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
-------
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
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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.
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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
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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)
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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
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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
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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
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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).
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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.
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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
-------
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
American Public Health Association, American
Water Works Association, and the Water
Pollution Control Federation. 1989. Standard
methods for the examination of water and
wastewater. 17th edition. APHA, Washington,
"•. DC. ' " : ' •' ' ' ; '
ASTM. 1990. Standard guide, for collection,
storage, characterization, and manipulation of
sediments for lexicological testing. American
Society of Testing and Materials. ASTM
Designation E1391-90. • >
Baudo, R. 1990. Sediment sampling,, mapping,
and data analysis, pp. 15-60. In: R. Baudo,
J.P. Giesy, H. Muntau, Sediments: Chemis-
try and Toxicity of In-Place Pollutants.
Buchanan, J.B. 1984. Sediment analysis. In:
Methods for the Study of Marine/Benthos.
IBP Handbook No. 16, 2nd edition. NA.
Holme and A.D. Mclntyre (eds.). Blackwell
Scientific Publications, Oxford, UK.
Burton, G.A., Jr. 1989. Quality assurance project
plan for "A multi-assay/multi4est site evalua-
tion of sediment toxicity." U.S. Environmen-
tal Protection Agency, Great Lakes National
Program Office.
Crecelius, E. 1990. Quality assurance project
plan for "Assessment and remediation of
contaminated sediments (ARCS) assistance."
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Sediment Classification Methods Compendium
Prepared by Battelle/Marine Sciences Labora-
tory for the Environmental Protection Agency,
Great Lakes National Program Office.
Delbert, S.B, and T.H. Starks. 1985. Project
summary sediment sampling quality assurance
user's guide. EPA-600/4-85-048. Prepared
by Environmental Research Center, University
of Nevada, Las Vegas, for U.S. Environmen-
tal Protection Agency, Office of Research and
Development, Environmental Monitoring
Systems Laboratory, Las Vegas, Nevada, May
1985.
Di Toro, D.M., J.D. Mahony, DJ. Hansen, KJ.
Scott, W. Burry, M.B. Hicks, S.M. Mayr, and
M.S. Redmond. 1990. Toxicity of cadmium
in sediments: The role of acid volatile sul-
fides. In: Environmental Toxicology and
Chemistry. In press.
Downing, LA. and F.H. Rigler, eds. 1984. A
manual on methods for the assessment of
secondary productivity in fresh waters.
Second edition Blackwell Scientific Publica-
tions. .
EPRI. 1986. Speciation of selenium and arsenic
in natural waters and sediments. Vol. 2.
Prepared by Battelle Pacific Northwest Labo-
ratories for the Electrical Power Research
Institute. EPRIEA-4641.
Folk, R.L. 1968. Petrology of sedimentary rocks.
University of Texas, Austin, TX.
Higgins, T.R. 1988. Techniques for reducing the
costs of sediment evaluation. Tech. Note
EEDP-06-2. U.S. Army Engineer Waterways
Experiment Station, Vicksburg, Mississippi.
Holland, A.F., ed. 1990. Environmental monitor-
ing and assessment program, near coastal
program plan for 1990: Estuaries. Environ-
mental Research Laboratory, U.S. Environ-
mental Protection Agency.
IJC. 1988. Procedures for the assessment of
contaminated sediment problems in the great
lakes. Report to the Water Quality Board of
the International Joint Commission by the
Sediment Subcommittee and its Assessment
Work Group, International Joint Commission,
Windsor, Ontario Canada, December, 1988.
l£arr, J.R., KD. Fausch, P.L. Angermeier, P.R.
Yant, and I.J. Schlosser. 1986. Assessing
biological integrity in running waters: A
method and its rationale. Illinois Natural
History Survey, Special Publication 5. Sprin-
gfield, IL.
Klemm, DJ., PA. Lewis, F. Fiulk, and J.M.
Lazorchak. 1990. Macroinvertebrate fieldand
laboratory methods for evaluating the biologi-
cal integrity of surface waters. U.S. Environ-
. mental Protection Agency, Office x>f Research
and Development, EPA/600/4-90/030.
Lee, H. n, B.L. Boese, J. Pellitier, M. Winsor,
' D.T. Specht, and.R.C. Randall. 1989. Guid-
ance manual: Bedded sediment bioaccumulat-
ion tests. U.S. Environmental Protection
Agency, Pacific Ecosystems Branch, Bioacc-
umulation Team, Newport, Oregon. EPA 60-
O/x-89-302. ERLN-N111.
Leonard, E. 1991. Standard operating procedures,
for total organic carbon analysis of sediment
samples, U.S. Environmental Protection
Agency, Office of Research and Development,
Environmental Research Laboratory, Duluth,
Minnesota.
NOAA. 1989. Standard analytical procedure Of
the noaa national analytical facility. 2nd ed.
NOAA Tech. Mem. NMFC F/NWC-92,
1985-1986.
Ocean Dredged Material Disposal Program. 1991.
Testing and reporting requirements for ocean
disposal of dredged material off southern
California under marine protection, research
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
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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.,
-------
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
-------
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.
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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.
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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.
3-10
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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
-------
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..
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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-
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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
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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.
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Toxicol. Chem. 7: 1013-1020.
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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
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Wiederholm, T., A.M. Wiederholm, and G.
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Sediment Classification Methods Compendium
Milbrink. 1987. Bulk sediment bioassays with
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mental Program Office, Washington, DC.
4-10
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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;
-------
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
-------
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
-------
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
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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.
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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.
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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-
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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
-------
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
-------
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
7-3
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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-
7-4
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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
7-5
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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
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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. •'..,'
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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
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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
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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-
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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
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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
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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).
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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
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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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^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
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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
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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
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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
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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.
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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
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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
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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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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
-------
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
-------
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
-------
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
-------
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
-------
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);
-------
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.
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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.
20-15
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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
-------
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
-------
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
-------
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
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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
-------
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.
22-20
-------
-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
-------
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
-------
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:
12-3
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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
-------
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
-------
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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23-4
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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
13-13
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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
-------
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;
-------
Sediment Classification Methods Compendium
f
f
E
e
-5
a
I
S
g
e
+
e
14-2
-------
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
-------
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
-------
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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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
-------
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'
-------
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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