC program
(USEPA, 1983; Delbert and Starks, 1985).

    Accuracy is defined as the difference between
     a measured value and the assumed or expect-
     ed value.  Accuracy in percent is 100 minus
     the total error, which is composed of bias and
     random errors.
    Bias is the systematic distortion of a measure-
     ment process that adversely affects the repres-
     entativeness of the results.  Bias can result
     from the basic sampling design, the kind of
     equipment used to collect the samples, the
     sample-handling procedures,  and poor recov-
     ery of the analyte. Because bias is systematic,
     its magnitude can be predicted if proper QA
     procedures are being used in the field and
     laboratory.
     Comparability is  the measure of  confidence
     one has in being able to compare one data set
     with another.  Comparability is increased if
     similar  field and  laboratory methods  were
 :    used and  decreased if different or unknown
     (undocumented) methods were used. Compa-
     rability between different laboratories can be
     evaluated through the use of inter-laboratory
                                                                                              2-5

-------
 Sediment Classification Methods Compendium
    comparisons, or "round-robin" studies, where-
    in standardized samples are analyzed by each
    of the participating laboratories.

    Completeness is the amount of valid data
    obtained (i.e.,  that met QA/QC acceptance
    criteria)  compared to the planned  amount.
    Completeness is usually expressed as a per-
    centage.

    Data quality refers to the sum of all features
    and characteristics of the data that determine
    its capability to satisfy the objectives of the
    data collection.

    Data quality indicators are quantitative statis-
    tics and qualitative descriptors that are used to
    interpret  the degree of acceptability or utility
    of data to  the user.  Data quality indicators
    include bias, precision, accuracy, comparabili-
    ty, completeness, and representativeness.

    Data quality objectives (DQO) are statements
    of the overall uncertainty that a decision-
    maker is willing to accept in results or deci-
    sions derived from the data, and they provide
    the framework for the data-collection effort.

    Duplicate  samples are  two samples  taken
    from  and representative of the same popula-
    tion and carried through all the same steps of
    sampling, storage, and analysis in an identical
    manner.

    Field blank is a  clean sample (i.e.,  distilled
    water) carried to the sampling site, exposed to
    sampling  conditions,  and returned to  the
    laboratory  and treated as  an environmental
    sample. Field blanks  are used to try to assess
    contamination problems caused by conditions
    in the field, including contamination of the
    sampling device, sample containers, shipping
    containers, etc.

    Measurement error is the difference between
    the true sample values and the reported values
    and  can  occur during analysis,  data entry,
    database  manipulation, or other steps.
Method sensitivity/method  detection  limit
defines the lower limits of reliable analysis of
a particular parameter inherent in the use of a
particular test method. The method detection
limit is the minimum  concentration  of a
substance that can be measured with 99 per-
cent confidence that the analyte concentration
is greater than zero in a particular medium (40
CFR Part 136, Appendix B).

Precision is the degree of consistency among
duplicate/replicate measurements.

Quality assurance  is an integrated program
for ensuring the reliability  of monitoring and
measurement data.   It includes the well-de-
fined plans and procedures for how to ensure
the production of sufficient data of known and
documented quality, including monitoring how
well QC procedures are actually being imple-
mented.

Quality control is the routine application of
procedures for obtaining prescribed standards
of performance in the monitoring and  mea-
surement process. It is the actual implementa-
tion of the QA plan, effected  through  mea-
surements of data quality through the use of
blanks, spikes, etc. Quality control consists of
both internal and external checks including
repetitive measurements, internal test samples,
interchange of technicians and equipment, use
of independent methods to  verify findings,
exchange of samples  and standards  among
laboratories,  and use  of standard reference
materials.

Random error is nonsystematic (and, there-
fore, unpredictable) error that can occur dur-
ing any part of the sample  collection,  han-
dling, and analysis.  Hopefully, random errors
are normally  distributed with a mean of zero
so that the overall  evaluation  will not  be
affected even though individual measurements
will be affected.

Representativeness is the degree to which the
data accurately  and precisely  represent the
2-6

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                                            2—QA/QC, Sampling, and Analytical Considerations
   parameter or condition being sampled.  Repre-
 -  sentativeness is affected by sampling design
   (e.gi, number of samples, method of selecting
   sampling sites), as well as analytical sampling
   accuracy and precision.

   Sampling error is the difference between the
   sampled value and the true value, and  is a
   function of natural spatial and temporal vari-
   ability and sampling design. It also includes
   error  due  to 'improperly  selected/collected
   samples or improperly gathered measurements.
   Sampling  error is more difficult to  control
   than the other  type  of error, measurement
   error, and  typically accounts for most of the
   total error.

   Uncertainty is the total variability in sampling
   and analysis including systematic error (bias)
   and random .error.

   Duplicates, spikes, and blanks are all used to
assess  the  quality of the  data, to identify any
systematic  problems, and to isolate the sources of
such problems.
2.4 SOURCES AND SIGNIFICANCE OF
    MONITORING ERROR
          !     ,   •          .-..-'.     -
    To increase  the  accuracy, precision,  and
representativeness of the data collected in a sedi-
ment assessment study, it is important to be aware
of and minimize two  types of error that can be
introduced into sediment contaminant concentra-
tion data: bias and scatter.   Sources of bias  in
sediment studies include the actual heterogeneity
in the  distribution of contaminants in the sedi-
ments, the sampling design (number of samples,
method for selecting sampling sites), the sampling
method, the sample preparation procedures, and
the testing methods.                .    '•'.".
    Factors that tend to make sediment contami-
nants  distribute  themselves   heterogeneously
include the differences in the density of the bulk
 contaminant (e.g., sinking versus floating); differ-
 ences in the affinity of the contaminant for parti-
cles as a function of particle size, organic carbon
content,  etc.;  particle sorting as  a function of
water currents and particle size; lateral mixing of
water and sediments as a function of  flow or
distance downstream of the sources; resuspension;
bioturbation; and biouptake.
    The objective of a  well-designed sampling
program is to minimize  the introduction of data
artifacts associated with the sampling plan, sample
collection, sample preparation, and sample analy-
sis while revealing the actual contaminant concen-
tration profile in space as a function of  time.  A
plan that requires preferential sampling  of areas
that are devoid of aquatic life will likely be biased
toward high toxicant concentrations, resulting in
an unrepresentative  horizontal  spatial  sediment
contaminant profile.  Artifactual variability can be
introduced if the number and size of the samples
are inappropriate to the scale of the system under
investigation,  yet the sampling size  has to  be
balanced against cost.
    With respect to bias due to sampling method, if
 certain , core samplers are used to quantify the
vertical distribution of a  sediment contaminant,  for
 example, the actual vertical profile is likely to be
 distorted because the absolute vertical relationship of
 contaminant concentrations is lost due to differential
 compression of the sample .during coring. Another
 example of sampling method bias occurs when a
 grab sampler is used to collect the surficial sediment
 sample.  The potential disproportionate loss of fine
 particles from the grab during the drop, closing, and
 withdrawal phases  of sampling  can result in an
 underquantificatiori  of  the  contaminant surficial
 concentration  if the contaminant is preferentially
 concentrated on the fines.
     Regarding sample preparation bias, a sample
 preparation-procedure that transforms,  loses,   or
 destroys one member of a homologous series (e.g.,
 PCBs, PCDDs, or PCDFs) will not only result in
 an underquantification  of the total concentration
 for that toxicant category, but will also misrepre-
 sent  the -relative  proportions  of the isomers.
 Analytical method bias can result from the inabili-
 ty to  separate complex mixtures into'  individual
 constituents (interference),  thus resulting in the
 misidentification or misquantification  of a toxi-
  cant;  from differences in the sensitivity of the
                                                                                               2-7

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 Sediment Classification Methods Compendium
 detector for a particular pollutant over the range of
 concentrations encountered in the sediment (non-
 linear responses); or from poor or varying recov-
 ery of the analyte.
     Analytical variability arises primarily from the
 compounded uncertainty associated with  the
 tolerance on each of the components and steps of
 the wet or electronic methods of sample prepara-
 tion (aliquot selection, weighing, drying, grinding,
 sieving, etc.) and analysis.
 2£ COMPONENTS OF A QUAIJTY
    ASSURANCE PROJECT PLAN

    As mentioned previously, a QA Project Plan
 clearly documents the participants' responsibilities;
 what will be done; why  it is  being done; the
 desired  accuracy,  precision,  completeness,  and
 representativeness of the resulting data; who will
 report what information to whom; and what will
 be done in the event something goes wrong.
 Rather than attempting to describe the actual
 components of a QA Project Plan in any detail
 here, an example of the table of contents from a
 recent plan is presented in Figure 2-1.  In addi-
 tion, actual QA Project Plans from projects similar
 to the one being planned can be extremely useful
 in  suggesting the important issues  to  consider.
 For detailed guidance on  preparing QA Project
 Plans, one should refer to Interim Guidelines and
 Specifications for Preparing  Quality Assurance
 Project  Plans  (USEPA,  1980).    Some  good
 examples of actual sediment assessment Quality
 Assurance Project Plans include Burton (1989),
 Crecelius  (1990), and  Valente  and  Schoenherr
 (1991).
2.6 SAMPLE COLLECTION AND
    HANDLING

2.6.1  Sampling for Physical and Chemical
       Analyses

2.6.2.1 Sample Collection Methods

  . The most appropriate  device for a specific
study depends on the study objectives, sampling
 conditions, parameters to be analyzed, and cost-ef-
 fectiveness of the sampler.  There are basically
 three types;of devices used to collect sediment
 samples:  dredges,  grab samplers, and corers
 (Baudo, 1990).
    A dredge1 is  a vessel that is dragged across
 the bottom of the surface being sampled, collect-
 ing a composite of surface sediments and associat-
 ed benthic fauna.  Dredge  samplers  are more
 commonly used to sample sediments in marine
 waters than in fresh water.  This type of sampler
 is primarily used for collecting indigenous benthic
 fauna rather than.samples for analyses or assays.
 Because the sample is mixed with the overlying
 water, no pore water studies  can be  made of
 dredged samples.  Additionally, because the walls
 of the dredge are typically nets, they act as a sieve
 and only the coarser material is trapped, resulting
 in  the loss of fine sediments  and water-soluble,
 compounds (ASTM, 1990).   Results of dredge
 sampling are considered qualitative in nature since
 it is  difficult to  determine the actual surface
 sampled by the'dredge.
    Grab  samplers  have jaws that close by  a
 trigger mechanism upon impact with the bottom
 surface.   Grab samplers offer  the  advantage of
 being able to collect a large amount of material in
 one sample, but they have the disadvantage of
 giving  an unpredictable depth of penetration.
 Grab  samplers are recommended when  sampling
 is being performed for routine  dredging projects
 because the sediments are continually disrupted by
 marine traffic, homogenizing the sediments that
 have accumulated  since the last dredging (Plumb,
 1981).
    A core sampler is  basically a tube that is
 inserted  into the sediment by  various means to
 obtain a cylinder  or box sample of material at
'known depths. Corers can be simple, hand-oper-
 ated devices used by scuba divers, or they can be
    'Although grab samplers are sometimes referred to as "dredges,"
in  this document grab samplers are distinguished from dredge
samplers in that the grab samplers sample a discrete volume of
surface sediments in an area defined by the opening size of the
sampler's jaws, as opposed to the dredge sampler, which collects a .
composite of bottom .sediments as ill is dragged across the bottom.
2-8

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                                      2—QA/QC, Sampling, and Analytical Considerations
Cover Page (w/Approval Signatures)
Title Page
introduction      .
Table of Contents
Ust of Tables
List of Figures
List of Appendices
List of Acronyms and Abbreviations
Glossary

PROJECT DESCRIPTION
1.1 Introduction
1.2 Project Scope
1.3 Data Quality Objectives
1.4 Sample Network Design and Rationale
1.5 Project Implementation

PROJECT ORGANIZATION AND
RESPONSIBILITY
2.1 Organization
2.2 Authority and Responsibility
2.2.1   Project Oversight
2.2.2   Reid Activities
2.2.3   Laboratory Analyses
2.2.4  Other Regulatory Personnel
2.3  Project Communication

 QUALITY ASSURANCE OBJECTIVES
 3.1  Quality Assurance Documents
 3.2  Project Quality Assurance Objectives
 3.3  Field Measurement Quality Objectives
 3.3.1   Navigation
 3.3.2  Sample Collection Parameters
 3.3,3  Water Column Measurements
 3.4 Laboratory Data Quality Objectives
 3.5 Macrobenthic  Community Assessment
     Quality Assurance Objectives
 3.6 Computer Model Quality Assurance
     Objectives          .

 SAMPLE COLLECTION AND
 HANDLING PROCEDURES
 4.1  Sample Containers
 4.1.1  Volume and Type
 4.1.2 Quality Control and Storage
4.2 Sampling Procedures
4.2.1  Selection and Decontamination of
      Equipment
4.2.2  Sampling Methods
4.2.3  Collection of Sample
4.2.4  Sample Volume, Preservation, and
      Holding Times
4.2.5  Reid-Generated Waste Disposal
4.3 Sample Packaging and Shipment

SAMPLE DOCUMENTATION AND CUSTODY
5.1 Reid Procedures
5.1.1  Sample Labeling
5.1.2  Reid Logbooks
5.1.3  Reid Chain of Custody
5.1.4  Transfer of Custody
5.2 Laboratory Procedures
5.2.1  Sample Scheduling and Management
5.2.3  Sample Receipt and Handling
5.2:4  Log Books and Chain of Custody
5.2.5  Sample Disposal
5.3 Rnal Evidence File
5,3.1  Contents
5.3.2  Custody Procedure

CALIBRATION PROCEDURES AND
FREQUENCY
6.1  Reid Measurements     ;
6.1.1   Records and Traceability of Standards
6.1.2  Initial and Continuing Calibration
       Procedures                    ,
6.1.3  Conditions to Trigger Recalibration
6.2  Physical and Chemical Laboratory
     Analyses of Sediment
6.2.1   Records and Traceability of Standards
6.2.2  Preparation and Storage of Standards
6.2.3  Initial and Continuing Calibration
       Procedure
6.2.4  Conditions to Trigger Recalibration
6.3  Biological Effects Tests - Water Quality
     Monitoring
 6.3.1  Records and Traceability of Standards
 6.3.2  Initial and Continuing Calibration
       Procedure
 6.3.3  Conditions to Trigger Recalibration
                   Figure 2-1. Contents of a Quality Assurance Project Plan
                                                                                     2-9

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Sediment Classification Methods Compendium
7   MEASUREMENT PROCEDURE
    7.1  Reid Measurements
    7.1.1  Navigation
    7.1.2  Sample Collection Parameters
    7.1.2.1   Sediment
    7.1.2.2   Fish
    7.1.2.3   Benthic Organisms
    7.1.3  Water Column Measurements
    7.2  Chemical Analysis of Sediment
    7.2.1  Sample Preparation Methods
    7.2.2  Sample Extract Cleanup Methods
    7.2.3  Analytical Methods
    7.3  Other Sediment Analyses
    7.4  Biological  Effects Tests
    7.5  Macrobenthic Community Assessment
    7.6  Model Calculations

8   INTERNAL QUALITY CONTROL CHECKS
    8.1  Sample Collection
    8.2  Reid Measurements
    8.3  Chemical Analyses of Sediment
    8.4  Other Analyses of Sediment
    8.5  Biological  Effects Tests
    8.6  Macrobenthic Community Assessment
    8.7  Computer Model Calculations

9   DATA REDUCTION, VALIDATION, AND
    REPORTING
    9.1  Reid Measurements
    9.2  Laboratory Data
    9.2.1  Internal Data Reduction
    9.2.2  Data  Reporting Requirements
    9.2.3  External Data Validation
    9.3  Macrobenthic Community Assessment
    9.4  Computer Model Calculations
         i
10  PERFORMANCE AND SYSTEM AUDITS
    10.1  Audit Scheduling and Planning
    10.2 Internal Audits
    10.2.1  Reid Activities
    10.2.2  Laboratory Activities
    10.2.2.1  System
    10.2.2.2 Performance
   10.3 External Audits
   10.3.1  Reid Activities
   10.3.2  Laboratory Activities
   10.3.2.1  System
   10.3.2.2  Performance
   10.4 Audit Reports o

11 PREVENTIVE MAINTENANCE
   11.1 Reid Equipment
   11.2 Sample Collection Equipment
   11.3 Laboratory Instruments
   11.4 Computer Hardware and Software

12 SPECIRC ROUTINE PROCEDURES TO
   ASSESS DATA USABILITY    .
   12.1 Sample Collection
 ,  12.2 Reid and Laboratory Data
   12.2.1  Data Quality Indicators
   12.2.1.1  Sensitivity
   12.2.1.2  Precision
   12.2.1.3  Accuracy
   12.2.1.4  Completeness
   12.2.2  Other Data Review
   12.3 Macrobenthic Community Assessment
   12.4 Computer Model Calculations

13 CORRECTIVE ACTIONS
   13.1 Introduction
   13.2 Equipment Failures
   13.3 Procedural Problems
   13.4 Sample Custody Failures
   13.5 Documentation Deficiencies
   13.6 Data Anomalies
   13.7 Performance Audit Failures
   13.8 System Audit Failures

14 QUALITY ASSURANCE REPORTS TO
   MANAGEMENT
   14.1 Project-Specific Final Reports
   14.2 Deviation and Corrective Action Memos
   14.3 Internal and External Audit Reports

15 REFERENCES
                Rgure 2-1. Contents of a Quality Assurance Project Plan. (Continued)
2-10

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                                            2—QAIQC, Sampling, and Analytical Considerations
large, costly, motor-driven mechanisms that can
collect samples from great depths. A few types of
corers include a gravity corer, which uses weights
attached to the head of the sampling tube to push
the tube into the sediment; a piston corer, which
is similar to a gravity corer but also has a piston
inside the tube that remains stationary during sedi-
ment penetration and creates a vacuum that helps
pull the sampler into the sediment; a vibra-corer,
which is like a gravity corer except with.a vibrat-
ing head attached to  enhance penetration; and  a
multiple corer, which is an array of plastic tubes
attached to a frame, allowing for the collection of
several samples at the  same location.  Because
gravity corers can compact the sample and distort
the vertical profile, a piston corer or vibra-corer is
recommended to minimize sample compaction.
The corer that disturbs the sediments the least is
a box corer. Instead of being cylindrical, it is a
large box-shaped sampler that is deployed inside
a frame. After the frame  is brought to rest on the
bottom, heavy weights lower the open-ended box
into  the sediment. A  bottom door then swings
shut upon retrieval to prevent sample loss.   The
advantages of the box corer include its ability to
collect a large amount of sample with the center
 of the sample virtually undisturbed.   Corers are
not generally recommended for use in sandy sedi-
 ments  since  they have  difficulty retaining the
 sample upon withdrawal.
     A comparison of the general characteristics of
 various commonly used sediment-sampling devic-
 es for chemical, physical, arid biological studies is
 given in Baudo (1990); Plumb  (1981); Downing
 and Rigler (1984); and ASTM (1990).
 2.6.2  Sample Handling, Containers,
        Preservation, and Holding Times

 2.6.2.1 General Requirements

     Proper handling of the samples is essential to
 preserve the sample integrity and the validity of
 the results.  Mishandling of samples at any stage
 of the sample-collection process could distort
 analytical results, wasting the effort  and expense
 of the sampling survey.  Some of the basic con-
siderations in sediment sample handling include
the following (Plumb, 1981):

    •  It is essential that noncontaminated sam-
       pling devices are used and that obvious
       sources of contamination such as exhaust
       fumes from the collecting ship, lubricating
       drilling fluids, and powder from surgical
 -      gloves be eliminated.

    •  Sampling devices should be washed be-
       tween samples with an appropriate series
       of cleansers* and solvents to prevent cross-
       contamination from  one sample to the
       next:

    •  Analysis for different parameters requires
        different storage containers to  ensure
        noncontamination and to prevent degra-
        dation  of the sample.  Basic  rules for
        containers include using  plastic or  glass
     '  containers for metal  analysis, glass con-
        tainers for organic analysis, and glass or
        plastic for inorganic analysis.  Since no
        set guidelines have been determined for
        sediment sampling, a .good general rule to
        follow is to use containers recommended
        for water testing.

     •  A reliable and identifiable sample-labeling
        process should be used.

     »  Sampling containers should be filled to
        capacity, allowing only enough air  space
        for  possible expansion  of the sample
        resulting from the preservation technique
        (e.g.,  freezing) to eliminate or greatly
        reduce oxidation of the sample  (USEPA/-
        USACE,  1991).  Sample  containers for
        volatile organics analyses should be filled
         completely, allowing no headspace.

     Preservation methods are intended to maintain
 the integrity of the sample by limiting the deterio-
 ration or alteration of  a specified  parameter by
 hydrolysis, oxidation,  and/or biological activity
 while the sample awaits analysis.  Methods are
 basically limited to pH control, chemical addition
                                                                                             2-11

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 Sediment Classification Methods Compendium
 or fixation, sample extraction or  isolation,  or
 temperature control. Preservation steps should be
 initiated immediately after collection of the sample
 since significant alteration of the sample can occur
 in the first few hours after sampling.  Immediate-
 ly after collection, sediment samples are typically
 kept on ice or refrigerated.  Upon arrival at the
 laboratory, samples are usually preserved by
 drying, freezing, or  cold storage (ASTM, 1990).
    The type of preservation required will depend
 on the parameters being tested.  For example, if
 the sediment is to be tested for both bulk metals
 and particle size, either two  samples should be
 collected  or the sample should be split, since it is
 recommended that samples for bulk metal analysis
 be preserved by dry ice  and  stored at less  than
 -20°C, whereas samples to be analyzed for particle
 size should be refrigerated  at 4°C  (USEPA/
 USAGE,  1991).  For this  reason, it is essential to
 know which tests are to be performed, or poten-
 tially performed,  on the  samples in advance to
 allow for additional  sample collection or splitting
 of samples as needed to comply with differing
 sampling, handling, and preservation requirements.
    Freezing appears to be the generally preferred
 method for preserving sediment samples for most
 chemical  analysis, although  sediments to be used
 for particle size determination, volatile organics,
 and toxicity testing should not be frozen (ASTM,
 1990).

 2.6.2.2 Requirements for Specific Analyses

    There are basically  four  ways to analyze
 chemical  and physical parameters of sediments:
 bulk analysis, standard elutriate test,  fractionation
 procedures, and physical analysis.  Brief descrip-
 tions of these types of analyses follow, along with
 any special sample handling procedures, contain-
 ers, or preservation techniques needed.
    Bulk  analysis allows one to evaluate the total
 concentration  of a parameter  within a  sediment
 sample or the toxicity of the whole sediment
 Most chemical parameters are evaluated by bulk
 analysis.  In general, the collection container  and
 preservation and storage method are dependent on
 the parameter to be tested. Bulk analysis samples
 can be stored wet, air-dried, or frozen.  If trace
 organic constituents are to be analyzed, a glass.
 container should  be used  to  store the sample.
 When preserving and storing samples, one needs
 to  take into consideration that other parameters
 could change as a result of oxidation, volatiliza-
 tion, or chemical instability (Plumb, 1981).
    Elutriate tests indicate the ability of chemical
 constituents to migrate from the solid phase to the
 liquid phase.  An  elutriate sample is prepared by
 mixing or shaking sediment and water in  pre-
 scribed proportions for a prescribed period of time
 and separating  the  liquid  fraction  by filtration
 and/or centrifugatibn!   The liquid  fraction, the
 elutriate, is then analyzed by  methods used for
 analysis of water samples. Sediments to undergo
 elutriate testing should be stored wet, at 4°C, in
 airtight containers and should be tested as soon as
 possible following sample collection.   If  trace
 organic analyses are to be performed, glass con-
 tainers with Teflon lids are required for storage
 (Plumb, 1981).
  .  Fractionation procedures provide information
 on  the distribution of constituents.  The samples
 are extracted multiple  times using a series of
 extractants  and procedures,  thereby  isolating
 specific pollutants or classes of pollutants.  Pore
 water extraction is a form of fractionation where-
 by  the interstitial water in the whole sediment
 sample is extracted by squeezing or centrifugation.
 The resulting water sample can be used in chemi-
 cal and biological tests.  To date, fractionation has
 been  used primarily for research.   As a result,
 most  agencies  do not subject  their sediment
 samples to fractionation procedures (Plumb, 1981).
 However, some fractionation tests,  such as the
 toxicity identification evaluation (TIE), a fraction-
 ation procedure to isolate the toxic component of
 a sample,  are beginning  to be used to make
 decisions regarding regulatory actions and remedi-
 al approaches since  they  can be used to assess
 which pollutants are responsible for the toxicity
 observed in a sediment. Samples to be analyzed
lor fractionation should be stored wet, at 4°C, and
 in airtight containers. Testing procedures should
 start as soon as possible after  sample collection
 (Plumb, 1981).
    Physical analysis  provides information on
 particle size,  color, texture, and mineralogical
2-22

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                                           2—QA/QC, Sampling* and Analytical Considerations
characterization  and  includes tests  for  cation
exchange capacity, particle size, pH, temperature,
salinity, oxidation reduction potential, total volatile
solids, and specific gravity.  Samples to undergo
physical analyses may be stored wet, at 4°C, or
frozen, depending on  the parameter to be tested.
Some  of these parameters  (e.g., pH) should be
analyzed immediately upon collection.
    The  1991  Green Book  (USEPA/USACE,
1991) suggests the use of a grab sampler or corer
for collection of sediment samples and offers the
following general guidelines for preservation, and
handling and sample  sizes needed for sediment
samples  collected  for  chemical  and physical
testing:
    Bulk metals should be stored in nonreactive
containers, such as high-density polyethylene, and
analyzed as soon as possible.
    Bulk organics, including PCBs,  pesticides,
and high-molecular-weight hydrocarbons, should
be contained  in - solvent-rinsed  glass jars  with
Teflon lids, preserved by dry ice, and stored at
less than -20°C in the dark. The samples can be
stored for up to 10 days. Approximately 475 mL
of sample should be collected.
    Samples to be analyzed  for  total organic
carbon (TOC) should be preserved by dry ice and
stored at less than -20°C. They can be kept for an
undetermined amount of time.
    Sediments for particle size testing should be
kept refrigerated at 4°C in any jsealed container
and can be kept for an undetermined amount of
time.

2.63  Minimum Parameters to Be Tested

    Sampling efforts are performed with a variety
 of objectives in mind, and therefore the minimum
 chemical and physical parameter testing require-
ments vary between studies or programs. Howev-
 er, some chemical and physical  parameters seem
 to be common to several programs. They include
 particle or grain size, total organic carbon, heavy
 metals, acid volatile sulfides, polycyclic aromatic
 hydrocarbons,   polychlorinated  biphenyls,   and.
 pesticides.   Unionized  ammonia must  also be
 measured, taking into account its sensitivity to pH
 and temperature, both of which are affected by
sample manipulation.  When testing sediment
samples from estuarine or marine environments,
the analysis methods chosen must address salinity
since this can alter the analytical results (USEPA/-
USACE, 1991).
    Particle or grain size analysis is a physical
parameter that determines  the  distribution  of
particle sizes.  Methods for particle size analysis
aye suggested in Folk (1968), Buchanan (1984),
Plumb (1981),  ASTM (1990), and Tetra Tech
(1985). Plumb  (1981) suggests that analysis will
usually require  two or more methods, depending
on the range of particle  sizes encountered.  He
gives a detailed account  of the use of sieves in
conjunction  with electronic particle counters or.
sieves and pipet analysis.  Testing and Reporting
Requirements for Ocean  Disposal  of Dredge
Material off Southern  California under Marine
Protection, Research and Sanctuaries Act Section
103 Permits (Ocean Dredged Material Disposal
Program, 1991) recommends the method given in
Plumb (1981) for analysis of particle size.
    Total organic carbon (TOC) is an important
indicator of bioavailability for nonionic hydropho-
bic organic  pollutants. When analyzing for this
parameter, it is essential that the sample be stored
in a glass or plastic-container and that all air
bubbles be removed  from the sample before it is
sealed and stored. The method given in Plumb
(1981) is commonly recommended (Tetra Tech,
1985).   Plumb (1981)  suggests using  sample
ignition, which uses  a hydrochloric acid wash to
separate the inorganic and  organic carbon, or
differential  combustion,  which uses  thermal
combustion to  separate the two carbons by their
different  combustion temperature ranges.   The
 1991 Green Book recommends that the analytical
method to test for TOG be based on high-tempera-
ture combustion rather than on chemical oxidation.
Additionally, it recommends using sulfuric acid
rather than  hydrochloric  acid rinse.  Testing and
Reporting Requirements for Ocean Disposal of
Dredge Material  off Southern California under
Marine Protection, Research and Sanctuaries Act
Section  103  Permits recommends EPA  Test
 Method No. 9060 for  TOC determinations.  The
 method  recommended by EPA for use in apply-
 ing organic carbon-normalized sediment quality
                                                                                           2-13

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Sediment Classification Methods Compendium
criteria for nonionic hydrophobic organic chemi-
cals uses catalytic combustion and nondispersive
infrared detection (Leonard, 1991).
   Metals are found naturally occurring in the
environment, but an excess of metals can be an
indication of anthropogenic contamination. The
most commonly used method to analyze sediments
for metals is atomic absorption spectrophotometry.
Plumb (1981) details the use of the direct-flame
atomic absorption method for all metals except
arsenic, mercury, and selenium. For these metals,
he recommends using  arsine generation, cold
vapor technique, and  digestion/flameless atomic
absorption  or hydride generation, respectively.
The 1991 Green Book points out that the concen-
tration of salt in marine or estuarine samples may
cause interference in analysis for metals. There-
fore, the approach of an acid digestion  followed
by atomic absorption spectroscopy should  be
coupled with an appropriate technique to control
this interference. The 1991 Green Book recom-
mends USEPA (1986) for analysis of mercury and
EPRI (1986) for the analysis  of selenium and
arsenic.  Testing and Reporting Requirements for
Ocean Disposal of Dredge Material  off Southern
California under Marine Protection, Research and
Sanctuaries Act Section 103 Permits recommends
the following EPA Test Methods: cadmium (Nos.
7130, 7131); hexavalent chromium  (Nos. 7190,
7191); copper (No. 7210); lead (Nos. 7420,7421);
mercury (No. 7471); nickel (No. 7520); selenium
(Nos. 7740,7741);^ silver (No. 7760); and zinc
(No. 7950).
   Acid volatile sulfides (AYS) have been found
to be closely related to the toxicity of sediment-
associated metals (Di Toro et al., 1990).  AVS
have  been found  to  be  important in binding
potentially bioavailable metals, thereby  reducing
their toxicity.  The approved method is given in
USEPA (1991).
   Polyaromatic  hydrocarbons  (PAHs) are
semivolatile organic priority pollutants, a number
of which are potential carcinogens. Plumb (1981)
details the methods of  methanol extraction/UV
analysis and ethanol extraction/UV spectrophotom-
etry to analyze for this  parameter.  Testing and
Reporting Requirements for Ocean  Disposal of
Dredge Material off Southern  California under
Marine Protection, Research and Sanctuaries Act
Section  103  Permits recommends EPA Test
Method Nos. 8100,8250 and 8270 for analysis of
PAHs.
    Polychlorinated biphenyls (PCB) are chlori-
nated organic compounds that were once used for
numerous purposes including as a dielectric fluid
in electrical transformers. Desirable properties of
PCBs include low flamrnability, nonconductivity,
and nonreactivity. However, PCBs do not break
down readily and they bioaccumulate in the
environment.  The 1991 Green Book offers gas
chromatography/eledron-capture detection (GC/
ECD) methods as the primary tool for the analysis
of PCBs, or the use of GC/MS using selected ion
monitoring (SIM).  They do not recommend the
traditional methods of PCB analysis, which quan-
tify  PCBs as arochlor  mixtures.   Testing and
Reporting Requirements for Ocean Disposal of
Dredge  Material off Southern  California  under
Marine Protection, Research and Sanctuaries act
Section  103 Permits recommends the use  of the
methods described in  Tetra Tech  (1986) and
NOAA (1989) for analysis of PCBs.
    Pesticides are man-made  compounds pre-
dominantly used in agriculture to control crop-
damaging insects.  Some pesticides, especially
halogenated compounds, persist in the  environ-
ment and can contaminate the food chain.  Plumb
(1981) details the method of hexane extraction in
preparation for  testing, for  organophosphorus
pesticides.  The 1991 Green Book recommends
using GC/ECD  or GC/MS to analyze for  chlori-
nated pesticides. Testing and Reporting Require-
ments for Ocean Disposal of Dredge Material off
Southern California  under Marine Protection,
Research and Sanctuaries Act Section 103 Permits
recommends  EPA Test  Method  No.  8080  to
analyze  for pesticides.
    For  analyses of volatile organic pollutants
and  semivolatile organic pollutants,  the 1991
Green Book recommends the methods described
by  Tetra  Tech (1986),  which  should always
include the use of capillary-column GC or GC/MS
techniques. For volatiles, a purge-and-trap method
is used, followed by GC/MS analysis according to
U.S, EPA Method 624 or U.S. EPA Method 1624,
Rev. B,  Ref. 3 (Tetra Tech, 1986).
2-14

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                                            2-—QA/QC, Sampling, and Analytical Considerations
    As stated previously, the minimum set of
parameters tested in sediments varies and is based
on the sampling objectives of the program. Listed
below are several examples of minimum data sets
required by specific programs.
    The 1991 Green Book recommends that all
sediment  samples  be analyzed  for TOG, PAHs,
grain size, total solids/water content, and specific
gravity. The remaining parameters to be sampled
are compiled from the priority pollutants list based
on historical testing data, potential contaminants
due to known industries in the area, and a general
knowledge of the area  to be sampled.
    Testing  and  Reporting Requirements  for
Ocean Disposal of Dredge Material off Southern
California under Marine Protection, Research and
Sanctuaries Act Section 103 Permits has  very
specific  parameters and methods  required for
materials to be disposed of off the coast.   Re-
quired analyses for physical parameters include
grain size, total solids/water content, and specific
gravity.   Chemical  analyses includes 9 metals,
 ammonia, arsenic, total sulfides, acid volatized
 sulfides (AVS), 11 pesticides including total pesti-
 cides, 9 organic compounds, all PCB  congeners,
 individual totals of tetra-, penta- and hexa-chlorb-
 biphenyl isomers, and 17 PAHs.
     The  EPA  Environmental  Monitoring  and
 Assessment Program-Near  Coastal  (EMAP-NG)
 established guidelines  identified in its Near Coast-
 al Program Plan for 1990: Estuaries (Holland,
 1990) for sediment sampling for determination of
 contaminant levels. They include sample collec-
 tion by means of a Young-modified Van  Veen
 grab and, initially,  analyzing  the NOAA Status
 and Trends suite  of contaminants, which include
 chlorinated pesticides, PCBs,  PAHs, major ele-
 ments, and toxic metals.  EMAP-NC, with the
 assistance  of other programs,  plans to refine the
 list  of  contaminants  to include pesticides and
 herbicides and other toxic chemicals.

 2.6.4  Sampling for  Benthic  Community
        Structure in Fresh Water

      Macrobenthic organisms  play  an important
 role in marine, estuarine, and freshwater lotic and
  lentic ecosystems.   As major secondary con-
 sumers, they  represent  an  important  linkage
 between primary  producers and higher trophic
 levels for both planktonic and detritus-based food
 webs.  They are a significant food source for
 juvenile fish arid crustaceans and may improve
 water quality by filter-feeding of paniculate matter
 (Holland, 1990).  Benthic populations also repre-
 sent diverse taxa and can serve as  sentinels for
 environmental stress.  Benthic organisms access
 all  aspects of the aquatic habitat with varying
 feeding strategies, reproductive modes, life history
 characteristics, and physiological  tolerances to
 environmental conditions.  Most benthic organ-
 isms have  limited mobility and  cannot avoid
 environmental stressors. As a result, the responses
 of some species serve as indicators of changes in
 sediment quality  (Holland,  1990).  This section
 will detail  specific procedures  and precautions
 necessary for proper  conduct of benthic sample
 collection and handling in freshwater, marine, arid
. estuarine ecosystems.

 2.6.4.1 Sample Collection Methods

      It is  helpful to  consult Macroinvertebrate
 Field and laboratory Methods for Evaluating the
 Biological Integrity of Surface Waters (Klemm et.
  al, 1990), which thoroughly addresses methodolo-
  gy.   State  environmental regulatory programs
  should have a Quality Assurance  Program Plan
  describing the field methods and standard operat-
  ing  procedures  for  collecting and evaluating
  benthic macroirivertebrates.   This information
  should be  obtained to  ensure acceptance  and
  comparability of study results with those obtained
  by the state agency.  If this information  is not
  available, then field methods and standard operat-
  ing procedures  from  other existing  programs
  should be used.
  .In soft freshwater sediments, the most com-
  mon method used to collect benthos is with a grab
  sampler such as a Ponar (15 x 15  cm or 23 x 23
  cm) or Ekman grab sampler (15 x 15 cm, 23 x 23
  cm, or 30 x 30 cm), each of which provides a
  quantitative sample based on the surface area  of
  the sampler. The smaller of the sampler sizes are
  most  commonly  used for freshwater studies
  because  of their relative ease of manipulation.
                                                                                             2-15

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 Sediment Classification Methods Compendium
 The Ekman grab sampler is not as effective  in
 areas of vegetative debris but is much lighter than
 the Ponar and easier to use in softer substrates.
 Artificial substrates (Hester-Dendy using several
 3-inch plates and spacers attached by an eyebolt,
 or substrate/rock-filled baskets) provide consistent
 habitat for the benthos to colonize  in both soft-
 bottomed  and stony areas.  Artificial substrates
 can be used in almost any water body and have
 been  successfully  used  to standardize results
 despite  habitat  differences  (Ohio  EPA,  1989;
 Rosenberg and Resh, 1982; and Resh and Jackson,
 1991).
     A variety of methods for sampling benthos  in
 hard-bottomed lotic systems are available, includ-
 ing artificial substrates.    If  quantification by
 sediment or sampler surface area is needed,  a
 Surber-type square-foot sampler with a Standard
 #30-mesh   (0.589-mm  openings) can  be  used
 (Klemm et al, 1990). The traveling kick-net (or
 dip-net) method, also using a #30-mesh net, can
 be used to quantify the sample collected by the
 amount of time spent sampling and the approxi-
 mate surface area sampled (Pollard, 1981; Pollard
 and Kinney,  1979).  The Surber-type and kick
 methods can each be used to provide consistent,
 reproducible  samples, but both  are limited  to
 wadable streams.  The Surber sampler's optimal
 effectiveness is limited to riffles, whereas kick-net
 or dip-net  samplers can be effectively used in all
 available habitats.  Although dip-net  samplers
 have been effectively used to sample riffles and
 other relatively shallow habitats to determine taxa
 richness, presence of indicator organisms, relative
 abundances, similarity between sites, and other
 information, they do not  provide  definitive esti-
 mates of the number of individuals or biomass per
 surface area.


 2.6.4.2 Sample Handling  and Preservation

    The following decisions will need to be made
 once the sample  collection  method is. chosen:
 (1) whether samples will  be picked from debris
 and  sorted in the field, (2) which  preservative
 should be used, (3) whether a stain (rose bengal)
 or other material will be added to the sample to
 facilitate separating the organisms from debris,
 (4) the type of sample containers and labeling of
 the  containers required,  and (5) the  mode of
 transportation of the samples to their destination.
 Many of these decisions are based on professional
 preference or the required logistics of the study.
    Sorting of the benthos from debris and preser-
 vation are fully discussed by Klemm et al. (1990).
 American Public Health Association et al. (1989)
 and Klemm et al (1990)  defined the benthos by
 what is retained on a standard #30 sieve.  How-
 ever, some types of Chii onomidae and other small
 benthos pass  through a #30-mesh sieve but are
 retained by a #40-mesh  sieve.  It has been recom-
 mended  that samples  should  first  be passed
 through a #30-mesh sieve.  Then the  materials
 washed through should  be passed through a #40-
 mesh sieve, and the materials retained in both
 sieves should be sorted (Ohio EPA, 1989).  Once
 the material is washed through the sieves, the
 organisms should be separated from the vegetation
 and other debris in a white enamel pan. As the
 materials are separated,  the organisms can be
 placed  in different vials for the major  taxa.
 Preservation with either formalin  or 70 percent
 ethanol is common. . Although formalin  is an
 excellent fixative,  the  human  health  concerns
 associated with its use require extreme caution and
 adequate ventilation. Many programs rely oil 70
 percent ethanol as a fixative and preservative.
    A practical technical reference  that details
 procedures  for cost-effective  biological assess-
 ments of lotic systems has been developed. Rapid
Bioassessment Protocols for Use in Streams and
Rivers: Benthic  Macroinvertebrates  and Fish
 (Plafkin et al, 1989) presents three berithic rapid
bioassessment protocols (RBPs) and  two fish
 RBPs, with a progressive order of increasing rigor
in evaluation within each series for each class of
organisms.
   The RBPs are based on integrated assessments
that compare physical conditions of habitat (e.g.,
physical structure, flow regime) and biological
measures of reference conditions. These reference
conditions are derived after systematic monitoring
of sites that represent the natural range of varia-
tion in  water  chemistry, habitat, and biological
condition.
2-16

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                                            2-H3/VQC Sampling, and Analytical Considerations
    The  functional and structural components
evaluated for aquatic communities comprise eight
metrics for benthic RBPs and  12 metrics for the
fish RBPs.   Examples of  metrics for benthic
communities include the following: taxa richness,
the modified Helsenhoff Biotic Index (summarizes
overall pollution (tolerance of the benthic arthropod
community  with  a single .value; this  index was
modified to include nonarthropod species as well),
ratio of scraper and filtering collector functional
feeding groups, ratios of the number of organisms
in  the EFT  (Ephemeroptera,  Plecoptera,  and
Trichoptera) to  the  number  of Chironomidae
present,  and community similarity indexes.  The
fish protocol is  based on the  index of biotic
integrity (IBI) or a fish community  assessment
approach developed by Karr  et al. (1981).  As
with the approach of metrics in the benthic evalu-
ations, the metrics of the fish protocol represent
differing sensitivities.

2.6.5  Sampling for Benthic Community
       Structure in Marine and Estuarine
       Waters

     Historically,  regional monitoring  programs
 have used benthic community studies as an effec-
 tive indicator of the extent of pollution impacts on
 marine  and estuarine ecosystems, as well as the
 effectiveness of management actions.  In addition,
 information on changes in benthic population and,
 community parameters due to sediment character-
 istics can be used to distinguish natural variation
 from changes due to human activities (Holland,
 1990).

 2.6.5.1  Sample  Collection Methods

     Three  grab  samples, are collected for benthic
 species composition,  abundance, and biomass.
 Additional sediment grabs are collected for chemi-
 cal analyses and for use in acute toxicity tests. To
 minimize the possibility of biasing results, benthic
 biology grabs should not be collected consecutive-
 ly, but rather  interspersed  among  the chemis^
 tryAoxicity grabs. While a biology grab is being
 processed (sieved), grabs should be  collected for
  chemistryAoxicity (Holland,  1990).
    A 1/2s"m2, stainless steel, Young-modified Van
Veen grab sampler may be used to collect sedi-
ments for benthic analyses. The sampler is con-
structed entirely  of stainless steel and has been
coated  with Kynar (similar to Teflon) and  is,
therefore,  appropriate  for collecting sediment
samples for both biological and chemical analyses.
The top of the sampler is hinged to allow for the
removal of the top layer of sediment for chemical
 and toxicity analyses.  This gear is relatively easy
 to operate and requires little specialized training.
 To minimize the chance  of sampling the exact
 same location twice," the  boat should be moved
 5 meters downstream after three grabs have been
 taken, whether successful or not (Holland, 1990).

 2.6.5.2  Sample Handling and Preservation

     Grab samples to be used in the assessment of
 macroinvertebrate  communities are" processed by
 first extracting a core sample from the sampler.
 The depth of sediment at the middle of the sam-
 pler should be at least 7 cm. Descriptive informa-
 tion on the grab is recorded;  The depth to the
 black layer of sediment within the core, the redox
 potential discontinuity (RPD), is measured in the
 field.  The sample is then extruded from the core
 tube to fill a whirl pac bag, labeled, and recorded.
 The sample should be refrigerated  at 4°C, not
 frozen (Holland, 1990).
      The remainder of the grab is processed for
 benthic community  analysis.  The sediments are
  transferred into a basin and then into a 0.5-mm
  mesh sieve.  The sieve is agitated to wash away
  sediments and  leave organisms, detritus,  sand
  particles,  and pebbles larger than 05 mm.  A
  gentle flow of water over the sample is acceptable,
.. but forceful jets of water should  be  avoided
  because  they can cause mechanical damage to
  fauna. The organisms are rinsed and transferred
  from the sieve  into a jar and covered in  seawater
  with MgCl added.  This "relaxes" the organisms,
  reducing damage from addition of the preservative
  (Holland, 1990).  Ten percent buffered formalin is
  used to fix and preserve samples. After 30 min-
  utes in the relaxant, formalin with a small amount
   of borax should be added to each sample jar. The
  jar is filled to the rim with seawater to  eliminate
                                                                                             2-17

-------
 Sediment Classification Methods Compendium
 any air space, eliminating the problem of organ-
 isms sticking to the cap during shipment.  Prior to
 sieving the next sample, the sieve is rinsed and
 brushed thoroughly to prevent cross-contamination
 of samples.                  "

 2.6.6 Sampling for Bioassays and Toxiciity
       Testing

    Environmental impacts on marine ecosystems
 are primarily assessed and monitored using the
 tools outlined in the 1991 Green Book.  The 1991
 Green Book is used to make decisions regarding
 the suitability  of dredged material for  ocean
 dumping.  EPA and the USAGE have shown that
 the greatest potential for environmental impact
 from dredged material disposal is on the benthic
 environment since benthic organisms burrow into
 and  are exposed to  sediments  and associated
 contaminants for extended periods of time.  The
 1991 Green Book uses whole sediment bioassays
 to evaluate potential impacts of dredged sediments
 and, in concert with the identification of contami-
 nants of concern through chemical analysis, serves
 to determine the extent and type of bioavailability.
 In addition, sediment toxicity tests can be used to
 assess spatial and temporal changes in toxicity in
 contaminated areas, rank sediments based on their
 toxicity to benthic organisms, and define cleanup
 goals for contaminated areas.  This section will
 highlight some of the collection and handling
 methods of sediments for toxicity testing and
 whole sediment bioassays,

 2.6.6.1 Sample  Collection, Handling, and
       Preservation

    The  sediment environment is composed of
 many microenvironments, redox gradients, and
 interacting physicochemical and  biological  pro-
 cesses.  Many of these characteristics  influence
 sediment toxicity  and bioavailability to benthic
 and planktonic organisms, microbial degradation,
 and chemical sorption.  Maintaining the integrity
 of a sediment sample during its removal, transport,
 storage, and testing in the laboratory is extremely
 difficult.   Any disruption of  this environment
 complicates interpretations of treatment effects,
 causative factors, and in situ comparisons (ASTM,
 1990).
     Sample handling,  preservation, and storage
 techniques have to be  designed to minimize any
 changes in composition of the sample by retarding
 chemical and/or biological activity and by avoid-
 ing  contamination.   Sufficient  sample volume
 must be collected to perform the necessary analy-
 ses, partition the samples for  respective storage
 requirements, and archive portions of the sample
 for possible later  analysis.    Core sampling is
 recommended to best maintain  the integrity of the
 sediment for studies' of sediment toxicity, inter-
 stitial  waters,  microbiological  processes,  and
 chemical  fate.   Subsampling, compositing, or
 homogenization of sediment  samples  may  be
 necessary depending  on  the  study objectives.
 Subsamples of  the inner core area may be taken
 since this area is more likely to retain its  integrity
 and depth profile and not be contaminated by the
 sampler.  The loss of sediment integrity and depth
 profile is an important consideration, as are chang-
 es in chemical  speciation through  oxidation and
 reduction resulting in volatilization, sorption, or
 desorption; changes in biological activity; com-
 pleteness  of mixing;  and  sampling container
 contamination (ASTM, 1990).
     Subsamples of the top  1  or 2 cm  may be
 collected with a nonreactive sampling tool (e.g.,
 polytetrafluoroethylene  (PTF)-lined  calibration
 scoop). Some studies may require a composite of
 single sediment samples, which usually consist of
 three to five grab samples.  Subsamples should be
 collected with a Teflon  paddle,  placed in a nonre-
 active bowl or  pan, and stirred until the texture
 and color appear uniform.  The sediments should
 be removed and partitioned for chemical and AVS
 analysis.   Samples should  completely  fill  the
 storage containers, leaving no airspace.  If the
 sample is to be frozen, just  enough air space
 should  be allowed for  expansion to take place.
 The labeling system should be  tested prior to use
• in the field, making sure that labels can withstand
 soaking, drying, and  freezing without becoming
 detached or illegible (USEPA/USACE, 1991).
     Maintaining clean and uncontaminated sam-
 pling equipment between samples is necessary. It
 is important to clean the sampling device, scoop,
2-18

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                                            2—QA/QC, Sampling, and Analytical Considerations
spatula, and/or mixing bowls between sites.   A
suggested cleaning procedure includes a soap-and-
water wasli followed by an organic solvent rinse
(ASTM, 1990).                    ,
    The choice of sample containers for sediment
should consider the type of sediment, storage time,
chemical sorptioh, and sample composition.  For
sediments containing organics, brown borosilicate
glass containers with Teflon lid liners are optimal,
whereas plastic or polycarbonate  containers  are
recommended  for  metal-containing  sediments.
PTF or high-density polyethylene containers  are
relatively inert and are  suggested for use with
samples contaminated  with  multiple chemical
types (ASTM, 1990).
     Sediment samples for biological testing should
be press-sieved.through a  1-mm mesh screen to
- remove all living organisms from the sediment
 prior to testing.   Other matter retained  on  the
 screen  with the  organisms, such as  shell frag-
 ments,, gravel, and debris, should be recorded and
 discarded.  Sediment samples for use in bioassays
 should be well mixed.
     Since the first few hours are the most critical
 to  changes in the sample, preservation steps
 should be  taken  Immediately  upon  sediment:
 collection.  There is no universal  preservation or
 storage technique, and a technique for one group
 of  analyses may interfere with  other  analyses.
 Problems can be overcome by collecting sufficient
 sample volume  to  use  specific preservation or.
 storage techniques for specific analytes or tests on
 subsamples.   Preservation, whether by refrig-
 eration, freezing, or addition of chemicals, should
 be  accomplished in the field whenever possible.
 If final preservation techniques cannot be imple-
 mented in the field, samples should be temporarily
 preserved in a manner that retains the integrity of
 the sample.  Sediment  samples  for biological
 analysis should be preserved at 4°C, never frozen
 or  dried.   Field refrigeration  is  easily  accom-
 plished with coolers and  ice; however, samples
 should be segregated from melting ice or cooling
 water.
     Storage containers can be the same as the
 transport containers, and where sediments contain
 volatile compounds, transport and storage should
 be in airtight PTF or glass  containers with PTF-
lined screw,caps.  Exposure of sediments to air
should also be prevented in the handling of AVS--
containing sediments.  AVS is the reactive sulfide
pool that can reduce metal toxicity by binding
metals in anoxic sediments.  Oxidation of these
sediments can either increase toxicity by disassoci-
ation of the AVS-metal complex and precipitation
of the metal species,  or reduce toxicity if the
AVS-metal complex  should volatilize  (ASTM,
1990).     -
    It has been found that sediments can be stored
at 4°C without significant alterations in toxicity.
Completion of testing within a 2-week storage
period  is  recommended, but limits on storage
time will depend  on sediment and contaminant
characteristics (ASTM, 1990).            :
 2.7 REFERENCES

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    and sanctuaries act, section 103 permits.
 Ohio  EPA.   1989.  Biological  criteria for  the
    protection of aquatic life: Volume in. Stan-
    dardized biological field sampling and labora-
    tory methods for assessing fish and macroinv-
    ertebrate communities.  Division of  Water
    Quality Planning and Assessment, Ecological
    Assessment Section, Columbus, Ohio.
Plafkin, J.L.; M.T.  Barbour, K.D. Porter, S.K.
    Gross, and R.M. Hughs. 1989. Rapid bioass-
    essment protocols  for use  in streams  and
    rivers:  Benthic macroinvertebrates and fish.
    U.S. Environmental Protection Agency, Office
    of Water, EPA/444(440)/4-39-001, Washing-
    ton, DC.
Plumb, R.H., Jr. 1981.  Procedure for handling
    and chemical analysis of sediment and water
2-20

-------
                                           2—QA/QC, Sampling, and Analytical Considerations
   samples.  Tech.  Rep.  EPA/CE-81-1.  Pre-
   pared by Great Lakes Laboratory, State Uni-
   versity College at Buffalo, NY, for the U.S.
   Environmental  Protection Agency/Corps of
   Engineers Technical Committee on Criteria
   for Dredged and Fill Material. Published by
   theU.S. Army Engineer Waterways Experi-
   ment Station, Vicksburg, Mississippi.
Pollard, J.E.   1981.    Investigator  differences
   associated with a kicking method for sampling
   macroinvertebrates.    J.  Freshwater  Ecol.
    1:215-224.
Pollard, J.E., and W.L. Kinney.  1979.  Assess-
    ment of macroinvertebrate monitoring tech-
 ,   niques in an energy development area: A test
    of the  efficiency of three macroinvertebrate
    sampling methods in the White River.  U.S.
    Environmental Protection Agency, Office of
    Research and Development, Las Vegas, NV.
    EPA-600/7-79/163.
Resh,  V.H., and J.K. Jackson.   1991.   Rapid
    assessment approaches to biomonitoring using
    benthic macroinvertebrates.  In:  Freshwater
    Biomonitoring  and Benthic Macroinverte-
    brates. D.M. Rosenberg and V.H? Resh (eds.).
    Chapman and Hall, New York Press.
 Rosenberg, D.M., and V.H. Resh. 1982. The use
    of artificial substrates in the study of freshwa-
    ter benthic macroinvertebrates. In:  Artificial
    Substrates.  J. Cairns, Jr. (ed.).  Ann Arbor
    Science Publisher, Ann Arbor, MI.
 Ryti,  R.T., and  D.  Neptune.  1991.  Planning
    issues for superfund site remediation.   Haz-
    ardous Material Control,  November/Decem-
    ber, 1991. pp. 47-53.         ,'.-."
 TetraTech. 1985. Summary of U.S. EPA-appr-
     oved methods, standard methods,  and other
     guidance for 301(h)  monitoring  variables.
     Final Report, EPA Contract No. 68-01-6938.
 Tetra Tech. 1986.  Analytical methods for U.S.
     EPA  priority pollutants  and 301(h) pesti-
     cides in estuarine and marine sediments.
    Final Report, EPA Contract No.  68-01-69-
   .38!  '•''  '          ,;   .-'     ;  '
USEPA/USACE.  1991.  Evaluation of dredged
    material proposed for ocean disposal-testing
    manual.   U.S.    Environmental Protection
    Agency and U.S.  Army Corps of Engineers.
USEPA.  1980.  Interim guidelines and speci-
    fications for preparing  quality  assurance
    project plans, U.S. Environmental Protection
    Agency, Office of Monitoring-Systems and
    Quality Assurance, Office of Research and
    Development, Publication Number QAMS-
    005/80y December 29, 1980.
USEPA.  1983.  Guidelines and specifications for
    preparing  quality  assurance  program  plans.
    U.S.    Environmental  Protection Agency,
    Office of Research and  Development, Quality
    Assurance Management Staff.
 USEPA.  1986.  Test methods for evaluating solid
    waste.  U.S. Environmental Protection Agen-
    cy, -Office of  Solid Waste  and Emergency
    Response, Washington, DC.
 USEPA.    1991.,  Draft  analytical  method  for
    determination of acid volatile sulfide (AVS) in
    sediment, proposed technical basis for estab-
    lishing sediment  quality criteria for nonionic
    organic chemicals using equilibrium partition-
    ing, U.S. Environmental Protection Agency,
     Criteria and Standards  Division, Washington,
•'•  '  DC.    ,        .".'•.'        "   •'•
 USEPA.  undated.     Cost-Efficient  sampling
     schemes  for marine  benthic  communities.
     U,S.   Environmental Protection  Agency,
     Environmental Research Laboratory - Narrag-,
     ansett and Environmental Research Laboratory
     Newport, Publication Number ERLN-N156.
 Valente,R.,  andJ. Schoenherr.  1991. Environ-
     mental monitoring and assessment program,
     near coastal Virginian  Province, quality assur-
     ance project plan.  Environmental Research
     Laboratory, U.S.  Environmental Protection
     Agency.
                                                                                           2-21

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          CHAPTER 3
Bulk  Sediment Toxicity  Test Approach

Nelson Thomas                                        ,_,,.,
U S  Environmental Protection Agency, Environmental Research Laboratory
6201 Congdon Blvd., Duluth MN 55804, (218) 720-5702

Janet O. Lamberson and Richard C. Swartz
U S  Environmental Protection Agency, Pacific Ecosystems Branch, EHL-N
2111SE Marine Science Dr., Newport, OR 97365-5260, (503)867-4031
    In the bulk sediment toxicity test (BSTT)
 approach, test organisms  are  exposed in  the
 laboratory to sediments collected in the field. To
 measure toxicity,  a specific biological endpoint
 is used to assess the response of the organisms to
 the sediments.  The bulk sediment toxicity ap-
 proach is a descriptive method and; cannot be
 used by itself to generate sediment quality crite-
 ria.
 3.1   SPECIFIC APPLICATIONS

 3.1.1  Current Use

     Sediment toxicity testing has been applied in
 dredged  material disposal  permit  and  other
 regulatory  programs in  the  following  ways
 (USEPA/USACE, 1991).

     •  To determine potential biological hazards
        of dredged material intended for disposal
        in an aquatic environment;

     •  To evaluate the effectiveness of various
        dredged material management actions;

     •  To indicate the spatial  distribution of
        toxicity in contaminated areas, the rela-
        tive degree of toxicity, and the changes
        in toxicity along a gradient of pollution
        or with respect to distance from pollutant
        sources  (Scott  and  Redmond,  1989;
        Swartz et al., 1982, 1985b);

      • To reveal temporal changes in toxicity
         (i.e., by sampling  the  same locations
      over time or by assaying layers of buried
      sediment in core samples) (Swartz etal.,
      1986, 1991);

   •  To reveal hot spots of contaminated sedi-
      ment for further investigation (Chapman,
      1986); and

   •  To rank sediments based on  toxicity to
      benthic organisms and to define cleanup
      boundaries of small or large problem
      areas of contaminated sediment, i

   BSTT integrates interactions among complex
mixtures of contaminants that may be present in
the field.  Many classes of chemical contami-
nants, including  metals,  polycyclic aromatic
hydrocarbons (PAHs), polychloririated biphenyls
(PCBs), dioxins, and chlorinated pesticides can
contribute to toxicity in effluents and sediments
(Chapman et al., 1982).  The BSTT measures the
total toxic effect of all contaminants, regardless
of their physical and chemical composition.
3.1.2 Potential Use

    By itself, BSTT cannot generate chemical-
specific toxic effects data, but it can determine
toxicity.   Used in conjunction  with  toxicity
identification evaluation procedures (Ankley et
al., 1990) such as those described in Chapters 5,
 10,  and  il, BSTT could help identify causal
toxicants. To generate sediment quality criteria,
.the  procedure  must be  combined with other
 methods of estimating sediment quality such as
 the  Triad (Chapman,  1986b; Chapman et al.,

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 Sediment Classification Methods Compendium
 1987; see Chapter 10) and the Apparent Effects
 Threshold (AET) approach (Tetra Tech,  1986;
 FIT, 1988; see Chapter 11). BSTT will be most
 valuable in verifying other methods  used to
 develop sediment quality criteria.
 3.2 DESCRIPTION

 3.2.1  Description of Methods

    The toxicological approach involves exposing
 test organisms to sediments.  The chemical com-
 position of the sediments, which may be complex,
 need not be known.  At the  end of a specified
 time period, the response of the test organisms is
 examined  in  relation to a specified biological
 endpoint (e.g.,  mortality,  growth, reproduction,
 cytotoxicity, alterations in development or respira-
 tion rate).  Results are then statistically compared
 with  control  and  reference sediment results  to
 estimate sediment toxicity.
3.2.1.1 Objectives and Assumptions

    The objective of BSTT is to derive toxicity
data that can be used to predict whether the test
sediment will be harmful to benthic biota. It is
assumed that the behavior of chemicals  in test
sediments in the laboratory is similar to that in
natural in situ sediments. The effects of various
interactions (e.g., synergism, additivity,  antagon-
ism) among chemicals in the field or in dredged
materials can be predicted from laboratory results
without  measuring total or bioavailable concen-
trations of potentially hundreds of contaminants in
the test sediment (Swartz et al, 1989) and without
a priori knowledge of specific pathways of inter-
action between sediments and test organisms
(Kemp and Swartz, 1989). One of the strengths
of this test is to integrate the effects of all contam-
inants. However, the effect of individual contami-
nants cannot be determined by BSTT, therefore
limiting  its Use in source control.  This method
can be used for all classes of sediments and any
chemical contaminants,  but not to answer cause-
and-effect questions.
 3.2.1.2 Level of Effort

     Implementation of this procedure requires a
 moderate amount of laboratory effort.  A variety
 of toxicity test procedures (see Methods below)
 have been developed and are fairly straightforward
 and well documented.

 3.2.1.2.1 Type of Sampling Required'

     It is recommended that bulk sediments be
 collected for analysis of total solids, acid volatile
 sulfide, grain size, and total and dissolved organic
 carbon (ASTM, 1990a).   Bulk and interstitial
 concentrations of chemicals of interest can be
 determined in subsamples of the sediment added
 to  the toxicity test chambers  to  enhance  the
 interpretation of toxicity results.  However, meth-
 ods for sampling interstitial water have not been
 standardized (ASTM, 1990b). Sediment variables
 such as pH and Eh should also be monitored.

 3.2.1.2.2 Methods

    The American Society for Testing and Materi-
 als (ASTM) has developed standard guidelines for
 several BSTTs (ASTM, 1990a, 1991). The most
 commonly used of these partial life cycle tests
 feature the marine amphipods Rhepoxynius abroni-
 us, Eohaustorius estuarius,Ampelisca abdita, and
 Grandidierella japonica  (ASTM,  1990a);  Hie
 freshwater/estuarine amphipod Hyalella azteca
 (ASTM, 1990c); and the freshwater chironomid
 species  Chironomus tentans and  C.  riparius
 (ASTM, 1990c). Brief generalized descriptions of
 these tests are given below.
    BSTTs with, the two freshwater chironomid
 species are  functionally very similar, differing
 only in the age of the organisms with which the
test is initiated and the duration of the test. Both
 C.  tentans and  C.  riparius are available from
various aquatic toxicology laboratories and com-
mercial sources, and both  species are cultured
easily in a laboratory setting.  Toxicity tests are
initiated by adding C. riparius <3 days old or C.
tentans 10-14 days  old (second instar)  to test
chambers that contain bulk sediment with over-
lying water in various ratios (e.g., 6  water: 1
5-2

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                                                                            3—BSTT Approach
sediment; Giesy et. al, 1988).  The length of the
test also varies  with  the biological endpoint of
interest and the species used.  If the biological
endpoint of interest is growth and survival of the
larvae, the test is terminated after 10-14 days by
sieving the C. riparius or C.  tentans from the
sediment.  It also is possible to conduct the test
until the adults emerge, which will occur (depend-
ing on temperature) in approximately 30 days for
C. riparius and 20-25 days for C tentans. Toxi-
city test  procedures  with C.  riparius  and C.
tentans are given in more detail in Adams 'et al.
(1985), Nebeker et al. (1984), Qiesy et al. (1988),
Ingersoll and Nelson (1989), and ASTM (1991).
    Partial life-cycle toxicity tests with tie firesh-
water/estuarine  amphipod  H.  azteca  and  bulk
sediments have been conducted in a number of
laboratories. H. azteca are available from various
aquatic toxicology  laboratories and commercial
sources and can be cultured easily in a laboratory.
Toxicity tests are initiated by adding juveniles <7
days old to test chambers that contain bulk  sedi-
ment with overlying water in various ratios  (e.g.,
4 water:! sediment; Ingersoll and Nelson, 1989).
The length of the test can range from slO  days
(short-term partial life-cycle test) to 30 days (long-
term partial life-cycle test) (Nebeker et al., 1984;
Ligersoil and Nelson, 1989).  Depending on the
length  of the test, biological  endpoints include
survival, behavior,  growth,  and  reproduction:
More detailed, descriptions of toxicity  test proce-
dures are given by Nebeker et al. (1984), Nebeker
and Miller (1988), Ingersoll and Nelson (1989),
and ASTM (1991).
     Partial life-cycle toxicity tests with the marine
 amphipods Rhepoxynius abronius, Eohaustorius
 estuarius, Ampelisca abdita, and Grandidierella
japonica and bulk sediments have been used for
 some time (Swartz et al., 1985a). Amphipods and
 bulk sediments generally are collected from the
 field and acclimated to laboratory conditions for
 2-24 days before toxicity testing.  The tests are
 initiated by adding immature or adult amphipods
 to test chambers that contain bulk sediment with
 overlying water in various ratios.  The length  of
 the test generally is felO days, and the biological
 responses monitored consist of behavioral effects
 (e.g., emergence from the  sediment, ability  to
burrow in clean sediment after exposure to test
sediment) and mortality. More detailed descrip-
tions of the toxicity test procedures are given by
Swartz etal. (1985a), DeWitt et al. (1989), Nipper
et al. (1989), Scott and Redmond (1989), ASTM
(1990a), and the Puget Sound Estuary Program
(1991).  Chronic test procedures for marine and
estuarine amphipods  are  under development at
several laboratories.   Other test procedures for
marine and estuarine  polychaetes, pelecypods,
shrimp,  and fish are  described in the USEPA/
USAGE (1991) and Reish and LeMay (1988)
manuals  for  testing  dredged  materials before
disposal.

32.1.23 Types of Data Required

     The physical and chemical data  described
above under Section 3.2.1.2.1, Type of Sampling
Required, are needed to interpret the test results.
The required biological data (which vary by test)
may  include  mortality and  various  sublethal
 effects (e.g., changes in  growth, reproduction,
 respiration rate, behavior, or development). These
 data can be compared to control  and reference
 data to determine the occurrence of biological
 effects (ASTM, 1990a). Dilution experiments m
 which uncontaminated sediment is added to test
 sediment collected from the field can be used to
 calculate  LCa values, ECjo values, no-effect
. concentrations,   and   lowestrobservable-effect
 concentrations (Swartz et al., 1989).

 3.2.1.2.4  Necessary Hardware and. Skills

     In general, only readily available and inexpen-
 sive field and laboratory equipment is needed,
 procedures are fairly simple and straightforward,
 and a minimum of training is necessary to detect
 endpoints through toxicity tests: Interpretation of
 the toxicity data (chemical and biological) requires
 a higher degree of skill and training.  Chemical
 sampling methods are generally simple and rou-
 tine, although analysis of chemical samples re-
 quires specialized training and equipment Some
 biological effects tests also require specialized
 training, handling, and facilities.
                                                                                              3-3

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Sediment Classification Methods Compendium
3.2.1.3 Adequacy of Documentation

    Various sediment toxicity test procedures have
been developed and well documented for testing
field sediments (ASTM 1990a; Chapman 1986a,
1988; Lamberson and  Swartz, 1988;  Melzian,
1990; Puget Sound Estuary Program (PSEP) 1991;
Swartz, 1987; Thompson^* al., 1989; USEPA/
USAGE,  1991).   Although  standardization  of
methodology  is  progressing,  intercalibration
among laboratories and better field validation are
needed.
3.2.2 Applicability of Method to Human
      Health, Aquatic Life, or Wildlife
   •  Protection

    The  BSTT  approach is  suitable  only for
protection of aquatic life.  Sediment toxicity test
procedures incorporate a direct measure of sedi-
ment biological effects and can be used to predict
biological  effects  of contaminated  sediments
before approval  of state and federal  disposal
permits.  These procedures can be used to assess
the toxicity of sediments in the natural  environ-
ment and to predict the effects of these sediments
on resident aquatic life.  Combined with  other
approaches  such  as  the AET  and  the Triad
approaches  (Chapman,  1986b),  BSTTs  can be
used to establish sediment quality criteria. Use of
the most sensitive species within a benthic com-
munity as a test organism will serve to protect the
structure and function of the entire  ecosystem
(Becker et al., 1990).
3.23 Ability of Method to Generate
      Numerical Criteria for Specific.
      Chemicals

    The BSTT approach cannot be used by itself
to generate sediment quality criteria. Instead it
must be combined with chemical measurements
and other  data to generate information on the
effects of individual contaminants. Both the Triad
and the AET  approaches rely on bulk sediment
toxicity data to derive numerical criteria. BSTTs
in conjunction with  sediment quality criteria
derived from equilibrium  partitioning (USEPA,
1980; Swartz et al., 1990) can also be used in
assessments of potentially contaminated sediments
(see'   Chapter   6,  Equilibrium   Partitioning
Approach).
33 USEFULNESS
33.1  Environmental Applicability
                   »
3.5.1.1 Suitability for Different Sediment Types

    The sediment toxicity test approach is suitable
for any type of sediment.   In  some  cases, the
physical or chemical properties of the test sedi-
ment, such as salinity or grain size, may limit the
selection of organisms that can be used for testing
(Ott,  1986; DeWitt et al,  1989).  Appropriate
controls or statistical models (DeWitt et al., 1988)
for sediment  properties may be necessary  to
discriminate chemical toxicity from conventional
effects. In establishing sediment quality criteria,
the effects of features of the sediment itself, such
as grain size, must be recognized (DeWitt et al.,
1988). Data can be normalized to such factors as
organic carbon or acid volatile sulfide (DiToro et
al, 1990, 1991; Nebeker et dl, 1989) and thus
can be applied to any sediment.  However, nor"
malization techniques are in the developmental
stage (see Chapter 6, Equilibrium Partitioning
Approach).

3.3.1.2 Suitability for Different Chemicals or
       Classes of Chemicals

    BSTT is the only currently available approach
that directly measures the biological effects of all
classes  of chemicals,  including  the  combined
interactive (additive,  synergistic, antagonistic)
toxic effects  among  individual  chemicals  in
mixtures of contaminants usually found in  field
sediments (Plesha et al, 1988;  Swartz et al,
1989).  Bioaccumulative chemicals can be eval-
uated if the length of the test is extended to ensure
adequate exposure of the test organism.
3-4

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                                                                            3—BSTT Approach
3.3.1.3 Suitability for Predicting Effects on
       Different Organisms

    Theoretically, any organism can be  used in
sediment toxicity testing.  To protect a biological
community and to predict the effects of contami-
nated  sediments  on  different  organisms,  test
organisms should be selected on the basis of their
sensitivity to contaminants, their ability.to with-
stand laboratory handling, and their  ability to
survive  in  control  and  reference   treatments
(DeWitt et al., 1989; Reish and  LeMay, 1988;
Shubaefa/., 1981).   In  tests to determine the
effects of contaminated sediments on a particular
biological community, the test species  selected
should be among the most sensitive found in the
community of interest, or should  be comparably
sensitive.  Test species should include more than
one type  of organism to ensure a range of sensi-
tivity to various types of contaminants (Becker et
al, 1990).

3.3.1.4 Suitability for In'Place Pollutant Control

     Sediment toxicity testing can be used directly
to monitor in-place  pollution.  As discussed  in
 Section 3.2.1.1,  sediment toxicity testing can be
 used to determine the extent of the problem area,
 monitor temporal and spatial trends, detect the
 presence of unsuspected hot spots, assess the need
 for remedial  actions, and monitor  changes  in
 toxicity after remediation. Such tests can also be
 used as a cost-effective and rapid screening tool
 for in situ pollutant reconnaissance surveys and in
 a priori simulations of proposed remedial actions
 to test the effectiveness of capping or other reme-
 dial alternatives.

 3.3.1.5 Suitability for Source Control

     Bulk field sediment  toxicity testing can  be
 used ,to  identify suspected  sources of  sediment
 pollution. Field reconnaissance surveys can reveal
 hot spots near contaminant sources,  and a map
 showing contours of sediment toxicity values can
 reveal gradients that identify point and nonpoint
.sources (Swartz et al., 1982).   Toxicity testing
 cannot be used by itself to verify reductions in the
mass loading of chemicals that might be expected
as a result of source,control.  However, the bio-
logical effects of source control can be represented
through the use of BSTT,

33.1.6 Suitability for Disposal Applications

    BSTT has been  used widely in regulatory
programs to determine the toxicity. of material1
before disposal (Reish and LeMay, 1988; USEPA/
USAGE,  1991).  The potential hazard to benthic
organisms at the disposal site (which is  deter-
mined by making comparisons'with the "refer-
ence" sediments  collected near the disposal site)
can be predicted from laboratory  toxicity. test
results. Sediment toxicity tests also can be used
to  monitor  conditions  at the disposal site both
before and after a disposal operation.
 3.3.2 General Advantages and Limitations
 3.3.2.1 Ease of Use

     Most sediment toxicity test procedures  are
 simple to use, requiring limited  expertise and
 standard inexpensive laboratory equipment (PSEP,
 1991). Only a few sublethal effects tests require
 specialized training. Field sampling requires only
 readily   available  equipment  and   standard
 procedures (ASTM, 1990b).

 3.3.2.2 Relative Cost

     Individual laboratory toxicity tests and field
 sampling are cost-effective because they require
 limited  expertise  and  inexpensive  equipment.
 Such costs generally range from $150 to $500 per
 sampling replicate. Laboratory sediment toxicity
 testing is a comparatively inexpensive and cost-
 effective method of monitoring the field  distri-
 bution of sediment toxicity because it integrates
 the effects of all toxic contaminants,  does not
 require individual chemical measurements,  and
 does not require  time-consuming  analysis  of
 benthic community structure.
                                                                                               3-5

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 Sediment Classification Methods Compendium
3.3.2.3 Tendency to Be Conservative '

    Sediment toxicity tests can be made as sensi-
tive or  as  conservative (i.e.,  environmentally
protective)  as  necessary  through  selection  of
biological endpoints and species of test organism.
Reliance on mortality  as an endpoint may  be
underprotective, while some sublethal endpoints
(e.g., enzyme inhibition) may be overprotective.

3.3.2.4 Level of Acceptance

    BSTT is widely accepted by the scientific and
regulatory  communities and  has been tested and
contested in court.  Field sediment toxicity test
results  have been published  widely  in peer-
reviewed journals  and incorporated into other
measures of sediment quality such as the AET
and the Triad approaches.   Standard guides for
sediment toxicity testing continue to be developed
by ASTM (1990a, 1990b, 1991), and field sedi-
ment toxicity testing is incorporated into most
dredged  material  disposal regulatory programs
(PSEP, 1991; Reish and LeMay, 1988; USEPA/
USAGE, 1991).  Toxicity testing in general has
long been the basis for water quality  criteria,
dredged  material  testing, effluent testing, and
discharge monitoring.                .    •

33.2.5 Ability to Be Implemented by
       Laboratories with Typical Equipment
        and Handling Facilities

    Sediment toxicity  test methods  are  easily
implemented by laboratories with typical equip-
ment using inexpensive glassware and procedures
requiring little specialized training, although the
interpretation of some  sublethal biological end-
points  may require some degree of training and
experience.  Field sediment sample  collection
procedures are routine.

3.3.2.6 Level of Effort Required to Generate
       Results

    This procedure consists of field sampling and
a laboratory toxicity test.  Compared to an exten-
sive survey of chemical concentrations or benthic
community structure analysis, the level of effort is
relatively small.

3.3.2.7 Degree to Which Results Lend
       Themselves to Interpretation

    Biological responses to toxic sediment can be
easily  interpreted.  Generally, data fit "pass-fail1*
criteria (i.e., the result is either above, or below a
predetermined acceptance level) or the result is
compared statistically  to control  and reference
results to determine whether there is a toxic effect
Little expert guidance is required for interpretation
of mortality  data although chronic or sublethal
effects might require some explanation.

3.3.2.8 Degree of Environmental Applicability

    As noted in Section 3.3.1.1, the sediment
toxicity test approach applies to a wide range of
environmental conditions and sediment types. The
effects of various  sediment  properties such  as
grain size and organic content can be addressed
experimentally  with appropriate uncontaminated
controls.                            .

3.3.2.9 Degree of Accuracy and Precision

    Because the sediment toxicity test is a labora-
tory-controlled experiment, its results have a high
degree of accuracy, precision, and repeatability.
3.4 STATUS

3.4.1  Extent of Use

    Sediment toxicity  tests are widely used  in
research and regulatory programs in both marine
and freshwater systems (ASTM, 1990a, 1991), as
described  in Section 3.2.1.1.  Sediment toxicity
tests also  are incorporated into the evaluation of
applications for dredged material disposal permits
and.are used to assess the toxicity  of sediments
subject to regulatory decisions. BSTTs are used
to investigate the mechanisms of sediment toxicity
to benthic organisms (Kemp and Swaitz, 1989;
Swartz et  al, 1988).
3-6

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                                                                          3—BSTT Approach
3.4.2 Extent to Which Approach Has Been
     Field-Validated

    Field validation  of BSTT  includes several
publications in peer-reviewed literature (Chapman,
1986b; Plesha et al, 1988; Swartz et al,  1982,
1986, 1989).  As  more data become available,
results can be compared with available informa-
tion on contaminant concentrations in sediment in
areas where biological effects have been observed.
The effects of interactions among contaminants, as
well as  the effects of  nonchemical  sediment
variables, must be  taken into consideration when
attempts are made at field validation (DeWitt et
al,  1988;  Swartz  et al.,  1989).  As noted in
Section 3.2.1.3, better field validation of predicted
effects is needed.
 3.43  Reasons for Limited Use

    BSTT has been widely used in research and
 regulatory programs (see Section 3.4.1, Extent of
 Use).     -.,.••-.      ,      '
 3.4.4  Outlook for Future Use and Amount of
       Development Yet Needed

    The outlook for future use of sediment toxi-
 city tests is promising where direct measurement
 of biological effects of toxicants in sediments is
 desired, especially where the effects of chemical
 interactions' are of interest.   Development and
 standardization  of  biological testing  methods
 should continue, especially for tests using species
 locally available in geographic areas that have not
 been represented such as tropical and arctic re-
 gions.  More emphasis should be placed on the
 development of procedures  to measure chronic
 effects. Methods should be compared and stand-
 ardized among laboratories, and results should be
 field-validated to establish their ability to predict
 biological effects on populations and communities
 in the field. As more toxicity tests are conducted
 and  the  results subject to  a quality assurance
 review, results should be  compiled in  a central
 database so that comparisons can be made among
 species, methods, and laboratories.
3.5 REFERENCES

Adams, W.J., R.A. Kimerle,  and R.G. Mosher.
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Ankley, G.T., A. Katko, and  J.W. Arthur. 1990.
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ASTM. 1990a. E 1367-90. Guide for conducting
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    PA.           .           .         '.'  .' ,
ASTM. 1990b. E 1391-90. Guide for collection,
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ASTM.  1990c. E 1383-90, Guide for conducting
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 Becker, D.S., G.R. Bilyard, and T.C Ginn.  1990.
    Comparisons between sediment bioassays and
    alterations of benthic macroinvertebrate assem-
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    ment Bay,  Washington.  Environ.  Toxicol.
    Chem. 9: 669-685.
 Chapman, P.M. 1986a. Sediment bioassay tests
    provide data necessary  for  assessment and
    regulation. In:  Proceedings of the Eleventh
   . Annual Aquatic Toxicology Workshop;  Tech-
    nical Report 1480.  Green,  G.H. and K.L,
    Woodward (eds.).  Fish. Aquat. ScL, pp. 178-
    197.
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 Sediment Classification Methods Compendium
Chapman, P.M.  1986b.  Sediment quality criteria
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   Environ. Toxicol. Chem. 5: 957-964.
Chapman, P.M.  1988.  Marine sediment toxicity
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   American  Society for Testing  and Materials,
   Philadelphia, PA. pp. 391-402.
Chapman, P.M., G.A. Vigers, MA. Farrell, R.N.
   Dexter, E.A. Quinlan,  R.M. Kocan, and M.
   Landolt. 1982.  Survey of biological effects of
   toxicants upon Puget Sound biota.  1. Broad-
   scale toxicity survey. NOAA Technical Memo-
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Chapman, P.M., R.N. Dexter, and E.  R. Long.
   1987. Synoptic measures of sediment contami-
   nation, toxicity and infaunal community com-
   position (the sediment quality  triad) in San
   Francisco  Bay.  Mar.  Ecol. Prog.  Ser. 37:
   75-96.
DeWitt, T.H., G.R. Ditsworth, and R.C. Swartz.
   1988. Effects of natural sediment features on
   the phoxocephalid  amphipod,  Rhepoxynius
   abronius:  Implications  for sediment toxicity
   bioassays.  Mar. Environ. Res. 25: 99-124.
DeWitt, T.H., R.C.  Swartz, and J.O. Lamberson.
   1989. Measuring the toxicity of estuarine sedi-
   ments. Environ. Toxicol. Chem. 8:1035-1048.
DiToro, D.M., J.D. Mahony, DJ. Hansen,  KJ.
   Scott,  M.B.  Hicks,  S.M.  Mayr,  and  M.S.
   Redmond. 1990. Toxicity  of cadmium in sedi-
   ments:  the role of acid volatile  sulfide.
   Environ. Toxicol. Chem. 9: 1487-1502.
DiToro, D.M., J.D. Mahony, DJ. Hansen,  KJ.
   Scott, A.R. Carlson,  and  G.T.  Ankley. 1991.
   Acid volatile sulfide predicts the acute toxicity
   of cadmium and nickel in  sediments. Environ-
   mental Science arid Technology.           '
Giesy, J.P.,  R.L.  Graney,  J.L. Newsted,  CJ.
   Rosiu, A. Benda, R.G. Kreis, and FJ. Horvath.
   1988. Comparison of three sediment bioassay
   methods   using   Detroit   River  sediments.
   Environ. Toxicol. Chem. 7: 483-498.
Ingersoll,  C.G., and M.K. Nelson.  1989. Solid-
   phase sediment toxicity testing with the fresh-
   water invertebrates: Hyalella azteca (Amphi-
   poda) and Chironomus riparius (Diptera). In:,.'
   Aquatic Toxicology Risk Assessment: Proceed-
   ings of the Thirteenth Annual Symposium,
   ASTM STP; American Society for Testing and
   Materials, Philadelphia, PA.        '
Kemp, P.F., and R.C. Swartz. 1989. Acute toxicity
   of interstitial and particle-bound cadmium to a
   marine infaunal  amphipod. Marine  Environ.
   Res. 26: 135-153.
Lamberson, J.O., and R.C. Swartz. 1988. Use of
   bioassays in determining the toxicity of sedi-
   ment to benthic organisms, Chapter 13, In:
   Toxic Contaminants and Ecosystem Health: A
   Great Lakes Focus.  Evans, M.S. (ed.).  John
   Wiley and Sons, New York, NY. pp.  257-279.
Melzian,  B.D. 1990.  Toxicity  assessment of
   dredged materials:  acute and chronic toxicity
   as determined by bioassays and  bioaccum-
   ulation tests. In:  Proceedings of the Inter-
   national    Seminar  on the  Environmental
   Aspects of  Dredging Activities,   Goubault
   Impremeur, Nantes; France, pp. 49-64.
Nebeker, A.V., MA. Cairns, J.H. Gakstatter, K.W.
   Maleug, G.S. Schuytema, and D.F. Krawczyk.
   1984. Biological methods for determining toxicity
   of contaminated freshwater sediments to inverte-
   brates. Environ. Toxicol. Chem. 3: 617-630.
Nebeker, A.V. and C.E. Miller. 1988. Use of the
   amphipod crustacean Hyalella azteca in fresh-
   water and  estuarine  sediment  toxicity tests.
   Environ. Toxicol. Chem. 7: 1027-1034.
Nebeker, A.V., G.S. Schuytema, W.L. Griffis, J A.
   Barbitta,  and  LA.  Carey. 1989. Effect of
   sediment   organic  carbon  on  survival  of
   Hyalella azteca exposed-to DDT and endrin.
   Environ. Toxicol. and Chem. 8: 705-718.
Nipper, M.G.,  DJ.  Greenstein, and S.M. Bay.
   1989. Short- and long-term sediment toxicity
   test methods with the amphipod Grandidierdla
   japonica. Environ. Toxicol.  and Chem. 8:1191-
   1200.
Ott, F.S.  1986. Amphipod sediment bioassays:
   Effect of grain size, cadmium, methodology,
   and variations in animal sensitivity on interpre-
   tation of experimental data.  Ph.D. dissertation.
   University of Washington, Seattle, WA.
Plesha,  P.O., J.E. Stein,  M.H. Schiewe, B.B.
   McCain, and U. Varanasi.  1988. Toxicity of
3-8

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                                                                         3—BSTT Afrproach
   marine sediments supplemented with mix-
   tures of selected chlorinated and  aromatic
   hydrocarbons  to the  infaunal  amphipod,
   Rhepoxynius abronius. Mar:  Environ.  Res.
   25:85-97.
Puget Sound Estuary Program. 1991. Recommend-
   ed guidelines for conducting  laboratory bio-
   assays on Puget Sound sediments. Draft report
   prepared for U.S. Environmental  Protection
   Agency, Region 10, Office of Puget Sound,
   Seattle, WA.
PTI  Environmental Services.   1988.   Sediment
   quality values refinement: Tasks .3 and 5 -
   1988 update and evaluation of the Puget Sound
   AET. PTI Environmental Services, Bellevue,
   WA.
Reish, D.J.,  and J.A. Lemay.  1988. Bioassay
   manual  for  dredged  materials.    Contract
   DACW-09-83R-005. U.S.  Army  Corps of
   Engineers, Los Angeles District, Los Angeles,
   CA.        .••      ,   -        '     .   .-
Scott, KJ., and M.S. Redmond. 1989. The effects
   of a contaminated dredged material on labora-
   tory populations of the tubicolous amphipod,
   Ampelisca abdita. In: Aquatic Toxicology and
   Hazard Assessment:  Vol 12. U. M. Cowgill
   and L. R. Williams (eds.).   ASTM STP 1027.
   American  Society for Testing and Materials,
   Philadelphia, PA.
 Shuba, PJ., S.R. Petrocelli, and R.E.  Benfley.
    1981.  Considerations  in  selecting bioassay
    organisms for determining the potential en-
    vironmental impact of dredged material.  Tech-
    nical Report  EL-81-8. U.S.  Army Engineer
    Waterways Experimental  Station, Vicksburg,
    MS.
 Swartz, R. C. 1987. Toxicological methods for
    determining the effects of contaminated sedi-
    ment on  marine organisms, pp. 183-198. In:
    Fate and Effects of Sediment Bound Chemicals
    in Aquatic Systems.   K. L. Dickson, A.W.
    Maki,  and W. A. Brungs  (eds.). Pergamon,
   f Press, New York.
 Swartz, R.C., WA. DeBen, KA. Sercu, and J.O.
    Lamberson.  1982., Sediment toxicity and the
    distribution of amphipods in Commencement
    Bay,  Washington, USA.  Mar. Poll. Bull. 13:
    359-364.             •'"'.•,.-•'
Swartz,  R.C., WA. DeBen, J.K.P. Jones, J.O.
   Lamberson, and FA. Cole. 1985a.   Phoxo-
   cephalid amphipod bioassay for marine sedi-
   ment toxicity.  In:  Aquatic Toxicology  and
   Hazard Assessment. R:D. Cardwell, R, Purdy
   and R.C. Banner (eds,).   ASTM STP 854,
   pp. 284-307.  American Society for Testing
   and Materials, Philadelphia, PA.
Swartz, R.C., D.W. Schults, G.R. Ditsworth, W A.
   DeBen, and JFA.  Cole.  1985b.   Sediment
   toxicity,  contamination, and  macrobenthic
   communities  near a  large  sewage  outfall.
   pp. 152-175.   In:   Validation and Predict-
   ability of Laboratory Methods for Assessing
   the Fate arid Effects of Contaminants in Aquat-
   ic Ecosystems.  T.P. Boyle (ed.). ASTM  STP
   865. American Society for Testing and Mater-
   ials, Philadelphia, PA.
 Swartz. R.C., FA. Cole, D.W. Schults, and WA.
   DeBen. 1986. Ecological changes on the Palos
   Verdes Shelf  near a large sewage outfall:
   1980-1983. Mar. Ecol. Prog. Ser. 31: 1-13.
 Swartz, R.C., P.F. Kemp, D.W. Schults, and J.O.
   Lamberson. 1988. Effects of mixtures of  sedi-
   ment contaminants on  the marine infaunal
   amphipod, Rhepoxynius abronius.   Environ.
   Toxicol. Chem. 7: 1013-1020.
 Swartz,  R.C., P.F. Kemp, D.W. Schults,  G.R.
   Ditsworth, and R.O.  Ozretich. 1989. Acute
   toxicity of sediment from Eagle Harbor, Wash-
   ington, to the infaunal amphipod Rhepoxynius
   abronius. Environ. Toxicol. and Chem. 8: 215-
   222.'
 Swartz, R.C., D.W. Schults, T.H. DeWitt,  G.R.
    Ditsworth, and J.O. Lamberson. 1990. Toxicity
    of fluoranthene in sediment to marine amphi-
    pods: a test of the equilibrium partitioning
    approach to sediment quality criteria. Environ.
    Toxicol. and Chem. 9: 1071-1080.
 Swartz, R.C., D.W- Schults, J.O. Lamberson, RJ.
    Ozretich, and J.K. Stall." 1991. A lexicological
    record in cores  of  contaminated sediment
    Mar. Environ. Res.
 Tetra Tech, Inc.  1986. Eagle Harbor preliminary
    investigation.   Final  Report EGHB-2, TC-
    3025-003. Tetra Tech, Inc., Bellevue, WA.
 Thompson, B.E..S.M. Bay, J.W. Anderson, J.D.
    Laughlin, D.J. Greenstein, and D.T. Tsukada.
                                                                                           3-9

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 Sediment Classification Methods Compendium
    1989.  Chronic effects of contami-nated sedi-
    ments on the urchin Lytichinuspictus. Environ.
    Toxicol. and Chem. 8: 629-637.
 USEPA.  1980. Water quality criteria for fluor-
    anthene.    U.S.  Environmental  Protection
    Agency, Washington,  DC.
USEPA/USACE. 1991. Evaluation of dredged
   material proposed for  ocean disposal-testing
   manual. EPA-503-8-91/001.  U.S. Environ-
   mental Protection Agency and U.S.  Army
   Corps of Engineers, Washington, DC.
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          CHARTER 4
Spiked-Sediment  Toxicity Test Approach

Janet O. Lamberson and Richard C. Swartz
U.S. Environmental Protection Agency, Pacific Ecosystems Branch, ERL-N
2111 Southeast Marine Science Dr., Newport, OR 97365-5260
(503) 867-4031
    The toxicological approach to generating sedi-
ment quality criteria uses concentration-response
data from sediments spiked in the laboratory with
known concentrations of contaminants. Sediments
are spiked to establish cause-and-effect relation-
ships between chemicals and adverse biological
responses (e.g., mortality, reduction in growth or
reproduction, physiological changes). Individual
chemicals or other potentially toxic substances can
be tested alone or in combination to determine
toxic concentrations of contaminants in sediment.
This approach can be used to generate  sediment
quality  criteria or to validate  sediment quality
criteria generated by other approaches.
 4.1  SPECIFIC APPLICATIONS

 4.1.1 Current Use

     The spiked-sediment  toxicity  test (SSTT)
 approach is in the research stage.  Although the
 procedures used resemble those used to generate
 water quality criteria, the influence of the variable
 properties of sediment makes generating quality
 criteria values much more complex.
     Where LCjo values and chronic effects data
 are available for chemicals  in sediments (see
 Section 4.3.2.3), they can be used to identify
 concentrations of chemicals in sediment that are
 protective of aquatic life. The predictive value of
 sediment  quality  criteria generated  by  this
 approach should be tested by comparing them
 with field  data  on chemical  concentrations in
 natural sediments and observed biological effects.
 However, interim laboratory-derived criteria can
 be implemented before field validation.
4.1.2 Potential Use

   This method can be used to address empirical-
ly the problem of interactions  among complex
mixtures of contaminants that are almost always
present in the field (Swartz et al., 1988, 1989).
Chemical-specific data can be  generated for a
wide variety of classes of .chemical contaminants,
including  metals,  PAHs,  PCBs,  dioxins,  and
chlorinated pesticides.  Both acute and chronic
criteria can be established, and the approach is
applicable to both marine and freshwater systems
(Tetra Tech,   1986; Battelle, 1988),  However,
unless  the sediment  factor that normalizes  for
bioavailability  is known, this procedure must be
applied to every sediment (i.e., a value derived for
one sediment may not be applied with predictable
results  to  another  sediment  with  different
properties).        •
 4.2 DESCRIPTION
 4.2.1 Description of Method

    The toxicological approach involves expos-
 ing test organisms to sediments that have been
 spiked with known quantities of potentially toxic
 chemicals or mixtures of compounds.  At the
 end of a  specified time peripdj the response of
 the test organism is examined in relation to a
 biological endpoint  (e.g.,  mortality, growth,
 reproduction, cytotoxicity, alterations in devel-
 opment or respiration rate).   Results are then
 statistically compared with results from control
 or reference sediments to identify toxic concen-
 trations of the test chemical.

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 Sediment Classification Methods Compendium
4.2.1.1  Objectives and Assumptions  •

    The objective of this approach is to derive in
the laboratory concentration-response values that
can  be used to predict the concentrations  of
specific chemicals harmful to resident biota under
field conditions.    The  effects  of the  inter-
actions—synergism,   addijivity,  antagonism—
among  chemicals in the  field can be  predicted
from laboratory results with sediments spiked with
combinations of chemicals.  This method can be
used for all classes of sediments and any chemical
contaminant.   The bioavailable component  of
contaminants in sediment can be determined by
this method, and an a priori knowledge  of specif-
ic pathways of interaction between sediments and
test organisms is not necessary. Any method of
expressing the bioavailability of contaminants in
sediment can be used with sediment toxicity tests,
including the "free" interstitial concentration and
normalization  to organic  carbon,  acid volatile
sulfide, and other sediment properties.
    Data generated by this method may be diffi-
cult to interpret if the normalizing factor for
bioavailability is  unknown.  If the normalization
factor is  known, this  method  can be  used  to
validate sediment  quality  criteria  generated by
other approaches.  It is assumed that laboratory
results for a  given  sediment and overlying water
represent biological effects of similar sediments in
the field, and that  the behavior of chemicals in
spiked sediments is similar to that in natural, in
situ sediments.


4.2.1.2  Level of Effort

    Implementation of this procedure requires  a
moderate  to  considerable amount  of laboratory
effort.  The various toxicity  test procedures that
have been developed are generally   straightfor-
ward and well  documented  (Lamberson and
Swartz, 1988; Melzian,  1990;  Nebeker  et al,
1984; Swartz et al., 1989; PSEP,  1991).  How-
ever, many individual tests would be required to
generate an extensive database of sediment quality
values for a large number of chemicals,  chemical
combinations, and sediment types.
 4.2.1.2.1 Type of Sampling Required

     Collection  of sediments from the field  is
 required.   Depending on the particular study
 objectives, the sediments may be clean (uncontam-
 inated) sediments from a control area, uncontami-
 nated reference sediments for  comparison with
 similarly contaminated sediments, or contaminated
 sediments to be spiked with known concentrations
 of chemicals in  a test for interactions among
 contaminants.   Sufficient sediment must be col-
 lected to provide samples for chemical analysis,
 spiking, and reference or controls (i.e., sediment
 for statistical comparison with spiked sediment).
 Depending on the experimental design, the follow-
 ing controls might be required: sediment from the
 collection site for test animals (or culture sediment
 for laboratory-cultured animals), positive controls
 with a reference toxicant, carrier controls, and
 reference sediment controls for natural sediment
 features that may affect test animals, such as grain
 size distribution (DeWitt et al.,  1988).

 4.2.1.2.2 Methods

    Various methods of  adding  chemicals  to
 sediment (spiking sediments) have been used.  In
 general, the chemical  is  either added to the sedi-
 ment and mixed in (Birge  et al., 1987; Ditsworth
 et al, 1990; Francis et al.,  1984) or added to the
 overlying water (Hansen and Tagatz, 1980; Kemp
 and Swartz, 1988) or to a sediment slurry (Lan-
 drum, 1989; Oliver, 1984; Schuytema etal., 1984)
 and allowed  to equilibrate with  the  sediment.
 Sediments are spiked  with  a range of concentra-
 tions to generate LCjo  data or to determine a
 minimum concentration at which biological effects
 are observed.                             .
    The  effect  of sediment  contaminants  on
 benthic biota is determined either by exposing
 known numbers of individual benthic test organ-
 isms to the sediment for a  specific length of time
' (Swartz et al.,  1985) or by exposing  larvae of
 benthic species to the sediment in flowing natural
 waters  (Hansen and Tagatz, 1980).   Biological
 responses  are determined  at the end of the test
 period using response criteria that include mortali-
 ty, changes in growth  or reproduction, behavioral
4-2

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                                                                           4—SSTT Approach
or physiological  alterations, or  differences in.-,
numbers and species of larvae in contaminated
versus control sediments.

4.2.1.2,3  Types of Data Required

    Spiked  sediments,  as well  as reference or
control sediments, must be analyzed for total
solids, grain size, and total and dissolved organ-
ic  carbon.   The concentrations of toxicants
added to sediment must be determined in stock
solutions as well as in the test  sediment.  Bulk
and interstitial levels of the spiked chemicals in
the test sediment must be determined throughout
a concentration range  at least at the beginning
and at the  end of the toxicity  test. However,
methods for sampling interstitial water have not
been  standardized.  If sediment properties that
 control availability, such as acid volatile sulfides
 or dissolved or total, organic  carbon, change
 during exposure, measurements must  be taken
 before,  during and at the  end  of the  exposure
 period.  In  addition, these changes must be taken
 into account in interpreting the data. Sediment
 parameters such as  pH and Eh should also be
 monitored.
     Biological and chemical data are compared
 statistically with control or reference data to
 determine  the occurrence of biological effects,
 and can be used to calculate LC50 values, EC50
 values,  no-effect  concentrations,  or  lowest-
 observable-effect concentrations. Establishment
 of the maximum acceptable toxicant concentra-
 tion  requires  data  from a chronic or  life-cycle
  test.
     Data correlating observed  biological effects
  with chemical concentrations in spiked sediment
  can be used to calculate probit  curves for deriva-
  tion  of biological effect level values (e.g.,,ECso).
  Data from several  species of test organisms can
  be ranked, and the lowest contaminant concen-
  trations that affect the most sensitive species can
  be used to establish sediment quality criteria that
  will protect the entire benthic community and
  associated aquatic ecosystem.  This approach has
  regulatory and  scientific precedence  in the
  development of water quality  criteria.
4.2.1.2.4  Necessary Hardware and Skills

    Most toxicity test procedures require a mini-
mum of specialized hardware and level of skill.
In general, only readily available and inexpensive
laboratory equipment is  needed, procedures  are
fairly simple and straightforward, and a minimum
of training  is necessary to detect and interpret
biological endpoints. Although analysis of chemi-
cal  samples requires specialized training and
equipment,  the  chemical sampling methods  for
spiked-sediment toxicity  are generally simple and
routine. Some bioldgical effects tests also require
specialized  training and experience, especially to
interpret the results.


4.2.1.3 Adequacy of Documentation

     Various acute sediment toxicity test proce-
 dures have been  developed and are well docu-
 mented for testing  freshwater and  marine field
 sediments (Chapman, 1986,1988; Lambersoh and
 Swartz,  1988;  Melzian,  1989;  Swartz,   1987).
 Although only  a few of these procedures have
 been used with  laboratory-spiked sediments, most
 of the established methods could be used with
 laboratory-prepared sediments as well as with field
 sediments.         •
     In contrast to acute tests, there  are relatively
 few procedures for testing the chronic effects of
 contaminated sediments on benthic  invertebrates.
 Life-cycle  test methodology has  been presented
 for the amphipods Ampelisca abdita (Scott and
 Redmond,  1989), Hyalella azteca (ASTM, 1990c;
 Borgmann and  Munawar,  1989),  and  Grand-
  idierella lutosa  and G. lignorum  (Connell and
 Airey, 1982); the polychaetes Neanthes arenaceo-
  dentata  (Pesch,  1979) and Capitella capitata
  (Chapman and  Fink,,  1984); freshwater oligo-
  chaetes (Wiederholm et al., 1987); and species of
  Daphnia and Chironomus (ASTM, 1991; Nebeker
  et al, 1988). Chronic exposures to most sensitive
  life stages are also inherent in the  benthic recol-
  onization  procedure (Hansen and Tagatz, 1980).
  Further research is needed to develop and validate
  methodology for other species.
                                                                                              4-3

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  Sediment Classification Methods Compendium
 4.22 Applicability of Method to Human
       Health, Aquatic Life, or Wildlife
       Protection

     Spiked-sediment toxicity tests  incorporate a
 direct measure of sediment biological effects. This
 approach is the only method that can quantify the
 interactive effects of combinations of contaminants
 directly.
     When chemical concentrations in tested biota
 are measured after a spiked-sediment toxicity test,
 uptake of contaminants by benthic organisms (bio-
 accumulation) can be predicted.  As an important
 component of food webs in aquatic ecosystems,
 benthic organisms can contribute toxicants accumu-
 lated from contaminated sediments to higher levels
 of the aquatic food web and ultimately affect human
 health. Sediment quality criteria and bioaccumula-
 lion studies using sediment toxicity test methods can
 help to set limits on the disposal of toxic sediments
 and predict uptake of toxicants into food webs. If
 this approach is combined with chemical analysis of
 sediment samples and BSTT, these  limits  can be
 used to define areas from which food species should
 not be harvested or consumed  or  where direct
 contact with contaminated sediments can be hazard-
 ous to human health.
     Bioaccumulation studies and sediment  quality
 criteria established using data from SSTT  with
 several benthic species can also be used to  protect
 benthic communities and aquatic species that feed
 on the benthos.  Assuming that a sufficient  mix of
 taxonomic groups is used, a sediment quality criteri-
 on based on  the responses of the most sensitive
 species within a benthic community can be devel-
 oped.   This criterion can then be  employed to
 protect the structure and function of  the entire
 ecosystem (Hansen and Tagatz, 1980).

 4.23 Ability of Method to Generate Numerical
      Criteria for Specific Chemicals

    Laboratory tests with the SSTT approach can be
 used to  measure the effects of specific chemicals in
 various types of sediments directly and to establish
 unequivocal analysis of causal effects. Test condi-
 tions allow this method to determine the effects of
 individual chemicals or  mixtures of chemicals on
 benthic biota (Plesha  et al., 1988; Swartz et a/.,.
 1988,  1989), establish pathways of toxicity,  and
 provide specific effects concentrations (e.g. LQo,
 ECjo, no-effect concentration).  The influence of
 various physical characteristics of the sediment on
 chemical toxicity also can be determined (DeWitt et
 al, 1988; Ott 1986). The available data represent
 concentrations  at which toxicity occurs rather than
 numerical sediment quality criteria.  Recent spiked
 sediment studies have  provided data that can be
 useful in setting preliminary sediment criteria levels
 based on equilibrium partitioning models and water
 quality values (Swartz et al., 1990).
     Concentration-response data have been gener-
 ated using SSTT for a variety of chemicals, includ-
 ing metals and organic compounds.  Specific data
 are available for phenanthrene, fluoranthene, zinc,
 mercury, copper, cadmium,  hexachlorobenzene,
 pentachlorophenol, Arodor 1242 and 1254, chlor-
 dane,  DDE, DDT,  dieldrin,  endosulfan, endrin,
 sevin, creosote, and kepone (Adams et al., 1985;
 Cairns et al.r 1984; DeWitt et al., 1989; Kemp and
 Swartz, 1989; McLeese and Metcalfe,  1980;  Mc-
 Leese et al., 1982; Nebeker et al, 1989; Swartz ~et
 al., 1986, 1988,  1989;  Tagatz et al, 1977, 1979,
 1983; Word et al, 1987).  Concentrations of non-
 ionic organic compounds are usually normalized to
 sediment organic carbon  or  acid volatile sulfide
 (DiToro et al,  1990, 1991; Nebeker et al, 1989):
 Normalizing factors for  other  compounds   in
 sediment currently are being researched.
4.3 USEFULNESS
4J.I  Environmental Applicability

43.1.1  Suitability far Different Sediment Types

    The SSTT approach is suitable for any type of
sediment  This  approach also can be  used to
establish the bioavailable component of the sedi-
ment responsible for the observed toxicity. The
effects of various physical properties of the sedi-
ment  on chemical toxicity  can be  determined
experimentally.  In some cases, the physical or
chemical properties of the test sediment such as
salinity or  grain  size may  limit the species of
organisms  that can be used  for testing, and a
4-4

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                                                                              4—SSTT Approach
substitute species must be used (DeWitt  et al.,
1988,1989). When establishing sediment quality
criteria, the effects of adverse physical or chemical
properties of the sediment itself must be reflected.
When factors controlling bioavailability  (e.g.,
organic carbon, acid volatile sulfide) are known,
data can be normalized to such factors, and the
approach applied to any sediment type.

4.3.1.2 Suitability for Different Chemicals or
        Classes of Chemicals

    A major  advantage of the SSTT method is
that it is suitable for all classes of chemicals.  In
addition, it is the only approach currently avail-
able that can empirically determine the interactive
effects among individual chemicals in mixtures of
contaminants  usually found  in real-world sedi-
ments (Swartz etal, 1988, 1989). This approach
 also can be used to provide experimental valida-
 tion of sediment quality criteria generated by other
 approaches.

 4.3.1.3  Suitability for Predicting Effects on
         Different Organisms

    , Theoretically, any organism can be used  in
 SSTT.  To protect a biological community and to
 predict the effects of a toxicant on different organ-
 isms, test organisms should be selected based on the
 following criteria: (1) then- sensitivity to contami-
 nants,  (2) their ability  to withstand  laboratory
 handling, and (3) their ability to survive in control
 treatments.   Tests  to  determine the effects  of
 toxicants  on a  particular biological  community
 should use the most sensitive species found in the
 community or, a species with comparable sensitivity.

 4.3.1.4 Suitability for In-Place Pollutant Control

      SSTT can be used to develop sediment quality
  criteria, which will then be used to determine the
  extent of the problem area.  It also can be used to
  monitor temporal and spatial trends and to assess the
  need for  remedial action.  Criteria can be used in
  setting target cleanup levels and in post-cleanup
  monitoring of actual contaminant levels.
4.3.1.5 Suitability for Source Control

    SSTT can be combined with wasteload alloca-
tion models and used in source control to establish
maximum allowable effluent concentrations or mass
loadings of single chemicals and mixtures of chemi-
cals.

4.3.1$  Suitability for Disposal Applications

    SSTT can be used to  predict the biological
effects of contaminants before approval of dredged
material disposal or sewage outfall permits.
            "  '         '- '   J
432 General Advantages and Limitations

4.3.2.1  Ease of Use

    Most  sediment toxicity test procedures are
 simple  to use, require  limited expertise, and use
 standard laboratory equipment.  Some of the sub-
 lethal-effects tests require specialized training.

 4.32.2 Relative Cost
          - •        ' '       ' '             \     -  .
     The cost of individual toxicity tests is relatively
 low because such tests require limited expertise and
 inexpensive equipment. (See Chapter 3, Bulk Sedi-
 ment Toxicity Approach.)  The costs to implement
 this approach as a regulatory tool would be compar-
 atively high because SSTT requires the collection of
 sediment  chemistry data for comparison to data
 established by the sediment toxicity test method.
 The cost of developing a large toxicological data-
 base would be relatively high because of the large
 number pf individual chemicals and sediments that
 would have, to be tested. Generating the chemical
 and toxicological  data necessary to  establish a
 sediment quality criterion for one chemical by this
 method is estimated to cost $100,000.

 43.23 Tendency to Be Conservative

      Laboratory-controlled SSTT experiments pro--
 vide a high  degree of accuracy.  The tests ara
  controlled sufficiently  to  give  an estimate  of the
  toxicity  of individual chemicals in sediment Lab-
  oratory bioassays, especially acute toxicity tests, are
                                                                                                 4-5

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 Sediment Classification Methods Compendium
 inherently limited in their ability to reflect all of the
 ecological processes through which sediment con-
 taminants may affect benthic ecosystems in the field.

 4.3.2.4  Level of Acceptance

     SSTT methods, which follow the procedures
 and rationale used to develop water quality cri-
 teria, are easily interpreted/technically acceptable,
 and legally  defensible.    The  procedures and
 resulting data have been accepted and published in
 peer-reviewed journal  articles, and some proce-
 dures have been incorporated into standard guide-
 lines  by  ASTM's  subcommittee on  sediment
 toxicology (ASTM, 1990a,  1990c).

 4.3.2.5 Ability to Be Implemented by
        Laboratories with typical Equipment
        and Handling Facilities

     SSTT methods are implemented easily  by
 laboratories  with typical  equipment,  requiring
 inexpensive glassware and little specialized train-
 ing.   Spiking  sediments  may  require  special
 handling facilities for preparing stock solutions of
 highly toxic substances, arid the interpretation of
 some sublethal biological endpoints may require
 some degree of training and experience.

 4.3.2.6 Level of Effort Required to Generate
        Results

    This procedure consists of a  laboratory toxi-
 city test and requires a moderate amount of effort
 to begin and end an  experiment.  The data gener-
 ated must be compiled, and some calculations
 must be made to derive concentration-response
 relationships.  The  generation of chemical and
 biological data required for a large database of
 sediment quality values based on this approach
 would require a relatively high level of effort.

 4.3.2.7 Degree to Which Results Lend
        Themselves to Interpretation

    Sediment toxicity tests applied to spiked sedi-
 ments provide an unequivocal analysis of cause-
 and-effect relationships between toxic chemicals
 and biological responses. Because the procedures •
 follow the rationale used in the development of
 water quality criteria, the methods are legally
 defensible. Toxicity tests have long been accepted
 by both the public and the scientific community as
 a basis for water quality  criteria and dredged
 material testing.

 4.3.2.8 Degree of Environmental Applicability

    The SSTT approach is applicable to. a wide
 range of environmental conditions and sediment
 types.- The confounding effects of sediment vari-
 ables such as grain size and organic content can
 be addressed experimentally by using toxicity test
 methods  or can  be addressed by using normal-
 ization equations (DeWitt et al. 1988).  A major
 advantage jaf SSTT is the ability to predict inter-
 active effects of chemical mixtures such as those
 found in field sediments.                  ,  .

 4,3.2.9 Degree of Accuracy and Precision

    Because the SSTT is a laboratory-controlled
 experiment, results have a high degree of accuracy
 and precision.  The procedure produces a direct
 dose-response data set for individual chemicals in
 sediment.   Sediment  criteria generated by this
 approach must be field-validated.        "   .
4.4  STATUS

4.4.1 Extent of Use

    SSTT procedures are under development in
several laboratories.  Spiking procedures, as well
as biological test procedures, are currently being
standardized by ASTM's sediment  toxicology
subcommittee (ASTM, 1990b).

4.4.2 Extent to Which Approach Has Been
      Field-Validated

    Although some results have been published,
spiked-sediment toxicity test values have not been
well .validated in the field, (Plesha et al., 1988;
Swartz et al., 1989).  As more data and criteria
4-6

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                                                                          4-^SSTT Approach
values become available, they can be compared
with existing information on contaminant levels in
sediment in areais where biological effects have
been observed. The effects of interactions among
contaminants, as well as the effects of nonchemi-
cal sediment variables, must be considered during
Held validation (DeWitt et al., 1988; Swartz et at.,
1989).

4.4.3 Reasons for Limited Use
                              f

    Although some data have been generated and
compared to field conditions,  the approach is still
in the developmental stage in several laboratories,
and a relatively large expenditure of effort will be
needed to generate  a large database.  To date,
there have been  few comparisons of methods and
species sensitivity, and few chronic toxicity tests
have been developed.

 4.4.4  Outlook for Future Use and Amount of
       Development Yet Needed

     The outlook for future use of SSTTs or other
 sediment  toxicity   tests  is  promising  where
 accurate, direct  dose-response data are desired, or
 where the effects of chemical interactions need to
 be examined.  Development of sediment-spiking
 and biological-testing .methods should  continue,
 methods should be compared and  standardized
 among laboratories, and results should be field-
 validated  to  establish their ability to predict
 biological effects in sediments. As more toxicity
 tests are conducted, results  should  be  compiled
 in a central database so  that comparisons  can
 be  made among species, methods, and laboratories
 and so  that sediment  quality criteria can  be
 developed..
  4.5  REFERENCES

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ASTM. 1990a. E 1367-90. Guide for conducting
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,   Society for Testing and Materials, Philadelphia,
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'ASTM. 1990b. E 1391-90. Guide-for collection,
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 ASTM. 1990c. E 1383-90. Guide for conducting
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   Water and Environmental Technology, ,Vol.
    11.04.  American Society for  Testing  and
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 Battelle- 1988.  Overview of methods for assess-
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 Birge, W.J., J.  Black, S. Westerman, and  P.
    Francis. 1987. Toxicity of sediment-associated
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    procedures, pp. 199-218. In:  Fate and Effects
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 Borgmann,  U., and M. Munawar. 1989. A new
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 Cairns, MA., A.V. Nebeker, J.H.  Gakstatter, and
    W.L. Griffis, 1984. Toxicity of copper-spiked
    sediments to freshwater invertebrates. Environ.
    Toxicol. Chem. 3: 435-445.
  Chapman, P.M. 1986.  Sediment bioassay tests
    provide data necessary  for  assessment and
    regulation. In:   Proceedings of the Eleventh
   .. Annual Aquatic Toxicology Workshop.  G.H.
     Green and K.L. Woodward (eds.).  Technical
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 • Sediment Classification Methods Compendium
    Report 1480. Fish. Aquat. Sci., pp. 178-197.
 Chapman, P.M. 1988. Marine sediment toxicity
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    Winter, C.I.  Weber, 'and  L.  Fradkin (eds.).
    ASTM STP 976, American Society for Testing
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 Chapman, P.M., and R.  Fink.  1984.  Effects of
    Puget Sound sediments and their elutriates on
    the life cycle of Capitella capitate.   Bull.
    Environ. Contamination Toxicol. 33: 451-459.
 Connell, A.D., and D.D. Airey. 1982. The chronic
    effects of fluoride on the estuarine amphipods
    Grandidierella lutosa and G. lignorum. Water
    Res. 16: 1313-1317.
 DeWitt, T.H., G.R. Ditsworth, and R.C.  Swartz.
    1988.  Effects of natural sediment features on
    the phoxocephalid  amphipod, Rhepoxynius
    abronius: Implications for  sediment toxicity
    bioassays.  Mar. Environ. Res. 25: 99-124.
 DeWitt, T.H., R.C. Swartz, and J.O. Lamberson.
    1989.  Measuring  the toxicity of  estuarine
    sediments. Environ. Toxicol. Chem. 8: 1035-
    1048.                         .
 DiToro, D.M., J.D.  Mahony,  DJ. Hansen, KJ.
    Scott,  M.B.  Hicks,   S.M. Mayr,  and  M.S.
    Redmond. 1990. Toxicity of cadmium in sedi-
    ments:  the role  of acid volatile sulfide.' En-
    vironmental  Toxicology  and  Chemistry  9:
    1487-1502.
 DiToro, D.M., J.D. Mahony,  DJ. Hansen, KJ.
    Scott, A.R. Carlson,  and G.T.  Ankley. 1991.
    Acid volatile sulfide predicts the acute toxicity
    of cadmium and nickel in sediments. In press.
    Environmental Science and Technology.
 Ditsworth, G.R., D.W. Schults, and J.K.P. Jones.
    1990.  Preparation of benthic  substrates for
    sediment toxicity  testing. Environ.  Toxicol.
    Chem. 9: 1523-1529.        ;
 Francis, P.C., WJ Birge, and  JA. Black. 1984.
   Effects of cadmium-enriched sediment on fish
   and amphibian embryo-larval  stages.   Eco-
   toxicol. and Environ. Safety 8: 378-387.
 Hansen, D.J., and M.E. Tagatz.  1980. A labora-
   tory test for assessing  impacts of substances
   on  developing communities of benthic est-
   uarine  organisms,  pp 40-57.   In:  Aquatic
    Toxicology.  J.G. Eaton, P.R. Parrish and A.C.
    Hendricks (eds.). ASTM STP 707. American
    Society for Testing and Materials, Philadelphia,
    PA.
 Kemp, P.P., and R.C. Swartz. 1989. Acute toxicity
    of interstitial and particle-bound cadmium to a
    marine infaunal amphipod. Mar. Environ. Res.
    26: 135-153.
 Lamberson, J.O., and R.C. Swartz. 1988.  Use of
    bioassays in  determining the toxicity of sedi-
    ment to benthic organisms,  pp. 257-279. In:
    Toxic Contaminants and Ecosystem Health;
    Evans, M.S., (ed.) A Great Lakes Focus. John
    Wiley and Sons, New York, NY.
 Landrum, P.F. 1989. Bioavailability and toxico-
    kinetics of polycyclic aromatic hydrocarbons
    absorbed   to  sediments- for  the  amphipod
    Pontoporeia  hoyi. Environ. Sci. Technol. 23:
    588-595.                      ,
 McLeese, D.W.,  and CD.  Metcalfe.    1980.
    Toxicities of eight organochlorine compounds
    in sediment and seawater to Crangon septem-
    spinosa. Bull. Environ. Contain. Toxicol. 25:
    921-928.
 McLeese, D.W., L.E. Burridge, and J. Van Dinter!
    1982.  Toxicities of five organochlorine com-
    pounds in water and sediment to Nereis virens.
    Bull. Environ. Contain. Toxicol. 28: 216-220.
 Melzian,  B.D.  1990.  Toxicity assessment of
    dredged materials: acute and chronic toxicity
    as determined by bioassays and bioaccumula-.
    tion tests,  pp. 49-64. In:  Proceedings of the
    International  Seminar on the Environmental
    Aspects of Dredging Activities,   Goubault
    Impremeur, Nantes, France.
 Nebeker, A.V., MA. Cairns, J.H. Gakstatter, K.W.
    Malueg, G.S.  Schuytema, and D.F. Krawczyk.
    1984.  Biological  methods  for determining
   toxicity of contaminated freshwater sediments
   to invertebrates. Environ. Toxicol. Chem. 3:
   617-630.
 Nebekerj A.V., S.T. Onjukka, and MA. Cairns.
    1988. Chronic effects of contaminated  sedi-
   ment on Daphnia magna and Chironomus
   tentans. Bull. Environ. Contam. Toxicol. 41:
   574-581.
Nebeker, A.V., G.S. Schuytema, W.L. Griffis, JA.
   Barbitta, and  LA.  Carey.  1989.  Effect of
4-8

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                                                                         4—SSTT Approach
   sediment organic .carbon  on  survival  of
   Hyalella azteca exposed to DDT and endrin.
   Environ. Toxicol. Chem. 8: 705-718.
Oliver,  E.G.  1984.  Biouptake of chlorinated
   hydrocarbons from laboratory-spiked and field
   sediments by oligochaete worms.  Environ. Sci.
   and Technol. 21: 785-790.
Ott, F.S.  1986.  Amphipod sediment bioassays:
   Effect of grain size, cadmium, methodology,
   and variations in animal sensitivity on interpre-
   tation of experimental data. Ph.D. dissertation,
   University of Washington, Seattle, WA.
Pesch, C.E. 1979.  Influence of three sediment
   types on copper toxicity  to  the  polychaete
   Neanthes arenaceodehtata.  Marine Biol. 52:
   237-245.
Plesha, P.D.,  J.E.  Stein, M.H.  Schiewe, B.B.
   McCain, and U. Varanasi.  1988. Toxicity of
   marine sediments supplemented with mixtures
   of selected chlorinated and aromatic hydrocar-
   bons to the infaunal amphipod, Rhepoxynius
   abronius. Mar. Environ. Res. 25: 85-97.
 Puget Sound Estuary Program 1991. Recommend-
   ed guidelines  for conducting laboratory bio-
   assays on Puget Sound sediments. Draft report
   prepared for  U.S. Environmental Protection
   Agency, Region X,  Office of  Puget Sound,
   Seattle, WA.
 Schuytema, G.S., P.O.  Nelson, K.W. Malueg,
   A.V. Nebeker, D.F. Krawczyk,  A.K. Ratcliff,
   and J.H. Gakstatter. 1984. Toxicity of cadmi-
   um in water and sediment slurries to Daphnia
    magna.  Environ. Toxieol. and Chem. 3: 293-
    308.           '':'.-.
 Scott, K.J., and M.S. Redmond. 1989. The effects
    of a contaminated'dredged material on labora-
    tory populations of the tubicolous amphipod,
   Ampelisca abdita. In:  Aquatic Toxicology and
    Hazard Assessment:   Vol. 12  ASTM STP
    1027. U.M. Cowgill and L.R. Williams, (eds.).
    American Society for Testing  and  Materials,
    Philadelphia,  PA.
 Swartz, R.C. 1987.  Toxicological methods for
    determining the effects of contaminated sedir
    ment  on marine organisms pp. 183-193. In:
    Fate and Effects of Sediment Bound Chemicals
    in  Aquatic Systems.  K.L. Dickson, A.W.
    Maki,  and W.A.  Brungs, (eds.).   Pergamon
   Press, New York.
Swartz, R.C., D.W. Schults, G.R. Ditsworth, WA.
   DeBen, and FA. Cole.  1985. Phoxocephalid
   amphipod bioassay for marine sediment toxi-
   city.  pp 284-307. In:  Aquatic Toxicology
   and Hazard Assessment: Proceedings of the
   Seventh Annual Symposium. R.D. Gardwell,
   R. Purdy and R.C. Bahner (eds.).  ASTM STP
   854, American Society for Testing and Materi-
   als, Philadephia, PA.
Swartz, R.C., G.R. Ditsworth, D.W. Schults, and
   J.O. Lamberson.. 1986. Sediment toxicity to a
   marine infaunal amphipod: cadmium and its
   interaction with sewage sludge. Mar. Environ.
   Res: 18: 133-153.
Swartz, R.C., P.F. Kemp, D.W. Schults, and J.O.
   Lamberson.   1988.  Effects  of mixtures of
   sediment contaminants oh the marine infaunal
   amphipod, Rhepoxynius abronius.   Environ.
   Toxicol. Chem. 7: 1013-1020.
Swartz,  R.C., P.F.  Kemp, D.W. Schults, GJL
   Diteworth, and R.J. Ozretich. 1989.  Toxicity
   of sediment from Eagle Harbor, Washington to
   the infaunal amphipod, Rhepoxynius abronius.
   Environ. Toxicol. Chem. 8: 215-222.
 Swartz,  R.C., D.W.  Schults,  T.H. DeWitt, G.R.
   Diteworth, and J.O. Lamberson. 1990. Toxicity
   of fluoranthene .in sediment to marine amphi-
   pdds: a test of the equilibrium partitioning
   approach to sediment quality criteria. Environ.
   Toxicol. Chern. 9: lOyi-1080.
 Tagatz, M.E., J.M. Ivey, and H.K. Lehman. 1979.
   Effects of sevin on development of experimen-
    tal estuarine  communities.   J. Toxicol.  and
    Environ. Health 5: 643-651.
 Tagatz, M.E.,  J.M. Ivey, J.C. Moore, and M.
    Tobia. 1977.  Effects of pentachlorophenol on
    the development of estuarine communities.  J.
    Toxicol; and Environ. Health 3: 501-506.
 Tagatz, M.E., G.R. Plaia, C.H. Deans, and E.M.
    Lores. 1983. Toxicity of creosote-contamuiat-
    ed sediment to field- and laboratory-colonized
    estuarine benthic communities.   Environ.,
  :  Toxicol. Chem. 2: 441-450.
 Tetra Tech, Inc.  1986.  Development of sediment
    quality values for Puget Sound.  Task 6 Final
    Report. Tetra Tech, Inc., Bellevue, WA.
 Wiederholm, T.,  A.M.  Wiederholm, and  G.
                                                                                           4-9

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 Sediment Classification Methods Compendium
    Milbrink. 1987. Bulk sediment bioassays with
    five  species  of  fresh-water  oligochaetes.
    Water, Air and Soil Pollut. 36: 131-154.
 Word, J.Q.,  JA. Ward, L.M.  Franklin, V.L
    Cullinan and S.L. Kiesser. 1987. Evaluation of
    the equilibrium partition theory for estimating
the toxicity of the nonpolar  organic compound
DDT  to  the  sediment  dwelling   organism
Rhepoxynius abronius.  Report prepared for U.S.
Environmental  Protection Agency,  Criteria  and
Standards Division, Battelle Washington Environ-
mental Program Office, Washington, DC.
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           CHAPTER 5
Interstitial  Water Toxicity Identification

Evaluation  Approach
Gerald Ankley and Nelson Thomas
U.S. Environmental Protection Agency, Environmental Research Laboratory
6201 Congdon Boulevard, Duluth, MN 55804
(218) 720-5702
    The  interstitial water toxicity  approach is a
multiphase procedure for assessing sediment toxicity
using interstitial  (pore) water.  The use of pore
water for sediment toxicity assessment is based on
the strong correlations between contaminant concen-
trations in pore  water and observed exposure of,
benthic macrqinvertebrates to sediment-associated
contaminants (Adams  et al.,  1985; Swartz et al,
 1985; 1988; 1990; Connell et al., 1988; Khezovich
and Harrison, 1988; USEPA, 1989a; DiToro et al.,
 1990), as well as correlations between the actual
 toxicity of pore  water and bulk sediments  to epi-
benthic or benthic species (Ankley et al., 1991a).
 The approach combines the quantification of pore
 water toxicity with toxicity identification evaluation
 (TIE) procedures to identify and quantify chemical
 components responsible for sediment toxicity (U.S.
 Environmental Protection Agency, 1988;  1989b;
 1989c, 199la).  TIE involves the use of toxicity-
 based fractionation  procedures  to identify toxic
 compounds in aqueous samples containing mixtures
 of chemicals (Burkhard and Ankley, 1989). In the
 interstitial water toxicity method^ TIE procedures are
 implemented in three phases to characterize the
 nature of the pore water toxicants),  identify the
 suspect toxicants), and confirm identification of the
 suspect toxicants).
macroinvertebrates.  Although  the methods were
developed largely with freshwater species, they are
generally applicable to, and are currently being used
with, marine organisms  as well.  The procedures
have proven to be successful in identifying acutely
toxic substances in more than 90 percent of the
samples to which they  have been applied (e.g.,
Ankley et al, 1990a, 1991b;  Kuehl et al., 1990;
Amato et al, 1991; Norberg-King et al, 1991;
Schubauer-Berigan and Ankley, 1991; Ankley and
Burkhard, 1992).

S.L2  Potential Uses

    The use of  pore water as a fraction to assess
sediment toxicity, in conjunction with associated
TIE procedures, can provide data concerning specif-
ic compounds responsible for toxicity of contaminat-
ed  sediments.  These data could be critical to the
success of remediation of toxic sediments, including
the control of inputs of contaminants.
    In spite of  existing uncertainties  in preparing
and using pore water to assess sediment toxicity, the
ability to identify specific toxicants responsible for
acute toxicity in contaminated sediments makes pore
water an important test fraction. Thus this method,
 in  conjunction  with other sediment' classification
methods, could  prove to be extremely valuable.
  5.1  SPECIFIC APPLICATIONS

  5.1.1 Current Use
     The  TIE  procedures described herein were
  developed over the last 4 years using municipal and
  industrial effluents from more than 50 locations, as
  well as sediment samples from more than 10 differ-
  ent sites. They have been used with several aquatic
  species including cladocerans, fishes, and epibenthic
 5.2 DESCRIPTION
 5.2.1  Description of Method
     The interstitial water toxicity method involves
 three major steps:

     •  Isolation of pore water from sediment
      ,  samples;

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Sediment Classification Methods Compendium
    •  Performance  of toxicity tests  on pore
       waters; and

    •  Application of TIE procedures to pore
       water fractions.

    Pore water  can be isolated from sediment
samples by compression (squeezing) techniques,
displacement of water from sediment via the use
of inert gases, centrifugation, extraction via dialy-
sis, and micro-syringe sampling (Knezovich et al.,
1987; Knezovich and Harrison, 1988; Sly, 1988;
USEPA, 1991b).  The most representative pore
water samples may be obtained using the latter
two procedures.  However, the resulting sample
volumes are too small to be useful for toxicity
tests and associated TIE work.  Centrifugation has
been used  in a number of studies evaluating the
toxicily of sediment  pore water (Giesy  et al.,
1988;  Swartz et al.,  1989; Hoke et al., 1990;
Ankley et  al.,  1990a; Schubauer-Berigan and
Ankley, 1991) and comparative studies at Duluth,
as well as other laboratories, indicate that centrifu-
gation is a reasonable technique for pore water
preparation (Schults et al., 1991; U.S. Environ-
mental Protection Agency, 1991b). Regardless of
the techniques chosen for pore water isolation, the
method should not involve filtration either during
or  after  preparation  (Schubauer-Berigan  and
Ankley, 1991; USEPA, 1991b).
    After preparation of pore water, toxicity tests
can be performed using either standard test species
(e.g., USEPA, 1985a, 1985b) or various types of
epibenthic  or benthic  organisms amenable to
toxicity testing in aqueous samples (Ankley et a/.,
1991a; USEPA, 1991b).  In samples exhibiting
acute toxicity, it is then possible to directly apply
the TEE procedures described below in Section
5.2.1.2.2.

5.2.1.1 Objectives and Assumptions

    The objective of the interstitial water toxicity
method is to derive chemical-specific toxicity data
in the laboratory that can be used to assess sedi-
ment  toxicity in field situations.  With this ap-
proach, it  is possible to quantify toxicity in a
sample and potentially to identify chemical com-
ponents  responsible for toxicity.   The  major
assumption in using this method is that the com-
pounds that are toxic to test organisms in the pore
water, as it is isolated in the laboratory, are the
same compounds that cause toxicity in sediments
in situ.

5.2.1.2 Level of Effort

  •  Implementation of this method requires a
moderate amount  of  laboratory effort, both to
perform toxicity tests and to conduct TIE studies.
The effort expended in the TIE studies will be
proportional to the complexity  of analyses  re-
quired for the identification of suspected toxicants.

5.2.1.2.1  Type of Sampling Required

    Bulk sediment must be  obtained and pore
water  prepared  from the  sediments.   Routine
measurement of certain chemical components of
the pore water should be conducted.   These
measurements should include  (but are not limited
to) pH, hardness, alkalinity, salinity (where appro-
priate), dissolved oxygen, sulfides, and ammonia.
Certain of these variables,  in particular pH, also
should be monitored in the bulk sediment.

5.2.1.2.2 Methods

    The framework for existing TIE procedures is
summarized below.   Greater detail (e.g.; with
respect to all possible results that could be gener-
ated)  is  available elsewhere  (USEPA,  1988,
1989b, 1989c), as are specific methods for per-
forming sediment TIEs (USEPA, 1991b).
    Toxic sediment samples can potentially con-
tain thousands of chemicals,  and usually only a
handful are responsible for  the observed toxicity.
The goal of the TIE method is to identify quickly
and cheaply the  chemicals  causing  toxicity.
However, components causing toxicity can vary
widely, and potential toxicants include cationic
metals, polar and nonpolar organics, and anionic
inorganics, as well as ammonia or hydrogen
sulfide.  In addition, when  multiple toxicants are
present,  it must  be possible to determine the
proportion of the overall  toxicity due  to each
toxicant.
5-2

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                                                              5—Interstitial Water TIE Approach
    After preparation of pore water and  perfor-
mance of initial toxicity tests, the first step in the
TIE is to separate toxic from nontoxic components
in the pore water sample. To isolate the toxicants,
sample manipulations and subsequent fractionation
techniques are used in combination with toxicity
tests (toxicity tracking).  Each fractionation step
consists  of manipulations  to identify  the physi-
cal/chemical properties of the sample toxicants,
thereby enabling selection of the "correct"  analyti-
cal  technique for  detecting,  identifying,  and
quantifying  the  toxicants in  the  manipulated
samples.  Because there may be significantly
fewer chemical components in the manipulated
samples than in the original sample, the task of
deciding which component is causing the toxicity
is much easier. The toxicity-based TIE approach
enables  direct relationships to  be established
between toxicants and measured analytical  data
because toxicants are tracked through all sample
 fractionations, using the  most relevant  detector
 available, the organism.  Establishing this relation-
 ship ultimately results in highly efficient TIEs.
     With the toxicity-based  TIE approach, detec-
 tion of synergistic and antagonistic interactions, as
 well as matrix effects, for the toxicants is possible
 via toxicity tracking.  A priori knowledge of the
 toxicants' behavior in the aqueous phase is not
 required.
     The TIE approach is divided into three phases.
 Phase I consists of methods to identify the physi-
 cal/chemical nature of the constituents causing
 acute toxicity.  Phase II describes fractionation
 schemes and analytical methods to identify the
 toxicants, and Phase III presents procedures to
 confirm that the suspected toxicants are the cause
 of toxicity.

 Phase I: Toxicant Characterization—-In Phase
 I, the physical/chemical properties of toxicants are
 characterized by performing manipulations to alter
 or render biologically unavailable generic classes
  of compounds with similar properties.  Toxicity
 tests, performed in conjunction with the manipula-
  tions, provide information on the nature of the
  toxicants.   Successful completion  of  Phase I
  occurs when both the  nature of the components
  causing toxicity, as well as their consistent pres-
ence in a number of samples, can be established.
After Phase  I, the toxicants can  be tentatively
categorized as having chemical characteristics of
cationic metals, nonpolar organics, polar organics,
volatiles,  oxidants, and/or substances  whose
toxicity is pH-dependent.
    An overview of the sample  manipulations
employed in  Phase I is shown in Figure 5-1.  Not
shown in Figure 5-1, but performed on all sam-
ples, are routine water chemistry measurements
including  pH, hardness, conductivity, and dis-
solved oxygen.  These routine measurements are
needed for designing sample manipulations and
interpreting test data. The manipulations shown in
Figure 5-1 are usually sufficient  to characterize
toxicity  caused by a single chemical.   When
multiple toxicants are present, various combina-
tions of the Phase I manipulations will most likely
be required for toxicant  characterization.
     Many of the manipulations in Phase I require
samples that have been pH-adjusted.  The adjust-
ment of pH is a powerful tool for detecting cation-
 ic and anionic  toxicants since  their behavior is
 strongly influenced by pH.  By changing pH, the
 ratio of ionized to un-ionized species in  solution
 for a chemical  is changed significantly.   The
 ionized and un-ionized species  have different
 physical/chemical properties as well as toxicities.
 In Phase I, pH manipulations are used to examine
 two different questions:

     •   Is the toxicity different at various pHs?

     •   Does changing the pH, performing  a
         sample manipulation, and then readjusting
         to ambient pH affect toxicity?

 The graduated pH test examines the first question,
 and the pH adjustment, aeration, filtration, and
 solid phase  extraction (SPE) manipulations exam-
 ine the second.
      In the graduated pH test, the pH of a sample
 is adjusted within a physiologically tolerable range
 (e.g., pH 6.0, 7.0, and 8.0) before toxicity testing.
 In many instances,  the un-ionized form of a
 toxicant  is  able to cross biological membranes
 more readily than the  ionized form arid thus is
 more toxic.  This test is designed primarily for
                                                                                                5-3

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 Sediment Classification Methods Compendium
                 TOXIC AQUEOUS SAMPLE
         Oxidant
        Reduction
            Aeration
Acid
     PH|
                    Base
          i
                                      EDTA
                                    Ghelation
                             C 18 Solid Phase
                                Extraction
              Acid
              pH,
              Base
      Filtration
               pH Adjustment
                                        1
                                     Graduated pH
                                         Test
   Acid
PH,
      Base
                                      _._n
Acid
pH,
Base
pH6   pH7   pH8
Rgure 5-1. Overview of the Phase 1 Toxicity Characterization Process.
The ambient pH of the sample is indicated as pH,.
5-4

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                                                             5—Interstitial Water TIE Approach
ammonia, a relatively common toxicant  whose ;
toxicity  is  extremely pH-dependent  (USEPA,
1985c). However, different pH values can strong-
ly affect the toxicity of many common ionizable
pesticides, and also may influence the bioavailr
ability and toxicity of certain heavy metals and
surfactants (Campbell and Stokes, 1985; Doe et
al., 1988).
    Aeration  tests  are  designed to determine
whether toxicity is attributable to volatile,  oxidiz-
able, or sublatabie  compounds.  Samples at pH,
(ambient pH), pH 3, and pH 11 are sparged with
air for 1 h, readjusted to pH, and tested for toxici-
ty.  The different pH values affect the ionization
state of polar toxicants, thus making them more or
less volatile, and also affect the redox potential of
the system. If toxicity is reduced by air sparging
at any of the pH values, the presence of volatile or
oxidizable  compounds may  be suggested.   To
distinguish the former from the latter  situation,
 further experiments are performed using nitrogen
 rather than air to sparge the samples.  If toxicity
 remains the same,  oxidizable materials are impli-
 cated; if toxicity is again reduced, volatile com-
 pounds are suspect.  The pjl at which toxicity is
 reduced is important.   If nitrogen sparging  de-
 creases toxicity at pHD neutral volatiles are pres-
 ent; if toxicity decreases  at pH 11.0 or  pH 3.0,
 basic and acidic volatiles, respectively, are impli-
 cated.   An  additional process  through which
 aeration can remove sample toxicants is sublation,
 which is the movement  of compounds  through
 aqueous solutions at the surface of the air bubbles,
 often followed by deposition on the aeration glass-
 ware.    Compounds  that  exhibit  this  behavior
 include resin acids and surfactants; in some in-
  stances  it may be possible to implicate the pres-
  ence of sublatabie compounds by rinsing the
  aeration glassware with clean laboratory dilution
  water and testing this fraction  (Ankley et al.,
  199Gb).
      Filtration provides information concerning the
  amount of  tpxicity associated with  filterable
  components. In this test, samples at pHD pH 3.0,
  and pH 11.0 are passed through l-/on glass fiber
  filters, readjusted to pH,, and tested for toxicity.
  Reductions  in toxicity  due to filtration  could be
  related to factors  such as  decreased  physical
toxicity, rather than chemical toxicity (Chapman
et al.,  1987), or removal of particle-bound toxi-
cants, which could be important, particularly if
filter-feeding organisms such as cladocerans are
the test species.
    Reversed-phase, solid-phase extraction (SPE)
is designed to determine the extent of toxicity due
to compounds that, are relatively nonpolar at pHe
pH 3.0, or pH 9.0.  This test, in conjunction with
associated Phase .n analytical  procedures, is an
extremely powerful TIE tool.   In this procedure,
filtered sample aliquots at pH,, pH 3.0, and pH 9.0
are passed through small columns packed with an
octadecyl (Qg) sorbent.  At pHfc the C18 sorbent
will remove neutral compounds such as certain
pesticides (Junk and Richard, 1988).  By shifting
ionization equilibria  at the low and  high pH
values, the  SPE column  also can be used to
 extract organic  acids and  bases (Wells and Mi-
 chael, 1987).   During extraction, the  resulting
 post-column  effluent is collected and  tested for
 toxicity to determine whether the manipulation
 removed toxicity and/or whether the capacity of
 the column was exceeded.  Following this, the
 column is eluted with solvents, such as methanol,
 which then can be tested for recovery of toxicity.
 If sample toxicity is decreased and subsequently
 recovered in solvent elutibns, a nonpolar toxicant
 would be suspected.
     The presence of toxicity due to cationic metals
 is tested through additions of ethylenediaminetet-
 raacetic acid (EDTA), a strong chelating agent that
 produces nontoxic complexes with many metals.
 As with SPE chromatography, the  specificity of
 the EDTA test for a class-of ubiquitous toxicants
 makes it a powerful TIE tool.  Cations chelated by,
 EDTA include certain forms of aluminum, barium,
  cadmium,  cobalt, copper, iron, lead, manganese,
  nickel, strontium, and zinc (Stumm and Morgan,
  1981).  EDTA does not complex anionic forms of
  metalSj and only weakly chelates certain cationic
  metals,  such as silver, chromium, and thallium
  (Stumm and Morgan, 1981).
     The oxidant  reduction test is designed to
  determine the  degree of toxicity associated with
  chemicals reduced, or in some instances chelated,
  by sodium thiosulfate. The  toxicity of oxidahts
  such as chlorine, bromine, iodine, and manganous
                                                                                               5-5

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  Sediment Classification Methods Compendium
  ions  is  neutralized by sodium thiosulfate,  and
  metals such as copper, cadmium, and silver are
  chelated  and rendered biologically  unavailable
  (Hockett and'  Mount, 1990).   Because sodium
  thiosulfate, like EDTA, has low toxicity to most
  aquatic  organisms,  a relatively wide range of
  concentrations  can be tested.

  Phase II:  Toxicant Identification—Initial labo-
  ratory work in Phase II focuses on isolation of
  the toxicants using  chemical fractionation tech-
  niques with toxicity tracking. The ideal isolation
  process would create a subsample that contains
  one chemical,  the toxicant.  Depending on  the
  nature of the toxicants, wide differences in  the
  techniques, as  well as  in  the  amount of  effort
  required for fractionation, will occur.
     In general, after fractionation, instrumental
  analyses are performed on the toxic subsamples,
  and lists of identified chemicals are assembled
  for each  subsample. For each chemical in a list,
 an  LQo value  is  obtained, usually from  the
 literature or occasionally from structure activity
 models (Institute  for Biological and Chemical
 Process Analyses, 1986). By comparing concen-
 trations of the identified chemicals to their LCjo
 values, a list of suspect toxicants is made. This
 list is then refined by actually determining LQ,,
 values for the suspects using the TIE test species.
 If only one toxicant is present, it should be easily
 identified.  For samples with multiple toxicants,
 identification becomes significantly more pro-
 tracted since interactions among toxicants may
 need to be examined. If none of the suspected
 toxicants appears to explain the toxicity, the true
 toxicants were  probably  not  detected  during
 instrumental analysis. Usually, additional separa-
 tion and  associated  concentration  steps are re-
 quired  to increase the analytical sensitivity for
 toxicant identification.
    The information obtained in Phase I provides
 the  analytical  roadmarks  for  performing the
 fractionation and identification tasks in Phase  II.
 To illustrate the relationship between Phase I
 data and the analytical approaches employed  in
 Phase  II, results for two  typical Phase  I  TIE
 evaluations are presented in Table 5-1.   The
 Phase II methods and approaches appropriate for
  these examples are discussed below.
     In the first sample in Table 5-1, SPE reduced
  toxicity.  In Phase II, the SPE column is eluted
  with  graded,  increasingly nonpolar methanol/-
  water solutions, and toxicity testing is performed
  on each fraction (Burkhard  et al, 1990).  Al-
  though solvents other than methanol are routinely
  used in analytical work with C18 chromatography
 i columns, the low toxicity of methanol to aquatic
  organisms (e.g.,  LQ,, il.5  percent.for  clado-
  cerans) makes it a solvent of choice for toxicity
  tracking in the fractions. If no toxicity occurs in
  the fractions,  the toxicants  have been  lost and
  further characterization (Phase  1) work is  re-
  quired.  If toxicity occurs in the fractions, Phase
 II methods feature concentration of the toxic
 methanol/water  fractions;  high  performance
 liquid chromatography fractionation of the con-
 centrate (again with a C18/methanol/water solvent
 system),  with  concurrent toxicity testing of the
 fractions; and, ultimately, identification of sus-
 pected toxicants in the toxic fractions via  gas
 chromatography/mass spectroscopy.  For pore
 water TIE, toxicity caused by high log  k^, mon-
 polar organics is often not elutable with metha-
 nol. In these cases, it is useful to elute the SPE
 column with a  less polar solvent (i.e., methylene
 chloride) (Schubauer-Berigan and Ankley, 1991).
    In the second sample, both EDTA additions
 and SPE reduced  toxicity.   The  reduction  of
 toxicity by EDTA strongly suggests the presence
 of toxic levels of cationic metals. Thus, Phase II
 procedures would include both metal  analyses
 and the concentration, fractionation, and identi-
 fication techniques  described for nonpolar  or-
 ganics in the first example. If analyses  identify
 specific metals at concentrations high enough to
 cause toxicity, various mass balance procedures
 can be used to define  the portion of the sample
 toxicity due to the suspected metals  and the
 portion of the toxicity due to the suspect non-
 polar compounds.
    Only a very small subset of possible Phase I
 results is shown in Table 5-1, particularly when
 one considers that  three of the tests (aeration,
 filtration,  SPE) are conducted at three different
pH values.  A complete  discussion of the types
of Phase I results that may be encountered and
5-6

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                                                              5—Interstitial Water TIE Approach
                      Table 5-1.  Phase 1 Characterization Results and Suspect
                              Toxicant .Classification for Two Samples.

Phase (Test
Oxidant reduction • • .
: EDTA addition
Graduated pH test
pH adjustment
Filtration
Aeration
SPE
Methanol fractions
Suspected toxicant classification
Sample '
One
NR*
NR .
NR

NR
NR
Rb
r
Nonpolar organics
j
Two
NR
R
NR
NR
NR
NR
R
T
Nonpolar organics/heavy
metals
   •NR = No reduction in toxicrty.
   "R = Reduction in toxicrty.
   T = Toxicity recovered.
subsequent Phase II strategies  that  could be
implemented is beyond the scope of this review.

Phase III: Toxicant Confirmation—After Phase
II identification procedures implicate  suspected
toxicants, Phase III is initiated to confirm that the
/suspects are indeed the true toxicants. Confirma-
tion is perhaps the most critical step of the/TIE
because procedures used in Phases I and II may
create artifacts that could lead to erroneous con-
clusions about the toxicants.  Furthermore, there is
a possibility that substances causing toxicity are
different from sample to sample within a suppos-
edly homogeneous geographic region;   Phase HI
 enables both situations to be addressed.  The tools
used  in  Phase in include correlation,  relative
 species sensitivity,  observation  of symptoms,
 spiking, and mass balance techniques.  In most
 cases, no single Phase III test is adequate to con-
 firm suspects as the true toxicants; it is necessary
 to use multiple confirmation procedures.
    In the correlation approach, observed toxicity
is  regressed against expected  toxicity due to
measured concentrations of the suspected toxicants
in samples  collected over time or from several
sites within a location.  For the correlation ap-
proach to succeed, temporal or spatial variation
has to be wide  enough  to provide  a range of
values adequate for meaningful analyses.  To use
the correlation approach effectively when there are
multiple suspect  toxicants,  it  is  necessary to
generate data" concerning the additive, antagonistic,
and synergistic effects of the toxicants in ratios
similar to those found in the samples.  These data
also are needed for the spiking and mass balance
techniques described below.
    The relative sensitivity of different test species
can be used to evaluate  suspected toxicants.  K
two or more species  exhibit markedly different
sensitivities to a suspected toxicant in standard
reference tests, and the same patterns in sensitivity
are seen  with the toxic pore water sample, this
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  Sediment Classification Methods Compendium
 provides evidence for the validity of the suspect
 being the true toxicant
     Another Phase ni procedure is observation of
 symptoms (e.g.,  time to mortality) in poisoned
 animals.  Although this approach does not neces-
 sarily provide support for a given suspect, it can
 be used to provide evidence against a suspected
 toxicant  If the symptoms observed in a standard
 reference test  with a suspected toxicant differ
 greatly  from those 'observed with the pore water
 sample  (which contains similar concentrations of
 the suspected toxicant), this is strong evidence for
 a misidentification.
     Confirmatory evidence also can be  obtained
 by spiking  samples with the suspect toxicants.
 While  the results  may not be conclusive,  an
 increase in toxicity by the same proportion as the
 increase in concentration of the suspect toxicant in
 the sample suggests that the suspect is correct To
 obtain a proportional increase in toxicity from the
 addition of a suspect toxicant when in fact it is
 not the  true toxicant,  both the true and suspect
 toxicants would have to have very similar toxicity
 levels and their  effects  would also have to  be
 additive.
    Mass  balance calculations can be  used  as
 confirmation steps when toxicity can be at  least
 partially removed from the pore water sample, and
 subsequently recovered.   This  approach can  be
 useful in instances  when SPE removes  toxicity.
 The  methanol  fractions eluted  from the  SPE
 column  are evaluated individually for toxicity;
 these toxicities are summed and then compared to
 the total amount of toxicity lost from the sample.
    Other techniques, including alteration of water
 quality  characteristics (e.g.,  pH, salinity) in a
 manner designed to affect the toxicity of specific
 compounds, and  analysis  of body burdens  of
 suspected toxicants in exposed animals, also can
 be useful confirmation steps.

 5.2.123   Types of Data Required

    In addition  to the  routine measurements de-
 scribed above, biological response data, either acute
 or chronic, will be  obtained. Specific data collected
will depend on the  choice of test organism and
 endpoints.   If the TIE process  is  initiated, the
 researcher will first obtain  data concerning the
 physical/chemical characteristics of the toxicants in
 the pore water, followed by actual identification of
 toxic  compounds, and standard determination of
 their concentrations in the toxic samples (see Sec-
 tion 5.2.1.2.2 above).

 5.2.1.2.4   Necessary Hardware and Skills

    Pore water preparation and toxicity test proce-
 dures  are fairly straightforward  and require com-
 monly available equipment and facilities. Many of
 the TEE procedures also require only routine facili-
 ties. However, certain TIE techniques require some
 degree of  advanced  analytical capability  (e.g.,
 atomic absorption spectroscopy, gas chromatogra-
 phy/mass spectroscopy).  Similarly, although many
 of the routine toxicity tests require relatively  little
 training, certain of the TIE procedures, in particular
 some  of the chemical analyses, require advanced
 technical expertise and experience.

 5.2.13 Adequacy of Documentation

    The theoretical basis for using pore water to
 assess toxicity  appears to be scientifically sound,
 and pore water has been used for sediment toxicity
 evaluation (Adams et al., 1985; Swartz et at., 1985,
 1988,1990; Knezovich and Harrison, 1988; Connell
 et al., 1988; Giesy et al., 1988; USEPA, 1989a;
 Ankley et al,  1990a, 1991a, 1991b; Hoke et dl,
 1990;   Schubauer-Berigan and   Ankley,   1991).
 Toxicity tests that can be used are in many instances
 well-documented, standard procedures (U.S. EPA,
 1985a;  1985b).  The TEE techniques  involved,
 including those specifically for sediments, have been
 documented (USEPA, 1988, 1989b, 1989c, 1991a,
 1991b). Also, sediment TEEs with pore water have
 been  successfully demonstrated  (Ankley  et  al.,
 1990a,  1991b;  Schubauer-Berigan et  al., 1990;
 Schubauer-Berigan and Ankley, 1991).

 5.22    Applicability of Method to Human
        Health, Aquatic Life, or Wildlife
        Protection

   ..This method can be used to  predict acute and
chronic (i.e., growth or reproductive) effects of toxic
5-8

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                                                              5—Interstitial Water TIE Approach
sediment on aquatic organisms  and can identify
toxicants responsible for observed effects. The data
generated thus can be used to, design sediment
remediation programs that would have an optimal
likelihood of success.  These procedures are not
suitable, however,  for .evaluating human  health
effects  or protecting wildlife, and they cannot be
used to address bioconcentratable toxicants.
        \     ,        ,
5.23 Ability of Method to Generate
      Numerical Criteria for Specific
      Chemicals                      .      .

     Pore water toxicity assessment, in conjunction
with successful TIE procedures, can be used to
generate numerical criteria for toxic compounds in
sediment pore water  because  the toxicants  are
.actually identified.  However,  it must be estab-
 lished that compounds identified as being toxic to
test organisms  in  the laboratory  are the same
 compounds  (both  in  form  and concentration)
 responsible  for  toxicity to organisms in field
 situations. This relationship can be evaluated both
 through biosurveys (possibly in conjunction with
 analysis  of contaminant residues in organisms
 collected from the field), and laboratory toxicity
 tests in which benthic organisms perceived to be
 affected  in  contaminated sediments  in situ  are
 exposed to toxicants identified in the pore water.
 Both types of data also would be required  for any
 sediment classification method based on toxicity
 or chemical analyses.
 5.3 USEFULNESS
 5.3.1 Environmental Applicability

 5.3.1.1  Suitability for Different Sediment Types

     The pore water toxicity assessment approach
 is suitable for any sediment from which adequate
 quantities of pore water can be isolated. In typical
 sediments, 20 50 percent of the volume of the
 bulk sediment sample is pore water.  For a com-
 plete Phase I characterization with a test species of
 relatively small body size (e.g., cladocerans, larval
 fishes), approximately 1.5 L of pore water is re-
 quired.   This translates  into a bulk sediment
requirement of 3-8 L.  Bulk  sediment volumes
needed for Phase n identification will, of course;
be dependent on the toxicants present in the pore
water, but typical  volumes required would be
expected to range from 1 to 20 L.

5.3.1.2 Suitability for Different Chemicals or
       Classes of Chemicals               ,

    This  approach appears to  be  suitable  for
various nonpolar organics, cationic metals, and
ammonia (Adams et al., 1985; Swartzeffl/., 1985,
1988, 1990;  Knezovich and  Harrison,  1988;
Connell  et al., 1988; USEPA, 1989a;  Ankley et
al.,  1990a, 1991b; DiToro et al., 1990).  The
applicability of the approach to toxicants such as
polar organics or extremely lipophilic compounds
has  yet to be established.  Also, the TIE proce-
dures enable the evaluation of interactive (addi-
tive, synergistic,  antagonistic)  effects  among
various toxicants present in pore water samples.

5.3.1.3 Suitability for Predicting Effects on
        Different Organisms .

     If the TIE procedures successfully! identify
specific toxicants responsible for sediment toxici-
ty,  the  impacts of these toxicants on  various
species of concern can be easily predicted, provid-
 ed that there are data concerning the  toxicity of
the  identified  compounds to  these species.  Al-
 though  toxicity  data may hot be available for
 certain benthic species, once suspect toxicants are,
 identified, it would be possible to generate tbxicity
 data for specific species of concern.

 5.3.1.4 Suitability for In-Place Pollutant Control

     The pore  water  toxicity  assessment method
 and associated TIE procedures could prove to be
 a powerful tool for in-place pollutant control. Be-
 cause sediment toxicants are actually identified, it
 is possible to design remediation plans for toxi-
 cants from point sources or controllable nonpoint
 sources, and to routinely monitor the success of
 these plans through continued assessment of pore
 water for toxicity and specific chemical toxicants.
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 Sediment Classification Methods Compendium
 5.3.1.5 Suitability for Source Control .

     Because the potential exists for  identifying
 specific sediment toxicants,  this method is ideal
 for point source control, as well as controllable
 nonpoint source inputs.

 53.1.6 Suitability for Disposal Applications

     As stated  above, because specific sediment
 toxicants can be identified, it would be possible to
 identify potential hazards of contaminated sedi-
 ments to aquatic organisms before disposal opera-
 tions, such  as those associated with dredging
 (Ankley et al, 1991c).
 53.2  General Advantages and Limitations

 53.2.1 Ease of Use

    Pore water  preparation,  routine  chemical
 analyses, toxicity tests, and certain  of the TEE
 procedures  are reasonably straightforward and
 require  relatively  little  technical expertise or
 extensive laboratory  facilities.   Because  it is
 possible to work with aqueous samples, many of
 the standard toxicity tests developed for toxicity
 assessment of surface waters and effluents can be
 used,  in addition to tests with  various benthic
 species (e.g., USEPA,  1985a,  1985b). However,
 interpretation of results  of certain of the  TIE
 procedures, as well  as  analytical support for the
 TIE work, requires advanced training and experi-
 ence.  Also, several TIE analyses require highly
 sensitive analytical instrumentation for procedures,
 such as atomic absorption spectroscopy and gas
 chromatography/mass spectroscopy.

 5.3.2.2 Relative Cost

    Cost of the actual toxicity  test procedures is
 relatively low.  Cost of the TIE  procedures  will
 vary depending on the nature of the toxic com-
 pounds;  certain toxicants (e.g.,  pesticides), are
 more costly to identify and quantify  than others
 (e.g., ammonia). Also, identification and determi-
 nation  of the effects  of  multiple toxicants in
 samples  costs more than  the identification of
 single toxicants. Thus, cost analysis for the TIE
 portion of the toxicity assessment is case-specific.

 5.3.2.3 Tendency to Be Conservative

    Depending on the species  used  and the end-
 point evaluated, pore water toxicity tests can be as
 conservative as  desired!   However, acute pore
 water toxicity tests described for sediment TIE are
 not meant to represent chronic or bioaccumulation
 endpoints.

 5.3.2.4 Level of Acceptance

  .. The theoretical basis of pore water toxicity
 assessment is sound (Adams et al., 1985; Swartz
 et al 1985, 1988, 1990; Knezovich and Harrison,
 1988;  Connell  et al.,  1988;   USEPA,  1989a;'
 DiToro et al, 1990; Ankley et al, 1991a). The
 most important, advantage of using pore water as
 a sediment test fraction, however, is the fact that
 it enables the application of recently developed
 TIE procedures  for  the  identification of toxic
 compounds in aqueous samples containing com-
 plex mixtures of chemicals (USEPA, 1988,1989b,
 1989c,  1991a,  1991b).  TEE  procedures have
 proven to be extremely powerful  tools for work
 with both complex effluents and  sediment pore
 water (Ankley et al,  1990a, 1991b; Kuehl et <«/.,
 1991; Amato et al,  1991;  Norberg-King et al^
 1991;  Schubauer-Berigan  and Ankley,  1991;
 Ankley and  Burkhard, 1992).   The ability  to
 identify specific  compounds responsible for ithe
 toxicity of contaminated sediments clearly could
 be critical to the success of remediation.

 5.3.2.5 Ability to Be Implemented by
       Laboratories with Typical Equipment
       and Handling-Facilities

    Pore water preparation, toxicity  test  proce-
 dures, and certain of the TEE methods are easily
 implemented by laboratories with typical equip-
ment and a moderate degree of expertise.  Inter-
pretation of some TEE results requires additional
technical training and experience,  and certain  of
the analytical procedures associated  with TEE
5-10

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                                                             5—Interstitial Water TIE Approach
work require both specialized training and analyti-     ry-controlled experiments, results obtained are
cal instrumentation:                                statistically accurate and precise.
5.3.2.6 Level of Effort Required to Generate
       Results               ;

    This procedure consists of field sampling,
preparation of  pore water, toxicity tests, and
various  TIE  procedures.   Depending on  the
results of the TIE work, the level of effort ex-
pended to obtain potentially important data can
be relatively small.

5.3.2.7 Degree to  Which Results Lend
        Themselves to Interpretation

     Biological  responses (i.e., toxicity) can be
 easily interpreted,  and when properly performed,
 the results of the TIE procedures can be straight-
 forward and easily interpreted; however, this is
 dependent on the  complexity of the sample and
 the number of compounds contributing to sample
 toxicity.

 5.3.2.8 Degree of Environmental Applicability

     Pore  water toxicity  assessment  and  TIE
 procedures  are applicable to virtually all envi-
 ronmental  conditions  and  sediment  types.
 Moreover, a wide variety  of test organisms can
 be evaluated  with this approach.   However,
 although data indicate  that the toxicity and/or
 bioaccumulation of a variety of contaminants are
  correlated with their pore water concentrations,
 there is no guarantee that this relationship exists
  for all types of contaminants. For example, a
  potentially important  route  of  exposure  for
  highly lipophilic  compounds is thought to be via
  ingestion of contaminated particles.  This  route
  is not addressed using pore water  exposures.
  Finally, existing  TIE procedures are applicable
  for acutely toxic samples,  and  thus generally
  would not be useful for identifying chronically
  toxic sediment contaminants.

  5.3.2.9 Degree of Accuracy and Precision

      Because the  procedures consist of laborato-
5.4 STATUS

5.4.1  Extent of Use
                     •          ,      '    -
    Various toxicity tests have been widely ap-
plied  to the evaluation of both freshwater and
marine sediments, and pore water is merely one of.
the possible fractions that can be tested. Theoreti-
cally,  pore  water-appears to be appropriate for
sediment toxicity assessment and there have been
many examples of its use for this purpose (Adams
et  al., .1985;  Swartz et al, 1985, 1988,  1990;
Giesy et al, 1988; Knezovich and Harrison, 1988;
Connell et al,  1988; USEPA, 1989a;  Ankley,
1990a, 1991a, 1991b; DfToro et al, 1990; Hoke
et  al,  1990; Schubauer-Berigan  and  Ankley,
 1991).  The TIE  procedures (USEPA,  1988,
 1989b, 1989c, 1991a, 1991b) although developed
 only relatively recently, already are widely used in
both research and regulatory programs.

 5.4.2  Extent to Which Approach Has Been
       Field-Validated  ,

    Because the procedure is relatively new, there
 has been little field validation.  This area requires
 research, not  only for the pore water TIE methods
 described herein, but for virtually any other sediment
 method involving toxicity tests or chemical analyses.

 5.4.3  Reasons  for Limited Use

     Various sediment toxicity tests have been widely
 used; however, relatively few studies have evaluated
 pore  water toxicity.-  This is primarily because Ihe
 theoretical basis for using pore water  has only
 recently been critically evaluated.  For this reason,
 ihere are no standard melhods for pore water prepa-
 ration. Systematic TIE procedures for toxic aqueous
  samples have only recently been developed and thus
  have not yet been widely. applied to the area of
  sediment toxicity assessment  Because current TIE
  procedures cannot be used with bulk sediment sam-
  ples, pore water appears to be the best fraction with
                                                                                              5-11

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  Sediment Classification Methods Compendium
  which  to  attempt  to  identify  specific  sediment
  contaminants responsible for acute toxicity.


  5.4.4  Outlook for Future Use and Amount of
        Development Yet Needed

     The outlook for this approach is extremely
  promising because it is the'only method currently
  available that enables the identification of specific
  compounds responsible for sediment toxicity with
  some degree of certainty.  This information could
  be critical to the success of remediation.  Howev-
  er, as with all of the existing sediment methods,
  further development is needed, particularly in .the
  following areas:

     »   The development of standard arid scientif-
         ically sound techniques for  pore water
         isolation;

     •   Further characterization  of relationships
         between sediment toxicity in situ and the
         toxicity  of  sediment  pore water in the
         laboratory for different classes of comp-
         ounds; and

     «  The development of TIE procedures  to
        identify chronically toxic compounds  in
        aqueous samples.

    Research in all these areas is ongoing at ERL-
 Duluth.

    For more information please contact:

    Gerald Ankley and Nelson Thomas
    U.S. Environmental Protection Agency
    Environmental Research Laboratory
    6201 Congdon Boulevard
    Duluth, MN  55804
    (218) 720-5603

    Mary K. Schubauer-Berigan
    AScI Corporation
    6201  Congdon Boulevard
    Duluth, MN 55804
    (218) 720-5619
  5.5 REFERENCES


  Adams, W.J.,  R.A. Kimerle,  and R.G. Mosfaer.
    1985. Aquatic safety assessment of chemicals
    s6rbed to sediments, pp. 429-453.  In: Aquatic
    Toxicology and Hazard Assessment:  Seventh
    Symposium. R.D. Cardwell, R. Purdy, and R.C.
    Banner (eds.). ASTM.STP854. American Soci-
    ety for Testing and Materials, Philadelphia, PA.
  Amato,  J.R.,  D.I.  Mount, E J. Durban,  M.T.
,   Lukasewycz, G.T. Ankley,  and E.D. Robert.
    1991.   An example of the identification of
    diazinon as a primary toxicant in  an effluent.
    Environ. Toxicol. .Chem. In press.
  Ankley,  G.T. and L.P. Burkhard.  1992. 'Identifi-
    cation of surfactants as toxicants in a primary
    effluent.  Environ. Toxicol. Chem.  .Submitted.
  Ankley,  G.T., A. Katko, and J.W. Arthur.  1990a.
    Identification of ammonia as an important sedi-
    ment-associated toxicant in the lower Fox River
    and Green Bay, Wisconsin.  Environ. Toxicol.
    Chem. 9:313-322.
  Ankley,  G.T., M.T.  Lukasewycz, G.S.  Peterson,
    and DA. Jenson.   1990b. Behavior of surfact-
    ants   in   toxicity  identification  evaluations.
    Chemosphere. 21:3-12.
 Ankley, G.T., M.K. Schubauer-Berigan, and J.R.
   Dierkes.  1991a.  Predicting the toxicity of bulk
   sediments to  aquatic organisms with aqueous
   test fractions:   Pore water  versus  elutriate.
   Environ. Toxicol. Chem.  In press.
 Ankley, G.T., GX. Phipps, P.A. Kosian, DJ. Han-
   sen, J.D.  Mahony, A.M. Cotter, E.N. Leonard,
   J.R. Dierkes,  DA. Benoit, and V.R.  Mattsora.
   1991b. Acid volatile sulfide as a factor mediat-
   ing cadmium and nickel  bioavailability in
   contaminated  sediments.   Environ.  Toxicol.
   Chem.  In press.
 Ankley, G.T., M.K. Schubauer-Berigan, and RA.
   Hoke.   1991c.  Use of toxicity identification
   evaluation techniques to identify dredged mate-
   rial disposal options:  A proposed'approach.
   Environ. Management. In press.
 Burkhard, L.P., and  G.T.  Ankley.    1989.
   NETAC's  toxicity-based approach to identify
   toxicants.   Environ. Sci.  Technol.   23:1438-
   1443.
5-22

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                                                            5—Interstitial Water TIE Approach
Burkhard,   L.P.,  EJ.  Durban,  and   M.T.
  Lukasewycz.  1990.  Identification of nonpolar
  toxicants in effluent  using toxicity-based frac-
  tionation with gas  chromatography/mass spec-
  trometry.  Anal. Chem. 63:277-283.
Campbell,  P.G.C.,  and P.M.  Stokes.   1985.
  Acidification and toxicity of metals to aquatic
  biota.  Can. J. Fish. Aq. Sci. 42:2034-2049.,
Chapman,  P.M., J.D.  Popham,  J. Griffin, D.
  Leslie, and J. Michaelson.  1987. Differentia-
  tion of physical from chemical toxicity in solid
  waste fish bioassays.  Water  Air Soil  Pollut.
  33:295-308.                    .
 Connell, D.W., M.  Bowman, and D.W. Hawker.
   1988. Bioconcentration of chlorinated hydrocar-
  bons  from sediment by oligochaetes.   Eco-
   toxicol. Environ. Safety 16:293-302.
 DiToro, D.M.,  J.D. Mahony, D.J. Hansen, KJ.
   Scott, M.B. Hicks, S.M. Mays, and M.S. Red-
   mond.  1990.  Toxicity of cadmium in sedi-
   ments:    the  role  of  acid  volatile  sulfide.
   Environ. Toxicol. Chem. 9:1489-1504.
 Doe, K.G. W.R. Ernst, W.R. Parker, G.R J. Julien,
   and PA. Hennigar.  1988. Influence of pH on
   the acute lethality  of fenitrothion, 2,4-D and
   aminocarb and some pH-altered sublethal effects
   of aminocarb on rainbow trout (Salmo gaird-
'   neri). Can. J. Fish. Aq. Sci. 45:287-293.
  Giesy,  J.P., R-L.  Graney,  J.L. Newsted, C.J.
   Rosiu, A. Benda, R.G. Kreis, and F.J. Horvath.
    1988. Comparison of three sediment bioassay
    methods  using Detroit  River sediments.  En-
    viron. Toxicol. Chem. 7:483-498.
  Hockett, J.R.,  and D.R. Mount.  1990.  Use of
    metal chelating agents  to differentiate among
    sources  of toxicity. Eleventh Annual  Meeting
    of the Society of Environmental Toxicology and
    Chemistry, Abstract, p. 162.
  Hoke, RA., J-P. Giesy, G.T. Ankely, J.L. New-
    sted, and J.R. Adams.   1990.   Toxicity of
    sediments from western Lake Erie and  Maumee
    River at Toledo, Ohio, 1987:  Implication for
    current dredged  material  disposal  practices. J.
    Great Lakes Res.  16:457-4,70.
  Institute for Biological and Chemical Process Analy-
    ses.  1986.  User manual for QSAR system.
    Montana State University, Bozeman, MT.
Junk, GA., and J. J.Richard. 1988. Organicsin
  water:  Solid phase extraction on a small scale.
  Anal. Chem. 60:451-454.
Knezovich, J.P., and F.L. Harrison. 1988. The
  bioavailability  of  sediment sorbed  chlorp-
  benzenes to larvae of tiie midge Chironomus
  decorus.    Ecotoxicol.     Envuron.  Safety
  15:226-241.
Knezovich, J.P.,  F.L. Henderson, and R.iG. Wil-
  helm.  1987. The bioavailability of sediment-
  sorbed organic chemicals:  A review. Water Air
  Soil Pollut 32:233-245.
Kuehl, D.W., G.T. Ankley, L.P. Burkhard, and
  DA. Jensen.  1990. Bioassay directed charac-
  terization of the acute toxicity of  a creosote
   leachate. Hazardous Waste Hazardous Mater.
   7*283-291
. Norberg-King, TJ., EJ. Durban, G.T. Ankley, and
   E. Robert  1991. Application of toxicity identi-
   fication evaluation procedures-to the ambient
   waters of the Colusa Basin Drain:  Environ.
   Tox. and Chem. In press.
 Schubauer-Berigan, M.K., J.R. Dierkes, and G.T.
   Ankley.  1990.  Toxicity identification evalua-
   tions  of contaminated sediments in the Buffalo
   River, NY and Saginaw River, MI.   National
   Effluent Toxicity /Assessment Center Rep. No.
   20-90.   Environmental Research  Laboratory,
   Duluth, MN.
  Schubauer-Berigan, M.K.,   and G.T. Ankley.
    1991.  The contribution of ammonia, metals,
    and nonpolar organic compounds to the toxicity
    of sediment interstitial water from an Illinois
    River tributary.  Environ.  Toxicol. Chem. In
    press.                        .         •
  Shults, D.W., L.M. Smith, S.P. Ferraro, FA. Rob-
    erts,  and C.K. Poindexter.  1991. A comparison
    of methods for measuring trace organic com-
    pounds and metals in interstitial water. Water
    Res.  In press.
  Sly, P.G. 1988:  Interstitial  water quality of lake
    trout spawning habitat.  J.  Great Lakes Res.
     14:301-315.
  Stumm, W., and J.J. Morgan.  1981.  Aquatic
     chemistry - An introduction emphasizing chemi-
     cal equilibria in natural waters. John Wiley and
     Sons, New York.'. 583 pp.
                                                                                             5-23

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  Sediment Classification Methods Compendium
 Swartz, R.C., G.R. Ditsworth, D.W., Schults, and
   J.O. Lamberson. 1985.- Sediment toxicity to a
   marine infaunal  amphipbd:  Cadmium and  its
   interaction with sewage sludge.  Mar. Environ.
   Res. 18:133-153.
 Swartz, R.C., P.F.  Kemp, D.W. Schults, and J.O.
   Lamberson.  1988. Effects of mixtures of sedi-
   ment contaminants on the marine infaunal
   amphipod Rhepoxynius abronius.    Environ.
   Toxicol. Chem. 7:1013-1020.
 Swartz, R.C.,  P.P. Kemp, D.W.  Schults, G.R.
   Ditsworth, and R.J. Ozretich.   1989.  Acute
   toxicity of sediment from Eagle Harbor, Wash-
   ington, to the infaunal  amphipod Rhepoxynius
   abronius.  Environ. Toxicol. diem.  8:215-222.
 Swartz, R.C., D.W. Schults, T.H. DeWitt, G.R.
   Ditsworth, and J.O. Lamberson.  1990, Toxi-c-
   ity of fluoranthene in sediment to marine amph-
   ipods:  A test of the equilibrium partitioning
   approach to sediment quality criteria. Environ.
   Toxicol. Chem. 9:1071-1080.
 USEPA.   1985a.  Methods for  measuring the
   acute  toxicity  of effluents to freshwater  and
   marine organisms.   EPA/600/485-013.   U.S.
   Environmental  Protection Agency, Cincinnati,
   OH.
 USEPA. 1985b.  Short-term methods for estimat-
   ing the chronic toxicity of effluents and receiv-
   ing   waters   to   freshwater   organisms.
   EPA/600/4-85-014.   U.S. Environmental Pro-
   tection Agency, Cincinnati, OH.
 USEPA. 1985c.  Ambient water quality criteria
   for ammonia - 1984. EPA/440/5-85-001. U.S.
   Environmental Protection Agency, Duluth, MN.
 USEPA.  1988.  Methods for aquatic  toxicity
   identification evaluations:   Phase I  toxicity
   characterization procedures. EPA/600-3-88/034.
   U.S. Environmental Protection Agency, Duluth,
   MN.                          -   .  '•
 USEPA.    1989a.    Equilibrium  partitioning
   approach to generating sediment quality criteria.
   EPA/440/5-89/002.     U.S.   Environmental
   Protection Agency, Washington, DC.
 USEPA.  1989b.  Methods for aquatic  toxicity
   identification evaluations:   Phase II  toxicity
   identification procedures.  EPA/600-3-88/035.
   U.S. .Environmental1 Protection Agency, Duluth,
   MN.         '
 USEPA.  1989c.  Methods for aquatic toxicity
,   identification evaluations:  Phase  III  toxicity
   confirmation procedures.  EPA/600-3-88/036,
   U.S. Environmental Protection Agency, Duluth,
   MN.              .      ••                  :
 USEPA.   1991a.  Methods for aquatic  toxicity
   identification evaluations:   Phase I  toxicity
   characterization  procedures.   Second  edition.
   EPA-600/6-91/003.  Environmental Research
   Laboratory, Duluth, MN.
 USEPA.  1991b. Methods for sediment toxicity
   identification  evaluations.  National Effluent
   Toxicity Assessment Center Rep. No. 08-91
   Environmental Research Laboratory,  Duluth,
   MN.                           ,  •   .'
 Wells,  M.LM., and J.L.   Michael.    1987,.
   Reversed-phase solid-phase extraction for aque-
   ous environmental sample preparation in herbi-
   cide residue analysis.   J. Chromatogr.  Sci.
   25:345-50.
5-14

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             CHAPTER 6
              i  .     ...    .              ~        ..-''..
  Equilibrium Partitioning Approach

  Christopher S. Zarba
  U.S. Environmental Protection Agency
  401 M Street, SW (WH-586), Washington, DC 20460
  (202)260-1326
   ,   The equilibrium partitioning (EqP) approach
   focuses on predicting the chemical interaction
   among sediments,  interstitial water (i.e.,  the
   water between sediment particles), and contami-
   nants.   Based on correlations with toxicity,
   interstitial water concentrations of contaminants
   appear to be  better predictors of  biological
   effects than  do bulk sediment concentrations.
   The EqP method for generating sediment quality
   criteria is based on predicted contaminant con-
   centrations  in interstitial water.   Chemically
   contaminated sediments are  expected to cause
   adverse biological effects if the predicted inter-
   stitial water^concentration for a given contami-
   nant exceeds the chronic water quality criterion
   for that contaminant.
 i  6.1 SPECIFIC APPLICATIONS

       Specific applications of EqP-based sediment
   quality criteria are under development.   The
   primary use of EqP-based sediment criteria will
   be to identify and prevent risks associated with
   contaminants.  Because the regulatory needs
   vary widely among and within U.S. EPA offices
   and  programs,  EqP-based   sediment  quality
   criteria will be used in a variety of ways.
'  .  -  - EqP-based  numerical  sediment  quality
   criteria would likely be used directly to assess
   risk and would be applied in a tiered approach.
   In tiered applications, concentrations of sediment
    contaminants that exceed   sediment  quality
    criteria would be considered as causing unac-
    ceptable impacts. Further testing may or may
    not be required, depending on site-specific and
    program-specific conditions. Sediment contami-
    nants at concentrations less  than the sediment
    criteria  would not be,of concern.  However,
    sediments would hot be considered safe in cases
where they are suspected to contain other con-
taminants at concentrations above  safe levels,
but for which no sediment criteria exist.
    Synergistic, antagonistic, or additive effects
of multiple contaminants in the sediments may
also be of concern.  Additional testing in other
tiers of the evaluation approach, such as bio-
assays, could be required to determine whether
the sediment is safe.   It is likely tiiat such
testing would incorporate site-specific consider-
ations.         '

6.1.1  Current Use

    Specific regulatory uses of EqP-based sedi-
ment quality criteria are under development and
will be articulated in the Contaminated Sediment
Management Strategy.  The Science Advisory
Board (SAB) has completed the review  of this
approach for nonionic organic contaminants.
Based on the findings of this review, the method
will be used for developing national sediment
 quality criteria.  (The first five sediment  quality
 criteria will be proposed in the Federal Register
 shortly for public  comment.)  At the present
 time, the criteria are for the protection of ben-
 thic organisms.  The methodology for develop-
 ing sediment criteria for metal contaminants will
 be presented to the SAB for review in 1993.
The range of potential applications of the EqP
 approach is large because the approach accounts
 for contaminant bioavailability and can be used
 to evaluate most sediments.
     Draft  sediment criteria values have been
 developed for a variety of organic compounds
 using the EqP approach.  In pilot studies  at a
 variety of contaminated sediment sites at which
 site., characterization and evaluation activities
 were undertaken, the draft criteria were used in
 the following ways:

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 Sediment Classification Methods Compendium
    •   Identify extent of contamination;

    •   Assess the risks or potential risks associ-
        ated with the sediment contamination;

    •   Identify responsible parties and the need
        for source controls; and

    •   Identify the environmental benefit associ-
        ated with a variety of remedial options.

    In addition, a number of states have used draft
EqP-based sediment criteria to evaluate the poten-
tial effects  of sediment contaminants  found in
aquatic habitats.
6.1.2   Potential Use

    Potential applications of  the EqP  approach
include a variety of ongoing activities conducted
by the U.S. EPA.  EqP-based sediment  quality
criteria  could  play a major role in the identifi-
cation, monitoring, and cleanup of contaminated
sediment sites on a national basis. This is  true, in
part, because EqP-based SQC establish a direct
cause-and-effect relationship between a contami-
nant concentration and biological impacts. They
could also be used to ensure that uncontaminated
sites remain uncontaminated. In some cases, such
sediment criteria alone will be sufficient to iden-
tify and establish cleanup levels for contaminated
sediments.  In other cases, it will be necessary to
supplement the sediment criteria with biological
sampling, testing, or other types of analysis before
a decision can be made.
    EqP-based sediment criteria will be particular-
ly valuable at sites where sediment contaminant
concentrations are gradually increasing. In such
cases, criteria will permit an assessment of the
extent to which unacceptable contaminant concen-
trations are being approached or have  been ex-
ceeded.  Comparisons of field measurements to
sediment criteria will be a reliable method for
providing an early warning of a potential problem.
Such an early warning would provide an opportu-
nity  to  take  corrective action before adverse
impacts occur.
     Although sediment criteria developed using
 the EqP approach are similar in many ways to
 existing water quality criteria, their applications
 may differ substantially.  In most cases, contami-
 nants in the water column need only be controlled
 at the source to eliminate unacceptable adverse
 impacts.  In  contrast, contaminated  sediments
 often have been in place for quite some time, and
 controlling the source of  that pollution  (if the
 source still exists) will not be sufficient to allevi-
 ate the problem.   Safe removal, treatment,  or
 disposal of contaminated sediments can also be
 difficult and expensive.   For this reason,  it is
 anticipated that EqP-based sediment criteria will
 rarely be used  as  mandatory  cleanup  levels.
 Rather, they will  likely be used to predict  or
 identify the degree and spatial extent of problems
 associated with contaminated areas, and thereby
 facilitate regulatory decisions.
 6.2 DESCRIPTION

 6.2.1   Description of Method

     Concentrations of contaminants in the intersti-
 tial water correlate very closely  with toxicity,
 whereas concentrations of contaminants bound to
 the sediment particles do not. The EqP method
 for generating sediment criteria involves predicting
 contaminant concentrations in the interstitial water
 and comparing those concentrations to quality
 criteria. If the predicted sediment interstitial water
 concentration for a given contaminant, exceeds its
 respective chronic water quality criterion, then the
 sediment would be  expected to cause adverse
 effects.
     The processes that govern the partitioning of
 chemical contaminants among sediments, inter-
• stitial water, and biota are better understood for
 some  kinds  of chemicals than for others.  Con-
 centrations of sulfjdes and organic carbon have
 been identified as primary  factors that control
 phase associations, and therefore bioavailability,
 of trace metals  in sediments.  However, models
 that can use these factors to predict research are
 not fully developed.  Mechanisms that control
 the partitioning of polar organic compounds are
6-2

-------
                                                                              6—EgP Approach
 also poorly understood. Polar organic contami-
 nants, however, are not,generally considered to
 be a significant problem in sediments.  Parti-
 tioning of nonionic organic compounds between
 sediments and interstitial water is highly corre-
 lated with the  organic carbon content of sedi-
 ments.  Also, the toxicity of nonionic organic
 contaminants in sediments is highly dependent
 on their interstitial water concentrations. Conse-
 quently, to  date, the EqP approach is well
 developed for nonionic  organic contaminants
 and is in the process of development  for trace
 metals.
     Interstitial   water  concentrations  can  be
 calculated using partition coefficients for speci-
 fic nonionic organic chemicals and criteria con-
 tinuous concentrations from WQC documents.
 The sediment  quality criterion for a specific
 chemical is defined as the solid phase concentra-
 tion that will result in an uncomplexed intersti-
' tial water concentration equal  to  the chronic
 water quality criterion for that  chemical.  The
 rationale for using water quality criteria as the
 effect concentrations for  benthic organisms is
 that the sensitivity range for benthic organisms
 appears to be similar to the sensitivity  range for
 water column organisms.   Moreover, partition
  coefficients for a wide variety of contaminants
 are available.
     The  calculation procedure for  nonionic
  organic contaminants is as follows:
                rSQC=KpxcWQC
  where:         .
          cWQC = Criterion continuous concen-
                    tration
          rSQC =   Sediment  quality  criterion
                    (fig/kg sediment)
          Kp    =   Partition coefficient  for the
                    chemical  (L/kg  sediment)
       .             between sediment and water.


  Although the method for  developing sediment
  criteria for nonionic organic contaminants has
  been identified,  continuous refinement of the
  methodology is expected.    .»
6.2.1.1 Objectives and Assumptions

  , Three principal assumptions underlie use of
the EqP-based approach  to  establish sediment
quality criteria:

    •  For sediment-dwelling  organisms, the
       uncomplexed Mterstitial water concentra-
       tion of a chemical correlates with ob-
       served biological effects across sediment
       types, and the concentration  at  which
       effects are observed is the same as that
       observed in a water-only exposure.

    •  Partitioning models permit calculation of
       uncomplexed interstitial water concentra-
       tions of the chemical phases of sediments
       controlling availability.

    •   Benthic  organisms exhibit a range of
        sensitivities to chemicals that is similar to
        the range of sensitivities  exhibited by
        water column organisms.

 Data exist supporting each of these assumptions.

 6.2.1.2 Level of Effort

 6.2.1.2.1   Type of Sampling Required

    Sufficient sediment  chemistry  sampling is
 required to adequately characterize the area of
 concern. Total organic carbon concentrations are
 also needed, preferably for each sampling station.

 6.2.1.2.3   Types of Data Required

    Analyses are needed to determine the concen-
 trations  of the  contaminants of concern  in the
 sediment (bulk sediment analysis) and the concen-
 trations  of organic carbon hi the sediment

 6.2.12.4  Necessary Hardware and Skills

     The investigator must be  able to design an
 appropriate sampling study, conduct bulk sediment
 analyses, operate a pocket calculator, and under-
 stand developed values and what they protect.
                                                                                              6-3

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 Sediment Classification Methods Compendium
 6.2.1.3 Adequacy of Documentation

     The method is very well documented (see
 Section 6.5).
 6.2.2   Applicability of Method to Human
        Health, Aquatic Life, or Wildlife
        Protection

    At the present' time  SQC do  not address
 bioaccumulative impacts to aquatic life, wildlife,
 and human health.  Efforts are under way to
 derive criteria protective of these endpoints.


 6,23   Ability of Method to Generate
        Numerical Criteria for Specific
        Chemicals

    The EqP method  generates numerical criteria
 for a number of nonionic organic chemicals.  A
 methodology for developing sediment criteria for
 metal  contaminants is being developed.   Draft
 criteria to be proposed in the Federal Register
 were developed for endrin, phenanthrene, fluor-
 anthene, dieldrin, and acenaphthene.  It is expect-
 ed that three to five additional sediment criteria
 will be issued each subsequent year.
    Methods for developing sediment criteria for
 metal contaminants are under development and are
 expected to be reviewed by the SAB in 1993.
63 USEFULNESS

63.1   Environmental Applicability

    One of the principal reasons for selecting the
EqP approach is that it is applicable in a wide
variety of aquatic systems, which is a prerequisite
for the development of national sediment quality
criteria.

5.3.1.1 Suitability for Different Sediment Types  .

    Although aspects of the EqP method are still
under development, it is expected that sediment
  criteria for nonionic contaminants developed using
  this approach will be applicable to all types of
  sediments found in  both freshwater and marine
  environments with organic carbon concentrations
  aO.2 percent organic carbon. Additional work is
  needed to clarify the best use of the EqP approach
  for sediments with less than 0.2 percent organic
  carbon.

  6.3.1.2 Suitability for Different Chemicals* or
         Classes of Chemicals

     The EqP method for developing sediment
  criteria has been modified for different types of
  contaminants. Nonionic, ionic, and metal contam-
  inants all interact  with  sediment  particles in
  different ways, and partitioning models have to be
  modified to account for these differences.   The
  technical approach for developing sediment cri-
  teria for nonionic organic contaminants has been
  well developed  and  is under peer review.   The
  technical approach for developing sediment cri-
  teria for metal contaminants is under development
  and is expected  to undergo peer review in 1993.
  Ionic contaminants are not believed to cause major
  problems hi sediments, but work plans for sedi-
  ment criteria development methods for  these
  compounds have been written.

  6.3.1.3 Suitability for Predicting Effects on
        Different Organisms

    As indicated above (see Section 6.2.1), the
 EqP  approach is based on  predicted interstitial
 water concentrations of nonionic organic  con-
 taminants, and comparisons of these concentra-
 tions with chronic water quality criteria. Typi-
 cally, water quality criteria are based on toxicity
 information (e.g., median lethal or median effec-
 tive concentrations) for a wide number of species
 and are set low enough to be protective of at least
 95 percent of the species tested.  Consequently,
- exposure levels that are  predicted using the EqP
 approach can be compared with a range of toxic
 effects values that are representative of the differ-
 ent kinds pf organisms  on which  water quality
 criteria are based.
6-4

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                                                                              6-r-EgP Approach
6.3.1.4 Suitability for In-Place Pollutant
       Control

    The  EqP method is suitable for in-place
pollution control because  it  can be used  to
identify  locations where concentrations of indi-
vidual contaminants are causing adverse effects.
Target cleanup levels can be identified, and the
success of cleanup activities can be determined.

6.3.1.5 Suitability for Source Control

    The  EqP  method  is  suitable for  source
control.  This method predicts the concentration
of a contaminant above which adverse impacts
are  likely.   A  direct  measure of  biological
effects is not needed to identify safe levels.

6.3.1.6 Suitability for Disposal Applications

    The EqP method is  suitable for  predicting
the  effects  that  contaminated  sediments  may
have if moved  to an  aquatic site.   It is not
applicable to  contaminated sediments that are
disposed of at upland sites.
 6.3.2  General Advantages and Limitations

     The  EqP  approach  offers  the following
 advantages:

     •  It is consistent with existing water qual-
        ity criteria;

     •  It establishes a cause-and-effect relation-
        ship;

     •  It  relates risks to  specific substances,
        and it  can be used to identify  probable
        species at risk;

     *  It is applicable across all types of sedi-
        ments  and in all types of aquatic envi-
        ronments, including lentic, lotic, marine,
        and estuarine environments;

     •   Only site-specific chemistry data are needed;
   •   Site-specific or station-specific sedunent
       criteria can be calculated as soon as sedi-
       ment chemistry data are available;

   •   It incorporates the large quantities of data
       that were used in  the development of
       water quality criteria;

 " •   It can be incorporated into existing regu-
       latory mechanisms with little or no need
   .    for additional staffing or skills;

   •   The  equilibrium partitioning  theory on
       which it is based is well developed;

   •   It can be modified easily to accommodate
       site-specific circumstances;

   •  It can be used with additional  develop-
       ment to identify risks to  humans  and
       wildlife that may occur as a result of
       bioaccumulation; and  .

   •  It identifies the degree of sediment  con-
x      lamination and permits an  assessment of
       whether contaminant  concentrations are
       approaching an effects level.

The EqP approach  is limited in  the following
ways:                                   .     ^

    •  Sediment criteria developed using this ap-
       proach do not address possible  synergis-
       tic,  antagonistic, or  additive effects  of
       contaminants;

    •  Interim and draft sediment criteria pres-
       ently exist  for only 12  contaminants at
       this time;

    •  The  technical approach for  developing
        sediment criteria for metal contaminants is
        still under development;

    •   Sediment quality  criteria  for  nonionic
        chemicals apply to sediments that have an
        organic carbon concentration iO .2 percent;
        and
                                                                                              6-5

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 Sediment Classification Methods Compendium
     "  Sufficient water-only toxidty data do not
        exist for all contaminants of concern.

 6.32.1 Ease of Use

     The calculation of site-specific sediment criteria
 is relatively easy, provided that sediment chemistry
 data adequately characterizing the site, a partition
 coefficient, and water quality criteria protective of
 the desired organism are available.

 632.2 Relative Cost

     Because site-specific biological data are not
 needed,  the costs associated with this method
 depend primarily  on the  cost of collecting site-
 specific chemistry data.

 6,3.23 Tendency to Be Conservative

     Sediment criteria  are derived using the chronic
 water  quality criteria  as  effect levels.  Hence, the
 levels of protection afforded by sediment criteria are
 similar to those of water quality criteria.  In general,
 water quality criteria are deemed to be protective of
 95 percent of the organisms most of the time. Each
 SQC is bracketed with levels of uncertainty.

 6.3.2.4 Level of Acceptance

    The  EqP approach  and  its use in  deriving
 sediment quality criteria are the result of the efforts
 of many scientists who represent a variety of federal
 agencies,  industries, environmental  organizations,
 universities, U.S. EPA laboratories, state agencies,
 and  other institutions.  These scientists were in-
 volved in the selection of the EqP approach for
 generating sediment criteria and have also  played a
 role  in development of the method. Papers that
 discuss various  aspects  of this effort have been
 presented at scientific conferences.

 6.3.2.5 Ability to Be Implemented by Laboratories
       with Typical Equipment and Handling   •
       Facilities
           »

    No special laboratory facilities or requirements
 are needed.  Sediment chemistry analysis is all that
 is required.
63.2.6 Level of Effort Required to Generate
        Results

    The necessary level of effort varies substan-
tially from site to site and is dependent on many
factors.  Compared with other methods, the EqP
method generates results quickly and more cost-
effectively. No site-specific biological data are re-
quired.    '

6.3.2.7 Degree to Which Results Lend
        Themselves'to Interpretation

    All sediment evaluation  procedures  require
some level of interpretation. However, a sediment
criterion that  is bracketed with an  appropriate
degree of uncertainty can provide pertinent infor-
mation.  For example,  sediment chemistry data
that identify concentrations below the conservative
effect level for a particular contaminant could be
deemed safe for that contaminant. A contaminant
concentration above the upper uncertainty level
could be identified immediately as contaminated,
and some degree of contamination  could be
assigned to  those sediments for the individual
contaminant.  Sediments whose  concentration of
a particular contaminant falls within the degrees of
uncertainty could require more detailed interpreta-
tion and possibly additional testing.

6.3.2.8 Degree of Environmental Applicability

    EqP-based sediment quality  criteria can be
applied directly  to any contaminated sediment
containing &0.2 percent  organic carbon and non-
ionic chemicals for which  criteria are available.
Extensive data 'analysis and site-specific biological
data are not required to  use sediment criteria
developed  using this  method.   (In some cases
these attributes may nonetheless be desirable.) As
a result, the EqP method can be considered envi-
ronmentally applicable in some cases.  Because a
wide variety of contaminated sediment sites exist,
absolute   statements   regarding  environmental
applicability are difficult to make. However, the
EqP method would be appropriate in many situa-
tions to predict bioavailability, bioaccumulation,
and biological effects.
6-6

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                                                                             6—EqP Approach
6.3.2.9 Degree of Accuracy and Precision

    Each sediment criterion value developed using
the EqP method will have an associated degree of
uncertainty, which  will  vary from  criterion to
criterion.  The principal uncertainties associated
with sediment criteria developed using the EqP
method are those associated with partition, coeffi-
cients.  Hence, each developed sediment criterion
should  be and is bracketed with  uncertainty,
thereby providing decision-makers with a greater
understanding of the meaning of the developed
values.                -                   .
 6.4 STATUS

     The method for developing sediment criteria
 for nonionic  organic contaminants  has  been
 developed and has been reviewed by the SAB on
 two separate occasions.  Guidelines and guidance
 on the  regulatbry use of  sediment criteria are
 under development.  The method for developing
 sediment criteria for metal  contaminants is being
 investigated and results are promising. The metals
 method is expected to be sufficiently well devel-
 oped for peer review by 1993.


 6.4.1   Extent of Use

     Specific regulatory uses for EqP-based sedi-
 ment quality  criteria  are  being developed.   A
 formal framework for the application of sediment
 criteria is not expected until EPA  completes  its
 effort to develop a  contaminated sediment man-
 agement strategy.  The range of potential applica-
 tions is very large because  the need  for evaluating
 potentially contaminated sediments arises in many
  contexts.

      Interim sediment criteria values were devel-
  oped for a variety of organic compounds.  These
  values  were used in a pilot, study at a number of
  sites where  site characterization and evaluation
  activities were conducted.  The interim criteria
  were used in three ways:
   •   To identify the extent of contamination
       and responsible parties;                 '  ,

   •   To assess the risks associated with sedi-
       ment contamination; and

 • •   To  identify the environmental  benefits
       associated with a  variety  of  remedial
       options.    ,

   . A number of States have used interim  and
draft sediment criteria to  evaluate the potential
effects of several contaminants found in sediments
in state waters.  The methodologies for deriving
sediment criteria  have been used in a variety of
situations including  the  evaluation  of dredged
material, Superfund  site  assessments,  and the
identification of  appropriate  cleanup levels for
contaminated sediment sites.       - ,        .


6.4.2   Extent to Which Approach Has Been
       Field-Validated

    Considerable effort has been made by EPA to
use field sites as part of the  criteria  validation
effort and to aid in designing regulatory programs.
Table 6-1 lists ongoing and completed studies
where SQC are being used to directly support
sediment activities.   In  addition to these  sites*
there are other sites and situations (completed,
 ongoing, and planned) where the  EqP is, being
 applied to field situations.  Although these efforts
 are not involved with criteria development efforts,
 they  do provide valuable data on the appropriate-
 ness  of the EqP.
    It needs to be understood, however, that  "field
 validation" does not describe a specific experimen-
 tal protocol.   The idea  is to  find a site that is
 contaminated with a single  chemical  and deter-
 mine whether the benthic populations are degraded
 when the SQC is exceeded.  However, there are
 practical difficulties.  Such a field site contamin-
 ated with only one  chemical must be found, and
 there can be no ongoing sources of the chemical
  since the exposure should be only from the sedi-
  ment.  A gradient of chemical concentration that
  spans the SQC concentration is necessary.   The
                                                                                               6-7

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 Sediment Classification Methods Compendium
                        Table 6-1. Ongoing and Completed Studies Using SQC.
Location
Hurtsvilto, AL
KeweenawLIc
Steilacoom Lk
Fox River
Fox River
Foundry Cove
Calumet River
Nationwide
New Bedford Harbor
Narragansett Bay
Colonization ExpL
Colonization Expt
San Diego Bay
Lauritzen Canal
Nationwide
Nationwide
Chemical
DDT/DDD/DDE
Cu
Cu •
PCS bfoaccumulation
Metal bioaccumulatJon
Cd,Ni
Sediment partitioning
Comparison of toxicity test and benthic community disnjp-
tiontoSQC
BioaccumJaJion
Bioaccumulaiion
6 chemicals
3 chemicals to test SQC
PAHs ' -
DDT
Comparison of SQC chemicals to STORET
Comparison of SQC chemicals to NOAA National Status
and Trends data
Status
Ongoing
Submitted for publication
Submitted for publication
Submitted for publication
In preparation
Published
In preparation
Ongoing
Published
Published
Published
Ongoing
Ongoing
Ongoing
In the documents
In the documents
sediment type must be essentially uniform in the
gradient so that only chemical concentration  is
changing.  The benthic population must be plenti-
ful enough so that population degradation can be
observed as the SQC is exceeded. In spite of the
difficulties, major field efforts are presently under
way.
    An intermediate  level of field validation  is
provided by the benthic colonization experiments.
The experimental design is described above.  The
populations that develop are determined entirely
by natural recruitment.  The uniformity of sedi-
ment type is  guaranteed by  the experimental
'design. The experiments last from 2 to 4 months
so  that the sediment can properly be called  a
"natural" sediment.  Three benthic colonization
experiments have been performed using spiked
sediments.  The data analysis, which is partially
complete,  indicates that  the experiments  are
consistent with the SQC for the chemicals being
tested.
    A third type of field validation is proceeding
as well. It is based on the- notion that although it
is not possible to prove  the validity of SQC
(continual accumulation of evidence in favor of its
validity does not guarantee that all evidence will
always be supportive),  it is possible to prove that
it is invalid.  If sediments are collected and the
state of the benthic population is evaluated relative
to control sites from  the same region,  there are
6-8

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                                                                            6—EcjP Approach
                           Table 6-2. SQC Field Validation Truth Table.
             Bmrthie Impact
                                         SQC Not ExcMded

                                          Other Chemicals
                                                                        SQCExcMcM
                        Consistent
            No Benthic Impact
                                            Consistent
                                                                         Invalidates
four possibilities, which are arranged as a truth
table in Table 6-2.
    The correlation of the presence or lack of
benthic impact with exceeding or not exceeding
the SQC is consistent but not proof of causality.
The observation of benthic impact where the SQC
is not exceeded can be attributed to the impact of
other chemicals. However, if the SQC is exceed-
ed, with a proper accounting for the uncertainty of
SQC, and no benthic impact is observed, then the.
SQC is invalidated. The collection of these data
is an ongoing part of the SQC development effort
Analysis to date suggests that these data do not
invalidate the SQC.

6.43  Reasons for Limited Use

    The  EqP method is not commonly used for
the following reasons:

     • The method was developed only recently,
        and sufficient time has not elapsed for the
        principles to be understood and used by
        others.

     •  Final criteria have not been issued.

     •  Guidance  and  technical  support  docu-
        ments are in draft form and will be issued
        along with final criteria.

 6.4.4   Outlook for Future Use and Amount
        of Development Needed

     This method is the only procedure for deriva-
 tion of  sediment quality criteria  that is  generic
across sediments, accounts for bioavailability of
chemicals, and relates effects to specific chemi-
cals.   Therefore, EqP-based  sediment  quality
criteria will be used much as water quality criteria
are being used to define environmentally accept-
able concentrations.   Sediment quality criteria,
along with sediment toxicity tests analogous to
water quality criteria and whole-effluent toxicity
tests, will play a major role in EPA's management
of contaminated sediment.                .
6.5 REFERENCES

USEPA. April 1989. Brief ing 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.  •
USEPA. February 1990, Report of the Sediment
  Criteria Subcommittee of the Ecological Process-
  es and Effects  Committee - Evaluation of the
  equilibrium  partitioning approach for assessing
  sediment quality.  A Science Advisory Board
  Report.
 USEPA.  August 1991.  Analytical method for
  determination of acid volatile sulfide in sediment
  (final draft). Office of Science and Technology,
  Health and Ecological Criteria Division,
 USEPA.   August 1991.   Technical basis for
  establishing sediment quality criteria for non-
  ionic chemicals using equilibrium partitioning.
  Office of Science and Technology, Health and
  Ecological Criteria Division.
 USEPA.   November  1991.  Proposed sediment
  quality criteria for  the  protection  of benthic
                                                                                             6-9

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 Sediment Classification Methods Compendium
  organisms:  Acenapththene (draft).   Office of
  Science and Technology, Health and Ecological
  Criteria Division.
USEPA.   November  1991.   Sediment quality
  criteria for the protection of benthic organisms:
  Dieldrin (draft).  Office of Science and Technol-
  ogy, Health  and Ecological Criteria Division.
USEPA.    November 1991.  Sediment  quality
  criteria for the protection of benthic organisms:
  Endrin (draft). Office of Science and Technolo-
  gy, Health and Ecological Criteria Division.
USEPA.. ' November 19911   Sediment quality
  criteria for the protection of benthic organisms:
  Fluoranthene  (draft).  Office of Science and
  Technology, Health  and  Ecological  Criteria
  Division.
USEPA.   November  1991.  Sediment quality
  criteria for the protection of benthic organisms:
  Phenanthrene  (draft).  Office of Science and
  Technology, .Health  and  Ecological  Criteria
  Division.
6-10

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         CHAPTER 7
Tissue Residue  Approach


U.S.PEnvironmental Protection Agency, Environmental Research Lab-Duluth
6201 Congdon Boulevard., Duluth,  MN  5D804
(218) 720-5553, FTS 780-5553

Anthony R. Carlson                             .„     •-,.,  ^ n I.M,
U.S. Environmental Protection Agency, Environmental Research Lab-Duluth
6201 Congdon Boulevard., Duluth, MN  55804
(218) 720-5523, FTS 780-5523

US^nvironmental Protection Agency,  Environmental Research Lab-Newport
Marine Science Drive, Newport, OR  97365
-(503) 867-4042
     In  the tissue  residue  approach,  sediment
 chemical concentrations that will result in accept-
 able residues in exposed biotic tissues are deter-
 mined.   Concentrations of unacceptable tissue
 residues may be derived from toxicity  tests per-
 formed during generation of chronic water quality
 criteria, from  bioconcentration factors derived
 from the  literature or generated  by experimen-
 tation, or by comparison with human health risk
 criteria associated with consumption of contami-
 nated  aquatic organisms.  The tissue  residue
 approach generates numerical criteria and is most
 applicable for nonpolar organic and organometallic
 compounds.
  7.1  SPECIFIC APPLICATIONS

  7.1.1   Current Use

      Tissue residues of chemical contaminants in
  aquatic organisms, particularly fish, are frequently
  used as measures of water quality in both fresh-
  water and  marine systems.   The  tendency  of
  organisms to bioaccumulate chemicals from water
  and food is one of the factors used in establishing
  national  water quality criteria (WQC) for the
  protection of aquatic life (Stephan et at., 1985):
  Nonpolar organic  chemicals, which may  bio-
  accumulate to levels toxic to organisms or render
  organisms unfit for human food, generally will
also be found as sediment contaminants. Hydro-
phobic organic chemicals preferentially distribute
into  organic  carbon  in  sediment and lipid  in
aquatic biota.  The tissue residue approach has
been used recently to establish Ihe amount  of
reduction of  2,3,7,8-tetrachlorodibenzo-p-dioxin
(TCDD) concentration in Lake Ontario sediments
necessary to attain acceptable TCDD levels in fish
(Cook et al.,  1990).  The  acceptable sediment
TCDD concentration is being used as a sediment
criterion to determine the remedial action neces-
sary to reduce the incremental loading of TCDD
from the Hyde Park Superfund site to Lake Ontar-
io (Carey et al, 1989). Tissue residues of benthic
organisms have also been used in some regulatory
actions, such as the assessment of bibaceumulation
potential of dredged materials (USAGE, 1991).

7.1.2  Potential Use

    Although tissue residues have been used more
 commonly to determine the potential for bioaccu-
 mulation of chemical contaminants from sediments
 arid dredged materials, they also provide an excel-
 lent measure of "effective exposure dose": a mea-
 sure of an organism's actual exposure over time to
 a pollutant of concern. This exposure measure may
 be related to  the dose expected at the water quality
 criterion or  related directly to the potential for
 producing chronic toxic effects. Given the ability to
 measure or predict tissue residues resulting from

-------
 Sediment Classification Methods Compendium
 exposures in contaminated sediment systems, it is
 possible  to establish  sediment criteria  based on
 residue-toxicity effects relationships.   These rela-
 tionships can provide  a basis for sediment criteria
 that are free of uncertainties normally associated
 with organism  exposures and sediment contaminant
 bioavailability.  This is especially true when in situ
 measurements  provide the basis for the sediment
 residue link to the residue-toxic effect relationship.
     One  example  of  tissue  residue-toxic effects
 linkage is the  relationship between the  failure of
 Great Lakes lake trout (Salvelinus namaycush) to
 reproduce and bioaccumulation  of  TCDD  and
 non-ortho substituted PCBs (Mac, 1988).  Labora-
 tory studies have shown significant  mortality of
 larvae when lake trout ova contain as little as 50 ppt
 2^,7,8-TCDD  (Cook  et al., 1990; Walker et al.,
 1991).  This residue level is found in Lake Ontario
 lake trout that have not successfully reproduced
 naturally  for many years.  On the basis of TCDD
 toxic equivalents  for  organochlorine components
 having the same mode of toxic action, residues in
 lake trout from Lakes Ontario and Michigan may
 provide a measure of the  reduction  in sediment
 contamination necessary to reduce fish tissue con-
 centrations to a threshold presumed to allow repro-
 duction.  The  same  approach can be  used for
 benthic organisms, which may have greater intersite
 variations in residue levels than do fish because of
 benthic   organisms'   closer  association   with
 sediments.
7.2 DESCRIPTION

7.2.1    Description of Method

    The tissue residue approach involves the estab-
lishment of safe sediment concentrations for individ-
ual chemicals or classes  of chemicals by deter-
mining the sediment chemical concentration that will
result in acceptable tissue residues.  This process
involves two steps:  (1) linking toxic effects to resi-
dues (dose-response relationships) and (2) linking
chemical residues in specific organisms to sediment
chemical  contamination concentrations  (exposure
relationships).    Methods  to derive unacceptable
tissue residues include at least three approaches:
     •   The   water   quality   criterion-residue
         approach;

     •   The experimental approach; and

     •   The human health approach.

 Each of these approaches is described briefly below.

 Water Quality Criterion-Residue Approach—A
 rapid approach for determining acceptable concen-
 trations of tissue  residues  involves  establishing
 maximum   permissible   tissue    concentrations
 (MPTCs)  expected for organisms at  the chronic
 water  quality criterion  concentration  previously
 established for a specific pollutant.  MPTCs, when
 not available through residue measurements obtained
 with toxicity tests used for water quality criteria, can
 be obtained by multiplying the water quality criteri-
 on by an appropriate bioconcentration factor (BCF)
 obtained from  the literature.   When a reliable
 empirical BCF is not available, the BCF may be
 predicted from an octanol-water partition coefficient
 or a bioconcentration kinetic  model.   Thus, the
 absence of a water quality criterion for a  chemical
 does not eliminate this approach as long as appropri-
' ate chronic toxicity test data are available for the
 species of interest.

 Experimental  Approach—Tissue  residue-toxic
 effects linkages can be established through a series
 of chronic dose-response  experiments  or  field
 correlations. Although this approach  has the advan-
 tage of directly determining the tissue residue-toxic
 effects linkages, it can be relatively time consuming
 and costly to implement for  a large  number of
 pollutants.   The experimental approach should be
 used to  test the assumptions of the water quality
 criterion-residue approach and to  supplement the
 existing tissue residue-toxic effects database.   The
 experimental work can be closely coupled with the
 experiments conducted under  die bulk sediment
 toxicity test approach to deriving sediment quality
 criteria (see Chapter 3, Bulk Sediment Toxicity Test
Approach).

 Human Health  Approach—Human  health  risk
 from consumption  of  freshwater fish  or seafood
7-2

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                                                                   7—Tissue Residue Approach
may be used as the criterion  for tissue residue
acceptability. A sediment quality criterion for a
specific compound can be derived by establishing
an acceptable human risk level  (e.g., an excess
human cancer risk of IxlO'5) and then back-calcu-
lating to the sediment concentration that would
result in tissue residues associated with this level
of risk. The human health approach can generate
sediment  quality criteria lower for carcinogenic
compounds (e.g., PCBs, dioxins, benzo(a)pyrene)
than those criteria derived from ecological end-
points.
     The choice of method to determine a quantita-
tive tissue residue-sediment contamination level
relationship  depends on the specific pollutants,
 organisms, arid water systems  of concern, as well
 as the regulatory approach (e.g., remedial action,
 wasteload  allocation, Superfund  enforcement).
 The linkage between organism residue and sedi-
 ment chemical concentration  can be made from
 site-specific measurements of sediment-organism
 partition coefficients (Kuehl etal, 1987); fugacity
 or  equilibrium partitioning model (Clark  et al,
 1988); predictions of organism residues; or pharm-
 acokinetic-bioenergetic  model  predictions  of
 organism residues that  result from uptake from
 food   chain,   water,   and  sediment   contact
 (Thomann, 1989).  The residue approach works
 best for aquatic ecosystems that are at or close to
  steady state with respect to the  distribution  of
 chemicals between biotic and abiotic components.
  Steady-state conditions are  common for most
  sediment contaminants because of their persistence
  and  tendency  to  exert  long-term  rather  than
  episodic bioaccumulation and toxic effects.

  7.2.1.1  Objectives and Assumptions

      The objective of this approach is to generate
  numerical sediment quality criteria based  on
  acceptable levels of chemical contaminants in
  sediment-exposed  biota.   This objective is
  parallel to that  of the water quality criteria,
  except that organism residues provide measures
  of exposure to chemical contaminants rather than
  water concentrations of contaminants.  By using
   tissue  residues  rather  than  interstitial water
   concentrations to measure, dose, as in the equi-
librium partitioning approach (Chapter 5), this
method does not require that the organism be at
thennqdynamic equilibrium with respect to the
sediment contamination level. The site-specific
residue approach is powerful because it does not
require knowledge of bioavailability relation-
ships  for each  organism in the  system.   All
interaction  pathways between  sediment  and
organisms are incorporated in the determination
of organism-to-sediment contamination ratios.
These can be expressed on the basis of sediment
organic carbon-organism lipid for hydrophobic
organic chemicals.- It is assumed that reduction
in sediment  contaminant concentrations over
time (e.g., as a result of remedial actions, waste-
 load allocations) will result in parallel reduction
 in exposure, aquatic organism  residues,, and,
 consequently, the potential for toxic effects.  It
 is further assumed that data on residue-to-toxici-
 ty relationships can be obtained from laboratory
 exposures of organisms when such data are not
 already available and that the route of exposure
 responsible for residue accumulation does not
 influence the residue-toxicity relationships.

 7.2.1.2 Level of Effort

     Relatively  little effort would be required to
 generate  preliminary  sediment  quality  criteria
 using MPTCs calculated  from  existing .water
 quality criteria and BCFs.  In  the  absence of
  appropriate water quality  criteria or BCFs,  the
  level of  effort depends on the availability of
  tissue residue  action levels and the  complexity
  of the sediment contaminant mitigation approach
  to be used. Relatively little effort is required to
  determine the degree to which sediment contam-
  ination concentrations must be reduced  for
  single chemicals in well-rmixed systems where
  fish  residues  are uniformly unacceptable for
  human consumption.   Much  more  effort is
  required for systems having sediment contamina-
   tion  "hot spots" where resident aquatic organ-
   isms are eliminated or reduced in number due to
   a complex  mixture of sediment  contaminants.
   Another  complexity  that could increase the
   required level of effort is the presence of sedi-
   ment contaminants that are readily  metabolized
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  Sediment Classification Methods Compendium
 to chemicals of greater toxicity that are responsi-
 ble for the observed adverse effects.  In some
 cases, residue-toxic effects data would incorpo-
 rate the effects of toxic metabolites.

 7.2.1.2.1   Type of Sampling Required

     Surface sediment samples must be analyzed
 for chemical  contaminants  of interest.  Inter-
 stitial  water composition does not need to  be
 determined because the residues in biota are
 related to bulk sediment chemical composition.
 Sediment  characteristics such  as grain  size,
 organic carbon content,  and metal binding ca-
 pacity are useful for defining sediment-to-biota
 relationships for different sites within an ecosys-
 tem. Biota sampling for residue analysis should
 include sensitive organisms when toxic  effects
 are a concern or, in the absence of  sensitive
 organisms, organisms whose residues will serve
 as biomarkers for establishing safe sediment
 contaminant levels.

 7.2.1.2.2   Methods

    The tissue residue approach, as discussed in
 Section 7.2, depends on determining residues in
 aquatic organisms that are unacceptable  on the
 basis of toxicity to the organism or unsuitability
 for human or animal consumption as food.  The
 linkage of sediment contaminant concentrations
 to organism residues is possible through a num-
 ber   of  approaches   including   site-specific
 measurements,  equilibrium  partitioning-based
 predictions, and steady-state food chain models.
 The choice of a specific approach depends on
 the  chemical  of concern, the availability of
 residue-toxic effects  data, the  contamination
 history (in-place pollutant problem versus  a
 continuing or projected sediment contamination
 problem), and the characteristics of the impacted
 ecosystem.  The construction of comprehensive,
 systematic strategies for  all potential sediment
 contamination  assessments will  be achieved
 through further research and development.
    Toxicity  identification  evaluation  (TIE)
procedures  (see  Chapter  5)  complement  the
tissue-residue approach.  The TIE approach is
 especially useful if sediment assessment begins
 without knowledge of the sediment contaminants
 that are causing toxicity or unacceptable residues
 in biota.   The  absence  of benthic species  or
 failure of fish eggs to hatch may be attributable
 to acutely toxic, but non-residue-forming, chemi-
 cals (e.g., ammonia) in sediments.  TIE proce-
 dures can distinguish between potential metal,
 nonpolar organic, polar organic, and inorganic
 .chemical sources of toxicity in sediment pore
 waters or elutriates.  These procedures enable a
 more complete assessment of the significance of
 residue-associated tqxicity in the system.
     Once  potentially  toxic,  bipaccumulative
 contaminants are identified, either in sediment or
 in aquatic organisms associated through expo-
 sure to sediments, the toxicological significance
 of site-specific  sediment-to-biota  contaminant
 partition factors  can be assessed.  Conservative
 generic sediment quality criteria can be generat-
 ed from residue-toxicity relationships by assum-
 ing equilibrium partitioning between the binding
 fractions of organisms and sediments (e.g., lipid
 and sediment organic carbon for nonpolar organ-
 ic chemicals).

 7.2.1.2.3    Types of Data Required

    The tissue residue method requires identifi-
 cation of chemicals in the sediment  that are
 likely to be associated with chronic environmen-
 tal effects.  An indirect method for identifying
 such chemicals and their  locations is to screen
 aquatic organisms for residues as in the National
 Dioxin Study (USEPA, 1987b) or the National
 Study  of Chemical Residues in Fish (USEPA,
 1992),  sponsored by EPA's Office of Water
 Regulations and Standards.  When toxicity data
 are not available,  either laboratory dose-response
 experiments  or  quantitative structure-activity
predictions can be used to establish  the toxico-
logical significance of the tissue residues. Field
surveys that indicate the absence of sensitive
organisms in contaminated sediment areas are
useful for establishing sediment quality criteria,
especially if  interspecies  sensitivities to the
chemicals of concern are  known.  Tissue resi-
dues associated  with  no-effect and  lowest-
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                                                                  7—Tissue Residue Approach
 observable-effect  concentrations  are  needed
 when the sediment criterion is not based on a
, human health standard.

 7.2.1.2.4   Necessary Hardware arid Skills

     Sediment and tissue analyses require com-
 monly available chemical analytical capabilities.
 Some chemicals require advanced instrumental
 analytical techniques,  such as high resolution
 gas chromatography/mass spectrometry.

 7.2.1.3 Adequacy of Documentation

     The use of tissue residues to establish sedi-
 ment criteria on the basis of human health ef-
 fects associated with ingestion of  contaminated
 fish has been documented. Methods for using
 tissue residue-toxicity relationships to establish
 sediment criteria, although scientifically sound,
 have not been extensively documented.  The
 various methods for predicting tissue residues in
 benthos and fish have been well documented.

  7.2.2   Applicability of Method to Human
         Health, Aquatic Life, or  Wildlife
         Protection

     Tissue residue  measurements are directly
  applicable to human risk assessment  when the
  aquatic organism is used as  human food.  Be-
  cause of  this relationship, the tissue  residue
  method provides  a direct link between human
  health and sediment criteria development,  tis-
  sue residues for wildlife and aquatic organisms
  can be used to assess sediment  toxicity when
  there is an established exposure  linkage to the
  sediment;  The tissue residue approach is most
  advantageous for sediment  contaminants that
  adversely impact organisms such as fish that are
  not in direct contact with the sediment or  its
  interstitial water. The tissue residue approach is
  well  suited to evaluating sediment  quality  in
  systems that have aquatic food chain connections
  from benthos to  birds   experiencing eggshell
  thinning or genotoxic effects. The tissue residue
  concentration serves as a quantitative measure of
  sediment contaminant bioavailability, which may
differ as  a function  of ecosystem, sediment,
water, food chain, and species characteristics.

7.2.3   Ability of Method to Generate
       Numerical Criteria for Specific
       Chemicals

    The tissue residue approach can be used to
generate site-specific numerical criteria for non-;
polar organic chemicals such as PCDDs, PCDFs,
and PCBs.   Tissue residues of aldrin/dieldrin
(USEPA;  1980a)  and endrin (USEPA,  1980b)
have been used to establish water quality criteria
on the basis of human health risks. The DDT and
PCB water quality criteria  are based on  toxic
effects in birds and animals as a function of fish
residues (USEPA, 1980c, 1980d). Tissue residues
of organometallic chemicals  such as  methyl
.mercury (USEPA, 1984) and elements such as
 selenium (USEPA,  1987a)  have been used to
 establish  water quality criteria and/or to predict
 toxic effects.  There,is some evidence to indicate
 that metal residues in sediment-dwelling aquatic
 organisms can reflect both metal bioavailability
 and potential  metal toxicity.  Thus, tissue residue-
 toxicity relationships for some elements could be
 used as an adjunct to the interstitial water equilib-
 rium partitioning approach.
  13 USEFULNESS                        V
  7.3.1   Environmental Applicability

  7.3.1.1  Suitability for Different Sediment Types

     There is no limitation to the suitability of this
  approach for different sediment types since the
  method is sensitive to bioavailability differences
  among sediments.  When pelagic organisms are
  used to assess sediment quality, sediment variabi-.
  lity in the water body tends to be averaged.

  73.1.2 Suitability for Different Chemicals or
         Classes of Chemicals

      This approach is most applicable to nonpolar
  organics and organometallics that bioaccumulate,
  are slowly metabolized, and exert chronic toxic
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  Sediment Classification Methods Compendium
 effects or present risks to human health.  This
 approach also could work well for chemicals that
 are  metabolized by the organism to nontoxic
 forms since the parent compound residue reflects
 this change  in toxic potential.   In some  cases
 residues of known metabolites, which are more
 toxic than the  parent compound,  can be used to
 establish residue-toxic effects relationships (Krahn
 et al,  1986).   The approach is  not  useful  for
 assessing sediment toxicity associated with non-
 lesidue-forming toxic chemicals such as ammonia,
 hydrogen sulfide, and polyelectrolytes.   Since
 there is evidence that metal  residues in some
 sediment-dwelling organisms  are indicative  of
 both  metal bioavailability  and potential metal
 toxicity,  sediment  quality  criteria for  metals
 should be aided by a site-specific tissue residue
 approach.   However, when  biological species
 sequester metals in  a nonbiologically  available
 form, tissue residue-toxicity effects linkages may
 be obscured.  The suitability of the method for
 evaluating additive,  synergistic,  or antagonistic
 effects  associated with  complex mixtures  of
 sediment  contaminants depends on the develop-
 ment of chemical mixture toxic  dose-response
 relationships  where dose is  indicated by tissue
 residue levels.

 73.13 Suitability for Predicting Effects on
        Different Organisms

    The tissue  residue approach should not be
 limited by species unless organism residues cannot
 be obtained or toxic effects cannot be determined
 through water quality criteria or bioassays.  The
 key  species problem is identification of sensitive
 species for the sediment contaminants of concern.
 When adequate comparative  toxicity data exist,
 residues from tolerant organisms may be used to
 infer sediment criteria for sensitive organisms that
 are not found in association with the sediment
 because of toxic effects.

 7.3.1.4 Suitability for In-Place Pollutant Control

   Evaluation of the association of site-specific ;
tissue residues  with  sediment toxic  chemical
concentrations provides an established method for
  in-place  pollutant assessment  for both  human
  health and ecological risks. Comparison of tissue
  residues in field-collected organisms to the MFTC
  would be a direct estimate of ecological risk. The
  use of resident or caged biota for bioaccumulation
 "potential and toxicity assessments is useful for
  detection of the most toxic sediments or monitor-
  ing of changes in toxicity following remedial
  action.   By  weighing  the  relative toxicity  of
  bioaccumulated pollutants (e.g., by using "dioxin
  equivalents"), evaluation of tissue residue concen-
  trations can help identify the pollutants most likely
  responsible for toxicity and their additive contribu-
  tion to total sediment toxicity.  This information
  could then be used to design the most appropriate
  and cost-effective mitigation response.

  7.3.7.5 Suitability for Source Control

     The tissue residue approach is well suited for
  establishing source control.  Comparison  of the
  existing  or  predicted tissue residue levels with
  MPTCs generates a quantitative estimate  of the
  extent to which a given sediment exceeds or .is
  below a sediment quality criterion.  In conjunction
 with physical transport models, this information
  can then be used directly to determine acceptable
 discharge limits, wasteload allocations,  or the
 types of remedial procedures required to achieve
 acceptable tissue residue levels. The Lake Ontario
 TCDD-Hyde Park  Superfund case example de
 scribed in Section 7.1.1 demonstrates the suitabili-
 ty of this approach for establishing source con-;
 trols.   The site-specific nature of this approach
 provides strong support for establishing controls
 on existing point and nonpoint  sources of sedi-
 ment contamination.

. 7.3.1.6 Suitability for Disposal Applications

    When site-specific sediment-biota contaminant
 partition coefficients are unavailable, such as for
 evaluation of proposed disposal operations, the
 residue approach can be applied  by predicting
 benthic tissue residues from steady-state toxico-
 kinetic bioaccumulation models or by conducting
 laboratory bioaccumulation tests on the dredged
 material.  If adverse effects on fishes, wildlife, or
7-6

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                                                                    7—Tissue Residue Approach
human health are of concern at such disposal sites,
it would then be necessary to apply a trophic
transfer  or equilibrium partitioning  model  to
predict tissue residues in these higher trophic
levels.  When the disposal site already.has sedi-
ments containing the contaminants of concern,
residues in existing biota may be used to predict
residue levels and toxic effects that would result
from additional disposal of similarly contaminated
dredged material.

7.3.2   General Advantages and Limitations

7.3.2.1 Ease of Use

     The application  of sediment  quality criteria
derived from tissue residues for assessing pelagic
or benthic ecological effects is fairly direct.  The
measured  or predicted sediment concentration
would simply be compared to the sediment quality
 criterion derived from MPTCs.  The development
 of a tissue residue toxicity database from laborato-
 ry bioassays would allow convenient access to the
 required biological effects endpoints.  Chemical
 analyses of sediment, total organic carbon, and
 tissue samples for assessing existing conditions
 require routine analytical  chemistry  capabilities
 that do not present unique problems.  One poten-
 tial difficulty when using tissue residues in field-
 collected  benthos to assess in-place sediments is
 the difficulty  in  obtaining  sufficient benthic
 biomass for chemical analysis. This problem can
"be avoided by conducting laboratory bioaccumula-
  tion tests  on field-collected sediment or by placing
  caged benthic organisms in the field.

  7.3.2.2 Relative Cost

      Costs associated with further development of
  the generic tissue residue approach for sediment
  quality  criteria  include  (1) development  of a
  residue-toxicity relationship database and (2) vali-
  dation of the relationships between the MPTC and
  chronic impacts on aquatic organisms for different
  chemical classes of sediment contaminants.  The
  cost of applying the method to  a particular site,
  however, depends on the number of sediment and
  biota samples to  be analyzed, the availability of
residue-toxicity relationship data, and the difficul-
ty in identifying sensitive organisms.  The estab-
lishment of a sediment criterion based on fish
residue levels acceptable for protection of human
health generally  results in low analytical costs
when  only  a few  reference sediment  sites are
needed to characterize the system of concern.

7.3.2.3 Tendency to Be Conservative   J

    This approach  does not tend to  be either
conservative or liberal for prediction of ecological
effects unless the system responds in a nonlinear
manner to reductions in sediment  contaminants.
In the case of nonlinearity,  the tendency would
probably be toward conservatism because of the
greater bioavailability of more recently introduced
sediment contaminants.   When human health
 endpoints are used to generate sediment quality
 criteria, the criteria may be more strict than neces-
 sary to protect resident biota.

 7.3.2.4 Level of Acceptance

     The tissue residue approach is accepted as a
 basis for regulatory decisions such as the estab-
 lishment of water quality criteria for the protection
 of aquatic life and its uses.  The direct prediction
 of chronic toxic effects from measured or predict-
 ed tissue residues requires Validation before it can
 be widely endorsed. Since sediment contaminants
 tend  to be long-term exposure problems and can
 bioaccumulate, residues should be acceptable for
 sediment criteria  development.  This approach
 should  be acceptable for identifying sediments
 associated with a degree of exposure which ex-
 ceeds that  indicated as deleterious in  previous
  experiments.

  7.3.2.5 Ability to Be Implemented by
         Laboratories with Typical Equipment
         and Handling Facilities

      The tissue residue approach requires analyses
  of only sediment  and tissue residues when poten-
  tially toxic sediment contaminants are known and
  residue-toxicity relationship data are available. If
  extensive laboratory work is needed to determine
                                                                                                7-7

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 Sediment Classification Methods Compendium
 chemical residue-chronic toxicity dose-response
 relationships  for sensitive  species,  specialized
 aquatic toxicology capabilities are required.  In
 theory,  residue-toxicity-based  MPTCs  can  be
 obtained for all chemicals subject to water quality
 criteria development.

 7.3.2.6 Level of Effort Required to Generate
        Results

    The level of effort depends on the number and
 nature of sediment contaminants, the  complexity
 of the contaminant distribution  pattern,  and the
 regulatory application of the method. Some cases
 will require relatively few analyses of tissue and
 sediment residues and no toxicity testing to apply
 the method (e.g., to  remedial action decisions,
 wasteload allocations).

 73.2.7 Degree to Which Results Lend
        Themselves to Interpretation

    Tissue residues  that exceed concentrations
 considered  safe for  human  exposure   through
 seafood  consumption require no interpretation
 wlien used to set residue-based sediment  criteria.
 However, the degree of interpretation may be very
 large when  evaluating  ecotoxicological  effects
 attributed to site-specific measurements of sedi-
 ment-to-biota chemical partitioning. This interpre-
 tation problem exists for all sediment classification
 methods when  applied on  a site-specific basis.
 The presence of unacceptable residues in indicator
 organisms  resident in or linked to an  area  of
 sediment  contamination  can  be  used   without
 elaborate interpretation to determine compliance of
 sediments with sediment quality criteria.

 7.3.2.8 Degree of Environmental Applicability

    The  use of site-specific  tissue residues  as
 quantitative  exposure  biomarkers   eliminates
uncertainties  associated with  chemical bipavail- .
ability; exposure duration, frequency, and  magni-
tude; and toxicokinetic/bioenergetic factors. When
the tissue residue approach is applied on a  generic
basis to generate  sediment criteria  for different
chemicals,  these uncertainties can be partially
 addressed through classification of sediments and
 exposure environments.

 7.3.2.9 Degree of Accuracy and Precision
    \ '      '
    Sediment and tissue residue chemical concen-
 trations can be determined accurately and precise-
 ly  for most  chemicals.   Most  uncertainties  in
 sediment/organism partition coefficients are due to
 biological variability. Accuracy and precision can
 be maximized through site-specific investigations
 of biological  factors- that  influence  organism
 linkage to sediment (through food chain, water, or
 direct contact) and through refinement of residue-
 toxicity relationships.  ,
 7.4 STATUS
 7.4.1   Extent of Use

    Use of tissue residues to establish sediment
 criteria on the basis of human health effects has
 been documented. Tissue residues have also been
 used to derive water quality criteria for the protec-
 tion of aquatic life and wildlife connected to the
 aquatic food chain.  Tissue residue-toxicity data
 that may be used for deriving numerical sediment
 quality criteria for some chemicals already exist in
 water quality criteria documents, fish consumption .
 advisories, and the peer-reviewed literature.  Much
 aquatic toxicology work in progress or planned for
 the future could produce the  necessary data if
 residue-based dose measurements are incorporated
 into research plans.

 7.4.2    Extent to Which Approach Has Been
        Field-Validated

    Sediment TCDD  contamination limits have
been established for Lake Ontario on the basis of
fish tissue residues.' This use of tissue residue to
generate  sediment  criteria  has  been  validated
through a steady-state  model  (Endicott et  al.,
1989)  and a laboratory  bioaccumulation  study
(Cook et  al., 1989) that demonstrated  a  linear
relationship  at  steady-state  between  sediment
contaminant  concentration and bioaccumulated
7-8

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                                                                  7—Tissue Residue Approach
TCDD in lake trout, regardless of route of uptake.
Declines in DDT residues in fish and birds since
its use was banned are associated with declining
surficial  sediment concentrations in  the  Great
Lakes, the Southern California Bight, and else-
where.   Although other examples of studies
validating the residue approach for single chemi-
cals are available, its use for complex mixtures of
chemicals  in sediments to predict  acceptable
contaminant concentrations with ecosystem protec-
tion in mind has not been validated.

7.4.3   Reasons for Limited Use

     Use of the tissue residue approach has been
 limited for the following reasons:

     •  This method is in a developmental stage
        and  has not been  formally adopted by
        EPA.

     •  Aquatic toxicology has only recently pro-
        gressed to an understanding of residue-
        based dose-response relationships for sedi-
        ment contaminants.

     • Regulatory agencies, including EPA, have
        not yet become committed to systematic
         establishment and application of sediment
         criteria methods.

     •  The available and potentially available
         residue-based toxicity data have not been
         collated  into  a database  for potential
         sediment criteria users.

  7.4.4  Outlook for Future Use and Amount
         of Development Yet Needed

     This method can  be implemented  with a
  minimal amount of effort in many cases, especial-
  ly where a single chemical or  lexicologically
  related family of chemicals is of concern. Guid-
  ance documents should be written and reviewed..
  Tissue residue criteria should be  accumulated
  systematically for  a database.  The use of  this
  method  in combination  with other sediment
  classification methods should be considered. Field
validation  of  residue-based  ecological  effects
predictions is essential. All sediment assessment
methods should be developed with concern for
identification of and application to those chemi-
cals in the aquatic environment that are long-term
sediment  contaminants having chronic toxicity
potential.                         •'..,'
 7.5  REFERENCES

 Batterman, A.R., P.M. Cook, KB. Lodge, D.B.
    Lothenbach, and B.C. Butterworth. In press.
    Methodology used for a laboratory determina-
    tion of relative contributions of water, sedi-
    ment and  food chain routes  of  uptake  for
    2,3,7,8-TCDD bioaccumulation by lake trout
    in Lake Ontario.  Chemosphere.
 Carey, A.E., N.S. Shifrin, and A.C. Roche. 1989.
    Lake Ontario TCDD bioaccumulation study
     final report.  Chapter 1: introduction, back-
     ground,  study description and chronology.
     Gradient   Corporation,   Cambridge,  MA.
     17 pp.
 Clark, T., K Clark, S. Pateson, D. Mockay,  and
     R.J. Norstrom.  1988.  Wildlife  monitoring,
     modeling  and fugacity.  Environ. Sci. Tech-
     nol. 22:120-127.
 Cook, P.M., A.R. Batterman,  B.C. Butterworth,
     KB. Lodge, and S.W. Kohlbry. 1990. Labo-
     ratory study of TCDD bioaccumulation by
     lake trout from Lake Ontario sediments, food
     chain and water.  In:  Lake  Ontario TCDD
     Bioaccumulation Study - Final Report, Chap-
     ter 6. U.S. Environmental Protection Agency,
     Region II, New York.
  Endicott, D:, W. Richardson, and  D. DfToro.
      1989.  Lake Ontario TCDD modeling report
     U.S. Environmental Protection Agency* Large
     Lakes Research Station, Environmental Re-
      search Laboratory   Duluth, Grosse He, MI.
      94 pp.
  Krahn,  M.M., L.D. Rhodes,  M.S. Myers,  L.K
   .  Moore, W.D.  MacLeod, and D.C. Malins.
      1986.  Associations between metabolites of
      aromatic  compounds in bile  and the occur-
      rence of hepatic lesions in English sole (Paro-
      phrys velulus) from Puget Sound, Washington.
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 Sediment Classification Methods Compendium
    Arch. Environ. Contam. Toxicol. 15:61-67.
 Kuehl, D.W., P.M.  Cook, A.R.  Batterman,  D.
    Lothenbach, and B.C. Bufterworth.   1987.
    Bioavailability of polychlorinated dibenzo-p-
    dioxins and dibenzofurans from contaminated
    Wisconsin River sediment to  carp. Chemo-
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 Mac, MJ. 1988.  Toxic substances and survival
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    Contaminants and Ecosystem  Health.  M.S.
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 Stephan, C.E., D.I.  Mount, DJ. Hansen,  J.H.
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    1985.   Guidelines  for deriving  numerical
    national water quality criteria  for the protec-
    tion of aquatic  organisms and  their uses.
    PB85-227040.  National Technical Informa-
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 Thomann, R.V. 1989. Bioaccumulation model of
    organic chemical distributions in aquatic food
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 USAGE.  1991.  Influence of sediment potential
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    fined Disposal Facility.  Environmental Ef-
    fects of Dredging Notes - EEDP-02-16, U.S.
    Army Corps of Engineers.  U.S. Army Engi-
    neer Waterways Experimental Station, Vicks-
    burg, MS.
 USEPA. 1980a.  Ambient water quality criteria
    for aldrin/dieldrin. EPA 440/5-80-019. NTIS
    number PB81-117301.  U.S. Environmental
    Protection Agency, Washington, DC.
USEPA. 1980b.  Ambient water quality criteria
    forendrin. EPA 440/5-80-047.  NTIS number
    PB81-117582. U.S. Environmental Protection
    Agency, Washington, DC..
 USEPA.  1980c.  Ambient water quality criteria
    for DDT.  EPA 440/5-80-038.  NTIS number
    PB81-117491. U.S. Environmental Protection
    Agency, Washington, DC.
 USEPA.  1980d.  Ambient water quality criteria
    for polychlorinated biphenyls. EPA 440/5-80-
    068.   NTIS  number PB81-117798.   U.S.
    Environmental Protection Agency, Washing-
    ton, DC.                    .
 USEPA. 1984. Ambient water quality criteria for
    mercury.  EPA 440/5-84-026.  NTIS number
    PB85-227452. U.S. Environmental Protection
    Agency, Washington, DC.
 USEPA.  1987a.  Ambient water quality criteria
    for selenium.  EPA 440/5-87-006.  NTIS
    number PB88-142237.  U.S. Environmental
    Protection Agency, Washington, DC
 USEPA.   1987b.  The  national  dioxin study.
    Tiers 3,5,6, and 7. EPA 440/4-87-003.  U.S.
    Environmental Protection Agency, Office of
    Water Regulations and Standards, Washing-
    ton, DC.                               .
 USEPA.    1992.   National study  of chemical
    residues in fish.   2 vols.  EPA 823-R-92-
    008a,b. U.S. Environmental Protection Agen-
    cy, Office of Science and Technology, Stan-
    dards and Applied Science Division, Washing-
    ton, DC.
Walker, M.K., J.S. Spitbergen, J.R.  Olson, and
    R.E, Perterson.  1991.  2,3,7,8-Tetrachloro-
    dibenzo-p-dioxin  toxicity  during early life
    stage development of lake trout (Salvelinus
    namaycush).  Can. J. Fish. Aqua. Sci. 48:875;
7-70

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          .CHAPTER 8
Freshwater Benthic  Macroinvertebrate

Community Structure  and Function

Wayne S.Davis
U.S. Environmental Protection Agency Region V, Environmental Sciences Division
77 West Jackson (SQ-14J), Chicago, IL 60604
312/FTS 886-6233                                ,

Joyce E. Lathrop                 "   '
College of DuPage, Division of Natural Sciences
22nd at Lambert Road, Glen Ellyn, IL 60137
    The community, or assemblage, structure and
function of benthic macroinvertebrates is used exten-
sively to evaluate the quality of water resources and
characterize causes and sources of impacts in lotic
(flowing water) and lentic (standing water) freshwater
ecosystems. (Marine benthic community structure is
discussed in Chapter 9.) Benthic macroinvertebrates
are relatively  sedentary organisms that inhabit or
depend  on the sedimentary environment for their
various life functions. Therefore, Ihey are sensitive to
both long-term and short-term  changes in habitat,
sediment, and water quality. This chapter discusses
assessment of benthic macroinvertebrates to determine
sediment quality in conjunction with an integrated
approach for  assessing  the quality of the water
resources. This integrated approach uses sediment
chemistry, sediment toxicity, habitat quality, and ben-
thic  macroinvertebrate  community  (assemblage)
 structure and  function to evaluate sediment quality,
 similar  to the approaches now used to  evaluate
 surface  water quality.  The structural assessment
 relates to the  numeric taxonomic distribution of the
 community, and the functional  assessment involves
 trophic  level  (feeding group)  and morphological
 assessment.   This chapter addresses the  specific
 benthic community  assessment methods  that are
 available, or  being developed, to complement the
 chemical and toxicological portions of the sediment
 quality  assessment

 8.1 SPECIFIC APPLICATIONS

'8.1.1   Current Use
     Freshwater  benthic macroinvertebrate commu-
 nities are used in the following ways to assess the
quality of the water resource (sediments, water, and
habitat):                            ..,..'

    •  Identification of the quality of ambient
       sites through a knowledge of the pollution
       tolerances and life history requirements of
       berathic macroinvertebrates;

    »  Establishment of  criteria  and  standards
       based .on  community  performance at
       multiple  reference sites throughout an
       ecoregion or other regionalization categor-
       Comparison of the quality of reference (or
       least impacted)  sites with test (ambient)
       sites;

       Comparison of the quality of ambient
       sites with historical data to identify tem-
       poral trends; and

       Determination of spatial gradients of con-
       tamination for source characterization.
 8.1.1.1 Ecological Uses

    Benthic macroinvertebrate community (assem-
 blage) structure and function assessments/have
 many different applications. Site-specific knowl-
 edge of surface  water  quality; habitat quality,
 sediment chemistry, and sediment toxicity provide
 the best context in  which to interpret  benthic

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Sediment Classification Methods Compendium
community assessment data.   The objectives of
each particular study should determine the types of
related data necessary.  Alone, benthic macroin-
vertebrates can be used to screen  for  potential
sediment contamination .based on spatial gradients
in community structure, but they should not be
used  alone to definitively determine sediment
quality.  Benthic macroinvertebrate data must be
integrated with other available data to determine
sediment quality. Benthic macroinvertebrate often
provide the most important piece of information on
sediment  quality.   Care must be  exercised to
collect representative samples to minimize prob-
lems with data interpretation due to natural varia-
tions.  For example, collections should not be
made after floods or other physical disturbances
that may physically alter or remove benthic assem-
blages.
    Benthic macroinvertebrate community structure
and function have been used extensively to charac-
terize freshwater ambient conditions and impacts
from  various sources.   Documented changes in
benthic community structure have resulted from
crude oil exposure in ponds and streams (Rosen-
berg and Wiens, 1976; Mozley, 1978; Mozley and
Butler, 1978; Cushman, 1984; Cushman and Goy-
ert, 1984) and heavy metal  contamination of lake
sediments and streams (Winner et al., 1975,1980;
Wentsel et al., 1977; Moore et al., 1979; Wieder-
holm,  1984a,  1984b;  Waterhouse  and Farrell,
19S5). Benthic macroinvertebrates have been used
extensively to identify organic enrichment in lentic
systems (Cook and Johnson, 1974; Krieger, 1984;
Rosas et al., 1985) and lotic systems (Richardson,
1928; Gaufin.and Tarzwell, 1952; Hynes, 1970;
Hilsenhoff, ,1977,  1982, 1987, 1988).  Benthic
community responses to pesticides (van Dyk et al.,
1975; Webb, 1980; Penrose  and Lenat,  1982;
Yasuno et al., 1985), acid- and mine-stressed lotic
environments  (Simpson,  1983;  Armitage  and
Blackburn, 1985), thermally stressed water bodies
(Grossman et al., 1984), and urban and highway
runoff impacts (Smith and Kaster, 1983; Dupuis et
al., 1985; Denbow and Davis, 1986) have also
been documented.  Chironomidae (midge) larvae
were even found to transport substantial amounts
of  PCBs  from  contaminated sediments to the
terrestrial environment (Larsson, 1984).
8.1.1.2 Regulatory Uses

    Assessment of benthic macroinvertebrate com-
munity (assemblage) structure and/or function has
been used as a regulatory tool for a number of
years (Davis, 1990). In 1987, USEPA hosted the
First National Workshop on Biological Monitoring
and Criteria (USEPA, 1988a, 1988b), which ad-
dressed the use of benthic macroinvertebrates, as
well as fish, in  EPA and State regulatory pro-
grams.  This workshop formally initiated EPA's
efforts toward development and implementation of
"biological criteria"'based on benthic  macroin-
vertebrate, fish, and habitat assessments.   These
biological criteria, which have been predominantly
based on the macroinvertebrates, are designed to
determine whether a specific water body or water
body segment is meeting its  designated use for
aquatic life (i.e.,  water quality standards).
    EPA requires the  development of biological
criteria and  adoption  by  States into  their water
quality standards by September 30,1993  (USEPA,
1991a, 1990b).  This  requirement has been sup-
ported by a  formal policy (USEPA, 1990c), pro-
gram guidance (USEPA, 1992a), and  technical
guidance and support documents (USEPA, 1991a,
1991b,  1991c, 1991d,  1991e;  1992b, 1992c).
Several States currently use  benthic  macroin-
vertebrates as a regulatory tool, either alone or in
combination  with  other  ecological  parameters
(Ohio  EPA,  1990,  USEPA,  1991c,  199le).
USEPA also supports the use of benthic macroin-
vertebrates as a primary environmental  indicator
for surface waters  that EPA should use to track
compliance with Clean Water Act objectives (Abe
et al., 1992; USEPA, 1990d, 1990e).
    Under the Clean Water Act, as amended in
1987, benthic macroinvertebrates are used for the
following:

    •  Measurement of the restoration and main-
       tenance of biological integrity in surface
       waters (section 101);

 •   •  Development of water quality criteria based
       on biological assessment methods when nu-
       merical criteria for toxicity have not been
       established [section 303(c)(2XB)];
8-2

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                    8—Freshwater Berithic Macroinvertebrate Community Structure and function
   •  Production of guidance and criteria based
      on biological monitoring and assessment
      methods [section 304(aX8)];

   •  Development of improved measures of the
      effects of pollutants on biological integrity
      (section 105);

   •  Production of guidelines  for evaluating
      nonpoint sources (NFS) [section 304(f)];

   •  Listing of waters lhat cannot attain desig-
      nated  uses  without  additional   NFS
      controls (section 319);

   ,•  Listing of waters unable to support bal-
       anced   aquatic   communities   [section
       304(1)];

   •   Assessment of lake  trophic states and
       trends (section 314);

   •   Production  of  biennial  reports, oh the
       extent to which waters support balanced
       aquatic .communities  [section  305(b)];
       and,

    •  Determination of the effect of dredge and
       fill   disposal   on   balanced   wetland
       communities (section 404).

    Benthic  macroinvertebrates   and biological
criteria have also  been used to  evaluate on-site
and off-site ecological  impacts  from hazardous
waste sites.    Environmental assessment  of a
Superfund site is done in accordance with EPA's
responsibility  to  protect public  health and  the
environment under the  Comprehensive  Environ-
mental Response, Compensation, and Liability Act
of 1980 (CERCLA) as amended by the Superfund
Amendments  and Reauthorization Act of 1986
(SARA). The regulation that enables EPA to carry
out its responsibilities under CERCLA/SARA is
the National Contingency Plan (NCP).
    The  NCP calls  for  the identification  and
mitigation of environmental impacts of these sites
and the  selection  of  remedial  actions that are
"protective   of environmental   organisms  and
ecosystems." Federal and state laws and regula-
tions that aid in this process are potentially "appli-
cable or relevant and appropriate requirements"
(ARARs).   Compliance  with these  laws and
regulations  increasingly requires that the  site's
ecological effects be evaluated and measures be
taken to mitigate those adverse effects.
    The Clean Water Act, as amended by the
1987 Water Quality Act, is another ARAR and
major federal regulation that  requires the  main-
tenance and restoration of the  chemical, physical,
and biological integrity of the Nation's waters.
Most Superfund sites potentially affect surface
waters and  need to  be assessed for both on-site
and off-site effects.  A detailed discussion  of the
legal and technical requirements for environmental
assessments at Superfund sites can be found in
EPA's Risk Assessment Guidance for Superfund:
Environmental  Evaluation   Manual   (USEPA,
1989a).  As EPA focuses on watershed and water
body  impacts regardless  of the programmatic
sources and causes, the use of benthic macroin-
vertebrates  for  assessing  the health  of surface
water systems will increasingly become important

 8.1.2   Potential Use

    The use of benthic macroinvertebrates to assess
 sediment contamination will  be  most successful
 when combined  with  sediment  chemistry and
 toxicjty results, as  in the "integrated" Sediment
 Quality Triad approach (see Chapter 10). Benthic
 macroinvertebrates  will  best indicate in-place
 pollutant control needs  through a site-specific
 knowledge of surface water quality, habitat quality,
 and  sediment chemistry and toxicity*   Habitat
 quality assessments will help establish reasonable
 expectations for benthic community structure and
 function. Alone, benthic macroinvertebrates can be
 used to screen for potential sediment contamination
 and  source identification by displaying  spatial
 gradients in community structure, but they should
 not be uised alone to definitively determine sedi-
 ment quality or to develop chemical-specific guide-
 lines.   Benthic macroinvertebrate data must  be
 integrated  with other available data to determine
 sediment  quality as well as the quality of the
 overall water resource.
                                                                                              8-3

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 Sediment Classification Methods Compendium
 8.2 DESCRIPTION

 8.2.1  Description of Method

     The  benthic  macroinvertebrate community
 structure and function assessment  involves  the
 following steps:

     (1) Establishment of data  quality objectives,
        selection of sample sites and frequency of
        collection  in  Quality Assurance Program
        Plan;

     (2) Collection of benthic macroinvertebrates in
        the field (artificial or natural substrates);

     (3) Sorting the organisms from debris (field or
        laboratory);

     (4) Identification to the lowest taxon necessary
        (varies depending on the study objectives);

     5)  Multimetric or composite index quantifica-
        tion (e.g., taxa richness, number of individ-
        uals,  indicator organism count,  structural
        indexes  and ratios, functional character-
        istics of taxa);

     (6) Assessment of the relationship with other
        environmental  measurements   including
        numeric  habitat quality assessment (e.g.,
        correlations,  habitat  requirements)  and
        expectations;

     (7) Comparison with a local or regional  "refer-
        ence"  site  (e.g., similarity indexes, non-
        parametric analyses); and

     (8) Complete  documentation  of the  study
        methods,  results,  database  management,
        and discussion of the relevance of the data.

8.2.1.1  Objectives and Assumptions

    The primary objective of benthic macroinverte-
brate community (assemblage) structure and func-
tion analyses is to provide data and information to
assist  in  determining  the  quality  of the  sedi-
 ment/water environment.  This determination can
 then be used for the purposes described above in
 Section 8.1 (Specific Applications).
     It is assumed that benthic macroinvertebrates
 can provide consistent and accurate assessments of
 sediment/water quality at a given sample location or
 water body.  Specifically, the following assump-
 tions are implicit in this objective:

     •   The benthic macroinvertebrates are rela-
         tively sedentary, especially compared  to
         fish communities, and they depend on the
         sedimentary (or benthic) environment for
         their life functions.

     •   Chemical and physical perturbations of the
         sediments or bottom waters affect benthic
         macroinvertebrates since they are depen-
         dent on the benthic environment for com-
         pletion of their life  cycles, and they are
         therefore sensitive to changes in sediment
         and water quality.

     •   Benthic   macroinvertebrates   physically
         interact with the sediments  to cause chem-
         ical exchange between the sediment  and
         the overlying water,  and therefore tend to
        reflect sediment quality as well as water
         quality.

     •  Minimum  habitat quality  exists  below
        which the community structure and func-
        tion will perform poorly regardless of the
        chemical  contaminants  present  or  not
        present.

     •  The optimal use of benthic macroinverte-
        brates as sediment quality indicators is as
        part of an integrated sediment quality as-
        sessment   approach  using    sediment
        chemistry, sediment  toxicity, and benthic
        community structure  and function.

    Equally important assumptions apply to actual
benthic macroinvertebrate sampling strategy, collec-
tion, identification, data reduction, interpretation of
results, and report preparation.  It is assumed that
all U.S. EPA-supported studies have an adequate
8-4

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                     8—Freshwater Benthic Macroinvertebrate Community Structure and function
Quality Assurance Project Plan (QAPP) and that all
benthic  macroinvertebrate  community data  are
reproducible and collected in a manner to minimize
data interpretation problems with natural variations;
the methods must be consistent within each study.
Specific QA procedures'that should be established
early  in  benthic  macroinvertebrate  community
studies include the following:

    •   Rationale for sample location selection;

    •   Sample collection methods, sorting,  and
        storage procedures;             <

    •   Taxonomic proficiency evaluations using
        either U.S. EPA check-samples from Cin-
        cinnati-ERL  or  state-developed check-
        samples, in addition to voucher collec-
        tions from each study  area and  a list of
        the taxonomic references used;

     •  Multimetric data analysis techniques used
        to objectively assess the data, including
        the  structural and functional measures;
        and

     •  Nonparametric or parametric (as appropri-
        ate) statistical methods used to  compare
        site results.

     Each Regional U.S.  EPA  Quality Assurance
 Office can provide the details of QAPP require-
 ments.   Further discussion of quality  assurance
 measures can be found in Klemm et al. (1990),
 Bode (1988), Ohio  EPA (1989b), and  Stribling
 (1991).

 8.2.1.2 Level of Effort

     The level of effort required to conduct fresh-
 water   benthic   macroinvertebrate  community
 studies  is  comparable  with  chemical/physical
 water quality measurements and bioassays and has
 been thoroughly discussed in Plafkin et  al. (1989)
 and Ohio EPA (1990a).  However, rapid benthic
 community assessment techniques can range from
  1 to 5 hours per site if  laboratory identifications
  are not required (Plafkin et al., 1989). As expect-
ed, the greatest time expenditure is in the travel to
and from the site and in the sorting and identifica-
tion of the organisms.
    Separating the organisms from  debris  and
sorting the organisms mto taxonomic categories
can take up  to 15 hours per sample, with an
additional 12 hours for identification,  for very
enriched sites with high numbers of individuals
among several taxa.  In such extreme situations,
subsampling may be preferred.  More  typically,
the time spent would be about 3 hours for sorting
(more  time for dredge and artificial  substrate
samples and  less time for  dip-net  samples),
 2 hours for preparing the samples (e.g., clearing
 and then mounting the chironomids on microscope
 slides),  and 6 hours for identifying the organisms
 to the lowest possible taxonomic level. An exper-
 ienced  taxonomist with appropriate keys  may
 average only 2-4 hours per site. This typical time
 equates to about 11 hours per site after the sam-
 ples have  been collected.   These estimates are
 only a  general'guide to the time it  may take to
 perform the identifications and are meant to help
 assess potential or actual project costs.

 8.2.1.2.1    Type of Sampling Required

     The specific sampling methods to be used are -
 dictated by the study needs.  Debate will continue
 regarding the use of "quantitative" and "qualita-
 tive" sampling methods,  but  each method  is
 acceptable contingent upon how well it will satisfy
 study objectives, reproducibility of the data, and
J consistency  of collection.    Typically, benthic
 macroinvertebrate data are  quantified  by the
 surface area of the  sampler or sediment being
  collected.  However, benthic macroinvertebrates
  can be quantified in other ways depending on the
  objectives  of the study.   For example,  if the
  objective is to determine the number and types of
  taxa in a study area, rather than the  number  of
  individuals within each taxon, then using a dip-net
  in various habitats within the study area until no
  new taxa are encountered could be  considered
  quantitative with relation to the number of taxa
  and tune expended.  Examples of programs using
  data quantified by methods other than surface area
  of the sampler or substrate include those described
                                                                                               8-5

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 Sediment Classification Methods Compendium
 by Pollard (1981), Hilsenhoff (1982^ 1987,1988),
 Cummins and Wilzbach (1985), Bode and Novak
 (1988),  Cummins (1988), Kite  (1988),  Lenat
 (1988),  Maret  (1988),  Penrose  and  Overton
 (1988),  Plafkin  et  al (1989), and Shakelford
 (1988).   The  success  of each sampling  effort
 depends on a thorough understanding of the data
 quality objectives of that stu/ly and the implemen-
 tation of a quality assurance program.

 8.2.1.2.2 Methods

    EPA (Klemm et al., 1990) recently published
 Macroinvertebrate Field and Laboratory Methods'
 for Evaluating the Biological Integrity of Surface
 Waters, which thoroughly addresses methodology.
 Most  state environmental  regulatory  programs
 have a Quality Assurance Project Plan describing
 the field methods and standard operating proce-
 dures for collecting and evaluating benthic macro-
 invertebrates (Bode,  1988; Illinois EPA,  1987;
 Ohio  EPA,  1989a,  1989b).   This information
 should be  obtained  to ensure acceptance and
 comparability of study results with those obtained
 by die state agency.  If  this information is not
 available, then field methods and standard operat-
 ing procedures  from  other  existing  programs
 should be used.  Since several different collection
 and analysis methods  are  used throughout the
 country  depending on water body type (lotic vs.
 lentic), habitat type, substrate type, and familiarity
 with specific methods, it is not practical to recom-
 mend any single sampling method.  The general
 quality assurance requirements the use of any one
 particular method is that the method produce data
 that are reproducible, consistently used within the
 program, and applicable by different investigators
 (Klernm et al, 1990).

 Methods Summary—In soft freshwater sediments
 the most common method used to collect benthos
 is with a grab sampler such as a Ponar (15 x 15
 cm or 23 x 23 cm) or Ekman dredge (15 x 15 cm,
 23 x 23 -cm, or 30 x 30  cm), each  of which
 provides a quantitative  sample based  on  the
'surface area of the sampler.  The smaller of the
 surface area  sizes are most commonly used for
 freshwater studies because of their relative ease of
 manipulation.  The Ekman dredge is not as effec-
 tive  in areas of vegetative debris, but is much
 lighter than the Ponar and easier to use in softer
 substrates.   Artificial  substrates (Hester-Dendy
 using several 3-inch plates and spacers attached by
 an eyebolt; or substrate/rock-filled baskets) pro-
 vide a consistent habitat for the benthos to colo-
 nize in  both soft-bottomed and  -stony  areas.
 Artificial substrates can be used in  almost  any
 water body and have been successfully used to
 standardize  results despite habitat  differences
 (APHA et fl/.,'1989; DePauw, 1986; Hester  and
 Dendy, 1962; Ohio EPA, 1989b; Rosenberg  and
 Resh, 1982, 1991), but the major drawback to
 using the artificial substrates is the 4- to 8-week
 period for  instream  colonization.   This would
 require at least two visits for each study site—one
 to place the samplers and one to remove them.
    A variety of  methods for  sampling hard-
 bottomed lotic systems are available. Colonization
 of substrates and comparisons of the artificial and
 natural substrate methods  have  been described
 (Beckett and Miller, 1982; Chadwick and Canton,
 1983; Grossman and Cairns, 1974; Lenat, 1988;
 Ohio EPA, 1989b; Peckarsky, 1986; Plafkin et al.,
 1989; Shepard, 1982).  If quantification by sedi-
 ment or sampler surface area is needed, a Surber-
 type  square-foot sampler  (Surber, 1937,  1970)
 with a #30-mesh (0.589-mm openings) can be
 used. The traveling kick-net (or dip-net) method,
.also using a #30-mesh net, can be used to quantify
 the sample'collected by the amount of time spent
 sampling and the approximate surface area sam-
 pled  (Pollard, 1981; Pollard and Kinney, 1979).
 The Surber-type and kick-net methods can each be
 used to provide consistent, reproducible samples,
 but  both are limited  to  wadable  streams. The
 Surber sampler's optimal  effectiveness is limited
 to riffles, whereas kick- or  dip-net  sampling  can
 be used in all available habitats. Although dip-net
 samplers  have been effectively used to sample
 riffles and  other  relatively shallow  habitats to
 determine  taxa richness,  presence of  indicator
 organisms, relative abundances, similarity between
 sites, and other information, they do not provide
 definitive estimates of the number of individuals
 or biomass per surface area.
8-6

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                     8—.f reshwater Benthic MacroinvertArate Community Structure and Function
    For sediment evaluations of lotic systems, a
combination of artificial substrate (e.g., Hester-
Bendy) and natural substrate (dip-net) sampling is
recommended. This combination allows compari-
son of the benthos communities independent of
habitat so that sediment/water quality effects can
be better assessed.

Sampling Strategy—Sampling strategies  have
been addressed by Klemm et al. (1990), Millard
and Lettenmaier (1986),   Plafkin et al. (1989),
Rosenberg and Resh (1991), Sheldon (1984), and
USEPA 1990b, 1990c). Special monitoring strate-
gies have been prepared for EPA's Environmental
Monitoring and  Assessment Program (EMAP),
which  employs  a probabilistic  sample  design
(USEPA, I991f); the intensive watershed surveys
of the U.S. Geological  Survey  (Leahy  et  al.,
 1990); and forestry activities in the Pacific North-
west (USEPA, 1991g).  Regardless of the study
 objectives for regulatory use, reference (least-
 impacted) sites  will be required  for comparison
 with the results from test (ambient) sites.  Refer-
 ence sites can be established on a site-specific or
 regional basis.  It is1 preferable to use a regional-
 ization approach because the level  of confidence
 in the results is greater using an increased number
 of reference sites, which  allows for a verification
 that  the  sites truly are least-impacted reference
 sites.  Regionalization (ecoregions, watersheds)
 has been successfully used in a number of State
 programs to support biological criteria develop-
 ment for benthic macroinvertebrates (Gallant et
 al.,  1989, Ohio EPA, 1990, Arkansas  DPCE,
 1987, Hughs et al., 1990, USEPA, 1991c,1991e).
     When using site-specific reference sites  to
 detect spatial differences in sediment/water quali-
 ty, or to characterize sources of pollution, the best
 strategy is to collect samples in similar habitats
 upstream and downstream of suspected pollution
 sources  or other  areas  of interest for ambient
 monitoring such as high-quality or wild and scenic
 streams (USEPA, 1992b).  A minimum  of two
 upstream sites  and three downstream sites of the
  suspected pollutant source(s) should be sampled;
  however, many programs are limited to only one
  upstream site and one or two downstream sites. If
  habitats vary too widely, then artificial substrates
should be placed at each site, with multihabitat
dip-net sampling done when the substrates are
placed instream and retrieved, to complement the
artificial substrate data.
    To best detect temporal trends, a fixed station
network should  be established near  the area of
interest and  sampled consistently at least one
season each  year.  A .reference location should
also be sampled at the same times to ensure that
differences found in the results can be attributed
to changes in water quality near the site.   It is
strongly recommended that a set of reference sites
be developed within each ecdregion  (or by other
regionalization methods) and that those reference
sites be sampled seasonally to better understand
site-specific seasonal variability. Sampling should
be done each year during similar  flow conditions
and should not  be conducted  for at least 1 or 2
weeks after a major rainfall because of the poten-
tial  for  physical  disturbances of- the substrate
resulting  in potentially lower biological integrity
ratings.
     Seasonal distributions are always a  concern
 for  ensuring the  collection  of a representative
 sample.  Therefore, routine sampling or monitor-
 ing is  optimal  during  the seasons  indicated in
 Plafkin et al. (1988), and long-term monitoring
 should strive for consistent sampling seasons. The
 benthic macroinvertebrate discussion group at the
 1987 National Workshop on Instream Biological
 Monitoring and Criteria agreed that the biological-
 ly  optimal  time  of  year for sampling in lotic
 systems was during the latter part of the season(s)
 that demonstrate a stable base-flow (normal flow)
 and temperature regime (Davis and.Simon 1988).

 Sample  Replication—Sample  replication is a
 component of a good Quality Assurance Program
 Plan. Recommendations and discussion regarding
 sample replication can be found in Plafkin et al.
 (1989), Klemm etal. (1990), and USEPA (1992b).
 Statistical derivation of  the  number of samples
 required to decrease the variability of the data
 have been discussed by Green (1978), Merritt et
  al. (1984), Resh  and Price (1984), and Klemm et
  al. (1990). These methods generally rely on a
  prior knowledge of the  variability of  the data.
  This prior knowledge is  often not available nor
                                                                                              8-7

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 Sediment Classification Methods Compendium
 practical  to  obtain  from a programmatic view
 (e.g.,  the  cost  of initial sampling to estimate
 variability and required number of replicates may
 be prohibitive). Another problem with statistically
 determining the number of samples  needed is the
 assumption that the data follow a specific distribu-
 tion such as normal or lognormal,  which is not
 necessarily true for biological samples. Also, the
 variability,  as  measured by  the  variance  or
 standard deviation,' could be  different for each
 descriptive index analyzed (number of taxa versus
 number of individuals, etc.).

 Field  Methods—Field sampling methods have
 been adequately addressed by many manuals,
 including the new USEPA macroinvertebrate field
 and laboratory manual (Klemm et al.,  1990), the
 ASTM methods for sampling  benthos (ASTM,
 1988), Ohio EPA's Field Methods Manual (Ohio
 EPA,  1989b), Standard Methods (APHA et al.,
 1989), USEPA's Rapid Bioassessment Protocols
 (Plafkin  et al.,  1989), and USEPA's Superfund
 Field Compendium (USEPA, 1987). The follow-
 ing  decisions will need to be made once  the
 sample gear is chosen:

    •  Whether samples will be  picked from
       debris and sorted in the field;

    •  Which preservative should be used;

    •  Whether a stain (rose bengal) will be
       added to the sample to facilitate separat-
       ing the organisms from debris;

    •  Whether the samples need to be shipped
       and whether they require  a chain-of-cus-
       tody form (as in Superfund samples); and

    •  The type of sample containers and label-
       ing of the containers required.

 Sorting—There are many discussions elsewhere
 of techniques for sample sorting and preparation
 of slides  for identification. Klemm et al.  (1990),
 Merrill et al. (1984), Pennack (1978) and APHA
 et al. (1989) offer excellent guidance for sample
 sorting.    Hynes (1970,  1971) stated that the
 earlier stages of benthos are retained by a 0.2-mm
 mesh size (approximately the size of a #75 stan-
 dard sieve), and APHA et al. (1989) and Klemm
 et al. (1990) defined the benthos by a mesh size
 of  0.595  (standard sieve #30), which' is now
 standard practice. However, some types of Chiro-
 nomidae and other small benthos pass through the
 #30-mesh sieve but are be retained by the #40-
 mesh sieve.  It is therefore recommended that
 samples be passed through a #30-mesh sieve and
 that the materials washed  through be passed
, through a #40-mesh sieve; the material retained in
 both sieves should 'then be sorted (Ohio  EPA,
 1989b). Once the material is washed through the
 sieves the organisms should be separated from the
 vegetation and other debris in a white enamel pan.
 As the materials are separated, the organisms can
 be placed in different vials for the major taxa.

 Taxonomy—The level to which the taxonomy
 should be taken is dependent on the objectives of
 the study.  For a system reconnaissance or screen-
 ing survey,  it is generally  not necessary  to go
 beyond the family level (Hilsenhoff, 1988; Illinois
 EPA, 1987; Plafkin etal, 1989; Resh, 1988). For
 studies  attempting to  identify  designated use
 impairment and/or evaluate impacts from a speci-
 fic  source,  the recommended minimum level  of
 taxonomic detail should follow the list by Ohio
 EPA (1989b). Ohio EPA has successfully imple-
 mented numeric biocriteria based on this taxonom-
 ic detailing.  This strategy is to expend the  effort
 to differentiate those taxa which are better  water
 resource quality indicators  and for which  taxo-
 nomic keys  and expertise are readily available.
 The level of taxonomic detailing must be consis-
 tent within the program and applied for each study
 site. Species-level identifications for all organisms
 are  not necessary for a successful program, and
 they.commonly depend on the availability of local
 keys.   General keys  available for genus-level
 identifications include  Merritt  and  Cummins
 (1984) for  insects, Peckarsky et al.  (1990) for
 insects and other invertebrates, Pennack (1978,
 1989) for all common invertebrates, Wiederholm
 (1983) for midges, and Klemm (1985) for annelids
 (oligochaetes and leeches).  Klemm et al. (1990)
 provide an excellent list of taxonomic references
8-8

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                     8—Freshwater Benthic Macroinvertdrate Community Structure and Function
for other general and specific uses. Regional U.S.
EPA or state biologists should be  contacted to
determine which of the hundreds of other taxo-
nomic keys are available  for specific taxa, both
nationally and regionally.

8.2.1.2.3   Types of Data Required

    The types of data analyses that are required to
meet program objectives directly affect the types
of data required.  A list  of the families of taxa
present may be sufficient to meet some program
objectives.  Under other  circumstances, species-
fevel  taxonomy and enumerations  may  be re-
quired.  The necessary data required to conduct
different types of analyses can be obtained from
the following discussion of data analysis methods.
     One of the most inconsistent and perplexing
 aspects of a freshwater benthic macroinvertebrate
 community assessment is the numeric representa-
 tion and analysis of the data collected. Structural
 community  measures such  as richness  values,
 diversity and  blotic indexes,  and enumerations
 have been used  almost  exclusively.  Indicator
 organisms have been used to establish many of the
 biotic indexes but also have the  potential  to
 differentiate among types of impacts.  Recently,
 functional community measures based on feeding
 groups such as shredder, collector,  scraper, and
 predator  (Cummins  and  Merritt,  1984)  have
 gained wider application and acceptance due to
 their sensitivity in detecting system perturbation
 on food resources.  Sediment and water quality
 assessments based on the benthic macroinverte-
 brate community should use a complementary mix
 of both structural and functional measures. It is
 strongly recommended  that a multimetric tech-
 nique be used (Plafkin  et al., 1989; Ohio EPA,
  1990a) so any single index value or observation
  will not substantially influence the results. Dis-
  cussions of various data analysis  techniques can
  be  found  in Hawkes  (1979),  Cairns  (1981),
  Klernm et al. (1990),  Washington  (1984), and
  Resh and Jackson (1990).

   Composite  Indexes—Composite or multimetric
   indexes combine selected structural or functional
   measures, or  "metrics," in a cumulative scoring
system, as was  done with the Index of  Biotic
Integrity (IBI) for the fish community (Karr et al,
1986). These composite, or multimetric, indexes
are highly recommended and are among the most
used  assessment techniques for development of
biological criteria for both benthic macroinverte-
brates and fish.          ,
    Karr and Kerans (1992) provide an outstand-
ing discussion of the process of developing met-
rics proposed for use in an invertebrate IBI. They
evaluated 28 potential metrics for inclusion and
have eliminated 10 from further consideration.
The metrics fall into three categories:  taxa rich-    •
ness  and community composition, trophic and
 functional feeding group, and abundance.
    Ohio EPA (1989b, 1990a) successfully devel-
 oped a similar  index for invertebrates  using the
 following  10   structural metrics, adjusted for
 drainage area size with each ecoregion, to derive
 a final Invertebrate Community Index (ICI) score:

    (1)  Total number of taxa;

    (2)  Jbtal number of mayfly taxa;

    (3)   Total number of caddisfly taxa;

    (4)   Total number of dipteran taxa;

    (5)   Percent mayflies;

    (6)   Percent caddisflies;

    (7)   Percent Tribe Tanytarsini midges;

    (8)  Percent other dipterans and non-insects;

    (9)  Percent tolerant organisms; and

    (10)  Total number of qualitative EPT taxa.

      The ICI score is part of Ohio EPA's numeric
  biocriteria for designated use attainment, and it
  was developed using artificial and natural sub-
  strate data for 232 "least-impacted" reference sites.
  A statistical validation of the ICI using a factor
  analysis  technique showed  high  correlations
  between the factor analysis scores and the ICI
                                                                                               8-9

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   Sediment Classification Methods Compendium
  scores and little redundancy between the metrics
  (Davis and Lubin, 1989).
      U.S. EPA (Plafkin et al, 1989) developed a
  composite  index  for rapid assessments in lotic
  systems using the following two functional and six
  structural metrics:

      (1) Taxa richness;

      (2) Modified Hilsenhoff biotic index;

      (3) Ratio of scrapers and filtering collectors
         (functional);

      (4) Ratio of EPT and  Chironomidae abun-
         dances;

      (5) Percent contribution of dominant taxon;

      (6) EPT index;

     (7) Community similarity index; and

     (8) Ratio of shredders  to  total number of
         organisms  (functional).

     These Rapid Bioassessment Protocols (RBPs)
 recommend conducting single-habitat (riffle) dip-
 net sampling. The scores are based on a percent-
 age of the metric values found at a reference site,
 rather  than  comparison of  the results based  on
 "optimal" values for each metric. Modifications to
 the RBPs can include use of multiple reference
 sites. The RBPs are flexible and can be modified
 for different geographical locatipris, as evidenced
 by the use of different metrics in Arkansas (Shak-
 elford,  1988) and New York (Bode and Novak,
 1988).  The success of the RBPs is in the use of
 the composite index  for rapid assessments that
 allows  for three levels  of taxonomic work (i.e.,
 order, family, or genus/species levels). Order and
 family taxonomy do not require laboratory taxono-
 my and may be done in the field.   The RBPs
 normally use single-habitat (riffle) sampling and a
 100-organisrn count in the field.  However, they
 can be adapted for most program uses; for exam-
 ple, by employing  multihabitat sampling and/or
 various count limitations. To be applicable to a
  state's program,  the  RBPs should  undergo a
  rigorous validation effort within that state.

  Diversity Indexes — When diversity indexes were
  introduced, they were used widely because of their
  ability to reduce the complex benthic community
  measurements into a single value that could be
  used by nonbiologist decision-makers.  Diversity
  indexes are based on measuring the distribution of
  the  number of individuals among the different
  taxa, and use methods that result in enumerations
  by surface area.   The most common diversity
  index used for water "quality studies is the Shan-
  non,  or Shannon-Wiener Index  (Shannon and
  Weaver, 1949) as shown below:
          Shannon's H1
 where:
      n =
      s =
Total number of individuals in the
itt taxon
Total number of individuals
Total number of taxa.
 (Washington (1984) provides a good explanation
 of how the index derived the name Shannon-
 Wiener Index rather than Shannon-Weaver Index.)
 Theoretically, higher community diversity indi-
 cates better water quality (Wilhm, 1970).  How-
 ever, low diversity may be caused by factors other
 than water quality impacts, such as extremes in
 weather (floods  or  droughts), poor habitat, or
 seasonal fluctuations. Although diversity indexes
 such as the Shannon-Wiener Index still remain in
 widespread use (Washington, 1984), their limita-
 tions in accurately addressing a variety of pertur-
 bations has decreased  their reliability (Cooke,
 1976; Hilsenhoff, 1977; Hughs, 1978; Chadwick
 and Canton, 1984; Washington,  1984; Mason et
 al., 1985; Resh,  1988).  Kaesler et al.,  (1978)
 demonstrated that the popular Shannon's  Index
was actually not the preferred index for aquatic
ecology studies  and recommended the use of
Brillouin's (1962) Index.  Resh  (1988) reported
that diversity indexes showed varied results in de-
8-10

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                     8—Freshwater Benthic Macroinvert&rate Community Structure and Function
tecting changes in water quality and that they are
not the optimal measures of water quality.  How-
ever,  diversity indexes  can  provide additional
information as to the community composition and
should be reported if the data are available.  Reli-
ance on these indexes as the only, or predominant,
measure on which water pollution  control deci-
sions are based is  not valid. Washington (1984)
provides an outstanding review of the history and
uses of diversity indexes.

Biotic  Indexes—Biptic  indexes  use pollution
tolerance scores for each taxon, weighted by the
number of individuals assigned to each tolerance
value. If desired, relative abundance measures can
be used in biotic indexes.   An example of a
widely used biotic index  (Hilsenhoff, 1977,1982)
is as follows:     ,
             Biotic Index =    —
 where:
n  =
               Number of individuals in taxon i
               Tolerance value assigned to taxon
               *•'•.-'.
               Total number of individuals in the
               sample.
     Tolerance values can be found in Hilsenhoff
 (1987) or can be generated by regional-specific
 knowledge of the organisms' tolerances. Typical
 ranges of organism index values are 0-5, 0-10, or
 0-11, with  the higher numbers indicating greater
 tolerance to pollutants.  Community indexes are
 generally limited to lotic systems  impacted  by
 organic enrichment (Woodiwiss, 1964;  Chandler,
 1970; Hilsenhoff, 1977; Murphy, 1978; DePauw
 et  al.,  1986)  or  other  general  perturbations
 (Hawkes,  1979).   Biotic  indexes  based  on a
 specific  population,  rather than  community, are
 addressed in the "Indicator Organisms" discussion
 below.   Although the first widely applied biotic
 index  in this country was developed by  Beck
 (1955) for  Florida streams, the Hilsenhoff Biotic
 Index (Hilsenhoff, 1977, 1982) has gained great
popularity and has been updated to revise  the
scoring system from a range of 0-5 to 0-11 (Hil-
senhoff, 1987) and to include a family-level biotic
index (Hilsenhoff, 1988).  Because the biotic
indexes rely heavily on known pollution tolerances
of the taxa, Washington (1984), Mason et al.
(1985), and Hawkes  (1979) preferred the biotic
indexes over  the  diversity indexes for water
quality assessments.  The success of the Hilsen-
hoff Biotic Index prompted use of the index, or
modifications of it, in several state programs (e.g.,
Wisconsin, Illinois, New York, North Carolina)
and EPA (Plafkin et al, 1989) programs.  Unfor-
tunately,  tolerance values are not available for
many taxa because they tend not to exhibit water
 quality  preferences,   and  the  assessments  are
 generally limited to organic enrichment.  Wash-
 ington (1984)  provides an outstanding review of
 the history and uses of these indexes.

 Indicator Organisms—Indicator organisms have
 played a key  role in the development of biotic
 indexes for both lotic and lentic systems.  One of
 the first classifications based  on indicator organ-
 isms was done in the Illinois River by Richardson
 (1928).  Simpson and Bode (1980), Bode and
 Simpson (1982), and Rae  (1989), among many
', others, used Chironomidae as indicator organisms
 for  a variety of  toxicants in  stream  systems.
 Hawkes  (1979)  provides an excellent review of
 the use of benthic macroinvertebrates for stream
 quality assessments, and Wiederholm (1980) does
 the  same for  lake systems.   Data analyses for
 benthic macroinvertebrates in lentic systems have
 not been as progressive as those in lotic systems
 with regard to composite indexes and have relied
 extensively, on  enumerations, diversity indexes,
 richness values, and indicator organisms (Fitchko,
 1986).  Howmiller  and Scott (1977),  Krieger
 (1984), and Lauritsen et al. (1985) used oligo-
 chaete communities  to establish a Great Lakes
 trophic index.  Lafont (1984)  also used oligo-
 chaetes  to indicate  fine sediment  pollution.
 Brinkhurst et al. (1968) and Winnell and White
 (1985) used  chironomids  to develop a similar
 index for the  Great Lakes, and  Courtemanch
 (1987) classified Maine lakes using chironomid
 larvae similar to the studies of Saether (1979) and
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  Sediment Classification Methods Compendium
 Aagaard (1986) in European lakes.   Hart and
 Fuller (1974) presented pollution ecology data for
 a  number  of freshwater  benthic  macromverte-
 brates,  as  did U.S. EPA's  pollution  tolerance
 information series on Chironomidae (Beck, 1977),
 Trichoptera (caddisflies) (Harris and Lawrence,
 1978), Ephemeroptera (mayflies) (Hubbard and
 Peters, 1978*), and Plecoptera (stpneflies) (Surdick
 and Gaufin,  1978).   Washington (1984) also
 reviewed population-based biotic indexes.


 Richness   Measures—Richness  measures  are
 based on the presence or absence of selected taxa.
 Commonly used measures include the total num-
 ber of taxa, the number of EPT (Ephemeroptera,
 Plecoptera, and Trichoptera),  and the number of
 families.  The higher the richness value is, the
 better the quality of the system. Richness mea-
 sures have been shown to have low variability and
 high accuracy in identifying impact (Resh, 1988)
 and should be applied in each study.


 Enumerations—Enumerations involve - obtaining
 a sample  quantified by surface area  to obtain
 specific abundances  of  each  taxon.  Examples
 include the number of total individuals, number of
 EPT individuals, ratios of number of individuals
 within a taxon to the total number of individuals
 (Ohio EPA, 1989a; Resh, 1988), and ratios of the
 number of individuals within one taxonomic group
 (e.g., EPT) to the number of individuals within
 another  taxonomic group  (e.g.,  Chironomidae)
 (Plafkin et al, 1989; Resh; 1988). Interpretation
 of the enumeration ratios can be difficult without
 prior validation.   Most possible enumerations
 comparing individual taxa to the total number of
 individuals are done for many studies, although
 the results may not be presented.  The percent
 contribution of the individuals within a taxon at a
 sample  site can  be  compared with the  percent
 contribution at the reference sites to detect  a
 change  in  community structure.  Resh (1988)  '
 concluded that the seven common enumerations he
 tested had-extremely high variability and unac-
 ceptably low accuracy in detecting various  im-
 pacts, and he suggested that they are not as useful
 for detecting environmental change as richness
  measures or the family biotic index. Although the
  measures Resh (1988) used may not be optimal
  for widespread use, they may still provide insight
  .into  changes  in the community structure.  Ohio
  EPA (1989a)  has successfully used enumerations
  for the percentage of mayflies, caddisflies, Tany-
  tarsini midges, tolerant  organisms, and "other"
  dipterans combined with  non-insect individuals as
  a basis for their state biocriteria.


  Similarity Indexes—Community similarity index-
  es measure the similarity between benthic  com-
  munities at a reference and a study site, with high
  similarity  indicating little  change, or  impact,
  between the  two  sites.   The  use  of similarity
  indexes has been reviewed by Brock (1977) and
  Washington (1984).   The simplest indexes  to
  apply are those which use only the types of taxa
  found, not the abundance of the organisms within
  each  taxon. The Jaccard Index (1908) and Van
  Horn's Index  (1950) are  examples of the simpler
  indexes.  Van Horn's Index, used by Ohio EPA
  (1989b), is as follows:
             Similarity (c)  =
 where:
     a   =

     b   =

     w   =
Number of taxa collected at one
site
Number of taxa collected at the
other site
Number of taxa common to both
stations.
 A value over 6.5 or 7.0 indicates good similarity.
 Plafkin et al. (1989) use the Jaccard Index in the
 rapid bioassessment  protocols  (RBPs).  Other
 indexes  such as  the  percent similarity  (Brock
. 1977) and the Bray-Curtis (1957) use the abun-
 dance of organisms.

 Functional Information—Community function
 measurements based on habitat, trophic structure,
 and other ecological measures were described by
 Kaesler  et al. (1978) and used by Rooke and
8-12

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                     8—Freshwater Benthic MacrotHVertetirate Community Structure and Function
Mackie (1982a) as the  "ecological  community
analysis" (EGA).  Rooke  and Mackie (1982b)
reported the EGA to provide more information on
environmental  quality than diversity  or biotic
indexes, but the EGA was very time-consuming
and not practical for rapid assessments. However,
Cummins and  Wilzbach (1985) and  Cummins
(1988) describe a rapid assessment method based
on sampling coarse paniculate organic matter and
determining the functional feeding  groups de-
scribed in Merritt and  Cummins (1984).  This
method is recommended in EPA's RBPs (Plafkin
et al, 1989). Rabeni et al (1985) also described
the  usefulness  of a functional feeding group
approach to provide a "more ecologically sound
classification of water quality" during their devel-
opment of a biotic index for papdr mill impacts.
Another useful measure of function is observations
of the incidence of morphological deformities in
benthic macroinvertebrates, similar to the observa-
tions made for Karr's Index  of Biotic Integrity
(IBI)  for fish  (Karr  et  al, 1986).  Deformities
have been associated with exposure of metals and
 organic compounds to Ghironomidae (Cushman,
 1984; Cushman and Goyert, 1984; Wiederholm,
 1984b; Warwick, 1985; Warwick et al., 1987) and
 Trichoptera (Simpson, 1980; Petersen and Peter-
 sen, 1983). Karr and Kerans (1992) are develop-
 ing an invertebrate IBI and  have evaluated 10
 trophic and functional  feeding  group  metrics.
 This promising work is  continuing.

 Statistical  Approaches—Various  statistical
 approaches have been applied to determine wheth-
 er the benthic community at a study site varies
 from that at a reference or other site. An excel-
 lent discussion of this issue appears in Klemm et
 al.  (1990) and USEPA (1992b). Depending «n
 the chosen endpoints of the study, rigorous statis-
 tical analysis may not be necessary.  For,instance,
 if the endpoint is the number of taxa or richness
 measures, the variability is generally quite low and
 accuracy quite high. In this case, the differences
 between  two  communities would  need to  be
 evaluated based on study objectives.  A "statisti-
 cal" difference between two communities will not
 always indicate whether more subtle changes in
 community composition are occurring or whether
mitigation may be warranted before a statistical
change  occurs.   Sometimes when that  change
occurs,  it is too late to protect the  community.
USEPA (1992b) has an outstanding discussion on
applying uncertainty to  decision-making.   The
same data evaluation procedures apply to both the
marine  and freshwater systems.  The reader  is
referred to the statistical discussion in Chapter 9
(marine benthic community structure).
    Bivariate and multivariate analysis are often
applied in benthic studies to define relationships
between and among; variables. Examples of these
analyses include analysis of variance (ANOVA),
correlations,  regressions   (including  multiple
regressions), and the two-sample t-test.  A major
drawback to these methods is the assumption that
the data follow a statistical distribution such as a
normal or lognormal distribution.  This assump-
tion is often invalid when dealing with biological
populations and communities.
    Alternatively, nonparametric analyses may be
 conducted.  Such analyses  are not based on as-
 sumptions about a specific distribution of the data.
 Examples of such tests include the chi-square test,
 binomial tests, rank correlations, or tests compara-
 ble to  the t-test such as the Mann-Whitney test
 Whichever statistical methods are employed, all
 data assumptions  must  be  clearly stated  and
 objectives known.

 8.2.1.2.4   Necessary Hardware and Skills

     The hardware  needed  for field , collection
 includes samplers (e.g., dredges, dip-nets), sieves*
 benthic  macroihvertebrate  containers,  forceps,
 white  enamel  pans,  ethanol preservative, and
 appropriate personal gear (e.g., hip boots or chest-
 waders, life vest if needed, and fust aid kits). For
 the  laboratory, standard  biological laboratory
 equipment should be available, such as micro-
 scopes (both dissecting and compound), forceps,
 microscope slides and cover slips, ethanol, potassi-
 um hydroxide, mounting media, and sieves.  A
 personal computer (containing a 20-MB or larger
 hard drive) is important for storing  and analyzing
 the data.
     Trained benthic macroinvertebrate field biolo-
  gists  and taxonomists are  needed for benthic
                                                                                             8-13

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 Sediment Classification Methods Compendium
 community assessments. At least one should be
 proficient at identifications beyond  the family
 level.  That taxonomist should remain involved
 until the proficiency of the identifier in reaching
 family-level identifications is ensured. A mini-
 mum of a Master of Science degree in a related
 discipline is usually required for the taxonomist to
 have  learned  the, necessary skills.   However,
 adequate training is commonly available through
 taxonomy courses and workshops that can provide
 the necessary  proficiency without an advanced
 degree.  A demonstration of proficiency by accur-
 ately identifying a check sample prepared by U.S.
 EPA or a state agency is important.  A trained
 benthic  ecologist is necessary to compile  and
 interpret the data.  Although it would be ideal if
 the benthic  ecologist had a rigorous statistical
 background, consultation with a statistician should
 be adequate.

 8.2.13 Adequacy of Documentation

    There is ample documentation of both  field
 methods and analytical techniques.  The Journal of
 the North American Benthological Society is a prime
 source of this information, as is technical exchange
 at professional  meetings.  Furthermore, there is a
 large volume of published and unpublished material
 that documents  use of this method (USEPA 1992d,
 1991e, 1990g, 1989f, 1988a).

 8.22   Applicability of Method to Human
        Health, Aquatic Life, or Wildlife
        Protection

    This method is directly applicable to the protec-
 tion of aquatic life since it is based on direct mea-
 surements of benthic macroinvertebrates.   This
 method is directly applicable to the protection of
 those aquatic organisms (e.g., fish) and wildlife that
 directly feed on benthic macroinvertebrates (e.g.,
 small  mammals and  wading shorebirds).   It is
 indirectly applicable to other wildlife that depend on
 benthos at other levels  in the food chain.  This
 method is also indirectly applicable to the protection .
 of human health since benthic macroinvertebrates
 can serve as indicators of toxicant impacts that may
 affect humans via bioaccumulation pathways.
 8.2.3  Ability of Method to Generate
        Numerical Criteria for Specific
        Chemicals

     This method is  used in  conjunction  with
 sediment toxichy and chemistry data to charac-
 terize toxicant impacts and assist with determin-
 ing the appropriate levels at which the toxicants
 should be controlled.  By itself, however, this
 method would not be used to generate chemical-
 specific criteria.
 8.3 USEFULNESS

 8.3.1   Environmental Applicability

     Benthic macroinvertebrates have been rou-
 tinely used to assess environmental quality in a
 variety of geographical areas and ecoregions, as
 was discussed in Section 8.1.

 8.3.1.1 Suitability for Different Sediment Types

    Assessment of the freshwater benthic macro-
 invertebrate community structure is well suited
 for evaluating different sediment types since the
 benthos inhabit all substrates (Merrit and Cum-
 mins,  1984).   Comparisons should  be made
 among benthic communities of similar substrate
 since different types and numbers of organisms
 will inhabit different types  of substrates.

 8.3.1.2 Suitability for Different Chemicals or
        Classes  of Chemicals

    Benthic macroinvertebrate communities  are
 routinely used to assess potential impacts caused
 by many different chemicals or classes of chemi-
 cals.  In addition to the uses described in Sec-
 tion 8.1.1.1 of this chapter, many benthic organ-
 isms are used to indicate stresses from specific
 chemicals or classes of chemicals (Brinkhurst et
 al., 1968; Hart and Fuller, 1974; Saether, 1979;
 Simpson and Bode, 1980;  Wiederholm,  1980;
Bode and Simpson, unpublished; Winnell and
White,  1985;  Aagaard,  1986;  and  Fitchko,
 1986).
8-14

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                     8—Freshwater Benthic Macromvert&rate Community Structure and Function
8.3.1.3 Suitability for Predicting Effects on
       Different Organisms

    The  use of benthic macroinvertebrates as
indicator organisms has akeady been discussed.
Benthic  macroinvertebrates can be used to pre-
dict the  effects on other aquatic organisms be-
cause if the benthic macroinvertebrate communi-
ty is impacted, then the impact is likely to be, or
akeady has been, detrimental to other organisms.

8.3.1.4 Suitability for In-Place Pollutant Control

     Benthic macroinvertebrates will best indicate
 in place pollutant control needs  through a site-
 specific knowledge of  surface water  quality;,
 habitat quality, and sediment chemistry and toxici-
 ty   Alone, the benthic macroinvertebrates can be
 used to screen  for potential sources of sediment
 contamination based on spatial gradients in com-
 munity  structure, but  they should not  be used
 alone to definitively determine sediment quality or
 to  develop  chemical-specific guidelines.   The
 benthic data must be integrated with other avail-
 able data to  determine sediment quality using a
 "weight-of-evidence" approach.

 83.1.5  Suitability for Source Control

     Benthic macroinvertebrates have been exten-
 sively used for source characterization and control in
 many of  the state and U.S. EPA monitoring pro-
 grams involving spatial surveys upstream and down-
  stream of suspected sources (Ohio EPA, 1987; Bode
  and Novak, 1988; Courtemanch and Davies, 1988;
  Fiske,  1988; Maret,  1988; Penrose and Overton,
  1988;  Shakelford,  1988; USEPA, 1991c, 1988a,
  1988b; Fandrei,  1989).  If a detrimental change is
  detected in the benthic macroinvertebrate community
  and that change can be attributable to a source, then
  control measures can  be implemented through the
  NPDES permit program. Many states aggressively
  pursue this action.

  8.3.1.6 Suitability for Disposal Applications

      The discussion presented in Section 9.3.1.6 of
  Chapter  9 (marine benthic macroinvertebrate com-
munity  structure)  is  applicable to  fresh water.
Recently benthic community assessments have been
required by U.S. EPA Region V, as stated in the
Draft Interim Guidance for the Design and Execu-
tion of Sediment Sampling Efforts Relating to Navi-
gational Maintenance Dredging in Region V- May
1989 (USEPA, 1989d).   In this guidance, benthic
macroinvertebrate assessments are advised for areas
that are suitable for open-lake disposal or for sedi-
ments that are difficult to characterize. All benthic
community assessments will be made in concert with
sediment chemistry and toxicity evaluations.

 8.3.2   General Advantages and Limitations

     The advantage of using  the benthic macroin-
 vertebrates  community  assessment approach  to
 determining sediment quality is that it provides an
 economical and accurate indication of the health of
 the system under study, and it is based on direct
 observation rather than theoretically derived data.
 The major limitation is the difficulty in relating the
 findings to the presence of individual chemicals and
 specific concentrations of those chemicals for numer-
 ic in-place pollutant management.  This method
 should be integrated with sediment chemistry and
 toxicity information.

 83.2.1 Ease of Use

      The equipment requirements for benthic surveys
  is minimal and inexpensive compared to those for
  chemical/physical analyses or even toxicity tests.
  The organisms are easy to obtain, but difficult to sort
  and identity.   All  materials  needed for benthic
  assessments are easily obtained through chemical and
  biological supply companies and require no special
  mechanical setup or calibration.

  8.3.2.2 Relative Cost

      The cost for benthic macroinvertebrate assess-
  ments is economical compared to that for chemis-
  try or  lexicological  evaluations.   Ohio  EPA
  (1990a) provided a cost of about $700 to conduct
  a benthic assessment at one sample site. Howev-
  er, this cost included overhead (e.g., rent, office
  equipment),  all  travel expenses, time spent in the
                                                                                                8-15

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  Sediment Classification Methods Compendium
 field, and report preparation. Ohio EPA conducts
 artificial substrate (composite of five substrates)
 sampling in addition to natural substrate (multi-
 habitat)  sampling at each  site.   Their cost of
 $1,099 ($824 for artificial substrates and $275 for
 qualitative samples) was quite economical com-
 pared  to chemical/physical  testing ($1,653) or
 bioassay testing ($3,000 to $12,000) for each site.
 Plafkin et al. (1989) discussed staff requirements
 for sample collection and analysis.
     The  most expensive items are the samplers
 and the  microscopes to identify the organisms.
 However, most state programs and  contractors
 have this equipment available for other program
 needs. The fieldwork can be conducted during the
 time it takes to collect  a sediment sample.  The
 most time-consuming aspect is  the  laboratory
 sorting and identifications, which may average 11
 hours per site.  However, this process compares
 favorably with the amount of time required to set
 up and run a toxicity test or to prepare and ana-
 lyze chemical variables.

 5.3.2.3 Tendency to Be Conservative

     The  benthic  macroinvertebrate  community
 assessment provides a conservative measure, since
 the community is responding to both temporal and
 spatial perturbations. There are few chances, if any,
 of obtaining a result indicating a high-quality com- ,
 munity when an impact occurs.  Because of influ-
 ences other than sediment/water quality, it is more
 common to observe an impacted community when
 there is no sediment/water quality impact. Although
 the  primary  focus is  on community-level infor-
 mation, changes in individual populations could also
 be addressed. However, the ecological significance
 of population changes may not be evident until the
 community is affected..
    In a  review of surface water chemistry  and
 benthic macroinvertebrate community assessments
 over 800 water body segment sites in Ohio, biocri-
 teria based on benthic macroinvertebrates were more
 sensitive (conservative) indicators  of water quality
 (Ohio EPA, 1990b).  In  49.5 percent of the seg-
 ments, the benthic and fish  assessment revealed
 impacts not detected  by  chemical  water quality
 standards violations. In 47.4 percent of the sites, the.
  chemical and biological assessment supported one
  another. Only 2.8 percent of the sites did not have
  a biological impact when the chemistry indicated
  that there would be one.

  8.3.2.4 Level of Acceptance

      Benthic macroinvertebrate community assess-
  ments of sediment/water quality have been used in
  freshwater   systems   since   the   early   1900s
  (Richardson, 1928). Most of the methods employed
  today have been widely accepted for use, although
  the use of function measurements is not as well
  documented.   Perhaps the single most important
  demonstration of the level of acceptance of benthic
  assessments is the growing regulatory use and estab-
  lishment of numerical biological criteria in state
  water quality standards.

  8.3.2.5 Ability to Be Implemented by Laboratories
         with Typical Equipment and Handling
        Facilities                         .-  .

     The only special pieces of equipment required
 are the samplers and sieves, which are easily ob-
 tained from biological supply warehouses.  Most
 biological laboratories  will  have dissecting and
 compound microscopes, chemical reagents, micro-
 scope slides and cover slips, forceps, and any other
 materials needed.  The laboratory's capability to
. identify benthic macroinvertebrates is less common.
 Taxonomy  is not a widespread skill and is more
 likely to be  found in consulting firms than in analy-
 tical laboratories.

 8.3.2.6 Level of Effort Required to Generate
        Results

     Depending on the study objectives and level
 of effort needed,  results  can  be generated in
 written form in as little as 1 day (Plafkin et al.
 1989) or in several months. For example, Ohio
 EPA processes over 500 individual benthic sam-
 ples  each year,  identifies the organisms,  and
 prepares reports for regulatory use in less than 1
 year, with fewer than three full-tune employees in
 their benthic macroinvertebrate unit. The critical
 period is the turnaround time for the taxonomy.
8-16

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                      8—Freshwater Benthic Macrotnvertebrate Community Structure and Function
With artificial substrates,  an additional 6-week
colonization  period is  required; unless a  rapid
assessment or moderate sized study is done, a
written report including interpretation of results
will typically require between  6 months  and 1
year.        .

8.3.2.7 Degree to Which Results Lend
        Themselves to Interpretation

    It is  never advisable  to  have an individual
without training in benthic ecology interpret benthic
data. Once the benthic  ecologist provides a report
with recommendations, the results can be  easily
implemented into a  management  strategy.   Al-
though several numerical indexes that appear simple
to use are available, data interpretation reliejs on all
of the information generated for a study, including
chemical, physical, and lexicological measurements,
as well as indicator organisms and function mea-
sures.

8.3.2.8 Degree of Environmental Applicability

     Benthic macroinvertebrate community structure
 and function is used extensively to evaluate sedi-
 ment and water quality .and characterize impacts in
 lotic and lentic freshwater ecosystems.

 8.3.2.9 Degree of Accuracy and Precision

     Since benthic macroinvertebrates are measured
 directly, this method is highly accurate for charac-
 terizing sediment/water quality effects on aquatic
 life.  There is little chance, if any, that  a high-
 quality community  will  be indicated  when an
 impact actually occurs (Type II error with a null
 hypothesis of no  community  change). Because of
 influences other than sediment/water quality, it is
 more common to observe an  impacted community
 when  there is no sediment/quality impact (Type I
 error-with  a null  hypothesis  of no community
 change).  For environmental pollution  control, a
 Type n error is much  more serious than a Type I
 error,  which is conservative.  To reduce the possi-
 bility of a Type II error, the data (including chem-
 istry and toxicity) must be interpreted by a trained
 benthic .ecologis;,    Resh (1988)  and  USEPA
(1992b) reviewed the levels of accuracy and preci-
sion for several of the data analysis techniques.
    To ensure as much accuracy and precision in
the data as possible, a detailed Quality Assurance
Program Plan should be established and followed.
Careful and consistent field and laboratory proto-
,cols are necessary. It is also necessary to sample
during optimal conditions, which can minimize the
effects of natural variations in the data. However,
the natural variability,  especially seasonal, is re-
duced when using  a community-level  approach.
rather than a population-level approach.
 8.4  STATUS

     Sections 8.1.1 (Current Uses) and 8.3 (Useful-
 ness) describe the status of the discipline.

 8.4.1   Extent of Use

     This method is widely used in both regulatory
 and  nonregulatory  sediment  arid water quality
 programs: It has been used to assess impacts due
 to organic enrichment and a variety of chemical
 classes in both lotic and lentic systems.  Benthic
 macroinvertebrate community assessments are the
 most widely used instream biological measures in
 state water quality programs.

 8.4.2   Extent to Which Approach Has Been
         Field-Validated

     Since it is an  in situ study, field validation
 occurs when  the approach can  consistently  and
 accurately assess environmental quality.   Most
 benthic studies employ reference stations and rely
 on other environmental data to validate the method.
 The documentation provided in this paper should
 •present adequate documentation  of the  method's
 validity.

 8.4.3   Reasons for Limited Use

      Benthic macroinvertebrate community assess-
 ments are very  common in freshwater systems
 because of their relatively low cost and high infor-
 mation output         .
                                                                                               8-17

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  Sediment Classification Methods Compendium
 8.4.4   Outlook for Future Use and Amount
         of Development Yet Needed

     The outlook  for the future use of benthic
 macroinvertebrate community structure and func-
 tion in sediment quality assessment is very good
 because of the recognition  that benthic macro-
 invertebrates provide substantial information that
 the* chemistry  and toxicity data alone  cannot
 provide. With the Clean Water Act mandate to
 maintain and restore biological integrity, benthic
 community assessments can help determine wheth-
 er sediment quality is  impairing the designated
 uses and biotic integrity.  With the increasing
 reliance on numerical biocriteria, additional sedi-
 ment quality problems will be identified.  The area
 where development is most needed is in combin-
 ing benthic community assessments with chemical
 and toxicological data in  an integrated approach
 for assessing  sediment  quality.  Lot addition,  the
 functional measures, which also hold much prom-
 ise for sediment assessments, need to be validated
 more thoroughly.
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                     8—Freshwater Benthic Macroinvertebrate Community Structure and function
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                     8—Freshwater Benthic Macroinvertebrate Community Structure and Function
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 Wihlm, J.L.  1970.  Range of diversity in benthic
     macroinvertebrate populations. J. Wat. Pollut.
     Control Fed.  42:R221-224.
 Winnell, M.H., and  D.S. White.  1985.  Trophic
     status of southeastern Lake Michigan based on
     the Chironomidae (Diptera).  J. Great Lakes.
     Res.  11:540-548.
 Winner, R.W., J.S. Van Dyke, N.  Caris, and M.P.
     Farrell.   1975.   Response of  a macroinverte-
    brate fauna to a  copper gradient in an experi-
    mentally-polluted Stream.    Verb.  Internal.
    Verein. Limnol.  19:2121-2127.
 Winner,  R.W.,  M.W. Boesel, and  M.P. Farrell.
    1980. Insect community structure as an index
    of heavy-metal pollution  in lotic ecosystems.
    Can. J. Fish. Aquat Sci. 37:647-655.
Woodiwiss, F.S. 1964.  The biological system of
    stream classification used by the Trent  River
    Board, Chem. Lid. 11:443-447.
Yasuno, M., Y. Sugaya, and T. Iwakuma.   1985.
    Effects of insecticides on the benthic communi-
    ty in a model stream. Environ.  Pollut (Ser. A)
    38:31-43.
8-26

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         ^CHAPTER 9
Marine Benthic Community  Structure

Assessment
Betsy Strlplln, Gary Braun, and Gordon Bllyard
Tetra Tech, Inc.                             nnnn>,
11820 Northup Way, Suite WOE, Bellevue, WA 98005
(206)822-9596
    Benthic  communities are communities  of
 organisms that live in or on the sediment. Inmost
 benthic community structure assessments, primary
 emphasis is placed on determining the species that
 are present and  the distribution  of individuals
 among those species. These community attributes
 are emphasized  largely  for  pragmatic reasons.
 Although it is relatively simple to collect, identify,
 and  enumerate benthic. organisms, it is very
 difficult to determine first-hand the spatial distri-
 butions of species  and individuals within  the
 benthic habitat, or the  functional interactions that
 occur among the resident organisms or between
 the resident organisms and  the abiotic habitat.
 Hence, information on  benthic community compo-
 sition and abundance is typically used in conjunc-
 tion with information in the scientific literature to
 ,infer the distributions of species and individuals in
 three-dimensional space and the functional attri-
 butes of the community. Because all of the major
  structural and functional attributes  of  benthic
  communities are affected by sediment quality ^in
  generally predictable  ways, benthic  community
  structure assessment is a valuable tool for evaluat-
  ing sediment quality  and its effects  on  a major
  biological component of marine, estuarine, and
  freshwater ecosystems.
      Benthic habitats may be broadly divided into.
  hard-bottom  habitats and soft-bottom  habitats.
  Many types of each exist in marine, estuarine, and
  freshwater  ecosystems.   Hard-bottom  habitats
  include rocky shorelines and bottoms of lentic and
  lotic systems, rocky intertidal and subtidal habitats.
  in marine and estuarine systems, and coral reefs.
  Soft-bottom habitats  include mud and sand habi-
  tats in marine, estuarine, and freshwater systems-
  marine, estuarine, and freshwater macrophyte
  beds;  freshwater  wetlands; and  estuarine  salt
marshes.  Each of these habitats requires different
sample collection  methods and different survey
design considerations.   The  emphasis of this
chapter  is  on assessments of marine benthic
community structure in soft-bottom habitats as an
indicator of sediment quality. Freshwater benthic
invertebrate community structure is discussed in
Chapter 8.                                  ,
 9.1  SPECIFIC APPLICATIONS

    Assessment of benthic community structure is
 an in situ method that can be used alone, as part
 of other approaches [e.g.j Sediment Quality Triad
 (see Chapter 10) and Apparent Effects Threshold
 (AET) (see Chapter 11)], or in combination with
 other sediment assessment techniques (e.g., sedi-
 ment toxicity bioassays). It is commonly used in
 three ways to assess impacts to benthic communi-
 ties and sediment quality:

     •  To compare test and reference stations,
        for the purpose of determining the spatial
        extent and magnitude of .such impacts;

     • To identify spatial gradients of impacts;
        and
     • To identify temporal trends at the same
        locations through time.

     By definition, benthic communities include all
  organisms living on or in the bottom substrate.
  For  practical reasons,  assessments of  benthic
  community structure in soft sediments usually rely
  on the macrofauna (i.e., organisms retained on a
  1.0-  or 0.5-mm sieve) and to a lesser extent the
  meiofauna (i.e., multicellular organisms that pass

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 Sediment Classification Methods Compendium
 through a 1.0- or 0.5-mm sieve). Reasons for the
 more limited use of meiofauna are twofold:

     •   Although they may be sampled quantita-
         tively, their  small size makes  working
         with them difficult, and the taxonomy of
         many of the groups (e.g.,  nematodes) is
         not well known.

     •   The functional attributes of the  various
         meiofaunal taxa are poorly known, and it
         is therefore difficult to interpret  the im-
         portance of the presence or absence of the
         various taxa in relation to  environmental
         quality.   (For example, knowledge of
         meiofaunal taxa that respond positively or
         negatively to  organic enrichment of  the
         sediments is extremely limited.)

 Difficulties in quantitatively sampling other size
 classes of benthic organisms such as the mega-
 fauna (i.e.,  large  organisms that  are typically
 measured in centimeters) and the microfauna (i.e.,
 microbes) usually preclude them from consider-
 ation in  assessments of benthic community struc-
 ture. Furthermore, although the functional impor-
 tance of sediment microbes has been studied, their
 structural and functional characteristics have not
 been used as indicators of sediment quality.


 9.1.1 Current Use

    Assessments  of benthic community structure
 have been used to  describe reference conditions,
 baseline  conditions, and the effects of natural and
 anthropogenic disturbances. Selected examples of
 current uses of this approach are provided below.
    Organic Enrichment—Pearson and Rosen-
 berg  (1978)  performed an  extensive  review  of
 benthic  community succession  in relation  to
 organic enrichment of marine and estuarine sedi-
 ments.  Based on that review, they developed a
 generalized model of structural community chang-
 es (i.e., numbers of species, abundances, biomass)
 in relation to  organic enrichment, and identified
 opportunistic and pollution-tolerant species that
 are indicative of organic enrichment.  Concepts
 developed by Pearson and Rosenberg (1978) have
subsequently been used by many investigators to
 assess the degree of organic enrichment that has
 occurred in a variety of soft-bottom habitats. For
 example, Dauer and Conner (1980) assessed the
 effects of sewage inputs on  benihic polychaete
 populations  in a Florida  estuary  by collecting
 information  on the total number of individuals,
 total biomass,  and average number of species.
 They compared the sewage-affected .site with a
 reference  site  and examined the response  of
 individual species to  organic .enrichment.   In
 another study in Florida, Grizzle (1984) identified
 indicator species based on life history responses to
 organic enrichment and other  physicochemical
 changes.  The taxa identified as indicator species
 in enriched areas were generally characterized by
 opportunistic life history strategies.  Vidakovic
 (1983) assessed the influence of domestic sewage
 on the density and distribution of meiofauna in the
 Northern Adriatic  Sea.  He concluded that raw
 domestic sewage did not have a negative influence
 on the density and distribution of meiofauna, but
 the nematode/copepod ratio (Parker,  1975) indicat-
 ed that these stations were under  stress.
    Contamination Due to Toxic Metals and
 Metalloids—Rygg (1985a, 1986) assessed benthic
 community structure in Norwegian fjords where
 the disposal of mine tailings had resulted in metals
 contamination  of  the  sediment.   His  studies
 showed an inverse relationship between concentra-
 tions of metals in  the sediment and the species
 richness and  abundance of the  benthic macro-
 invertebrate fauna.  Bryan et al. (1987) examined
 population distributions of  the  oyster  Ostrea
 edulisy, the polychaete Nereis diversicolor, and
 the cockle Cerastoderma  edule in relation to
 wastes from  metals mining in the Fal Estuary. ,
 They concluded that the distribution of species is
 dependent on their ability to tolerate copper and
 zinc,  and on the capabilities of a population to
 develop a resistance to metals and thereby main-
 tain their original distribution range.
    Contamination Due to Toxic Organic Com-
 pounds—Toxic organic compounds are frequently
 associated with municipal discharges,  industrial
 effluents, and storm drains. These discharges may
 also result in organic enrichment and contamina-
 tion  by metals or metalloids.   The following
benthic studies provided evaluations of sediment
9-2

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                                            9—Marine Benthic Community Structure Assessment
quality in areas primarily affected by toxic organic
compounds:
    •   Creosote  contamination.;   Tagatz ef flJ.
        (1983) examined the benthic communities
        that colonized uncontaminated sediments
        and sediments contaminated  with three
        different concentrations of creosote (177,
        844, and 4,420 ng/g) in field and labora-
        tory aquaria to  assess  the  effects  of
        marine-grade  creosote  on   community
        structure.   Numbers of individuals and
        numbers  of  species  in  field-colonized
        communities were significantly lower in
   •     all three creosote-contaminated sediments
        than in the controls.  In the laboratory-
      ,  colonized  communities  only  the two
        higher creosote concentrations  had re-
        duced numbers of individuals and species.
       - Distribution of individuals within species
        was similar for the  laboratory and field
        'assemblages of animals.
     •  Oil contamination. Elmgreh et al. (1983)
        determined that acute effects of the Tsesis
        oil spill were noted after 16 days on both
        the macrofauna and meiofauna.   Initial
        recovery was noted 2 yr after the spill.
        However, the,authors predicted that com-
        plete recovery would require  at least 5 yr.
        Jackson et al. (1989) investigated the ef-
         fects  of spilled  oil on  the  Panamanian
         coast and found that shallow  subtidal reef
         corals and the infauna of seagrass beds
         had   experienced   extensive  mortality.
         After  1.5 yr, only some  of the organisms
         in areas exposed to the open  sea  had
         recovered.  Clifton et al.  (1984) per-
         formed field experiments in Willapa Bay,
         Washington, and found that  oil in the
         sediments modified the burrowing behav-
         ior of infaunal benthos.

      Dredging and Construction-Related Activi-
  ties—Swartz  et  al.  (1980) examined  species
  richness   and species  abundances  just  before
  dredging occurred in Yaquina Bay, Oregon, and
  for 2 yr after dredging.   Benthic community
  recolonization was followed from the appearance
of opportunistic taxa through their replacement by
less tolerant taxa, Rhoads et al (1978) examined
the influence of dredge-spoil disposal on benthic
infaunal  succession in  Long  Island Sound by
classifying species into groups based on their ap-
pearance in a disturbed area. They suggested that
the "equilibrium community is less productive
than a pioneering stage" and suggested  that pro-
ductivity  may be  enhanced  through  managed
disturbances.   The abundance of  polychaetes,
molluscs, and crustaceans is currently used to help
assess potential  biological effects  of dredged   ,
material disposal by the Puget Sound Dredged
Disposal Analysis Program (SAIC, 1991; Striplin
ef of., 1991).                            .
    Natural  Disturbances—Most   studies  of
natural disturbances have assessed the recovery of
benthic communities after the disturbance  (e.g.,
 large storms and associated wave activity, oxygen
 depletion, salinity reductions,  El  Nino).   For
 example,  Dobbs and  Vozarik (1983) sampled
 stations before and after  Storm David and ob-
 served that the number of species decreased after
 the storm.  They also documented changes in the
 rank order of the dominant  taxa.   Santos and
 Simon (1980) examined defaunation of benthic
 communities before, during,  and  after annual
 hypoxia in Biscayne Bay. They documented that
 recolonization  occurs   fairly  rapidly  after the
 defaunation period.  Oscillations in macrobenfhic
 populations in the shallow waters of the Peruvian
 coast were examined by Tarazona et al (1988).
 Fluctuations in density, biomass, species composi-
 tion, and diversity were attributed to the El Nino  .
 of 1982-1983.
     Assessment of benthic community structure is
 also  used as  a component  of other sediment
 quality assessment tools.  Along with sediment
 chemistry .and sediment toxicity bioassays, it is
  one of three components of the Sediment Quality
 Triad (see Chapter 10). It is also a component of
  the  Apparent  Effects  Threshold  approach (see
  Chapter 11).


  9.1.2 Potential Use
      To  date,  benthic community assessments
  performed to  evaluate  sediment quality have
                                                                                              9-3

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 Sediment Classification Methods Compendium
 focused on the relationships between community
 variables  (e.g.,  numbers  of species, total abun-
 dance, biomass) and measures of sediment quality
 (e.g., organic content, concentrations of chemical
 contaminants). Only for organic enrichment have
 individual species been identified that are indica-
 tive of various degrees of sediment alteration [see
 for example Pearson and Rosenberg (1978), Word
 et al, (1977)].  Moreover, for  only a very few
 species has the autecological relationship between
 organic enrichment  of the  sediments  and  an
 individual species been explored.  [For example,
 Fabrikant (1984) explored the autecology of the
 bivalve mollusc Parvilucina tenuisculpta in rela-
 tion to organic enrichment of the sediments in the
 Southern California Bight.]  A tremendous poten-
 tial  exists, however,  for identifying species that
 are  indicative  (by  their  persistence, enhanced
 abundance, reduced  abundance, or absence)  of
 sediment contaminants at various concentrations.
 The identification of such taxa will not be simple
 because of the complex ecological interactions that
 occur within benthic communities,  and because
 sediments  are  frequently contaminated with  a
 mixture of chemicals. A first  step  in this process
 might be to attempt to identify species or suites of
 species  that could be used to separate the effects
 of sediment organic enrichment from sediment
 contamination by toxic substances.
    Another potential use of benthic community
 assessments  would  be to  predict  recovery  of
 benthic habitats following the  execution of reme-
 dial  actions at contaminated sites. To date, it has
 not been possible to use extant  benthic community
 structure to predict  recovery  because the  only
 model that relates benthic  community structure to
 sediment quality [i.e., the Pearson and Rosenberg
 (1978) model] is not quantitative.  Quantification
 of this model and the development of quantitative
 models  for  other sediment contaminants will be
 required before  benthic community  assessments
 can be used to predict sediment quality. A valu-
 able byproduct  of such  models would be the .
ability to predict the capacity of the remediated
area to support higher trophic level organisms that
forage on  benthic organisms, including commer-
cially and recreationally harvested demersal fishes.
 9.2  DESCRIPTION

 9.2.1  Description of the Method
    An assessment of benthic community structure
 typically  involves a field survey  that includes
 replicated sampling  at each station; sorting  and
 identification of the organisms to species or lowest
 possible taxon; analyses of the numbers of taxa,
 numbers of individuals, and sometimes biomass in
 each sample;  and identification of the dominant
 taxa.  Results of the field survey are then inter-
 preted in conjunction with other sediment vari-
 ables  (e.g., sediment  grain  size,  total  organic
 carbon) that were collected concurrently with the
 benthic samples.

 9.2.1.1 Objectives and Assumptions

    The   objective of  the benthic community
 structure  approach is to identify degraded  and
 potentially degraded sediments by examining the
 communities  of   organisms  that inhabit  those
 sediments. This empirical approach assumes the
 following:

    •  Because  benthic infauna are  generally
       sedentary, benthic community structure
       reflects the chemical and physical envi-
       ronment at the sampling location.

    • Benthic community  structure  may  be
       altered in a predictable manner over time
       and space by chemical or physical distur-
       bances.

    •  The execution  of proper data collection
       and analysis methods can reduce natural
       variability of benthic  infaunal data and
       enable the detection of trends in sediment
       quality.

9.2.1.2 Level of Effort

   The level of effort required to assess benthic
community structure is relatively high. Regardless
of the analytical  methods, a field survey is  re-
quired to collect the organisms.  The sorting and
identification process is labor-intensive and usu-
9-4

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                                            9—Marine Benthic Community Structure Assessment
ally expensive. Program objectives will determine
whether the data analyses are simple or complex.

9.2.1.2.1 Type of Sampling Required

The type of sampling required to collect benthic
organisms  is dependent on the objectives of the
sampling program and on the area under study.
Usually, the  objective  of a benthic sampling
program is to study the characteristics of and the
variation in the benthic community that occupies
specific sampling stations.  In this case, all organ-
isms present in  the sediment at that location are
sampled together:  those that normally reside in
the surface few centimeters of sediment and those
that normally reside deeper in the sediment (e.g.,
5-15 cm below the surface).  In some instances, a
sampling program may have  a different objective.
For example, sampling for the Benthic Resources
Analysis Technique (BRAT) (Lunz and Kendall,
 1982) involves  collecting box core samples and
 determining the biomass (and possibly the com-
 munities)  present in specific sediment strata (i.e.,
 0-2 cm, 2-5 cm, 5-10 cm, and 10-15 cm below the
 sediment surface). In that technique, the benthic
 data are compared with  the benthic organisms
 consumed by bottom-dwelling fish (as determined
 by  gut  content analyses of fish captured in the
 same area) to  determine the food value of the
 benthos.
     Characteristics of the area under study also
 influence the type of sampling..  In  intertidal or
 littoral environments where  sampling stations can
 be occupied by walking to  the site, samples are
 usually collected using a  hand-held  corer.  If
 stations are located in subtidal areas, then remote
 sampling from a vessel is performed using a box
 corer or grab, sampler.  Sediment grain size may
 influence  final selection  of the sampler.  Some
 samplers (i.e., many box corers) perform poorly in
  sandy sediments, whereas others (i.e., van Veen
  grab, Smith-Mclntyre grab)  perform adequately in
  a greater range  of  sediment types  (i.e., fine to
  medium  sand,  silt, silty clay).  Methods and
  equipment for sampling infaunal communities are
  further described in several publications (Word,
  1976; Swartz, 1978; Eleftheriou and Holme, 1984;
  Nalepa et al.,  1988).  Blomqvist (1991) provides
an  extensive  review of  quantitative  .sampling
methods,  including  a  detailed bibliography of
pertinent papers.
    Program objectives and knowledge of benthic
communities in the study  area  will  influence
selection of the sieve size through which sediment
samples will be washed.  It is important that the
sieve mesh sizes be appropriate for the community
under study (e.g., 64/an for meiofauna, 0.5 or 1.0
mm.for macrofauna).  Generally, the chances of
retaining most macrofauna species and individuals
(and therefore increasing  sampling accuracy) are
improved by the use of  a  finer  mesh {but, see
Bishop and Hartley, 1986).  However, sieve size
is, an important determinant of the cost and. level
of  effort necessary to obtain  quantitative  data.
Very little difference in the field processing time >.
exists between use  of  a  0.5-mm  and  a 1.0-mm
sieve when sieving sediments finer than coarse
sand, but laboratory analyses are much more tune-
 consuming when the smaller mesh is used because
 it retains more abiotic materials and many smaller
 organisms.

 9.2.1.2.2 Methods

     Methods for collecting data on benthic com-
 munity structure are divided into three categories:
 program design,  field methods,  and laboratory
 methods.   Each  of these categories  is briefly
 discussed below.
     Program design  includes the  selection  of
 station locations,  level  of replication,  type  of
 sampler, screen size, data analysis methods (dis-
 cussed later), and quality  assurance/quality control
 (QA/QC) procedures.   The selection of station
 locations will directly  influence the usefulness of
 the resulting data. Stations that will be compared
 to one  another  (including  reference stations)
 should be situated in areas with similar hydro-
 graphy, water depth,  and grain size to minimize
 the  natural  variability  in benthic  community
 composition that can be attributed to these factors;
 However, such station placement is  not always
  attainable because of altered grain size distribu-
  tions that often result from contaminant sources.
      Selection of the number of replicates is an
  important component of program design because
                                                                                               9-5

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  Sediment Classification Methods Compendium
 the accuracy  and precision with  which benthic
 community variables are estimated depend in part
 on the size of "the sample (including all replicates).
 For example,  the abundance of a single taxon is
 generally a less accurate descriptive variable than
 is the abundance of the total taxa because of the
 greater variability typically associated with one
 taxon  in comparison with the sum of all taxa.
 The total area sampled  among the replicates at
 each station should be large enough to estimate a
 given variable within the limits of accuracy and
 precision that are acceptable to meet study objec-
 tives. A single sample may be useful for general
 distributional or trends analyses (Cuff and Cole-
 man, 1979), but the inherent patchiness of benthic
 communities  makes  collection of a  sufficient
 number of replicate samples (a minimum of 3-5,
 depending on study objectives and sampler area)
 necessary  to  ensure  statistical reliability  (see
 Elliott,  1977).   Within  a study  area, adequate
 sample size may be determined by maximizing the
 number of species collected or by minimizing the
 error associated with the mean for the variable in
 question (Gonor and  Kemp,  1978).  Additional
 research on replication is presently being conduct-
 ed by EPA in Newport, Oregon, under the direc-
 tion of S. Ferraro (Swartz, R.C., 15 March 1989,
 personal communication).
    Power analysis can assist in determining the
 appropriate number of replicates. A power analy-
 sis includes consideration of the minimum detect-
 able difference in selected biological  variables
 (i.e., the minimum difference in mean values of a
 variable at  several stations that can be detected
 statistically, given a certain level  of variability
 about those mean values) and the  power of the
 statistical test to be used.  The power of the test is
 especially important because it  defines the proba-
 bility of correctly detecting experimental effects
 (e.g., differences in biological  variables among
 sampling stations).   For a specified  variance
 associated with a biological variable, the statistical
 power of a test  and the  minimum detectable
 difference among sampling areas can be expressed
 as a function of sample size.   The allocation of
 sampling resources  (stations,   replication,  and
 frequency) can  then be determined with regard to
available resources,  practicality of  design, and
  desired  sensitivity of the subsequent analyses.
  Discussions  and examples of this approach are
  found in Winer (1971), Saila et al. (1976), Cohen
  (1977), Moore and McLaughlin (1978); Bros and
  Cowell (1987), Ferraro  et al (1989), Kronberg
  (1987), Tetra Tech (1987),  Self and Mauritsen
  (1988), and Vezina (1988).
     A potential drawback to use of power analysis
  is that it requires a priori knowledge of variability
  in the benthic communities that will be studied.
  If such variability  is not known and cannot be
  estimated, then the  number of replicates will
  probably  reflect either  funding  limitations  or
  generally approved sampling methods.  For exam-
  ple,  Eleftheriou and  Holme  (1984) and Swartz
  (1978) recommend that an  area  of 0.5 m2 be
  sampled to assess species composition in coastal
  and  estuarine regions.  Most studies of benthic
  community structure routinely involve five repli-
  cate  0.1-m2 grab samples.  A single 0.1-m2 grab
  sample may be sufficient to obtain "useful descrip-
  tive  information"  for use  in cluster analyses
  (Word, 1976).   However, a  single sample pre-
  cludes direct estimates of within-group variance
  for statistical analyses.  Because individuals are
  distributed logarithmically among the species of a
 benthic community (Preston, 1948; Sanders, 1968;
 Gray and Mirza, 1979), species collected in the
 second  and  successive replicates that were not
 collected in  any of the previous replicates most
 often will be numerically "rare."  Note that "rare"
 is not synonymous with "unimportant." Hence, a
 single 0.1-m2 sample is generally not adequate to
 characterize  benthic  community structure  and
 function. In general, five 0.1-m2 grab samples are
 recommended for determining benthic community
 structure,  unless evaluation of site-specific data
 (i.e.,  a power analysis) indicates that sufficient
 sensitivity can be obtained with fewer samples, or
 that a greater number  is required due to extreme
 spatial  heterogeneity.   (Note;that at least three
'samples are   required for parametric  statistical
 analyses.)
    Another aspect of program design is selection
 of the appropriate degree of navigational accuracy.
 For baseline  or  distributional  studies, repeatable
 station  location may not be a high priority, and
 methods such as Loran  C may be  sufficient
9-6

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                                            3—Marine Benthic Community Structure Assessment
However, for monitoring programs where reoccu-
pation of exact stations is important (e.g., disposal
site  monitoring),  a more accurate positioning
.method (e.g., an  electronic distance-measuring
device or Mini-Ranger) may be required.
    A quantitative sampling device and an appro-
priate mesh size must be selected to ensure that
size classes of organisms appropriate for assessing
sediment quality are collected.  Selection of a
sampler and sieve are discussed above, in Section
9.2.1.2.1.
     Field and laboratory  methods must  be con-
ducted according  to rigorous QA/QC protocols.
Field methods  include collecting, sieving,  and
preserving the  samples.   Samples are typically
preserved in a solution of 10  percent  buffered
 formalin for  at least 24 h. Laboratory  methods
 include rinsing the formalin solution from  the
 samples within 7-10 days, followed by storage in
 70 percent ethanol.  Samples are sorted under a
 dissecting microscope during which all organisms
 are removed from the samples and placed in vials
 for identification and  enumeration of individual
 taxa.  The time required to  sort and identify a
 benthic sample varies greatly depending  on the
 sieve size, sample area, and sediment composition.
 Sorting may take as  little as  1 h for  a  0.1-m
 sample sieved through a 1.0-mm screen, or as
 much  as 12 h  if wood chips or other debris are
 present.  The time needed to identify organisms in
 a sample depends on the number of organisms
 (which is a function of sieve size, habitat, or
 degree of contamination) and number of taxa
 present. The number  of hours needed to identify
  organisms in a sample may range from  1  to over
  10 h.                      ,-.--'.
      In addition to the collection of samples for
  analysis of benthic community structure, separate
  sediment samples should be collected at all sta-
  tions  for conventional sediment chemistry vari-
  ables (e.g., sediment organic content, sediment
  grain size distribution).   Because organic carbon
  content and sediment grain  size naturally affect
  the composition of benthic communities, measure-
  ment of these variables will  assist in determining
  whether benthic communities are  affected by
  reduced sediment quality.
9.2.1.2.3 Types of Data Required

    The two primary structural attributes of any
benthic community are the distribution of species
and individuals in three-dimensional  space, and
the distribution of individuals among species and
higher taxa. Given an understanding of these two
structural attributes, it is possible to infer function-
al-attributes of the benthic community, including
trophic  relationships,   primary and  secondary
productivity, and interactions between the resident
biota and the abiotic habitat.  The  data required
for analysis of the structural and functional attrib-
utes include the  number  of taxa (identifications
should be to the lowest taxonomic level possible),
the abundance of each taxon, biomass (depending
on program objectives), and conventional sediment
chemistry variables.  However^ collection of the
appropriate data does not ensure proper evaluation
of the structural and  functional attributes.  The
selection and implementation of data analyses are
 equally  important, and are discussed in the re-
 mainder of this section.  The  data analyses pre-
 sented in this section address primarily structural
 components of benthic communities. However,
 functional attributes can be inferred from many of
 those structural attributes.
     Various types of data  analyses  are .used to
 describe benthic community structure, depending
 on the objectives  of  the particular program.
 However, several descriptive values are common
 to most program objectives.  All organisms col-
 lected in each sample are enumerated (i.e., total
 abundance), and abundances of major taxonomic
 groups are usually summarized. Depending on the
 level of identification, abundances of individual
 taxa, numbers of taxa, and lists and abundances of
 pollution-tolerant and pollution-sensitive taxa in
 each sample may be developed. Biomass of major
 taxonomic groups and total biomass are sometimes
 reported.  The  composition of the numerically
  dominant  taxa are analyzed when species level
 •identifications are performed. In addition, descrip-
  tive indexes such as diversity [the distribution of
  individuals among species; see Washington (1984)
  for additional definitions of diversity], evenness
  (the evenness with which individuals are distribut-
  ed among taxa), and dominance (the degree to
                                                                                               9-7

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 Sediment Classification Methods Compendium
 which one or a few species dominate the commu-
 nity) are usually calculated.         ,
    Most  programs  evaluate the  temporal or
 spatial differences in benthic community structure.
 Typically, comparisons of one or more indexes are
 made at the same station over time and compared
 to  a baseline value,  or  comparisons are made
 between stations in a study area and stations in a
 reference area. If an adequate number of samples
 is collected (i.e., three or more), statistical tests
 such as t-tests or Analysis of Variance (ANOVA)
 (or  their  nonparametric  analogues)  are often
 performed to determine whether significant spatial
 or  temporal  differences  exist  among  benthic
 communities.
    Besides univariate (i.e., single-variable) statis-
 tical analyses, multivariate (i.e., multiple-variable)
 analyses are frequently performed (e.g., Boesch,
 1977; Green and Vascotto, 1978; Gauch, 1982;
 Shin, 1982; Long and Lewis, 1987; Ibanez and
 Dauvin, 1988; Nemec and Brinkhurst, 1988a,b;
 Stephenson and Mackie,  1988).   Multivariate
 analyses  include classification   methods  (i.e.,
 grouping similar stations into clusters) and ordina-
 tion methods [i.e., representing sample or species
 relationships as faithfully  as  possible in a low-
 dimensional (two-four dimensions) space].  [See
 Gauch  (1982) for an overview  of multivariate
 methods.] Multivariate techniques group data and
 display them on a two-dimensional plot or dendro-
 gram so that stations exhibiting similar communi-
 ties  are located  closer to one another than to
 stations with dissimilar communities. The numeri-
 cal and graphical results can then be compared
 with  physical and chemical  data collected con-
 currently  to  determine whether  those  variables
 correlate with trends in benthic communities. A
 commonly used classification technique involves
 first computing a matrix of similarity indexes that
 represent the degree of similarity  in species com-
 position between two stations. Commonly used
 similarity indexes include Bray-Curtis,  Canberra
 metric, and  Euclidian distance  indexes.   The
 similarity matrix  is then entered into a clustering
 algorithm ' (e.g.,  pair-wise averaging,  flexible
sorting) to produce a dendrogram depicting simi-
larities among stations. Commonly used ordina-
tion   techniques  include  principal  components
 analysis, detrended correspondence analysis, and
 discriminant function analysis.   Bernstein  and
 Smith (1986)  developed  an index  of benthic
 community change along pollution gradients that
 is derived from results of ordination analysis.  The
 index (called Index 5) is  a measure of change
 from reference conditions.
    Benthic  community surveys generate large
 data  matrices.  These data matrices are often
 reduced  by • the  elimination of  certain species
 (Boesch, 1977) prior to performing multivariate
 analyses.  A variety of methods exist for reducing
 data matrices (see Stephenson et al., 1970, 1972,
 1974; Day et al., 1971; Clifford and Stephenson,
 1975).
    Both  parametric  statistical tests  and multi-
 variate analyses may involve data transformations.
 Transformations  of the original  data  may be
 necessary for one or  more of the following rea-
 sons:

    • Benthic data sets are usually characterized
       by large abundances of a few species and
       small abundances of many species;

    •. The  distribution of, individuals among
      , species tends  to be lognormal; and

    • Sampling  effort  may  be  inconsistent
       (Boesch, 1977).

The two basic types of transformations are strict
transformations  and   standardizations.     Strict
transformations  are alterations  of the original
values (e.g., species abundances) without reference
to the range of values within the data. Commonly
used transformations are square root, logarithmic,
and arcsine (Sokal and Rohlf, 1981). Standardiza-
tions  are alterations that depend on some property
of the data under consideration. A common stan-
dardization is the conversion of values to percent-
ages.
    Benthic data are transformed to better meet
the assumptions of parametric tests (e.g., normali-
ty,  homogeneity of variances).   In multivariate
analyses, data are  often transformed using  loga-
rithms [e.g., log (x+1)] because of the presence of
zero scores.  This transformation is also applied
9-8

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                                            9—Marine Venthic Community Structure Assessment
when population variance estimates are positively
correlated with mean values (Sokal and Rohlf,
1981). Clifford and Stephenson (1975) discuss in
detail the effects of transformations on commonly
used resemblance measures.
    Benthic community structure is usually com-
pared with  chemical and physical data that are
collected concurrently.  These comparisons may
take the form of simple linear correlations, corre-
lations with cluster groups, or correlations using
multivariate techniques  such  as discriminant
analyses. Multiple discriminant analysis attempts
to isolate groups of similar stations so that vari-
ables responsible for the separation of groups can
be identified. Results may be used to determine
whether differences in community structure are
due to variations in sediment grain size, variations
 in other physical characteristics of the environ-
 ment, or changes in sediment quality  due to toxic
 substances  or organic materials.;
     The use of different  methods and analyses
 may result  in different interpretations of the same
 data.   For  example, use of the same data with
 different standardization methods in  a classifica-
 tion analysis can yield very different results (Aus-
 tin and Grieg-Smith, 1968). Generally, the more
 analyses that are conducted on the data, the higher
 the probability of interpreting the data accurately.

 9.2.12.4 Necessary Hardware and Skills

     The hardware  needed to  perform a benthic
 community assessment  is fairly common and
 should be  readily available. Equipment includes
 field collection gear (e.g., sampling vessel, appro-
 priate sampler, sieves, sample storage containers,
 buffered fixative) and standard biological laborato-
  ry equipment (e.g., microscopes, sieves, hydrome-
  ters or pipets,  and a balance). More specialized
  equipment includes a muffle furnace  for determin-
  ing total volatile solids concentrations, a taxonom-
  ic reference collection, and a taxonomic reference
  library-   Computer  equipment  and appropriate
  software are required to make studies cost-effec-
  tive.   A  microcomputer is sufficient for most
  analyses,  but  some  complicated  multivariate
  analyses may require the use of a minicomputer or
 " mainframe computer.
    Trained benthic taxonomists are required to
ensure accurate identifications.  Some computer
programming and some level of data management
are usually required. A trained benlhic ecologist
is required to synthesize and interpret the data.
However, the amount of training depends on the
required level of interpretation.   For example,
interpretation of  several  multivariate  methods
would require a higher level of training than inter-
pretation of descriptive indexes.

9.2.1.3 Adequacy of Documentation

    Many different approaches and methods are
used to analyze benthic data, some of which have
thek origins in  classical terrestrial  community
ecology.  Because analysis of benthic community
structure is a relatively old assessment tool, liter-
 ally thousands of papers have been written about
the method.  Several books and protocols have
 also been developed to describe field and laborato-
 ry techniques {e.g., Holme and Mclntyre (1984),
 Puget Sound Protocols (Tetra Tech, 1986b), U.S.
 EPA 301(h) protocols  (Tetra  Tech,  1986a)].
 However, a comprehensive document that de-
 scribes standardized procedures for analyzing and
 interpreting  benthic community data  is lacking.
     The most commonly used  interpretive ap-
 proaches include measures of diversity and classi-
 fication. Sometimes a general consensus exists  on
 the  best techniques  to use within an  approach
 (e.g., widespread use of Shannon-Wiener diversity
  index, although there is debate as to whether this
  is  a suitable index  for environmental  impact
  analysis). Despite this consensus, studies do not
  necessarily  follow a specified format.  Program
  objectives tend to dictate the types of hypotheses
  posed and analyses used. Many relatively new and
  exciting  approaches  have been proposed   for
  assessing benthic community structure.  However,
  most are relatively untested and  are not widely
  used  [e.g., benthic  resource analysis  technique
  (Lunz and Kendall,  1982), abundance-biomass
  comparison (Warwick,  1986; Warwick  et  al.,
  1987), infaunal trophic index (Word, 1978,1980),
  nematode-.copepod ratio (Amjad and Gray, 1983;
  Lambshead, 1984; Shiells  and Anderson, 1985;
  Raffaeffi, 1987), lognonnal distribution (Gray  and
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   Sediment Classification Methods Compendium
  Mirza,  1979),  Index 5 (Bernstein  and  Smith,
  1986)]. Each of these methods has shown prom-
  ise in some situations, but more testing and vali-
  dation are needed before any can gain universal
  acceptance.
      Very  few  assessments of  the  information
  gained from analyses of data at the species level
  vs. the major taxa level have been  undertaken.
  Warwick (1988) evaluated  the results of ordina-
  tions run on various hierarchical levels of taxo-
  nomic data for five data sets.  Three of the data
  sets were of macrofauna (from Loch Linne, Clyde
  Sea, and Bay of Morlaix); one was of nematodes
  from the Clyde Sea; and the last was of copepods
  from Oslofjord that  were subjected to different
  levels of particulate organic material. He reported
  that in none of those five  cases was there any
  substantial loss  of information at the family level,
  and that in two cases the sample groupings related
  more closely to the gradient of  pollution at the
  phylum level than at the species level.  Warwick
  tentatively suggested that "anthropogenic effects
  modify community composition at a higher taxo-
  nomic level than natural environmental variables,
  which influence the  fauna  more  by species re-
  placement."  Warwick's paper appears to be the
  only published work to support the use  of higher
  taxonomic groups for analysis purposes. In cases
  where  only  major taxa level data have been
  collected (e.g., PTI and Tetra Tech, 1988), it has
  been difficult to determine differences in commu-
  nity structure between impacted areas and refer-
  ence areas, and to establish causes of community
  alterations.  Although it would be a cost-saving
  approach, use of higher taxonomic levels to assess
 benthic communities is currently not an  accepted
 approach in the United States.
 9.2.2  Applicability of Method to Human
       Health, Aquatic Life, or Wildlife
       Protection

    The assessment of benthic community struc-
 ture is directly  applicable to the protection of
 aquatic  life.   Because benthic  organisms are
, aquatic, assessments of benthic community struc-
 ture provide a direct measure of the condition of
  aquatic life.  Furthermore, because benthic organ-
  isms are consumed by other aquatic  organisms
  (e.g., fish),  assessing the condition of benthic
  communities provides information on other aquatic
  organisms.
     Assessment of benthic community structure is
  both directly applicable to the protection of some
  wildlife (e.g., wading shorebirds that feed on the
  benthic infauna) and indirectly applicable to the
  protection of  other wildlife  (e.g.,  fish-eating
  wildlife). A substantial decrease in abundance of
  benthic organisms may result in the loss of food
  and a reduction in the value of certain habitat to
  wildlife.  For example, distributions of demersal
  fishes have been shown to be affected by changes
  in the composition of benthic infaunal communi-
  ties (e.g., see Kleppel et al.,  1980), as has  the
  distribution  of the  starfish Astropecten  verilli
 (Striplin, 1987).
     Assessment of benthic community structure
 may be directly or indirectly applied to the protec-
 tion of human health. When changes in commu-
 nity structure are caused by the presence of toxic
 contaminants, the bioaccumulation of those con-
 taminants in more tolerant species may sometimes
 be postulated.  Those contaminated benthic  in-
 fauna may directly affect human health if they are
 ingested (e.g., shellfish contamination), or may
 indirectly affect human health if contaminants are
 transferred through the food web to humans (e.g.,
 consumption of. contaminated demersal fish).
 9.2.3 Ability of Method to Generate
      Numerical Criteria for Specific
      Chemicals

    Benthic  community structure as a  stand-
 alone  assessment   method   cannot  presently
 generate numerical  criteria for specific chemi-
 cals, nor is it likely that it will without extensive
 research.  However, it is an integral component
 of other methods that generate numerical criteria
 (e.g., Apparent  Effects  Threshold,  Sediment
 Quality Triad).   The  great number of factors
 influencing benthic  community structure at a
given site generally precludes isolation of chem-
ical-specific effects.
 9-10

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                                            9—Marine Benthic Community Structure Assessment
93 USEFULNESS

    Assessment of benthic community structure
has become a valued tool for determining sedi-
ment quality.  It is recognized as the only in situ
measure that provides information on changes in
ecological relationships among species that inhabit
potentially contaminated sediment. Its usefulness
will continue both as an assessment method on its
own and as a component of other sediment quality
assessment tools.
 9.3.1 Environmental Applicability

    This  method  is  applicable in a variety  of
 environments.  As a tool for assessing sediment
 quality, it has been used to assess the effects of
 known or suspected contaminants (e.g., industrial
 or municipal discharges, oil spills). The results of
 such studies reveal the geographic extent of the
 problem area and the type and severity of contam-
 ination.   "
 9.3.1.1 Suitability for Different Sediment Types

     Benthic community structure is well suited for
 assessing spatial and temporal effects of chemical
 contamination and/or organic  enrichment in  a
 variety of sediment types.  However, to the extent
 possible, benthic communities occupymg different
 types of sediment should not be compared.  Con-
 siderable research has shown that the structure of
 benthic communities in coarse sediments  differs
 from that  in fine sediments (see Rhoads and
 Young, 1970; Rhoads and Boyer, 1982). Briefly,
 species recruiting into soft, silty sediments must
 be able to  tolerate the deposition of  fine particu-
 late material.  These environments tend to be
 inhabited by subsurface deposit-feeding organisms,
 whereas sandy environments tend to be inhabited
 by both surface suspension-feeding species and
 subsurface-dwelling  species.    Therefore,  the
 experimental design  of a benthic survey must
 reflect  that  the  functional attributes of benthic
 communities in silty and sandy  environments
  fundamentally differ.
    When reference stations are used as the basis
for determining differences in community structure
between  nonimpacted  and potentially impacted
stations,  the  reference and test stations should
exhibit, to the extent possible, similar sediment
characteristics (as well as similar water depths
because benthic communities naturally vary by
depth). However, it is Hot always possible for the
reference and test stations  to have sediment that
has a similar composition; for example, dredged
material at a dump site may have different charac-
teristics  than native sediment  surrounding  the
dump site. If the experimental design is based on
sampling the same stations through time to assess
temporal change, then presumably sediment grain
size would remain constant. If the objective is to
sample along a  potential gradient  of  chemical
contamination or organic enrichment, then all
stations should have similar grain sizes and water
depths.   However, this is not always possible
because the source of contamination may alter the
natural grain size distribution of the sediments.
     Benthic community structure is also a suitable
 technique for assessing the presence of anaerobic
 sediments caused by poor flushing  or excessive
 organic loading.  The success of this approach will
 once again hinge on comparing benthic communi-
 ty structure between  stations with similar  grain
 sizes and water depths.

 9.3.1.2 Suitability for Different Chemicals or
         Classes of Chemicals

     Analysis of benthic community structure is
 frequently used to determine effects of chemicals
 present in the sediment.  However,  it is not used
 as a method to quantify the relative concentrations
 of individual chemicals or  classes  of chemicals
 present in sediment. Although individual species
 may react to'certain chemicals, these reactions are
 not quantifiable at the community level.  The
 Apparent Effects Threshold approach (Chapter 10)
 incorporates changes in abundance  of major taxa
 for specific chemicals.
     Benthic communities respond predictably to
 general categories of contamination. For example,
 metals  contamination  of sediments  results  in
  decreased species diversity (Rygg, 1985a, 1985b,
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 Sediment Classification Methods Compendium
 1986).  Organic enrichment, which leads  to  an
 increase in the food supply, generally results in
 increased  diversity and  abundance  at  slight  to
 moderate  levels  of  enrichment  (Pearson and
 Rosenberg, 1978; Rygg; 1986). However, beyond
 some level of organic enrichment, diversity and
 abundance decrease with continued organic load-
 ing (Pearson and Rosenberg, 1978).  In an area
 receiving  both  organic  enrichment and  toxic
 contaminants, it may be difficult to distinguish the
 effects of these forms of pollution from each
 other.  Additional research  is greatly needed  to
 help separate the effects of multiple sources  of
 contaminants.
 9.3.1.3 Suitability for Predicting Effects on
        Different Organisms

    Changes in benthic communities that result
 from the presence of organic enrichment or chemi-
 cal pollutants  may be useful indicators of the
 potential effects of that pollution on predators of
 the infauna (see Kleppel, 1982; Striplin, 1987).
 However, using benthic community  structure to
 predict specific effects on potential predators (such
 as benthic-feeding fish or shorebirds)  may  be
 difficult.   Information  on trophic relationships,
 competition, and predation is often not available.
 The capability to predict the effects of altered prey
 communities on predators may  improve  with
 research on these topics.  Factors such as food
 quality, distribution of  food, interactions among
 species, and distribution of prey will all be impor-
 tant components of this research.
9.3.1.4 Suitability for In-Place Pollutant
        Control

    Benthic community structure has not  been
used to set sediment quality goals or criteria for
polluted marine sediments.  Benthic communities
naturally express sufficient spatial and temporal
variability to eliminate them from consideration as
a goal  or  criterion-setting  variable.   However,
benthic communities are an integral part of other
approaches to  assess sediment quality (see Chap-
 ters 10, and 11, and 12) in which benthic commu-
 nity structure is the only in situ biological mea-
 sure.                           ..     .  •   '
 9.3.1.5 Suitability for Source Control

     Benthic community assessments can provide
 valuable information for certain aspects of source
 control.   Benthic communities can assist the
 identification  of  outfalls that discharge  toxic
 chemicals or high organic loads.  Depending on
 the  nature of the material being discharged, ben-
 thic communities may be diverse and abundant if
 the  material is  organically enriched or may be
 depauperate if the material has high levels of toxic
 contaminants.  Because benthic communities are .
 not currently useful for identifying specific chemi-
 cals or classes of chemicals present in the sedi-
 ment, they .are  of Jimited value for identifying
 specific sources of contaminants.
     Following  the control of sources, benthic
 community structure may be used to monitor long-
 term recovery of the receiving environment (Tetra
 Tech, 1988). It is not recommended as an indica-
 tor of the immediate effects of controlling sources
 because the sediment will remain contaminated
 until the sediment is actively remediated, or until
 bioturbation and  natural deposition of  uncon-
 taminated participates  dilute the  contaminated
 sediment. Furthermore, this assessment technique
 would be  useful only in areas where other uncon-
 trolled sources would not obscure sediment recov-
 ery due to the controlled source.  Where source
 control has occurred, or is planned on a regional
 level, establishment of one or more stations for the
 analysis of long-term trends in benthic community
 structure is recommended as a method of monitor-.
 ing regional sediment recovery. The concentration
 and type of the contaminants and the hydrodynam-
 ics of the study area will govern the length  of
 time over which recovery will occur (Perez, K.,
 1 May 1989, personal communication).


 93J. 6  Suitability for Disposal Applications

   Regulations concerning biological testing of
sediment that is  dredged under sections 401 and
9-12

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                                             9—Marine Benthic Community Structure Assessment
 404 of the Clean Water Act do not include assess-
 ments of benthic  community structure.  Benthic
 communities inhabit  only the upper layers of
 sediment that will be dredged.  Because sediment
 quality near the sediment surface may not reflect
 sediment quality throughout the depth of sediment
 to be dredged, benthic communities are unable to
 provide information that is suitable for assessing
 the  entire  volume  of sediment that  will be
 dredged. Chemical analyses, laboratory biqassays,
 and bioaccumulation studies  can, however, be
 used to  assess sediment  quality throughout the
 dredging depth. Section 102 of the Marine Pro-
 tection Research and Sanctuary Act does call for
 monitoring of benthic community  structure in
 areas where dredged material is  disposed.
      The International Joint Commission  (UC)
 recommends use of benthic communities to deter-
 mine whether areas of concern exist in sediments
 that require dredging (UC, 1988a, 1988b).  How-
  ever, they do not discuss whether benthic commu-
  nity structure would be used to determine the suit-
  ability  of dredged material for open-water dis-
,  posal.
    , Analysis of  benthic  community  structure  is
  appropriate for postdisposal monitoring of con-
  fined and unconfined disposal sites and for moni-
  toring recovery of areas  that were dredged. As
  part of the Puget Sound Dredged Disposal Analy-
  sis (PSDDA) postdisposal monitoring program,
  benthic community structure is used to monitor
  the potential transport of disposed material away
  from the disposal site (SAIC, 1991; Striplin et al,
  1991). The purpose of this aspect of the monitor-
  ing program  is  to determine whether benthic
  communities are altered near the disposal site and,
  if  so, whether  the changes are due  to  offsite
  migration of the  disposed  material.    Benthic
   community structure was also incorporated into
   the proposed monitoring program  for confined
   aquatic disposal sites to confirm recolonization of
   the clean sediment cap and to monitor cap integ-
   rity  at  the  Commencement  Bay Nearshore/
   Tideflats Superfund site in Tacoma, Washington
   (Tetra Tech, 1988). As described earlier, Swartz
   et al. (1980) documented recovery in Yaquina
   Bay, Oregon, following dredging.  Rhoads et al.
(1978) suggested that periodic disturbance such as
dredging  and  disposal  may enhance  benthic
productivity.


93.2 General Advantages and Limitations

    General advantages of using benthic commu-
nity  structure  to  determine sediment  quality
include  its inherent capability  to  provide an
ecological basis for evaluation of sediment quality.
It is an empirical rather than a theoretical ap-
proach.  However* -as with  most assessment tech-
niques involving field studies, the evaluation of
benthic communities is costly and time-consum-
ing.  The information gained is often not suitable
for specific management decisions because of the
lack of numerical management  criteria  and the
method's inability to identify specific chemicals
responsible for an impact.  However, the tech-
nique has been incorporated into other predictive
 techniques  (see  Chapters  10, 11, and  12) that
 provide information more easily used by resource
 managers.

 9.3.2.1 Ease of Use

     Assessments  of benthic community  structure
 require field  collections, extensive laboratory
 work,  and data  analysis and  interpretation by
 trained benthic ecologists.  It is difficult to argue
 that the method is easy to use,  especially  in
 comparison to other methods that rely  on estab-
 lished criteria.   However, the  use of benthic
 community structure as a sediment quality assess-
 ment tool is widely accepted, and trained benthic
 ecologists are available throughout the country.
 By using highly experienced individuals to con-
 duct the field, laboratory, and data analysis work,
 potential problems (such as generating  "noisy"
  data that obscure real trends, or arriving at differ-
  ent interpretations using the same data) should not
  occur.                                     •

  9.3.2.2 Relative Cost

      The relative cost of conducting an assessment
  of benthic communities  is less than the cost to
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  Sediment Classification Methods Compendium
  develop and  implement other sediment  quality
  assessment techniques  such as  the Apparent
  Effects Threshold  and  equilibrium partitioning
  approaches.   However, once sediment  quality
  values have been generated, the relative  cost of
  conducting a  benthic survey is greater than the
  cost of analyzing sediment for contaminant con-
  centrations and comparing those data to the values
  to determine sediment quality. Sediment toxicity
  bioassays are generally less costly than analysis of
  replicated  benthic samples.  Because the Triad
  approach  requires synoptic analyses of sediment
  chemistry, sediment toxicity, and benthic commu-
  nities, it is more costly to implement than  simply
  an  analysis of benthic  communities.   It also
  provides  broader information from   which  to
  determine sediment quality.
     The objectives of benthic community assess-
  ment programs strongly influence cost by dictating
  the number of stations and number of replicates
 per station. The cost per replicate is relatively
 high (i.e., $400-$1,000), but varies greatly depend-
 ing on the size of the area  sampled,  the  screen
 size, the level of the taxonomic identifications,
 and the environment sampled.

 9.3.2.3 Tendency to Be Conservative

     Benthic community structure is a moderately
 conservative measure  of sediment  quality.  Be-
 cause benthic  community structure reflects the
 collective  response  of all species, responses of
 individual species that are susceptible to degrada-
 tion in sediment quality may not be obvious at the
 community level because of the lack of response
 in other species that are more tolerant of environ-
 mental degradation. Changes to numerous species
 or dominant species must occur before changes at
 the community level are evident. If assessments
 of sediment quality were made using individual
 species instead of communities, they  could be
 either conservative by relying on sensitive species
 or not conservative by relying on tolerant species.

 9.3.2.4 Level of Acceptance

    Benthic community assessments have been
used as a sediment  quality assessment tool fqr
  several decades in North America, Europe, and
  Australia, as well as in South Africa, China, and
  Japan. The method has gained widespread accep-
  tance because of its inherent capability, to assess
  sediment quality at the community level, thereby
  documenting  ecological response  to sediment
  perturbations.
     Many methods may be used to analyze ben-
  thic community data, as discussed above. Some
  of these methods have gained far wider acceptance
  than have  other,  sometimes  newer,  approaches.
  The most  widely  accepted  types of analyses
  include measures  of abundance, numbers  of taxa,
  diversity, similarity, community classification, and
  the abundance  of sensitive and tolerant species.
  Other analytical methods include the lognonnal
  distribution (Gray and Mizra, 1979),  the use of
  major taxa instead of species-level data (Warwick,
  1988), and the Infaunal Trophic Index  (Word,
  1978, 1980).  Each of these may be appropriate
 for certain types'of perturbations, but have yet to
 gain widespread acceptance.


 9.3.2.5 Ability to Be Implemented by
        Laboratories with Typical Equipment
        and Handling Facilities  .
    Many laboratories either  have the essential
 equipment  for  conducting benthic  community
 surveys  or can readily obtain  this  equipment
 However, locating qualified taxonomists to over-
 see the sorting and to identify  the organisms may
 be difficult! Taxonomists require several years of
 training and experience before they are considered
 experts in their respective taxonomic fields. They
 also require access to a reference  museum of
 verified organisms to assist in their identifications.
 A thorough taxonomic library containing original
 descriptions of species is also an integral compo-
 nent of taxonomic laboratories.

 93.2.6 Level of Effort Required to Generate
       Results

   The level of  effort required  to  conduct  a
benthic  community survey is  dependent  on the
objectives of the program, which may affect the
number of stations, number of replicates per sta-
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                                            9—Marine Benthic Community Structure Assessment
tion, taxonomic level of the identifications, and
data analysis  procedures.   Regardless  of those
objectives, a field effort is required; the samples
must be sorted, identified, and enumerated; and
the resulting data must be analyzed. This process
typically requires several  months, but it is not
unusual for it to  require a full year for a very
large sampling effort, or for 'a program in which
the samples require large sorting or identification
times.  For example, the sorting time for samples
collected from deep water silt and clay may be
1-2 h, whereas  that for samples from shallow
sandy'sites might be 4-6 h because shallow sandy
areas typically contain more abiotic material,.  If
wood chips are present in the sample, then  the
 sorting time can easily exceed 12 h, depending on
 the volume of wood chips.

 9.3.2.7 Degree to Which Results Lend
         Themselves to Interpretation

     The interpretation of benthic community data
 requires an expert who is familiar with the natural
 history 9f the fauna and the statistical techniques
 that are routinely used to analyze the data.  Inter-
 pretation of the many data points generated by this
 approach may require many weeks before mean-
 ingful trends are recognized.  The inherent vari-
 ability of benthic communities has so far prevent-
 ed the development of specific benthic criteria for
 use in assessing pollutant-related trends in sedi-
 ment quality.

  9.3.2.8 Degree of Environmental Applicability

      The assessment of benthic community struc-
  ture is a direct measure  of the  environmental
  effects of pollutants and, as, such, is highly appli-
  cable as a method to assess sediment quality.  Its
  applicability lies in its ability to provide informa-
  tion on  the effects of pollutants on ecological
  processes within the sedimentary environment

  9.3.2.9 Degree of Accuracy and Precision

       Provided that sufficient funding is available to
  collect and process the necessary numbers of
replicate samples, analysis of benthic community
structure is accurate (defined as how well the data
represent true field conditions) and  precise (de-
fined as the consistency and reliability  of the
samples).  The resulting data are obtained directly.
from the populations under study. Other sediment
quality assessment  methods  described in  this
compendium are  not direct  measures of  field
conditions and therefore are less likely to be as
accurate and precise.
  .  Many factors in the design of a benthic com-
munity survey directly  influence the degree of
accuracy  and  precision  of the resulting  data.
These factors include station placement, number of
replicates,  appropriateness of  reference areas,
 sampler,  sieve mesh  size,  sampling  interval,
 quality of taxonomy, and the type and quality of
 the data analysis.  The best way to ensure high
 degrees of accuracy and precision is to conduct a
 pilot study in the area of interest prior to design-
 ing a major field survey. The pilot survey will
 provide information on variability within benthic
 communities,  which then  directly affects the
 required number  of replicates and  station place-
 ment The analysis of data from a pilot study
 may also help generate different hypotheses that
 may alter the sampling and analysis plans to better
 define the communities.
  9.4 STATUS

      Many methods to assess sediment quality rely
  on benthic community structure as a measure of
  potential ecological effects of pollutants.  Benthic
  community structure has been incorporated into
  programs with vastly different objectives because
  the resident biota are sensitive indicators of many
  kinds of environmental perturbations.  Aspects of
  the status  of benthic  community structure as a
  sediment quality assessment tool are discussed in
  this section.
   9.4.1 Extent of Use

      Assessment of benthic community structure
   has been a valued tool in marine, estuarine, and
                                                                                               9-15

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   Sediment Classification Methods Compendium
  freshwater environments  for several decades.
  Many  of the  early programs examined benthic
  communities from an academic viewpoint.  Since
  the 1970s, benthic community structure has been
  used as a measure of sediment quality.  Since then
  this method has been used to determine the effects
  of  municipal   effluents,  industrial  discharges,
  eulrophication, organic enrichment, oil spills, and
  mine tailings disposal  (see Section 9.1.1).  It has
  also been used to  determine the suitability of
  sediments for dredged  material disposal, to  monir
  tor dredged material disposal sites, and to monitor
  recovery of impacted areas following the cessation
  of contaminant loading.


  9.4.2 Extent to Which Approach Has Beea
       Field-Validated

     Because benthic community structure is an in
  situ sediment quality assessment tool, it does not
  require additional field validation.

  9.43 Reasons for Limited Use

     Although conducting studies of benthic  com-
  munity  structure is a common  practice, the cost
  and  amount of time required to generate usable
  results may prevent the method from being imple-
 mented by all who could benefit from its use. In
 fact, the method has  been deleted  from  some
 programs due solely  to cost (Bilyard, 1987).  In
 some situations, costs and time have been reduced
 by taking the  identifications only to the major
 taxonomic  level.   This reduction  of taxonomic
 detail frequently reduces the usefulness of the
 information (Warwick,  1988), which exacerbates
 a perception by some resource managers that the
 data  are too variable to  be useful.   Detecting
 trends within benthic  data is not a simple process.
 However, the proper design and implementation of
 a field survey will radically increase the probabi-
 lity of producing valuable data and results.

 9.4.4 Outlook for Future Use  and Amount of
      Development Yet Needed

  •  The  outlook for  the  future use of benthic
community structure as a sediment quality assess-
   ment tool is particularly bright because of the
   continuing  development of  new  data  analysis
   methods by researchers in North America and
   Europe.  The objective of these methods is gener-
   ally to reduce cost or variability within the data by
   relating" aspects of the distributions of organisms
   or organism biomass to specific kinds of environ-
   mental perturbations.   Gray  and Mirza  (1979)
   determined  that  the  lognormal  distribution  of
   individuals was altered in a predictable manner in
   the presence of slight  organic pollution. A more
   recent method for detecting pollution effects on
   marine benthic communities is the species abun-
   dance/biomass comparison (ABC) method devel-
  oped by Warwick (1986).  This method proposes
  that  the  relationship  between the number of
  individuals among species and the distribution of
  biomass among species changes in a predictable
  manner  in the presence of  organic  pollution.
  Beukema (1988) evaluated the ABC method in an
  intertidal  habitat in  the Dutch Wadden Sea and
  determined that the method "cannot be applied to
  tidal flat communities  without reference to long-
  term and  spatial series of control samples."  Yet
  another benthic community assessment method
  that remains Under development  is the Infaunal
  Trophic Index proposed  by Word (1978,  1980).
  That method is based on changes in the feeding
  ecology of benthic infauna in relation to organic
  enrichment.  The  Benthic Resource Assessment
  Technique,  developed by Lunz  and  Kendall
  (1982), quantifies the effects of changes in benthic
  communities on fish resources.  Although  the
  BRAT technique is  not  a direct  assessment of
 benthic community structure, it provides important
 information on .the relationships among benthic
 communities and  higher level predators, and
 describes how those relationships may change in
' the presence of pollutants.
    A radically different approach  to interpreting
 long-term changes in benthic community structure
 involves use of a sediment profile  camera.  Rhp-
 ads and Germane (1986) developed  the RE-
 MOTS® (remote ecological mapping of the sea-
 floor) system. They use a vessel-deployed sedi-
 ment-profile camera to  photograph vertical sec-
 tions of the  sediment.   Although  REMOTS®
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                                           9—Marine Benthic Community Structure Assessment
cannot determine the species composition of the
benthic community, it can document relationships
between  organisms and sediment.   Rhoads  and
Germane  (1986) characterized the successional
stages of benthic communities and suggested that
mapping these stages will permit the detection of
changes  in benthic communities.   When  this
information is collected as part of a preliminary
survey, it can be used to assist in the design of a
cost-efficient  benthic  community  survey  for
obtaining geochemical and biological information.
    Additional research is needed on some funda-
mental aspects of benthic community assessment.
These include the development of guidelines for
the identification of reference sites or reference
values and additional studies into the usefulness of
 identifying infauna to various taxonomic levels.
 U.S. EPA is presently examining some aspects of
 these questions through  the  Clean  Water Act
 section 301(h) program^ including examination of
 the degree of variability in benthic communities in
 contaminated and reference areas, development of
 a quantitative definition of "balanced indigenous
 populations," and assessment of the effects  of
 overlapping  contaminant sources   on  benthic
 infaunal communities.
     The sediment profile camera has been used for
 a variety of other purposes including assessing the
 relationships between sediment quality and eutro-
 phication (Day et al., 1987; Revelas et al,  1987;
 Rhoads, D.C., 1 May 1989, personal communica-
 tion), monitoring the perimeter of dredged materi-
 al  disposal sites (Rhoads, D.C,  1 May  1989,
 personal communication; Diaz, R J.,  1 May 1989j,
 personal communication),  and  evaluating the
  overwintering habitat of blue crabs in Chesapeake
  Bay (Schaffner and Diaz, 1988).  With further
  research, the sediment profile camera may be used
  for  other  applications  concerning  aspects  of
  benthic community structure and sediment quality.
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Striplin,  P.L.   1987.  Resource utilization  by
   Astropecten verrilli along gradients of organic
   enrichment. M.  Sc. Thesis.  California State
   University at Long Beach,-Ijong Beach, CA.
   108 pp. -f appendices.  .
Swartz, R.C. 1978.  Techniques for sampling and
   analyzing the marine  macrobenthos.   EPA
   600/3-78-030.  U.S. Environmental Protection
   Agency, Corvallis, OR. 27 pp.
Swartz, R.C, WA. DeBen, F.A. Cole,  and L.C.
   Bentsen.  1980. Recovery of the macrobenthos
   at a dredge site in Yaquina Bay, Oregon, pp.
   391-408.  In:  Contaminants  and  Sediments,
   Vol. 2.  R. Baker (ed.). Ann Arbor Science,
   Ann Arbor, ML
Swartz, R.C.  15 March 1989.  Personal com-
   munication (phone by Ms.  Betsy  Day, Tetra
   Tech,  Inc., Bellevue, WA regarding  status of
   replication  study  using  samples  collected
   during the Everett  Harbor Action  Program
   survey). U.S. Environmental Protection Agen-
   cy,  Newport, OR.
Tagatz, M.E., G.R. Plaia, C.H. Deans, and E.M.
   Lores.  1983.  Toxicity of creosote-contami-
   nated  sediment to field-and laboratory-colo-
   nized estuarine benthic communities.  Environ.
   Tox. Chem. 2:441-450.
    9-20

-------
                                           g—Marine Benthic Community Structure Assessment
Tarazona, J., H. Salzwedel, and W. Arntz. 1988.
   Oscillations of macrobenthos in shallow waters  -„
   of the Peruvian  central coast induced by El
   Nino  1982-83. J. Mar. Res. 46:593-611.
Tetra Tech.  1986a.  -Quality assurance/quality
   control (QA/QC) for 301(h) monitoring  pro-
   grams:   guidance  on field  and  laboratory
   methods. Prepared for the U.S. Environmental
   Protection  Agency,  Office  of Marine  and
   Estuarine Protection, Marine Operations Divi-
   sion,  Washington,  DC.   Tetra Tech,  Inc.,
   Bellevue, WA.
 Tetra Tech. 1986b.  Recommended protocols for
   measuring selected environmental variables in
   Puget Sound. Prepared  for the Puget Sound
   Estuary Program, U.S. Environmental Protec-
   tion Agency, Region X, Seattle, WA.  Tetra
   Tech, Inc., Bellevue, WA.
 Tetra Tech.  1987.  Technical support document
    for ODES statistical power analysis.  Prepared
    for  Marine  Operations  Division,  Office of
    Marine  and  Estuarine  Division,   Office of
    Marine and Estuarine Prqtection, U.S. Environ-
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    Bellevue, WA.  34 pp. + appendices.
 Tetra Tech.  1988.  Commencement Bay near-
    shore/tideflats feasibility study.  Prepared for
    Washington Department of Ecology and U.S.
    Environmental Protection Agency.  Tetra Tech,
    Inc., Bellevue, WA.
 Tilley, S., D. Jamison, J. Thornton, B. Parker, and
    J. Malek. 1988.  Management plans technical
    appendix. Prepared for  Puget Sound Dredged
    Disposal Analysis. U.S. Army Corps of Engi-
    neers, Seattle, WA.
  Vezina, A.F.  1988.  Sampling variance and the
     design  of quantitative surveys of the marine
     benthos. Mar. Biol. 97:151-155.
  Vidakovic,J.  1983. The influence of raw domes-
     tic  sewage  on density and distribution of
     meiofauna.  Mar. Poll. Bull. ,14:84-88.
Warwick, R.M. 1986. A new method for detect-
   ing pollution effects on marine macrobenthic
   communities.  Mar. Biol. 92:557-562.
Warwick, R.M-  1988.  The level of taxonomic
   discrimination required  to, detect  pollution
   effects on marine benthic communities.  Mar.
   Poll. Bull. 19:259-268.
Warwick, R.M.,  T.H. Pearson, and Ruswahyuni.
   1987. Detection of pollution effects on marine
 "  macrobenthos:  further evaluation of the spe-
   cies abundance/biomass  method. Mar. Biol.
   95:-193-200.    .
Washington, H.G.  1984.  Diversity, biotic, and
   similarity  indices.   A  review  with special
   relevance to aquatic ecosystems.  Water Res.
    18:653-694.
 Winer, B J. 1971. Statistical principles in experi-
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    New York, NY.
 Word,J.Q. 1976. Biological comparison of grab
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    Water Research Project Annual Report. South-
    em California Coastal Water Research Project,
    El Segundo, CA.               ."...-.
 Word, J.Q.   1978.  The infauhal trophic index.
    pp. 19-39. In: Coastal Water Research Project
    Annual Report for 1978.  Southern California
    Coastal Water Research Project, El Segundo,
    CA.
 Word, J.Q.  1980.  Classification of benthic inver-
    tebrates  into infaunal trophic index feeding
    groups,   pp. 103-121.    In:  Coastal Water
    Research Project. Biennial Report of the years
    1979-1980.  W. Bascom (ed.).  Southern Cali-
    fornia Coastal Water Research Project, Long
    Beach, CA.
  Word, J.Q., B.L. Myers, and AJ.  Mearns.  1977.
    Animals that are indicators of marine pollution.
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    Annual  Report.  Southern California Coastal
     Water Research Project, El Segundo, CA.
                                                                                            9-21

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          CHAPTER 10
Sediment  Quality  Triad Approach

Peter M. Chapman
E.V.S. Consultants Ltd.
195 Pemberton Avenue, North Vancouver, BC, Canada V7P 2R4
Phone (604) 986-4331, FAX (604) 662-8548
    The Sediment Quality Triad (Triad) approach
is an effects-based approach to describe sediment
quality.  It typically incorporates measures of
sediment chemistry, sediment toxicity, andbenthic
infauna communities, although other variables can
be used. This combination method  is both de-
scriptive and numeric. It is most commonly used
to describe sediment  qualitatively, but has also
been used to generate chemical-specific sediment
quality  criteria  (Chapman, 1986, 1989;  Long,
1989). One application of the Triad approach, the
Apparent Effects Threshold (AET), is described in
detail in the following chapter (Chapter 11).
 10.1  SPECIFIC APPLICATIONS

•10.1.1 Current Use

    The Triad approach can be used to determine
 the extent of pollution-induced degradation of
 sediments in  a non-numerical, multiple-chemical
 mode (e.g. Chapman e'tal., 1986, 1987a, 1991a;
 Chapman and Power, 1990; Chapman, 1990). It
 can also be used to determine numerical sediment
 quality criteria directly (e.g. Chapman, 1986,
 1989) and, through  manipulations, to determine
 AET values (see Chapter 11).
    The  AET is  .only  one  possible. method of
 evaluating triad  data and  is directed solely at
 determining  numeric sediment quality values
 (Chapman et al, 1991b, 1991c).  The triad ap-
 proach has been used in marine coastal waters on
 the west coast of North America (e.g., Puget
 Sound, San Francisco Bay, and Vancouver Harbor,
 Canada), in  the Gulf of Mexico,  in freshwater
 environments including the Great Lakes, and in
 the North Sea (Long and Chapman, 1985; Chap-
 man, in  press; .Chapman et al., 1986, 1987a, in
 press; Chapman and Power, 1990; Cross et al.,
 1991, in review).   Current uses  of the  Triad
approach are summarized in Table  10-1  and
discussed  in  Section  10.3,1,  Environmental
Applicability.
10.1.2  Potential Use

    The Sediment Quality Triad approach can also
be used to meet the following objectives:

    •  To identify problem areas of sediment
       contamination  where pollution-induced
       degradation is occurring;

    •  To prioritize and rank degraded areas and
       their environmental significance; and

    •  To predict where  such degradation will
       occur based on levels of contamination
       and toxicity.
    The Triad approach can be used in any
 her .of situations and is not restricted to aquatic
 sediments. For example, it can be used in water
 column work with phytoplankton and in terrestrial
 hazardous waste dump studies with other organ-
 isms of  concern.  Other uses are described in
 Section 10.3.1.  A complete description of the
 Triad in tie context of integrated assessments is
 provided in Chapman et al., 1991b.
 10.2 DESCRIPTION
                                           )
 10.2.1 Description of Method

     The Triad approach consists of three com-
 ponents (Figure 10-1):

     •  Sediment chemistry-—to measure chemical
        contamination;

-------
 Sediment Classification Methods Compendium
                    Table 10-1. Current Uses of the Sediment Quality Triad Approach.
Use
Prioritize areas for remedial
actions
Determine size of areas
Verify quality of reference areas
Determine contaminant concen-
trations always associated with
effects
Describe ecological relationships
between sediment properties and
biota at risk
Comment
Most common usage to date
Assuming increasing importance
Assuming increasing importance •
Common usage; can result in numerical
sediment quality criteria and setting of
standards
Along with setting standards and criteria,
provides for proactive approach to envi-
ronmental protection
General Locations
Where Implemented*
PS, GM, SF, VH, FW
PS
PS
PS, NS
PS, VH, FW, NS
     "PS - Puget Sound, various locations (Long and Chapman, 1985).
     GM = Gulf of Mexico, oil platform (Chapman era/., 1991 a; Chapman and Power, 1990).
     SF=  San Francisco Bay, various locations (Chapman era/., 1986,1987a).
     VH -  Vancouver Harbor. Canada, various locations (Chapman etal., 1989; Cross eta/., 1991; Cross era/., in
          review).
     FW = Various freshwater environments (Malueg et a!., 1984; Chapman unpublished data; Rogers, North Texas
          State, unpublished data; Wiederholm era/., 1987).                                          •
     NS =  North Sea (Chapman, in press; Chapman ef a/., in press). .
     »  Sediment bioassays—to measure toxicity;

     •  In situ biological variables — to measure
        in situ alteration (e.g., a change in benthic
        community structure).

    The three components provide complementary
data. No single component of the Triad approach
can  be used to predict the measurements of the
other components.  For instance, sediment chemis-
try provides information on contamination but not.
on biological effects. Sediment bioassays provide
direct evidence of sediment toxicity.  However,
the laboratory  conditions under which bioassays
are  conducted may not  accurately reflect field
conditions of exposure to toxic chemicals. In situ
alteration of resident biota measured by infauna
community analyses provides direct evidence of
contaminant-related effects in the environment, but
only if confounding effects not related to pollution
(e.g., competition,  predation, recruitment cycles,
sediment type, salinity, temperature, recent dredg-
ing) can be excluded.  In. particular, because the
toxicity of a chemical substance in sediments may
vary with its concentration and with the conditions
within a specific sediment, the importance of any
particular concentration of a chemical or suite of
chemicals in sediments  cannot be determined
solely from chemical measurements.  Sediment
conditions include grain size, organic content, pH,
Eh,  chemical  form,   and  presence  of  other
chemicals.
    The three components of the Triad approach
integrate chemical and biological response'data.
They also provide the strongest  evidence  for
identifying pollution-induced degradation.  For
instance,  if there  are  high levels  of sediment
contamination, toxicity, and biological alteration,
the burden  of  evidence indicates  degradation.
Conversely, low levels of sediment contamination,
toxicity,  and biological alteration indicate non-
degraded  conditions.   Conclusions  that  can be
drawn from intermediate responses  are listed  in
Table 10-2.
10-2

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                                                                       10—Triad Approach
                                          BULK
                                  SEDIMENT CHEMISTRY
                          SEDIMENT
                         BIOASSAYS
Figure 10-1. Conceptual Model of the Sediment Quality Triad.                                .  .
The Triad co/nb/nes dafa from chemistry, toxicity bioassays, and in situ sfud;es. Che/n/sfty and
fc/bassay estimates are based on laboratory measurements with field-collected sediments. In situ
studies generally include, but are not limited to, measures ofbenthic community structure. Areas
where the three facets of the Triad show the greatest overlap (in  terms of either positive or
negative results) provide the strongest data for determining sediment quality criteria.
10.2.1.1  Objectives and Assumptions

    The objectives of the Triad approach are to
independently measure sediment contamination,
sediment toxicity,  and biological alteration, and
then use the burden of evidence to assess sediment
quality based on all three sets of measurements.
    The following  assumptions apply:

    •  The approach allows for (1) the interac-
       tions between  contaminants in complex
       sediment mixtures (e.g., additivity, antag-
       onism, synergism);  (2) the  actions of
       unidentified toxic chemicals; and (3) the
       effect of environmental factors that influ-
       ence biological responses (including toxi-
       cant concentrations).

   •   Selected chemical contaminant concentra-
       tions are appropriate indicators of overall
       chemical contamination.

   •   Bioassay test results and values of select-
       ed benthic community structure variables
       are appropriate indicators  of biological
       effects.

   These components are presently often treated
in an additive manner,  with each having equal
                                                                                         10-3

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  Sediment Classification Methods Compendium
         Table 10-2. Possible Conclusions Provided by Using the Sediment Quality Triad Approach.
Possible
Outcome
1.
2.
3.
4.
5.
6.
7.
8.
Contamination
+
•
+
•
-
+
-
+
Toxicity
+
.-
-
+
-
+
+
-
Alteration
+
-
-
-
+
-
+
+
Possible Conclusions
Strong evidence for pollution- induced degradation
Strong evidence for absence of pollution-induced
degradation
Contaminants are not bioavailable
Unmeasured chemicals or conditions exist that
have the potential to cause degradation
Alteration is probably not due to toxic chemical
contamination
Toxic chemicals are stressing the system
Unmeasured toxic chemicals are causing degrada-
tion
Chemicals are not bioavailable or alteration is not
due to toxic chemicals
    N- * Measured difference between test and control or reference conditions.
    - «  No measurable difference between test .and control or reference conditions.
 weight because there is insufficient information
 available to assign weightings.

 10.2.1.2 Level of Effort

     Ideally, the Triad approach would be based on
 the use of synoptic data.  Sediments for analysis
 of toxicity should come from the same composited
 homogenate,  as originally detailed by Chapman
 (1988), ideally from field rather than solely labor-
 atory  test replicates.   Benthic infauna samples
 should be collected at the same  sampling loca-
 tions.   Chemistry and bioassay  sediments  are
 collected (usually by remote grab), transferred to
 a solvent-rinsed glass or stainless  steel bowl, and
 thoroughly  homogenized by stirring with a glass
 or stainless steel spatula until textural and color
 homogeneity  are achieved.   The homogenized
 sediments are then placed in appropriate sampling
 containers.  In  general, chemistry and  bioassay
•samples should include field rather than laboratory
 replication.  Benthic infaunal samples are collected
 at the same location.   In the absence of initial
 sampling  to  determine  the  optimum  level  of
 replication at a site, five field replicate benthic
 samples are recommended per station (see Chapter
 8,  Methods).   Coincident rather than  synoptic
 sampling is possible (e.g., Long and Chapman,
 1985); however, spatial heterogeneity in sediment
 contamination and toxicity make such data diffi-
 cult to interpret (Swartz et al, 1982).
    Adequate quality QA/QC measures  must be
 followed in all  aspects of the study, from field
 sampling  through laboratory  analyses and data
 entry.  Detailed QA/QC procedures are available
 through international (e.g., Keith et al., 1983) and
 regional publications (e.g., Tetra Tech, 1986a).
    The first  component of the Triad  involves
 identification and quantification of inorganic and
 organic contaminants present in the sediments.
 Chemical   analytes measured are generally re-
 stricted by equipment, technology, and the avail-
 ability of funds and facilities. Local concerns and
 existing data also affect target analytes measured.
 Cost, if a  factor,  must be balanced  against the
need for an analytical  database sufficiently large
20-4

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                                                                            10—Triad Approach
                  Table 10-3. Example Analytes and Detection Limits for Use in the
                     Chemistry Component of Sediment Quality Triad Approach.

Analyte
Conventionals (mg/kg, dry)
Grain size
TOC"
Sulfides
Acid volatile sulfide (AVS)b
Inorganics (mg/kg, dry)
Arsenic
Iron
Chromium ,
Copper
Cadmium
Lead
Mercury
Nickel ,. -
Silver
Selenium
Zinc


Organics (fig/kg, dry)
LPAHV
Benzo(a)pyrene
Benzo(e)pyrene
Benzo (a) anthracene
Chrysene
Dibenzoanthracene
Fluoranthene
Pyrene
Detection
Limit

n/a
n/a
0.5
n/a
0.05
2.5
1.0
0.5
0.05
0.05
0.01
1.0
0.05
0.05
0.5



5
10
10
10
10
16
5
5

Analyte

Biphenyl
Perylene
Coprostemol
Ammonia
op'-DDD
op'-DDE
op'-DDT
pp'-DDD
pp'-DDE
pp':DDT
Dieldrin
Heptachlor
Hexachlorobenzene
Lindane
Mirex
PCBs"
PCP*
TCP'









Detection
Limit '

5
5
10
0.5
0.15
0.25
0.15
0.15
0.10
0.10
0.10
0.10
0.10
o.is
0.10
2.5
1.0
1:0









The detection

   VTOC =
   " AVS =
   CLPAH  =

   d PCBs =
   •PCP =
   'TCP =
           limits are the instrumental estimates.  Actual detection limits may be higher because of matrix effects.

             total organic carbon.
             AVS  methodology  is described by the U.S. EPA (1991); modifications are expected.   Contact
             Christopher Zarba at (202) 475-7326 to obtain latest protocols.
             low-molecular-weight polycylic aromatic hydrocarbons (includes acenaphthene, anthracene, naphthalene
             and methylated naphthalenes, fluorene, phenanthrene, and methylated phenanthrenes).
             polychlorinated biphenyls.
             pentachlorophenol.                        ;
             tetrachlorophenol.                          "                 •                         :
to allow determination of the presence (or ab-
sence) of known toxicants of concern.
    An example of some of the types and classes
of compounds required to provide a  reasonable
characterization of chemical  contamination is
shown in Table 10-3.
                                                        Total organic carbon and grain size are mea-
                                                    sured to provide a basis for normalizing the data
                                                    to different types of sediments.   Acid  volatile
                                                    sulfides (AVS) provide information for determin-
                                                    ing metals availability from sediments.  Copro-
                                                    stanol,  an indicator of human waste,  can  be
                                                                                               10-5

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  Sediment Classification Methods Compendium
                            Table 10-4. Possible Static Sediment Bioassays.
Bloassay
Marine Waters
Rhepoxynius abroniutf
(adult amphipod}
Bivalve Larvae '
development
Neanthes sp.
(juvenile polychaetes)
Fresh Waters
Hyalella azteca
(adult amphipod)
Daphnia magna
(water flea)
Chironomus tentans
(juvenile insect)
Estuarine Waters
Eohaustorius estuarius
(adult amphipod)
Duration

10 days
48 hours
20 days

10 days
10 days
25 days

10 days
Endpolnt

Survival, avoidance
Survival, development
Survival, growth

Survival, avoidance
Survival, reproduction
Survival, growth
- .
Survival, avoidance
Amount of
Sediment
Required (L)

1.5
0.5
2.0 '

1.5
0.5
1.5

1.5
   * Note: Other options Include but are not necessarily restricted to Ampelisca abdita, Corophium volutator. Gran-
         didieneltajaponlca, Foxfphalus xiximeus.                             •
measured  to differentiate  sewage inputs from
industrial inputs.
    The second Triad component involves identifi-
cation and quantification of toxicity  based on
laboratory tests using field-collected sediments.
Ideally, one would test  the toxicity of the sedi-
ments to all ecologically  and commercially impor-
tant  fauna  living in  or  associated  with  the
sediments. For logistical reasons, a small number
of bioassays is conducted to cover as wide a range
as possible of organism type, life cycle, exposure
route, and feeding type.  The number of tests
undertaken is affected by the same constraints as
those mentioned for sediment chemistry analyses.
    Possible static sediment bioassays that provide
a reasonable characterization of  the degree of
toxicity are shown in  Table 10-4.  Obvious omis-
sions from this list include full life-cycle chronic
tests, and genotoxic or cytotoxic response tests.
Such tests merit consideration for inclusion when
proven accepted methods become available (e.g.,
Long and Buchman, 1989).
    The  final  Triad  component  involves the
evaluation of in situ biological alteration.  Gener-
ally, this  component  is provided by benthic in-
fauna community data because benthic organisms
are relatively sessile and location-specific.  Histo-
pathology of bottom fish has also been used for
this Triad component (Chapman, 1986), but for
areawide rather than site-specific studies, because
these fish are relatively mobile. Several variables
in  combination are effective in characterizing
benthic  community  structure   for  the  Triad
approach: numbers of taxa, numerical dominance,
10-6

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                                                                            10—Triad Approach
 total abundance, and percentage composition of
 major taxonomic groups. In the marine environ-
 ment, this  last category includes  any or all ot
 polychaetes,  amphipods, molluscs, and echmo-
 derms   In the freshwater environment, oligo-
 chaetes,  chironomids,  and other  major insect
 groups would fit into the last category.
     Sediment chemistry,  ^toxicity, and benthic
 infauna data are combined in the Triad approach
 to assess the degree of degradation of each station
 and of each site (see Figure 10-1).  All data are
 compared  on  a  quantitative  basis and can  be
 normalized to reference site values by converting
 them to ratio-to-reference (RTR) values as de-
  scribed  by  Chapman et al. (1986,  1987a) and
  Chapman  (1990).   The  reference  site chosen
  (either a priori or a posteriori) is generally  the
  least contaminated site  of- those sampled, and
  ideally its sediment and other characteristics (e.g.,
  water depth) would be  similar to those of the
  other sites.  To determine RTR values, the values
  of specific variables (e.g., normalized concentra-
  tion of a  particular metal, percent mortality in a
  particular bioassay, number of taxa) are divided by
  the corresponding reference values.  This process
  normalizes the data so that they can be compared
   even when, for instance, there are large differences
   in the units of measurement.  The reference site
   may be a single station (whose RTR value is 1.0
   by  definition) or an area containing several sta-
   tions for which data are averaged.
       The  RTR  criterion  is based but does  not
   depend on the assumption that the  reference site
   concentrations   are  indicative  of  reference  or
   background  conditions.   The degree to which
   chemical concentrations are elevated above the
   mean reference  concentrations at a selected site is
   used as the criterion for selecting chemicals most
    likely to be anthropogenically enriched and of
    concern.  An  index of contamination  can be
    calculated for each station by separately determin-
,    ing RTR values for groups of similar chemicals
    (e.g., metals, PAH, chlorinated organics) and then,
    assuming additivity, combining these values as a
    single mean chemistry RTR value. Similarly,, an
    index of toxicity can be calculated by combining
    bioassay  RTR values as a  single  mean value.
    Finally, an index of biological alteration can be
calculated in the same manner as is toxicity, using
benthic community structure data. The indexes of
contamination can be used to rank stations. These
summary  ranks can also  be compared with the
ranks generated using the sediment bioassay and
infaunal data.                              .
    The composite RTR  values for each Triad
component can also provide useful visual indexes.
These values can be plotted  on scales  with  a
common  origin and placed at  120 degrees from
each other such that each of the three values
becomes  the vertex  of a triangle.  The  relative
 degree of degradation  is derived by  calculating
 and comparing the areas of the triangles for each
 station or site. Examples of such triaxial plots are   -
 shown in Figure 10-2, for the eight possible situa-
 tions shown in Table 10-2.   These  plots also
 provide  a visual guide to the characteristics  of
 background or reference stations. Because refer-
 ence data usually involve a site containing more
 than one  reference station,  RTR comparisons
 should also be made against individual reference
 stations.  Alden (1992) provides a  method  for
 determining confidence limits^  for such triaxial
 plots  Non-RTR methods of Triad data analysis
  are outlined in Section 10.2.1.2.3, Types of Data
  Required.

  10:2.1.2.1 Type of Sampling Required

      As  described, synoptic sampling is preferred
  for all three Triad components. Any reasonable
  sampling procedure can  be  used if it provides
  suitable sediment samples for quantifying sedi-
  ment  contamination,  toxicity,  and  biological
  alteration.  To date, studies have  used remote
  samplers such as a 6.1-m2 Van Veen grab operated
  from a vessel,

   10.2.1.2.2 Methods

       Typical variables  included in  the chemical
   analyses and  sediment  bioassays  are  listed in
   Tables 10-3 and 10-4, respectively.  Details for
   benthic infauna analyses are provided in Chapter
   8  Although unit costs vary, costs are generally
   oil the order of $1,500 for three separate replicated
   (n=5) sediment bioassays, $1,500 for unreplicated
                                                                                                 30-7

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  Sediment Classification Methods Compendium
                      TOXICITY
                      Toxicmr
                                            1+*
                                         ALTERATION
                                            1+x
                                         ALTERATION
                                           1+a
                                        ALTERATION
                                                           i+x CONTAMINATION
                                                          1+x CONTAMINATION
                                                          1 +x CONTAMINATION
Rgure 10-2. Sediment Quality Trial Triaxial Plots for the Eght Possible Situations Shown in Table 10-2.
7770 Sediment Quality Triad determined, in the example situation,
to p.Sffo^T6eC//n Tab,le 1°'2' Toxicity> co^mi^^^d
to Ratlo-to-References values as described by Chapman et al.
                                          *-
10-8

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                                                                           10—Triad Approach
chemical analyses, and $2,500 for replicated (n=5)
benthos.

10.2.1.2.3 Types of Data Required

    Standard measurements of chemistry, toxicity,
and biological  alteration are required.  These
measurements can then be combined, as described
above. Detailed data calculations and analyses are
as follows:

Data Calculations - Benthic Data

     m   Calculate/determine endpoints
         • taxa richness
         • total abundance
         • numerical dominance
         • species diversity
         • mean abundances of all species of ma-
          jor taxa  (e.g. polychaetes, amphipods,
          chironomids, oligochaetes)

     •  Cluster Analysis
         • e.g., using mean numbers of individuals
          per taxa present at each station tested.

  Data Calculations - Chemistry

      •  Bulk concentration  normalized to  dry
         weight       ,

     - •  Organic carbon normalized concentration
         of organic compounds

      »  Normalize to percent fines, sand, silt, and
         clay

      •  AVS normalized concentration of metals
         (DiToro et al, 1990; DeWitt et al., 1990)

       •   Summarize means, standard  deviations,
          ranges for each parameter at each site.

  Data Calculations - Bioassay

       •  Between station differences in mean re-
          sponse, ANOVA, multiple  comparison
          tests.
    •  Paired comparison with control response.

    •  Comparison of mean response with lower
       prediction  limit (LPL) (DeWitt et al.,
       1988); this comparison addresses possible
       grain-size effects on amphipods,

Non-RTR Methods of Triad Data Analysis

 •   The traditional reduction technique of calculat-
ing RTRs  (by translating resultant measures to
proportions of comparable values obtained for the
reference site) has the following problems (Cross
et al., 1991; Cross et al., in review):

    •  Substantial loss of information during the
        conversion of multivariate data into single
        proportional indexes;

     •  Loss of any spatial relational information;

     •  Inability to statically assess significance
         of spatial impacts; and

     •   Requirement of an appropriate reference
         station.

     In addition, Triad results could be strongly
 influenced by the presence of unmeasured toxic
 contaminants  that may or may  not covary with
 measured chemicals (Chapman, 1990).  The RTR
 approach is useful in specific situations and with
 defined limitations; however, the following op-
 tions are useful for reducing or removing the
 problems identified.

  Ranking—In addition to RTRs, rankings can also
  be assigned to biological, chemical, and lexicolog-
  ical data for statistical comparisons of the data.
  Using  the  chemistry data  as  an example,  the
  sample with the lowest level  of  a chemical is
  scored as 1 and the highest is scored with a
  number that is equal to the number of time peri-
  ods  or samples that are to be ranked.  Tied data
  should be scored by calculating an average of the
  tied ranks.  Each site will have a rank for each
  biological, chemical, and toxicological parameter.
  Ah overall mean rank for each site can be calcu-
                                                                                               10-9

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  Sediment Classification Methods Compendium
  lated using each of the parameters.  This effective-
  ly determines how each site compares to each of
  the other sites.
     Average ranks for  biological,  chemical, and
  lexicological data can also be calculated and can
  be compared using Kendall's  coefficient of con-
  cordance (Zar,  1984).   High  concordance will
  indicate that biological, chemical, and toxicologi-
  cal parameters are changing in the same direction
  (improving or degrading). Low concordance will
  indicate that biological, chemical, lexicological
  data are changing independently of each other.

  Multivariate  Analysis—Multivariate  analysis
  comprises data matrix preparation,  analysis inde-
  pendent of the Triad components, analysis concur-
  rent with the Triad components, and Mantel's test
  Each  of these is  briefly  described here.

 Data Matrix Preparation

    For each Triad component, data are standard-
  ized to common units where possible and incorpo-
 rated  into separate  matrices  for  analysis and
 interpretation.

 Benthos:    Data are abundance of each taxon
             per grab sample; transformed to log
             (x+1).

 Chemistry:  Values less than the detection limit
             are omitted to maintain the integrity
             of the matrix. Remaining data are
             log-transformed.

 Bioassay:    Because of ihe number of indepen-
             denl bioassays and differing  end-
             poinls (e.g., mortality,  avoidance,
             reburial, elc.), these data cannot be
             standardized to  common  internal
             units.    Various  transformations
             (arsine square rool, log, elc.) may
             be used as required.

Independent Analysis of the Triad Components

 •  Each matrix is analyzed separately to deter-
mine environmental impact as provided by each
  independent approach. Community classification
  analysis may be performed for each data matrix
  using  cluster  analysis.   "Boot-slrapping11  tech-
  niques  developed  by  Nemec  and  Brinkhurst
  (1988a, 1988b) can be used to tesl whelher clus-
  ters of samples  differ  significanlly from  each
  olher.

  Concurrent Analysis of the Triad Components

     The ecological ordination technique, principal
  components analysis.(PCA), can be used to exam-
  ine  relationships between  benthos  community
  slruclure,  toxicology,  and ihe  physical-chemical
  attributes of the bottom  sedimenls, (Cross et al.,
  1991, in review).   PCA is used to reduce the
 multidimensionality of the benthos dala, creating
 two variables (principal  component or PC) from
 the original matrix of  many  variables (taxon
 abundances).  These PCs can then be correlated
 with PCs derived from physical-chemical data or
 bioassay  results, or with individual physical  or
 chemical parameters.  High  correlations among
 PCs from the  three Triad components indicate
 agreement or concordance of impact assessments.
     Correlations  of PCs from  benthic data  (or
 bioassay dala) wilh individual chemical parameters
 can be used to assess or develop sediment quality
 criteria.  The impacts  associated with existing
 criteria can be expressed as a PC score for benthic
.dala, calculated from a regression of Ihese scores
 on  chemical  concenlrations.  Sediment  quality
 criteria could also be developed by predicting the
 chemical concentration associated with a signifi-
 cant impact on  the benthic community, provided
 that "significant impact"  could be unequivocally
 associated wilh a particular PC score or range of
 scores.

Mantel's Test

    Another method that can be used to determine
whether different components  of the Triad are
related is Mantel's test (Mantel,  1967; Legendre
and Fortin, 1989).  Mantel's test uses a random-
ization procedure that calculates the probability
that two distance matrices are more similar than
would be expected by chance alone. Multivariate
30-20

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                                                                           20—Triad Approach
(or univariate) distance between each of the sites
(observations) can be calculated using data from
each component of the Triad.  For example, to
develop a distance matrix based on toxicity test
results, each of the toxicology variables would be
used to develop the distance.  Similar matrices
would be calculated  for benthos and  chemistry
data.                                       .
    The randomization procedure ensures that the
relationships between two distance matrices are
real and not spurious. The distance between two
stations (A and B) is always partially related to
the distance between these two and other stations
(e.g., A and C, B and C).  Mantel's test avoids the
possibility of spurious correlations by calculating
 correlations between the two matrices based on
 random samples, and comparing the actual correla-
 tion with the distribution based on the random
 samples.

 10.2.1.2.4 Necessary Hardware and Skills

     Appropriate sampling equipment and trained
 field  and laboratory personnel are required for
 chemical analyses, toxicity testing,  and benthic
 infaunal analyses.   Although the  equipment re-
 quired can be both costly and sophisticated, it is
 commonly necessary for sediment contamination
 investigations.  The necessary.equipment, facili-
 ties,  and expertise are generally available through
  a wide variety of government, university, commer-
  cial, and private facilities.

  10.2.1.3 Adequacy of Documentation

      Documentation for use  of this  method  is
  provided by Long and Chapman (1985), Chapman
  (1986,  1989, 1990), and Chapman et al. (1986,
   1987a,  1991a,  1991b).  Other investigators have
   also successfully applied this method (cf. Chap-
   man et al.,  1991c).

   10.2.2  Applicability of Method to Human
          Health, Aquatic Life, or Wildlife
          Protection

       This  approach is directly applicable to the
   protection of aquatic life.  To date, only benthic
invertebrates and fish have been used to assess in
situ  biological effects  and  sediment  toxicity.
Protection of aquatic  life may indirectly protect
wildlife (e.g., wading birds feeding  on  benthos)
and humans (e.g., via consumption of aquatic life).
The approach can be directly applicable to human
health and wildlife protection if the Triad compo-
nents  are  redirected  towards  issues  such  as
bacterial contamination and toxic contaminant bio-
accumulation.  For instance, Triad could be used
in  three  ways to address bacterial problems:
(1) measure bacterial contamination in water or
sediment, (2) measure bacterial diseases or con-
centrations  in tissues, and  (3) perform laboratory
tests  to  quantify  relationships between  sedi-
 ment/water concentrations and  effects.   Toxic
 contaminant bioaecumulation could be  addressed
 by these uses of the Triad  approach: (1) measure
 toxic contaminant concentrations  in  water or
 sediment, (2)  measure bioconcentration/biomag-
 riification in tissues, and  (3) perform laboratory
 tests  to determine effects related  to bioconcen-
 tration and biomagnification.


  10.2.3  Ability of Method to Generate
         Numerical Criteria for Specific
         Chemicals
        i '      -        '.                 '

      The Triad approach has been used to generate
  criteria for three  contaminants: lead, PAH, and
  PCBs  (Chapman, 1986).  These criteria were
  developed in Puget Sound by examining large data
  sets to identify  contaminant areas and  concen-
  trations lhat were associated with no or minimal
  biological effects. The criteria fall within a factor
  of 2 to 10 of values generated for these contami-
  nants by  the screening-level  concentration (see
  Chapter 11,  Section ll.l.l-)>  *e A^ apProach
  (see Chapter 11), and laboratory toxicity methods
  (Chapman et al., 1987b).  As detailed by Chap-
  man (1989), the AET application of the Triad
  concept provides criteria for benthic infauna and
   each bioassay conducted, whereas the latter conv
  bines all bioassay  and in  situ biological  effects
   data to provide a single value, interpretation, or
  ' analysis.  However, there  has been little work
   since Chapman  (1986)  on  development  of the
                                                                                              10-12

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    Sediment Classification Methods Compendium
   Triad approach for the production of numerical
   sediment quality criteria separate from AET.
   10.3 USEFULNESS

   103.1  Environmental Applicability

      Although the Triad approach is both labor-
   intensive and expensive, its strengths render  it
   extremely cost-effectiye for the level of infor-
   mation provided,   First,  it provides  empirical
   evidence of sediment quality (based on observa-
   tion,  not theory).  Second, it  allows  ecological
   interpretation of physical, chemical, and biological
   properties (i.e., interpretation of how these relate
   to the real environment). Third, it uses a prepon-
   derance-of-evidence approach rather than relying
   on single measurements  (i.e.,  all  the data  are
   considered). Because of the comprehensive nature
   of Triad studies, additional follow-up studies are
  usually not necessary. Finally, the data generated
  by the Triad approach can  be  used to generate
  effects-based classification indexes.
      The Triad approach enables investigators to
  estimate  the size of degraded and  nondegraded
  areas.  It also provides a test of the quality of
  reference areas (i.e., do contamination or biologi-
•  cal  effects occur?).  Standards  in  the form of
  sediment quality criteria (Chapman, 1986,  1989;
  PIT, 1988a, 1988b) can be set from  the contami-
  nant concentrations that are always associated with
  effects, using the AET application of the Triad.
  The Triad approach also provides the information
  necessary to describe the ecological relationships
  between sediment properties and biota at risk from
  sediment contamination.
     The Triad approach has been used in dredging
 studies to support dredged material disposal  siting
 and disposal  decisions (Chapman, unpublished).
 In multiplying the relative degree of degradation
 at a site by the volume of sediment to be dredged,
 investigators can compare different sites, provided
 that the same reference area is used.  This  com-
 parison helps investigators  determine  whether
 dredging  will affect useful  habitat or result in
 material unacceptable for ocean disposal.  Similar-
   ly, potential disposal sites can be compared with
 ,  each other and with the material to be dredged,
   and then compared to acceptability criteria for
   various uses and options. This application of the
   Triad approach  replaces similar but less useful
   comparisons based solely on the  total mass  of
   chemical contaminants to be dredged.
      In areas where benthic communities have been
   eliminated or  drastically changed  because, of a
   natural event (e.g.,  storms, oxygen depletion) or
  physical anthropogenic impact (e.g., recent dredg-
  ing, boat scour), the other two Triad components
  (sediment chemistry and toxicity) provide informa-
  tion when conventional univariate  approaches
  would prove deficient.  Such cases emphasize the
  need to use knowledge of an area in making any
  type of environmental  assessment,  including the
  Sediment Quality Triad.
      The Triad  approach can be used to discern
  and  ultimately to  monitor  regional  trends  in
  sediment quality.  Such information is necessary
  to delineate areas that  are excessively  contami-
  nated with toxic chemicals affecting the biota and,
  therefore, most in need  of remedial  action. Pilot
  studies  of this nature  have  been conducted  in
  Puget Sound and San Francisco Bay  (Long and
  Chapman, 1985; Chapman, 1986; Chapman etal.,
  1986, 1987a) and in Europe (e.g., Chapman, in
 press; Chapman et al, in press).
 3.1.1   Suitability for Different Sediment
        Types

    The Triad approach can be used with  all
 sediment types, including sands, muds, aerobic
 sediments, and anaerobic sediments.  It includes
 sediment characterization with physical parameters
.[e.g., grain size, acid volatile sulfides (AVS), and
 total organic carbon (TOC)] that may be important
 in interpreting the Triad compounds.  For exam-
 ple, caution must be used in  interpreting the
 results of toxicity  tests  in sediments that remain
 anaerobic in  the  laboratory  despite aeration.
 Specifically, organisms  will die  from lack  of
 oxygen,  making it difficult to  distinguish that
mortality from toxicity due to high concentrations
of contaminants.
10-12

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                                                                           10—Triad Approach
10.3.1.2  Suitability for Different Chemicals or
         Classes of Chemicals
       •
    The  Triad  approach can  be used  with all
chemicals or classes of chemicals, provided that
bioassay organisms and tests are appropriate for
all chemicals. For this reason, a battery  of bioas-
say tests Is recommended.  Caution must be used
when testing sediment extracts that may be specif-
ic to certain chemical classes.  Interpretation of
the  results  must  be  restricted  to  only  those
chemicals.                                ..   - .
                    i     -        .

10.3.1.3 Suitability for Predicting Effects on
         Different Organisms

    Application of the Triad  approach can  be
limited by the organisms in the environment if the
in  situ effects are determined primarily by the
same species used in the bioassay tests. In other
words, all biological effects data are based on a
single species. In^such cases, independence of the
infaunal community  analyses  and bioassay test
results cannot be assumed. Hence, more than one
bioassay test is recommended.  Ideally, the tests
would include a wide variety  of organisms, life
 stages, feeding types, and exposure routes.

 10.3.1.4 Suitability for In-Place Pollutant Control

     The Triad approach provides a comprehensive
 approach to in-place pollutant control because it
 allows for assessment of all potential interactions
 between chemical mixtures and the environment.
 This method is comprehensive because it includes
 the measurements of multiple chemicals as well as
 the potential toxic effects of both measured and
 unmeasured chemicals.

 10.3.1.5 Suitability for Source Control

     The Triad approach is as  suitable  for source
 control as  it is for in-place pollutant control. It
 can be an  environmental  complement to toxicity
 reduction evaluation (TRE) programs that involve
 chemical and toxicity investigations of sediments,
 and effluents and other discharges.
10.3.1.6  Suitability for Disposal Applications

    The Triad approach has been used for disposal
applications, including Navy Homeporting work in
San Francisco  Bay.   In that  study, the Triad
approach clearly separated potential dredge sites
from one another in terms of the relative level of
pollution. Although the Triad was not used in the
final decision because of other  considerations,
decision-makers were able to use  information
provided by the Triad to compare the suitability of,
dredging and disposal options.

10.3.2 General Advantages and Limitations

     The following are the major advantages of the
Triad approach:

     •  Combines three separate components to
        provide   a   preponderance-of-evidence
        approach;

     »   Does not require a,priori  assumptions
         concerning the  specific  mechanisms of
         interaction between organisms and toxic
         contaminants;

     •   Can be used to develop sediment quality
         values (including criteria) for any mea-
         sured contaminant  or  a combination of
         contaminants, including  both acute and
         chronic effects;

     •   Provides empirical  evidence of sediment
         quality;

     •   Can be used for any sediment type;

     •   Allows ecological interpretation  of  both
         physical-chemical and biological proper-
         ties; and

     .•   Does not usually required follow-up when
         a complete study is conducted.

     The following are the major limitations to the
  Triad approach:
                                                                                            10-13

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 Sediment Classification Methods Compendium
     •  Statistical  criteria have not been  fully
        developed for use with the Triad approach
        (but see Section 10.2.1.2.3, Types of Data
        Required);

     •  Rigorous  criteria  for calculating single
        indexes from each of the sediment chem-
        istry,  bioassay, and in situ biological
        effects data sets have not been developed
        (but may not be required);

     »  A large database is required;

     •  If the approach is used to determine sin-
        gle-chemical criteria, results could  be
        strongly influenced  by, the presence of
        unmeasured toxic contaminants that may
        or  may ' not  covary  with  measured
        chemicals;

     •  Methods for sediment  bioassay  testing
        need to be standardized;

     •  Sample collection, analysis, and  inter-
        pretation are labor-intensive and  expen-
        sive; and

     •  The choice  of  a  reference site is often
        made  without adequate  information  on
        how degraded the site may be.

10.3.2,1 Ease of Use

    The Triad approach is relatively easy to use
and understand. The concept is straightforward.
A high level of chemical and biological expertise
is required to obtain the data for the three separate
Triad components. However, many laboratories or
groups of laboratories  possess  the  required
expertise.

103.2.2 Relative Cost

    Relative cost can be evaluated in either dollars
or environmental damage. The Triad approach
may not prevent environmental damage, but it can
be used to identify contaminated areas for future
remediation. In terms of dollars, the Triad  ap-
 proach requires substantial resources to be imple-
 m'ented properly, although step-wise, tiered use of
 Triad components is possible.  Measured against
 the potential environmental damage due to toxic
 contamination  and the costs of remediation, the
 Triad approach can be extremely cost-effective.

 10.3.23 Tendency to Be Conservative

    The Triad approach provides  objective data
 with which to determine and sometimes to predict
 environmental  damage.    Its predictive ability
 allows for, but does not require, conservatism on
 the part of the  decision-makers.

 10.3.2.4 Level of Acceptance

    The Triad approach is gaining a high level of
 acceptance in various parts of North America and
 in  Europe (Forstner et al.,  1987;  Chapman, in
 press). In addition, Canada has conducted Triad
 studies in Vancouver to determine the suitability
 of  this approach for implementation of the new
 Canadian Environmental Protection Act (Gross et
 al., 1991; Cross et al., in review).

 10.3.2.5 Ability to Be Implemented by
         Laboratories with Typical Equipment
         and Handling Facilities

    All aspects of the Triad approach (i.e., benthic
 infaunal studies, sediment chemistry analyses,
 sediment toxicity bioassays) can be conducted by
 any competent,  specialist laboratory that is reason-
 ably well equipped.  The major requirements are
 adequate QA/QC procedures for chemical mea-
 surements; appropriate detection limits;  and, for
biological  analyses, taxonomic experts and a
taxonomic  reference library or museum.

10.3.2.6 Level of Effort Required to Generate
        Results

    Different levels of effort will generate differ-
ent levels of results.  For instance, results can be
generated by simply measuring one or two chemi-
cals, determining the number of infauna present,
and conducting a single sediment toxicity bioas-
10-24

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                                                                          10—Triad Approach
say.  However, the applicability of these results
may be severely limited. Consequently, multiple
chemicals including inorganic and organic com-
pounds should be measured, and in situ biological
alteration and sediment toxicity should be mea-
sured multiple times.  Although it is  possible to
use  previously  collected  nonsynoptic data to
derive results in a "paper" study (Long and Chap-
man,  1985),  fieldwork and synoptic  sampling
generate the most useful results.

10.3.2.7 Degree to Which Results Lend
         Themselves to Interpretation

     Beyond the general conclusions noted in Table
 10-2,  expert judgment is required to implement
 and interpret the Triad approach. In particular, the
 definition of "minimal" and "severe" biological
 effects is required to establish chemical-specific
 criteria. The Triad approach reflects the complex-
 ity  of the issues that must be addressed to assess
 environmental quality.

 10.3.2.8 Degree of Environmental Applicability

     As discussed, the Triad  approach  has an
 extremely  high degree of environmental  applies
 bility (see Section 10.3.1).
                  I
  10.3.2.9 Degree of Accuracy and Precision

     The  accuracy and precision of the  Triad
  approach have not been quantitatively determined.
  It  is  expected to have a high degree of accuracy
  and precision, although these parameters will vary
  with those of the constituent components.
   10.4  STATUS
   10.4.1  Extent of Use

      •Development of the formalized Triad concept
   has occurred relatively recently (Long and Chap-
   man, 1985; Chapman, 1986, 1990;  Chapman et
   al, 1986, 1987a, 1988, 1991a).   The Triad ap-
   proach has been used directly to establish  sedi-
   ment  quality criteria  (Chapman,   1986)  and,
through date manipulations, to determine  AET
values for sediment quality criteria (Tetra Tech,
1986a; PTI, 1988a, 1988b).
    The Triad has been used to identify spatial
and temporal trends of pollution-induced degrada-
tion.  Indexes developed using the Triad approach
can be numeric (as described in Chapter 11 for the
AET application of the Triad concept) or primarily
descriptive (see Figure 2, Chapman et al, 1987a).
In  either case, the Triad approach provides an
objective identification of sites where contami-
nation is causing discernible harm (cf. Power et
 10.4.2 Extent to Which Approach Has Been
        Field-Validated

     Because the Triad approach measures in situ
 biological alteration in the field, field validation is
 an   integral   part  of  each  complete  Triad
 investigation.

 10.4.3 Reasons for Limited Use
            /                    •  -••
     As previously described, the Triad approach is
 being used in the United  States, Canada, and
 Europe for marine, estuarine, and freshwater areas.
 It is not being used in small projects because of
 the  cost  and  expertise   required  for   full
  implementation.

  10.4.4 Outlook for Future Use and Amount
         of Development Yet Needed

      The  following areas of the Triad approach
      require development:

       •  Determining the appropriateness of  the
          various endpoints of different bioassays,
          selected chemical contaminants, selected
          measures of benthic community structure,
          and other potential measures of in  situ
          biological alteration;

       n  Determining  the  appropriateness  of an
          additive treatment of the data (e.g., sum-
          ming bioassay responses  to provide a
          single index for toxicity);

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  Sediment Classification Methods Compendium
      »  Further development of statistical criteria;

      •  Development of rigorous criteria for de-
         termining, where and if appropriate, com-
         posite indexes for each of the three Triad
         components; and

      •  Continued standardization of methods for
         sediment toxicity bioassays.

 Even without development  of these areas, the
 Triad approach  provides valuable information.
 The argument has been made (Chapman et al.,
 1986, 1987a)  that the  Triad approach  provides
 objective information on which to judge the extent
 of pollution-induced degradation.  For this reason
 the Triad approach will likely be used much more
 widely in future.
 10.5 REFERENCES

 Alden, R. W. II. 1992. Uncertainty and sediment
     quality assessments: I. Confidence limits for
     the Triad. Environ. Toxicol. Chem. 11:637-
     644.
 Chapman, P.M.  1986.  Sediment quality criteria
     from the Sediment Quality Triad - an exam-
     ple. Environ. Toxicol. Chem. 5: 957-964.
 Chapman, P.M.,  R.N. Dexter,  S.F.  Cross, and
     D.G.  Mitchell.   1986.   A field trial of the
     Sediment Quality Triad in San Francisco Bay.
     NOAA Technical Memorandum NOS OMA
     25.  National Oceanic and Atmospheric Ad-
     ministration, San Francisco, CA. 127 pp.
 Chapman,  P.M., R.N. Dexter, and E.R. Long.
     1987a. Synoptic measures of sediment con-
     tamination, toxicity and infaunal community
    structure  (the Sediment Quality Triad) in San
    Francisco Bay.  Mar. Ecol. Prog. Ser. 37:75-
    96.
 Chapman, P.M., R.C. Barrick, J.M. Neff,and R.C.
    Swartz. 1987b. Four independent approaches
    to developing sediment quality criteria yield
    similar values for model contaminants.  En-
    viron. Toxicol. Chem. 6:723-725.
Chapman, P.M. 1988. Marine sediment toxicity
     tests,  pp. 391-402.  In:  Chemical and Bio-
     logical  Characterization  of  Sludges, Sedi-
     ments, Dredge Spoils, and Drilling Muds. J.J.
     Lichtenberg, F.A Winter, Clr Weber, and L.
     Fradkin (eds.). ASTM STP 976. American
     Society  for Testing and  Materials, Philadel-
     phia, PA.
 Chapman, P.M.   1989*   Current approaches to
     developing sediment quality criteria. Environ.
     Toxicol. Chem. 8: 589-599.
 Chapman, P.M., CA McPherson, and K.R. Mun-
     kittrick.  1989.. An assessment of the ocean
     dumping tiered  testing approach using the
     Sediment Quality Triad.  Unpublished report
     prepared for Environmental Protection Cana-
     da.  E.V.S.  Consultants, North  Vancouver,
     BC, Canada.                   ,
 Chapman,  P.M., and E.A. Power.  1990.  Sedi-
     ment toxicity evaluation. American Petroleum
     Institute Publication No. 4501. 209 pp.
 Chapman,  P.M.   1990."  The Sediment Quality
     Triad  approach  to  determining  pollution-
     induced  degradation.   Sci. Total Environ.
     97/8:815-825.
 Chapman,  P. M..  In  press.  Pollution status of
     North Sea sediments—An international scien-
     tific study. Mar. Ecol. Prog. Ser.
 Chapman, P.M., R.N. Dexter, H.A Andersen, and
     B.A. Power.   1991a.   Evaluation of  effects
     associated with an oil platform,  using the
     Sediment Quality  Triad.  Environ. Toxicol.
     Chem.  10:407-424.
 Chapman, P. M., E. A. Power, and G. A. Burton,
    Jr. 1991b. pp. 313-340. Chapter 14: Integra-
    tive assessments in aquatic ecosystems. In:
    Contaminated Sediment Toxicity Assessment
    G. A.  Burton  Jr.  (ed.).  Lewis Publishers,
    Chelsea, Michigan.
Chapman, P.M., E.R.- Long, R.  C. Swarcz, T.H.
    DeWitt, and R. Pastorok.  1991c.   Sediment
    toxicity tests, sediment chemistry and benthic
    ecology do provide new insights into the
    significance and management of contaminated
    sediments - a reply to  Robert Spies. Environ.
    Toxicol. Chem. 10:1-4.
Chapman,  P.M., R.C.  Swartz,  B. Roddie, H.
    Phelps,  P.  van den Hurk  and  R.  Butler. In
    press.  An international comparison of sedi-
20-26

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                                                                         10—Triad Approach
   ment toxicity tests in  the North  Sea.  Mar.
   Ecol..Prog. Ser.
Cross, S.F., J.M. Boyd, P.M. Chapman, and R.O:
   Brinkhurst. 1991. A multivariate approach for
   defining spatial  impacts using the Sediment
   Quality Triad, p. 886. In: Proceedings of the
   17th Annual Aquatic  Toxicity Workshop,
   P.M.'Chapman, F.  S. Bishay, E. A. Power, K.
   Hall, L. Hardking. D. McLeavy, M. Nassichuk
   and W. Knapp (eds.). Can. Tech. Kept. Fish.
   Aquat. Sci.  1774.
Cross, S. F., J. M. Boyd, P. M. Chapman, and R.
   O. Brinkhurst. (In review).   A multivariate
   approach to assessing  the spatial extent of
   benthic impacts established using the Sedi-
   ment Quality Triad. Environ. Toxicol. Chem.
DeWitt,  T.  H., G.  R.   Distworth, and R. C.
    Swartz.  1988.  Effects of natural sediment
    features on survival of the  Phoxocephalid
    arriphipod,  Rhepoxynius  abronius.    Mar.
    Environ. Res. 24:99-124.
DeWitt, D. M., J. D. Mahony, D. J. Hansen, K. J.-
    Scott, M. B. Hicks, S. M. Mayr, and  M. S.
    Redmond,  1990.  Toxicity  of cadmium  in
    sediments:  the  role of > acid volatile sulfide.
    Environ. Toxicol.  Chem. 9:1487-1502.
DiToro, D.M., J.p. Mahony, DJ. Hansen, KJ.
    Scott,  M.B.  Hicks, S.M. Mayr,  and M.S.
    Redmond. 1900. Toxicity of cadmium in sedi-
    ments: The role of acid volatile sulfide. En-
    viron. Toxicol.  Chem. 9: 1487-1502.
Forstner, V.U., F.  Ackermann,  J. Alberti, W.
    Calmano, F.H.  Frimmel,  K.N. Kornatzki, R.
    Leschber, H. Rossknecht, U. Schleichert, and
    L. Tent. 1987.  Qualitatskriterien fur Gewas-
    sensedimente - Allgemeine Problematik und
    internationaler stand der Diskussion.  Wasser-
    Abwasser-Forsch  20:54-59.
 Keith, L.H., W. Crummett, J. Deegan, Jr., RA.
    Libby,  J.K.  Taylor, and G.  Wentler. 1983.
    Principles  of environmental analysis.  Anal.
    Chem. 55:2210-2218.
 Legendre, P and M.J.  Fortin. 1989. Spatial pattern
    and ecological analysis.  Vegetatio 80:107-
     138.                                 ,
 Long, E. R. 1989. The use of the  Sediment Qual-
     ity Triad in classification of sediment contam-
     ination, pp. 78-93. In: Marine Board, National
    Research  Council  Symposium/Workshop on.
    contaminated marine sediments.
Long, E.R., and M,F. Buchman. 1989. An evalu-
    ation of  candidate  measures of biological
    effects  for the National Status  and Trends
    Program.   NOAA Technical Memorandum
    105 pp. NOS OMA 45: National Oceanic and
    Atmospheric Administration, Rockmille, MD.
Long, E.R., and P.M. Chapman. 1985.  A sedi-
    ment quality triad:   measures of sediment
    contamination, toxicity and infaunal commun-
    ity composition in Puget Sound.  Mar. Poll.
    Bull. 16:405-41*5.
Malueg, K.W., G.S. Schuytema, D.F. Krawczyk,
    and J.H.  Gakstatter. 1984. Laboratory sedi-
    ment toxicity tests, sediment chemistry and
    distributions of benthic macroinvertebrates in
    sediments from the Keweenaw Waterway,
    Michigan.  Environ. Toxicol. Chem. 3:233-
    242.
Mantel, N. 1967. The detection of disease cluster-
    ing  and  generalized regression  approach.
    Cancer Res. 27:200-209.
Nemec,  A.F.L.,  and  R.O. Brinkhurst.  1988a.
    Using the bootstrap to assess statistical signifi-
    cance in  the cluster analysis of species abun-
    dance data.  Can.  J. Fish. Aquat Sci. 45:965-
    970.
 Nemec, A.F.L.,  and R.O. Brinkhurst. 1988b. The
    Fowlkes-Mallows statistic and the comparison
    of  two  independently  determined dendro-
    grams. Can  J. Fish. Aquat. Sci. 45:971-975.
 Power, E. A., K. R. Munkittrick, and P. M. Chap-
    man. 1991.  An ecological impact assessment
    framework  for decision making related to
    sediment quality,  pp.  48-64.  In:  Aquatic
    Toxicity  and Risk  Assessment: Fourteenth
    Volume.  M. A. Mayers and M. G. Barren
    (eds.). ASTM STP 1124. American Society
    • for Testing  and Material, Philadelphia, PA.
 PTI Environmental Services, Inc. 1988a.  Sedi-
    ment quality values refinement: Tasks 3 and
    5 -1988 update and evaluation of Puget Sound
    AET.  Unpublished report prepared for Tetra
    Tech, Inc. for the Puget Sound Estuary Pro-
     gram, EPA Contract No.  68-02-43441.  PTI
     Environmental Services, Inc., Bellevue, WA.
                                                                                          30-37

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 Sediment Classification Methods Compendium
 PTI Environmental Services, Inc. 1988b. Briefing
     report to the EPA Science Advisory Board:
     the  Apparent Effects Threshold  approach.
     Unpublished  report  prepared  for Battelle
     Columbus Division, EPA Contract No. 68-03-
     3534.    PTI Environmental  Services, Inc.,
     Bellevue, WA.
 Swartz, R.C., W.A. DeBen, ICA. Sercu, and J.O.
     Lamberson. 1982. Sediment toxicity and the
     distribution of amphipods in Commencement
     Bay, Washington, USA.   Mar. Poll.  Bull.
     13:359-364.
 Tetra Tech.  1986a.  Recommended protocols for
     measuring selected environmental variables in
     Puget Sound.  Prepared for the Puget Sound
    Estuary Program, U.S. Environmental Protec-
     tion Agency, Region X, Seattle, Washington,
    Tetra Tech, Inc., Bellevue, WA.
Tetra Tech. -1986b.  Development of sediment
    quality values for Puget Sound. Prepared for
    Resource Planning Associates and U.S. Army
    Corps of Engineers, Seattle District, for the
    Puget  Sound  Dredged  Disposal  Analysis
    Program. Tetra Tech, Inc., Bellevue, WA.
U.S. EPA. 1991. Analytical method of determina-
    tion of acid volatile sulfide in sediment. U.S.
    Environmental Protection Agency,  Criteria
    and Standards, Washington, DC.
Wiederholm, T., A-M. Wiedefholm, and G. Mil-
    brink. 1987.  Bulk sediment bioassays with
    five  species  of  fresh-water oligochaetes.
    Water, Air and Soil Pollut. 36: 131-154.
Zar, J. H.  1984. Biostatistical Analysis, 2d  ed.
    Prentice-Hall, Englewood Cliffs, NJ.
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          CHAPTER 11
Apparent  Effects Threshold Approach

John Malek                                  .   M      _]  .   v
Office of Puget Sound, U. S. Environmental Protection Agency Region X
1200 Sixth Avenue, Seattle, WA 98101                       ,    .'-,.-'•
(206) 553-1286
    In the Apparent  Effects Threshold (AET)
 approach, empirical data are used to identify
 concentrations of specific chemicals above which
 specific biological effects would always be expect-
 ed. Following the development of AET values for
 a particular geographic area, they can be used to
 predict whether statistically significant biological
 effects are  expected  at a station with known
 concentrations of toxic chemicals.
 11.1 SPECIFIC APPLICATIONS

 11.1.1  Current Use

     At present, the AET approach is being used
 by several programs as guidelines for the protec-
 tion of aquatic life in Puget Sound. These guide-
 -lines are the culmination of cooperative planning
 and scientific investigations that were initiated by
 several federal and state agencies in the early and
 mid-1980s.
     Three programs and applications of the AET
 approach are highlighted  below.   Notably, all
 these  programs  involve an  element  of direct
 biological testing in  conjunction with the use of
 AET values, in  recognition  of the fact that no
 approach to chemical sediment quality values is
  100 percent reliable in predicting adverse biologi-
  cal effects.   An underlying strategy in many of
  these programs was  to develop two sets of sedi-
  ment  quality values based primarily  on AET
  values:

      •   One set of values identifies low chemical
          concentrations below which . biological
          effects are improbable.

      •   A second set of values identifies higher
          chemical  concentrations  above  which
          multiple biological effects are expected.
The programs incorporate direct biological testing
in concentration  ranges between these two  ex-
tremes to serve as a "safety net" (i.e., to account
for the uncertainty of chemical predictions) for
potential adverse effects or anomalous situations
at "moderate" chemical concentrations.

Commencement Bay NearshorelTideflats
Superfund Investigation

    Commencement Bay is a heavily industrial-
ized harbor in Tacoma, WA. Recent surveys have
indicated  over 281  industrial activities in the
nearshore/tideflats area. Comprehensive shoreline
surveys have identified more than 400 point and
nonpoint  source discharges in the study  area,
consisting primarily  of seeps, storm drains, and
open  channels.  A  remedial investigation  (RI)
under Superfund,  started  in  1983,  revealed 25
major sources contributing to sediment contamina-
tion,  including  major chemical manufacturing,
 pulp mills,  shipbuilding and repair, and smelter
 operations.  Adverse biological effects were found
 in sediments adjacent to these sources.
     The AET approach was developed during the
 course of the RI to assess sediment quality using
 chemical and biological effects data [i.e., depres-
 sions in the number of individual benthic  taxa,
 presence of tumors and  other abnormalities in
 bottom fish, and several laboratory toxicity tests
 (amphipod mortality, oyster larvae abnormality,
 bacterial bioluminescence)].   AET values  were
 also used in the subsequent feasibility study (FS)
 to identify cleanup goals arid define volumes of
  contaminated sediment for remediation.  The AET
 values used in the FS  were generated from a
  reduced set of biological effects indicators, which
  comprised depressions in total benthic abundance,
  amphipod  mortality, oyster  larvae abnormality,
  and bacterial luminescence.

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   Sediment Classification Methods Compendium
  Puget Sound Dredged Disposal Analysis Program

      In  1985, the Puget Sound Dredged Disposal
  Analysis (PSDDA)  program was initiated to
  develop environmentally safe and publicly accept-
  able options for unconfined, open-water disposal
  of dredged material.  PSDDA  is  a cooperative
  program conducted under the direction of the U.S.
  Army Corps of'Engineers (Corps) Seattle District,
  U.S. EPA Region X, the Washington Department
  of Ecology (Ecology), and the  Washington  De-
  partment of Natural Resources  (WDNR).  AET
  values  were used  to develop chemical-specific
  guidelines to determine whether biological testing
  on  contaminated  dredged  material is  needed.
  Results of the biological testing help  determine
  suitable disposal alternatives.
     Above a specified chemical concentration (i.e.,
  the screening-level concentratipn or SLC) biologi-
  cal testing is required to determine the  suitability
  of dredged  material for  unconfined, open-water
  disposal.  Based primarily on AET values  for
  multiple biological indicators, a higher "maximum
  level concentration" was also  identified. Above
  this latter concentration, failure of biological tests
  is considered to be predictable.  However, an
  optional series of biological tests can be conducted
 under PSDDA  to demonstrate the suitability of
 such contaminated material  for unconfined, open-
 water disposal (Phillips et al., 1988).

 Urban Bay Toxics Action Program

    The Urban  Bay Toxics Action Program is a
 multiphase program to control pollution of urban
 bays in Puget Sound. The program includes steps
 to identify areas where contaminated sediments
 are  associated  with adverse biological effects,
 specify potential pollution  sources, develop  an
 action plan for source control, and form an action
 team for plan implementation.  Initiated in 1984
 by Ecology and U.S. EPA Region X's Office of
 Puget Sound, the program is a major component
 of the Puget Sound Estuary Program  (PSEP).
 Substantial participation has also been provided by
 the Puget Sound  Water Quality Authority  (Author-
 ity) and  other state  agencies and local govern-
ments. Major funding and  overall guidance for
  the program is provided by U.S. EPA Office of
  Wetlands, Oceans and Watersheds.
      In the PSEP urban bay program, AET values
  are used in conjunction with site-specific biologi-
  cal tests during the assessment of sediment con-
  tamination  to  define  and rank  problem  areas.
  Source control actions are well under way, but
  sediment remediation has not yet begun at any of
  the sites (PTI,  1988).

  11.1.2 Potential Use
                   *
      The AET approach to determining sediment
  quality can also be used as follows:

      •  To determine the spatial extent and rela-
        tive priority  of  areas  of contaminated
        sediment;

      •  To identify potential problem chemicals in
        impacted  sediments and,  as  a result, to
      .  focus   cleanup  activities  on  potential
        sources of problem contaminants;

     •  To define and prioritize laboratory studies
        for determining cause-effect relationships;
        and                     .

     •  With appropriate  safety factors or other
        modifications,  to  screen  sediments  in
        regulatory programs that involve extensive
        biological testing.

 Proposed regulations for sediment contamination
 are currently under review in Puget Sound. These
 regulations  may include use of AET values to
 develop  statewide  sediment quality standards.
 Ecology is currently developing a suite of sedi-
 ment management standards, as mandated by the
 Puget Sound Water Quality Authority (1988) in its
 1989 Management Plan. The proposed standards
 are based in part on AET values.  Development of
 these standards (Becker et al., 1989) relies heavily
 on the  past and ongoing  efforts  described in
 Section 11.1.1 and involves active participation by
Ecology, U.S. EPA, the Authority, WDNR, the
 Corps (Seattle District), and various public interest
groups.   The draft  regulation currently  under
23-2

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                                                                             11—AET Approach
development affects  only sediments  in Puget
Sound.  As additional data become available from
other locations, the adopted regulation will eventu-
ally  be broadened  and modified to include the
entire state.
11.2 DESCRIPTION

11.2.1 Description of Method

    AET values are derived using a straightfor-
ward algorithm that relates biological and chemi-
cal data from field-collected samples.  For a given
data set, the  AET for a given chemical  is the
sediment concentration above which  a particular
adverse biological effect (e.g., depressions in the
total abundance of indigenous benthic infauna) is
always statistically significant (PsO.05) relative to
appropriate reference conditions.  The calculation
of an AET  for each  chemical  and biological
indicator is conducted as follows:

    (1) Collect "matched" chemical and biological
        effects data-^-Conduct chemical and bio-
        logical  effects  testing on subsamples  of
        the same field sample. (To  avoid unac-
        countable losses of benthic organisms,
        benthic infaunal  and chemical  analyses
        are conducted on separate samples collect-
        ed concurrently at the same location.)

     (2) Identify "impacted" and  "nonimpacted"
        stations—Statistically test the significance
        of adverse biological effects relative to
        suitable reference conditions   for each
        sediment sample.   Suitable   reference
        conditions are established by  sediments
        exhibiting very low or undetectable con-
        centrations of any toxic chemicals, an ab-
        sence of other adverse effects, and physi-
        cal characteristics that are directly compa-
        rable with those of the test sediments.

     (3) Identify AET  using only "nonimpacted"
        stations—For each chemical, the AET can
        be identified for a given biological indica-
        tor as the highest  detected concentration
       among  sediment  samples that do not
       exhibit statistically significant effects. (If
       the chemical  is undetected in  all non-
       impacted samples, then no AET can be
       established for that chemical and biolog-
       ical indicator.)

   (4) Check for preliminary AET—-Verify that
       statistically  significant biological  effects
       are observed at a chemical concentration
       higher than the AET; otherwise, the AET
       should be regarded only as a preliminary
       minimum estimate.

   (5) Repeat Steps (l)-(4) for each biological
       indicator.

   The AET approach  for  a group of  field:col-
lected sediment samples is shown in Figure 11-1.
The  samples were collected at various locations
and were analyzed for (1) toxicity in a laboratory
bioassay and (2) the concentrations of a suite of
chemicals, including lead  and 4-methylphenol.
Based on the results of bioassays conducted on the
sediments from each station, two subpopulations
of all sediments are represented by bars in the
figure:           •

     •   Sediments that did not exhibit statistically
        significant  (P>0.05) toxicity relative to
        reference conditions ("nonimpacted" sta-
        tions) and

     •   Sediments   that  exhibited   statistically
        significant  (PssO.05) toxicity in bioassays
        relative to reference conditions ("impact-
        ed" stations).

     Over the observed range of concentrations for
these sediment samples (horizontal axis in Figure
 11-1), the sediments fall into two groups for each
chemical:

     •  At low to moderate  concentrations, signif-,
        leant sediment toxicity occurred  in some
        samples, but not in others.
                         \
     m  At  concentrations   above  an  apparent
        threshold  value,  significant  sediment
        toxicity occurred in all samples.
                                                                                              11-3

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Sediment Classification Methods Compendium
                                   Lead
                     SP-14
         IMPACTED    |
                          660 ppm
                on
                  ii INI MI M ii mi
         NONIMPACTED
RS-18
                                            on   a   a
                                       AET
I   I  I I I I MM  I  I  I I II III   I  I I I I I III   I  I I I I I III

1          10         100       1000      10000

           INCREASING CONCENTRATION
                                                              100000
                              4-Methylphenol
                    RS-18

           IMPACTED   \
                              3600 ppb     SP-14
                                            I
                   Q UlUUlHUlLOHIDCn nCaam • n rTTTli mmii
                      imim*i'.mm>——n
          NONIMPACTED
                                       an   a
                                            IQ
                                          AET
               I  I  i I tlilil   I I limn  i ,i i linn   I i IIIIIH  i  i i linn  i  i i Mini

               1        10       100      1000     10000    100000   100000

                        INCREASING CONCENTRATION         >
Figure 11-1. The AET approach for a group of field-collected sediment samples.
7770 AET approach applied to sediments tested for lead and 4-methylphenol concentrations and
toxicity response during bioassays.
11-4

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                                                                             11—AET Approach
    The AET value is defined for each chemical
as the highest concentration of that chemical in
the sediments that did not exhibit sediment toxici-
ty.  Above this AET value, significant sediment
toxicity was always observed  in  the  data set
examined.   Data are treated  in  this manner to
reduce the  weight given  to  samples in which
factors other than the contaminant examined (e.g.,
other contaminants, environmental variables) may
be responsible for the biqlogical effect.
    For each chemical, additionalAET values
could be  defined for other biological indicators
that were tested (e.gi, other bioassay responses or
depressions in  the abundances of certain indige-
nous benthic infauna).

11.2.1.1 Objectives and Assumptions

    The objective of the AET approach is  to
 identify concentrations of contaminants that are
 associated exclusively with sediments exhibiting
 statistically significant biological effects relative to
 reference sediments.  AET value generation is a
 conceptually simple process and incorporates the
 complexity  of biological-chemical interrelation-
 ships in  the environment without relying  on  a
 priori assumptions about the mechanisms of these
 interrelationships.  Although the AET  approach
 does not  require  specific  assumptions  about
 mechanisms of the uptake and toxic action  of
 chemicals, it does rely on more general assump-
 tions regarding  the  interpretation of  matched
 biological  and chemical data for  field-collected
  samples, as described below:

      •  For a given chemical, concentrations can
         be as high as the AET value and not  be
         associated with statistically  significant
         biological  effects (for  the indicator  on
         which the AET was based).
                     r'               '   ' •      '
      •  When biological impacts are observed at
          concentrations below an AET value for a
          given chemical, it  is  assumed  that  the
          impacts may be related to another chemi-
          cal, chemical interactive effects, or other
          environmental  factors (e.g.,   sediment
          anoxia).
    •  The AET  concept is  consistent with a
       relationship between  increasing concen-
       trations of toxic chemicals and increasing
       biological effects (as observed in laborato-
       ry exposure studies).         .

    The assumptions in interpreting environmental
data are demonstrated below with actual field data.
IFsing Figure 11-1 as an example, sediment from
Station SP-14 exhibited Severe toxicity, potentially
related to a greatly elevated concentrations of 4-
methylphenol (7,400 times reference levels). The
same sediment from  Station  SP-14 contained a
relatively low concentration of lead that was well
below the AET for lead (Figure 11-1).  Despite
the  toxic  effects  associated with  the sample,
sediments from many other  stations with  higher
lead concentrations than Station SP-14 exhibited
no  statistically  significant  biological effects.
These results were interpreted to suggest that the
effects at Station SP-14 were potentially associat-
ed with 4-methylphenol  (or  a substance  with a
similar environmental distribution) but were less
 likely to be associated with lead.  A converse
 argument can be  made  for  lead and 4-methyl-
 phenol in sediments from Station RS-18.
     Applied in this manner, the AET approach
 helps to  identify  measured chemicals that are
 potentially associated with  observed effects at
 each biologically  impacted  site  and  eliminates
 from  consideration chemicals  that are far less
 likely to be associated with effects (i.e., the latter
 chemicals have been observed at higher concentra-
 tions at other sites without associated biological
 effects).  Based  on  the results for lead and
 4-methylphenol, bioassay toxicity  at five of the
 impacted sites shown in the figure may be associ-
 ated with elevated  concentrations of  4-methyl-
 phenol, and  toxicity  at eight other sites  may be
 associated with elevated concentrations of lead (or
 similarly distributed contaminants).
     As illustrated by these results, the occurrence
  of biologically impacted stations at concentrations
  below  the AET  of  a single chemical does  not
  imply that AET values in general are not protec-
  tive  against biological  effects, only that single
  chemicals may not account for all stations with
  biological effects.   By developing  AETs  for
                                                                                               11-5

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  Sediment Classification Methods Compendium
 multiple chemicals,  a high  percentage of  all
 stations with biological effects are accounted for
 with the AET approach (see Section 11.3.2.9 and
 USEPA, 1988).
     AETs can be expected to be more predictive
 when developed from a large, diverse database
 with wide ranges of chemical concentrations and
 a wide diversity of measured chemicals. Data sets
 that have large concentration gaps between sta-
 tions and/or do not cover a wide range of concen-
 trations  must be scrutinized carefully (e.g.,  to
 discern  whether chemical concentrations in. the
 data set  exceed reference concentrations) to deter-
 mine whether AET generation is appropriate.

 21.2.1.2 Level of Effort

 11.2.1.2.1 Type of Sampling Required

    Collection of field data for initial generation
 of AETs is a labor-intensive and capital-intensive
 process.    The exact  level  of sampling effort
 required depends on the amount and variety of
 data collected (e.g., the number of samples collect-
 ed, the diversity of biological indicators that are
 tested, and the range of chemicals measured).
 One means of minimizing these costs is to com-
 pile existing data that meet appropriate  quality
 assurance criteria. There are no definitive require-
 ments  for the size and variety  of  the  database,
 although a study of the predictive abilities of the
AET approach with Puget Sound data (Barrick et
al., 1988) resulted in the following recommend-
ations for data collection:

    •  Collect or compile chemical and biologi-
       cal effects data from 50 stations or more
       (and from suitable reference areas).

    •  Bias the positioning of stations to ensure
       sampling of various contaminant sources
     • (e.g., urban environments with a range of
       contaminant sources and, preferably, with
       broad  geographic  distribution)  over a
       range of contaminant concentrations (pref-
       erably  over at least 1-2 orders of magni-
       tude).

    •  Conduct chemical tests for a wide range
       of chemical classes (e.g., metals, nonionic
         organic  compounds, ionizable organic
         compounds).  To  generate AETs  on an
         organic  carbon-normalized  basis,  total
         organic carbon (TOC) measurements are
         required in all sediments.

     •  Ensure that detection limits of <100 ppb
         (lower if possible) are attained for organic
         compounds.  High detection limits (i.e.,
  (       insensitive analyses) can obscure  the
         occurrence of chemicals at low to mod-
         erate concentrations;  as noted previously,
         only detected data are used in AET calcu-
         lations.   Metals are naturally occurring
         substances, and most metals concentra-
         tions typically  exceed routine  detection
         limits.                                 -

     The only strict requirement for field sampling
 of data for AET generation  is the collection  of
 "matched" chemical and biological data (as de-
 scribed at the  beginning of Section  11.2.1).
 Matched data sets should be used to reduce the
 possibility that uneven  (spatially variable) sedi-
 ment  contamination could result  in associating
 biological and chemical data that are  based on
 dissimilar sediment samples.   Because the toxic
 responses of stationary organisms (e.g., bioassay
 organisms confined to a test sediment, or infaunal
 organisms largely confined to a small area) are
 assumed to be affected by direct association with
 contaminants in the surrounding environment, it is
 considered essential that chemical and biological
 data be collected from nearly identical subsamples
 from a given station.

 11.2.1.2.2 Methods

    Methodological details  for the  generation of
AET values  are  described  at the  beginning of
Section 11.2.1.

11.2.1.2.3 Types  of Data Required

    Two fundamental kinds of data analysis are
required for AET generation:

    •   Statistical analysis of the significance of
        biological  effects relative  to  reference

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                                                                           11^-AET Approach
       conditions (i.e., classification of stations
       as impacted or nonimpacted  for  each
       biological indicator) and
    •  Generation  of  an  AET value f6r each
       chemical and biological indicator (essen-
       tially a process of ranking stations based
       on chemical concentration).
    Additional kinds of data analysis needed for
AET generation are  quality  assurance/quality
control (QA/QC) review of biological and chemi-
cal data, and evaluation of the appropriateness of
reference area stations. These topics have been
described elsewhere (e.g., Beller et al.,  1986;
Barrick et al., 1988).
    The AET method does not intrinsically require
a specific method of statistical analysis for deter-
mination of significance  of biological  effects
relative  to reference conditions.  Existing Puget
Sound  AETs  have relied largely   on pairwise
t-tests; details of statistical analyses performed for
the generation of Puget Sound AET have been  de-
scribed elsewhere (USEPA, 1988; Barrick et  al.,
 1988; Beller et al., 1986).   For example,  the
 following steps were used to determine the statisti-
 cal significance of amphipod mortality  bioassay
 results  (Swartz  et al.,  1985) in field-collected
 sediments:

     •   All replicates from  all stations  in  the
         reference area used for  each study were
         pooled, and a mean bioassay response and
         standard deviation were calculated.

     •   Results from each  potentially  impacted
         site were then compared statistically with
         the reference  conditions using pairwise
         analysis.                            '

      •  The Fffi« test (Sokal and Rohlf, 1969) was
         used to test for homogeneity of variances
         between each pair of mean values.

      B   If  variances  were  homogenous, then a
          t-test was used to compare the two means.

      D   If variances were not homogenous, then an
          approximate t-test (Sokal and Rohlf, 1969)
         .was used to compare the two means.
    •  Statistical significance was tested with a.
       pairwise error rate of 0.05 to ensure con-
       sistency among studies of differing sam-
       ple sizes.
    Data analyses that have been applied to other
biological  indicators are  described elsewhere
(Beller et al., 1986; Barrick et al, 1988). Nota-
bly, comparisons  to  reference conditions were
somewhat more complicated for benthic infaunal
abundances than  for sediment bioassays.   For
benthic infaunal comparisons, reference data for
each potentially impacted site were categorized so
that comparisons were made with samples collect-
ed during the same season, at a similar depth, and
whenever possible,  in  sediments with  similar
particle  size characteristics (i.e., percentage of
particles <64 urn) as those of the potentially im-
pacted site. In this manner, statistical comparisons
were normalized  to account for the influence of
three  of the major  natural variables known to
 influence the abundance and distribution of ben-
 thic macroinvertebrates.  All benthic data were
 also log-transformed so that data distributions
 conformed to  the assumptions of the parametric
 statistical tests that were applied.  Additional data
 treatment methods presented elsewhere (Barrick et
 al., 1988) are not discussed further herein, because
 they are not considered intrinsic to the AET ap-
 proach, but rather are options, to address poten-
 tially unusual  matrices or biological conditions.

                            "  '   • •           5
  11.2 1.2.4 Necessary Hardware and Skills

     The primary skills  required for AET genera-
  tion are related to the development of the biologi-
  cal/chemical database. Expertise in environmental
  chemistry is  required to evaluate  chemical date
  quality, and the meed for normalization of chemi-
  cal data  and related  factors.   Biological and
  statistical expertise are required for the determina-
  tion of statistical significance, For benthic data in
  particular, evaluation  of appropriate reference
  conditions and  knowledge of benthic taxonomy
  and ecology .are necessary.
      Computers are recommended for the efficient
  generation of AET Values.  A menu-driven data-
  base (SEDQUAL) has been developed for U.S.
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   Sediment Classification Methods Compendium
  EPA Region X that is capable of a number of data
  manipulation  tasks,  including  the  following:
  (1) storing chemical and biological data, (2) calcu-
  lating AET values, (3) comparing a specified set
  of  AET to stored sediment  chemistry  data  to
  identify  stations  at  which  adverse biological
  effects are or are not predicted, and (4) based on
  such comparisons, calculating the rate of correct
  prediction of biological impacts. The SEDQUAL
  system, which  requires an IBM-AT  compatible
  computer with a hard disk, has been documented
  in detail in a user's manual (Nielsen, 1988).  The
  SEDQUAL database currently includes stored data
  from Puget Sound (over 1,000 samples, not all of
  which have biological and chemical data).

  1L2.L3 Adequacy of Documentation

      Various  aspects of the AET  approach have
  been extensively documented  in reports prepared
  for U.S. EPA and other regulatory agencies, as
  listed below and in the reference list:

      •   Generation of Puget Sound AET values
         and evaluation of their predictive ability
         (Seller et al., 1986; Barrick et al, 1988);

     •   Data used to generate Puget Sound AET
         values (appendices of Seller et al, 1986
         and field surveys cited in Seller et  al,
         1986 and Barrick et al, 1988);

     •   Briefing report to the U.S. EPA Science
        Advisory Board (USEPA, 1988); and

     •  Policy implications of effects-based ma-
        rine sediment criteria (PTI, 1987).

 11.2.2 Applicability of Method to Human
       Health, Aquatic Life, or Wildlife
       Protection

    The AET approach has been designed for use
 in evaluating potential adverse impacts to aquatic
 life associated with  chemical  contamination  of
 sediments. By empirically determining the associ-
 ation between chemical contamination and adverse
 biological effects, predictions can be made regard-
 ing the levels of contamination that are always
 associated  with  adverse  effects  (i.e.,  the AET
  values). These critical levels of contamination can
  then be used to develop guidelines for protecting
  aquatic life (e.g., sediment quality values). AETs
  can be developed for any kind of aquatic organism
  for which biological responses to chemical toxicity
  can be measured.  The protectiveness of the AET
  can therefore be ensured by evaluating organisms
  and biological responses with different degrees of
  sensitivity to  chemical toxicity.  For  example,
  evaluations of metabolic changes (i.e., usually a
  very sensitive biological response) in a pollution
  sensitive species  would  likely result  in AET
  values  that are lower and more protective than
  evaluations of mortality, (i.e.,  generally a less
  sensitive response)  in a more pollution-tolerant
  species. The protectiveness of AETs can also be
  ensured through the application of "safety factors."
  For example, to be protective of chronic biological
  responses, a factor based on an acute-chronic ratio
  could be applied to AETs developed on the basis
  of acute biological responses.

  11.23 Ability of Method to Generate
        Numerical Criteria for Specific
        Chemicals

     The AET approach is not intrinsically limited in
 application to specific  chemicals or chemical groups.
 In general, the approach can be used for chemicals
 for which data are available.  However, when using
 a specific data set to generate AETs, it is preferable
 that AET generation be limited to chemicals with
 wide  concentration  ranges  (e.g.,  ranging  from
 reference concentrations to concentrations near direct
 sources) and/or with appropriate detection frequen-
 cies (e.g., greater than 10 detections). A partial list
 of chemicals for which AETs have been developed
 is presented in Table 11-1.
113    USEFULNESS

11 J.I Environmental Applicability
113.1.1  Suitability for Different Sediment Types

    The  AET approach  can be applied to any
sediment type in saltwater or freshwater environ-
ments for which biological tests can be conducted.
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                                                                            11—AET Approach
        Table 11-1. Selected Chemicals for Which AETs Have Been Developed in Puget Sound.
  Antimony
  Arsenic
  Cadmium
  Chromium
                                     ORGANIC COMPOUNDS
   Low-Molecular-Welght PAHs
   Naphthalene
   Acenaphthylene
   Acenaphthene
   Fluorene
   Phenanthrene
   Anthracene
   2-Methylnaphthalene
   Phthalates
   Dimethyl phthalate
   Diethyl phthalate,
   Di-n-butyl phthalate
   Butyl benzyl phthalate
   Bis (2-ethylhexyl) phthalate
   Di-n-octyl phthalate
   ••^—i   *~
   Pesticides

   p.p'-DDE
   p.p'-DDD
   p.p'-DDT
High-Molecular-Weight PAHs

Fluoranthene
Pyrene
Benz(a)anthracene
Chrysene  :
Benzofiuorantheries
Benzo(a)pyr,ene
IndenoCl^.a-c.dypyrene
Dibenzo(a,h)anthracene
Benzo(g,h,i)perylene
   	~—^—	—
Total PCBs
 Miscellaneous Extractables

 Benzyl alcohol
 Benzoic acid
 Dibenzbfuran
 Hexachlorobutadiene
 N-Nitrosodiphenylamine
1,3-Dichlorobenzene
1,4-Dichlorobenzene
1,2-Dichlorobenzene
1,2,4-Trichlorobenzene
Hexachlorobenzene (HCB)
Phenols
Phenol
2-Methylphenol
4-Methylphenol
2,4-Dimethylphenol
Pentachlorophenol
 Volatile Organic*
 Tetrachloroethene
 Efriylbenzerie
 Total xylenes
 By normalizing chemical concentrations to appro-
 priate sediment variables (e.g., percent organic
 carbon),  differences between different sediment
 types can  be  minimized  in the generation of
 AETs.  In practice, identification of unique or
 atypical sediment matrices is important in deter-
 mining the general applicability of AET values
 generated from a specific set of data.
     Differences in  physical  characteristics (e.g.,
 grain size, habitat exposure)  are one major factor
 that may account for stations not meeting predic-
. tions based on existing AET values.  In. Puget
 Sound studies, for example, fine-grained sediments
                   dominated stations that had significant amphipod
                   mortality that had not been predicted, and coarse-
                   grained sediments dominated stations  that had
                   significant depressions in benthic infauiia that had
                   not been predicted by benthic AETs  (Barrick et
                   al., 1988).

                   11.3.1.2 Suitability for Different Chemicals or
                            Classes of Chemicals

                       There are no constraints  on  the  types of
                   chemicals for which AETs can be developed. An
                   AET can be developed for any measured chemical
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  Sediment Classification Methods Compendium
  (organic or inorganic) that spans a wide-concentra-
  tion range in the data set used to generate AETs.
  The availability of a wide diversity of chemical
  data increases the probability that toxic agents (or
  chemicals that covary  in the environment with
  toxic agents)  can be'included in interpreting
  observed biological impacts.
     To date, AETs have been developed for over
  60 chemicals frequently detected in the environ-
  ment, including 16 polycyclic aromatic hydrocar-
  bons (PAHs); several alkylated PAHs and related
  nitrogen-, sulfur-, and oxygen-containing hetero-
  cycles; polychlorinated biphenyls (PCBs) (reported
  as total PCBs); 5 chlorinated benzenes; 6 phthalate
  esters;  3  chlorinated  hydrocarbon  pesticides;
  phenol and 4  alkyl-substituted and chlorinated
 phenols;  10 metals and metalloids; 3  volatile
 organic compounds; and 5 miscellaneous extract-
 able substances.   Data for other miscellaneous
 chemicals that  were less frequently detected or
 analyzed for in the Puget Sound area were also
 evaluated for their potential use in developing
 AETs (e.g., resin acids and .chlorinated phenols in
 selected sediments from areas influenced by pulp
 and paper mill activity).
    AETs have been  developed  for  chemical
 concentrations normalized to sediment dry weight
 and sediment organic carbon content (expressed as
 percent of dry weight sediment).  Using a 188-
 sample data set from Puget Sound, AETs were
 also developed for data normalized to fine-grained
 particle content (expressed as the percent of  silt
 and clay, or <63-nm particulate material,  in dry
 weight of sediment). These latter AET values did
 not appear to offer advantages in predictive reli-
 ability over the more commonly used dry weight
 and TOC normalizations (Beller et al, 1986).

 113.1.3  Suitability for Predicting Effects on
         Different Organisms

    The AET approach  can be used to  predict
 effects on any life stage of any marine or aquatic
 organism  for which a  biological  response  to
 chemical toxicity can be determined. Because the
 approach is empirical, relying on direct measure-
ment of  the chemical concentrations associated
with samples exhibiting adverse effects, the results
 are directly applicable to predicting effects on the
 organisms used to generate the AET. The results
 can also be used to predict  effects  on nontarget
 organisms by ensuring that the organisms used to
 generate an AET are either representative of the
 nontarget  organisms or  are more  sensitive to
 chemical toxicity than  those organisms.   For
 example, AETs generated for a species of sensi-
 tive amphipod might be considered as protective
 of the chemical concentrations associated with
 adverse effects in other species of equally or less
 sensitive amphipods.  At the same time, these
 AET might be considered protective of most other
 benthic macroinvertebrate taxa because they are
 based  on  a member  of  a  benthic taxon (i.e.,
 Amphipoda) that is considered to  be sensitive to
 chemical toxicity (Bellan-Santini,  1980).   By
 contrast, AETs generated  for a pollution-tolerant
 species such as the polychaete Capitella capitata
 (cf.  Pearson  and Rosenberg, 1978)  might be
 considered  representative  for  other  pollution-
 tolerant species, but not protective for most other
 kinds of benthic macroinvertebrates.

11.3.1.4 Suitability for In-Place Pollutant Control

    In remedial action programs, assessment tools
such as the AET approach can be used to address
the following specific regulatory needs:

    •   Provide a preponderance of evidence for
        narrowing a list of  problem chemicals
        measured at a site;

    •   Provide  a predictive  tool for cases in
       which  site-specific  biological  testing
       results are not available;

   •  Enable designation  of  problem   areas
       within the site;

   •  Provide a consistent  basis on which to
       evaluate  sediment  contamination and to
       separate  acceptable from  unacceptable
       conditions;

   •  Provide an environmental basis for trig-
       gering sediment remedial action; and
33-30

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                                                                            11—AET Approach
    m  Provide a reference point for establishing
       a cleanup goal.

Because AET values are derived from sediments
with multiple contaminants, they incorporate the
influence of  interactive effects in  environmental
samples. The ability to incorporate the influences
of chemical mixtures, either by design or default,
is an advantage  for the assessment of in-place
pollutants.

113.1.5 Suitability for Source Control

    The AET approach is well suited for identify-
ing problem  areas. Because specific cause-effect
relationships are not proven for specific chemicals
and biological effects, remedial actions should not
be designed  exclusively for a  specific chemical.
(This caution applies to all approaches because of
the complex mixture of contaminants in environ-
mental samples.) The link between problem areas
and potential sources  of contamination is estab-
 lished by  analysis of concentration gradients of
 contaminants in  these  problem  areas  and  the
 presence  and  composition  of contaminants in
 sediments and source materials.   The AET ap-
 proach provides  a means of narrowing the list of
 measured  chemicals that should be considered for
 source control and provides .supportive, evidence
 for eliminating chemicals from consideration that
 appear to  be present at a concentration tob low to
 be  associated with adverse  biological  effects.
 Reduction of the overall  contaminant load to a
 problem area such that all measured chemicals are
 below their  respective AETs is predicted to result
 in mitigation of the adverse biological effects. It
 is possible  that such source controls may.be
 effective  because of the concomitant removal of
 an unmeasured contaminant.

 11.3.1.6  Suitability for Disposal Applications

     The evaluation of potential biological impacts
  associated with the disposal of dredged material is
  an important component  in  the designation of
  disposal sites and review of disposal permits for
  dredged material, AET values provide a prepon-
  derance of  evidence in determining a "reason.to
believe" that sediment contamination could result
in adverse biological effects.  Hence, the AET
approach is a useful tool for assessing the need for
biological testing during the evaluation of disposal
alternatives.   It is assumed that AET  values
generated for in-place sediments provide a useful
prediction of whether  adverse biological  effects
will occur in  dredged material after disposal at
aquatic sites.
 11.3.2  General Advantages and Limitations

 11.3.2,1 Ease of Use

    In this section, "use" is treated as both genera-
 tion and application.  The ease of generating AET
 values depends on the status of the data to be used
 for AET generation (i.e., whether field data have
 been collected and whether statistical significance
 has been determined for biological indicators). It
 is recommended that a search for existing data be
 conducted as part of determining the need for
 collecting new samples. The existing database of
 matched biological and chemical data  from Puget
 Sound comprises over 300 samples.  Collection of
 new field data (e.g., for application outside of
 Puget Sound) would require a considerable expen-
 diture of effort, as would the statistical analysis of
 a large number of samples.  HdweVer, if data are
 available  and statistical analyses have been  per-
 formed, the generation of AET values is very easy
 with the SEDQUAL database (described in Sec-
 tion 11.2.1.2.4).  The menu-driven system allows
 for a considerable amount of flexibility in choos-
 ing stations and biological indicators to be includ-
 ed in AET generation^  Application of AET (i.e.,
 comparison of AET values to chemical concentra-
 tions in field samples)  is also very easy when
 using  SEDQUALj provided that the field  data
 have been computerized.  Application  of AET
 values to chemical data presented  in existing
 • literature is also  straightforward.

  11.3.2.2  Relative Cost

      The cost of developing AET values can span
  a wide range, depending on the stage of database
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  Sediment Classification Methods Compendium
  development and the numbers and kinds of chemi-
  cals  and biological indicators used.  The least
  costly means of developing the values is to use
  existing chemical and biological information, thus
  minimizing  the  expenses associated  with  field
  sampling  and  laboratory  analyses.  (Selective
  sampling to confirm whether existing AET values
  are applicable would still be useful.) The histori-
  cal database could be based on the pooled results
  from various studies conducted in a region, pro-
  viding that each study passed QA/QC performance
  criteria and satisfied the prerequisites of the AET
  approach (e.g., matched chemical and biological
,  measurements and  the  ability to  discriminate
  adverse biological effects).
     If the historical database is judged inadequate
  to generate AETs for a region, then the costs of
  field measurements of chemical concentrations in
  sediments and associated biological  effects must
  be incurred to develop the database.  These costs
  can vary substantially, depending on the chemicals
  and biological indicators evaluated.  Costs would
  be minimized if evaluations were  based on  a
  limited range of chemicals and a single, inexpen-
  sive biological test.  It  is recommended that the
  approach be based on a relatively .wide range of
  chemicals, and if possible, several kinds of biolog-
  ical indicators.
     The  existing database for the Puget Sound
 region is based on a wide range of chemicals (i.e.,
 U.S. EPA priority pollutants  and other selected
 chemicals) and four kinds of biological indicators.
 The costs for developing  AETs varied consider-
 ably among the four indicators.  For example,
 laboratory costs for the  least expensive indicator
 (Microtox bioassay) were approximately $200 per
 station,  whereas costs  for the most expensive
 indicator (abundances of benthic macroinverte-
 brates) were as high as $1,800 per station. There-
 fore, within the  existing database, the range of
 costs for biological testing spanned almost 1 order
 of magnitude.
     Once AET values have been generated, use of
 these values to predict the occurrence of biological
 effects is relatively inexpensive.  Chemical data
 may be  compared to AET values by using the
 SEDQUAL database  or  through, manual data
manipulations.                   ''    ,
  11.3.2.3 Tendency to Be Conservative

     The empirical, field-based nature of the AET
  approach precludes definitive a priori predictions
  of its tendency to be either over- or underprotec-
  tive of the  environment.   The occurrence  of
  biologically impacted  stations at concentrations
  below the AET of a given chemical (see. Figure
  11-1) may appear to be underprotective. Howev-
  er, the occurrence of impacted stations at concen-
  trations below the AET of a single chemical does
  not imply that AETs in general are not protective
  against biological effects, only that single chemi-
  cals may not account for all stations with biologi-
  cal effects.  If AETs are developed for multiple
  chemicals, the approach can  account  for a high
 percentage  of stations with  adverse  biological
 effects.
     To date, AETs have been developed for acute
 sediment bibassays of mortality" in  adult am-
 phipods, developmental  abnormality  in  larval
 bivalves, and metabolic alterations in bacteria. All
 of  these organism/endpoiht  combinations  are
 considered to be sensitive to chemical  toxicity.
 AETs have also been generated for in situ reduc-
 tions  in the  abundances  of  benthic macro-
 invertebrates. Because these reductions  incorpo-
 rate chronic (i.e., long-term) exposure to contami-
 nants, they  can also be considered as sensitive
 measures of the effects of  chemical  toxicity.
 However, a more protective approach would be to
 use the lowest of the four kinds of AET for each
 chemical as  the concentration on which predic-
 tions are made.  Alternatively, the protectiveness
 of any kind of AET could be modified by devel-
 oping sediment quality values based on "safety
 factors" applied to existing AETs.

 11.3.2.4 Level of Acceptance

    The AET approach has been  accepted  by
 several federal and state agencies in the Puget
 Sound region as one tool in providing guidelines
 for regulatory decisions. U.S. EPA has used AET
values to develop sediment quality values with
which to evaluate the potential toxicity of contam-
inated sediments in urban bays. PSDDA has used
AET values as a todl to develop chemical guide-
11-32

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                                                                           11—AET Approach
lines for determining whether biological testing is
necessary for dredged  sediments proposed  for
unconfmed, open-water disposal.  Ecology has
used  AET  to develop sediment  management
standards.  These standards were promulgated by
the State of Washington and  approved by EPA
Region X in 1991. The standards are being used
by a number of water quality programs (e.g.,
source control, remediation).
    Several  major  characteristics influence  the
acceptability of the AET approach.  The most
attractive characteristic  of the .approach  is proba-
bly the reliance on empirical information based on
field-collected sediments or indigenous organisms,
and exposure of laboratory  test organisms to
environmental samples.    A second  attractive
feature of the approach is the setting of an AET at
the chemical concentration  in the data set above
which adverse biological effects are always ob-
served.  This characteristic provides consistency
that, with a representative database used to gener-
 ate AETs, enhances the preponderance of evidence
 of adverse effects in the environment.  The AET
values can  be  updated as new information is
 collected.  The AET approach can also be applied
 to an existing database in new regions,  providing
 certain prerequisites are met by the database (e.g.,
 synoptic measurement of chemical and biological
 data, and QA/QC guidelines).
      A  limitation of the AET approach is that
 field-based  approaches do not  directly assess
 cause-effect relationships.  Because sediments in
 the environment are often contaminated with a
 complex mixture of chemicals, it is difficult when
 using field-collected sediment for any approach to
 relate observed biological effects  to  a single
 chemical. The approach also requires selection of
 appropriate normalized chemical data to address
 the bioavailability of contaminants  to organisms.
 Organic carbon normalization may be most appro-
 priate for honpolar organic contaminants based on
 theoretical considerations.  In addition, nonprotec-
 tive AETs could be generated if unusual matrices
  (e.g., slag) that anomalously restrict bioavailability
  are included in  the database used to generate the
 AETs, or if biological test results are  incorrectly
  classified.  Recommended data treatment guide-
  lines for chemical and biological data are dis-
cussed by Barrick et al. (1988).  The AET ap-
proach was  reviewed by the U.S. EPA Science
Advisory Board (SAB, 1989),  which noted the
method had "major strengths  in  its ability to
determine biological effects and assess .interactive
chemical effects."

11.3.2.5 Ability to Be Implemented by
        Laboratories with Typical Equipment
         and Handling Facilities

    If applicable data  do not already exist,  the
development of AET values requires a relatively
extensive amount of field sampling and laboratory
analysis.   The chemical analyses required  for
development of AET represent standard analytical
procedures.   A laboratory with  appropriately
trained staff should be able to conduct the neces-
sary  benthic community  analyses and sediment
bioassays.  Specific methods for performing the
chemical and biological tests that were used to
develop Puget  Sound  AET are detailed in the
Puget Sound Protocols (Tetra Tech, 1986). These
 efforts can  be minimized by using historical data
whenever possible. Once  AETs are developed,
 their routine implementation is relatively easy. In
 addition, they can be easily updated as additional
 data become available.

 11.3.2.6 Level of Effort Required to Generate
          Results

     As noted in Section 11.3.2.1, the SEDQUAL
 database facilitates AET generation and applica-
 tion.  After field  data have been collected, the
 most  tune-consuming task  is  data entry  and
 verification. Entry of chemical and biological data
 for 50 samples requires roughly 16 person-hours
 (assuming  75  chemicals have been measured and
 biological  effects are being coded simply  as
  "impacted" or "nonimpacted").  Generating a set
  of AET values for a given biological indicator, 75
  chemicals, and 50 stations takes approximately
  0.75-1 h of computer time on  SEDQUAL (and
  about 5 min of labor to set up the analysis). To
  compare a set of AET (for 75 chemicals) to a 50-
  sample set of field data takes approximately 0.5-
  0.75  h  of computer  time on SEDQUAL (and
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  Sediment Classification Methods Compendium
 roughly 5 min of labor to set up the analysis).
 SEDQUAL is capable of comparing any kind of
 chemical sediment  criteria to field data, but re-
 quires that the numerical criteria be entered in the
 database.

 113.2.7 Degree io Which Results Lend
         Themselves to Interpretation

     The manner in which the AET approach can
 be used to interpret matched biological and chemi-
 cal  data  from field-collected sediments  is de-
 scribed in Section 11.2.1.  As noted previously,
 the  use of AET can help investigators eliminate
 chemicals from further consideration (as the cause
 of an observed effect);  however, the approach
 cannot identify specific cause-effect relationships.
 Because the AET approach is empirical, it is not
 well suited to identifying specific toxic agents or
 elucidating mechanisms of biological uptake and
 metabolism.  However, certain general relation-
 ships could be examined on an a posteriori basis
 with the AET approach (e.g., testing the relative
 importance of different ways  of  normalizing
 chemical concentration data in predicting adverse
 biological effects).
    A number of  environmental factors, may
 complicate the interpretation of the data.   Al-'
 though the AET concept is simple, the generation
 of AET values based on  environmental data
 incorporates  many complex biological-chemical
 interrelationships.   For example, the AET  ap-
 proach incorporates the net effects of the follow-
 ing factors that may be important in field-collected
 sediments:

    •   Interactive effects  of  chemicals  (e.g.,
        synergism, antagonism, and additivity);

    •   Unmeasured chemicals and other unmea-
        sured, potentially-adverse variables; and

    •   Matrix effects  and bioavailability (i.e.,
       phase associations between contaminants
        and sediments that affect bioavailability of
       the contaminants, such as the incorpora-
       tion of PAH in soot particles).

   The AET  approach cannot quantify the indi-
vidual  contributions of interactive effects, unmea-
 sured chemicals, or matrix effects in environmen-
 tal samples, but AET values may be influenced by
 these factors.   AET values are expected to be
 reliable predictors of adverse effects that could
 result from the influence of these environmental
 factors if the samples used to generate AETs are
 representative of samples for which AET predic-
 tions are made. Alternatively, isolated occurrenc-
 es of such environmental factors in a data set used
 to generate AETs may limit the predictive reliabil-
 ity of those AET values. If confounding environ-
 mental factors render- the  AET approach unreli-
 able, then this should be evident from validation
 tests in which biological effects are predicted in
 actual environmental samples.
     A more detailed discussion of the interpreta-
 tion of AETs  and the confounding effects of
 environmental  factors is presented in U.S. EPA
 (1988).

 11.3.2.8 Degree of Environmental Applicability

     The AET  approach has a high  degree of
 environmental applicability based on its reliance
 on chemical and biological measurements made
 directly on environmental  samples.  Such infor-
 mation provides tangible  evidence  that various
 chemical  concentrations either are  or  are  not
 associated with adverse  biological effects  in
 typically complex environmental settings.
    The environmental  applicability of the AET
 approach has been quantified for the four kinds of
 AET developed for Puget Sound by evaluating the
 reliability with which each kind of AET predicted
 the  presence or absence of adverse  biological
 effects  in  field samples  collected  from Puget
 Sound (USEPA, 1988).  The overall reliability of
 the  four tests  ranged from  85 to  96 percent,
 indicating that all four kinds  of AETs were rela-
 tively  accurate  at predicting  the  presence or
 absence of effects for samples  from the existing
 database.  This high level  of reliability suggests
 that AETs have  a relatively high degree of envi-
 ronmental applicability in Puget Sound, and it has
 been a primary factor  in  the  use  of the AET
 approach by agencies in the Puget Sound region.
AET values generated for Puget Sound have also
been used as examples of effects-based sediment
22-14

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                                                                              11—AET Approach
criteria to provide an initial estimate of the magni-
tude of potential problem areas in coastal regions
of the United States for the U.S. EPA Office of
Policy Analysis (PTI, 1987).

11.3.2.9 Degree of Accuracy and Precision

    In this section, accuracy is considered to be
the ability of AET to predict biological effects and
precision  represents  the  expected  variability
(uncertainty range) for a given AET value for a
given data set.
    In previous evaluations of the AET approach
and other sediment  quality  values using field-
collected data, the accuracy of the approach was
defined by two qualities:

     •  Sensitivity  in  detecting  environmental
        problems (i.e., are all biologically impact-
        ed sediments identified by the predictions
        of the chemical sediment criteria?)

     •  Efficiency  in  screening  environmental
        problems (i.e.,  are only biologically im-
        pacted sediments identified by the predic-
        tions of the chemical sediment  criteria?).

 Sensitivity is defined  as  the  proportion of all
 stations exhibiting adverse biological effects that
 are correctly  predicted using  sediment  criteria.
 Efficiency  is defined  as  the  proportion of all
 stations  predicted to  have adverse   biological
 effects  that  actually are  impacted.    Ideally, a
 sediment criteria approach should be efficient as
 well as sensitive. For example, a sediment criteria
 approach that sets values  for  a wide range of
 chemicals near their analytical detection limits will
 likely be conservative  (i.e., sensitive) but ineffi-
 cient.  That is, it will predict a large percentage of
 sediments  with biological effects.  It will also
 predict impacts at many stations where there are
 no biological effects, but chemical concentrations
 are slightly elevated.  The concepts of sensitivity
 and efficiency are illustrated in Figure 11-2.
     The overall reliability of any sediment criteria
 approach addresses both sensitivity and efficiency.
 This measure is defined as the proportion of all
 stations for which correct predictions were made
for either  the  presence or absence of adverse
biological effects:
Overall
rttiatilitj
  ..
AH stations ccmetlj pn&cttd
                  Total number of stations tvehiatetl
High reliability results from correct prediction of
a large percentage of the impacted stations (i.e.,
high sensitivity, few false negatives) and correct
prediction of  a large percentage  of the  non-
impacted stations (Le., high efficiency, few flake
positives). An assessment of AET reliability was
recently conducted using a large database compris-
ing samples from 13 Puget Sound embayments
(Barrick et al, 1988). These evaluations suggest
that the AET approach is relatively sensitive for
the biological indicators tested and also relatively
efficient.  For example, 68-83 percent sensitivity
and 55-75 percent efficiency were observed when
AETs generated from a  ISS-sample data set were
evaluated with an independent  146rsample data
set. The ranges of sensitivity and efficiency cited
 above  represent the ability of  benthic infaunal
AET values to  predict  statistically  significant
 depressions in the abundances of benthic infauna
 in field-collected samples and the  ability of am-
 phipod mortality bioassay AET values to. predict
 statistically  significant  mortality  in  bioassays
 conducted on field-collected sediment.
     Precision of the AET approach has not been
 as  intensively investigated as  accuracy.   AET
 values  are  the result  of  parametric statistical
 procedures (i.e., determination of the significance
 of biological effects relative to reference condi-
 tions) and nonparametric methods (e.g., ranking of
 stations by concentration), and thus are not amena-
 ble to the routine definition of confidence inter-
 vals.  However, the degree of  AET precision  is
 considered to depend on the following factors:
      •  The concentration range between the AET
         (determined by a nonimpacted station)
         and the next highest concentration that is
         associated with a statistically significant
         effect;
      Hi   Classification error associated with the
         statistical significance of biological indi-
                                                                                               11-15

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 Sediment Classification Methods Compendium
                      B
                                                                   PREDICTED
                               CORRECTLY PREDICTED
           SENSITIVITY- C/B x 100 = 5/8 x 100 = 63%
           EFFICIENCY = C/A x 100 = 5/7 x 100 = 71%
              FOR A GIVEN BIOLOGICAL INDICATOR:

                A ALL STATIONS PREDICTED TO BE IMPACTED
                B ALL STATIONS KNOWN TO BE IMPACTED
                C ALL STATIONS CORRECTLY PREDICTED TO BE IMPACTED
 figure 11-2. Measures of reliability (sensitivity and efficiency).
       cator results (i.e., whether  a  station is
       properly classified as impacted or non-
       impacted, as related to Type I and Type H
       statistical error);
    •  The weight of evidence  or number of
       observations supporting a  given AET
       value; and
    »  The analytical error associated with quan-
       tification of chemical results.
Detailed discussion of these factors is provided in
Beller et al, (1986).
    One approach used in Puget Sound to estimate
the uncertainty range around the AET value was
to define the lower limit as the concentration at
the nonimpacted station immediately below the
AET and to define the upper limit as the concen-
tration at the impacted station immediately above
the AET.  These limits are based largely on
probabilities of statistical classification error. For
data sets with large concentration gaps between
stations, such uncertainty ranges will be wider and
precision will be poorer than for data sets with
more continuous distributions.  The number of
22-26

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                                                                             1—AET Approach
stations used to establish an AET would be ex-
pected to have a marked effect on AET uncertain-
ty because small data sets would tend to have less
continuous distributions of chemical concentra-
tions  than  large data sets.  Based on  analyses
conducted with Puget Sound data, the magnitude
of the AET uncertainty for 10 chemicals or chemi-
cal groups that are commonly detected is typically
less than one-third to one-half of the value of the
AET itself (considering both amphipbd mortality
bioassay and benthic infaunal AET data). Based
on quality assurance information for these data,
analytical error is probably a minor component of
overall precision, particularly for  metals.
 11.4 STATUS

 11.4.1. Extent of Use

     The AET approach is used by several agencies
 and sediment management programs in the Pacific
 Northwest to provide guideline values for regula-
 tory decisions.  The State of Washington has
 developed sediment management standards primar-
 ily using the AET approach but also including
 equilibrium partitioning values.  These standards
 were promulgated by the State and approved by
 EPA, Region X, in 1991 and are currently being
 implemented  in  a variety  of programs.   The
 standards  are  the culmination  of  cooperative
 planning and scientific investigations by several
 federal and state agencies throughout the 1980's,
 including:

      •  Superfund investigations at  Commence-
         ment Bay and Eagle Harbor;

      •  Puget Sound Dredged Disposal Analysis
         (PSDDA);

      •  Urban Bay Toxics Action Program; and

      •"•' Puget Sound Water .Quality Authority
         Management Plan.
                         \                 •   •
  A key result of these efforts has been the recogni-
  tion by regulators of two separate levels of sedi-
ment contamination and has led to the develop-
ment of two sets of sediment quality values. This
separation in management use of sediment values
arose from the sensitivity and efficiency concepts
of reliability previously discussed.  This manage-
ment decision was made because it was deter-
mined  that none of the available approaches for
developing sediment quality values would result in
100 percent sensitive and 100 percent efficient
values.  Different strategies have been used by
different  programs.for  use of  AET-generated
values.   In general, the lowest AET  (termed
LAET) for any of die biological tests  is used to
establish  the  lower level where there  is  little
concern of sediment contamination (e.g., the goal
for remedial actions).  The AET approach has
developed higher chemical levels (termed HAET),
above which adverse effects are predicted for all
the biological tests.  In most regulatory programs,
 direct biological testing is allowed to resolve the
 differences in predictions of these two  sets of
 sediment quality values (i.e., prediction of adverse
 biological  effect  by highly  sensitive sediment
 quality values, which at lower chemical concentra-
 tions  are not predicted by highly efficient sedi-
 ment  quality values).   To date, such sediment
 quality values  developed were for and used in
 marine and estuarine environments. The State of
 Washington and  EPA, Region X, are gathering
 chemical and biological data to potentially develop
 companion values for freshwater sediments.
     Other  efforts  are  under way outside  Puget
 Sound and the Pacific Northwest  to develop sedi-
 ment quality values using the AET approach.  These
 include California and the Great Lakes region in the
 United States, and the countries of Canada, New
 Zealand, and Australia internationally.
  11.4.2 Extent to Which Approach Has Been
         Field-Validated

      As described in U.S. EPA (1988), the reliabili-
  ty of AETs generated from Puget Sound data was
  evaluated with tests  of sensitivity and efficiency
  (defined in Section 11.3-2T9). Tests of the sensi-
  tivity and efficiency of the AET approach were
  carried out in several steps, as described below:
                                                                                             11-17

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  Sediment Classification Methods Compendium
         The  chemical database was subdivided
         into groups of stations that were tested for
         the same biological  effects indicators.
         Specifically, all chemistry stations with
         associated amphipod bioassay data were
         grouped together (287 stations), all chem-
         istry  stations  with  associated benthic
         infaunal data were grouped together (201
         stations), all chemistry stations with asso-
         ciated oyster larvae bioassay data were
         grouped  together (56 stations), and all
         chemistry stations with associated Micro-
         tox bioassay data were grouped together.
         (50 stations).  Stations  with more than
         one biological indicator were included in
         each appropriate group.

         The stations in each group were classified
         as impacted or nonimpacted based on the
         appropriate statistical  criteria (i.e., F^
         and t-tests at alpha = 0.05).

         Several tests of reliability were conducted
         at this point:

          •   Test 1: AET values (dry weight)
              were  generated with the  entire
              Puget Sound database available in
              1988, and sensitivity and efficiency
              tests were performed against  the
              same database for each biological
              indicator.

          •   Test 2:  The test described above
              was repeated in two parts: (a) using
              TOC-normalized AET values  for
              nonionic organic compounds and
              dry weight-normalized AET values
              for all other compounds (i.e., ioniz-
              able organic compounds, metals,
              and  metalloids),  and (b)  using
             TOC-normalized data for all chemi-
             cals. Test 2 allowed for a posterio-
             ri evaluation of the relative success
             of dry weight and TOC normaliza-
             tion for nonionic organic chemicals.

          •   Test 3:  Because the efficiency of
             the AET based on the entire Puget
             Sound database is  100 percent by
              constraint (as in Tests  1  and 2),
              predictive efficiency was estimated
              by the  following procedure.   For
              each biological  indicator, a single
              station  was sequentially  deleted
              from the total database, AETs were
              recalculated for the remaining data
              set,  and  biological  effects  were
              predicted for the  single  deleted
              station.   The predictive  efficiency
              was the cumulative result for the
              sequential deletions of single sta-
              tions. For example, the 287-sample
              database for amphipod bioassay
              results can  be -used  to  provide a
              286-sample  independent database
              for predicting (in sequence) effects
              on all 287 samples.

          •   Test 4:   In this test,  independent
              data sets were used to generate and
              test AETs to confirm the sensitivity
              and  efficiency  measurements in
              Tests 1 and 3. AETs (dry weight)
              generated with 188  stations from
              diverse geographic regions in Puget
              Sound were tested with a comple-
              tely independent set of 146 Puget
              Sound stations.

    In  addition,  the  influence  of geographic
location and other  factors on AET predictive
ability  were  examined (Barrick  et al.,  1988).
Further testing of Puget Sound AET values using
matched  biological/chemical  data  from  other
geographic areas is desirable before recommend-
ing direct application of the Puget Sound values in
other geographic regions.
       Reasons for Limited Use
    The AET approach is being increasingly used
outside of Puget Sound and the Pacific Northwest
to evaluate  and compare  different  classes  of
sediments and to develop 'bay-, site-, or region-
specific sediment quality values for a variety of
regulatory uses.  Because the approach is based on
empirical data,  direct application of values from
11-18

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                                                                          11—AET Approach
Puget Sound or  another area to a specific bay,
site, or region usually encounters some conflicting
or confounding data. Because regional reference
areas are used to determine the significance of
adverse biological effects in the AET approach,
the AET developed  for one region may be over-
protective or underprotective of the resources in
the other area. Additionally, the mix of chemicals
in one region's sediments may not be the same in
another region.  The use of the AET approach
and use of specific AET values should not be con-
fused.
    Development of site-specific AETs for other
geographic areas may require additional sampling.
Because many past studies were not multidiscipli-
nary,  measurements were often made only for
chemistry or biology rather than for both kinds of
information. In  such cases, there will be a limited
amount of appropriate historical data that can be
used to develop AETs. The integration or com-
 parison of AET data sets among different regions
 can also be restricted because appropriate biologi-
 cal indicators  for  generating  AETs may vary
 among regions.

 11.4.4 Outlook for Future Use and Amount of
        Development Yet Needed

     The following two approaches to AET devel-
 opment could be particularly beneficial in expand-
 ing the use of this  approach:

     •  Use  of laboratory cause-effect (spiking)
         studies to evaluate AET predictions on a
      ."  chemical-specific basis and

      •  Use of a large set of matched biological/
         chemical data from different geographic
         areas to test the predictive ability of AET
         and to  test the "precision" of AET values
         based on data sets from different areas.

      The AET approach was presented to (USEPA,
  1988) and reviewed by the U.S. EPA Science
  Advisory  Board (SAB, 1989).  The SAB noted
  major strengths and limitations of the method and
  provided recommendations that would improve the
  validity of the AET values.   The method was
considered to contain sufficient merit for use in
developing  location-specific  sediment  quality
values.  Because of the specificity of the method,
i.e., the empirical applications at specific locali-
ties, under specific environmental conditions, the
approach seemed less useful for development of
general, broadly applicable (i.e., national) sedi-
ment quality criteria.
 11J  REFERENCES

 Barrick, R.C., S. Becker, L. Brown, H. Beller, and
   R. Pastorok.  1988.  Sediment quality values
   refinement:   1988 update and evaluation of
   Puget Sound AET.  Volume I.  Final Report
   Prepared for Tetra Tech, Inc. and U.S. Environ-
   mental Protection Agency Region X, Office of
   Puget Sound.  PTI Environmental Services,
   Bellevue, WA.  74 pp. + appendices.
 Becker, D.S., R.P. Pastorok, R.C. Barrick, P.N.
   Booth, and L.A. Jacobs.  1989.  Contaminated
   sediments criteria report.   Prepared for. the
   Washington Department of Ecology, Sediment
   Management Unit.  PTI Environmental Servic-
   es, Bellevue, WA.  99 pp. + appendices.
 Bellan-Santini,  D.  1980.  Relationship between
   populations of amphipods and pollution. Mar.
   Poll. Bull. 11:224-227.
 Beller,  H.R.,  R.C. Barrick, and D.S. Becker.
    1986.  Development of sediment quality values
    for Puget Sound. Prepared for Resource Plan-
    ning Associates, U.S. Army Corps of Engineers,
    Seattle District, and  Puget Sound Dredged
   * Disposal Analysis Program.  Tetra Tech, Inc.,
    Bellevue WA.  128 pp. + appendices.
  Nielsen, D.  1988.  SEDQUAL users manual.
    Prepared for Tetra' Tech, Inc. and U.S. Environ-
    mental Protection Agency Region X, Office of
    Puget Sound.   PTI Environmental Services,
   r Bellevue, WA.
  Pearson, T.H.,  and R. Rosenberg. 1978.  Macro-
    benthic succession in  relation to organic en-
    richment and pollution of the marine environ-
    ment  Oceanogr.  Mar.  Biol. Annu. Rev. 16:
    229-311.
                                                                                           13-19

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  Sediment Classification Methods Compendium
  Phillips, K, P. Jamison, J. Malek, B. Ross, C.
    Krueger, J. Thornton, and J.  Krull.   1988.
    Evaluation  procedures  technical  appendix-
    Phase 1 (Central Puget Sound).  Prepared for
    Puget Sound Dredged Disposal Analysis by the
    Evaluation  Procedures Work Group.   U.S.
    Army Corps of Engineers, Seattle, WA.
  Puget  Sound  Water Quality  Authority.   1988.
    1989 Puget  Sound Water  Quality Management
    Plan.  Puget Sound Water Quality Authority,
    WA. 276 pp.                        .
  PTI.  1987.  Policy implications of effects-based
    marine sediment criteria.  Prepared for Ameri-
    can  Management Systems and U.S. Environ-
    mental  Protection Agency, Office of Policy
    Analysis.  PTI Environmental Services,  Belle-
    vue, WA.
 PTI.  1988.  Elliott Bay Action Program:  1988
    action plan.  Prepared for Tetra Tech, Inc.  and
    U.S. Environmental Protection Agency.  PTI
    Environmental Services, Bellevue, WA. 43 pp.
    +• appendices.
 Sokal,  R.R., and F.J. Rohlf.  1969.  Biometry.
   WJf. Freeman and Company, San Francisco,
    CA.  859pp.
 State of Washington, Department  of Ecology.
   1991. Chapter 173-204, Washington Adminis-
   trative Code, Sediment Management Standards.
   Olympia, WA.
 Swartz, R.C., WA. DeBen,  J.K. Phillips,  J.D.
   Lamberson, and FA. Cole.  1985.  Phoxoce-
   phalid amphipod bioassay for marine sediment
   toxicity. pp. 284-307. In: Aquatic Toxicology
   and Hazard Assessment:  Proceedings of the
   Seventh Annual Symposium. R.D. Cardwell, R.
   Purdy, and R.C. Banner (eds.). ASTM STP 854.
   American Society for Testing and Materials,
   Philadelphia, PA.
 Tetra Tech. 1986.  Recommended protocols for
   measuring selected environmental variables in
   Puget Sound.  Final report.  Prepared for U.S.
   Environmental Protection Agency, Region X,
   Office of Puget Sound, Seattle, WA.  Tetra
   Tech, Inc., Bellevue, WA.
USEPA, 1989. Science Advisory Board.  Report
   of the Sediment Criteria Subcommittee, Evalu-
   ation of the Apparent Effects Threshold (AET)
   Approach for Assessing  Sediment  Quality.
   SAB-EETFC-89-027. Office of the Administra-
   tor, Science Advisory Board, Washington, DC.
USEPA.  1988. Briefing report to the EPA Sci-
   ence Advisory Board. Prepared for Battelle and
   U.S. Environmental  Protection  Agency,  Re-
  gion X, Office of Puget Sound,  PTI Environ-
  mental Services, Bellevue, WA;  57 pp.
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         -CHAPTER 12
A  Summary of the  Sediment  Assessment
Strategy Recommended  by  the  International
Joint Commission
 Philippe Ross
 The Citadel, Department of Biology
 Charleston, SC 29409
 (803)792-7875
    The International  Joint Commission (UC)
 Sediment Subcommittee has published a document
 entitled Procedures for the Assessment of Contam-
 inated Sediment Problems in the Great Lakes (IJC,
 1988a).  An  overview of the IJC strategy for
 assessing contaminated sediments is provided in
 this chapter.   However, because it  would  be
 inappropriate to reproduce all, or substantially all,
 of the document in this chapter,  the interested
 reader is referred to the IJC (1988a) document
 itself for an explanation of details that are not
 provided herein.            ,
  12.1   SPECIFIC APPLICATIONS

  12.1.1  Current Use

     The IJC (1988a) document is intended as
  guidance for  the  assessment  of  contaminated
  sediments in the Great Lakes. Its first application
  is in a work plan for sediment investigations at
  Great Lakes areas of concern (AOCs, as identified
  by the  UC).   Section  118(c)(3)  of the Water
  Quality Act of 1987 calls for U.S. EPA's Great
  Lakes National Program Office to survey at least
  five AOCs as part of a 5^yr study and demon-
  stration program called ARCS (Assessment and
  Remediation of Contaminated Sediments).  The
  strategy recommended  by DC (1988a) will be
  applied through a series  of activities involving
  physical mapping and characterization, sampling,
  chemical analyses, toxicity testing,  and in situ
  community analysis.  The assessment began in
  1989 and was completed in 1991.  The ARCS
  program also  seeks to improve  upon the  DC
(1988a)  approach by  comparing  various test
methods and by evaluating cost-effective recon-
naissance and screening methods.

12.1.2  Potential Use

    Other AOCs will eventually be evaluated in
the process of developing remedial action plans.
It is possible that other Great Lakes harbors,
rivers, and estuaries will be added to the list of
AOCs, in which case remedial action plans would
have to be developed there..  In addition, the
guidance document could potentially be used to
assess suspected sediment contamination outside
the Great Lakes basin.
 12.2   DESCRIPTION
 12.2.1 Description of Method
 12.2.1.1    Objectives and Assumptions

    In response  to  the  need  for  a  common
 approach to  the assessment of contaminated
 sediments, the IJC's Sediment Subcommittee has
 developed a  strategy based on protocols that
 emphasize biological monitoring. The approach
 is intended for use in comprehensive assess-
 ments of areas (e.g., bays, harbors, rivers, other
 depositional zones) where sediment contamina-
 tion  and the need for remedial action are sus-
 pected. While the suggested strategy attempts to
 minimize the cost and expertise, the assessments
 are relatively large undertakings appropriate to
 situations where large-scale remedial actions
 might be contemplated.  In such cases, the cost

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  Sediment Classification Methods Compendium
  of conducting accurate assessments would be
 justified  if  the subsequent remedial  options
  could cost far more than the assessments. It was
  not the primary intent of the subcommittee to
  provide guidance for small-scale decision-mak-
  ing activities, such as sample-by-sample disposal
  of dredged material from navigation channels.
  Nevertheless, some of the  component methods
  described could be useful and cost-effective in
 this regard.  The first major assumption, there-
 fore, is that the scope of the study in question is
 sufficient  to  warrant a  large-scale integrated
 investigation.
     Another fundamental assumption is that the
 ultimate concern of a problem assessment focus-
 es on whether sediment contaminants are exert-
 ing biological stress or are being bioaccumu-
 lated. Accepting this assumption, it follows that
 adequate assessments of sediment quality should
 involve components of chemistry, toxicity, and
 infaunal  community  structure  (Chapman and
 Long  1983), a concept frequently referred to as
 the Sediment Quality Triad approach (see Chap-
 ter 9). The proposed strategy has the following
 objectives:

     •  To provide accurate assessments of spe-
        cific  problems  by  using  a modified
        "triad" approach, which integrates chem-
        ical, physical,  and biological  informa-
        tion;

     •   To perform tasks in a sequence so that
        the results from each technique can be
        used   to   reduce  subsequent  samp-
        ling requirements and costs;

     •   To provide adequate  proof of linkage
        between the contamination  and the ob-
        served biological impact;

     •   To  quantify problem  severity, thereby
        enabling intercomparisons between and
        within areas of investigation (thus allow-
        ing development of a priority list for
        remedial actions and the objective selec-
        tion of appropriate remedial options);
     •   To consider the effects on different species
         and different trophic levels, since biological
         impairment may occur in the water column
         and the sediments if resuspension occurs
         and since  there is no such thing as  the
         universal "most-sensitive species" (Cairns,
         1986).

     The ITC approach is an integrated strategy that
 provides the necessary  data to identify sediment-
 associated contamination as  the problem source,
 specify effects, rank* problem severity, and assist in   •
 the selection of remedial options. While the assess-
 ment portion of the document identifies  a set of the
 best currently available assessment tools (see Section
 12.2.1.22), it is assumed that decisions will be made
 based on the circumstances unique to each AOC
 There is no substitute for experience (expert judg-
 ment),  and it  is  also  assumed that appropriate
 expertise will be assembled before the  assessment
 study plan is formulated.


 12.2.1.2    Level of Effort

 12.2.1.2.1  Type of Sampling Required

    The ETC (1988a) approach involves two stages.
 Stage I, the initial assessment, is used for areas
 where an inadequate or outdated database exists.  In
 the DC document, Stage I is not subdivided, while
 Stage n is broken  into  Phases  I, H, HE, and JV
 Stage I uses only in situ assessment techniques and
 criteria:  a limited physical description of the area
 (e.g., basin size and shape,  bathymetry) and the
 sediments, bulk chemical analyses, resident benthic
 community organization (e.g., family-level identifi-
 cations), fish contaminant body burdens (one impor-
 tant species, selected by expert judgment), and
 external abnormalities on collected specimens. Any
 one  of  the following criteria provides sufficient
justification for proceeding to Stage II:

    •   Concentrations of metals above background
        levels in sediments;

    •   Concentrations  of hazardous  persistent
       organic compounds above best available
      , detection levels in sediments;
22-2

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                                                                              22—J/G Approach
   m   Concentrations  of  hazardous  persistent
       organic compounds above detection levels
       in fish or benthos;

   •   The absence of a healthy benthic commu-
       nity (e.g., absence of clean water organisms
       such as amphipods or mayflies, presence of
       'a community dominated by oligochaetes,
       the complete absence of invertebrates); and


    •  Presence of external abnormalities in fish.

    These conditions must be supported by evidence
that the observed situation is not due to a major
sediment perturbation, such as dredging or substrate
modification.                   :
    Available data  may preclude the need for a
Stage I assessment.  The cost arid effort that Stage I
entails should be avoided if there is already strong
evidence of a contamination problem.
    When   a probable  sediment  contamination
problem is identified,  either, through the initial
assessmentor from the examination of existing data,
then  Stage  II, the detailed assessment, should be
undertaken.  The detailed assessment consists of
four  phases, which together define the sediment
problem in the  most cost-effective manner.  The
phases are not inflexible protocols, butrather logical
groupings  of work units. The expert investigator
should be responsible for the final study design.
     In Phase I of Stage II, extensive information on
 the physical composition of the sediments is collect-
 ed. These data are used to define areas or zones of
 homogeneity within a study area.  Knowledge of
 these zones allows  sampling requirements  for
 Phase n to, be estimated.
     In Phase n of Stage n, the benthic community
 structure is examined to the lowest possible taxo-
 nomic level (e.g., species or variety), along with the
 surficial sediment chemistry (e.g., pH, total organic
 carbon, redox potential, metals, extractable organic
 compounds). Phase II results can be combined with
 Phase I date to reduce the sampling  effort  in the
 next phase.
      In Phase IE of Stage n, a battery  of laboratory
 bioassays  (e.g., Microtox,  algal,  daphnid, benthic
  invertebrate, fish, Ames test) are performed on a
smaller number of sediment samples than those in
the Phase H sample set  Since fresh sediment must
be collected for this phase, precision position-finding
equipment is required to'relocate previously sampled
sites. Phase m costs can be reduced by performing
acute lethality bioassays  on a  sediment sample
before proceeding to tests that measure chronic or
sublethal effects.  Also m Phase ffl, sediment cores
are  collected, dated,  and  sectioned  for  stratified
chemical analyses and bioassays. Finally, adult fish
are examined histopathplogically for internal (e.g.,
liver) tumors.  In relatively confined geographical
areas, Phases H and HI may be  combined because
further sampling may be more costly than conduct-
ing additional bioassays and relocating Phase H sets
 for  Phase  HI sampling may be difficult  in this
 case, Phase n sampling will include extra material
 for Phase BL  -
     In the fourth and final phase of Stage H sedi-
 ment dynamics (e.g., accumulation, resuspension,
 movement) and factors affecting them are quanti-
 fied. All of the foregoing information is necessary
 for the  selection of  appropriate remedial options.
 For example, depositional history, as revealed by
 sampling sediment cores, and sediment dynamics are
 critical  pieces of information hi the selection and
 cost evaluation of remedial options.
     Criteria that dearly indicate when some form of
 remedial action must be considered (based on the
 results  of Stage n) are essential.  Because of the
 absence of definitive sediment action criteria at time
 of writing, the criteria proposed by the IJC (1988a)
 are highly conservative,  following the language of
  the 1978  Great Dikes Water Quality Agreement as
  revised in 1987 (especially Annexes 1 and 12), in
  order to promote maximum protection and effective
  restoration of the Great Lakes ecosystem.  The UC
  (1988a) urges  that these criteria be reviewed regu-
  larly to ensure that -they continue to fulfill their
  intended purpose.

   12.2.1.2.2   Methods

      During Stage I, the  minimum amount of infor-
   mation necessary to assess potential problem sedi-
   ments is collected. A variety of physical* chemical,
   and biological measurements are recommended, as
   outlined below:
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  Sediment Classification Methods Compendium
      «  A geographical description of the area and
         its bathymetry is required.

      •  Sediment grain  size - Size analysis tech-
         niques based on settling velocity (Ameri-
         can  Society for Testing and  Materials,
         1964; Duncan and LaHaie,-1979) are re-
         commended.  The  sand fraction  is re-
         moved  by a 62-/on  sieve and analyzed
         separately from  the fine-grained material.

      •  Sediment water content-The water content
         can be determined during sample prepara-
         tion  for grain size and  other analyses by
         comparison of sample weights before and
         after either  freeze-drying  or oven-drying
         (Adams et al., 1980).

      •  Redox potential  (Eh) and  pH  should be
         measured [specific methods are not rec-
         ommended by IJC (1988a)].

     •  Organic carbon - It is recommended that
         total  sediment organic carbon be measured
         as described by Plumb (1981).

     •  Phosphorus - Two measurements are sug-
         gested: total phosphorus, as extracted from
         sediment by sodium carbonate fusion or by
         perchloric acid digestion, and bioavailable
         phosphorus, as estimated by NaOH extract-
         able phosphorus (Williams et al., 1980).

     •   Ten  metals  (lead, nickel,  copper,  zinc,
         cadmium,  chromium, iron, manganese,
         mercury, and arsenic) are recommended for
         routine analysis  at Great  Lakes  ADCs.
         Additional metal  analyses are left to the
        judgment of the investigator. An extraction
         procedure using a mix of hydrochloric and
        nitric acids (1:1) is  suggested  (Plumb,
         1981).

     •  Persistent organic compounds - The reader
        is referred to the  U.S. EPA (1984) proto-
        cols for broad scans and  analyses of indi-
        vidual compounds. When the strategy was
        written, no standardized chemical protocols
  for estimating bioavailability of trace organ-
  ic compounds were identified.

  External abnormalities in fish - The pres-
  ence of one or more external abnormalities
  is  often indicative  of anthropogenically
  induced stress or damage.  In the case of
  the brown bullhead, Ictalurus nebulosus,
  phenomena such as stubbed barbels, skin
  discoloration (melanoma), and skin tumors
  are  highly  correlated  with  liver  cancer
  incidence (Smith et al., 1988). It is recom-
  mended that locally occurring catfish (par-
  ticularly  /.  nebulosus) be  examined  for
  tumors, melanoma, blindness, and barbel
  abnormalities during a Stage I assessment.

  Contaminant body burdens  -  The benthic
 infauna are in continuous contact with the
 sediments, providing a direct measure of
 the specific relationship between localized
 sediment contaminant concentrations and
 bioavailability.  Carp are also regularly in
 contact with and ingest large quantities of
 sediments.  They represent a larger spatial
 and temporal integration of contaminants
 than do the benthic infauna. Collection of
 adult common carp (Cyprinus carpio)  for
 tissue residue analysis is  recommended.
 Three to five fish per replicate should be
 composited.  The number of  replicates is
 determined using variability estimates from
 monitoring programs (Schmitt et al, 1983)
 and a chosen level of precision, to calculate
 an idealized sample size (p. 247, Sokal and
 Rohlf, 1969). It is also recommended that
 the .most abundant benthic  invertebrate
 species (often oligochaete worms in con-
 taminated sediments) be sampled in early
 summer, prior to  thermal  stratification.
 Standard U.S. EPA methods  are suggested
 for tissue residue analysis. The problem of
 obtaining enough biomass for  analysis  (at
 least 1 g) is recognized.

Benthic  community structure  -   In   a
Stage I assessment, a preliminary analysis
of community structure  impairment  is
12-4

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                                                                             12—IJC Approach
       recommended.  A qualitative study with
       minimal replication and identification only
       to the family level is suggested.  Because
       it is important that rare taxa be sampled,
       simple techniques that  employ inexpen-
       sive equipment but take large samples are
       recommended.    This  approach should
       suffice  to identify, the existence  of  a
       stressed community for the  purposes of
       Stage I criteria  (see Section  12.2.1.2.1
       above).

    Phase II of the detailed assessment consists of
more focused analyses to supplement or complement
information obtained in Stage I.   Phase I  of the
detailed assessment focuses on physical mapping of
the environment.  The most important aspect of the
physical  assessment of  a suspected  contaminated
sediment deposit is  its three-dimensional mapping.
A rectangular grid pattern is recommended  for the
initial mapping operation. Concurrent with bottom
sampling at  grid intersections, echo-sounder and
side-scan sonar surveys should be  performed to
improve spatial resolution of sediment zones and
bottom features. Detailed surveys should  include
piston coring for stratigraphic resolution. The grid
 sampling results should be examined using cluster
 analysis (or similar techniques), which are  easy to
 interpret and functional with a small number of
 variables.  Basic information required in this phase
 includes geographic location, areal extent, thickness
 and total sediment  volume,  average  depths of
 overlying water, and the grain size properties of the
 deposit. Phase I results are used to select sampling
 sites for later phases.
      Phase n of the detailed assessment focuses on
 surficial sediment chemistry and benthic community
 structure. Based on the previous mapping of homo-
 geneous zones (Phase I), effort in Phase H can be
 expended in depositional areas and in those areas
 with finegrained  sediments.  Surficial chemistry
  sampling should be coincident with the sampling for
  detailed benthic community structure analysis. Total
  organic carbon, redox potential, pH, metals, and
  persistent organics should be measured. Investiga-
  tors are referred to Plumb (1981),  Williams et al.
  (1980), and U.S. EPA (1984)  for collection and
  analysis methods.
    Since the main objective of Stage H commu-
nity structure assessment is to examine subtle
distinctions  in  stress response, more  detailed
taxonomic data are  required in this phase than
were required in Stage I. In the study .design and
sample collection steps, investigators are urged to
follow the 10 principles of  sampling set forth by
Green (1979). Further guidance is given in Elliott
(1977) for critical factors such as  site selection,
sample numbers, sampling design, and data! analy-
ses.   To  help investigators  assess community
impact,  IJC,(1988a) provides a partial list  of
literature descriptions of normal nearshore com-
munities in habitats that most closely approximate
 Great Lakes AOCs.   A detailed  discussion of
 statistical methods is also included.
     Phase III of the detailed (Stage II) assessment
 consists of obtaining additional information con-
 cerning sediment toxicity (i.e., bioassays and fish
 histopathology) and stratigraphic characterization
 of sediment cores.   A suite of bioassays is pro-
 posed for toxicologicai evaluation of sediments:

     •  Microtox - an acute,  liquid-phase (elu-
         triate or pore-water) test with luminescent
         bacteria (Bulich, 1984);

      B  Algal  photosynthesis - an acute,  liquid-
         phase test using natural communities
         [algal fractionation  bioassay (Munawar
         and Munawar, 1987)] or the laboratory
         species Selenastrum capricornutum (Ross
          etal., 1988);

      «   Zooplankton life-cycle  tests  (Daphnia
          magna liquid and solid phases) monitor-
          ing growth and reproduction (Nebeker et
          al., 1984; LeBlanc and Surprenant, 1985);

       •  Chronic, solid-phase tests using the ben-
          thic  invertebrates  Chironomus  teritans
          (Nebeker  et al.,  1984),  Hyalella azteca
          (Nebeker   et al., 1984), or Hexagenia
          limbata (Malueg etal, 1983);

       •  A solid-phase fish bioaccumulation test.
          with the  fathead  minnow Pimephales
          promelas (Mac et al., 1984)
                                                                                                12-5

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   Sediment Classification Methods Compendium
      •  The liquid-phase (extract) Ames Salmo-
          nella/microsome assay, a bacterial muta-
          genicity test (Tennant et al, 1987).

  In addition to bioassays, histopathological exami-
  nations of indigenous adult fish (especially Icta-
  lurus nebufasus), focusing on preneoplastic  and
  neoplastic liver lesions (Couch  and Harshbarger,
  1985), are recommended.
      Also included in Phase m (Stage n) work are
  chemical analyses and dating of sediment cores.
  Isotopic (14C, a°Pb, S5Fe, 137Cs) and biostratigraphic
  [i.e., ragweed (Ambrosia) pollen] methods are both
  recommended for dating sediment cores.  This
  dating is necessary to establish  the three-dimen-
  sional configuration of the contaminated sediment
  mass and to assign a date to Hie sediment deposi-
  tional unit.
     In Phase IV (Stage II) of the detailed assess-
  ment, studies on sediment dynamics are necessary
  to determine the following:

     •   Potential water column  impacts through
         resuspension;

     •   Movement of contaminated sediment out
         of the AOQ

     »   The quality  and rate of new  sediment
         accumulation; and

     •  Vertical  and  horizontal redistribution  of
        sediments and their contaminant burdens
        within an AOC.

 This information is essential for the development
 and evaluation of a  remediation plan.   In the
 absence of practical predictive models, suspended
 sediment characterization (Poulton, 1987), shear
 strength measurements (Terzaghi and Peck, 1967),
 and resuspension studies (Tsai and Lick, 1986) are
 recommended.

 12.2.1.2.3 Types of Data Required

    The Stage I initial assessment should be
based on aberrant macrozoobenthic community
  structure  (ascertained from  family-level taxo-
  nomic  identification);  metals  concentrations
  above background levels  in  the surficial sedi-
  ments (ascertained from dating); hazardous per-
  sistent organic compound concentrations above
  detection  levels in carp,  benthos,  or surficial
  sediments; metals  concentrations  in  carp or
  benthos, established on a case-by-case basis; and
  presence  in  fishes of  external  abnormalities
  known to  have contaminant-related  etiologies.
    . The Stage II detailed assessment should be
  based on  a phased sampling of the physical,
  chemical,  and biological  aspects of  the  sedi-
  ments.  The  biological  impacts should be as-
  sessed with  both field  (benthic  invertebrate
  community structure and incidence of fish liver
  tumors) and  laboratory  (battery  of  selected
  bioassays)  methods.   The  phased  sampling
  approach will allow subsequent testing require-
  ments to be reduced.  When Phases I and II of
  Stage II have, revealed homogeneous  zones of
  sediment type and similar community structure,
  the number of Phase III samples can be appro-
  priately scaled down.  Impairment due to sedi-
  ment contamination and the probable need for
 remediation are established when the biomoni-
 toring results  from  the detailed   assessment
 demonstrate significant departures from controls.
    Each section of IJC (1988a) contains a de-
 tailed  discussion  of the statistical  procedures
 required, with references and examples.   The
 preferred method of interpretation is left to the
 expert investigator in many cases.

 12.2.1.2.4  Necessary Hardware and  Skills

    The  initial assessment, and to an  even
 greater degree the detailed assessment, requires
 a large array of field and laboratory equipment.
 Although none of the items recommended are
 unusual or inordinately sophisticated, one labo-
 ratory or field unit is unlikely to have all the
 required apparatus.   Specific suggestions for
 hardware and skills are provided by IJC (1988a).
 Because this approach is intended  for major
sediment  assessment  efforts,  several  groups
would probably have to be mobilized to contrib-
ute to the effort.
22-5

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                                                                            12—1)C Approach
122.1.3    Adequacy of Documentation

    Each  component method  described  in  DC
(1988a) is fully referenced in the text and accom-
panied by a separate bibliography.  Some methods
are more developed than others, and areas where
additional validation or calibration is needed are
clearly-identified in the text.

12.2.2  Applicability of Method to Human
        Health, Aquatic Life, or Wildlife
        Protection

    The IJC strategy includes  direct measures of
effects on benthic infauna and fishes and is thus
directly  applicable  to  aquatic biota.   Existing
sediment  assessment  methods  (e.g.,  Apparent
Effects Threshold', Sediment Quality Triad) could
be used to evaluate the results of the Stage II
detailed assessment and  to  determine whether
chemically contaminated sediments have affected
aquatic biota in the  vicinity of AOCs.  Although
the DC (1988a) strategy  was not designed  to
assess the effects of toxic chemicals on wildlife or
humans, the tissue residue data and the sediment
 chemistry data may  be useful  in preliminary
 evaluations of contaminant  exposure to these
 populations.   Wildlife exposure could  occur
 through consumption of chemically contaminated
 prey.   Human  exposure could  occur through
 consumption of chemically contaminated fish or
 through dermal absorption by direct contact with
 chemically contaminated sediments or water.

 12.23  Ability of Method to Generate
         Numerical Criteria for Specific
         Chemicals

     The document was designed to provide guid-
 ance to assessment programs.  Nevertheless, since
 chemical,  tpxicological,  and infaunal data are
 collected in the Stage II assessment, it is possible
 that these data could be used to develop chemical-
 specific criteria. For example, data from the Stage
 II assessment  could be used to develop empirical
 sediment quality values (e.g., AET values) that are
 protective of aquatic biota in locations other than
 the AOC under consideration.
123   USEFULNESS

123.1  Environmental Applicability
123.1.1    Suitability for Different Sediment
           Types

    The approach recommended in DC (1988a) is
suitable for any sediment type. Indeed, one of its
major objectives is to characterize and provide a
three-dimensional map of the contaminated sedi-
ment mass,  including physical,  chemical,  and
biological variables! The investigator is given the
flexibility to  choose the appropriate sampling
methods for the sediment  type or  types  in the
AOC under study.

12.3.1.2    Suitability for Different Chemicals
            or Classes of Chemicals

    The document is intended for situations where
contamination is suspected, but where the toxic
chemicals may or may. not be identified.  The
methods recommended by DC (1988a) are effec-
tive for most contaminants found in Great Lakes
sediments.   The broad-based nature of the ap-
proach contains sufficient flexibility to deal with
 anomalous situations.

 12.3.1.3    Suitability for Predicting Effects on
            Different Organisms

     The proposed strategy includes both laboratory
 testing and  analysis of  indigenous communities
 (i.e., fish, macrozoobenthos). In this way, labora-
 tory results (i.e., chemistry, toxicity) that can be
 compared to standard  conditions  and literature
 values may be placed in the context of empirically
 derived effects data from the site under investiga-
 tion.

 12.3.1.4     Suitability for In-Place Pollutant
             Control   .        .

     The guidance document was developed specif-
 ically  for  the  assessment  of in-place pollutant
 problems.  It is designed to fit into the framework
 of evaluating and choosing remedial options by
 providing an adequate database on which to base
                                                                                            12-7

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  Sediment Classification Methods Compendium
  such decisions.   A companion document  (UC,
  1988b)  provides guidance in the  selection of
  courses  of remediation.

  12.3.1.5   Suitability for Source Control

     The detailed assessment provides an adequate
  framework for  identifying hot spots, and for
  establishing significant differences  from back-
  ground conditions.  In some cases, the resultant
  maps may provide further evidence of contaminant
  sources  and  migration patterns,  using  spatial
*  autocorrelation  techniques.    Presumably,  such
  evidence could facilitate regulation of identified
  sources. However, source control is not a primary
  objective of the UC (1988a) strategy.

  12.3.1.6    Suitability for Disposal Applications

     Although the document was not intended for
  the use in decision-making related to the disposal
  of material from navigational dredging, the  data
  generated from an initial assessment could be used
  to make initial disposal decisions. Other practices
  for the assessment of dredged material may be
 more cost-effective, however.


 123.2  General Advantages and Limitations
 12.3.2.1  Ease of Use

     The proposed strategy is designed to be applica-
ble to the AOC under investigation.  It is intended
to flexible, relying on the judgment and experience
of those who apply it. A detailed assessment would
be practical only in cases where a major remedial
effort is contemplated.

12.3.2.2    Relative Cost

    The  Stage I  and II assessments are costly
compared to other less comprehensive methods of
assessing  sediment quality.  However, when com7
pared to the potential remedial costs, the assessment
costs are relatively small.  The sequential approach
is designed to reduce sampling, analysis, and  ex-
pense where possible.  In many cases, the Stage I
assessment need not be done. If it is clear that a
  sediment contamination  problem exists, then the
  investigators may  proceed directly  to  Stage n
  assessment. Alternatively, if the Stage I assessment
  produces no results of concern, then Stage II need
  not be undertaken.  The cost of a detailed assess-
  ment, although relatively high, is controlled some-
  what by the sequential approach  to data collection.
  No firm cost figures are currently available, but as-
  sessments planned for priority AOCs under Section
  118(cX3) of the Water Quality  Act of 1987 are
  projected to cost in the range of $500,000.  These
  costs are expected to vary from site to site.

  12.3.2.3    Tendency to Be Conservative

     The strategy is designed to be highly protective
  of the environment.  It combines chemical analysis,
  toxicity  testing, and  examination of indigenous
  communities to ensure that no significant effects are
  overlooked.  Because the application of criteria is
  left to the expert judgment of the investigator, the
 degree of conservatism in decision-making will be
 variable.

 12.3.2.4    Level of Acceptance

    The guidance document (UC, 1988a) does not
 describe a new method, but rather a combination of
 several types of methods, each widely accepted in
 its own sphere.  The strategy as  a whole is being
 used for the first time in 1989.

 12.3.2.5   Ability to Be Implemented by
           Laboratories with Typical Equipment
            and Handling Facilities

    None of the methods is particularly unusual or
 difficult,  but the  detailed  assessment requires  a
 breadth of expertise and resources that an individual
 organization may not possess.  The strategy will
 need to be implemented by drawing on a variety of
 expertise in a given' geographical area.

 12.3.2.6    Level of Effort Required to Generate
           Results

    The total level of effort for a detailed assess-
ment will be relatively high 'in most cases.  This
12-8

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                                                                           12—l]C Approach
strategy is most  suitable  for  major evaluation
projects.

12.3.2.7                          ,
Degree to which Results Lend Themselves to
Interpretation

    The actual statistical analysis and interpreta-
tion to generate effects conclusions are relatively
complex  and should be done  only by trained
investigators. Specific statistical  protocols are
not recommended. However, the reader is given
an array  of choices,  with  comments on  their
respective strengths and weaknesses. The ulti-
mate decision is left  to the investigator.  The
inclusion of chemical, lexicological, and in-
faunal  information  in  the database allows the
investigator to compare different types of indica-
. tors before making decisions.

 12.3.2.8    Degree  of Environmental
            Applicability

     One  of the strengths of a strategy that in-
 cludes in situ community analysis is that effects
 data have a high degree of environmental rele-
 vance. Site-relevant species can even be substi-
 tuted in  the bioassay battery  if necessary, and
 the body burden and.community  structure data
 are always site-specific.

 12.3.2.9    Degree of Accuracy and Precision

     The  strategy proposed by the  IJC (1988a) is
 not a  single method, but rather guidance for a
 study  design  containing  many  options  and
 decision points.  Overall precision or accuracy
- values would be impossible to calculate. Never-
 theless,  the criteria for selecting recommended
 protocols included a consideration of attainable
 precision.  In many sections,  the investigator is
 directed to choose the required level of precision
 for a given measurement during the study design
 process. The "accuracy" of an integrated strat-
  egy is difficult to assess, but the methods recom-
  mended by the IJC  (1988a) wenb chosen  for
  their relevance to the Great Lakes ecosystem..
12.4   STATUS
12.4.1  Extent of Use

    HC's  (1988a)  document  was published in
December 1988  and distributed in  early 1989.
The strategy is intended for the Great Lakes, and
was used for the first time in  1989.  Most of the
individual methods recommended are widely used
and accepted.   -

12.4.2 Extent to Which the Approach Has
       Been Field -Validated

    The first extensive field validation of the ap-
proach was conducted in 1989-1991 as part of the
ARCS program under section  118(c)(3) of the
Water Quality Act of 1987. The ARCS Sediment
assessment reports are expected to be released in
 1993.                    ,

 12.4.3  Reasons for Limited Use
•                   '      /       '     , '
    Most component protocols are  in wide use.
 Because the IJC (1988a) document describes a
 major effort  with an integrated approach, the
 ARCS program is the  only project where an
 undertaking using this approach has been initiated.

 12.4.4 Outlook for Future Use and      _  .-
        Development

     With the backing of both  signatories to the
 Great Lakes Water Quality Agreement, the docu-
 ment seems destined for widespread use in the
 Great Lakes  basin.  As methods progress, each
 section of the document will be updated.
  12.5    REFERENCES

  Adams, D.D., D.A. Darby, and R.J. Young. 1980;
     Selected analytical techniques for characteriz-
     ing the metal chemistry and geology of fine-
     grained sediments and interstitial water.  In:
     Contaminants andi Sediments. R A. Baker (ed.)
     Ann Arbor Sci. Pub., Inc. Ann Arbor, MI.
  American Society  for Testing  and Materials.
     1964.  Procedures  for testing soils.   ASTM,
                                                                                            12-9

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  Sediment Classification Methods Compendium
    Philadelphia, PA. 535pp.
 Bulich,AA. 1984. Microtox - a bacterial toxici-
    ty test with general environmental applications.
    pp. 55-64. In: Toxicity Screening Procedures
    Using Bacterial Systems.  D. Lin  and B.S.
    Dutka (eds.). Marcel Dekker, New York, NY.
 Cairns, J., Jr. 1986. The myth of the most sensi-
    tive species.  BioScience 36:670-672.
 Chapman, P.M., and E.R. Long.  1983. The use
    of bioassays as part of a comprehensive ap-
    proach to marine pollution assessment.  Mar.
    Pollut Bull. 14:81-84.
 Couch, JA., and J.C. Harshbarger. 1985. Effects
    of carcinogenic agents on aquatic animals: an
    environmental and experimental  overview.
    Env. Carcinogenesis Rev. 3:63-105.
 Duncan,  G.A.,  and G.G. LaHaie.   1979.   Size
    analysis procedures used in the sedimentology
    laboratory, NWRI. Env. Can. NWRI contribu-
    tion.  23 pp.
 Elliott, J.M. 1977. Some methods for the statisti-
    cal analysis  of samples of benthic inverte-
    brates. Scientific Publication No. 25. Fresh-
    water Biological Association.  160 pp.
 Green, R.H. 1979. Sampling design and statisti-
    cal methods for environmental biologists. John
    Wiley and Sons, New York, NY.  257 pp.
 UC.   1988a.  Procedures for the assessment of
    contaminated sediment problems in the Great
    Lakes. International Joint Commission, Wind-
    sor, Ontario.,  Canada.  140 pp.
 JJC.  1988b. Options for the remediation of con-
    taminated  sediments  in the Great Lakes.
    International   Joint  Commission,  Windsor,
    Ontario, Canada.  78 pp.
 LeBlanc,  GA.,  and  D J. Surprenant.  1985.  A
    method for assessing the toxicity  of contami-
    nated freshwater sediments,  pp. 269-283. In:
    Aquatic Toxicology and Hazard Assessment,
    Seventh Symposium.  R.D. Cardwell, R. Pur-
    dy, and R.C. Banner (eds.), ASTM STP 854.
    American Society for Testing  and Materials,
    Philadelphia,  PA.
 Mac, M J., CC Edsall, R J. Hesselberg, and R.E..
    Sayers, Jr.  1984.  Flow-through bioassay for
   measuring bioaccumulation of toxic substances
   from sediment EPA DW-930095-01-0.  U.S.
   Environmental Protection Agency, Chicago, XL.
    26 PP-
 Malueg, K.W., G.S. Schuytema, J.H. Gakstatter,
    and D.F. Krawczyk.  1983.  Effect of Hexa-
    genia on Daphnia response in sediment toxici-
    ty tests. Env. Toxicol. Chem. 2:73-82.
 Munawar, M., and I.F. Munawar.  1987.  Phyto-
    plankton bioassays for evaluating toxicity of in
   • situ sediment  contaminants.  Hydrobiologia
    149:87-105.
 Nebeker, A.V., MA. Cairns, J.H. Gakstatter, K.W.
    Malueg, and G.S. Schuytema.  1984.  Biologi-
    cal methods for determining toxicity of con-
    taminated  freshwater  sediments  to  inverte-
    brates.  Env. Toxicol. Chem. 3:617-630.
 Plumb, R.H., Jr. 1981.  Procedures for handling
    and chemical analysis of sediment and water
    samples.    Technical  Report EPA/CE-81-1.
    U.S.  Environmental Protection Agency/U.S.
    Army Corps of Engineers Technical Committee
    on Criteria for Dredged and Fill Material, U.S.
    Army Waterways Experiment Station, Vicks-
    burg, MS.  471 pp.      .
 Poulton, D J.  1987. Trace contaminant status of
    Hamilton Harbor. J. Great Lakes Res. 13:193-
    201.
 Ross, PJE., V. Jany, and H. Sloterdijk.  1988. A
    rapid bioassay  using the  green  alga Sel-
    enastrum capricornutum to screen for toxicity
    in St. Lawrence River sediments.  American
    Society  for  Testing  and  Materials.    STP
    988:68-73.
 Schmitt, CJ., MA. Ribick, J.L. Ludke, and T.W.
    May.  1983.  National pesticide  monitoring
    program: organochlorine residues in freshwa-
    ter fish, 1976-79.  Fish and Wildlife Service
    Res. Publ. No. 152.  U.S. Dept. of Interior,
   Washington, DC.
 Smith, S.B., M J. Mac, A.E. MacCubbin, and J.C
   Harshbarger.  1988.  External abnormalities
   and incidence of tumors in fish collected from
   three Great Lakes Areas of  Concern.  Paper
   presented at the 31st  Conference on  Great
   Lakes Research, McMaster University, Hamil-
   ton, Ontario.  May 17-20, 1988.
Sokal, R.R., and F.J. Rohlf.  19(59.  Biometry.
   WiH. Freeman and Co., San Francisco, CA.
Tennant, R.W., BM. Margolin,  D.D.  Shelby, E.
   Zeiger, J.K. Haseman, J. Spalding, W. Caspary,
12-10

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                                                                            12—1JC Approach
   M. Resnick, S. Stasiewicz, B. Anderson, and
   R  Minor.   1987.   Prediction of  chemical
   carcinogenicity in rodents from in situ genetic
   toxicity assays. Science 236:933-941.
Terzaghi, K., and R.B. Peck. 1967. Soil mechan-
   ics in engineering practice.  John Wiley and
   Sons, New York. 729 pp.
Tsai,  C.H.,  and W. Lick.   1986.   A portable
   device for measuring sediment resuspension.
   J. Great Lakes Res. 12:314.-321.
USEPA.  1984.  Guidelines establishing test pro-
   cedures for the analysis of pollutants under the
   Clean Water Act; final rule and interim final
   rule and proposed rule.  U.S. Environmental
   Protection Agency. Washington, DC. Federal
   Register Vol. 49, No. 209, Part Vffl.  pp. 1-
   210.
Williams, J.D.H., H. Shear,  and R.L. Thomas,
   1980.  Availability* to Scenedesmus quadri-
   cauda of different  forms of phosphorus in
   sedimentary materials in the Great  Lakes.
   Limnol. Oceanogr. 25:1-11.
                                                                                            12-11

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          CHAPTER 13
Summary of Sediment-Testing  Approach

Used  for  Ocean Disposal
David P. Redford
U.S. Environmental Protection Agency
499 South Capitol Street, SW (WH-556F), Washington, DC 20003
(202)260-9179
    The Evaluation of Dredged Material Pro-
posed  for  Ocean  Disposal—Testing  Manual
(USEPA/USACE, 1991) commonly referred to
as the "Green Book," was published in February
1991  by  the U.S. Environmental Protection
Agency (USEPA) and the U.S. Army Corps of
Engineers (USAGE). The Green Book contains
national guidance  for evaluating the suitability
of  dredged material for  ocean  disposal; it re-
places the guidance of  the  original  manual
(USEPA/USACE,  1977) that was  published by
USEPA and the USAGE in 1977.   The manual
stresses the use of bioassay and bioaccumulation
testing as evaluative tools, and it contains tech-
nical guidance on  the use of such tests.   The
following is a summary of the 1991 manual and
the approach used  by USEPA and the USAGE
to  determine  the suitability of dredged material
 for ocean disposal.  The manual willbe revised
 at  a  future date,  based  on the findings of an
EPA Science Advisory  Board (SAB) review
 (SAB, 1992), and changes will be made to the
 Ocean Dumping Regulations (referenced below).
 13.1   APPLICATION

     The  1991 USEPA/USACE  Green  Book
 provides  updated guidance  for dredging  appli-
 cants,  scientists,  and  regulators  to  evaluate
 dredged-material compliance with the 1977 U.S.
 Ocean Dumping Regulations [Title 40, Code of
 Federal Regulations (CFR), Parts 220-228]. The
 manual is applicable to all activities involving
 the transportation  of dredged material for the
 purpose of dumping it in ocean waters outside
 the baseline from  which the territorial  sea is
 measured. The guidance in  this manual is appli-
cable  to dredging operations  conducted under :
permits as well as to federal projects conducted
by the USAGE.  The procedures in this manual
do  not  apply  to  activities  excluded  by
40 CFR 220.1.                      ' .,  :
    It is important to note  that  the regulations
are legally binding and that the guidance provid-
ed in this manual is responsive to the specific
requirements of these regulations, but the manu-
al does not cany the force of law.   The docu-
ment simply provides guidance on evaluating the
potential environmental impact of dredged-mate-
rial ocean disposal.
    The manual is organized into tiers for effi-
cient  evaluation  of the suitability  of dredged
material for ocean  disposal.  Within the  tiers,
specific physical, chemical, and biological tests
are recommended.   To meet specific regional
needs, USEPA  Region  and USAGE District
offices  are  to  develop local agreements  and
manuals to implement the national  guidance in
the 1991 Green Book (such as using local spe-
cies in biological tests and screening for particu-
lar contaminants in chemical analyses).

 13.1.1   Current Use

    The  1991 Green Book replaces the 1977
 Green Book. USEPA Region and USAGE Dis-
 trict offices are developing local agreements and
 regional testing .manuals  that implement the
 1991 Green Book guidance and establish permit
 procedures for dredging  and dredged-material
 disposal.
    Projects  that  have  been  issued   under
 USAGE permits prior to the completion of the
 new local  agreement/manual for the area cov-
 ered by the project may continue to be evaluated

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 Sediment Classification Methods Compendium
 according to the 1977 guidance manual and the
 existing  local guidance.  New dredged-material
 disposal  projects,  projects that  have  not had
 sampling and analysis plans  approved prior to
 finalization   of the  • local  agreement/manual,
 should be evaluated under the updated guidance
 in the 1991 Green Book.  Ongoing projects that
 have been approved based'on 1977 Green Book
 guidance should be  reevaluated according to
 1991 Green  Book guidance and the new local
 agreement/manual  within  3  years  of  permit
 approval.

 13.1.2 Potential Use

    The  Green Book guidance, and revisions
 thereof,  will be applied  to dredged-material
 evaluations for the foreseeable future.
    The  manual will be revised at a future date
 based on (1) the findings of an EPA SAB  re-
 view (SAB,  1992),  (2)  technical  advances in
 assessing sediment contamination and marine
 environmental  impact, and (3) changes  to the
 Ocean Dumping Regulations.
13.2   DESCRIPTION

    Analysis of sediment to determine its suit-
ability for ocean disposal is conducted accord-
ing to the procedures in the 1991 Green Book.
The 1991 Green Book recommends procedures
that satisfy  section 103 of the Marine Protec-
tion,  Research, and Sanctuaries Act of 1972
(MPRSA), Public  Law 92-532. The MPRSA
was enacted to regulate ocean dumping of all
materials  that  might adversely affect human
health, the marine environment, or other legiti-
mate uses  of  the oceans.   In addition,  the
MPRSA   implements the Convention on  the
Prevention of Marine Pollution by  Dumping of
Wastes and Other Matter (London Dumping
Convention), of which  the United States is  a
signatory.  MPRSA section 103 specifies that
all  proposed operations involving the trans-
portation  and  dumping of dredged material
into ocean  waters must be  evaluated to  de-
termine the  potential environmental impact of
 such activities.   These environmental evalua-
 tions must  be in agreement with  the criteria
 published   in  40 CFR  Parts  220-228   and
 33 CFR Parts 320-330 and 335-338.
    Technical guidance  on  specific methods
 for testing dredged material is presented in the
 1991 Green Book,  If the results of the appro-
 priate tests show that the proposed  dredged
 material meets the  chemical- and biological-
 effects  criteria, and meets other requirements
 in the regulations, disposal of the material  at a
 designated ocean  dredged-material disposal site
 (ODMDS)  is supported.   If the  test results
 show that the material does  not meet the  cri-
 teria set forth in the regulations,  significant
 impact  on the ocean environment  is predicted.
 Significant  adverse  impact  may  include  ad-
 verse consequences  to the marine  ecosystem
 and negative  human-health effects  from  uses
 of the marine environment.
    The manual does not present guidance for
 the disposal of dredged material  that fails to
 meet the  regulatory  criteria.   Such disposal
 involves management  decisions  and case-spe-
 cific engineering  work (e.g., control of dump
 releases,  disposal-site  capping,   submarine
 burial,  and  predisposal  treatment)  that  are
 beyond  the scope of the document.

 13.2.1  Description of Method

    Integral to the  1991  Green  Book is  a
 tiered-testing   procedure    to   characterize
 dredged material  and predict its  impact on the
water-column  and  benthic  environment  at
 ODMDSs.   The  procedure was  developed by
USEPA and USAGE  personnel and  testing-
laboratory researchers, and is consistent with
the requirements of the Ocean Dumping Regu-
lations,   state-of-the-art   dredged-material
evaluation techniques,  and  the realities of the
testing  and permitting process  for new  and
existing projects.   Knowledge of local condi-
tions is both  recommended and necessary to
adapt the  national guidance in the manual to
specific dredged-material  projects.   USEPA
Regions and USACE  Districts  are presently
developing local  agreements/manuals to apply
33-2

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                                                 13—"Green Boole" Sediment-Testing Approach
the national guidance of the manual to specific
dredging and disposal areas.
    The tiered-testing  procedure in the Green
Book comprises four tiers, with decision points
at each tier (Figure  13-1). Each successive tier
provides increasing investigative  intensity  to
generate the information  for permitting deci-
sions on ocean disposal.
    The tiered-testing  procedure is constructed
to  determine  whether  the  dredged  material
meets the  limiting permissible concentration
(LPC), as  defined  in section  227.27  of the
Ocean Dumping Regulations. The LPC for the
liquid-phase concentration of dredged  material
in  the water column is the  concentration that,
after allowance for  initial mixing,  does not ex-
ceed applicable marine  water-quality  criteria
(WQC) or a toxicity  threshold of 0.01 of the
acutely toxic concentration.  The LPC of the
suspended particulate  and solid phases is the
concentration that will not cause unreasonable
 toxicity or bioaccumulation.
     The overall tiered-testing procedure is rela-
 tively flexible. The dredged-material evaluator
 can enter  and exit the testing procedures at any
 tier. However, to begin the evaluation in Tier
 II, III, or  IV, the existing data must satisfy the
 requirements of the earlier tier(s).  Additional-
 ly, Tier  II testing for  water-quality criteria
 (WQC) compliance is mandatory  if the water-
 column evaluation  cannot be completed within
 Tier I.   To  exit  any tier before reaching a
 decision  on  LPC compliance,  the  dredged-
 material evaluator  must select an option other
 than open-ocean disposal.
     In most cases,  determinations of LPC com-
 pliance can be made in Tier  I, II, or III:  In
 extraordinary  cases,  where  LPC compliance
 cannot be determined by Tier III, the dredged
 material  must be evaluated  under  Tier  IV.
 Tier IV tests  are case-specific investigations of
  potential  impact of the  dredged material at the
  ODMDS.  Significant  investment in the re-
  search and development of analytical methods
  is usually necessary to  conduct Tier  IV evalu-
  ations, and the applicant might select an alter-
  native to open-ocean  disposal  instead  of
  proceeding with Tier IV testing.  Similarly, an
applicant can  try to save time and money by
proceeding directly to Tier II, HI, or IV if it is
believed that  analysis in the earlier tiers will
not lead to a  definitive evaluation.  The  only
absolute requirement is that the dredged mate-
rial must comply with the regulations if it is to
be dumped at an ODMDS.  The tiered-testing
procedure facilitates this determination.

In summary, the 1991 Green Book

    •  Includes state-of-the-art  methods  to
       - determine the potential impact  of ma-
        rine-sediment disposal;

    •  Ensures adherence to the Ocean Dump-
        ing Regulations  (40 CFR  Parts  220-
        228);

    •  Incorporates existing  (and  valuable)
        regional expertise  and  guidance into
        the evaluation process; and

     •  Provides  for  National  consistency  in
        evaluating dredged  material for  ocean
        disposal.

 13.2.1.1    Objectives and Assumptions

     The objective  of  the tiered-testing proce-
 dure is to determine whether the water-column
 and  benthic LPC is  met  for  the proposed
 dredged material, as  defined  in  the Ocean
 Dumping Regulations. Three decision options
 are  possible as the dredged-material evaluator
 proceeds through the tiers.

      (1) The  LPC  is met;  the  ocean  disposal
         option is  supported; further evaluation
         is unnecessary.

      (2) The  LPC evaluation  is inconclusive;
         the ocean disposal  option  is  not sup-
        . ported; proceed to the next tier.

      (3) The LPC is not met; the ocean disposal
         option is  not supported; further evalua-
         tion is unnecessary.
                                                                                            13-3

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 Sediment Classification Methods Compendium
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                                                    13—"Green Book" Sediment-Testing Approach
     Both  the  water-column  and  benthic LPG
 considerations must be satisfactorily resolved for
 the open-ocean disposal option to be supported.
 An inconclusive evaluation in Tiers I-III,requires
 the  dredging  applicant  to conduct additional  ,
 testing in subsequent tiers, or to  decide not to
- ocean-dump.  However, a determination of LPC
 noncompliance does not necessarily exclude all
 possibilities for ocean disposal.   Management
 actions might be feasible  to make the dredged
 material meet the LPC. Management actions for
 dredged material that exceeds water-column or
 1 benthic LPC are not included in the Green Book
 because of  the  wide range of available options
 and the project-specific nature of such work.
     It is  assumed  that the users of the 1991
 Green Book are generally familiar with the need
 for  and  methods  of dredged-material - testing.
 The manual is not  a standalone document. The
 guidance  in the manual requires the evaluator to
 consult the  regulations frequently  (40 CFR Parts
 220-228 is  included  in the Green Book as Ap-
 pendix A) and  to have a  general understanding
 of material contained in the numerous  citations
 and references.  The guidance in the manual
 concentrates  on data  collection and  decision
 points, and it  only summarizes recommended ,
 field and laboratory procedures that can be used
 .to obtain data.  The user must refer to the origi-
 nal sources for most of the physical,  chemical,
 and biological testing procedures.

 13.2.1.2    Level of Effort

 Tier I:  Initial Assessments—Tier I is used to
  identify contaminants of concern and determine
  dredged-material LPC compliance through anal-
  ysis of existing physical,  chemical, and biologi-
  cal information.   For many  dredging  projects,
  there is a wealth of readily available information
  on the proposed dredged material  and on  the
  characteristics of the disposal site. This is espe-
  cially true  of areas that have historically under-
  gone maintenance  dredging  or  have been  the
  subject of  other studies, such as fishery  assess-
  ments.   The available information  for a given
  area might not be sufficient to reach a final LPC
  evaluation, but often there are accessible high-
quality data that can supplement  the results of
tests in subsequent tiers and  facilitate  reaching
an early decision with lowered expenditure of
time and resources.
    Whatever  the  source  of information  for
Tier I evaluations, the quality of  the data must
be  evaluated  and weighed  accordingly.   The
references in Chapter 13 of the manual, Quality-
Assurance  Considerations, should be  consulted
for guidance for evaluating the quality of data
obtained from different information sources.
    If the information set  compiled  in  Tier I is
complete, and  comparable to information that
would appropriately satisfy the LPC in Tier n,
in, or IV, a decision on regulatory compliance
be  completed without proceeding into  the next'
tiers. For compliance determination to be com-
pleted within Tier I, the weight of evidence of
the  collected  information  must convincingly
show that the dredged-material disposal  either
will or will not meet the LPC.
    Included in Tier I is  an assessment of the
three exclusionary criteria in  40 CFR 227.13(b):
(1) the  dredged material is predominantly sand,
gravel,  or rock from a high-energy area; (2) the
material is suitable  for beach nourishment; or
(3) the material  is  similar to the disposal site
and from  an area far removed  from  pollution
sources. If one or more of the above exclusion-
ary  criteria can be satisfied, the LPC is met for
the dredged material and no further evaluation is
required.  If none of the exclusionary  criteria is
met and the collected information is insufficient
to  reach a  definitive LPC  determination, the
evaluation process moves to Tier II.

Tier II:  Physical/Chemical Evaluations—Tier
 II consists of physical and chemical data evalua-
tion.  To determine marine WQC compliance, a
 numerical mixing model is used; to evaluate
 benthic-impact  potential  for nonpolar  organic
' compounds, a theoretical bioaccumulation poten-
 tial (TBP) calculation is  used.  The conceptual
 purpose of the tier is to  provide reliable, rapid
 screening  of impact potential without the need
 for further testing.  This  purpose is fulfilled for
 water-column evaluations, but at  present there is
 no  USEPA-approved single screening  procedure
                                                                                              13-5

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 Sediment Classification Methods Compendium
 for deposited sediment.  When technically sound
 sediment-quality criteria (SQC) are developed
 and  approved  for dredged-material  evaluation,
 they will be incorporated at this level.

 Her II:   Water-Column Physical/Chemical
 Evaluations—-The Tier II water-column  eval-
 uation for WQC compliance is a two-step pro-
 cess that includes the application of a numerical
 mixing model.  In Step 1, the model is used as a
 screen; all  of the  contaminants in the dredged
 material  are assumed  to  be released  into the
 water column during the disposal process. If the
 model predicts that the concentration of contam-
 inants of concern released into the water column
 is less than the applicable WQC and if no syner-
 gistic  effects   among  the contaminants  are
 suspected, the dredged material meets the water-
 column LPC and no further water-column evalu-
 ations are necessary.
    If LPC compliance  cannot be  shown in
 Step 1, Step 2 is conducted. In Step 2, chemical
 data from an elutriate test of the dredged materi-
 al  are run in the model.  Compared to the as-
 sumption of total contaminant release in the Step
 1 screen, the elutriate data applied in Step 2 are
 a more precise representation of the concentra-
 tion  of contaminants   that would actually be
 released into the water column during ocean dis-
 posal of dredged material.
    If the model  predicts in Step  2 that  any
 WQC are exceeded, the water-column LPC is
 not met (open-ocean disposal not supported). If
 there are WQC for all of the contaminants of
 concern, if no WQC are exceeded by the Step 2
 model, and  if no contaminant synergistic effects
 are suspected, the water-column LPC is met and
 no  further water-column evaluations are neces-
 sary (open-ocean disposal  supported).   If there
 are contaminants of concern without WQC  or if
 synergistic effects are  suspected, water-column
 toxicity and water-column LPC compliance must
 be evaluated in Tier ffl.                •

 Numerical Models for  Initial Mixing—Numer-
 ical models are used to evaluate dredged-mate-
 rial dilution  during the initial-mixing phase of
 ocean disposal,  as  defined in  the regulations.
  The  1991 Green  Book  recommends using the
  USAGE  Automated  Dredging  and  Disposal
  Alternatives  Management  System  (ADDAMS)
  models to evaluate initial  mixing  of  dredged
  material at ODMDSs.  ADDAMS models can be
  run on a  personal computer with a  minimum of
  hardware. The models account for the physical
  processes of dredged-material disposal at open-
  water disposal sites by  calculating the water-
  column concentrations  of dissolved contaminants
  and suspended sediments and the initial deposi-
  tion of material 'on the bottom. Three separate
  ADDAMS models address different methods of
  disposal:

     •  DIFJD    Disposal  from  an  instanta-
                   neous dump

     •  DIFCD    Disposal from a continuous
                   discharge

     •  DIFHD    Disposal   from   . a  hopper
                   dredge

     To evaluate initial mixing following .ocean
  disposal,  the appropriate model is  run  for the
  contaminant  requiring the greatest amount of
  dilution to meet the LPC. The models simulate
  movement of the  disposed  material as  it  falls
  through the water column, as  it is transported
  and diffused  by the ambient current, and as it
  spreads over  the  bottom.   The models have
  some  limitations; for example, the DIF1D model
 will not work for very  shallow disposal sites
 where the discharge time  from the barge exceeds
 the descent period to the  bottom. However, the
 models can simulate a wide range of disposal
 options.   USEPA  and the USAGE are in  the
< process of field-verifying  these models.
     Appendix 6 of the 1991 Green Book  is a
 summary of the ADDAMS models;  the comput-
 er diskettes that accompany the  manual  contain
 the models themselves.   ADDAMS modeling
 personnel  at the USAGE Waterways Experiment
 Station  (WES),  Vicksburg,   Mississippi,   are
 available to supply model updates, answer ques-
 tions,  and assist with the selection and running
 of the individual models.
13-6

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                                               23—"Green Book" Sediment-Testing Approach
         Table 13-1. 1991 Green Book Species for Water-Column and Benthic Evaluations
          Water-Column Species
      Crustaceans
        Mysids
            Mysidopsis sp.a
            Neomysis sp."
            Holmesimysis sp.'
        Shrimp
            Palaemonetes sp.
            Penaeus sp.
            Pandalussp.
        Crab
            Callinectes sapidus
            Cancer sp.

       Fish
            Menidia sp.a
             Cymatogaster aggregate?
          .   Cyprinodon variegatus
             Lagodon rhomboides
             Leiostomus xanthurus
             Citharicthys stigmaeus
             Leuresthes tenuis
             Coryphaena hippurus

       Zooplankton
         Copepods
             Acartia sp.a
         Mussel larvae
          .   Mytilus edulis*
         Oyster larvae
             Crassostrea virginicam
             Ostrea sp.8
         Sea-urchin larvae
             Strongylocentrotus
                  purpuratus
             Lytechinus pictus
Crustaceans
  Infaunal Annphipods
      Rhepoxyniussp.*
      Ampelisca sp.m
      Eohaustorius sp.'
      Grandiderella japonica
      Corophium insidiosum
  Mysids
      Mysidopsis sp.
      Neomysis sp.
      Holmesimysis sp.
  Shrimp
      Penaeus sp.
      Palaemonetes sp.
      Crangonsp.
      Pandalus sp.
      Sicyonia ingentis
  Crab'
       Callinectes sapidus
       Cancer sp.
 Fish
       Cleveland/a Jos
       Atherinops affinis
 Burrowing Polychaetes
       Neanthessp.m
       Nereis sp.*
       Nephthys sp.
       Glycera sp.
       Arenicola sp.
       Abarenicola sp.

 Molluscs
        Yoldia limatula
        Macoma sp.
        Nuculasp.
        Protothaca staminea
        Tapes japonica
        Mercenaria mercenaria
•Recommended test species.
                                                                                        13-7

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  Sediment Classification Methods Compendium
      The model output can present water-column
  contaminant  concentrations  in  milligrams  per
  liter.  These  concentrations are compared to the
  appropriate LPCs to determine compliance.

  Tier  II:  Benfhic  Physical/Chemical  Evalua-
  tions—As previously noted,  only benthic  effects
  attributed  to  nonpolar organic chemicals  in  the
  deposited sediment can be addressed in Tier II at
  the present time.  Nonpolar organic chemicals in-
  clude all organic compounds that do not dissociate
  or form ions.  These include chlorinated hydrocar-
  bon pesticides, other halogenated  hydrocarbons,
  polychlorinated  biphenyls (PCBs),  most polynu-
  clear aromatic hydrocarbons (PAHs), dioxins, and
  furans.   It does not include  polar organic com-
  pounds, organometals, and metals.  If all  of  the
  contaminants  of concern  in the dredged material
  are nonpolar  organic compounds, the  theoretical
 bioaccumulation potential (TBP) can be calculated
  for the dredged material  and the reference sedi-
 ment1 to determine benthic UPC compliance. The
 TBP calculation  is an environmentally conserva-
 tive screen, based on calculating the concentration
 of the nonpolar  organic chemical in the sediment,
 the total  organic-carbon  concentration, and the
 percent lipid content of an organism of interest. If
 the TBP of the dredged material is not statistically
 greater than that of the  reference  material, the
 LPC for the nonpolar organic contaminants is met.
 (Acute-toxicity  evaluations  must  be  performed
 under Tier HI unless sufficient toxicity information
 was obtained under Tier I.)
    If any of the  contaminants of concern are
 polar organic compounds  or have  suspected toxic
 components or if the dredged-material TBP ex-
 ceeds the reference-material TBP described above,
 the bioaccumulation evaluation for benthic impact
 by the dredged material must take place in Tier IE
 or TV.  The benefit of additional tests in Tier II to
 screen for benthic impact is recognized by USEPA
*A reference sediment is defined as a sediment, substantial-
ly free of contaminants, that is as similar as practicable to
the grain size of tha dredged material and the sediment at
the disposal site, and that reflects the conditions that would
eicist in the vicinity  of the disposal site had np dredged-
material disposal ever taken place, but had all other influ-
ences on sediment condition taken place.
  and the USAGE, and new, tests are under develop-
  ment and evaluation.   When the scientific and
  regulatory community  verifies  one  or more of
  these tests, they will be incorporated into-Tier n in
  a future Green Book revision.  Meanwhile, evalu-
  ation of benthic  impact that cannot  be made in
  Tier I must be completed in Tier III or IV.  .

  Tier III:   Biological Evaluations—Tier IE tests
  include (1) determination of water-column toxicity
  and (2) assessment of contaminant  toxicity and
  bioaccumulation. from the material to be dredged.
  The evaluations in this tier are based on the output
  from Tiers  I and  n and comprise  standardized
  bioassays with the organisms listed hi  Table 13-1.

 Her III:  Water-Column  Biological  Evalua-
  tions—Tier III water-column tests are acute tests
 that evaluate the toxicity of the  dissolved and
 suspended portions of the dredged material that
 remains in the water column after initial mixing.
 The bioassays are run if the Tier II evaluations
 are inconclusive,  e.g., if there are not applicable
 WQC for  all contaminants ,of concern or there is
 reason to  suspect synergistic effects among the
 contaminants. (See Tier II.)   The tests involve
 exposing fish, crustaceans, and zooplankton to a
 dilution series containing both  dissolved- and
 suspended-sediment components  of the dredged
 material.  A typical test monitors  organism mor-
 tality over a 96-h period.
     The results of the bioassays are used to calcu-
 late the LCjo concentration of the dredged material
 in the water column. The LPC for this evaluation
 is 1 percent  of the LQo outside the ODMDS
 during the  initial 4-h mixing period and anywhere
 in  the marine. environment  4 h  after disposal.
 Following  the determination of the LPC for the
 proposed dredged  material, the data are used to
 run  the numerical model (see  model discussion
 above) and determine LPC compliance.

 Her IH: Benthic Biological Evaluations—Ben-
 thic  evaluations in Tier in consist of toxicity and
bioaccumulation tests.  To conduct  these tests, the
 1991 Green  Book provides  laboratory guidance
on  sediment  preparation; treatment,  reference-,
and  control-sediment  tests; -replicates; organism
13-8

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                                                    13—"Green Book" Sediment-Testing Approach
handling; test-chamber conditions; QA considera-
tions; and data analysis.  The organisms used in
the tests are surrogates for  disposal-site species
and are used to estimate dredged-material effects.
The toxicity tests quantify mortality.  If the mor-
tality of the test species in  the dredged-material
bioassays is greater than the  allowable percentage
over the mortality in the reference-sediment bio-
assays, the  LPC is  not^met.   If, however, the
dredged-material tests tielow toe allowable per-
centage, or  the increased mortality is statistically
insignificant, the LPC is met.
    The bipaccumulatipn tests evaluate the poten-
tial of benthic organisms to  accumulate contami-
nants  from  the dredged material  in  their tissues.
At the conclusion of the tests, the tissues of the
organisms are analyzed for the contaminants of
concern that are identified in Tier I.
    Section 227.27  of  the Ocean Dumping Reg-
ulations requires that benthic bioassays be con-
ducted on  dredged  material with  filter-feeding,
deposit-feeding, and burrowing species.  Infaunal
arhphipods, such as Ampelisca sp. and Rhepoxy-
 nius sp., are sensitive bioindicators  and strongly
 recommended in the Green Book as the preferred
 species for toxicity tests.   Infaunal. amphipods
 filter-feed,  deposit-feed, and, to some extent, bur-
 row in the sediment, thereby fulfilling the three
 organism categories in the regulations. For bioac-
 cumulation evaluations, the manual recommends
 using a burrowing polychaete (e.g;, Neanihes sp.
 or Nereis sp.) and a deposit-feeding bivalve mol-
 lusc (e.g.,  Macoma sp. or  Yoldia  limatula).  In
 summary,  the manual  recommends that at least
 two species  be tested for  acute toxicity  and at
 least two other species  for bioaccumulation evalu-
 ation. Each set of test species should cover the
 three species types stipulated in the  regulations.
 The  ecological and economic  relevance  of the
 organisms  arid the practical aspects of using the
 species  in  the laboratory,  such as tolerance to
 grain-size  ranges  and  seasonal, availability,  also
 must  be  considered  when selecting the  test
 species.
      The Tier HI bioaccumulation evaluation com-
 pares the  contaminant level in the  tissues  of the
  organisms to two  criteria:  (1)  the  United States
  Food and Drug  Administration  (FDA)  Action
Levels for Poisonous or Deleterious Substances in
Fish and Shellfish for Human Consumption and
(2) the contaminant levels in organisms that are
exposed to the reference sediment.  Regardless of
the statistical comparison to the reference-material
test  organisms,  if the  level in  the tissues  of
dredged-material organisms  statistically exceeds
the FDA levels in any" category, the LPC is not
met.   If the  dredged-material results  are  lower
than the FDA action  levels and not statistically
greater than the reference material level, the LPC.
for bioaccumulation is satisfied. However, if bio-
accumulation  exceeds that found in the reference-
material tests, the test results must be evaluated
against case-specific criteria.   USEPA and the
USAGE develop  the  evaluative criteria case by
case from  local  technical information that ad-
dresses the bioaccumulation aspects of the benthic
criteria of section 227.13(cX3) of the regulations.
     At present, tests for chronic  sublethal expo-
sure to benthic  contaminants are being developed.
When the tests are approved  by USEPA,  they
will be incorporated in Tier ffl in future updates
 to the Green Book.

 Tier  IV:   Advanced Biological  Evaluations—
 Tier IV consists of bioassay  and bioaccumulation
 tests to evaluate the long-term benthic and water-
 column impact of dredged material.  Tests at this
 level are selected to address specific issues for a
 specific dredging operation that could not be fully
 evaluated in the earlier tiers.  Since these tests are
 case-specific and since they require significant
 tune  and money to complete, evaluative  criteria
 must be agreed on in advance by USEPA and by
 the  USAGE to  determine compliance with the
 LPC.
     Conducting Tier IV benthic testing is possible
 with current methods, but the 1991 Green Book
 emphasizes that this  tier is not intended for rou-
 tine  application.   Tier IV benthic tests consume
 significant resources  of the dredging  applicant
 and  of the regulatory authority, and a final non-
 compliancedetermination is still possible.  There-
 fore, the applicant must weigh the options and
 decide whether to perform Tier IV testing or to
  consider  an alternative that  does not  involve
  ocean dumping, such as upland disposal.  If the
                                                                                                13-9

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  Sediment Classification Methods Compendium
  applicant elects to proceed with Tier IV testing,
  the role of the regulatory  authority is to design
  tests that lead to  a definitive LPC evaluation for
  the project.
      Under Tier IV evaluations, bioaccumulation
  testing measures the steady-state body burden of
  contaminants of concern in tissues of organisms
  subjected to long-term laboratory exposures or in
  tissues of appropriately sampled field organisms.
  The contaminant  concentration in  the tissues  of
  dredged-material  test  organisms  is  compared
  against the appropriate FDA action  levels and
  against bioaccumulation data obtained from  or-
  ganisms that are exposed  to reference-material
  sediment.  If contaminant bioaccumulation in the
  dredged-material organisms  is less  than the FDA
  levels  but greater than the levels in the reference-
  material organisms, organisms are collected from
  the vicinity of the disposal  site and analyzed for
  the contaminants of concern.  If the contaminant
  bioaccumulation of the dredged-material  organ-
  isms is lower than the steady-state body burden
  of the  field-collected organisms, the LPC for bio-
  accumulation is met.  If field-collected organisms
 have contaminant levels lower than those of the
 dredged-material organisms, case-specific criteria
 are developed to make a final LPC compliance
 determination for bioaccumulation.

 13.2.1.2.1   Type of Sampling Required

     Section 8.0 of the  1991  Green Book,  Collec-
 tion and Preservation of Samples, provides gener-
 al  information on  sampling plans  and  sample
 handling, preservation, and storage.
    To  adequately  and efficiently  conduct  a
 dredged-material  evaluation,  a  comprehensive
 sampling plan should be in place before sampling
 begins. Sufficient amounts of sediment and water
 should  be collected to conduct the necessary eval-
 uations.   Careful  consideration of maximum
 allowable and recommended holding  times  for
 sediments, as well as the exigencies of resamp- -
 ling, should.be given careful consideration.  Ad-
 ditionally, sample size should be small enough to
 be  conveniently handled and  transported, but
 large enough to meet  the  requirements for all
planned analyses.  The overall confidence of the
  final LPC determination is based on the following
  three factors.

      •  Collecting representative samples;

      • „ Using  appropriate sampling  techniques;
         and

      •   Protecting or preserving the samples until
         they are tested.

     Table 13-2  shows the general sampling  re-
 quirements  to conduct  dredged-material  testing.
 Actual sampling requirements are project-specific
 and are determined during the development of the
 project plan, based on the guidance that is provid-
 ed in the 1991  Green Book and in local agree-
 ments/manuals.

 13.2.1.2.2  Methods

    As  described  in Section  13.2.1.2.1  above,
 only  existing information  is evaluated in Tier L
 This  requires the careful compilation and analysis
 of such information.  If the information cannot
 show that the proposed dredged material meets
 one  of  the exclusionary  criteria,  or  if  the
 information is insufficient to reach an  LPC deter-
 mination, physical, chemical, and biological infor-
 mation on the dredged material and the ODMDS
 must be collected in Tiers II and/or HI.
    Proper sample collection,  handling, and pres-
 ervation are critical to the accurate evaluation of
 Tier II and m test results.  Sampling methods are
 usually developed by individual testing laborato-
 ries and documented in standard operating proce-
 dure (SOP) documents.  Consistent use of SOPs
 in the field and  laboratory ensure that sampling
.and analytical errors are minimized.
    Methods necessary to conduct  toxicity  and
bioaccumulation  evaluations  may  include  the
following:

    •   Sieving;

    •   Combustion;

    •   Gravimetry;
33-30

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                                                 33— "Green Book" Sediment-Testing Approach
                           Table 13-2. Sample-Collection Requirements
1 'Bgg^g^ _ .^^^^s 1!^^^^^^^^^= r ~
rests Water Samples
Disposal Dredging
Site Site Control'
Her II
Water Column
Screen D
Elutriate n D
ner H
Benthic
Tier 111
Water Column Db ° n
Tier HI
Benthic
Her IV
Water Column D D D
HeriV
Benthic ' 	 ; .
Sediment Samples
Dredging Reference
Site Site
n
D
D D
D D
n
n o
—

Control*
n
; - . D - . ,
======
•May or may not have to be field-collected.
"Dilution water; disposal-site water, artificial water, or clean seawater.
        Gas chromatography (GC);

        Electron-capture detection (BCD);

        Mass spectrometry (MS);

        Graphite furnace atomic absorption spec-
        troscopy (GFAAS);

        Atomic absorption spectroscopy (AAS);

        Inductively  coupled  plasma (ICP)  tech-
        nique;

        96-h elutriate toxicity bioassays;

        10-day  whole-sediment   toxicity   bio-
        assays;
               whole-sediment bioaccumulation
       tests (for trace-metals analysis only); and

 '  •  28-day  whole-sediment bioaccumulation
       tests. •-..--'       /.   .   '   '

    Project-specific   methods   necessary   to
conduct  Tier  IV water-column and benthic
evaluations may include laboratory and/or field
evaluations of long-term  toxicity  or  bioac-
cumulation effects  of  the  dredged material,
such as the following:

    •   Population-survival assessments;

    •   Community-change assessments; and

    •   Reproduction assessments.
                                                                                            13-11

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   Sediment Classification Methods Compendium
  13.2.1.23   Types of Data Required

      As discussed in Sections 13.2.1.2.1-13.2.1.2.4
  above, data required to conduct the LPC evalua-
  tions may include the following:

      •  Physical sediment data;

      "  Organic-  and inorganic-chemistry  sedi-
         ment data;

      «  Organic-  and  inorganic-chemistry  sedi-
         ment-elutriate data;

      •  Physical-oceanography data;

      «  Bioassay data;

      •  Bioaccumulation data; and

      •  Held species data.

 13.2.1.2.4   Necessary Hardware and Skills

     The hardware  and skills necessary to. conduct
 1991 Green  Book.evaluations  are relatively spe-
 cialized.  Many federal, state, and contract labora-
 tories have capabilities to conduct most or all of
 the necessary evaluations.  However, to conserve
 time  and  resources, field  sampling,  laboratory
 work, data management, and analysis of the re-
 sults are often conducted by separate  organizations
 according   to  aptitude,   cost,  and  scheduling
 parameters.
     The general categories of capabilities neces-
 sary to  reach a Tier m  dredged-material  LPC
 compliance determination are the following:

     •  Regulation and literature research;

     •  Field sampling at the dredging site,  dis-
        posal site, and reference site;

     *  Physical analysis of sediment  samples;

     •  Trace-metal (chemical) analysis of water
        and sediment samples;
      •   Organic-compound (chemical)  analysis of
          water and sediment samples;

      •   Numerical  modeling   for  initial-mixing
          analysis;

      •   Toxicity bioassay testing of elutriate sam-
          ples;

      •   Toxicity bioassay testing of whole-sedi-
          ment samples;"                    ,

      •   Bioaccumulation testing;

      •   Chemical analysis of tissue samples;

      •   Statistical analysis of test results;

      •  Quality-assurance   implementation
         (throughout evaluation); and
       '         -       t '  :'     	
      •  Compliance determination.

 13,2.1.3    Documentation     "

     Throughout the 1991 Green Book, references
 are provided  for the recommended sampling and
 testing methods,  data  analyses, QA procedures,
 and additional testing guidance.  For convenience
 to manual users, a copy of the U.S. Ocean Dump-
 ing Regulations (40 CFR Parts 220-228) is includ-
 ed in the 1991 Green Book as Appendix A.
    Information  on  documentation  and  record-
 keeping   is interspersed  throughout the  testing
 guidance.  Records ensure that  all aspects of the
 field and laboratory work are documented so that
 the  resulting  data  may be  properly interpreted.
 Dredged-material test data may be rejected if their
 history cannot be confidently traced.

 13.2.2 Applicability of Method to Human
       Health, Marine Life, or Wildlife
       Protection

    The  effects-based guidance provided  in the
 1991  Green Book is directly applicable  to the
protection of  human health, marine life,  and
23-22

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                                                  13— "Green Boot" Sediment-Testing Approach
wildlife because it is based on determining LPC
compliance.  If the testing shows that either the
LPC  for the water-column  or  benthic environ-
ment will be exceeded,  ocean disposal for the
proposed dredged  material is not supported.  In
40 CFR 227.27(a), the  LPC  is defined as the
concentration of the liquid phase of the dredged
material that will not exceed  either  the  estab-
lished WQC or 1 percent of the acutely  toxic
concentration following  the initial-mixing phase
(initial mixing is defined in 40 CFR 229.29). In
40 CFR 227.27(b), the LPCs for the suspended
particulate and solid phases are defined as those
concentrations  ".  . .  that will  not  cause unrea-
sonable acute or chronic toxicity or other suble-
thal  adverse effects  based on bioassay results
using appropriately sensitive marine organisms
... or will' not  cause  accumulation  of toxic
materials in the human food chain."
     The tiered-testing procedure in the manual
 establishes a conservative, yet workable, deci-
 sion-making process for environmentally protect-
 ive  dredged-material  management    Dredged
 material that poses no risk of adverse impact is
 readily supported for ocean disposal early in the
 procedure (i.e., Tier I or II).  Dredged material
 that has  unknown impact potential is evaluated
 to the level required to make a definitive LPC
 compliance determination.  Only dredged mate-
 rial that is shown to meet both the water-column
 and benthic LPC through state-of-the-art analyti-
 cal  techniques  is  supported for  open-ocean
 disposal.

  13.23  Ability of the Testing to Generate
         Numerical Criteria for Specific
         Chemicals

      The physical, chemical,  and biological data
  generated by the Tier II, III, and IV tests can be
  used -to  field-validate  SQC  that are presently
  under development.  The state-of-the-art samp-
  ling and analytical  techniques contained in the
  1991 Green Book guidance will provide for in-
  creases in  mettiod reproducibility, confidence of
  the test data,  and utility  to  SQC research and
  development projects.
     USEFULNESS
13 J.I  Environmental Applicability

    The guidance  in  the  1991 Green Book is
suitable for  dredged  material  regulated  under
MPRSA because it is  based on biological-effects
testing, which takes  into account  synergistic,
antagonistic,  and additive  effects of all contami-
nants in the material.  This approach  includes
both  water-column  and  benthic impact,  and
assesses  both toxicity   and  bioaccumulation.
Adaptations of the guidance are  also being ap-
plied to nearshore and Great Lakes dredge dis-
posal projects, and the tiered testing framework
may serve as a model for sediment assessments
under  other  regulatory  and   nonregulatory
programs.

13.3.1.1    Suitability for Different Sediment
            Types

    Except for extremely  coarse-  or  angular-
 grain  sediments,  the tiered-testing  approach is
 suitable for  all sediment  types. The test organ-
 isms recommended in the manual are suitable
 for most medium- and fine-grain dredged mate-
 rial.   If the dredged material being  tested is
 composed  of very  coarse sediments,  or  the
 dredged  material has other physical properties
 that  are  potentially  incompatible with  lecom-
 mended  test species, alternative organisms may
 be used  if they meet 40  CFR  227.27(c) and are
 ecologically relevant to  the disposal site.  Al-
 ternative test organisms  may  also be necessary
 to avoid grain-shape insensitivities when using
 sediment-ingesting  organisms.    Noncontami-
 nant-related mortality has been linked on  at
 least one occasion to internal organism damage
 that was caused  by  highly angular sediment  of
 moderate grain size (Oakland Harbor  sediment;
 Word et  al, 1990).   Sample  handling  and
 chemical  extraction of very  coarse-sediment
 dredged  material   can  also cause  analytical
 problems.                      .
     In general,  few  analytical  problems  are
  caused by sediment type.  Grain-size problems
  occur rarely because (1) most  large-grainTsize
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  Sediment Classification Methods Compendium
  sediment contains few contaminants and meets
  the  LPC  in  either Tier  I or II,  and (2) the
  tiered-testing  procedure  is  relatively  flexible
  and allows for alternative evaluation methods.

  13.3.1.2    Suitability for Different Chemicals or
             Classes of Chemical Contaminants

     Since the guidance in  the 1991 Green Book
  uses  effects-based tests, it  does not rely on the
  explicit identification of contaminants for deci-
  sion-making.  However, the guidance is suitable
  for detecting  and quantifying a wide range of
  organic and inorganic chemicals.  In Tier I of the
  testing procedure, target analytes are determined
  for the proposed dredged material.  If contami-
  nation is  suspected,  but  specific  contaminants
  cannot be  isolated  in the Tier I evaluation, the
 manual recommends that the dredged material be
 scanned for a broad spectrum of contaminants. A
 list of 131  potential target analytes is provided in
 Table 9-1 of the  1991 Green Book, Priority Pol-
 lutant and  301(1) Pesticides Listed According to
 Structural Compound Class.
    Extensive guidance for laboratory analysis of
 organic and inorganic compounds is provided in
 Section 9 of the manual,  Physical  Analysis of
 Sediment and  Chemical Analysis of  Sediment,
 Water, and Tissue Samples.  Target analytes for
 the water and  tissue  analyses are  the same as
 those  for whole-sediment analyses.  Guidance is
 also  provided  in  Section  9 of the  manual  for
 minimizing salt interferences with the chemical
 analyses.

 13.3.1.3    Suitability for Predicting Effects on
            Different Organisms

    All four tiers of the tiered-testing procedure
 consider effects on marine organisms that are
 representative of organisms that are indigenous to
 ODMDSs and have known impact tolerances.  In
 Tier  I,  information  on the proposed .dredged
 material's .effect on  laboratory and indigenous
 species is- analyzed.  In Tier  II, the  theoretical
 bioaccumulation  potential  (TBP) for  nonpolar
inorganic contaminants in the dredged material is
calculated and compared against that of the refer-
  ence site.   In  Tier III, water-column  toxicity,
  benthic toxicity, and benthic bioaccumulation are
  determined  for  ecologically relevant  laboratory
  organisms.   In  Tier IV, case-specific bioassays
  and bioaccumulation studies are conducted  on
  laboratory and/or field organisms.

  13.3.1.4    Suitability for In-Place Pollutant
             Control

     The 1991  Green  Book was developed  to
  determine water-column and benthic LPC com-
  pliance for  proposed dredged  material,  not for
  in-place management of contaminated sediments.
  However, the physical, chemical, and  biological
  tests that are recommended in the tiered-testing
  procedure are readily adaptable  to nondredging
  management of sediments.
     The sediment  data  that  are  generated with
  the guidance in the manual must be of suffi-
  ciently high quality to  develop LPC determina-
 tions  for  the dredged  material.   If these data
 show that the dredged material  does  not meet
 the LPC for ocean disposal, the same data are
 readily adaptable to other sediment-management
 uses, including in-place  pollutant management.

 13.3.1.5    Suitability for Source Control

     The purpose of the detailed  sampling  and
 testing guidance in the  1991 Green Book is to
 fully characterize  the dredged material  that  is
 proposed for ocean disposal.  Although it is not
 the  intended purpose, this  characterization may
 be useful for controlling sources of contaminants
 that are entering the sediments.
    If portions of a proposed project exceed the
 LPC, it benefits the applicant to isolate the com-
 pliant  and noncompliant areas  to  economize
 management of the dredged material. For exam-
 ple, material that meets the LPC might be dis-
 posed  of at an QDMDS and material that does
 not  meet the LPC might  disposed  of upland.
 During the process of site characterization, con-
 taminant gradients and source locations might be
 identified (such  as occurred in  New  Bedford
 Harbor, Massachusetts) and remedial or enforce-
ment actions can be directed as appropriate.
33-34

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                                                  13—"Green Boot" Sediment-Testing Approach
13.3.1.6    Suitability for Disposal
           Applications

    As  discussed  in Section 13.1  above,  the
guidance  in the' 1991 Green Book is  used to
conduct  LPC evaluations,  which  are  in turn
used, to  support  ocean-disposal management
decisions.  The manual is not intended to pro-
vide guidance on  other disposal options  avail-
able  to  dredged-material  managers.    Some
ocean and nonocean disposal options  may re-
quire additional or  alternative analyses of the
dredged  material  to reach decision  points.
Numerous other guidance manuals on dredged-
material   management   are  available   from
USEPA and the USAGE.
 13.3.2 General Advantages and Limitations

 13.3.2.1    Ease of Use

     As discussed in Section 13.2.1  above,  the
 tiered-testing procedure is  relatively  flexible.
 The dredged-material  evaluator. can enter and
 exit the testing procedures at any tier.  Howev-
 er,  to begin the evaluation in Tier II, III, or IV,
 the data must  satisfy  the requirements  of the
 earlier tier(s).  The overall ease of use of the
 testing  procedure depends  on  the  evaluator's
 familiarity with the following:

     •  Federal   regulations    pertinent   to
         dredged-material testing and disposal;

     •  Sources  of existing  dredged-material
         (sediment-quality) information;

     •  Sampling design;

     •  Numerical modeling;

     •  Physical,  chemical,   and   biological
         testing;

      •  Statistical analysis; and

      •  Quality assurance.
13.3.2.2    Relative Cost

    Tiers  I,  II, III,  and  IV  are  ordered by
increasing complexity and cost. Tier I is rela-
tively  inexpensive  and consists solely  of as-
sembly and  analysis of  existing information.
Tier IV  can be very expensive, consisting of
case-specific   toxicity   and   bioaccumulation
analysis, including extensive  field and  labora-
tory studies.   However, significant time  and
resources can  be saved if the earlier tiers are
completed to  the maximum extent possible
before proceeding to the later tier(s).  For ex-
ample, an in-depth analysis of "grey literature"
(university reports, etc.) might show the possi-
ble existence  of "hot spots" within a project.
The sampling  plan could then be designed to
appropriately  sample these  areas  of  concern
during a single sampling event, thereby saving
the time and expense required to conduct addi-
tional sampling at a  later  time:   Similarly,
money and time will be saved if LPC  compli-
 ance for nonpolar organic contaminants can be -
 shown in the Tier II TBP  calculation rather
 than  in  the  Tier ffl laboratory  testing and
 analysis.
     As all  dredging  projects contain case-spe-
 cific components, it  is difficult to estimate the
 overall cost of a typical dredged-material anal-
 ysis.  USEPA and the USAGE predict  that the
 updated  methods in  the manual would  not
 cause a significant increase in evaluation ex-
 penses and actually might lead to lower testing
 costs because LPC  determinations  might be
 achieved earlier im the testing process, thereby
 making full-scale bipassay and bioaccumulatibn
 laboratory  tests  unnecessary.   Also, as the
. recommended analytical methods  become re-
 fined, market pressures will force costs lower.

 13.3.2.3    Tendency to Be Conservative

     As discussed in Section 13.2.2 above, the
 ,tiered-testing procedure  is  very  protective of
 human health and the marine environment.  It
  is a  sequential and comprehensive analysis of
  the  proposed  dredged  material's  biological
  effects, as  shown by previous studies, model-
                                                                                           33-15

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  Sediment Classification Methods Compendium
 ing, and laboratory testing.    However,  the
 tiered-testing procedure is an "expert system";
 that is, the product of the procedure (LPC com-
 pliance determination) is only as good as the
 information that is integrated into it.
     To reach  a defensible  and  ecologically
 sound  LPC evaluation, high-quality information
 is required. There is risk of an inaccurate com-
 pliance determination if incomplete  or inac-
 curate  information is used, or if  good informa-
 tion is misapplied.  The regulations and numer-
 ous references in the manual should be consult-
 ed, and well-trained and experienced evaluators
 should be involved  throughout  the decision-
 making process.

 13.3.2.4    Level of Acceptance

     The  1991   Green  Book is the  official
 USEPA/USACE  guidance manual  for  deter-
 mining the suitability of dredged material  for
 ocean disposal.  During the development of the
 updated manual, comments from USEPA and
 USAGE   Headquarters,   USEPA   Regions,
 USAGE Districts, other  federal  agencies, port
 authorities,  special-interest  groups, and   the
 general public were solicited,   received,  and
 addressed as appropriate.  In 1990, USEPA and
 the USAGE conducted a public meeting on  the
 document1 and held six  regional training ses- -
 sions2  on the updated   methods.   The final
 manual is  the  product  of extensive USEPA/
 USAGE dredged  material program experience,
 current state-of-the-art  testing  methods,  and
 review  by  a wide  array of individuals  and
 agencies.

 13.3.2.5   Ability to Be Implemented by
           Laboratories with Typical
           Equipment and Handling Facilities

    Many  evaluations recommended  in  the
 1991 Green Book, particularly for organic and
 chemical  analysis, require standard  laboratory
'Washington, DC
'Nanagansett, RI; Gulf Breeze, FL; Vicksburg, MS; Ney-
port, OR; San Francisco, CA; and Washington, DC.
 equipment  and handling facilities.  However,
 some  laboratories  have  difficulty  attaining
 accurate  and precise test results for low con-
 taminant concentrations.  Agency  and contract
 laboratories that  presently  do  not have the
 capabilities  to  conduct  precise analyses will
 have to make significant investments in equip-
 ment, personnel,  and  training.   It is  expected
 that contract laboratories will choose to special-
 ize in only a few methods to be efficient and
 competitive in  the diedged-material  testing
 market.   Quality  assurance  (QA) program de-
 velopment,  although not equipment-intensive, is
 also a necessary and significant  investment for
 testing laboratories.  QA programs are neces-
 sary  to  ensure  that sample and data integrity
 are of sufficient quality and defensible.

 13.3.2.6    Level of Effort Required to
            Generate Results

    The overall  level  of effort  necessary  to
 conduct dredged-material analysis is  compar-
 able to that required by the preceding guidance
 (1977 Green  Book).   The  level  of effort is
 relatively low in Tier I and relatively  high in
 Tiers III and IV.  •      ,                •

 13.3.2.7   Degree to Which Results Lend
            Themselves to Interpretation

    The analysis of raw data that are generated
 during the tiered-testing procedure  is relatively
 complex,  especially  for bioassay and  bio-
 accumulation test data.  Interpretation of results
 is specifically described  and  decision  points
 and  values  are clearly  defined in the  1991
 Green Book.  Section 13 of the manual, Statis-
tical Methods, presents guidance for handling
the following:

    •  Unequal  numbers  of experimental  ani-
       mals assigned  to  each treatment con-
  ,     tainer  or loss  of animals  during the
       experiment;

   •  Unequal  numbers of replications of the
       treatments (i.e., containers or aquaria);
13-16

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                                                 13—"Green Book" Sediment-Testing Approach
    »  Measurements  scheduled for  selected
       time intervals but actually performed at
       other times;

    •  Different  conditions  of salinity,  pH,
       dissolved  oxygen,  temperature,   etc.,
       among exposure chambers; and

    •  Differences in  placement conditions of
       the testing containers or in  the animals
       assigned to different treatments.

USEPA and the USAGE are presently develop-
ing software and additional guidance to facili-
tate data  interpretation  for  dredged-material
evaluations.        •''•_,

13.3.2.8    Degree of Environmental
           Applicability

    The  USEPA/USACE (1991) effects-based
approach used to evaluate marine sediments has
wide environmental and  regulatory applicabil-
ity. The approach uses test organisms that

    •  Are sensitive to impact;

    •  Are reasonable representatives of indi-
       genous ODMDS species;

    •  Fulfill  the species  categories required
       by 40 CFR 227.27(c,d);

    •   Have extensive test databases; and

    •   Are hardy enough to withstand labora-
        tory procedures.

 Alternative test species that meet the guidance
 in the 1991 Green Book may  be used to avoid
 testing  problems such as grain-size tolerance
 and seasonal availability.  Complete elucidation
 and quantification of all chemical components
 in the sediment are useful, but not required, for
 regulatory  decision-making.     The  overall
 approach  is environmentally  conservative and
 relatively economical.                      •
    One feature of the 1991 Green Book guid-
ance  posing environmental limitations  is the
numerical modeling that is used in Tier I and. n
water-column  evaluations.    The  ADDAMS
models are not suitable  for calculating  water-
column  impacts  at  disposal  sites  that are
extremely  shallow  (i.e.,  where  the discharge
period from the disposal vessel is longer  than
the descent time to the bottom).   Additionally,
there is some uncertainty about the applicabil-
ity of the models for extremely deep (>200 m)
ODMDSs.

13.3.2.9   Degree of Accuracy and Precision

    The  1991  Green Book guidance strongly
emphasizes the importance of a comprehensive
QA program to achieve sufficient data  quality
during the tiered evaluation process.  QA issues
are  addressed in  subsections throughout the
data-generation sections of the  manual, and
Section  13, Quality-Assurance  Consideration,
gives guidance on the  structure  and compo-
nents  of  QA  programs  and   data-quality
assessment.
    The general guidance for QA program de-
velopment  includes  information  on  field and
laboratory sample handling, personnel training,'
and documentation.  For chemical analyses, the
guidance recommends appropriate use of'meth-
od   blanks,  procedural   blanks,   matrix
spike/matrix-spike   duplicates   (MSSD),   and,
standard  reference materials (SRM) to deter-
mine accuracy and precision of the data.  For
biological testing, the  importance of control-
sediment tests, reference-site  tests, and refer-
 ence-toxicant testing is discussed.


 13.4   STATUS

 13.4.1  Extent of Use

     The 1991  Green  Book guidance will be  ap-
 plied to all evaluations for dredged material that is
 proposed for disposal outside the baseline of the
 territorial sea (non-state waters).  Until completion
 of ongoing work on a national  testing manual for
 disposal shoreward of the baseline of the territorial
                                                                                          13-17

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  Sediment Classification Methods Compendium
 sea (dean Water Act section 404 waters), portions
 of the Green Book guidance are also expected to
 be  applied  to  nearshore  and   internal-water
 dredged-material disposal projects  in  the  United
 States.

 13.4.2  Extent to Which the Approach Has
        Been Field-Validated

     Large portions of the tiered-testing procedure
 for  dredged material have been  field-validated
 since the publication of the original guidance in
 1977 by  ongoing  state and federal dredging pro-
 grams.      Several   large-scale,   long-term
 USEPA/USACE projects in the New England and
 West Coast regions have applied and improved on
 the methods in the 1977 manual.  The guidance in
 the 1991 Green Book contains methods proven for
 marine sediment analyses, developed for national
 testing consistency, and  organized into  tiers  for
 efficient compliance  determination.  The tiered
 approach  for environmental monitoring of aquatic ,
 ecosystems  is  strongly recommended  by the
 National Research Council (NRC, 1990).

 13.43  Reasons for Limited Use

    Only  extreme time and resource  constraints
 (national emergencies, etc) would limit the use of
 the guidance in the manual.  Most  of the recom-
 mended procedures are already widely applied.

 13.4.4 Outlook for Future Use and
       Development

    USEPA and  the USAGE will continue to
 support and apply the guidance in the manual both
 nationally  and regionally.  Ongoing public  and
 private research  and  development  of  evaluation
 methods will continue to expand federal and state
 dredging-program experience.
   The manual will be  revised  at  a future date
 based on (1) the findings of an EPA SAB review
 (SAB, 1992); (2) technical advances in assessing
 sediment contamination and marine environmental
 impact; and (3) changes  to the  Ocean Dumping
Regulations.
 13.5   REFERENCES

 NRC.  1990.  Managing troubled waters: The
     role of  marine environmental monitoring.
     National Research Council. National Acade-
     my Press, Washington, DC.  125 pp.
 SAB.  1992.  An SAB report  Review of a
     testing  manual  for  evaluation . of  dredged
     material proposed for ocean disposal.  Pre-
    pared by the Sediment Criteria Subcommit-
    tee of the Ecological Processes and Effects
     Committee,   USEPA   Science  Advisory
    Board, Washington, DC. .EPA-SAB-EPEC-
    92-014.
 USEPA/USACE.   1977. Environmental Protec-
    tion Agency/United  States Army. Corps  of
    Engineers Technical Committee on  Criteria
    for Dredged and Filled Material.  Ecological
    evaluation of proposed discharge of dredged
    material  into  ocean waters;  Implementation
    manual  for section  103  of Public Law 92-
    532 (Marine Protection, Research, and Sanc-
    tuaries Act of 1972).   July 1977  (second
    printing April 1978). Environmental Effects
    Laboratory, United  States  Army Engineer
    Waterways Experiment  Station, Vicksburg,
    MS.  24 pp + appendices.
USEPA/USACE.  1991.  Environmental Protec-
    tion Agency/United States Army  Corps  of
    Engineers.  Ecological  evaluation of pro-
    posed  discharge of  dredged material  into
    ocean waters.   January 1990.  United States
    Environmental  Protection Agency,  Office  of
    Marine and Estuarine Protection, Washing-
    ton, DC   20460.    USEPA-503-8-90/002.
    219 pp + appendices.
Word, J.Q., J.A. Ward, JA.  Strand, N.P. Kohn,
    and A.L. Squires.  1990.  Ecological eval-
    uation  of proposed  discharge of  dredged
    material from  Oakland Harbor into  ocean
    waters (Phase II of  42-Foot  Project).  Pre-
    pared  for United States Army  Corps  of
    Engineers.   U.S.  Department  of Energy
    Contract No. DE-AC06-76RLO 1830.  Sept-
    ember 1990.
13-18

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          CHAPTER 14
National Status  and Trends  Program

Approach

Edward R. Long
Coastal Monitoring and Bioeffects Assessment Division
National Oceanic and Atmospheric Administration
7600 Sand Pt. Way, NE, Seattle, WA 98115
(206) 526-6338                                         •      ' •
Donald D. MacDonald
MacDonald Environmental Sciences, Ltd.
2376 Yellow Point Road, R.R.#3, Ladysmith, BC, Canada VOR 2EO
    Sediment quality criteria based  on multiple
 methods have  been recommended  for  broad
 applications in the United States (USEPA/SAB,
 1989; Adams et al., in press). The approach used
 by the National Status  and Trends  Program
 (NSTP) of the National Oceanic and Atmospheric
 Administration  (NOAA)  to develop informal,
 effects-based guidelines involves the identification
 of the ranges in chemical concentrations associat-
 ed with biological effects based on a weight of
 evidence from many studies. In this approach, the
 data  for  many  chemicals are assembled from
 modeling, laboratory, and field studies to deter-
 mine the ranges in chemical concentrations that
 are rarely, sometimes, and usually associated with
 toxicity. The data from many of the studies of the
 individual approaches described elsewhere in this
 document are compiled and examined to develop
 no-effects, possible-effects, and probable-effects
 ranges (Figure 14-1).
  14.1  SPECIFIC APPLICATIONS

  14.14 Current Use

     The NSTP Approach was used initially to
  develop informal guidelines for use by the Nation-
  al Status and Trends (NS&T) Program (Long and
  Morgan, 1990; Long, 1992).  NOAA  analyzes
  sediments from numerous locations nationwide as
  a part of its monitoring program. The guidelines
  were developed as tools  for identifying locations
  in which there is a potential for toxicity to living
 resources for which NOAA is the federal steward.
 Areas  in  which chemical concentrations often
 exceeded the guidelines were identified as high
 priorities for investigations of toxicity with biolog-
 ical tests.                      :.
    Environment Canada evaluated many  candi-
 date approaches to the development of sediment
 quality guidelines and elected  to develop  its
 national guidelines  using the NSTP Approach
 (MacDonald and Smith, 1991; MacDonald et al.,
 1991).  The Florida Department of Environmental
 Regulation elected to use the  NSTP Approach to
 develop state sediment quality guidelines as a part
 of its sediment management strategy (MacDonald,
 1992).  The California Water Resources Control
 Board will use the NOAA guidelines in its initial
 evaluations of ambient chemical data. Following
 that  step, data from field  studies, laboratory
 bioassays, and equilibrium partitioning  models
 will be used to develop sediment quality objec-
 tives (Lorenzato et al., 1991). Finally, the Liter-
 national Council for Exploration of the Sea Study
 Group on the Biological Significance of Contami-
 nants in Marine Sediments has elected to adopt
 the NSTP Approach in the development of guide-
 lines for participating nations (Dr. Herb Windom,
 Working Group on Marine Sediments, ICES,
 personal communication).
  .   Guidelines developed with the  NSTP Ap-
" proach were used by NOAA to identify chemicals
 that occurred in concentrations that were suffi-
 ciently high to warrant concern and to  identify
 sampling sites and areas in which there was a
 potential for toxicity (Long and Morgan,  1990;

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 Sediment Classification Methods Compendium
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                                                                          34—NSTP Approach
Long et al, 1991; Long and Markel, 1992).  It
was  presumed that the potential for toxicity was
relatively high in areas where numerous chemicals
exceeded  the  upper bounds of the guidelines.
Likewise, it was assumedthat the  potential for
toxicity was relatively low in areas where none of
the chemical concentrations exceeded the lower
bounds of the guidelines. In those regions with
the  highest potential  for toxicityj NOAA has
implemented regional surveys of toxicity, using a
battery of biological analyses and tests.
    Also, NOAA has used the  guidelines in
: assessments and prioritization of hazardous waste
sites (Dr. Alyce Fritz, NOAA Hazardous Materials
Response  and  Assessment  Division,  personal
communication). Other agencies and consultants
have used the guidelines as a means of placing
ambient  chemical  data  into  perspective with
respect to the potential for toxicity (for example,
Squibb et al., 1991  for New York/New Jersey
Harbor;   Mannheim  and Hathaway,  1991  for
 Boston Harbor; Soule et al., 1991 for Marina Del
 Rey). The Florida Department of Environmental
 Regulation has used  the guidelines as informal
 tools for interpreting ambient chemical data and
 for identifying regional  priorities for  sediment
 quality management (MacDonald, 1992).


 14.1.2 Potential Use

     Potential uses of the guidelines are as follows:

      •  Identification of potentially toxic chemi-
         cals in ambient sediments;

      •  Ranking and prioritization of  areas and
         sampling sites for further investigation;

      •  Assessment of potential  ecological haz-
         ards of contaminated sediments;

      •  Design of  spiked-sediment bioassay  ex-
         periments;

      •   Description of the kinds of toxic effects
          previously associated with specific con-
          centrations of chemicals;
       Quantification of the relative likelihood of
       toxicity over specific ranges in chemical
       concentrations; and

       Identification of the need for sediment
       management initiatives.
14.2 DESCRIPTION
14.2.1  Description of Method

    The NSTP Approach involves a simple evalu-
ation of available data to identify three ranges in
concentrations for each chemical:
             /  -               .'     "'-•'
    •   No-Effects Range:  The range in'concen-
        trations over which toxic effects are rarely
        or never observed;                    ', •

    m   Possible-Effects  Range:  The range in
        concentrations over which  toxic effects
        are occasionally observed; and

     •   Probable-Effects Range:  The range in
        concentrations over which toxic effects
        are frequently or always observed.

    These ranges  are identified by evaluating
 information  from  numerous  studies in which
 matching  biological and chemical  data  were
 developed. The specific steps in the method are:

     (1) Compile matching chemical and biologi-
        cal data from laboratory spiked-sediment
        bioassays, equilibrium-partitioning mod-
        els,  and field studies and determine the
        chemical concentrations associated with
        no observed effects and those associated
        with adverse effects.

     (2) Enter the data into a  database, including
        the type of biological test performed, the
        adverse effects) measured, the chemical
        concentrations associated with observa-
        tions of either effects or no effects, the
        type of study method and approach, and
         the degree of concordance between the
                                                                                             14-3

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 Sediment Classification Methods Compendium
        measure of effects and the concentration
        of the chemical.

     (3) For those analytes for which sufficient
        data  exist,  prepare  data  tables sorted
        according to ascending chemical concen-
        trations.

     (4) Arithmetically determine the no-effects
        range, possible-effects range, and prob-
        able-effects range for each chemical.

     The steps taken to select and screen candi-
 date  data   sets  are  described  in  Section
 14.2.1.2.3.  The approach is intended to encour-
 age periodic updates as new data become avail-
 able.
     Two  slightly  different methods have been
 used to determine the three chemical ranges.
 First, two percentiles in the chemical concentra-
 tions associated with toxicity were derived by
 Long and Morgan (1990): the lower  10th per-
 centile and the  50th percentile  (median).  The
 lower 10th  percentile  was  identified as  the
 Effects Range-Low (ERL), and the median was
 identified as the Effects Range-Median (ERM).
 In their evaluation of the  ascending data tables,
 Long and Morgan  (1990) used only the chemical
 concentrations that  had been associated with
 toxicity (i.e., the "effects"  data).  The conceptual
 basis for this  approach and the three ranges are
 illustrated in Figure 14-2.
    Later, MacDonald (1992) identified the three
 ranges with a  method that used both the concen-
 trations associated with biological, effects (the
 "effects"  data)  and those associated with no
 observed effects (the "no-effects" data). In this
 method, a threshold effects  level (TEL)  was
 calculated first as the square root of the product
 of the lower 15th-percentile concentration asso-
 ciated with observations  of biological effects
 (the ERL) and the  SOth-percentile concentration
 of the no-observed-effects data (the NER.-M). A
 safety factor of  0.5 was applied to the TEL to
 define a No-Observable-Effects Level  (NOEL).
 Next,  a  Probable-Effects  Level   (PEL) was
 calculated as the square'root of the product of
 the SOth-percentile concentration of the effects
 data (the ERM) and the 85th-percentile concen-
 tration of the no effects data (the NER-M).
     Neither  of  these  methods  is preferred or
 advocated over the other.  The significant fea-
 ture of this approach is the use of a weight of
 evidence developed in the ascending tables, not
 in the specific method of using the data tables.
 In addition to the two methods described here,
 many  others could be applied to the ascending
 data tables to derive guidelines.  The method
 used by MacDonald (1992) considered both the
 "effects" and "no-effects" data,  whereas that of
 Long and Morgan (1990) used only the "effects"
 data. Different percentiles in the ascending data
 were used in the two methods.  Despite these
 differences in the methods, the agreement be-
 tween the NOELs and ERLs and between the
 PELs  and the ERMs was very good, usually
 Within a factor of 2.
    In both  documents,  the  lower of the two
 guidelines for each chemical was assumed to
 represent the concentration below which toxic
 effects rarely occurred. The range in concentra-
 tions between the two values was that in which
 effects occasionally  occurred.   Toxic  effects
 usually or frequently occurred at concentrations
 above  the upper guideline value.
    As an example, Figure 14-2 compares the
 frequency distribution of toxic  effects and no-
 effects  data  associated with  concentrations  of
 napththalene to  the ERL and ERM concentra-
 tions for naphthalene. Long and Morgan (1990)
 reported the ERL as  340 ppb dry wt. and  the
 ERM  as 2100  ppb dry wt.  for  naphthalene,
 based  on an  ascending data table of 49 data
 points.  These guidelines defined three ranges of
 chemical concentrations:  the no-effects range
 (0-340   ppb);   the   possible-effects   range
 (340-2100 ppb); and the probable-effects range
 (>2100 ppb). Only 10.5 percent of the chemical
 concentrations below the ERL were associated
with toxic effects,  suggesting  that toxicity  is
unlikely below  the  ERL concentrations.   In
contrast, 81 percent of the chemical concentra-
tions between the ERL and ERM values were
associated with the toxic effects  and 93 percent
of the data points were associated with toxicity
at concentrations above the ERM value.
14-4

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                             14^-NSTP Approach
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 Sediment Classification Methods Compendium
14.2.1.1   Objectives and Assumptions

    The objective of the NSTP. Approach is to
provide informal, effects-based guidelines that
are based on a weight of evidence and reported
as ranges in concentrations.  The guidelines are
based  on  chemical  concentrations  associated
with .measures of biological  effects,  thereby
providing lexicological and/or biological releva-
nce to the guidelines.  They are based on data
from multiple studies and research methods, thus
providing a weight of evidence. In recognition
of the variability in  the  kinds  of  data that are
available, they are presented as ranges, instead
of absolute values, thereby providing a flexible
interpretive tool with broad applicability.  They
are presented along with all of the  supporting
evidence in ascending tables, providing the user
an  interpretive framework for comparison with
ambient data.
    In  this approach it is assumed that the data
from all individual studies are equal in weight and
credibility, although they may have involved very
different methods and test endpoints.  It is as-
sumed that the methods used by  the individual
investigators  were reasonably accurate.   Most
important, it is assumed that as the concentrations
increase, the potential for toxicity also increases,
thereby providing a conceptual basis for identify-
ing the ranges in concentrations frequently associ-
ated with no  toxic effects and  those frequently
associated with toxic effects. The guidelines can
be formulated to  account for site-specific factors
that control bioavailability (see Section 14.3.1.1).


14.2.1.2   Level of Effort

14.2.1.2.1  Type of Sampling Required

    The NSTP Approach relies  on the use of a
database compiled from a wide variety of sedi-
ment quality assessments. The database currently
contains over  800 entries generated by the three
major approaches to the establishment of effects-
based guidelines: equilibrium-partitioning models;
laboratory spiked-sediment bioassays; and various
assessments of matching, field-collected, sediment
chemistry, and biological effects  data.  The NSTP
Approach was specifically designed to use existing
data, therefore eliminating or minimizing the need
for additional sampling.  However, evaluation of
the regional applicability of the guidelines could,
in some cases, require further site-specific investi-
gations, the magnitude of which could vary con-
siderably.

14.2.1.2.2   Methods

    The methods for deriving numerical sediment
quality guidelines using the NSTP Approach are
summarized in Section 14.2.1. Also, these meth-
ods are described by Long and Morgan (1990) and
MacDonald (1992).

14.2.1.2.3   Types of Data Required

    The NSTP Approach was intended to integrate
a diverse assortment of information into a single
database to support the derivation of  numerical
guidelines.  Consequently, data from  numerous
modeling,  laboratory, and field  studies  were
collated into one database.   Ideally, the database
used to establish guidelines should include entries
from  all  three  of  these types of  approaches.
Suitable data were available from a wide variety
of sources. While collection and analysis of these
data sets were labor-intensive, subsequent, incre-
mental updates of the database should be relative-
ly simple and inexpensive.
    The data compiled from numerous studies
were entered into the Biological Effects Database
for Sediments (BEDS) by MacDonald (1992).  All
of the compiled data were fully evaluated prior to
incorporation into the BEDS  to ensure internal
consistency in the database. The screening proce-
dures used to support the  development of  the
BEDS were designed to ensure that only relevant
and high-quality data were used  to derive  the
guidelines.  No subjective biases were employed
in screening the data; as many sources  of data
were included as possible.  Candidate data from
each study were evaluated to determine the  ac-
ceptability  of the experimental  design, the test
protocols, the analytical methods, and the statisti-
cal procedures that were used. Only data in which
there were matched measures of sediment chemis-
14-6

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                                                                           14—NSTP Approach
try, and biological  effects were  included.  The
database included only those data in which either
statistically  significant biological  results were
obtained or in which major  differences in the
biological results between samples were reported.
    The BEDS currently  includes over 800 data
entries,  mainly  data from  studies  performed
throughout  North America.  It  was  developed
jointly by NOAA, Florida Department of Environ-
mental Regulation,  Environment  Canada,  and
MacDonald Environmental Services Ltd.
    In the evaluation of candidate data from field
studies, only  those data  were used in which  at
least a 10-fold difference in the concentrations  of
at  least one  chemical among the samples  was
reported.  Once this criterion was met, the  data
from many of the field studies were evaluated to
determine the. mean chemical concentrations  in
toxic  samples (i.e.,  significantly different from
controls) and those in  nontoxic samples of  in
samples with relatively depauperate benthic com-
munities (i.e., those with low abundance or species
richness) versus those with more robust communi-
 ties. Further, those mean concentrations in biolog-
 ically affected samples that exceeded  by twofold
 or more the mean  concentrations in the back-
 ground, reference, or nonaffected samples were
 assigned an asterisk in the ascending tables.  The
 asterisks symbolized that a biological effect was
 noted  and that  there was a strong  association
 between the chemical gradient and the biological
 gradient. Concentrations associated with nontoxic
 reference conditions were  noted as "no  effects."
 Those in which there was no concordance between
 the measures of effects  and  chemical concentra-
 tions were noted as "no gradient"  or  "no concor-
 dance."  The concentrations derived in the model-
 ing and spikedTsediment bioassays were always
 assigned  asterisks.    The  concentrations  with
 asterisks were used as "effects" data by both Long
 and Morgan (1990) and MacDonald (1992).

  14.2.1.2.4   Necessary Hardware and Skills

     The primary skills required to derive guidelines
  are associated with the development of the database.
 Expertise is required to evaluate the suitability of the
 biological  and chemistry data, using the screening
 criteria.   This process requires  experience in the
 evaluation of sediment data and the methods that
 were used to develop the data.
     The database has been developed on a personal
 computer arid is readily transferable to other sys-
 tems, but requires knowledge of the use of a com-
 puter.  The database provides a means of storing
 and accessing all of the information that relates
 chemical concentrations to adverse biological  ef-
 fects. This information can be manipulated  in this
' environment or exported into other formats.

 14.2.1.3    Adequacy of Documentation

     The NSTP Approach was documented by Long
 and Morgan (1990), in whjch the approach was
 .peer-reviewed both within and outside NOAA.  A
 second printing of the document was issued in 1992,
 following farther review.  A synopsis of the  ap-
 proach was described in a scientific journal  (Long,
 1992).  The approach has been described orally in
 numerous  technical  and scientific forums.   Mac-
 Donald and  Smith  (1991) and MacDonald et aH.
 (1991) described the application of the approach in
 the development of guidelines  for Canada.  Mac>
 Donald (1992) described the use of the approach in
 a  statewide  sediment  management  strategy   for
 Florida.

  14.2.2  Applicability of Method  to Human
         Health, Aquatic Life, or Wildlife
         Protection

      The guidelines  are intended to provide  an  esti-
  mate of the potential for adverse biological effects of
  sediment-associated contaminants  on benthic organ-
  isms, based on a weight of evidence from analyses
  performed with multiple species and/or biological
  communities.- They accommodate and rely on the
  data from tests of acute and chronic toxicity and from
  analyses of benthic community structure. The guide-
  lines are based on data from many different areas and
  oceanographic regimes, thereby  broadening  their
  applicability.  Currently, the data entered  mto the
  BEDS are from only marine and estuarine areas.
      .The guidelines provide a means of numerically
  estimating the percent frequency of biological effects
  over the three ranges of concentrations. The ascend-
                                                                                               14-7

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 Sediment Classification Methods Compendium
 ing tables accompanying the guidelines also provide
 a supplementary basis for interpreting new ambient
 chemical data. Also, these tables provide a visual and
 statistical means of estimating the relative degree of
 certainty in the guidelines.
     The guidelines are not intended to be used for the
 protection of human life or wildlife. Rather, they are
 intended to be used in estimating the potential for
 adverse effects among benthic communities.

 1423   Ability of Method to Generate
         Numerical  Criteria for Specific
         Chemicals

     Long and Morgan (1990)  reported numerical
 guidelines for 41  chemicals, including  12  trace
 metals,  18   polynuclear  aromatic  hydrocarbons
 (PAHs),  and  11  synthetic  organic compounds.
 MacDonald (1992) developed guidelines for 9 trace
 metals, total PCBs, 13 PAHs, 3 classes of PAHs, and
 2 pesticides.
     Conceptually, guidelines derived  using  Ihis
 approach could be developed for any toxic chemical,
 provided sufficient  data  exist  and  provided the
 toxicity of the chemical is dose-responsive. Long and
 Morgan (1990) assigned a high degree of confidence
 to guidelines  for chemicals for  which data  existed
 from many different approaches, different regions,
 and in which there was a good agreement in the data
 from different studies. MacDonald (1992) calculated
 guidelines only for those chemicals for which  there
 was a minimum of 40 data points, after determining
 the minimum amount of data necessary to calculate
 reliable and consistent values.  These minimum data
 requirements were established by iteratively calculat-
 ing guidelines using data sets of increasing size  (e.g.,
 4 to 60 data points) and determining when the
 estimate of the guidelines stabilized.
143    USEFULNESS

143.1  Environmental Applicability

143.1.1    .' Suitability for Different Sediment Types

 •   The NSTP Approach can be applied equally to
any sediment type that occurs in freshwater, estuarine,
and marine environments.  Since the database that
 supports the guidelines contains information from a
 wide variety of sediment types, the resultant guide-
 lines are considered to be widely applicable.  An
 increasing amount of information suggests that the
 bioavailability,  and,  therefore,  toxicity,  of many
 contaminants is controlled by such factors as TOC,
 AVS, and grain size.  The BEDS currently  accom-
 modates the data for these  variables, and, conse-
 quently, the guidelines could be normalized to the
 appropriate  factors  that  control   bioavailability.
 However, insufficient information currently exists to
 derive guidelines 'that are expressed  in these terms.
 It is anticipated that future revisions of the guidelines
 will be expressed in these terms, thereby increasing
 their applicability.         •
     Partly to increase the suitability of the guidelines
 to different sediment types,  they are expressed as
 ranges in concentrations, not absolutes. These ranges
 provide a basis for evaluating chemical concentrations
 in the different types of sediments represented in the
 BEDS. Li addition, the ascending data tables used to
 generate the guidelines can be examined to calculate
 frequency  distributions of effects and no  effects
 within each range of concentrations. These frequency
 distributions can be used as estimates of the probabili-
 ty of toxic effects.

 14.3.1.2    Suitability for Different Chemicals
            or Classes of Chemicals

    The approach can be applied to a wide variety
 of chemicals for which analytical methods are
 available.   Thus  far,  numerical guidelines have
 been developed by Long and Morgan (1990) and
 by MacDonald (1992) for 43 and 28 chemicals or
 classes of  chemicals, respectively.   Data are
 included in the  BEDS for over 200 chemicals or
 classes of chemicals.  Guidelines could be devel-
 oped for all of these substances when sufficient
 information becomes available.
                                    '
 14.3.1.3    Suitability for Predicting Effects on
            Different Organisms

    Since  the   database compiled from  many
different studies  is based  on  tests or  analyses
performed with many different species, the guide-
lines are widely applicable to benthic organisms.
14-8

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                                                                            14—NSTP Approach
In addition, the species studied in each investiga-
tion is(are)  listed in  the database; therefore,
species-specific applicability can be evaluated by
the users; Furthermore, the ERL values often are
based on data from relatively sensitive species  or
life stages, and, therefore, can be used as guide-
lines suitable for the protection of sensitive spe-
cies.
 14.3.1.4
Suitability for In-Place Pollutant
Control
    Numerical  sediment  guidelines developed
 using the NSTP Approach can be used in a variety
 of ways as a tool in pollutant control. Specifical-
 ly, these assessment tools respond to regulatory
 requirements by:

     •   Providing a basis for evaluating existing
         sediment chemistry data and ranking areas
         of concern  and chemicals of concern in
         terms of their potential  for causing toxici-
         ty and

     •   Identifying the need for further investiga-
         tions, such as biological testing, to sup-
         port regulatory decisions.

     As is the case with all of the other approaches
 that rely on data collected in the field, the guide-
 lines derived using the NSTP Approach integrate
 information  obtained from studies  of  complex
 mixtures of contaminants  and thereby  consider
 their interactive effects. .  Consideration  of the
 effects of contaminant mixtures is an advantage in
 the assessment of in-place pollutants in real-world
 conditions. However, this approach also relies on
 and gives equal weight to the data from equili-
 brium-partitioning  models and laboratory spiked-
  sediment bioassays performed with single chemi-
  cals (see Section 14.2.1.1).'

  14.3.1.5   Suitability for Source Control

      A reasonable  amount of confidence in sedi-
  ment quality guidelines is needed to justify using
  them in source control actions.  Since the guide-
  lines  are developed with  a weight of evidence
compiled from many different studies> they pro-
vide a credible and defensible basis for evaluating
contaminants  in real-world  conditions. •  The
guidelines provide an efficient basis for identify-
ing  priority  chemicals  and .priority  areas  that
would benefit from source controls. In addition,
the ascending tables provide a basis for estimating
the probability of observing adverse effects at sites
of  interest, reducing the probability of effects
through source controls, and evaluating  the im-  ,
provements  in  sediment quality following  the
implementation of source control measures.

14.3.1.6    Suitability for Dredged Material
            Disposal Applications

     Neither  the numerical guidelines   nor  the
 frameworks that have been developed  for  their
 application are intended to replace accepted testing
 protocols  for dredged material  disposal evalua-
 tions Nonetheless, these guidelines can provide
 relevant  tools  for  estimating the potential for
 adverse biological effects of contaminants associ-
 ated with solid-phase sediments.

 143.2 General Advantages and Disadvantages

 14.3.2.1     Ease of Use

     The approach has the advantage of relying on
 existing data.  Therefore, guidelines can be devel-
 oped relatively quickly and easily.
      The original efforts by Long and Morgan (1990)
 and MacDonald (1992) to  assemble the databases
 used to develop the guidelines were labor-intensive.
 Numerous reports and data sets Were located, and a
 huge amount of data was entered into spreadsheets.
 However, these data now exist in a centralized,
 computerized database, the  BEDS.  Subsequent
 derivations  of guidelines based on iterative expan-
 sions of the BEDS database should  be relatively
  quick, easy, and inexpensive.
      The guidelines .are easily used and interpreted.
  Chemical data can be readily compared with the
  guidelines and with the ascending tables.  The fre-
  quency of occurrence of toxicity over the no^effects,
  possible-effects, and probable-effects ranges can be
  calculated and compared with  the chemical data.
                                                                                                14-9

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  Sediment Classification Methods Compendium
 Sediments in which numerous chemicals occur at
 concentrations that fall within the probable-effects
 ranges have a higher probability of being toxic than
 those in which most of the chemical concentrations
 are within the no-effects range. This type of simple
 interpretation makes the guidelines very easy to use.

 14.3.2.2    Relative Cost

     The original effort of Long and Morgan (1990)
 involved roughly one year of labor. The confirma-
 tion and expansion of the database by MacDonald
 (1992) involved more than another year of labor.
 The costs of subsequent iterations of the guidelines
 based on further expansions of the database would
 vary with the amount of data entered and the num-
 ber of chemicals.  The calculations of the guideline
 values themselves are very simple and quick. Also,
 the guidelines can be used very quickly and easily.
     If the necessary data are not available for entry
 into a database, then  the costs to generate them
 could be relatively high. - If initiated  de novo,
 modeling, bioassay, and field studies  necessary to
 generate sufficient data could vary considerably in
 costs and time, depending on the amount  of data
 needed.

 14.23  Tendency to  Be Conservative

    The predictive capabilities of  the guidelines
 have  not been independently quantified.   The
 protectiveness of the guidelines could  be  increased
 by considering  data only from chronic  sublethal
 endpoints or by applying a numerical safety factor,
 such as was applied in the Florida guidelines (Mac-
 Donald, 1992).  Also, the guidelines would become
 more conservative  if data were included only from
 areas in which toxicants were highly bioavailable.

 143.2.4    Level  of Acceptance

    The  NSTP Approach has been published by
 NOAA, following an  in-house and outside peer
 review. It has  been published in a peer-reviewed
 scientific journal.  The approach has been used by
 Environment Canada  and Florida Department of
 Environmental Regulation in the development of
 their respective guidelines.  It has been adopted by
 a committee of the International Council for Explo-
 ration of the Sea for use by member nations. The
 State of California has adopted a similar approach to
 the  development  of sediment quality objectives
 (Lorenzato et al., 1991).
     The numerical guidelines developed by use of
 the approach have been used by NOAA to compare
 and rank the potential for toxicity at monitoring sites
 nationwide, within San Francisco Bay, and within
 Tampa Bay,   Approximately 1500 copies of the
 report by Long and  Morgan (1990) have been
 distributed.  Users  of  the report  have compared
 ambient  concentrations with  the guidelines  in
 assessments of hazardous waste sites, analyses of
 prospective dredge material, evaluations of survey
 and monitoring data,  and estimates of ecological risk
 (for  example, Mannheim  and Hathaway,  1991;
 Soule et al, 1991; Squibb et al,  1991).  NOAA
 routinely  uses the guidelines in  its  estimates  of
 ecological risk at National Priority List hazardous
 waste sites.  The guidelines have been used as a
 basis for interpretation  of chemical data in court
 cases.                     •                 .

 14.3.2.5   Ability to Be Implemented by
           Laboratories with Typical
           Equipment and Handling Facilities

    The spreadsheets   and database needed  to
 generate the guidelines can  be prepared with a
 personal computer and need  not be very compli-
 cated.  Entry of data  into the database and the
 generation of the ascending tables are very simple.
 The calculations of the guidelines  can be per-
 formed manually, on a desk-top calculator or a
 personal computer.  The database can be supple-
 mented with, new data as they become available.
 Implementation of the approach can become more
 laborious and complicated if the necessary data
 must be generated de novo.

 14.3.2.6    Level of Effort Required to
           Generate Results

    As outlined in Section 14.3.2.2, the level of
 effort required in the development of the original
set of guidelines was relatively, high.  Subsequent
iterations  of  the guidelines for other purposes,
14-10

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                                                                            14—NSTP Approach
other chemicals, or for the same chemicals follow-
ing additions to the database would be relatively
easy.  Entry of  new  data  points  from spiked-
sediment   bioassays,    equilibrium-partitioning
models, or apparent effects thresholds  into the
database  would  require only  a  few  minutes.
Manipulation of raw matching  data from biologi-
cal and chemical analyses  performed in a field
study would require from a few hours to several
days, depending  on the  size of the data set, fol-
lowed by entry of the data points  into the data-
base.    .

14.3.2.7    Degree to Which ^Results Lend
            Themselves  to Interpretation

     The guidelines and  the ascending data tables
on which they are based can be used in a number
of ways.  First, the data  from analyses of ambient
samples can be  compared visually with the two
numerical guidelines  to determine whether the
 ambient concentrations exceed either of the guide-
 lines. Second, the ambient concentrations can be
 compared with the data  in the  ascending tables to
 determine the kinds of toxic effects that have been
 observed in previous studies: at the concentrations
 of concern.  Finally, the frequencies of toxicity in
 the  no-effects,  possible-effects,  and  probable-
 effects ranges can be used to predict the probabil-
 ity of toxicity associated with any contaminant
 concentration.
     The guidelines developed thus far with this
 approach do not account for the effects of factors
 that control bioavailability of the toxicants. This
 is not a weakness of the approach; rather, it is a
 weakness of the available data. Nevertheless, this
 weakness may  hinder  interpretation of ambient
 data with  the guidelines.   The BEDS database
 includes a provision for entering data from analy-
 ses of acid volatile sulfides and total organic
 carbon  (and  other potential normalizers)  and,
 therefore,  would lend  itself  to recalculation of
 guidelines normalized  to  these factors once the
  necessary data become available.
      An important strength of  this  approach is that
  it provides the user some flexibility in the use and
  interpretation of the guidelines.  All of the data
  are provided in ascending order for the user to see
and evaluate.  The degree of certainty in the data
can be assessed and judged by, the user. Ranges
in concentrations are provided,  instead of rigid,
single absolute values.
    One of the most  attractive features  of this
approach is the estimation of the probability of
biological effects, based on the frequency distribu-
tions of effects for each chemical.  For example,
the data in the BEDS  database indicate that only
5.8 percent of the chemical concentrations within
the no-effects range for cadmium (0 to 1 mg/kg)
determined by MacDonald (1992) were associated
with adverse  biological effects (Figure 14-3).
These data suggest that there is  a low probability
of observing adverse effects within this range.
Within the probable effects range for cadmium
(>75 mg/kg), roughly 68 percent of the database
 entries were  associated  with  adverse  effects.
 These data suggest that there is a relatively high
 probability of observing adverse effects within this
 range. Positive concordance between frequency of
 effects and chemical concentrations should inspire
 confidence in the guideline values.
     Evaluation of  the guidelines  for  mercury
 reveals that a lower level of confidence should be
 placed on the guidelines for this element.  The
 data in the BEDS database indicate that within the
 no-effects range (0  to  0.1  mg/kg),  roughly  7
 percent of the entries were associated with adverse
 effects (Figure 14-4). However, frequency distri-
 butions of effects are similar within  the possible-
 effects range (0.1 to 1.4 mg/kg) and the probable-
 effects range (>1.4 mg/kg), namely  30.1 percent
 and 33.3 percent, respectively.  Therefore, it is
 more difficult to adequately determine the unac-
 ceptable levels of mercury in sediments than with,
 say, cadmium.

  14.3.2.8   Degree of Environmental
             Applicability

      The guidelines are highly applicable to the
  interpretation of environmental data.  They are
  generated with data from environmentally realistic
  field  studies, as well  as theoretical  modeling
  studies  and controlled  laboratory  experiments.
  They are generated with data from many different
  regions in which the mixtures and concentrations
                                                                                              14-11

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Sediment Classification Methods Compendium
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           14-13

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  Sediment Classification Methods Compendium
  of chemicals differ and in which sedimentological
  properties differ.  They are generated with tests
  using different' species with different sensitivities
  to toxicants.  They are universally applicable in
  North America since they  were generated with
  data from many regions in the United States and
  Canada.  Confidence in the utility of the guide-
  lines is inspired by the weight of evidence from
  these multiple studies.

  14.3.2.9    Degree of Accuracy and Precision

     By iteratively adding and  removing different
  data sets  from the ascending  tables, MacDonald
  (1992) determined that a minimum of 40 data sets
  were needed  to develop  consistent  and reliable
  guidelines.  Clearly, some variability in the guide-
  lines is to be expected as data  are added or delet-
  ed, but, once the minimum amount of. data  is
  compiled, this variability appears to be minimal.
     MacDonald  (1992)  generally  doubled  or
  tripled the amount of data in the ascending tables
  compiled  by Long and  Morgan (1990) mainly
 with new data from field studies and laboratory
 spiked-sediment bioassays.    Also,  MacDonald
 (1992) considered only estuarine and marine data,
 thereby deleting the freshwater data included in
 Long and Morgan (1990).  The effects on the
 guideline concentrations of eliminating some data
 and adding a substantial amount of new  data are
 illustrated in Tables 14-1 and 14-2.  The ERL and
 ERM values,  based on the Long and  Morgan
 (1090) data tables and  the larger MacDonald
 (1992) tables, are compared by  using the methods
 of Long and Morgan (1990)  applied to both data
 sets.
    For 13 aromatic hydrocarbons, the average of
 the ratios between the two sets  of guidelines was
 IS (1.9 for the ERLs and 1.2 for the ERMs). For
 eight trace metals, the average of the ratios be-
 tween the  two sets of guidelines was 1.7.  The
 trace metals ERL values changed more than the
 ERM values (average ratios of 1.9  for the ERLs
 and 1.5 for the ERMs).
    Overall,  7 of the 23  ERL values  did not
 change and the ratios between the two  sets of
ERL values ranged from 1.0 to 9.4.  Also, 7 of
the 23 ERM values did not change.  Of the 46'
  values, 14 remained unchanged, 17 increased, and
  15 decreased. The overall mean factors of change
  were less than twofold for both trace metals and
  PAHs.  These observations suggest that the guide-
  lines are not terribly sensitive to the addition of
  new  data  once a minimum  amount has been
  compiled.  Also, they suggest that the guidelines
  originally developed by Long and Morgan (1990)
  generally  are  substantiated by  additional data
  compiled by MacDonald (1992).
     The accuracy of the guidelines in predicting
  toxicity has not yet been quantified.  However, in
  the Hudson-Raritan estuary, the concentrations of
  many chemicals quantified in previous  studies
  (Squibb et  al., 1992) frequently exceeded the
  ERM guidelines in the Arthur Kill  and rarely
  exceeded them in the lower Hudson River.  In a
  recent survey funded by NOAA, sediments from
  the Arthur Kill were extremely toxic to amphipods
  and other species, whereas  the sediments from the
 lower Hudson River were not toxic.
 14.4 •  STATUS
 14.4.1  Extent of Use

    The  NSTP  Approach  is  being  used  by
 NOAA's National Status and Trends Program, by
 Environment Canada, and by the Florida Depart-
 ment of Environmental Regulation.  A variation
 on the approach is being pursued by the  California
 Water Resources Control Board. Other  states and
 regional districts have inquired about the possible
 use of the approach.

 14.4.2  Extent to Which Approach Has Been
        Field-Validated

    Validations of  the guidelines  have not yet
been quantified. As described in Section 14.3.2.9,
the original  set  of guidelines generally were
substantiated by  the addition of  considerable
amounts of new data, largely from field studies
performed in many regions.  The concordance
between predictions  of toxicity with the guidelines
and actual observations of toxicity has been very
14-14

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                                                                               14—NSTP Approach
Table 14-1   Ratios Between the Guideline Values for Polynuclear Aromatic Hydrocarbons Determined with
       Data from Long and Morgan (1990) and Those Determined with Data from MacDonald (1992).
                    Total number of data points available are listed (with those
                          used to determine guidelines in parentheses).
I
Chemical
Analyte
=======
MacDonald
(1982)
======
Long and
Morgan
<1B90)
==============
Polynuclear aromatic hydrocarbon* (ppb d.w.)
Acenaphthene
ERL
ERM
Anthracene
ERL
ERM
Fluorene
ERL
ERM
2-methylnathphalene
ERL
ERM
naphthalene
ERL
ERM
phenanthrene
ERL
ERM
benzo(a)anthracene
ERL
ERM ' ...
benzo(a)pyfene
ERL
ERM
• chrysene
ERL
ERM
<- dibenzo(a,h)anthracene
ERL •
ERM
fluoranthene
ERL . .
ERM
pyrene "
ERL
ERM
total PAH
ERL
ERM
Mean change in PAH ERLs
Mean change in PAH ERMs
n=69(30)
16
500
n=88{46)
85.3
1100
n=95(48)
19
540
n=49(28)
70
670
n=97(44)
160
2100
n=101(51)
240
1500
n=81(43)
261
1600
n=89(44) '
430
1600
n=89(45)
384
2800
n=76(31)
63.4 -
260
nil 17(71)
600
5100
n=93(50)
665
2600
n=78(34)
4022
44.790


Overall mean change in PAH values
n=35(15)
150
650
n=39(26)
85
960 :
n=44(28)
' 315.
640
n=31(15) -
65
670
" n=50(28)
340
2100
n=49(34)
225
1380
ri=34(30)
230
1600
n=43(27)
400
2500
n=41(27)
400
2800
n=23(18)
•> 60
260
n=51 (33)
600
3600
n=43(28)
350
2200
. n=63{34)
4000
35,000
1.90
1f17 i
1.53
                                                                   Ratio Between
                                                                 Two Sets of Values
                                                                      2.0(2.0)
                                                                      9.4
                                                                      1.3
                                                                      2.3(1.8)
                                                                      1.0
                                                                      1.1
                                                                      2.2(1.7)
                                                                      1.8
                                                                      1.2
                                                                      1.6(1.9)
                                                                      1.1
                                                                      1.0
                                                                      1.9(1.6)
                                                                      2.1
                                                                      1.0
                                                                      2.1(1.5)
                                                                      1.1
                                                                      1.1
                                                                      1.9(1.4)
                                                                      1.1
                                                                      1.0
                                                                      2.1(1.6)
                                                                      1.1
                                                                      1.6
                                                                      2.2(1.7)
                                                                      1.0
                                                                      1.0
                                                                      2.3(1.7)
                                                                      1.1
                                                                      1.0
                                                                      2.3(1.8)
                                                                      1.0
                                                                      M
                                                                      2.2(1.8)
                                                                       1.9.
                                                                       1.2
                                                                       1.2(1.0)
                                                                       1.0
                                                                       1.3
   Values
Increased (+)
Decreased (-)
                                                                                                   14-15

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  Sediment Classification Methods Compendium
     Table 14-2.  Ratios Between the Guideline Values for Total PCBs and Trace Metals Determined with
        Data from Long and Morgan (1990) and Those Determined with Data from MacDonald (1992).
                   Total number of data points available are listed (with those
                          used to determine guidelines in parentheses).
Chemical
Analyte
PoIychtoriiMted BIphenyl
total PCS
ERL
ERM
Tr«c» Metals (ppm d.w.)
arsenic
ERL
ERM
cadmium
ERL
ERM
copper
ERL
ERM
chromium
ERL
ERM
lead
ERL
ERM
mercury
ERL
ERM
nickel
ERL
ERM
silver
ERL
ERM
zinc
ERL
ERM
Mean change In PAH ERLs
Mean change in PAH ERMs
MacDonald
(1892)
(ppb d.w.)
n=1 26(50)
22.7
180

n=143(27)
8.2
70.0
n=261(84)
1.2
9.6
n=221(76)
34.0
270
n=1 97(37)
81
370
n=210(73)
46.7
223
n=1 69(42)
0.15
0.71
n=169(19)
20.9
51.6
n=96(25)
1.0
3.7
n=214(74)
150
410


Overall mean change in metals values
Long and
Morgan
(1990)

n=77(33)
50
400

n=48(16)
33.0
85.0
n=1 06(36)
5.0
9.6
n=91(51)
70.0
390
n=76(21)
80
145
n=83(47)
35.0
110
n=76(30)
0.15
1.3
n=5€(18)
30
50
n=47(13)
1.0
2.2
n=79(46)
120
270


1.74
Values
Ratio Between Increased (+)
Two Sets of Values Decreased (-)

1.6(1.5)
2.2
2.2 .

3.0(1.7)
4.0
1.2 .
2.5(2.3)
4.2
1.0 . •
2.4(1.5)
2.0 -
1.4 - ,
2.6(1.8)
1.0
2.6 +
2.5(2.6)
1.3 . + °
2.0 +
-2.2(1.4)
1.0 - .
1.5
3.0(1.1)
1.4
1.0
2.0(1.9)
1.0 •
1.7 +
2.7(1.6)
1.25 +
1.5 +
.1.9
1.9

14-16

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                                                                          14—NSTP Approach
good thus far, but the degree of concordance has
not been quantified.  Additional opportunities to
field-validate the guidelines will be available in
future studies in Tampa Bay, the Hudson-Raritan
estuary, and southern California.

14.4.3  Reasons for Limited Use

    The NSTP  Approach initially was used by
NOAA to develop informal guidelines for internal
agency use.  Therefore, knowledge of and access
to the guidelines was limited. As interest in the
guidelines increased,  they  were released in  a
government document with a limited distribution.
Therefore, the main  reason for the limited use of
the approach has been the limited awareness of its
existence. Furthermore, the equilibrium-partition-
ing approach to national  criteria and  the most
successful regional approach to criteria (apparent
effects thresholds in Washington) have received
considerable attention. Moreover, the guidelines
thus  far have  not  considered the potential for
bioavailability or bioaccumulation because of a
 lack of data.

 14.4.4  Outlook for Future Use and Amount
         of Development Yet Needed

     There is significant potential for the expanded
 use of the NSTP Approach.  Canada, Florida, and
 California  currently are using the approach  to
 develop their respective guidelines.   Since the
Approach relies on existing  data, other region-
 specific guidelines could  be developed easily,
 using the data available  from specific  regions.
 The approach  can be used to validate criteria
 developed with other single-method approaches.
 The database can be accessed for specific regions
 or for fresh, estuarine, or marine waters.
     Several types of data are needed to further
 develop the approach.  First, additional data are
 needed from studies in which TOC, grain  size,
 and acid volatile sulfides were measured. Second,
 additional data are needed from spiked-sediment
 bioassays to establish cause-effect relationships.
 Third,  additional  data  are needed  from  field
  studies in which very strong chemical gradients
  were  observed.  These  studies should include
measures of the toxicity and chemical contamina-
tion of bulk sediments and pore water.  They
would benefit from toxicity identification evalua-
tions to identify the causative agents responsible
for  the observed  biological  effects (Ankley,
1989). A number of large field surveys are under
way and being planned by NOAA and will lead to
additional data to be  included in the database.
Once these additional data are available,  they
could be entered into  the database and  used to
develop updated or new guidelines.
 14.5    REFERENCES

 Adams, W.J., R. A. Kimerle, and J. W. Barnett,
    Jr. In press. Sediment quality and aquatic
    life assessment.  Envir. Sci. and Technol.
 Ankley, G.  1989.  Sediment toxicity assessment
    through  evaluation of the toxicity of intersti-
    tial water.  Environmental Research Labpra-
    tory-Duluth. U.S. Environmental Protection
    Agency, Duluth, MN.  27 pp.
 Long, E. R.  1992.  Ranges  in chemical concen-
    trations  in sediments associated with adverse
    biological effects. Mar.  Pollu. Bull.  24 (1):
    38-45.
 Long, E.R., D.  MacDonald, and C. Caimcross.
     1992. Status and trends in toxicants and the
    potential for their biological effects in Tampa
     Bay, Florida.   NOAA  Tech. Memo.  NOS
     OMA 58. National Oceanic and Atmospheric
     Administration, Seattle, WA. 77 pp.
 Long, E.R., and R. Markel.  1992. An evaluation
     of the  extent  and  magnitude of  biological
     effects associated with chemical contaminants
     in San  Francisco Bay,  California.  NOAA
     Tech. Memo. NOS  OMA 64.  National Oce-
     anic and Atmospheric Administration, Seattle,
     WA. 86pp.
 Long, E.R., and L.G. Morgan. 1990.  The poten-
 ;    tial for biological effects of sediment-sorbed
     contaminants tested in the National Status and
     Trends Program. NOAA Tech. Memo. NOS
     OMA 62. National Oceanic and Atmospheric
     Administration, Seattle, WA.  175 pp.
  Lorenzato,  S.G.,   A.   J. Gunther,  and  J.  M.
     O'Connor.  1991.  Summary of a workshop
                                                                                            14-17'

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  Sediment Classification Methods Compendium
     concerning sediment quality assessment and
     development of sediment quality objectives.
     California State Water  Resources Control
     Board, Sacramento, CA.  32 pp.
 MacDonald, D.D.   1992.   Development  of an
     integrated approach to  the  assessment of
     sediment quality in Florida.   Prepared for
     Florida Department of Environmental Regula-
     tion.  MacDonald Environmental Services,
     Ltd.  Ladysmith, British Columbia. 114 pp.
 MacDonald, D.D., and S.L. Smith. 1991.  A
     discussion paper on the derivation and  use
     of Canadian sediment quality guidelines for
     the protection of freshwater and marine
     aquatic life. Prepared for Canadian Council
     of Ministers of the Environment.  Environ-
     ment Canada. Ottawa.
 MacDonald, D-D., S.L. Smith, M.P. Wong, and P.
     Mudroch. 1991.  The development of Canadi-
     an marine environmental quality guidelines.
     Report  prepared  for  the Interdepartment
     Working  Group on Marine Environmental
     Quality Guidelines and the Canadian Council
     of Ministers of the Environment. Environment
     Canada.  Ottawa, Canada.  50 pp.  .
 Mannheim, F.T., and J.C. Hathaway.  1991. Pol-
    luted sediments in Boston Harbor-Massachus-
  ,  etts Bay:   Progress  report on  the  Boston
    Harbor data management file.  U.S. Dept. of
    the Interior, Geological Survey Open File
    Report 91-331.   USGS, Woods Hole, MA.
    18pp.
Soule, D.F., M.  Oguri, and B.H. Jones.   1991.
    Marine Studies of San Pedro Bay, California,
    Part 20F. The. marine environment of Marina
    Del Rey. October 1989 to September 1990.
    Submitted  to Department  of  Beaches and
    Harbors, County of Los Angeles. University
    of Southern  California, Los Angeles, CA.
    206pp.
Squibb, K. S., J. M.  O'Connor, and T.J.  Kneip.
    1991.  New York/New Jersey Harbor Estuary
    Program. Module 3.1:  Toxics characteriza-
    tion report. Prepared for U.S. Environmental
    Protection Agency, Region 2. NYU Medical
    Center, Tuxedo, NY. 65 pp.
USEPA/SAB.  1989.  Evaluation of the apparent
    effects threshold (AET) approach to assessing
    sediment quality.  U.S. Environmental Protec-
    tion Agency Science Advisory Board.  Report
    of the Sediment Criteria Subcommittee. U.S.
    EPA SAB-EETFC-89-027.  16pp.
14-18

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