United States June 1989
Environmental Protection
Agency
Watershed Protection Division
Final
Classification
Methods Compendium
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Draft Final Report
SEDIMENT CLASSIFICATION
METHODS COMPENDIUM
by
U.S. Environmental Protection Agency
Portions of this document were prepared by
Tetra Tech, Inc., under the direction of
Michael Kravitz, U.S. EPA Work Assignment Manager
June 1989
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CONTENTS
Page
LIST OF FIGURES ix
LIST OF TABLES x
ACKNOWLEDGMENTS xi
CHAPTER 1. INTRODUCTION 1-1
1.0 BACKGROUND 1-1
2.0 OBJECTIVE 1-2
3.0 OVERVIEW 1-2
CHAPTER 2. BULK SEDIMENT TOXICITY TEST APPROACH 2-1
1.0 SPECIFIC APPLICATIONS 2-1
1.1 Current Use 2-1
1.2 Potential Use 2-2
2.0 DESCRIPTION 2-3
2.1 Description of Method 2-3
2.2 Applicability of Method to Human Health, Aquatic Life,
««. MI* i *4i •: £« a..«*„,.* -• ~- n 7
wi niiuiiic r i \j lev. u i un C. I
2.3 Ability of Method to Generate Numerical Criteria for
Specific Chemicals 2-7
3.0 USEFULNESS 2-8
3.1 Environmental Applicability 2-8
3.2 General Advantages and Limitations 2-10
4.0 STATUS 2-13
4.1 Extent of Use 2-13
4.2 Extent to Which Approach Has Been Field-Validated 2-13
4.3 Reasons for Limited Use 2-13
4.4 Outlook for Future Use and Amount of Development Yet
Needed 2-13
5.0 REFERENCES 2-14
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CHAPTER 3. SPIKED-SEDIMENT TOXICITY TEST APPROACH 3-1
1.0 SPECIFIC APPLICATIONS 3-1
1.1 Current Use 3-1
1.2 Potential Use 3-2
2.0 DESCRIPTION 3-2
2.1 Description of Method 3-2
2.2 Applicability of Method to Human Health, Aquatic Life,
or Wildlife Protection 3-6
2.3 Ability of Method to Generate Numerical Criteria for
Specific Chemicals 3-7
3.0 USEFULNESS 3-8
3.1 Environmental Applicability 3-8
3.2 General Advantages and Limitations 3-10
4.0 STATUS 3-13
4.1 Extent of Use 3-13
4.2 Extent to Which Approach Has Been Field-Validated 3-13
4.3 Reasons for Limited Use 3-14
4.4 Outlook for Future Use and Amount of Development Yet
Needed 3-14
5.0 REFERENCES 3-14
CHAPTER 4. INTERSTITIAL WATER TOXICITY APPROACH 4-1
1.0 SPECIFIC APPLICATIONS 4-1
1.1 Current Use 4-1
1.2 Potential Use 4-2
2.0 DESCRIPTION 4-2
2.1 Description of Method 4-2
2.2 Applicability of Method to Human Health, Aquatic Life,
or Wildlife Protection 4-16
2.3 Ability of Method to Generate Numerical Criteria for
Specific Chemicals 4-16
3.0 USEFULNESS 4-17
3.1 Environmental Applicability 4-17
3.2 General Advantages and Limitations 4-19
11
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4.0 STATUS 4-21
4.1 Extent of Use 4-21
4.2 Extent to Which Approach Has Been Field-Validated 4-22
4.3 Reasons for Limited Use 4-22
4.4 Outlook for Future Use and Amount of Development Yet
Needed 4-22
5.0 REFERENCES 4-23
CHAPTER 5. EQUILIBRIUM PARTITIONING APPROACH 5-1
1.0 SPECIFIC APPLICATIONS 5-1
1.1 Current Use 5-2
1.2 Potential Use 5-3
2.0 DESCRIPTION 5-4
2.1 Description of Method 5-4
2.2 Applicability of Method to Human Health, Aquatic Life,
or Wildlife Protection 5-7
2.3 Ability of Method to Generate Numerical Criteria for
Specific Chemicals 5-8
3.0 USEFULNESS 5-9
3.1 Environmental Applicability 5-9
3.2 General Advantages and Limitations 5-11
4.0 STATUS 5-15
4.1 Extent of Use 5-16
4.2 Extent to Which Approach Has Been Field-Validated 5-16
4.3 Reasons for Limited Use 5-17
4.4 Outlook for Future Use and Amount of Development Yet
Needed 5-17
5.0 DOCUMENTS 5-18
CHAPTER 6. TISSUE RESIDUE APPROACH 6-1
1.0 SPECIFIC APPLICATIONS 6-2
1.1 Current Use 6-2
1.2 Potential Use 6-2
2.0 DESCRIPTION - 6-3
2.1 Description of Method 6-3
2.2 Applicability of Method to Human Health, Aquatic Life,
or Wildlife Protection 6-9
iv
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2.3 Ability of Method to Generate Numerical Criteria for
Specific Chemicals 6-10
3.0 USEFULNESS 6-10
3.1 Environmental Applicability 6-10
3.2 General Advantages and Limitations 6-14
4.0 STATUS ' 6-17
4.1 Extent of Use 6-17
4.2 Extent to Which Approach Has Been Field-Validated 6-17
4.3 Reasons for Limited Use 6-18
4.4 Outlook for Future Use and Amount of Development Yet
Needed 6-18
5.0 REFERENCES 6-19
CHAPTER 7. FRESHWATER BENTHIC MACROINVERTEBRATE COMMUNITY STRUCTURE
AND FUNCTION 7-1
1.0 SPECIFIC APPLICATIONS . 7-2
1.1 Current Use 7-2
1.2 Potential Use 7-5
2.0 DESCRIPTION ' 7-6
2.1 Description of Method 7-6
2.2 Applicability of Method to Human Health, Aquatic Life,
or Wildlife Protection 7-28
2.3 Ability of Method to Generate Numerical Criteria for
Specific Chemicals 7-28
3.0 USEFULNESS 7-28
3.1 Environmental Applicability 7-28
3.2 General Advantages and Limitations 7-30
4.0 STATUS 7-35
4.1 Extent of Use 7-35
4.2 Extent to Which Approach Has Been Field-Validated 7-35
4.3 Reasons for Limited Use 7-36
4.4 Outlook for Future Use and Amount of Development Yet
Needed 7-36
5.0 REFERENCES 7-36
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CHAPTER 8. MARINE BENTHIC COMMUNITY STRUCTURE ASSESSMENT 8-1
1.0 SPECIFIC APPLICATIONS 8-2
1.1 -Current Use 8-3
1.2 Potential Use 8-7
2.0 DESCRIPTION .8-8
2.1 Description of Method 8-8
2.2 Applicability of Method to Human Health, Aquatic Life,
or Wildlife Protection 8-20
2.3 Ability of Method to Generate Numerical Criteria for
Specific Chemicals 8-21
3.0 USEFULNESS 8-21
3.1 Environmental Applicability 8-22
3.2 General Advantages and Limitations 8-26
4.0 STATUS 8-31
4.1 Extent of Use 8-31
4.2 Extent to Which Approach Has Been Field-Validated 8-32
4.3 Reasons for Limited Use 8-32
4.4 Outlook for Future Use and Amount of Development Yet
Needed . 8-32
5.0 REFERENCES 8-34
CHAPTER 9. SEDIMENT QUALITY TRIAD APPROACH 9-1
1.0 SPECIFIC APPLICATIONS 9-1
1.1 Current Use 9-1
1.2 Potential Use 9-2
2.0 DESCRIPTION 9-2
2.1 Description of Method 9-2
2.2 Applicability of Method to Human Health, Aquatic Life,
or Wildlife Protection 9-15
2.3 Ability of Method to Generate Numerical Criteria for
Specific Chemicals 9-16
3.0 USEFULNESS 9-16
3.1 Environmental Applicability 9-16
3.2 General Advantages and Limitations 9-20
VI
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4.0 STATUS 9-24
4.1. -€xtent of Use 9-24
4.2 Extent to Which Approach Has Been Field-Validated 9-24
4.3 Reasons for Limited Use 9-24
4.4 Outlook for Future Use and Amount of Development Yet
Needed 9-25
5.0 REFERENCES " 9-25
CHAPTER 10. APPARENT EFFECTS THRESHOLD APPROACH 10-1
1.0 SPECIFIC APPLICATIONS 10-1
1.1 Current Use 10-1
1.2 Potential Use 10-4
2.0 DESCRIPTION ' 10-5
2.1 Description of Method 10-5
2.2 Applicability of Method to Human Health, Aquatic Life,
or Wildlife Protection 10-16
2.3 Ability of Method to Generate Numerical Criteria for
Specific Chemicals 10-16
3.0 USEFULNESS 10-17
3.1 Environmental Applicability 10-17
3.2 General Advantages and Limitations 10-22
4.0 STATUS 10-33
4.1 Extent of Use 10-33
4.2 Extent to Which Approach Has Been Field-Validated 10-35
4.3 Reasons for Limited Use 10-37
4.4 Outlook for Future Use and Amount of Development Yet
Needed 10-37
5.0 REFERENCES 10-38
CHAPTER 11. A SUMMARY OF THE SEDIMENT ASSESSMENT STRATEGY RECOMMENDED
BY THE INTERNATIONAL JOINT COMMISSION 11-1
1.0 SPECIFIC APPLICATIONS 11-1
1.1 Current Use 11-1
1.2 Potential Use 11-2
vn
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2.0 DESCRIPTION . - 11-2
2.1 Description of Method 11-2
2.2 Applicability of Method to Human Health, Aquatic Life,
or Wildlife Protection 11-14
2.3 Ability of Method to Generate Numerical Criteria for
Specific Chemicals 11-14
3.0 USEFULNESS 11-15
3.1 Environmental Applicability 11-15
3.2 General Advantages and Limitations 11-16
4.0 STATUS 11-19
4.1 Extent of Use 11-19
4.2 Extent to Which Approach Has Been Field-Validated 11-19
4.3 Reasons for Limited Use 11-20
4.4 Outlook for Future Use and Amount of Development Yet
Needed 11-20
5.0 REFERENCES 11-20
VI 11
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FIGURES
Number Page
4-1 Overview of the Phase I toxicity characterization process 4-7
9-1 Conceptual model of the Sediment Quality Triad 9-3
9-2 Triaxial plots of eight possible outcomes for Sediment
Quality Triad results 9-14
10-1 The AET approach applied to sediments tested for lead and
i-methylphenol concentrations and toxicity response during
bioassays 10-7
10-2 Measures of reliability (sensitivity -and efficiency) 10-31
IX
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TABLES
Number Page
1-1 Sediment quality assessment methods 1-3
1-2 Structure of sediment quality assessment method chapters 1-6
4-1 Phase I characterization results and suspect toxicant
classification for two effluents 4-12
9-1 Current uses of the Sediment Quality Triad approach 9-4
9-2 Possible conclusions provided by using the Sediment Quality
Triad approach 9-6
9-3 Example analytes and detection limits for use in the
chemistry component of Triad 9-9
9-4 Possible static sediment bioassays 9-11
10-1 Selected chemicals for which AET have been developed in
Puget Sound 10-18
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ACKNOWLEDGMENTS
This compendium was prepared by the U.S. Environmental Protection Agency,
Sediment Oversight Technical Committee. Chaired by Dr. Elizabeth Southerland
of the Office of Water Regulations and Standards, the committee has represen-
tation from a number of Program Offices in Headquarters and the Regions.
The methods represented here were written by the following authors (also
listed at the beginning of their respective chapters):
• Gerald Ankley, Anthony R. Carlson, Phillip M. Cook, Wayne S.
Davis, Catherine Krueger, Janet Lamberson, Henry Lee II,
Richard C. Swartz, Nelson Thomas, and Christopher S. Zarba
(U.S. EPA)
• Gordon R. Bilyard, Gary M. Braun, and Betsy Day (Tetra Tech,
Inc.)
• Peter M. Chapman (E.V.S. Consultants, Ltd.)
• Philippe Ross (Illinois Natural History Survey)
• Joyce E. Lathrop (Stream Assessments Company).
Critical reviews of portions of this document were provided by the following
U.S. EPA persons: Gerald Ankley, Carol Bass, Dave Cowgill, Philip Crocker,
Shannon Cunniff, Kim Devonald, Cynthia Fuller, Ray Hall, David Hansen,
Nicholas Loux, Menchu Martinez, Brian Melzian, Ossie Meyn, James Neiheisel,
Dave Redford, Greg Schweer, Richard Swartz, Nelson Thomas, Mark Tuchman,
Gerald Walsh, Al Wastler, Howard Zar, and Chris Zarba.
Assistance in preparation and production of the compendium was provided
by Tetra Tech, Inc. in partial fulfillment of EPA Contract No. 68-03-3475.
Dr. Karen Summers is Tetra Tech's Program Manager. Dr. Leslie Williams
served as Work Assignment Manager. Ms. Marcy Brooks-McAuliffe managed
editorial review and document production, and was assisted by Ms. Vicki
Fagerness, Dr. Jean Jacoby, Dr. Gary Pascoe, and Ms. Betsy Day. Ms. Mary
Bauchtel, also of Tetra Tech, provided technical assistance to the U.S. EPA
Work Assignment Manager, Michael Kravitz.
XI
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Introduction
CHAPTER 1. INTRODUCTION
1.0 BACKGROUND
Sediment management issues are of importance to many programs within the
U.S. Environmental Protection Agency (EPA). The ability to assess sediment
quality in a technically reliable and legally defensible manner is necessary
for effective sediment management. In the summer of 1988, the U.S. EPA
Office of Water Regulations and Standards (OWRS) formed two committees to
identify, coordinate, and provide guidance on activities relating to the
assessment and management of sediments contaminated with toxic chemicals: a
Sediment Oversight Technical Committee and a Sediment Oversight Steering
Committee. The goal of these committees is to facilitate decisions made at
various stages in the management process such as assessing sediment contam-
ination, deciding on the need for and type of management action, and
evaluating types of remediation. This document, prepared by the Sediment
Oversight Technical Committee, describes the various methods used to assess
sediment quality.
A number of approaches can be used to assess sediment contamination.
Many past approaches were based on comparing chemical concentrations in
contaminated areas to those in reference areas, and did not directly consider
biological effects. More recent approaches to evaluating sediment quality
have focused on determining relationships between sediment contaminant
concentrations and adverse biological effects. These approaches may be
applied to a variety of regulatory decisions, including identification of
problem areas, establishment of cleanup goals, development of discharge and
dumping permit criteria, and determination of monitoring requirements.
1-1
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Introduction
2.0 OBJECTIVE
This compendium is essentially an "encyclopedia" of methods that are
used to assess chemically contaminated sediments. It contains a description
of each method, associated advantages and limitations, and existing
applications. It is intended to serve as a common frame of reference to.
answer the "how clean is clean" question for particular sediments (i.e.,
does sediment contamination exist to a degree that warrants the evaluation
of need for further action?). Some of the methods in this compendium can
also be used as part of subsequent regulatory or remedial actions. It
should be pointed out that these methods are not at an equal stage of
development, and certain ones (or combinations) are more aopropriate for
specific management actions than are others. This document is not meant to
provide guidance on which method(s) to apply for specific situations, nor
how they can be used together as part of. a decision-making framework. Such
guidance will be forthcoming and will likely include both chemical and
biological methods in a tiered type of framework.
3.0 OVERVIEW
The sediment quality assessment methods described in this report can be
classified into two basic types: numeric or descriptive (Table 1-1).
Numeric methods are chemical-specific and can be used to generate numerical
sediment quality criteria. Descriptive methods are not chemical-specific
and cannot be used alone to generate numerical sediment quality criteria for
particular chemicals.
In addition, some of the approaches described in this report comprise
at least two methods *nd can be classified as combination approaches (Table
1-1). 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,
1-2
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TABLE i-i. S-3IMENT QUALITY ASSESSMENT METHODS
•M«nod (Chaoter)
Numeric :escnotive Conoination
Conceot
2ulk Sediment Toxicity (2.0)
Sf.iked-Sediment Toxicity (3.0)
Interstitial Water Toxicity (4.0) •
£oui 1 lorium Partitioning (5.0)
issue
(5.0)
3entmc Csmmumty
Structure (7.0)
"arine Sentnic Ccmtmnity Structurs (3.0)
Sediment Quality Triad (9.0)
iooarent Effects Thresnold (10.0)
Test organism are exposed to sediments which 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.
Oose-response relationships are established by
exposing test organisms to sediments that have been
spiked with known amounts of chemicals or mixtures
of chemicals.
Toxicity of interstitial water is quantified and
identification evaluation procedures are applied to
identify and quantify cremical comoonents resoons-
tble for sediment toxicity. The procedures are
implemented in three anases :o characterize
interstitial water toxtcity. icenttfy 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 tnat would corresoond to an
interstitial water concentration equivalent to the
U.S. EPA water quality criterion for the con-
taninant.
Safe sediment concentrations of specific chemicals
are established by determining the sediment
chemical concentration that will result in
acceptable tissue residues. Methods to derive
unacceptable tissue residues are based on chronic
•ater quality criteria and bioconcentration
factors, chronic dose-response exoenments or field
correlations, and human health risk levels from the
consumption of fresnwater fish or seafood.
Environmental degradation is measured by evaluating
alterations In fresnwater benthic community
structure.
Environmental degradation is measured by evaluating
alterations in marine Oenthic cornnunity structure.
Sediment chemical contamination, sediment toxicity,
and benthic infauna comnunity structure are
measured on the same sediment. Correspondence
between sediment cnermstry, toxicity. and biological
effects Is used to determine sediment concentrations
that discriminate conditions of minimal, uncertain,
and major biological effects.
An A£T is the sediment concentration of a contami-
nant above «nich statistically significant
biological effects (e.g., ampmpod mortality in
bioassays. depressions in the abundance of benthic
infauna) would always be expected. ACT. values are
empirically derived from paired field data for
sediment chemistry and a range of biological
effects indicators.
1-3
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"-3t£ 1-1. (Continued)
>:er-aticnal Joint Camtission (11.0)d
Contaminated sediments are a5'cased in two stages:
1) an initial assessment tha: is based on macro-
zoooenthic community structure and concentrations
of contaminants in sediments and biological
tissues, and 2] a detailed assessment that is based
an a onased samoling of the onysical. chemical, and
Biological ascects of the sediment, including
laboratory toxicity oioassays.
' "he IJC aooroacn is an examole of a sequential aooroacn. or "strategy" conOining a numocr of methods for the ourpose of
assessing contaminated sediments in the Great Lakes.
' 1-4
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Introduction
the Triad approach has been used in a descriptive manner to identify
contaminated sediments. However, it has also been used to generate criteria
for several chemical contaminants. The International Joint Commission (IJC)
approach presented at the end of this document (Chapter 11) is an example of
a sequential approach, or "strategy," combining a number of methods for the
purpose of assessing contaminated sediments in the Great Lakes.
Each sediment quality assessment method is presented as a separate
chapter. Each chapter is structured identically, as indicated in Table 1-2,
to facilitate comparisons among the various methods. Authors are listed at
the beginning of each chapter. A general description, application,
usefulness, and status of the method is then presented. A list of references
cited is provided at the end of each chapter.
1-5
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TABLE 1-2. STRUCTURE OF SEDIMENT QUALITY ASSESSMENT METHOD CHAPTERS
1.0 SPECIFIC APPLICATIONS
1.1 CURRENT USE
1.2 POTENTIAL USE
2.0 DESCRIPTION
2.1 DESCRIPTION OF METHOD
2.1.1 Objectives and Assumptions
2.1.2 Level of Effort
2.1.2.1 Type of Sampling Required
2.1.2.2 Methods
2.1.2.3 Types of Data Required
2.1.2.4 Necessary Hardware and Skills
2.1.3 Adequacy of Documentation
2.2 APPLICABILITY OF METHOD TO HUMAN HEALTH, AQUATIC LIFE, OR WILDLIFE
PROTECTION
2.3 ABILITY OF METHOD TO GENERATE NUMERICAL CRITERIA FOR SPECIFIC
CHEMICALS
3.0 USEFULNESS
3.1 ENVIRONMENTAL APPLICABILITY
3.1.1 Suitability for Different Sediment Types
3.1.2 Suitability for Different Chemicals or Classes of Chemicals
3.1.3 Suitability for Predicting Effects on Different Organisms
3.1.4 Suitability for In-Place Pollutant Control
3.1.5 Suitability for Source Control
3.1.6 Suitability for Disposal Applications
3.2 GENERAL ADVANTAGES AND LIMITATIONS
3.2.1 Ease of Use
3.2.2 Relative Cost
3.2.3 Tendency to be Conservative
3.2.4 Level of Acceptance
3.2.5 Ability to be Implemented by Laboratories with Typical
Equipment and Handling Facilities
3.2.6 Level of Effort Required to Generate Results
3.2.7 Degree to Which Results Lend Themselves to Interpretation
3.2.8 Degree of Environmental Applicability
3.2.9 Degree of Accuracy and Precision
1-6
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TABLE 1-2. (Continued)
4.0 STATUS
4.1 EXTENT OF USE
4.2 EXTENT TO WHICH APPROACH HAS BEEN FIELD-VALIDATED
4.3 REASONS FOR LIMITED USE
4.4 OUTLOOK FOR FUTURE USE AND AMOUNT OF DEVELOPMENT YET NEEDED
5.0 REFERENCES
1-7
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Bulk Sediment Toxicity
CHAPTER 2. BULK SEDIMENT TOXICITY TEST APPROACH
Nelson Thomas
U.S. Environmental Protection Agency
Environmental Research Laboratory
6201 Congdon Blvd.
Ouluth, MN 55804
(218) 720-5702
Janet Lamberson and Richard C. Swartz
U.S. Environmental Protection Agency
Environmental Research Laboratory - N. Pacific Division
Hatfield Marine Science Center
Newport, OR 97365
(503) 867-4031
In the bulk sediment toxicity approach, test organisms are exposed in
the laboratory to sediments that were collected in the field. A specific
biological endpoint is used to assess the response of the organisms to the
sediments (i.e., to measure sediment toxicity). The bulk sediment toxicity
approach is a descriptive method and cannot be used by itself to generate
sediment quality criteria.
1.0 SPECIFIC APPLICATIONS
1.1 Current Use
Sediment toxicity testing has been applied in the following ways in
dredged material disposal permit and other regulatory programs (U.S. EPA and
U.S. Army COE 1977).
• To determine potential biological hazards of dredged material
intended for disposal in an aquatic environment
• 2-1
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Bulk Sediment Toxicity
• To evaluate the effectiveness of various dredged material
management actions
• To indicate spatial distribution of toxicity in contaminated
areas, relative degree of toxicity, and changes in toxicity
along a gradient of pollution or with respect to distance
from pollutant sources (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 et al. 1986)
• To reveal "hot spots" of contaminated sediment for further
investigation (Chapman 1986a)
• To rank sediments based on toxicity to benthic organisms and
to define boundaries of small or large problem areas for
cleanup of contaminated sediment.
Bulk sediment toxicity testing integrates interactions among complex
mixtures of contaminants that may be present in the field. Many classes of
chemical contaminants, including metal s, PAHs, PCBs, dioxins, and chlorinated
pesticides, can contribute to toxicity in effluents and sediments (Chapman
et al. 1982). The bulk sediment toxicity test measures the total toxic
effect of all contaminants, regardless of physical and chemical composition.
1.2 Potential Use
By itself, bulk sediment toxicity testing cannot generate chemical-
specific toxic effects data but does determine toxicity. However, used in
conjunction with toxicity identification evaluation procedures such as those
described in Chapter 4, 9, and 10, bulk sediment toxicity testing could help
identify causal toxicants. The procedure must be combined with other
2-2
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Bulk Sediment Toxicity
methods of estimating sediment quality in order to generate sediment quality
criteria, such as the Sediment Quality Triad (Triad) (Chapman 1986b;
Chapman et al. 1987; see Chapter 9), and the Apparent Effects Threshold
(AET) approach (Tetra Tech 1986; PTI 1988; see Chapter 10). Bulk sediment
toxicity will be most valuable in verification of other methods used to
develop sediment quality criteria. This method is also useful in determining
acceptability for disposal options.
2.0 DESCRIPTION
2.1 Description of Method
The toxicological approach involves exposing test organisms to
sediments. The chemical composition 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 respiration rate). Results are then compared with control
and reference sediment results to estimate sediment toxicity.
2.1.1 Objectives and Assumptions--
The objective of this approach 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, antagonism) among
chemicals in the field or in .dredged materials can be predicted from
laboratory results without measuring total or bioavailable concentrations of
potentially hundreds of contaminants in the test sediment (Swartz et al. in
press) and without a priori knowledge of specific pathways of interaction
between sediments and test organisms (Kemp and Swartz in press). In that
one of the strengths of this test, is to integrate effects of all contami-
2-3
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Bulk Sediment Toxicity
nants, the effect of individual contaminants cannot be determined, therefore
limiting its use in source control. The method can be used ft>r all classes
of sediments and any chemical contaminant, but not to answer cause-and-effect
questions.
2.1.2 Level of Effort--
Implementation of this procedure requires a moderate amount of
laboratory effort. A variety of toxicity test procedures (see Section
2.1.2.2) have been developed and are generally fairly straightforward and
well documented.
2.1.2.1 Type of Sampling Required—It is recommended that bulk
sediments be collected for analysis of total solids, grain size, acid
volatile sulfide, and total and dissolved organic carbon. Bulk and
interstitial concentrations of chemicals of interest in the test sediment
can be determined in subsamples of the sediment added to the toxicity test
chambers to enhance the interpretation of toxicity results. However,
methods for sampling interstitial water have not been standardized.
Sediment variables such as pH and Eh should also be monitored.
2.1.2.2 Methods—There currently are several bulk sediment toxicity
tests under ballot by the American Society of Testing Materials (ASTM). The
most commonly used of these partial life cycle tests feature freshwater
chironomid species (Chironomus tentans, Chironomus riparius), the fresh-
water/estuarine amphipod Hyalella azteca, and the marine amphipod Rhepoxynius
abronius. Brief generalized descriptions of these tests are given below.
Bulk sediment toxicity tests 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 commercial sources, and both species are easily cultured in
2-4
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Bulk Sediment Toxicity
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 overlying water in various ratios (e.g.,
6 water:! sediment; Giesy et al. 1988). The length of the test als~b 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 (depending upon temperature) in around 30 days for
C. riparius and at 20-25 d for C. tentans. More detailed descriptions of
toxicity test procedures with C. riparius and C. tentans are given by Adams
et al. (1985), Nebeker et al. (1984), Giesy et al. (1988), and Ingersoll and
Nelson (1989).
Partial life cycle toxicity tests with the freshwater/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 easily cultured in a
laboratory setting. Toxicity tests are initiated by adding juveniles <7 days
old to test chambers that contain bulk sediment with overlying water in
various ratios (e.g., 4 water:! sediment; Ingersoll and Nelson 1989). The
length of the test can range from <10 days (short-term partial life cycle) to
30 days (long-term partial life cycle) (Nebeker et al. 1984; Ingersoll and
Nelson 1989). Depending upon the length of the test, biological endpoints
include survival, behavior, growth, and reproduction. More detailed
descriptions of toxicity test procedures are given by Nebeker et al. (1984),
Nebeker and Miller (1988), and Ingersoll and Nelson (1989).
Partial life cycle toxicity tests with the marine amphipod Rhepoxynius
abronius and bulk sediments have been used routinely for some time.
R. abronius and bulk sediments generally are collected from the field and
acclimated to laboratory conditions for some time (<14 days) before toxicity
testing. The tests are initiated by adding large immature and adult
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Bulk Sediment Toxicity
amphipods to test chambers that contain bulk sediment with overlying water
in various ratios. The length of the test generally is >10 days and the
biological responses monitored consist of behavioral effects (e.g.,
emergence from sediment) and mortality. More detailed descriptions of the
toxicity test procedures are given by Swartz et al. (1979, 1985), OeWitt et
al. (1988), and Robinson-et al. (in press). Other ASTM candidate species
for marine toxicity tests are Eohaustorius estuarius, Ampelisca abdita, and
Grandidierella japonica.
2.1.2.3 Types of Data Required—The physical and chemical data
described in Section 2.1.2.1 are needed to interpret the test results.
Biological data required, 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. Dilution
experiments in which uncontaminated sediment is added to test sediment
collected from the field can be used to calculate LC50 values, EC50 values,
no-effect concentrations, and lowest observable-effect concentrations.
However, standardized techniques with dilution (i.e., by sediment of similar
physical-chemical properties) have not been developed.
2.1.2.4 Necessary Hardware and Skills-- In general, only readily
available and inexpensive field and laboratory equipment is needed,
procedures are fairly simple and straightforward, and a minimum of training
is needed to detect endpoints through toxicity tests. Interpretation of the
toxicity (chemical and biological) data requires a higher degree of skill
and training. Chemical sampling methods are generally simple and routine,
although analyses of chemical samples requires specialized training and
equipment. Some biological effects tests also require specialized training,
handling, and facilities.
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Bulk Sediment Toxicity
2.1.3 Adequacy of Documentation--
Various sediment toxicity test procedures have been developed and are
well documented for testing field sediments (lamberson and Swartz 1988;
Swartz 1987). However, methods must be better standardized through use and
intercalibrated among laboratories, and most methods need better field
validation.
2.2 Applicability of Method to Human Health. Aquatic Life, or Wildlife
Protection
The bulk sediment toxicity. test approach is suitable only for protection
of aquatic life. Sediment toxicity test procedures incorporate a direct
measure of sediment biological effects, and can be used to predict biological
effects of contaminated sediments prior to approval of state or federal
permits. They can be used to assess the toxicity of sediments in the natural
environment and to predict the effects of these sediments on resident
aquatic life. Combined with other approaches, such as in the AET and the
Triad approaches, they can be used to establish sediment quality criteria.
2.3 Ability of Method to Generate Numerical Criteria for Specific Chemicals
The bulk sediment toxicity test approach cannot be used by itself to
generate sediment quality criteria, but must be combined with chemical
measurements and other data to generate information on effects of individual
contaminants. Both the Triad and the AET approaches rely on bulk sediment
toxicity data to derive numerical criteria. Bulk sediment toxicity tests in
conjunction with sediment quality criteria derived from Equilibrium
Partitioning (U.S. EPA 1980) (see also Chapter 5 herein) can also be used
in assessments of potentially contaminated sediments.
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Bulk Sediment Toxicity
3.0 USEFULNESS
3.1 Environmental Applicability
3.1.1 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
sediment, such as salinity or grain size, may limit the selection of
organisms that can be used for testing, and may also affect interpretation
of the data (Ott 1986; DeWitt in preparation). Appropriate controls for
sediment properties may be necessary to discriminate chemical toxicity from
conventional effects. In establishing sediment quality criteria, the
effects of toxic 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, and thus can be applied to any sediment. However,
normalization techniques are in the developmental stage (see Equilibrium
Partitioning, Chapter 5).
3.1.2 Suitability for Different Chemicals or Classes of Chemicals--
This is the only approach currently available that directly measures
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. in press). Bioaccumulative
chemicals can be evaluated if the length of the test is extended to ensure
adequate exposure of the test organism.
3.1.3 Suitability for Predicting Effects on Different Organisms--
Theore-tically, any organism can be used in sediment toxicity testing.
To protect a biological community and to predict the effects of'contaminated
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Bulk Sediment Toxicity
sediments on different organisms, test organisms should be selected on the
basis of their sensitivity to contaminants, their ability to withstand
laboratory handling, and their ability to survive in control and reference
treatments (Swartz et al. 1985a). 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 sensitivity to various
types of contaminants.
3.1.4 Suitability for In-Place Pollutant Control--
Sediment toxicity testing can be used directly to monitor in-place
pollution. As noted above in Section 1.1, sediment toxicity testing can be
used to determine the extent of the problem area, to monitor temporal and
spatial trends, to detect the presence of unsuspected "hot spots," to assess
the need for remedial actions, and to monitor changes in toxicity after
remedial action is taken. Such tests can also be used as a cost-effective
and rapid screening tool for in-place pollutant reconnaissance surveys, and
in a priori simulations of proposed remedial actions to test effectiveness of
capping or other remedial alternatives.
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" in the vicinity of 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 mass loading of chemicals that might be expected as a
result of source control. The biological effects of source control can be
represented, however, through the use of bulk sediment toxicity testing.
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Bulk Sediment Toxicity
3.1.6 Suitability for Disposal Applications--
Bulk sediment toxicity testing can be used in regulatory programs to
determine the toxicity of material prior to disposal. The potential hazard
to benthic organisms at the disposal site (which is determined by making
comparisons with the "reference" sediments collected near the disposal site)
can be predicted from laboratory toxicity test results. Sediment toxicity
tests can also be used to monitor conditions at the disposal site both
before and after a disposal operation.
3.2 General Advantages and Limitations
3.2.1 Ease of Use--
Most sediment toxicity test procedures are simple to use, requiring
limited expertise and standard inexpensive laboratory equipment. Only a
few sublethal effects tests require specialized training. Field sampling
requires only readily available equipment and standard procedures.
3.2.2 Relative Cost--
The cost of individual laboratory toxicity tests as well as field
sampling is low because of the limited expertise and inexpensive equipment
requirements. Costs generally range from S150 to S500 per replicate.
Laboratory sediment toxicity testing is a comparatively inexpensive and
cost-effective method of monitoring the field distribution of sediment
toxicity, because it integrates the effects of all toxic contaminants, does
not require individual chemical measurements, and does not require time-con-
suming analysis of benthic community structure.
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Bulk Sediment Toxicity
3.2.3 Tendency to be Conservative--
Sediment toxicity tests can be made as sensitive 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.
3.2.4 Level of Acceptance--
Bulk sediment toxicity testing is widely accepted by the scientific
and regulatory communities, and has been tested and contested in court.
Field sediment toxicity test results have been widely published in peer-
reviewed journals, and have been incorporated into other measures of
sediment quality such as the AET and the Triad approaches. Standard guides
for sediment toxicity testing are being developed by ASTM. Field sediment
toxicity testing is incorporated into most dredged material disposal
regulatory programs. Toxicity testing in general has long been the basis
for water quality criteria, dredged material testing, effluent testing, and
discharge monitoring.
3.2.5 Ability to be Implemented by Laboratories with Typical Equipment and
Handling Facilities--
Sediment toxicity test methods are easily implemented by laboratories
with typical equipment using inexpensive glassware and procedures requiring
little specialized training, although the interpretation of some sublethal
biological endpoints may require some degree of training and experience.
Field sediment sample collection procedures are routine.
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Bulk Sediment Toxicity
3.2.6 Level of Effort Required to Generate Results--
This procedure consists of field sampling and a laboratory toxicity
tests. Compared to an extensive survey of chemical concentrations or benthic
community structure analysis, the level of effort is relatively small.
3.2.7 Degree to Which Results Lend Themselves to Interpretation--
Biological responses to toxic sediment can be easily interpreted.
Generally data fit "pass-fail" criteria (i.e., the result is either above or
below a predetermined acceptance level) or the result is statistically
compared to control and reference results to determine whether there is a
toxic effect. Little "expert" guidance is required for interpretation of the
results.
3.2.8 Degree of Environmental Applicability--
As noted in Section 3.1, the sediment toxicity test approach is appli-
cable 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 experimentally addressed with appropriate uncontaminated controls.
3.2.9 Degree of Accuracy and Precision--
Since the sediment toxicity test is a laboratory-controlled experiment,
results have a high degree of accuracy, precision, and repeatability. The
procedure produces a direct biological response data set for individual
sediment samples.
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Bulk-Sediment Toxicity
4.0 STATUS
4.1 Extent of Use
Sediment toxicity tests are widely used in research and regulatory
programs in both marine and freshwater systems, as described in Section 1.1.
Sediment toxicity tests are also incorporated into evaluation of applications
for dredged material disposal permits, and are used to assess the toxicity
of sediments subject to regulatory decisions. Bulk sediment toxicity tests
are also used to investigate the mechanisms of sediment toxicity to benthic
organisms (Kemp and Swartz 1989).
4.2 Extent to Which Approach Has Seen Field-Validated
Field validation of bulk sediment toxicity testing includes several
publications in the peer-reviewed literature (Chapman 1986b; Plesha et al .
1988; Swartz et al. 1982, in press). As more data become available, results
can be.compared with available information 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. As noted in Section 2.1.3, better field validation of
predicted effects is needed.
4.3 Reasons for Limited Use
Bulk sediment toxicity testing has been widely used in research and
regulatory programs (see above).
4.4 Outlook for Future Use and Amount of Development Yet Needed
The outlook for future use of sediment toxicity tests is promising
where direct measurement of biological effects of toxicants in sediments is
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Bulk Sediment Toxicity
desired, especially where the effects of chemical interactions is of
interest. Development of biological testing methods should continue, and
more emphasis should be placed on the development of procedures to measure
chronic effects. Methods should be compared and standardized among labor-
atories, and results should be field-validated to establish their ability to
predict biological effects 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.
5.0 REFERENCES
Adams, W.J., R.A. Kimerle, and R.G. Mosher. 1985. Aquatic safety assessment
of chemicals sorbed to sediments, pp. 429-453. In: Aquatic Toxicology and
.Hazard Assessment: Proceedings of the Seventh Annual Symposium, ASTM STP
854. R.O. Cardwell, R. Purdy, and R.C. Bahner (Eds). American Society for
Testing and Materials, Philadelphia, PA.
Chapman, P.M. 1986a. Sediment bioassay tests provide data necessary for
assessment and regulation, pp. 178-197. In: Proceedings of the Eleventh
Annual Aquatic Toxicology Workshop, Nov. 13-15, 1984. G.H. Green and K.I.
Woodward (Eds). Vancouver, Canada. Tech. Rpt. 1480. Fish. Aquat. Sci.,
Chapman, P.M. 1986b. Sediment quality criteria from the sediment quality
triad: an example. Environ. Toxicol. Chem. 5:957-964.
Chapman, P.M., G.A. Vigers, M.A. Farrell, R.N. Dexter, E.A. Quinlan, R.M.
Kocan, and M. Landolt. 1982. Survey of biological effects of toxicants
upon Puget Sound biota. 1. Broad-scale toxicity survey. NOAA Technical
Memorandum OMPA-25, National Oceanic and Atmospheric Administration, Boulder
CO.
Chapman, P.M., R.N. Dexter, and E.R. Long. 1987. Synoptic measures of
sediment contamination, toxicity and infaunal community composition (the
sediment quality triad) in San Francisco Bay. Mar. Ecol. Prog. Ser.
37:75-96.
OeWitt, T.H., G.R. Oitsworth, and R.C. Swartz. 1988. Effects of natural
sediment features on the phoxocephalid amphipod, Rhepoxynius abronius:
Implications for sediment toxicity bioassays. Mar. Environ. 'Res. 25:99-124.
OeWitt, T.H., R.C. Swartz, and J.O. Lamberson. In preparation. Measuring
the toxicity of estuarine sediments. Submitted to Environ. Toxicol. Chem.
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Bulk Sediment Toxicity
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. Environ. Toxicol. Chem. 7:483-498
Ingersoll, C.G., and M.K. Nelson. 1989. Solid-phase sediment toxicity
testing with the freshwater invertebrates: Hyalella azteca (Amphipoda) and
Chironomus riparius (Oiptera). In: Aquatic Toxicology Ris.k Assessment:
Proceedings of the Thirteenth Annual Symposium, ASTM STP, American Society
for Testing and Materials, Philadelphia, PA.
Kemp, P.P., and R.C. Swartz. In press. Acute toxicity of interstitial and
particle-bound cadmium to a marine infaunal amphipod. Marine Environ.
Res.
Lamberson, J.O., and R.C. Swartz. 1988. Use of bioassays in determining
the toxicity of sediment to benthic organisms. pp. 257-279. In: Toxic
Contaminants and Ecosystem Health; A Great Lakes Focus. M.S. Evans (Ed).
John Wiley and Sons, Inc.,
Nebeker, A.V., M.A. 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.
Nebeker, A.V., and C.E. Miller. 1988. Use of the amphipod crustacean
HyaleJla azteca in freshwater and estuarine sediment toxicity tests.
Environ. Toxicol. Chem. 7:1027-1034.
Ott, F.S. 1986. Amphipod sediment bioassays: effect of grain size,
csdR!'!U.T, msthcdclogy, snd Variations in animal sensitivity on interpretation
of experimental data. Ph.D. Dissertation. University of Washington,
Seattle, WA.
Plesha, P.O., J.E. Stein, M.H. Schiewe, B.B. McCain, and U. Varanasi. 1988.
Toxicity of marine sediments supplemented with mixtures of selected
chlorinated and aromatic hydrocarbons to the infaunal amphipod, Rhepoxynius
abronius. Mar. Environ. Res. 25:85-97.
PTI Environmental Services. 1988. Sediment quality values refinement:
Tasks 3 and 5 - 1988 update and evaluation of the Puget Sound AET. PTI
Environmental Services, Bellevue, WA.
Robinsbn, A.M., J.O. Lamberson, F.A. Cole, and R.C. Swartz. In press.
Effects of culture conditions on the sensitivity of phoxocephalid amphipod
Rhepoxynius abronius in two cadmium sediments. Environmental Toxicology and
Chemistry.
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Bulk Sediment Toxicity
Swartz, R.C.
contaminated
A.W. Maki
1987. Toxicological methods for determining the effects of
sediment on marine organisms. pp. 183-198. In: Fate and
Effects of Sediment Bound Chemicals in Aquatic Systems.
and W.A. Brungs (Eds). Pergamon Press, NY.
K.L. Dickson,
Swartz, R.C., W.A. OeBen, and F.A. Cole. 1979. A bioassay for the toxicity
of sediment to marine macrobenthos. Journal of the Water Pollution Control
Federation 51:944-950.
Swartz, R.C., W.A. OeBen, K.A. Sercu, and J.O. Lamberson. - 1982. Sediment
toxicity and the distribution of amphipods in Commencement Bay, Washington,
USA. Mar. Poll. Bull. 13:359-364.
Swartz, R.C., W.A. OeBen, J.K.P. Jones, J.O. Lamberson, and F.A. Cole.
1985a. Phoxocephalid amphipod bioassay for marine sediment toxicity. pp.
284-307. In: Aquatic Toxicology and Hazard Assessment: Proceedings of the
Seventh Annual Symposium. R.O. Cardwell, R. Purdy, and R.C. Bahner (Eds).
ASTM STP 854, American Society for Testing and Materials, Philadelphia, PA.
Swartz, R.C., O.W. Schults, G.R. Oitsworth, W.A. OeBen, and F.A. Cole.
1985b. Sediment toxicity, contamination, and macrobenthic communities near
a large sewage outfall, pp. 152-175. In: Validation and Predictability of
Laboratory Methods for Assessing the Fate and Effects of Contaminants in
Aquatic Ecosystems. T.P. Boyle (Ed). ASTM STP 865, American Society for
Testing and Materials, Philadelphia, PA.
Swartz, R.C., F.A. Cole, O.W. Schults,
changes on the Palos Verdes Shelf near
Mar. Ecol. Prog. Ser. 31:1-13.
and W.A. OeBen. 1986. Ecological
a large sewage outfall: 1980-1983.
Swartz, R.C., P.F. Kemp, O.W. Schults, G.R. Ditsworth, and R.J. Ozretich.
In press. Toxicity of sediment from Eagle Harbor, Washington to the
infaunal amphipod, Rhepoxynius abronius. Environ. Toxicol. Chem.
Tetra Tech. 1986. Eagle Harbor preliminary investigation.
EGHB-2. Tetra Tech, Inc., Bellevue, WA.
Final Report
U.S. Environmental Protection Agency.
fluoranthene. U.S. EPA, Washington, DC.
1980. Water quality criteria for
U.S. Environmental Protection Agency and U.S. Army Corps of Engineers.
1977. Ecological evaluation of proposed discharge of dredged material into
ocean waters. U.S. Army Engineer Waterways Experiment Station, Vicksburg,
MS.
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Spiked-Sediment Toxicity
CHAPTER 3. SPIKED-SEDIMENT TOXICITY TEST APPROACH
Janet Lamberson and Richard C. Swartz
U. S. Environmental Protection Agency
Environmental Research Laboratory - N. Pacific Division
Hatfield Marine Science Center
Newport, OR 97365
(503) 867-4031
The toxicological approach to generating sediment quality criteria uses
concentration-response data from sediments spiked in the laboratory with
known concentrations of contaminants to establish cause-and-effect relation-
ships between chemicals and adverse biological responses (e.g, mortality,
reductions 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.
1.0 SPECIFIC APPLICATIONS
1.1 Current Use
The spiked-sediment toxicity test approach is still in the research
stage. Although the procedures resemble those used to generate water
quality criteria, the influence of variable properties of sediment makes the
generation of sediment quality criteria values much more complex.
Where LC50 values and chronic effects data are available for chemicals
in sediments (see Section 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
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Spiked-Sediment Toxicity
tested by comparing them with field data on chemical concentrations in
natural sediments and observed biological effects. Interim laboratory-
derived criteria, however, can be implemented prior to field validation.
1.2 Potential Use
This method can be used to empirically address the problem of interac-
tions among complex mixtures of contaminants that are almost always present
in the field. 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).
2.0 DESCRIPTION
2.1 Description of Method
The toxicological approach involves exposing 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 period, the
response of the test organism is examined in relation to a biological
endpoint (e.g., mortality, growth, reproduction, cytotoxicity, alterations in
development or respiration rate). Results are then statistically compared
with results from control or reference sediments to identify toxic concentra-
tions of the test chemical.
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Spiked-Sediment Toxicity
2.1.1 Objectives and Assumptions--
The objective of this approach is to derive concentration-response
values in the laboratory that can be used to predict concentrations of
specific chemicals that would be harmful to resident biota under field
conditions. The effects of interactions (i.e., synergism, additivity,
antagonism) among chemicals in the field can be predicted from laboratory
results with sediments spiked with combinations of chemicals. The 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 * priori knowledge of specific pathways of interaction between
sediments and test organisms is not necessary. Any method of expressing the
bioavailabil ity of contaminants in sediment can be used in conjunction with
sediment toxicity tests, including the "free" interstitial concentration and
normalizations to organic carbon and other sediment properties.
Data generated by this method may be difficult 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.
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 straightforward and well documented (reviewed
by Lamberson and Swartz 1988; Nebeker et al. 1984; Swartz 1987; Tetra Tech
and- EVS 1986). However, many individual tests would be required to generate
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Spiked-Sediment Toxicity
an extensive database of sediment quality values for a large number of
chemicals, chemical combinations, and sediment types.
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 (i.e., uncontaminated) sediments from a control
area, uncontaminated reference sediments for comparison with similar
contaminated sediments, or contaminated sediments to be spiked with known
concentrations of chemicals in a test for interactions among contaminants.
Sufficient sediment must be collected to provide samples for chemical
analysis, spiking, and reference or controls (i.e., sediment for statistical
comparison with spiked sediment). Depending upon the experimental design,
the following controls may be required: sediment from the collection site
for test animals, positive controls with a reference toxicant, carrier
controls, and controls for natural sediment features that may affect test
animals (e.g., grain size distribution).
2.1.2.2 Methods — Various methods of adding chemicals to sediment
(i.e., spiking sediments) have been used. In general, the chemical is either
added to the sediment and mixed in (Francis et al . 1984; Swartz et al.
1986b 1988; Birge et al. 1987), or added to the overlying water (Hansen and
Tagatz 1980; Kemp and Swartz 1988) or to a sediment slurry (Landrum et al.
in press; Oliver 1984; Schuytema et al. 1984) and allowed to equilibrate
with the sediment. Sediments are spiked with a range of concentrations to
generate LC50 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 organisms to the
sediment for a specific length of time (e.g., 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 mortality, changes in
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Spiked-Sediment Toxicity
growth or reproduction, behavioral or physiological alterations, or
differences in numbers and species of larvae that become established in
contaminated vs. control sediments.
2.1.2.3 Types of Data Required--Soiked sediments as well as reference
or control sediments must be analyzed for total solids, grain size, and
total and dissolved organic carbon. The concentrations of chemicals added
to sediment must be determined in stock solutions as well as in the test
sediment. Bulk and interstitial concentrations 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 total volatile
solids or metals, change during exposure, measurements must be taken before,
during, and at the end of the exposure period, and the 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 statistically compared with control or
reference data to determine the occurrence of biological effects, and to
calculate LC50 values, EC50 values, no-effect concentrations, or lowest-
observable-effect concentrations. Establishment of the maximum acceptable
toxicant concentration requires data from a chronic or life-cycle test.
Data correlating observed biological effects with chemical concentra-
tions in spiked sediment can be used to calculate probit curves for
derivation of biological effect level values (e.g., EC50, EC05). Data from
several species of test organisms can be ranked, and the lowest contaminant
concentrations 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.
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Spiked-Sediment Toxicity
2.1.2.4 Necessary Hardware and Skills—Most toxicity test procedures
require a minimum 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 needed to detect and interpret biological endpoints. Chemical sampling
methods are generally simple and routine, although analyses of chemical
samples requires specialized training and equipment. Some biological
effects tests also require specialized training and experience, especially
to interpret the results.
2.1.3 Adequacy of Oocumentation--
Various acute sediment toxicity test procedures have been developed and
are well documented for testing freshwater and marine field sediments
(reviewed by Swartz 1987; Lamberson and Swartz 1988). 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 life-cycle test
procedures for benthic invertebrates. Life cycle tests exist for the
amphipod Ampelisca abdita (Scott and Redmond in press), the polychaetes
Neanthes arenaceodentata (Pesch 1979) and Capital la capitata (Chapman and
Fink 1984), and freshwater oligochaetes (Wiederholm et al. 1987). Chronic
exposures to most sensitive life stages are also inherent in the benthic
recolonization procedure (Hansen and Tagatz 1980).
2.2 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 directly
quantify the interactive effects of combinations of contaminants.
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Spiked-Sediment Toxicity
When chemical concentrations in tested biota are measured after a
spiked-sediment toxicity test, uptake of contaminants by benthic organisms
(i.e., bioaccumulation) can be predicted. As an important component of food
webs in aquatic ecosystems, benthic organisms can contribute toxicants from
contaminated sediments to higher levels of the aquatic food web and
ultimately affect human health. Sediment quality criteria and bioaccumu-
lation 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. Combined with chemical analysis of sediment samples and bulk
sediment toxicity testing, these limits can be used to define areas where
food species should not be consumed, or where direct contact with con-
taminated sediments can be hazardous to human health.
Bioaccumulation studies and sediment quality criteria established using
data from spiked-sediment toxicity testing with several benthic species can
also be used to protect benthic communities and aquatic species that feed
upon the benthos. Assuming a sufficient mix of taxonomic groups, sediment
quality criteria based on responses of the most sensitive species within a
benthic community can be developed to protect the structure and function of
the entire ecosystem (Hansen and Tagatz 1980).
2.3 Ability of Method to Generate Numerical Criteria for Specific Chemicals
The spiked-sediment toxicity test approach can be used to directly
measure the effects of specific chemicals in various types of sediments in
the laboratory and to establish unequivocal analysis of causal effects.
Since test conditions are controlled, the method can be used to experi-
mentally determine the effects of individual chemicals on mixtures of
chemicals or aquatic biota (Plesha et al. 1988; Swartz et al . 1988, 1989),
to establish pathways of toxicity, and to provide specific effects concen-
trations (e.g., LC50, EC50, no-effect concentration). The influence of
various physical characteristics of the sediment on chemical toxicity can
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Spiked-Sediment Toxicity
also be determined (Ott 1986; OeWitt et al. 1988). The available data
represent concentrations at which toxicity occurs rather than numerical
sediment quality criteria. However, U.S. EPA is currently conducting a
research project that may provide insight into the minimum database
necessary to establish a sediment quality criterion for fluoranthene based
on toxicological data..
Concentration-response data have been generated using spiked-sediment
toxicity test methods for a variety of chemicals, including both metals and
organic compounds. Specific data are available for phenanthrene, fluor-
anthene, zinc, mercury, copper, cadmium, hexachlorobenzene, pentachloro-
phenol, Aroclors 1242 and 1254, chlordane, ODE, DOT, dieldrin, endosulfan,
endrin, sevin, creosote, and kepone (Adams et al. 1985; Cairns et al. 1984;
OeWitt et al. in press; Kemp and Swartz 1989; McLeese and Metcalfe 1980;
McLeese et al. 1982; Swartz et al. 1986b, 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 content.
Normalizing factors for metals and other chemicals are currently under
research.
3.0 USEFULNESS
3.1 Environmental Applicability
3.1.1 Suitability for Different Sediment Types--
The spiked-sediment toxicity test approach is suitable for any type of
sediment. This approach can also be used to establish the bioavailable
component of the sediment ' responsible for the observed toxicity. The
effects of various physical properties of the sediment on chemical toxicity
can be experimentally determined. In some cases, the physical or chemical
properties of the test sediment such as salinity or grain size may limit
the species that can be used for testing and a substitute species must be
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Spiked-Sediment Toxicity
used (DeWitt et al. in press). In 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) are known, data can be normalized to these factors and thus the
approach can be applied to any sediment type.
3.1.2 Suitability for Different Chemicals or Classes of Chemicals--
A major advantage of the spiked-sediment toxicity test method is that it
is suitable for all classes of chemicals. In addition, this is the only
approach currently available that can empirically determine the interactive
effects among individual chemicals in mixtures of contaminants such as are
usually found in real-world sediments (Swartz et al. 1988a). This approach
can be used to provide experimental validation of sediment quality criteria
generated by other approaches.
3.1.3 Suitability for Predicting Effects on Different Organisms--
Theoretically, any organism can be used in spiked-sediment toxicity.
To protect a biological community and to predict the effects of a toxicant on
different organisms, test organisms should be selected on the basis of their
sensitivity to contaminants, their ability to withstand laboratory handling,
and their ability to survive in control treatments. In tests to determine
the effects of toxicants on a particular biological community, the test
species selected should be among the most sensitive species found in the
community or comparably sensitive. If the most sensitive species are
protected, the entire community should be protected.
3.1.4 Suitability for In-Place Pollutant Control--
Spiked-sediment toxicity testing can be used to develop sediment
quality criteria that can then be used to determine the extent of the
problem area, in monitoring temporal and spatial trends, and to assess the
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Spiked-Sediment Toxicity
need for remedial action. Criteria can be used to set target cleanup levels
and in post-cleanup monitoring of acceptable contaminant levels.
3.1.5 Suitability for Source Control--
When combined with wasteload allocation models, spiked-sediment
toxicity tests can be used in source control to establish maximum allowable
effluent concentrations or mass loadings of single chemicals and mixtures of
chemicals.
3.1.6 Suitability for Disposal Applications--
Spiked-sediment toxicity tests can be used to predict biological
effects of contaminants prior to approval of dredged material disposal or
sewage outfall permits.
3.2 General Advantages and Limitations
3.2.1 Ease of Use--
Most sediment toxicity test procedures are simple to use, requiring
limited expertise and standard inexpensive laboratory equipment. Only a few
sublethal-effects tests require specialized training.
3.2.2 Relative Cost--
The cost of individual toxicity tests is relatively low because of the
limited expertise and inexpensive equipment requirements (see Chapter 2,
Bulk Sediment Toxicity Approach). Costs -o implement this approach as a
regulatory tool would be comparatively high, since it would require the
collection of sediment chemistry data for comparison to data established by
the sediment toxicity test method. The cost of developing a large tbxico-
logical database would be relatively high because of the large number of
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Spiked-Sediment Toxicity
individual chemicals and sediments that would have to be tested. The cost
of generating the chemical and toxicological data necessary to establish a
sediment quality criterion for one chemical by this method is estimated to
be $100,000.
3.2.3 Tendency to be Conservative--
Spiked-sediment toxicity tests, which are laboratory-controlled
experiments, provide a high degree of accuracy and precision. They are
sufficiently controlled to provide a true estimate of the toxicity of
individual chemicals in sediment. Biological endpoints, species, and life
stages of test organisms required for testing are analogous to water quality
criteria minimum data requirements. Laboratory bioassays, especially acute
toxicity tests, are inherently limited in their ability to reflect all of
the ecological processes through which sediment contaminants may affect
benthic ecosystems under field situations.
3.2.4 Level of Acceptance--
Spiked-sediment toxicity test methods, which follow the procedures and
rationale used to develop water quality criteria, 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 procedures are in the process of standardization by the American
Society of Testing and Materials (ASTM) subcommittee on sediment toxicology.
A regulatory strategy through which data generated by these test methods to
establish sediment quality criteria is under development.
3.2.5 Ability to be Implemented by Laboratories with Typical Equipment and
Handling Facilities--
Spiked-sediment toxicity test methods are easily implemented by
laboratories with typical equipment, requiring inexpensive glassware and
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Spiked-Sediment Toxicity
little specialized training. Spiking sediments may require special handling
facilities for preparing stock solutions of highly toxic substances, and the
interpretation of some sublethal biological endpoints may require some degree
of training and experience. In general, special expertise or elaborate
facilities are not required for the biological tests, although analyses of
chemical samples require special equipment and training, and quality
control procedures are essential.
3.2.6 Level of Effort Required to Generate Results--
This procedure consists of a laboratory toxicity test, and requires a
moderate effort to initiate and terminate an experiment. The data generated
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 large level of effort.
3.2.7 Degree to Which Results Lend Themselves to Interpretation--
Sediment' toxicity tests applied to spiked sediments provide an
unequivocal analysis of cause-and-effect relationships between toxic
chemicals and biological responses. Since the procedures follow the
rationale used in 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.
3.2.8 Degree of Environmental Applicability--
The spiked-sediment toxicity test approach is applicable to a wide range
of environmental conditions and sediment types. The confounding effects of
sediment variables such as grain size and organic content can be either
experimentally addressed using toxicity test methods, or addressed using
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Spiked-Sediment Toxicity
normalization equations (DeWitt et al. 1988). A major advantage of the
procedure is the ability to predict interactive effects of chemical
mixtures such as would be found in field sediments.
3.2.9 Degree of Accuracy and Precision--
Since the sediment toxicity test 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. Field validation of sediment criteria generated by this approach
is required.
4.0 STATUS
4.1 Extent of Use
Spiked-sediment toxicity test procedures are under development in
several laboratories. Spiking procedures, as well as biological test
procedures, are in the process of standardization by ASTM's sediment
toxicology subcommittee.
4.2 Extent to Which Approach Has Been Field-Validated
Spiked-sediment toxicity test values have not been well field-validated,
although some results have been published (Plesha et al. 1988; Swartz et
al. 1989). As more data and criteria values become available, they can be
compared with existing information 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 during field validation
(Swartz et al. 1989).
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Spiked-Sediment Toxicity
4.3 Reasons for Limited Use
The approach is still in the developmental stage in several labora-
tories, and although some data have been generated and compared to field
conditions, a relatively large effort will be needed to generate a large
database. There have been few comparisons of methods and species sen-
sitivity, and few chronic toxicity tests have been developed.
4.4 Outlook for Future Use and Amount of Development Yet Needed
The outlook for future use of 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 among species, methods, and laboratories can be
made and sediment quality criteria can be developed.
5.0 REFERENCES
Adams, W.J., R.A. Kimerle, and R.G. Mosher. 1985. Aquatic safety assessment
of chemicals sorbed to sediments, pp. 429-453. In: Aquatic Toxicology and
Hazard Assessment: Seventh Symposium. R.D. Cardwell, R. Purdy, and R.C.
Bahner (eds). ASTM STP 854. American Society for Testing and Materials,
Philadelphia, PA.
Battelle. 1988. Overview of methods for assessing and managing sediment
quality. Prepared for U.S. Environmental Protection Agency, Office of
Water, Office of Marine and Estuarine Protection, Washington, DC. Battelle
Ocean Sciences, Ouxbury, MA.
Birge, W.J., J. Black, S. Westerman, and P. Francis. 1987. Toxicity of
sediment-associated metals to freshwater organisms: biomonitoring pro-
cedures, pp. 199-218. In: Fate and Effects of Sediment Bound Chemicals
in Aquatic Systems. K.L. Oickson, A.W. Maki, and W.A. Brungs (eds).
Pergamon Press, N.Y.
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Spiked-Sediment Toxicity
Cairns, M.A., A.V. Nebeker, J.H. Gakstatter, and W.L. Griffis. 1984.
Toxicity of copper-spiked sediments to freshwater invertebrates. Environ.
Toxicol. Chem. 3:435-445.
Chapman, P.M., and R. Fink. 1984. Effects of Puget Sound sediments and
their elutriates on the life cycle of Capitella capitata. Bull. Environ.
Contain. Toxicol. 33:451-459.
DeWitt, T.H., G.R. Oitsworth, and R.C. Swartz. 1988. Effects of natural
sediment features on the phoxocephalid amphipod Rhepoxynius abronius:
implications for sediment toxicity bioassays. Mar. Environ. Res. 25:99-124.
OeWitt, T.H., R.C. Swartz, and J.O. Lamberson. In press. Measuring the
toxicity of estuarine sediments. Environ. Toxicol. Chem.
Francis, P.C., W.J Birge, and J.A. Black. 1984. Effects of cadmium-enriched
sediment on fish and amphibian embryo-larval stages. Ecotoxicol. Environm.
Safety 8:378-387. .
Hansen, D.J., and M.E. Tagatz. 1980. A laboratory test for assessing
impacts of substances on developing communities of benthic estuarine
organisms, pp. 40-57. In: Aquatic Toxicology. J.G. Eaton, P.R. Parrish,
and A.C. Hendricks (eds). ASTM STP 707. American Society for Testing and
Materials, Philadelphia, PA.
Kemp, P.F., and R.C. Swartz. 1988. Acute toxicity of interstitial and
particle-bound cadmium to a marine infaunal amphipod. Mar. Environ. Res.
26:135-153.
Lamberson, J.O., and R.C. Sv-artz. 1988. Uss cf bicassays in determining the
toxicity of sediment to benthic organisms. pp. 257-279. In: Toxic
Contaminants and Ecosystem Health: A Great Lakes Focus. M.S. Evans (ed).
John Wiley and Sons, Inc., New York, NY.
Landrum, P.F., W.R. Faust, and B.J. Eadie. In press. Bioavailability and
toxicity of a mixture of sediment associated chlorinated hydrocarbons to the
amphipod Pontoporeia hoyi. American Society of Testing and Materials,
Philadelphia, PA,
McLeese, O.W., and C.O. Metcalfe. 1980. Toxicities of eight organochlorine
compounds in sediment and seawater to Crangon septemspinosa. Bull. Environ.
Contam. Toxicol. 25:921-928.
McLeese, O.W., L.E. Burridge, and J. Van Ointer. 1982. Toxicities of five
organochlorine compounds in water and sediment to Nereis virens. Bull.
Environ. Contam. Toxicol. 28:216-220.
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Nebeker, A.V., M.A.
and O.F. Krawczyk.
contaminated freshwater
3:617-630.
Spiked-Sediment Toxicity
Cairns, J.H. Gakstatter, K.W. Malueg, G.S. Schuytema,
1984. Biological methods for determining toxicity of
sediments to invertebrates. Environ. Toxicol. Chem.
Oliver, B.G. 1984.
spiked and
21:785-790.
field sediments
Bio-uptake of chlorinated hydrocarbons from laboratory-
liments by oligochaete worms. Environ. Sci. Technol.
Ott, F.S. 1986. Amphipod sediment bioassays: effect of grain size,
cadmium, methodology, and variations in animal sensitivity on interpretation
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 arenaceodentata. Marine Biol. 52:237-245.
Plesha, P.O., 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 hydrocarbons to the infaunal amph.ipod Rhepoxynius
abronius. Mar. Environ. Res. 25:85-97.
Schuytema, G.S., P.O. Nelson, K.W. Malueg, A.V. Nebeker, O.F. Krawczyk, A.K.
Ratcliff, and J.H. Gakstatter. 1984. Toxicity of cadmium in water and
sediment slurries to Daphnia magna. Environ. Toxicol. Chem. 3:293-308.
Scott, K.J., and M.S. Redmond. In press. The effects of a contaminated
dredged material on laboratory populations of the tubicolous amphipod
Ampelisca abdita. In: Aquatic Toxicology and Hazard Assessment: Twelfth
Volume. U.M. Cowgill and L.R. Williams (eds). ASTM STP 1027. American
Society for Testing and Materials, Philadelphia, PA.
Swartz, R.C. 1987. Toxicological methods for determining the effects of
contaminated sediment on marine organisms. pp. 183-198. In: Fate and
Effects of Sediment Bound Chemicals in Aquatic Systems. K.I. Dickson, A.M.
Maki, and W.A. Brungs (eds). Pergamon Press, NY.
Swartz, R.C., W.A. OeBen, K.A. Sercu, and
toxicity and the distribution of amphipods
USA. Mar. Poll. Bull. 13:359-364.
J.O. Lamberson.
in Commencement
1982. Sediment
Bay, Washington,
Swartz, R.C., D.W. Schults, G.R. Oitsworth, W.A. OeBen, and F.A. Cole.
1985. Phoxocephalid amphipod bioassay for marine sediment toxicity. pp.
284-307. In: Aquatic Toxicology and Hazard Assessment: Proceedings of the
Seventh Annual Symposium. R.O. Cardwell, R. Purdy, and R.C. Bahner (eds).
ASTM STP 854. American Society for Testing and Materials, Philadephia, PA.
3-16
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Swartz, R.C., D.W. Schults, G.R. Ditsworth, W.A. OeBen, and F.A. Cole.
1985. Sediment toxicity, contamination, and macrobenthic communities near
a large sewage outfall, pp. 152-175. In: Validation and Predictability
of Laboratory Methods for Assessing the Fate and Effects of Contaminants in
Aquatic Ecosystems. T.P. Boyle (ed). ASTM STP 865. American Society for
Testing and Materials, Philadelphia, PA.
Swartz. R.C., F.A. Cole, D.W. Schults, and W.A. DeBen. 1986a. Ecological
changes on the Palos Verdes Shelf near a large sewage outfall: 1980-1983.
Mar. Ecol. Prog. Ser. 31:1-13.
Swartz, R.C., G.R. Oitsworth, O.W. Schults, and J.O. Lamberson. 1986b.
Sediment toxicity to a marine infaunal amphipod: cadmium and its interaction
with sewage sludge. Mar. Environ. Res. 18:133-153.
Swartz, R.C., P.F. Kemp, O.W. Schults, and J.O. Lamberson. 1988. Effects of
mixtures of sediment contaminants on the marine infaunal amphipod Rhepoxy-
nius abronius. Environ. Toxicol. Chem. 7:1013-1020.
Swartz, R.C., P.F. Kemp, D.W. Schults, G.R. Ditsworth, and R.J. Ozretich.
1989. Toxicity of sediment from Eagle Harbor, Washington to the infaunal
amphipod Rhepoxynius abronius. Environ. Toxicol. Chem. 8:215-222.
Tagatz, M.E., J.M. Ivey, and H.K. Lehman. 1979. Effects of sevin on
development of experimental estuarine communities. J. Toxicol. Environ.
Health 5:643-651.
Tagatz, M.E., J.M. Ivey, J.C. Moore, and M. Tobia. 1977. Effects of
pentachlorophenol on the development of estuarine communities. J. Toxicol.
Environ. Health 3:501-506.
Tagatz, M.E., G.R. Plaia, C.H. Deans, and E.M. Lores. 1983. Toxicity of
creosote-contaminated sediment to field- and laboratory-colonized estuarine
benthic communities. Environ. Toxicol. Chem. 2:441-450.
Tetra Tech. 1986. Development of sediment quality values for Puget Sound.
OACW67-85-0029, Work Order 0001C, TC3090-02; Task 6 Final Report. Prepared
for Puget Sound Dredge Disposal Analysis. Tetra Tech, Inc., Bellevue, WA.
Tetra Tech, and E.V.S. Consultants. 1986. Recommended protocols for
conducting laboratory bioassays on Puget Sound sediments. Prepared for U.S.
Environmental Protection Agency, Region 10, Office of Puget Sound. Tetra
Tech Inc., Bellevue, WA.
Wiederholm, T., A.-M. Wiederholm, and G. Milbrink. 1987. Bulk sediment
bioassays with five species of fresh-water oligochaetes. Water Air Soil
Pollut. 36:131-154.
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Spiked-Sediment Toxicity
Word, J.Q., J.A. Ward, L.M. Franklin, V.I. Cullinan, and S.I. Kiesser.
1987. Evaluation of the equilibrium partition theory for estimating the
toxicity of the nonpolar organic compound DDT to the sediment dwelling
organism Rhepoxynius abronius. Prepared for U.S. Environmental Protection
Agency, Criteria and Standards Division, Washington, DC. Battelle Washington
Environmental Program Office, Washington, DC.
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Interstitial Water
CHAPTER 4. INTERSTITIAL WATER TOXICITY APPROACH
Gerald Ankley and Nelson Thomas
U.S. Environmental Protection Agency
Environmental Research Laboratory
6201 Congdon Boulevard
Ouluth, MN 55804
(218) 720-5702
The interstitial water toxicity approach is a multiphase procedure for
assessing sediment toxicity using interstitial (i.e., pore) water. The use
of pore water for sediment toxicity assessment is based on the strong
correlations between contaminant concentrations in pore water and toxicity
(and/or bioaccumulation) of sediment-associated contaminants by benthic
macroinvertebrates (Adams et al. 1985; Swartz et al. 1985; Connell et al.
1988; OiToro 1988; Knezovich and Harrison 1988; Swartz et al. 1988; Giesy
and Hoke in press). The approach combines the quantitation of pore water
toxicity with toxicity identification evaluation (TIE) procedures to
identify and quantify chemical components responsible for sediment toxicity
(Mount and Ander$nn-Carnahan 1988a,b; Mount 1988). TIE involves recently
developed techniques for the identification of toxic compounds in aqueous
samples containing mixtures of chemicals. In the interstitial water
toxicity method, TIE procedures are implemented in three phases to charac-
terize pore water toxicity, identify the suspected toxicant, and confirm
toxicant identification.
1.0 SPECIFIC APPLICATIONS
1.1 Current Use
The TIE procedures described herein were developed over the last 3 yr
using municipal and industrial effluents from more than 30. sites. They have
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Interstitial Water
been used with several aquatic species including cladocerans and fishes,
and they can be used with any type of benthic species that is amenable to
toxicity testing in aqueous phases. 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. This success
rate was achieved with a sample size of greater than 60 municipal and
industrial effluents, surface water samples, and sediment fractions,
including pore water and elutriates.
1.2 Potential Use
The use of pore water as a fraction to assess sediment toxicity, in
conjunction with associated TIE procedures, can provide data concerning
specific compounds responsible for toxicity in contaminated sediments.
These data could be critical to the success of remediation of toxic
sediments.
In spite of existing uncertainties in using pore water to assess
sediment toxicity, the ability to identify specific toxicants responsible
for acute toxicity in contaminated sediments makes pore water a potentially
important sediment test fraction. Thus this method, in conjunction with
other sediment classification methods, could prove to be extremely valuable.
2.0 DESCRIPTION
2.1 Description of Method
The interstitial water toxicity method involves three major steps:
• Isolation of pore water from sediment samples
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Interstitial Water
• Performance of toxicity tests on pore waters
• 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 of bulk sediment, direct sampling of pore water
through the use of dialysis membranes, and micro-syringe sampling (Knezovich
et al. 1987; Knezovich and Harrison 1988; Sly 1988). The most representative
pore water samples probably are 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 toxicity of sediment pore water (Giesy et
al. 1988; Ankley et al. in press; Hoke et al. in preparation). However,
there has been no critical evaluation of the relative advantages and
disadvantages of the former three pore water preparation procedures in terms
of toxicity assessment. Consequently, it would be premature to recommend one
over another. With any of these pore water preparation techniques, care
must be taken to avoid loss of contaminants due to oxidation, change in pH,
or other interferences, during sample preparation.
After preparation of pore water, toxicity tests can be performed using
either standard test species (U.S. EPA 1985a,b) or various types of benthic
organisms amenable to toxicity testing in aqueous samples. In samples
exhibiting acute toxicity, it is then possible to directly apply the TIE
procedures described below in Section 2.1.2.2.
2.1.1 Objectives and Assumptions--
The objective of this approach is to derive toxicity data in the
laboratory that can be used to assess sediment toxicity in field situations.
With the interstitial water toxicity method, it is possible to quantify
toxicity in a sample and potentially to identify chemical components
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Interstitial Water
responsible for toxicity. The major assumption in using this method is
that the compounds that are toxic to test organisms in the pore water are
the same compounds that cause toxicity in sediments in situ.
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. Effort
expended in the TIE studies will be proportional to the complexity of
analyses required for the identification of suspected toxicants. .
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 measure-
ments should include (but are not limited to) pH, hardness, alkalinity,
salinity (where appropriate), dissolved oxygen, sulfides, and ammonia.
Certain of these variables, in particular pH, also should be monitored in
the bulk sediment.
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 generated) is available in Mount and Anderson-Carnahan
(1988a,b) and Mount (1988).
Toxic sediment samples can potentially contain 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. In addition, when multiple
toxicants are present, it must be possible to determine the proportion of
the overall toxicity due to each toxicant.
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Interstitial Water
After preparation of pore water and performance of initial toxicity
tests, the Initial step in the TIE process is to separate toxicants 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). This approach allows
the physical and chemical nature of the toxicants to be determined prior to
instrumental analysis. Consequently, the "correct" analyti-cal technique can
be selected for detecting as well as identifying the toxicants in the
subsample. In addition, significantly fewer chemical components are in the
subsamples as compared to the original sample, and thus, 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 relationship ultimately results
in highly efficient TIEs.
With the toxicity-based TIE approach, detection 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 physical/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
manipulations, provide information on the nature of the toxicants.
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Successful completion of Phase I occurs when both the nature of the
components causing toxicity, as well as their consistent presence 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 4-1. Not shown in Figure 4-1, but performed on all samples are
routine water chemistry measurements including pH, hardness, conductivity,
and dissolved oxygen. These routine measurements are needed for designing
sample manipulations, and interpreting test data. The manipulations shown in
Figure 4-1 are usually sufficient to characterize toxicity caused by a single
chemical. When multiple toxicants are present, various combinations of the
Phase I manipulations will most Itkely be required for toxicant characteri-
zation.
Many of the manipulations in Phase I require samples that have been
pH-adjusted. The adjustment of pH is a powerful tool for detecting cationic
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 examine
the second.
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OxWart
Reduction
Atritton [
I
Add
1
PH,
\
8**
Fftritton
Add
PHi
Toxic Aqu«oua Sa/npto
pH AdjuatTTMnt
f
Add
V
PH, &u
EDTA
Extrtctlco
Add
PH,
Qradu«t«d pH Teat
pH6
PH7
pH3
Figure 4-1. Overview of the Phase I toxicity characterization process.
The ambient pH of the sample is indicated as pH|.
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Interstitial Water
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. Generally, the un-ionized form of a toxicant is able to cross
biological membranes more readily than the ionized form and thus, is more
toxic. This test is designed primarily for ammonia, a relatively common
toxicant whose toxicity is extremely pH-dependent (U.S. EPA 1985c).
However, different pH values can strongly affect the toxicity of many common
ionizable pesticides, and also may influence the bioavailability and
toxicity of certain heavy metals (Campbell and Stokes 1985; Doe et al. 1988).
Aeration tests are designed to determine whether or not toxicity is at-
tributable to volatile or oxidizable compounds. Samples at pHj (ambient pH),
pH 3, and pH 11 are sparged with air for 1 h, readjusted to pHj, and tested
for toxicity. 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 is 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 implicated; if toxicity
is again reduced, volatile compounds are suspect. The pH at which toxicity
is reduced is important. If nitrogen sparging decreases toxicity at pHj,
neutral volatiles are present, whereas, if toxicity decreases at pH 11.0 or
pH 3.0, basic and acidic volatiles, respectively, are implicated.
Filtration provides information concerning the amount of toxicity
associated with filterable components. In this test, samples at pHj, pH 3.0,
and pH 11.0 are passed through 1-um filters, readjusted to pHj, ant, 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 toxicants, which could
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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 pHj,
pH 3.0, or pH 9.0. This test, in conjunction with associated Phase II
analytical procedures, is an extremely powerful TIE tool. In this procedure,
filtered sample aliquots at pHj, pH 3.0, and pH 9.0 are passed through small
columns packed with an octadecyl (C^g) sorbent. At pHj, the CIQ 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
Michael 1987). During extraction, the resulting post-column effluent is
collected and tested for toxicity in order to determine if the manipulation
removed toxicity and/or if the capacity of the column was exceeded. If
sample toxicity is decreased, a nonpolar toxicant would be suspected.
The presence of toxicity due to cationic metals is tested through
additions of ethylenediaminetetraacetic 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 metals, and only weakly chelates certain
cationic metals (e.g., silver, chromium, thallium) (Stumm and Morgan 1981).
Because EDTA nonspecifically binds a variety of cations, the appropriate
range of EDTA concentrations to use in the test is highly dependent upon
calcium and magnesium concentration (hardness) and salinity, as well as the
sensitivity of the test organism to EDTA.
The oxidant reduction test is designed to determine the degree of
toxicity associated with chemicals reduced by sodium thiosulfate. The
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toxicity of compounds such as chlorine, bromine, .iodine, and manganous ions
are neutralized by this treatment. 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 laboratory work in Phase II
focuses on isolation of the toxicants using chemical fractionation techniques
with toxicity tracking. The ideal isolation process would create a
subsample that contains one chemical, the toxicant. Depending upon 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 LC50 value is obtained,
usually from the literature or occasionally from structure activity models
(Institute for Biological and Chemical Process Analyses 1986). By comparing
concentrations of the identified chemicals to their LC50 values, a list of
suspect toxicants is made. This list is then refined by actually determining
LC50 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 protracted,
since interactions between 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
separation, combined with concentration steps is required 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 analytical approaches
employed in Phase II, results for two typical Phase I TIE evaluations are
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presented in Table 4-1. The Phase II methods and approaches appropriate for
these examples are discussed below.
In the first sample in Table 4-1, 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. Although
solvents other than methanol are routinely used in analytical work with
Cjg chromatography columns, the low toxicity of methanol to aquatic
organisms (e.g., LC50 >1.5 percent for cladocerans) 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 I) work is required. If toxicity occurs in the fractions, Phase II
methods feature concentration of the toxic methanol/water fractions, high
performance liquid chromatography fractionation of the concentrate (again
with a Cis/methanol/water solvent system) with concurrent toxicity testing
of the fractions, and ultimately, identification of suspected toxicants in
the toxic fractions via gas chromatography/mass spectroscopy.
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
mStal analyses and the concentration, fractionation, and identification
techniques described for nonpolar organics 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 nonpolar compounds.
Only a very small subset of possible Phase I results is shown in
Table 4-1, 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
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TABLE 4-1. PHASE I CHARACTERIZATION RESULTS AND SUSPECT
TOXICANT CLASSIFICATION FOR TWO EFFLUENTS
Effluent3
One
Two
Phase I Test
Oxidant reduction
EDTA addition
Graduated pH test
pH adjustment
Filtration
Aeration
SPE
NR
NR
NR
NR
NR
NR
R
NR
R
NR
NR
NR
NR
R
Suspected toxicant classification
Nonpolar
organics
Nonpolar
organics/
cationic metals
a NR • No reduction in toxicity.
R » Reduction in toxicity.
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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. Confirmation 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 conclusions about the
toxicants. Furthermore, there is a possibility that substances causing
toxicity are different from sample to sample within a supposedly homogeneous
geographic region. Phase III enables both situations to be addressed. The
tools used in Phase III include correlation, relative species sensitivity,
observation of symptoms, spiking, and mass balance techniques. In most
instances, no single Phase III test is adequate to confirm 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 approach to succeed, temporal or spatial variation has to be
wide enough to provide a range of values adequate for meaningful analyses.
In order to use the correlation approach effectively whan thsrs are multiple
suspected 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. If there are two or more species that 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 provides evidence for the validity of the
suspect being the true toxicant.
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Another Phase III procedure is 'observation of symptoms (e.g., time to
mortality) in poisoned animals. Although this approach does not necessarily
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 can also be obtained by spiking samples with the
suspected toxicants. While the results may not be conclusive, an increase
in toxicity by the same proportion as the increase in concentration of the
suspected toxicant in the sample suggests that the suspect is correct. To
get a proportional increase in toxicity from the addition of a suspected
toxicant when in fact it is not the true toxicant, both the true and
suspected 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.
2.1.2.3 Types of Data Required—In addition to the routine measure-
ments described above, biological response data, either acute •:- chronic,
will be obtained. Specific data collected will depend upon the choice of
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test organism. 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 Section 2.1.2.2 above).
2.1.2.4 Necessary Hardware and Skills — Pore water preparation and
toxicity test procedures are fairly straightforward, and require commonly
available equipment and facilities. Many of the TIE procedures also require
only routine facilities. However, certain TIE techniques require some degree
of advanced analytical capability (e.g., atomic absorption spectroscopy, gas
chromatography/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 an
advanced degree of technical expertise and experience.
2.1.3 Adequacy of Documentation--
The theoretical basis for using pore water to assess toxicity appears
to be scientifically sound, and thus, has been recommended for sediment
toxicity evaluation (Adams et al. 1985; Swartz et al. 1985; Knezovich and
Harrison 1988; Swartz et al. 1988; Connell et al. 1988; DiToro 1988; Giesy
and Hoke in press). Toxicity tests that can be used are well-documented,
standard procedures (U.S. EPA 1985a,b). The TIE techniques involved have
been documented and evaluated (Mount and Anderson-Carnahan 1988a,b; Mount
1988). Also, sediment toxicity assessment with pore water, including
toxicant identification, has been successfully demonstrated (Ankley et al.
in press).
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2.2 Applicability -of Method to Human Health. Aquatic LiFe. or Wildlife
Protection
This method can be used to predict biological effects of toxic
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 for evaluating human health effects or
protecting wildlife.
2.3 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 established 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 impacted 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.
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3.0 USEFULNESS
3.1 Environmental ADD!icability
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 complete 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 required. This translates into a bulk sediment
requirement of 3-8 L. Bulk sediment volumes needed for Phase II identifi-
cation will, of course, be dependent upon the toxicants present in the pore
water, but typical volumes required would be expected to range from 1 to
20 L.
3.1.2 Suitability for Different Chemicals or Classes of Chemicals--
This approach appears to be suitable for various water soluble nonpolar
organics, cationic metals, and ammonia (Adams et al. 1985; Swartz et al.
1935; Knezovich and Harrison 1988; Swartz et al. 1383; Connell et a1. 1388;
OiToro 1988; Ankley et al. in press). The applicability of the approach to
toxicants such as polar organics or extremely lipophilic compounds has yet
to be established. Also, the TIE procedures enable the evaluation of
interactive (additive, synergistic, antagonistic) effects among various
toxicants present in pore water samples (Mount and Anderson-Carnahan
1988a,b; Mount 1988).
3.1.3 Suitability for Predicting Effects on Different Organisms--
If the TIE procedures successfully identify specific toxicants
responsible for sediment toxicity, the impacts of these toxicants on various
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species of concern can be easily predicted, provided that there are data
concerning the toxicity of the identified compounds to these species.
Although toxicity data may not be available for certain benthic species,
once suspected toxicants are identified, it would be possible to generate
toxicity data for specific species of concern.
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. Because
sediment toxicants are actually identified, it is possible to design
remediation plans for toxicants 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.
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 sources.
3.1.6 Suitability for Disposal Applications--
As stated above, because specific sediment toxicants can be identified,
it would be possible to identify potential hazards of contaminated sediments
to aquatic organisms before disposal operations.
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3.2 General Advantages and Limitatinn<
3.2.1 Ease of Use--
Pore water preparation, routine chemical analyses, toxicity tests, and
certain of the TIE 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 utilized, in addition to tests with various benthic
species (U.S. EPA 1985a.b). However, interpretation of results of certain
of the TIE procedures, as well as analytical support for the TIE work,
requires advanced training and experience. At present, there are no set
protocols for the preparation of pore water, and there is uncertainty about
changes in pore water chemistry after extraction. Also, several TIE analyses
require highly sensitive analytical instrumentation for procedures, such as
atomic absorption spectroscopy and gas chromatography/mass spectroscopy.
3.2.2 Relative Cost--
Cost of the actual toxicity test procedures is relatively low. Cost of
the TIE procedures will vary depending upon the nature of the toxic
compounds; certain toxicants (e.g., pesticides) are more costly to identify
and quantify than others (e.g., ammonia). Also, identification and
determination of the effects of multiple toxicants in samples cost more
than the identification of single toxicants. Thus, cost analysis for the
TIE portion of the toxicity assessment is case-specific.
3.2.3 Tendency to be Conservative--
Depending upon tha species used and the endpoint evaluated, pore water
toxicity tests can be as conservative as desired.
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Interstitial Water
3.2.4 Level of Acceptance--
The theoretical basis of pore water toxicity assessment is sound
(Adams et al. 1985; Swartz et al. 1985; Knezovich and Harrison 1988; Swartz
et al. 1988; Connell et al. 1988; OiToro 1988; Giesy and Hoke in press). The
most important advantage of utilizing 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 complex mixtures of chemicals (Mount and Anderson-Carnahan
1988a,b). These procedures are not available for direct chemical analyses of
sediments. TIE procedures have proven to be extremely powerful tools for
work with complex effluents, and can be used with any type of acutely toxic
aqueous sample, including sediment pore water (Ankley et al.. in press).
The ability to identify specific compounds responsible for toxicity of
contaminated sediments clearly could be critical to the success of remedia-
tion.
3.2.5 Ability to be Implemented by Laboratories with Typical Equipment and
Handling Facilities--
Pore water preparation, toxicity test procedures, and certain of the TIE
methods are easily implemented by laboratories with typical equipment and a
moderate degree of expertise. Interpretation of some TIE results requires
additional technical training and experience, and certain of the analytical
procedures associated with TIE work require both specialized training and
analytical instrumentation.
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 upon the results of
the TIE work, the level of effort expended to obtain potentially important
data can be relatively small.
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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 are straight-
forward and easily interpreted by personnel with appropriate backgrounds.
3.2.8 Degree of Environmental Applicability--
Pore water toxicity assessment and TIE procedures are applicable to
virtually all environmental 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 is 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.
3.2.3 Degree of Accuracy and Precision--
Because the procedures consist of laboratory controlled experiments,
results obtained are statistically accurate and precise.
4.0 STATUS
4.1 Extent of Use
Various toxicity tests have been widely applied to the evaluation of
both freshwater and marine sediments, and pore water is merely one of the
possible fractions that can be tested. Theoretically, pore water appears
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to be appropriate for sediment toxicity assessment (Adams et al. 1985;
Swartz et al. 1985; Knezovich and Harrison 1988; Swartz et al. 1988; Connell
et al. 1988), and thus, it has been recommended as a suitable fraction for
the evaluation of sediment toxicity (DiToro 1988; Giesy and Hoke in press).
The TIE procedures (Mount and Anderson-Carnahan 1988a,b; Mount 1988),
although developed only relatively recently, already are widely used both in
research and regulatory programs.
4.2 Extent to Which Approach Has Been Field-Validated
Because the procedure is very new, there has been little field
validation. This area requires research, not only for the pore water
method described herein, but for virtually any other sediment classification
method involving toxicity tests or chemical analyses.
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 the theoretical basis for utilizing pore water has only
recently been critically evaluated. For this reason, there are no standard
methods for pore water preparation. 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 samples, pore water
appears to be the best fraction with which to attempt to identify specific
sediment contaminants responsible for acute toxicity.
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 which enables the identification of specific
compounds responsible for sediment toxicity with some degree of certainty.
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Interstitial Water
This information could be critical to the success of remediation. However,
as with all of the existing sediment classification methods, further
development is needed, particularly in the following areas:
• Development of standard and scientifically sound techniques
for pore water isolation
• Further characterization ;f relationships between sediment
toxicity in situ and the toxicity of sediment pore water in
the laboratory for different classes of compounds
• The development of TIE procedures to identify chronically
toxic compounds in aqueous samples (research in this area is
ongoing at ERL-Ouluth, primarily with complex effluents).
5.0 REFERENCES
Adams, W.J., R.A. Kimerle, and R.G. Mosher. 1985. Aquatic safety assessment
of chemicals sorbed to sediments, pp. 429-453. In: Aquatic Toxicology and
Hazard Assessment: Seventh Symposium. R.D. Cardwell, R. Purdy, and R.C.
Bahner (eds). ASTM STP 854. American Society for Testing and Materials,
Philadelphia, PA.
Ank1ey? G.T., A. Katko-, and J.W. Arthur. (In prsss). Identification of
ammonia as a major sediment-associated toxicant in the lower Fox River and
Green Bay, Wisconsin. Environ. Toxicol. Chem.
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, 0. Leslie, and J. Michaelson. 1987.
Differentiation 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 hydrocarbons from sediment by oligochaetes. Ecotoxicol.
Environ. Safety 16:293-302.
DiToro, O.M. 1988. Equilibrium partitioning approach to generating
sediment quality criteria. Report to the U.S. Environmental Protection
Agency Science Advisory Board, December 1988, Washington, DC.
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Doe, K.G. W.R. Ernst, W.R. Parker, G.R.J. Julien, and P.A. Hennigar. 1988.
Influence of pH on the acute lethality of fenitrothion, 2,4-0 and aminocarb
and some pH-altered sublethal effects of aminocarb on rainbow trout (Salmo
gairdneri). 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. Environ. Toxicol. Chem. 7:483-498.
Giesy, J.P., and R.A. Hoke.
bioassessment: rationale for
Lakes Res.
(In press). Freshwater sediment toxicity
species selection and test design. J. Great
Hoke, R.A., J.P. Giesy, G.T. Ankley, and J.L. Newsted. (In preparation).
Sediment toxicity assessment in the Maumee River and Lake Erie. Submitted
to J. Great Lakes Res.
Institute for Biological and Chemical Process Analyses. 1986,
for QSAR system. Montana State University, Bozeman, MT.
Junk, G.A., and J.J. Richard. 1988.
extraction on a small scale. Anal. Chem.
Organics in water:
60:451-454.
User manual
solid phase
Knezovich, J.P., and F.L. Harrison. 1988. The bioavailabil ity of sediment-
sorbed chlorobenzenes to larvae of the midge Chironomus decorus. Ecotoxicol.
Environ. Safety 15:226-241.
Knezovich, J.P., F.L. Henderson, and R.G. Wilhelm. 1987. The bioavail-
abil ity of sediment-sorbed organic chemicals: a review. Water Air Soil
Pollut. 32:233-245.
Mount, D.I. 1988. Methods for aquatic toxicity identification evaluations:
phase III toxicity confirmation procedures. EPA/600-3-88/036. U.S.
Environmental Protection Agency, Duluth, MN.
Mount, O.I., and L. Anderson-Carnahan. 1988a. Methods for aquatic toxicity
identification evaluations: phase I toxicity characterization procedures.
EPA/600-3-88/034. U.S. Environmental Protection Agency, Ouluth, MN.
Mount, O.I., and L. Anderson-Carnahan. 1983b. Methods for aquatic toxicity
identification evaluations: phase II toxicity identification procedures.
EPA/600-3-88/035. U.S. Environmental Protection Agency, Ouluth, MN.
Sly, P.G. 1988. Interstitial water quality of lake trout spawning habitat.
J. Great Lakes Res. 14:301-315.
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Interstitial Water
Stumm,' W., and J.J. Morgan. 1981. Aquatic chemistry - an introduction
emphasizing chemical equilibria in natural wate'rs. John Wiley 4 Sons, New
York, NY. 583 pp.
Swartz, R.C., G.R. Ditsworth, D.W. Schults, and J.O. Lamberson. 1985.
Sediment toxicity to a marine infaunal amphipod: cadmium and its interaction
with sewage sludge. Mar. Environ. Res. 18:133-153.
Swartz, R.C., P.P. Kemp, O.W. Schults, and J.O. Lamberson. 1988. Effects of
mixtures of sediment contaminants on the marine infaunal amphipod Rhepoxy-
nius abronius. Environ. Toxicol. Chem. 7:1013-1020.
U.S. Environmental Protection Agency. 1985a. Methods for measuring the
acute toxicity of effluents to freshwater and marine organisms. EPA/600/4-
85-013. U.S. EPA, Cincinnati, OH.
U.S. Environmental Protection Agency. 1985b. Short-term methods for
estimating the chronic toxicity of effluents and receiving waters to
freshwater organisms. EPA/600/4-85-014. U.S. EPA, Cincinnati, OH.
U.S. Environmental Protection Agency. 1985c. Ambient water quality
criteria for ammonia - 1984. EPA/440/5-85-001. U.S. EPA, Duluth, MN.
Wells, M.J.M., and J.L. Michael. 1987. Reversed-phase solid-phase
extraction for aqueous environmental sample preparation in herbicide
residue analysis. J. Chromatogr. Sci. 25:345-50.
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Equilibrium Partitioning
CHAPTER 5. EQUILIBRIUM PARTITIONING APPROACH
Christopher S. Zarba
U.S. Environmental Protection Agency
401 M Street S.W. (WH-585)
Washington, DC 20460
(202) 475-7325
The equilibrium partitioning (EP) approach focuses on predicting the
chemical interaction among sediments, interstitial water (i.e., the water
between sediment particles), and contaminants. Based on correlations with
toxicity, interstitial water concentrations of contaminants appear to be
better predictors of biological effects than do bulk sediment concentrations.
The EP method for generating sediment quality criteria is based on predicted
contaminant concentrations in interstitial water vs. chronic water quality
criteria. Chemically contaminated sediments are expected to cause adverse
biological effects if the predicted interstitial water concentration for a
given contaminant exceeds the chronic water quality criterion for that
contaminant.
1.0 SPECIFIC APPLICATIONS
Specific applications of EP-based sediment quality criteria are under
development. The primary use of EP-based sediment criteria will be to
identify risks associated with contaminants. Because the regulatory needs
vary widely among and within U.S. EPA offices and programs, EP-based
sediment quality criteria may be used in many different ways.
EP-based numerical sediment quality criteria would likely be used
directly to assess risk and applied in a tiered approach. In tiered
applications, concentrations of sediment contaminants that exceed sediment
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Equilibrium Partitioning
quality criteria would be considered as causing unacceptable impacts.
Further testing may or may not be required, depending on site-specific
conditions. Sediment contaminants at concentrations less than the sediment
criteria would not be of concern. However, sediments would not be considered
safe in cases where they are suspected to contain other contaminants 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 bioassays, could be required to determine
whether the sediment is safe. It is likely that such testing would
incorporate site-specific considerations.
1.1 Current Use
Specific regulatory uses of EP-based sediment quality criteria are
under development. The method is presently being reviewed by the U.S. EPA
Science Advisory Board to determine its suitability for generating sediment
criteria for non-ionic contaminants. This review should be completed prior
to establishing any formal framework for the application of sediment
criteria. (The EP approach was presented to the Science Advisory Board on 2
February 1989. Their report is expected in July 1989.) The range of
potential applications of the EP. approach is large because the approach
accounts for contaminant bioavailabil ity and can be used to evaluate most
sediments.
Interim sediment criteria values have been developed for a variety of
organic compounds using the EP approach. In a pilot study at six Superfund
sites at which site characterization and evaluation activities were
undertaken, the interim criteria were used in the following ways:
• Identify the extent of contamination
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Equilibrium Partitioning
• Assess the risks or potential risks associated with the
sediment contamination
• Identify the environmental benefit associated with a variety
of remedial options.
In addition, the State of New York has used interim EP-based sediment
criteria to evaluate the potential effects of sediment contaminants found
in aquatic habitats in that state.
1.2 Potential Use
Potential applications of the EP approach include a variety of ongoing
activities by the U.S. EPA. EP-based sediment quality criteria could play a
major role in the identification, monitoring, and cleanup of contaminated
sediment sites on a national basis. They could also be used to ensure that
uncontaminated sites remain uncontaminated. In some cases, such sediment
criteria alone will be sufficient to identify 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.
EP-based sediment criteria will be particularly 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 concentrations are being approached, or have been exceeded.
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 opportunity to take corrective action before adverse
impacts occur.
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Equilibrium Partitioning
Although sediment criteria developed using the EP approach are similar
in many ways to existing water quality criteria, their applications may
differ substantially. In most cases, contaminants 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 alleviate the problem. Safe removal,
treatment, or disposal of contaminated sediments can also be difficult and
expensive. For this reason, it is anticipated that EP-based sediment
criteria will rarely be used as mandatory cleanup levels. Rather, they will
be used to predict or identify the degree and spatial extent of problems
associated with contaminated areas, and thereby facilitate regulatory
decisions.
2.0 DESCRIPTION
2.1 Description of Method
Concentrations of contaminants in the interstitial water (i.e., the
water between the sediment particles) correlate very closely with toxicity,
whereas concentrations of contaminants bound to the sediment particles do
not. The EP 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, interstitial water, and biota are better understood for
some kinds of chemicals than for others. Concentrations of manganese oxide,
iron oxide, iron sulfide, and organic carbon are the primary factors that
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Equilibrium Partitioning
control phase associations, and therefore bioavailability, of trace metals
in sediments. However, models that can use these factors to predict
research and are not fully developed. Mechanisms that control the parti-
tioning of polar organic compounds are also poorly understood. However,
polar organic contaminants are not generally considered to be a significant
problem in sediments. Partitioning of non-ionic organic compounds between
sediments and interstitial water is highly correlated with the organic
carbon content of sediments. Also, the toxicity of non-ionic organic
contaminants in sediments is highly dependent on their interstitial water
concentrations. Consequently, to date, the EP approach is well developed
for non-ionic organic contaminants and is in the process of development for
trace metals.
Interstitial water concentrations can be calculated using partition
coefficients for specific non-ionic organic chemicals and criteria continuous
concentrations from WQC documents. The sediment quality criterion for a
specific chemical is defined as the solid phase concentration that will
result in an uncomplexed interstitial 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 non-ionic organic contaminants is as
follows:
rSQC - Kp * cWQC
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Equilibrium Partitioning
where:
cWQC « Criterion continuous concentration
rSQC - Sediment quality criterion (ug/kg sediment)
Kp - Partition coefficient for the chemical (L/kg sediment) between
sediment and water.
The method for calculating sediment quality criteria using the EP approach
for contaminants other than non-ionic organic contaminants is under
development.
2.1.1 Objectives and Assumptions--
Three principal assumptions underlie use of the EP-based approach to
establish sediment quality criteria:
• For sediment-dwelling organisms, the uncomplexed interstitial
water concentration of a chemical correlates with observed
biological effects across sediment types, and the concen-
tration at which effects are observed is the same as that
observed in a water-only exposure (see Document 18 in
Section 5.0)
• Partitioning models permit calculation of uncomplexed
interstitial water concentrations of the chemical phases of
sediments controlling availability
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Equilibrium Partitioning
• Benthic organisms exhibit a range of sensitivities to
chemicals that is similar to the range of sensitivities
exhibited by water column organisms (see Document 18 in
Section 5.0).
Data exist .supporting each of these assumptions.
2.1.2 Level of Effort-
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.
2.1.2.2 Types of Data Required—Analyses are needed to determine the
concentrations of the contaminants of concern in the sediment (bulk sediment
analysis), and the concentrations of organic carbon in the sediment.
2.1.2.3 Necessary Hardware and Skills—The investigator must be able
to design an appropriate sampling study, conduct bulk sediment analyses,
operate a pocket calculator, and understand developed values and what they
protect.
2.1.3 Adequacy of Documentation—
The method is very well documented (see Section 5.0).
2.2 Applicability of Method to Human Health. Aquatic Life, or Wildlife
Protection
Sediment criteria can be protective"of human health, aquatic life, and
wildlife. At present, only interim sediment criteria values that are
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Equilibrium Partitioning
protective of aquatic life have been developed. EP-based sediment criteria
are derived directly from water quality criteria. Sediment criteria derived
using water quality criteria are designed to be similar in their levels of
protection, and would be as protective of human health, aquatic life, and
wildlife as are water quality criteria.
2.3 Ability of Method to Generate Numerical Criteria for Specific Chemicals
The EP method generates numerical criteria for specific chemicals.
Interim sediment quality criteria have been developed for the following
chemicals:
PAH
Acenaphthene
Aniline
Phenanthrene
Pesticides
Chlordane
Chlorpyrifos
DDT
Dieldrin
Endrin
Ethyl
Parathion
Heptachlor
Heptachlor epoxide
Gamma-hexachlorocyclohexane (lindane)
PCBs.
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Equilibrium Partitioning
Techniques for developing sediment criteria for metal contaminants are under
development at U.S. EPA laboratories and by contractors.
3.0 USEFULNESS
3.1 Environmental Aoolicabilitv
One of the principal reasons for selecting the EP 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.
3.1.1 Suitability for Different Sediment Types--
Although aspects of the EP method are still under development, it is
expected that sediment criteria for non-ionic contaminants developed using
this approach will be applicable to all types of sediments found in both
freshwater and marine environments. Additional work is needed to clarify
the best use of the EP approach for sediments with less than 0.5 percent
organic carbon.
3.1.2 Suitability for Different Chemicals or Classes of Chemicals--
The EP method for developing sediment criteria has been modified for
different types of contaminants. Non-ionic, ionic, and metal contaminants
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 criteria for non-ionic organic contaminants
has been well developed and is under peer review. The technical approach
for developing sediment criteria for metal contaminants is under development
and is expected to undergo peer review in 1991. Ionic contaminants are not
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Equilibrium Partitioning
believed to cause major problems in sediments, but work plans for sediment
criteria development methods for these compounds have been written.
3.1.3 Suitability for Predicting Effects on Different Organisms--
As indicated above (see Section 2.1), the EP approach is based on
predicted interstitial water concentrations of non-ionic organic con-
taminants, and comparisons of these concentrations with chronic water
quality criteria. Typically, water quality criteria are based on toxicity
information (e.g., median lethal or median effective 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 EP approach can be compared with a range of toxic
effects values that are representative of the different kinds of organisms
upon which water quality criteria are based.
3.1.4 Suitability for In-Place Pollutant Control--
The EP method is suitable for in-place pollution control because it can
be used to identify locations where concentrations of individual contaminants
are causing adverse effects. Target cleanup levels can be identified, and
the success of cleanup activities can be determined.
3.1.5 Suitability for Source Control --
The EP 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.
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Equilibrium Partitioning
3.1.6 Suitability for Disposal Applications--
The EP 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.
3.2 General Advantages and Limitations
The EP approach offers the following advantages:
• It is consistent with existing water quality criteria
• It relates risks to specific substances and it can be used-to
identify probable species at risk
• It is applicable across all types of sediments and in all
types of aquatic environments, including lentic, lotic,
marine, and estuarine environments
• Only site-specific chemistry data are needed
• Site-specific or station-specific sediment criteria can be
calculated as soon as sediment 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 regulatory mechanisms
with little or no need for additional staffing or skills
• The equilibrium partitioning theory upon, which it is based is
well developed
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Equilibrium Partitioning
• It can be modified easily to accommodate site-specific
circumstances
• It can be used to identify risks to humans and wildlife that
may occur as a result of bioaccumulation
• It identifies the degree of sediment contamination, and
permits an assessment of whether contaminant concentrations
are approaching an effects level.
The EP approach is limited in the following ways:
• Sediment criteria developed using this approach do not
address possible synergistic, antagonistic, or additive
effects of contaminants
• Interim sediment criteria presently exist for only 12
contaminants
• The technical approach for developing sediment criteria for
metal contaminants is still under development
• Sediment quality criteria for non-ionic chemicals apply to
sediments that have an organic carbon concentration >0.5
percent
• Sufficient water-only toxicity data do not exist for all
contaminants of concern.
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Equilibrium Partitioning
3.2.1 Ease of Use--
The calculation of site-specific sediment criteria is relatively easy,
provided that sediment chemistry data adequately characterizing the site and
water quality criteria protective of the desired organism are available.
3.2.2 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.
3.2.3 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.
3.2.4 Level of Acceptance--
The EP 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 involved with the selection of the EP 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.
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Equilibrium Partitioning
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.
3.2.6 Level of Effort Required to Generate Results--
The necessary level of effort varies substantially from site to site,
and is dependent on many factors. Compared with other methods, the EP
method generates results quickly and more cost-effectively. No site-
specific biological data are required.
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 information. 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 fall within the
degrees of uncertainty would require more detailed interpretation, and
possibly additional testing.
3.2.8 Degree of Environmental Applicability--
EP-based sediment quality criteria can be applied directly to any
contaminated sediment containing >0.5 percent ionic carbon and non-ionic
chemicals for which criteria are available. Extensive data analysis and
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Equilibrium Partitioning
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 EP method can be considered environmentally
applicable in some cases. Because a wide variety of contaminated sediment
sites exist, absolute statements regarding environmental applicability are
difficult to make. However, the EP method would be appropriate in many
situations to predict bioavailability, bioaccumulation, and biological
effects.
3.2.9 Degree of Accuracy and Precision--
Each sediment criterion value developed using the EP 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 EP method are those associated with partition coeffi-
cients. Hence, each developed sediment criterion should be bracketed with
uncertainty, thereby providing decision-makers with a greater understanding
of the meaning of the developed values.
4.0 STATUS
The method for developing sediment criteria for non-ionic organic
contaminants has been developed and is currently being reviewed by the
U.S. EPA Science Advisory Board. Final comments are expected by July 1989.
Guidelines and guidance on the development and use of sediment criteria are
in early stages of 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-developed
for peer review by 1991.
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Equilibrium Partitioning
4.1 Extent of Use
Specific regulatory uses for EP-based sediment quality criteria have not
been established. A formal framework for the application of sediment
criteria is not expected until the U.S. EPA Science Advisory Board completes
its review. The range of potential applications is very large because the
need for evaluating potentially contaminated sediments arises in many
contexts.
Interim sediment criteria values were developed for a variety of organic
compounds. These values were used in a pilot study at six Superfund sites
where site characterization and evaluation activities were conducted. The
interim criteria were used in three ways:
• To identify the extent of contamination
• To assess the risks associated with sediment contamination
• To identify 'the environmental benefits associated with a
variety of remedial options.
The State of New York has also used interim sediment criteria to evaluate
the potential effects of several contaminants found in sediments in state
waters.
4.2 Extent to Which Approach Has Been Field-Validated
Field data were used to compare predicted effects with actual field
effects. The comparison was conducted by developing Screening Level
Concentration for various contaminants and organisms. A pilot field
verification study is underway in Puget Sound, where field sediments are
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Equilibrium Partitioning
being used to conduct laboratory experiments. Additional field verification
of this method is needed, and will be conducted in FY90.
4.3 Reasons for Limited Use
The EP method is not commonly used for the following reasons:
1. It has been developed only recently, and sufficient time has
not elapsed for the principles to be understood and used by
others
2. The U.S. EPA Science Advisory Board review of this method has
not been completed
3. The U.S. EPA has not yet developed and issued guidance on the
use of this method
4. The EP method has not yet been formally adopted by EPA.
4.4 Outlook for Future Use and Amount of Development Needed
This method is the only procedure for derivation of sediment quality
criteria that is generic across sediments, accounts for bioavailability of
chemicals, and relates effects to specific chemicals. Therefore, it is
likely that EP-based sediment quality criteria will be used much as water
quality criteria are being used to define environmentally acceptable
concentrations. Sediment quality criteria along with sediment toxicity
tests analogous to water quality criteria and whole effluent toxicity tests
could play major role in U.S. EPA's regulations of contaminated sediment.
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Equilibrium Partitioning
5.0 DOCUMENTS
0) Initial Evaluation of Alternatives for Development of Sediment Related
Criteria for Toxic Contaminants in Marine Waters (Puget Sound) 10/83 Phase I:
Development of Conceptual Framework Phase II: Development and Testing of the
Sediment-Water Equilibrium Partitioning Approach
1) Background and Review Document on the Development of Sediment Criteria
6/85
2) Sediment Quality Criteria Development Workshop 2/85
3) National Perspective on Sediment Quality 7/85
4) Elaboration of Sediment Normalization Theory for Nonpolar Hydrophobic
Organic Contaminants 1/86
5) Protocol for Sediment Toxicity Testing For Nonpolar Organic Compounds 2/86
6) An Activity-Based Model for Developing Sediment Criteria for Metals: I.
A New Approach 6/86
7) Sediment Quality Criteria Validation: Calculation of Screening Level
Concentrations from Field Data 7/86 Attachment: Recalculation of Screening
Level Concentrations for Nonpolar Organic contaminants in Marine Sediments
12/87
8) Guidance for Sampling of and Analyzing for Organic Contaminants in
Sediments 1/87
9) Sediment Quality Criteria for Metals: III Review of Data on Re-complexa-
tion of Trace Metals by Particulate Organic Carbon 1/87
10) Regulatory Applications of Sediment Criteria 6/87
11) Evaluation of the Equilibrium Partitioning Theory for Estimating the
Toxicity of the Nonpolar Organic Compound DDT to the Sediment Dwelling
Amphipod Rhepoxynius Abronius 8/87
12) Sediment Quality Criteria for Metals: IV Surface Complexation and
Acidity- Constants for Modeling Cadmium and Zinc Adsorption on to Iron
Oxides 8/87
13) Sediment Quality Criteria for Metals: II Review of Methods for Quanti-
tative Determination of Important Adsorbents and Sorbed Metals in Sediments
8/87
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Equilibrium Partitioning
14) Sediment Quality Criteria Methodology Validation: Uncertainty Analysis
of Sediment Normalization Theory for Nonpolar Organic Contaminants 11/87
15) Reconnaissance Field Study for Verification of Equilibrium Partitioning:
Nonpolar Hydrophobic Organic Chemicals 11/87
16) Sediment Quality Criteria for Metals: V Optimization of Extraction
Methods for Determining the Quantity of Sorbents and Adsorbed Metals in
Sediments 12/87
17) Interim Sediment Criteria Values for Nonpolar Hydrophobic Organic
Contaminants 5/88
18) Briefing Report to the EPA Science Advisory Board on the Equilibrium
Partitioning Approach to Generating Sediment Quality Criteria 1/89
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Tissue Residue
CHAPTER 6. TISSUE RESIDUE APPROACH
Anthony R. Carlson
U.S. Environmental Protection Agency,
Environmental Research Lab-Ouluth
6201 Congdon 81vd
Ouluth, MN 55804
(218) 720-5523
FTS 780-5523
Philip M. Cook
U.S. Environmental Protection Agency,
Environmental Research Lab-Ouluth
6201 Congdon Blvd
Duluth, MN 55804
(218) 720-5553
FTS 780-5553
Henry Lee II
U.S. Environmental Protection Agency,
Environmental Research Lab-Newport
Marine Science Drive
Newport, OR 97365
FTS 867-4042
In the tissue residue approach, sediment chemical concentrations that
will result in acceptable residues in exposed biotic tissues are determined.
Concentrations of unacceptable tissue residues may be derived from toxicity
tests performed during generation of chronic water quality criteria, from
bioconcentration factors derived from the literature or generated by
experimentation, or by comparison with human health risk criteria associated
with consumption of contaminated aquatic organisms. The tissue residue
approach generates numerical criteria and is most applicable for non-polar
organic and organometallic compounds.
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Tissue Residue
1.0 SPECIFIC APPLICATIONS
1.1 Current Use
Tissue residues of chemical contaminants in aquatic organisms,
particularly fish, are frequently used as measures of water quality in both
freshwater 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 al. 1985). Non-polar organic chemicals, which may bioaccumulate
to levels that are toxic to organisms or render the 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 the amount of reduction of 2,3,7,8-TCDO concen-
tration in Lake Ontario sediments that will result in attainment of
acceptable TCOO levels in fish (Cook et al. 1989). The acceptable sediment
TCDO concentration is being used as a sediment criterion to determine the
remedial action necessary to reduce the incremental loading of TCDD from the
Hyde Park Superfund site to Lake Ontario (Carey et al. 1989). Tissue
residues of benthic organisms have also been used in some regulatory actions,
such as the assessment of bioaccumulation potential of dredged materials.
1.2 Potential Use
Although tissue residues have been used more commonly to determine the
potential for bioaccumulation of chemical contaminants from sediments and
dredged materials, they also provide an excellent measure of "effective
exposure dose" - a measure of an organism's actual exposure ouer time to a
pollutant of concern. This exposure measure may be related to the dose
expected at the water quality criterion or directly to the potential for
producing chronic toxic effects. Given the ability to measure or predict
tissue residues resulting from exposures in contaminated sediment systems,
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Tissue Residue
it is possible to establish sediment criteria based on residue-toxicity
effects relationships. These relationships 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 failure of Great Lakes lake trout (Salvelinus namaycush) to
reproduce and bioaccumulation of TCDD and non-ortho substituted PCBs (Mac
1988). Laboratory studies have shown significant mortality of larvae when
lake trout ova contain as little as 50 ppt 2,3,7,3-TCDD (Walker et al.
1988). This residue level is found in Lake Ontario lake trout which have
not successfully accomplished natural reproduction for many years. On the
basis of TCDO toxic equivalents for organochlorine components having the
same mode of toxic action, residues in lake trout from Lake Ontario and Lake
Michigan may provide a measure of the reduction in sediment contamination
necessary to reduce fish tissue concentrations to a presumed reproductive
impairment threshold. The same approach can be used for benthic organisms
that may have greater inter-site variations in residue levels than do fish
because of their more intimate association with sediments.
2.0 DESCRIPTION
2.1 Description of Method
The tissue residue approach involves the establishment of safe
sediment concentrations for individual chemicals or classes of chemicals by
determining the sediment chemical concentration that will result in
acceptable tissue residues. This process involves two steps: 1) linking
toxic effects to residues (i.e., dose-response relationships), and 2) linking
chemical residues in specific organisms to sediment chemical contamination
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Tissue Residue
concentrations (i.e., exposure relationships). Methods to derive unaccept-
able tissue residues include at least three approaches:
• The water quality criterion-residue approach
• The experimental approach
• The human health approach
Each of these approaches is described briefly below.
Water Quality Criterion-Residue Approach--
A rapid approach for determining acceptable concentrations of tissue
residues involves establishing maximum permissible tissue concentrations
(MPTC) expected for organisms at the chronic water quality criterion concen-
tration previously established for a specific pollutant. MPTCs, when not
available through residue measurements obtained with toxicity tests used as
a basis for the water quality criteria, can be obtained by multiplying the
water quality criterion 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 coef-
ficient or a bioconcentration kinetic model. Thus, the absence of a water
quality criterion for a chemical does not eliminate this approach as long as
appropriate 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 advantage of directly determining the tissue residue-
toxic effects linkages, it can be relatively time-consuming and costly to
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Tissue Residue
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 the bulk sediment toxicity test approach to deriving
sediment quality criteria (see Chapter 2).
Human Health Approach-
Human health risk from consumption of freshwater fish or seafood 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 1x10"^)
and then back-calculating 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 for carcinogenic compounds
(e.g., PCBs, dioxins, benzo(a)pyrene) that are lower than those derived from
ecological endpoints.
The choice of method to determine a quantitative tissue residue-
sediment contamination level relationship depends on the specific pollutants,
organisms, and water systems of concern, as well as the regulatory approach
(e.g., remedial action, wasteload allocation, Superfund enforcement). The
linkage between organism residue and sediment chemical concentration can be
made from site-specific measurements of sediment-organism partition
coefficients (Kuehl et al. 1987); fugacity or equilibrium partitioning model
(Clark et al. 1988); predictions of organism residues; or pharmacokinetic-
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
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Tissue Residue
contaminants because of their persistence and tendency to exert long-term
rather than episodic bioaccumulation and toxic effects.
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 equilibrium partitioning approach (Chapter 5), this
method does not require that the organism be at thermodynamic equilibrium
with respect to the sediment contamination level. The site-specific residue
approach is powerful because it does not require knowledge of bioavailability
relationships 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 concen-
trations over time (e.g., as a result of remedial actions, wasteload
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-toxicity 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.
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
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Tissue Residue
BCFs, the level of effort depends on the availability of non-water quality
criteria residue criteria and the complexity of the sediment contaminant
mitigation approach to be used. Relatively little effort is required to
determine the degree to which sediment contamination concentrations must be
reduced for single chemicals in well-mixed systems where fish residues are
uniformly unacceptable for human consumption. Much more effort is required
for systems having sediment contamination "hot spots" where resident aquatic
organisms 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 sediment contaminants that are readily
metabolized to chemicals of greater toxicity that are responsible for the
observed adverse effects. In some cases, residue-toxic effects data would
incorporate the effects of toxic metabolites.
2.1.2.1 Type of Sampling Required—Surface sediment samples must be
analyzed for chemical contaminants of interest. Interstitial 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 capacity are
useful for defining sediment-to-biota relationships for different sites
within an ecosystem. 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.
2.1.2.2 Methods--The tissue residue approach, as discussed above in
Section 2.0, depends on determining residues in aquatic organisms that are
unacceptable on the basis of toxicity to the organism or unsuitabi1ity for
human or animal consumption as food. The linkage of sediment contaminant
concentrations to organism residues is possible through a number 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
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Tissue Residue
residue-toxic effects data, the contamination history (in-place pollutant
problem vs. a continuing or projected sediment contamination problem), and
characteristics of the impacted ecosystem. The construction of compre-
hensive, systematic strategies for all potential sediment contamination
assessments will be achieved through further research and development.
Toxicity identification evaluation (TIE) procedures- (see Chapter 4)
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, chemicals (e.g.,
ammonia) in sediments. TIE procedures can distinguish between potential
metal, non-polar 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 toxicity
in the system.
Once potentially toxic, bioaccumulative contaminants are identified,
either in sediment or in aquatic organisms associated through exposure to
sediments, the toxicological significance of site-specific sediment-to-biota
contaminant partition factors can be assessed. Conservative generic
sediment quality criteria can be generated from residue-toxicity relation-
ships by assuming equilibrium partitioning between the binding fractions of
organisms and sediments (e.g., lipid and sediment organic carbon for non-
polar organic chemicals).
2.1.2.3 Types of Data Reouired--The tissue residue method requires
identification of chemicals in the sediment that are likely to be associated
with chronic environmental effects. An indirect method for identifying such
chemicals and their locations is to screen aquatic organisms for residues as
in the National Dioxin Study or the National Bioaccumulation Study sponsored
by U.S. 'EPA (1987b) Office of Water Regulations and Standards. When
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Tissue Residue
toxicity data are not available, either laboratory dose-response experiments
or quantitative structure-activity predictions can be used to establish the
toxicological 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
residues associated with no-effect and the lowest observable effect
concentrations are needed when the sediment criterion is not based on a
human health standard.
2.1.2.4 Necessary Hardware and Skills--Sediment and tissue analyses
require commonly available chemical analytical capabilities. Some chemicals
require advanced instrumental analytical techniques, such as high resolution
gas chromatography/mass spectrometry.
2.1.3 Adequacy of Documentation--'
The use of tissue residues to establish sediment criteria on the basis
of human health effects 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.
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. Because of this
relationship, the tissue residue method provides a direct link between human
health and sediment criteria development. Tissue 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
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1
Tissue Residue
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 connec-
tions from benthos to birds experiencing eggshell thinning or genotoxic
effects. The tissue residue concentration serves as a quantitative measure
of sediment contaminant bioavailabil ity, which may differ as a function of
ecosystem, sediment, water, food chain, and species characteristics.
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 (U.S. EPA 1980a) and endrin
(U.S. EPA 1980b) have been used to establish water quality criteria on the
basis of human health risks. The DOT and PCB water quality criteria are
based on toxic effects in birds and animals as a function of fish residues
(U.S. EPA 19SOc,d). Tissue residues of organometallic chemicals such as
methyl mercury (U.S. EPA 1984) and elements such as selenium (U.S. EPA
19S7a) have been used to establish water quality criteria and/or 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
equilibrium partitioning approach.
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Tissue Residue
3.0 USEFULNESS
3.1 Environmental Applicability
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 variability in the water body tends to be
averaged.
3.1.2 Suitability for Different Chemicals or Classes of Chemicals--
This approach is most applicable to non-polar organics and organo-
metallics that bioaccumulate, are slowly metabolized, and exert chronic
toxic 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-residue 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
mrxtures of sediment contaminants depends on the development of chemical
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Tissue Residue
mixture toxic dose-response relationships where dose is indicated by tissue
residue levels.
3.1.3 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 due to toxic
effects.
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 MPTC 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 monitoring of changes in toxicity
following remedial action. By weighing the relative toxicity of bioaccumu-
lated pollutants (e.g., by using "dioxin equivalents"), evaluation of
tissue residue concentrations can help identify the pollutants most likely
responsible for toxicity and their additive contribution to total sediment
toxicity. This information could then be used to design the most appropriate
and cost-effective mitigation response.
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Tissue Residue
3.1.5 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 TCDO-Hyde Park Superfund case example described in Section
1.1 demonstrates the suitability of this approach for establishing source
controls. The site-specific nature of this approach provides strong support
for establishing controls on existing point and nonpoint sources of
sediment contamination.
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 toxicokinetic bioaccumulation models or by conducting laboratory
bioaccumulation tests on the dredged material. If adverse effects on fishes,
wildlife, or 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 sediments 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.
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Tissue Residue
3.2 General Advantages and Limitations
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 develop-
ment of a tissue residue toxicity database from laboratory 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 capabili-
ties that do not present unique problems. One potential 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
bioaccumulation tests on field- collected sediment or by placing caged
benthic organisms in the field.
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) validation of the relation-
ships 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 difficulty in identifying sensitive organisms. The establish-
ment of a sediment criterion based on fish residue levels acceptable for
protection of human health generally incurs low analytical costs when only a
few reference sediment sites are needed to characterize the system of
concern.
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Tissue Residue
3.2.3 Tendency to be Conservative--
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 stricter than necessary to protect resident biota.
3.2.4 Level of Acceptance--
The tissue residue approach is accepted as a basis for regulatory
decisions such as the establishment of water quality criteria for the
protection of aquatic life and its uses. The direct prediction of chronic
toxic effects from measured or predicted 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 exceeds that indicated as deleterious in previous experiments.
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 potentially toxic sediment contaminants are known and
residue-toxicity relationship data are available. If extensive laboratory
work is needed to determine 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.
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Tissue Residue
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).
3.2.7 Degree to Which Results Lend Themselves to Interpretation--
Tissue residues that exceed concentrations considered safe for human
exposure through seafood consumption require no interpretation when 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 sediment-to-biota chemical partitioning. This
interpretation 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 contami-
nation can be used without elaborate interpretation to determine compliance
of sediments with sediment quality criteria.
3.2.8 Degree of Environmental Applicability--
The use of site-specific tissue residues as quantitative exposure
biomarkers eliminates uncertainties associated with chemical bioavailability;
exposure duration, frequency, and magnitude; 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.
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Tissue Residue
3.2.9 Degree of Accuracy and Precision--
Sediment and tissue residue chemical concentrations can be determined
accurately and precisely 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.
4.0 STATUS
4.1 Extent of Use
Use of tissue residues to establish sediment criteria on the basis of
human health effects have been documented. Tissue residues have also been
used to derive water quality criteria for the protection 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 consump-
tion 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.
4.2 Extent to Which Approach Has Been Field-Validated
Sediment TCDO 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 TCDD in lake trout,
regardless of route of uptake. Declines in DDT residues in fish and birds
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Tissue Residue
since its use was banned are associated with declining surficial sediment
concentrations in the Great Lakes, the Southern California Bight, and
elsewhere. Although other examples of studies validating the residue
approach for single chemicals are available, its use for complex mixtures of
chemicals in sediments to predict sediment safe contaminant concentrations
with ecosystem protection in mind has not been validated.
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 U.S. EPA
• Aquatic toxicology has only recently progressed to an
understanding of residue-based dose-response relationships
for sediment contaminants
• Regulatory agencies, including U.S. 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.
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, especially where a single chemical or toxicologically related family
of chemicals is of concern. Guidance documents should be written and
reviewed. Tissue residue criteria should be accumulated systematically for
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Tissue Residue
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 appli-
cation to those chemicals in the aquatic environment that are long-term
sediment contaminants having chronic toxicity potential.
5.0 REFERENCES
Batterman, A.R., P.M. Cook, K.B. Lodge, D.B. Lothenbach, and B.C. Butter-
worth. In press. Methodology used for a laboratory determination of
relative contributions of water, sediment and food chain routes of uptake
for 2,3,7,8-TCDD bioaccumulation by lake trout in Lake Ontario. Chemosphere.
Carey, A.E., N.S. Shifrin, and A.C. Roche. 1989. Lake Ontario TCDD
bioaccumulation study final report. Chapter 1: introduction, background,
study description and chronology. Gradient Corporation. Cambridge, MA.
17 pp.
Clark, T., K. Clark, S. Pateson, 0. Mockay, and R.J. Norstrom. 1988.
Wildlife monitoring, modeling and fugacity. Environ. Sci. Technol.
22:120-127.
Cook, P.M., A.R. Batterman, B.C. Butterworth, K.B. Lodge, and S.W. Kohlbry.
1989. Laboratory study of TCDO bioaccumulation by lake trout from Lake
Ontario sediments, food chain and water. U.S. Environmental Protection
Agency, Environmental Research Laboratcry-Duluth, Duluth, MN. 112 pp.
Endicott, 0., W. Richardson, and 0. OiToro. 1989. Lake Ontario TCDD
modeling report. U.S. Environmental Protection Agency, Large Lakes Research
Station, Environmental Research Laboratory-Duluth, Grosse He, MI. 94 pp.
Krahn, M.M., L.D. Rhodes, M.S. Myers, L.K. Moore, W.O. MacLeod, and
D.C. Mai ins. 1986. Associations between metabolites of aromatic compounds
in bile and the occurrence of hepatic lesions in English sole (Parophrys
velulus) from Puget Sound, Washington. Arch. Environ. Contam. Toxicol.
15:61-67.
Kuehl, D.W., P.M. Cook, A.R. Batterman, 0. Lothenbach, and B.C. Butterworth.
1987. Bioavailability of polychlorinated dibenzo-p-dioxins and dibenzofurans
from contaminated Wisconsin River sediment to carp. Chemosphere 16:667-679.
Mac, M.J. 1988. Toxic substances and survival of Lake Michigan salmonids:
field and laboratory approaches, pp. 389-401. In:. Toxic Contaminants and
Ecosystem Health. M.S. Evans (ed). Wiley & Sons.
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Tis: e Residue
Stephan, C.E., D.I. Mount, O.J. Hansen, J.H. Gentile, G.A. Ch )man, and
W.A. Brungs. 1985. Guidelines for deriving numerical national wa .-r quality
criteria for the protection of aquatic organisms and their us ;. PB85-
227040. National Technical Information Service, Springfield,
VA.
Thomann, R.V. 1989. Bioaccumulation model of organic chemical di
in aquatic food chains. Environ. Sci. Technol. 23:699-707.
U.S. Environmental Protection Agency. 1980a. Ambient wat
criteria for aldrin/dieldrin. EPA 440/5-80-019. NTIS number P
U.S. EPA, Washington, DC.
U.S. Environmental
criteria for endrin.
Washington, DC.
U.S. Environmental
criteria for DOT.
Washington, DC.
Protection Agency.
EPA 440/5-80-047.
19805.
NTIS number
Ambient wat
PB81-117582.
Protection Agency. 19SOc. Ambient wat
EPA 440/5-80-038. NTIS number PB81-117491.
U.S. Environmental Protection Agency.
criteria for polychlorinated biphenyls.
PB81-117798. U.S. EPA, Washington, DC.
1980d. Ambient wat
EPA 440/5-80-068. f
U.S. Environmental Protection Agency,
for mercury. EPA 440/5-84-026.
Washington, DC.
1984. Ambient water quali
NTIS number PB85-227452.
ributions
* quality
1-117301.
" quality
U.S. EPA,
- quality
U.S. EPA,
r quality
IS number
/ criteria
J.S. EPA,
U.S. Environmental Protection Agency. 1987a. Ambient wat r quality
criteria for selenium. EPA 440/5-87-006. NTIS number F J8-142237.
U.S. EPA, Washington, DC.
U.S. Environmental Protection Agency. 1987b.
Tiers 3, 5, 6, and 7. EPA 440/4-87-003.
Regulations and Standards, Washington, DC.
The national di
U.S. EPA, Offic
in study.
of Water
Walker, M.K., J.S. Spitbergen, R.E. Peterson, R.D. Quiney, and R. Olson.
1988. Effects of 2,3,7,3-tetrachlorodibenzo-p-dioxin (TCDD) in ?arly life
stages of lake trout. p. 112. In: Abstract, Meeting of ciety for
Environmental Toxicology and Chemistry, Arlington, VA.
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Freshwater Macroirwertebrate Benthic Community
Structure and Function
CHAPTER 7. FRESHWATER BENTHIC MACROINVERTEBRATE
COMMUNITY STRUCTURE AND FUNCTION
Wayne S. Davis
U.S. Environmental Protection Agency Region V
Environmental Sciences Division
536 S. Clark Street, Chicago, IL 60605
(312 or FTS) 886-6233
Joyce E. Lathrop
Stream Assessments Company
P.O. Box 609
Villa Park, IL 60181
The community structure and function of benthic macroinvertebrates are
used extensively to evaluate water quality and characterize impacts in lotic
(flowing water) and lentic (standing water) freshwater ecosystems. (Marine
benthic community structure is discussed in Chapter--S). Benthic macro-
invertebrates are relatively sedentary organisms that inhabit or depend upon
the sedimentary environment for their various life functions. Therefore,
they are Sensitive to both lony-term and shuri-ieriji changes in sediment and
water quality. This chapter discusses assessment of benthic macroin-
vertebrates to determine sediment quality in conjunction with an integrated
approach for assessing the quality of the benthic environment. This
integrated approach utilizes sediment chemistry, sediment toxicity, and
benthic macroinvertebrate community 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
7-1
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
available, or being developed, to complement the chemical and toxicological
portions of the sediment quality assessment.
1.0 SPECIFIC APPLICATIONS
].1 Current Use
Freshwater benthic macroinvertebrate communities are used in the
following ways to assess sediment or water quality:
• Identification of the quality of ambient sites through a
knowledge of the pollution tolerances and life history
requirements of benthic macroinvertebrates
• 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 temporal trends
• Determination of spatial gradients of contamination for
source characterization.
1.1.1 Ecological Uses--
Benthic macroinvertebrate community structure and function assessments
have many different applications. Site-specific knowledge of surface water
quality, habitat quality, sediment chemistry, and sediment toxicity provide
the best context in which to interpret benthic community assessment data.
The objectives of each particular study should determine the types of
related data necessary. Alone, benthic macroinvertebrates can be used to
screen for potential sediment contamination based on spatial gradients in
7-2
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
community structure, but they should not be used alone to definitively
determine sediment quality. Benthic macroinvertebrates data must be
integrated with other available data to determine sediment quality. In a
"weight-of-evidence" approach, benthic macroinvertebrates may provide the
most important piece of information on sediment quality. Care must be
exercised to collect representative samples to minimize natural variations.
For example, collections should not be made after floods or"other physical
disturbances.
Benthic macroinvertebrate community structure and function have been
used extensively to characterize freshwater ambient conditions and impacts
from various sources. Documented changes in benthic community structure
have resulted from crude oil exposure in ponds and streams (Rosenberg and
Wiens 1976; Mozley 1978; Mozley and Butler 1978; Cushman 1984; Cushman and
Goyert 1984) and heavy metal contamination of lake sediments and streams
(Winner et al. 1975, 1980; Wentsel et al. 1977; Moore et al. 1979; Wiederholm
1984a, 1984b; Waterhouse and Farrell 1985). 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,
1 rt O7 1 flOO \ O/%M + k i /» j*«"t«vww* t*t 4 ¥ \t **arr\rmrae f n no c ^ i <
,-,,,. , -,.,... u^i i wi i i \» vwiiMiiwiiiwj i v; w p wi i .j w ^ ww p v> ^ - -
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).
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
1.1.2 Regulatory Uses--
Recently, benthic macroinvertebrate communities have been gaining use in
U.S. EPA (1988a,b) and state (Ohio EPA 1987a) regulatory programs as the
optimal measures of designated use attainment. They are suitable to
establish both narrative and numerical instream biological criteria. Among
the states implementing or developing instream benthic criteria are Arkansas
(Shakelford 1988), Florida (U.S. EPA 1988a), Maine (Courtemanch and Davies
1988), Minnesota (Fandrei, G., 1989, personal communication), Nebraska (Maret
1988), New York (Bode and Novak 1988), North Carolina (Penrose and Overton
1988), Ohio (Ohio EPA 1987a, 1987b), and Vermont (Fiske 1988). Under the
Clean Water Act, benthic macroinvertebrates are used for the following:
• Measurement of the restoration and maintenance of biological
integrity in surface waters (Section 101)
• Development of water quality criteria based on biological
assessment methods when numerical criteria~~for~ toxicity are
not established [Section 303(c)(2)(B)]
• Production of guidance and criteria based on biological
monitoring and assessment methods [Section 304(a)(3)]
• 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 that cannot attain designated uses without
additional NPS controls (Section 319)
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
• Listing of waters unable to support balanced aquatic
communities [Section 304(1)]
• Assessment of lake trophic states and trends (Section 314)
• Production of biennial reports on the extent to which waters
support balanced aquatic communities [Section 305(b)]
• Determination of the effect of dredge and fill disposal on
balanced wetland communities (Section 404).
Another important feature of this method is the ability to generate numerical
biological criteria for state water quality standards (Ohio EPA 1987a).
State development of biocriteria is being encouraged and supported by U.S.
EPA, and biocriteria policy and guidance documents are expected to be
published by the U.S. EPA Office of Water during FY89-90. In addition to the
Clean Water Act regulations, benthic macroinvertebrate community assessments
may be applied to Superfund evaluations of onsite and offsite impacts (U.S.
EPA 1989a,b). They may also be part of an applicable or relevant and
appropriate requirement (ARAR) if a state adopts standards for biocriteria
or a "no toxics in toxic amounts" narrative that utilizes benthrc macro-
invertebrates to determine compliance with those standards.
1.2 Potential Use
The use of benthic macroinvertebrates relating to sediment contamination
will be most successful when used with sediment chemistry and toxicity
results, as in the "integrated" Sediment Quality Triad approach (see Chapter
9). 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. Alone, benthic
macroinvertebrates can be used to screen for potential sediment contamination
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
and source identification by displaying spatial gradients in community
structure, but they should not be used alone to definitively determine
sediment quality or develop chemical-specific guidelines. Benthic macro-
invertebrate data must be integrated with other available data to determine
sediment quality using a "weight-of-evidence" approach.
2.0 DESCRIPTION
2.1 Description of Method
The benthic macroinvertebrate community structure and function
assessment involves the following steps:
1. Collection of benthic macroinvertebrates in the field
(artificial or natural substrates)
2. Identification to the lowest taxon necessary (varies depending
upon the study objectives)'
3. Quantification (e.g., taxa richness, number of individuals,
indicator organism count, structural indices and ratios,
functional characteristics of taxa)
4. Assessment of the relationship with other environmental
measurements (e.g., correlations, habitat requirements)
5. Comparison with a "reference" site (e.g., similarity indices,
nonparametric analyses)
6. Complete documentation of the study methods, results, and
discussion of the relevance of the data.
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
2.1.1 Objectives and Assumptions--
The primary objective of benthic macroinvertebrate community structure
and function analyses is to provide data and information to assist in
determining the quality of the sediment/water environment. This determi-
nation can then be used for the purposes described above in Section 1.0
(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 assumptions are implicit
in this objective:
• The benthic macroinvertebrates are relatively sedentary,
especially compared to fish communities, and they depend upon
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
dependent upon the benthic environment for completion 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 chemical exchange between the sediment and
the overlying water, and therefore tend to reflect sediment
quality as well as water quality
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
• The optimal use of benthic macroinvertebrates as sediment
quality indicators is as part of an integrated sediment
quality assessment approach utilizing sediment chemistry,
sediment toxicity, and benthic community structure and
function.
Equally important assumptions apply to actual benthic macroinvertebrate
sampling strategy, collection, identification, data reduction, interpretation
of results, and report preparation. It is assumed that all U.S. EPA-
supported studies have an adequate quality assurance program plan (QAPP) and
that all benthic macroinvertebrate community data are reproducible and
collected in a manner to minimize 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 Cincinnati-ERL or State-developed check-samples,
in addition to voucher collections from each study area and a
list of the taxonomic references used
• Data analysis techniques used to objectively assess the
data, including the structural and functional measures
• Nonparanietric or parametric (as appropriate) statistical
methods used to compare site results.
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
Each Regional U.S. EPA Quality Assurance Office can provide the details of
QAPP requirements.
2.1.2 Level of Effort--
The level of effort to conduct freshwater benthic macroinvertebrate
community studies is comparable with chemical/physical water quality
measurements and bioassays. However, rapid benthic community assessment
techniques can range from 1 to 5 h per site if laboratory identifications
are not required. As expected, the greatest time expenditure is in the
travel to and from the site and in the sorting and identification of the
organisms.
Separating the organisms from debris and sorting the organism can take
up to 15 h per sample, with an additional 12 h for identification for very
enriched sites with high numbers of individuals among several taxa. More
typically, the time spent would be about 3 h for sorting (more time for
dredge and artificial substrate samples and less time for dip-net samples),
2 h for preparing the samples (e.g., clearing and then mounting the
chironomids on microscope slides), and 6 h for identifying the organisms to
the lowest possible taxonomic level. An experienced taxonomist with
appropriate keys may average only 2-4 h per site. This typical time
equates to about 11 h per site after the samples 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.
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 "qualitative" sampling methods, but each method is
acceptable contingent upon how well it will satisfy study objectives,
reproducibility of the data, and consistency of collection. Although it is
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
not advisable to critically compare data collected with different sampling
techniques, any comparisons made should not involve taxa below the family
level. 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 upon the
objectives of the study. For example, if the objective is to determine the
number and types of taxa, rather than the number of individuals within each
taxon, then a study using a dip-net in various habitats within the site
would be considered quantitative. Examples of programs using data quantified
by methods other than surface area of the sampler or substrate include those
described by Pollard (1981), Hilsenhoff (1982, 1987, 1988), Cummins and
Wilzbach (1985), Bode and Novak (1988), Cummins (1988), Hite (1988), Lenat
(1988), Maret (1988), Penrose and Overton (1988), Plafkin et al. (1988), and
Shakelford (1988). The success of each sampling effort depends upon a
thorough understanding of the data quality objectives of that study and the
implementation of a quality assurance program.
In soft freshwater sediments, the most common method used to collect
benthic macroinvertebrates 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 upon the surface area of
the sampler. The smaller surface area samplers are most commonly used for
freshwater studies because of their relative ease of manipulation. The
Ekman dredge is not as effective in areas of vegetative debris, but is much
lighter than the Ponar and easier to use in softer s'ubstrates. Artificial
substrates (e.g., Hester-Oendy sampler using several 3-in plates and spacers
attached by an eyebolt, or substrate/rock-filled baskets) provide a
consistent habitat for the benthic macroinvertebrates 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 (Hester and Oendy 1962; Rosenberg and Resh 1982;
APHA et al. 1985; DePauw 1986; Ohio EPA 1987c). The major drawback to
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
artificial substrates is the 4-8 wk period required for instream coloni-
zation. At least two visits are required 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 (Grossman and Cairns 1974;
Beckett and Miller 1982; Shepard 1982; Chadwick and Canton 1983; Peckarsky
1986; Ohio EPA 1987c; Plafkin et al. 1988; Lenat 1988). If quantification
by sediment 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 sampled (Pollard and Kinney 1979; Pollard
1981). 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. Kick or
dip-net sampling techniques 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.
For sediment evaluations of lotic systems, a combination of artificial
substrate (e.g. Hester-Oendy) and natural substrate (dip-net) sampling is
recommended. This combination allows comparison of the benthic macroinverte-
brate communities independent of habitat, so that sediment/water quality
effects can be better assessed.
2.1.2.2 Methods—Most state environmental regulatory programs have a
QAPP describing the field methods and standard operating .procedures for
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
collecting and evaluating benthic macroinvertebrates (see previous discussion
in Section 2.1.1). 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
operating procedures from other existing programs should be used. Since
several different collection and analysis methods are used throughout the
country depending upon water body type (i.e., lotic vs. lentic), habitat
type, substrate type, and familiarity with specific methods, it is not
practical to recommend any single sampling method. The only general QA
requirement for the use of any one particular method is that the data be
reproducible, consistently used within the program, and applicable by
different investigators (U.S. EPA 1988a,b).
Sampling Strategy—Sampling strategies have been addressed by Sheldon
(1984), Mi Hard and Lettenmaier (1986), and Plafkin et al. (1988). To
detect spatial differences in sediment/water quality 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 (e.g., high quality or wild and
scenic streams). Preferably two upstream sites and three downstream sites
of the suspected pollutant sources 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. To complement the artificial substrate data, multi-
habitat dip-net sampling should be performed when the substrates are
deployed and retrieved.
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. Sampling should be done each year
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
during similar flow conditions and should not be conducted for at least
1-2 wk after a major rainfall.
Seasonal distributions are always a concern for ensuring the collection
of a representative sample. Therefore, routine sampling or monitoring is
optimal during the seasons indicated in Plafkin et al. (1988), and'long-term
monitoring should strive for consistent sampling seasons. Participants in
the benthic macroinvertebrate discussion group at the 1987 National Workshop
on Instream Biological Monitoring and Criteria agreed that the optimal time
of year for sampling in lotic systems was during the latter part of the
seasons 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 QAPP.
The following recommendations are somewhat arbitrary, but provide a
beginning to implementation of a QA program. When using a new method, at
least five replicate samples should be taken per site and analyzed by at
least two investigators before the techniques are applied to the program.
Coefficients of variation among the samples should be below 50 percent.
Although many investigators collect separate replicates and then composite
them into one sample, it should be standard practice to analyze each
individual replicate for at least 10 percent of the sample sites (or at
least one site per study) to check variability. For multihabitat dip-net
sampling within a site, the chance of not obtaining a representative sample
is greatly decreased, especially since enumeration methods are not likely to
be used to quantify the results. If multihabitat samples are routinely
composited, then for 10 percent of the sample sites (or at least one per
survey), two samples should be collected and analyzed by two investigators.
This approach would provide four replicates (two visits by two investigators)
from each site sampled.
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
Statistical derivation of the number of samples required to decrease the
variability of the data have been discussed by Weber (U.S. EPA 1973), Green
(1978), Merritt et al. (1984), and Resh and Price (1984). These methods
rely on prior knowledge of the variability of the data. This prior
knowledge is often not available or 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 distribution such as normal or log-normal,
which is not necessarily true for biological samples. Also, the variability,
as measured by the variance or standard deviation, would be different for
each descriptive index analyzed (e.g., humber of taxa vs. number of
individuals).
Field Methods — Field sampling methods are adequately addressed in many
manuals including the U.S. EPA (1973) biological field methods manual, the
ASTM (1988) methods for sampling benthic macroinvertebrates, Ohio EPA's
(1987c) field methods manual, Standard Methods (APHA et al. 1985), U.S.
EPA's rapid bioassessment protocols (R8P) (Plafkin et al. 1988), and U.S.
EPA's (1987) Superfund field compendium. The following decisions will need
to be made once the sample gear is chosen:
• Whether or not samples will be picked from debris and sorted
in the field
• Which preservative should be used
• Whether or not a stain (e.g., rose bengal) will be added to
the sample to facilitate separating the organisms from debris
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
• Whether or not the samples need to be shipped and whether or
not they require a chain-of-custody form (as in Superfund
samples)
• The type of sample containers.
Sieving of samples is addressed in the next paragraph. If sieving is
performed in the field, a #30-mesh sieve (0.595-mm openings) is recommended,
with the materials that pass through then sieved through a #40-mesh sieve
(0.425-mm openings). Separation of organisms and debris should be performed
in the laboratory. Depending on the desired level of effort for the study,
however, separation can be done in the field by placing the sample in a white
enamel pan to provide a bright background to see the organisms. Rose
bengal (200 mg/l) can be used to stain the organisms to facilitate separation
from detritus and other debris (APHA et al. 1985). For routine benthic
sampling, special fixatives are not necessary and the organisms should be
preserved in a 70 percent ethanol solution. Formalin, which is an excellent
fixative, is no longer recommended because of health .effects. The organisms
should always be collected in plastic, shatter-proof containers.
Sorting—There are many discussions elsewhere of techniques for sample
sorting and preparation of slides for identification. For example, Weber
(U.S. EPA 1973), Pennack (1978), Merritt et al. (1984), and APHA et al.
(1985) offer excellent guidance for sample sorting. Hynes (1970, 1971)
states that the earlier stages of benthic organisms are retained by a 0.2-mm
mesh size (approximately the size of a #75 standard sieve), and APHA et al.
(1985) and Weber (U.S. EPA 1973) define benthic organisms by a mesh size of
0.595 (standard sieve #30), which is now standard practice. However, some
types of chironomids and other small benthos pass through the #30-mesh sieve
but are 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
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
sieves should then be sorted (Ohio EPA 1987c). 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.
Midge Preparation—Slide preparation for the genus and species identifi-
cation of the chironomids can involve either wet mounts or permanent
mounts. In either case, the head capsule is the primary part of the
organism used in their taxonomy, and must be cleared of pigmentation for
best identification. Simpson and Bode (1980) and Ohio EPA (1987c) recommend
clearing the organisms in either a cold solution of 10 percent potassium
hydroxide (KOH) overnight, or a heated (not boiling) solution of the same
for no more than a few hours. Once the samples are rinsed in water (less
than 0.5 h), the organisms can be examined on a wet or permanent mount. The
wet mount uses only a few drops of water and a cover slip. The midges need
to be dehydrated in successive solutions of 70 percent and 95 percent
ethanol if a permanent mount is to be made. The prepared midges can then be
permanently mounted using a number of media, including Euparol before the
/
cover slip is applied. Great care should be taken to ensure the ventral
side of the head capsule is fully visible before making the mount permanent.
Generally, only the voucher specimens need to be permanent mounts, while the
other midges can be identified by using wet mounts. Weber (U.S. EPA 1973),
Pennack (1978), Simpson and Bode (1980), Merrit and Cummins (1984), and Ohio
EPA (1987c) provide detailed guidance on this subject.
Taxonomy--The level to which the taxonomic identifications should be
taken is dependent upon the objectives of the study. For a reconnaissance
or screening survey, it is generally not necessary to go beyond family
(Illinois EPA 1987; Hilsenhoff 1988; Plafkin et al. 1988; Resh 1988). For
studies attempting to identify designated use impairment or evaluate
impacts from a specific source, the minimum level of taxonomic detail should
follow recommendations of Ohio EPA (1987c). Ohio EPA has successfully
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
implemented numeric biocriteria based on this taxonomic detailing. The
focus is on differentiating taxa that are better water quality indicators
and for which taxonomic keys and expertise are readily available. The level
of taxonomic detailing must be consistent within the program and applied for
each sample site. Species-level identifications for all organisms are not
necessary for a successful program and they commonly depend upon the
availability of local keys. National keys available for genus-level
identifications include Merritt and Cummins (1984) for insects, Pennack
(1978) for all common invertebrates, and Klemm (1985) for annelids (oligo-
chaetes and leeches). Regional U.S. EPA or state biologists should be con-
tacted to determine which of the hundreds of other taxonomic keys are
available for specific taxa, both nationally and regionally.
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-level taxonomy
and enumerations may be required. The necessary data 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 representation
and analysis of the data collected. Structural community measures such as
richness values, diversity and biotic indices, and enumerations have been
used almost exclusively. Indicator organisms have been used to establish
many of the biotic indices 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 macroinvertebrate
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Freshwater Macroinvertebrate Benthic Community
. Structure and Function
community should use a complementary mix of both structural and functional
measures. Discussions of various data analysis techniques can be found in
Hawkes (1979), Cairns (1981), Washington (1984), and Resh (1988).
Diversity Indices—When diversity indices 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 indices 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 Shannon, or Shannon-Wiener index
(Shannon and Weaver 1949) as shown below:
s
Shannon's H'« Z ("i/n) In (n^/n)
where:
n^ - Total number of individuals in the i^n taxon
n - Total number of individuals
s - Total number of taxa.
[Washington (1984) provides a good explanation of how the index derived the
name Shannon-Wiener rather than Shannon-Weaver index]. Theoretically,
higher community diversity indicates better water quality (Wilhm 1970).
However, 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 indices such as the Shannon-Wiener
index still remain in widespread use (Washington 1984), their limitations in
accurately addressing ' a variety of perturbations 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)
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
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 indices showed varied
results in detecting changes in water quality and that they are not the
optimal measures of water quality. However, diversity indices can provide
additional information as to the community composition and should be reported
if the data are available. Reliance upon these indices as the only, or
predominant measure upon which water pollution control decisions are based
is not valid. Washington (1984) provides an outstanding review of the
history and uses of diversity indices.
Biotic Indices--Biotic indices utilize 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
indices. An example of a widely used biotic index (Hilsenhoff 1977, 1982)
is as follows:
s
Biotic Index - Z
where:
n^ » Number of individuals in taxon i
a, • Tolerance value assigned to taxon i
n « 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 indices 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 indices based on a specific
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
population, rather than community, are addressed in the "Indicator Organism"
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
(Hilsenhoff 1987) and to include a family-level biotic index (Hilsenhoff
1988). Because the biotic indices rely heavily on known pollution tolerances
of the taxa, Washington (1984), Mason et al. (1985), and Hawkes (1979)
preferred the biotic indices over the diversity indices for water quality
assessments. The success of the Hilsenhoff Biotic Index prompted its use,
or modifications of it, in several state programs (e.g. Wisconsin, Illinois,
New York, North Carolina). Unfortunately, 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.
Washington (1984) provides an outstanding review of the history and uses of
these indices.
Indicator Organisms — Indicator ^organisms have played a key role in the
development of biotic indices for both lotic and lentic systems. One of the
first classifications based on indicator organisms was done in the Illinois
River by Richardson (1928). Simpson and Bode (1980), Bode and Simpson
(unpublished), and Rae (1989), among many others, utilized 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 indices, and have relied extensively on enumerations, diversity
indices, 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 oligochaetes to indicate fine sediment pollution. Brinkhurst et
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
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
Aagaard (1986) in European lakes. Hart and Fuller (1974) presented
pollution ecology data for a number of freshwater benthic macroinvertebrates
as -did U.S. EPA's pollution tolerance information series on Chironomidae
(Beck 1977), Trichoptera (caddisflies) (Harris and Lawrence 1978), Ephemer-
optera (mayflies) (Hubbard and Peters 1978), and Plecoptera (stoneflies)
(Surdick and Gaufin 1978). Washington (1984) also reviewed population-based
biotic indices.
Richness Measures — Richness measures are based on the presence or
absence of selected taxa. Commonly used measures include the total number
of taxa, 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 measures 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 19S7a; Resh 1988), and ratios of the number of individuals within
one taxonomic group (e.g., EPT) individuals to number of individuals within
another taxonomic group (e.g., Chironomidae) (Plafkin et al. 1988; Resh
1988). Interpretation of the enumeration ratios can be difficult without
prior validation. Host 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
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Freshwater Macroinvertebrate Benthic Community
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(1988) concluded that the seven common enumerations he tested had extremely
high variability and unacceptably low accuracy in detecting various impacts,
and 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 optima.1 for widespread use, they may still provide
insight into changes in the community structure. Ohio EPA (1987a) has
successfully used enumerations for the percentage of mayflies, caddisflies,
Tanytarsini midges, tolerant organisms, and "other" dipterans combined with
non-insect individuals as a basis for their state biocriteria.
Similarity Indices—Community similarity indices measure the similarity
between benthic communities at a reference and a study site, with high
similarity indicating little change, or impact, between the two sites. The
use of similarity indices has been reviewed by Brock (1977) and Washington
(1984). The simplest indices to apply are those that 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
indices. Van Horn's Index used by Ohio EPA (1987c) is as follows:
Similarity (c) - 2w/(a+b)
where:
a - Number of taxa collected at one site
b • Number of taxa collected at the other site
w - Number of taxa common to both stations.
A value over 6.5 or 7.0 indicates good similarity. Plafkin et al. (1988)
utilize the Jaccard Index in the rapid bioassessment protocols (RBPs). Other
indices such as the percent similarity (Brock 1977) and the Bray-Curtis
(1957) utilize the abundance of organisms.
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Freshwater Macroinvertebrate Benthic Community
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Functional Information—Community function measurements based upon
habitat, trophic structure, and other ecological measures were described by
Kaesler et al. (1978) and used by Rooke and Mackie (1982a) as the "ecological
community analysis" (ECA). Rooke and Mackie (1982b) reported the ECA to
provide more information on environmental quality than diversity or biotic
indices, but the ECA 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 particulate
organic matter and determining the functional feeding groups described in
Merritt and Cummins (1984). This method is being used in the RBPs (Plafkin
et al. 1988). 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 development of a biotic index
for paper mill impacts. Another useful measure of function is observations
of the incidence of morphological deformities in benthic macroinvertebrates,
similar to the observations made for the 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 Chironomidae -(Cushman-1-984;-
Cushman and Goyert 1984; Wiederholm 1984b; Warwick 1985; Warwick et al.
1987) and Trichoptera (Simpson 1980; Petersen and Petersen 1983).
Composite Indices—Composite indices combine selected structural or
functional measures, or "metrics," in a cumulative scoring system, as was
done with the IBI for the fish community (Karr et al. 1986). Ohio EPA
(1987c) developed a similar index for invertebrates using the following 10
structural metrics, adjusted for drainage area size, to derive a final
Invertebrate Community Index (ICI) score:
1. Total number of taxa
2. Total number of mayfly taxa
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Freshwater Macroinvertebrate Benthic Community
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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
10. Total number of qualitative EPT taxa
The ICI score is directly related to Ohio EPA's numeric biocriteria for
designated use attainment, and was developed using artificial and natural
substrate 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 scores and little
redundancy between the metrics (Davis and Lubin 1989).
U.S. EPA (Plafkin et al. 1988) 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)
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Freshwater Macroinvertebrate Benthic Community
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4. Ratio of EPT and Chironomidae abundances
5. Percent contribution of dominant taxon
' 6. EPT index
7. Community similarity index
8. Ratio of shredders to total number of organisms (functional).
These RBPs are developed by conducting single-habitat (riffle) dip-net
sampling. The scores are based on a percentage of the metric values found
at a reference site, rather than comparison of the results based on "optimal"
values for each metric. U.S. EPA Region V is currently developing such
"optimal" metric values. The RBPs are flexible and can be modified for
different geographical locations, as evidenced by the use of different
metrics in Arkansas (Shakelford 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 taxonomy and may be done in the field. The RBPs normally use
single habitat (riffle) sampling and are limited to a 100 organism count in
the field. However, they can be adapted for most program uses, for example
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.
Statistical Approaches—Various statistical approaches have been
applied to determine whether the benthic community at a study site varies
from that at a reference or other site. Depending upon the chosen endpoints
of the study, rigorous statistical analysis may not be necessary. For
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
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 "statistical" difference between two communities will
not 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. The same data evaluation procedures apply to both the marine and
freshwater systems. The reader is referred to the statistical discussion
in Chapter 8 (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 statis-
tical distribution such as a normal or log-normal distribution. This
assumption is often invalid when dealing with biological populations and
f
communities.
Alternatively, nonparametric analyses may be conducted. Such analyses
are not based on assumptions about a specific distribution of the data.
Examples of such tests include the chi-square test, binomial tests, rank
correlations, or tests comparable 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.
2.1.2.4 Necessary Hardware and Skills — The hardware needed for field
collection includes samplers (e.g., dredges, dip-nets), sieves, benthic
macroinvertebrate containers, forceps, white enamel pans, ethanol pre-
servative, and appropriate personal gear (e.g., hip boots or chest-waders,
life-vest if needed, and first aid kits). For the laboratory, standard
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
biological laboratory equipment should be available, such as microscopes
(both dissecting and compound), forceps, microscope slides and cover slips,
ethanol, potassium 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 biologists and taxonomists are
needed for benthic 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 assured. A minimum 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 accurately
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.
2.1.3 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.
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Freshwater Macroinvertebrate Benthic Community
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2.2 Applicability of Method to Human Health. Aquatic •Lifgf or Wildlife
Protection
This method is directly applicable to the protection of aquatic life
since it is based on direct measurements 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 macro-
invertebrates (e.g., small mammals and wading shorebirds). It is indirectly
applicable to other wildlife that depend upon 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.
2.3 Ability of Method to Generate Numerical Criteria for Specific Chemicals
This method is used in conjunction with sediment toxicity and chemistry
data to characterize toxicant impacts and assist with determining the appro-
priate levels at which the toxicants should be controlled. However, by
/
itself, this method would not be used to generate chemical-specific criteria.
3.0 USEFULNESS
3.1 Environmental ADD!icabilitv
Benthic macroinvertebrates have been routinely used to assess environ-
mental quality in a variety of geographical areas and ecoregions as was
discussed in Section 1.0.
3.1.1 Suitability for Different Sediment Types--
Assessment of the freshwater benthic macroinvertebrate community
structure is well suited for evaluating different sediment types, since the
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
benthos inhabit all substrates (Men-it and Cummins 1984). Comparisons
should be made among benthic communities of similar substrate since
different types and numbers of organisms will inhabit different types of
substrates.
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 chemicals.
In addition to the uses described in Section 1.1.1 of this chapter, many
benthic organisms 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).
3.1.3 Suitability for Predicting Effects on Different Organisms--
The use of benthic macroinvertebrates as indicator organisms has
already been discussed. Benthic macroinvertebrates can be used to predict
the effects upon other aquatic organisms because if the benthic macroinverte-
brate community is impacted, then the impact is likely to be, or already
has been, detrimental to other organisms.
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 toxicity. Alone, the benthic
macroinvertebrates can be used to screen for potential sources of sediment
contamination based on spatial gradients in community structure, but they
should not be used alone to definitively determine sediment quality or
develop chemical-specific guidelines. The benthic data must be integrated
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
with other available data to determine sediment quality using a "weight-of-
evidence" approach.
3.1.5 Suitability for Source Control--
Benthic macroinvertebrates have been extensively used for source
characterization and control in many of the state and U.S. EPA monitoring
programs involving spatial surveys upstream and downstream of suspected
sources (Ohio EPA 1987a; Bode and Novak 1988; Courtemanch and Oavies 1988;
Fiske 1988; Maret 1988; Penrose and Overton 1988; Shakelford 1988; U.S. EPA
1988a,b; 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 NPOES permit program.
Many states aggressively pursue this action.
3.1.6 Suitability for Disposal Applications--
The discussion presented in Section 3.lV6 of Chapter- 3 (marine benthic
macroinvertebrate community structure) is applicable to fresh water.
Recently benthic community assessments have been required by U.S. EPA
(1989c) Region V, as stated in the Draft Interim Guidance for the Design and
Execution of Sediment Sampling Efforts Relating to Navigational Maintenance
Dredging in Region V - Hay 1989. In this guidance, benthic macroinvertebrate
assessments are advised for areas that are suitable for open-lake disposal
or for sediments that are difficult to characterize. All benthic community
assessments will be made in concert with sediment chemistry and toxicity
evaluations.
3.2 General Advantages and Limitations
The advantage of using the benthic macroinvertebrates community
assessment approach to determining, sediment quality is that it provides
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
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 numeric in-place pollutant management. This method should be
integrated with sediment chemistry and toxicity information.
3.2.1 Ease of Use--
The equipment requirements for benthic surveys is minimal and inex-
pensive 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.
3.2.2 Relative Cost--
The cost for benthic macroinvertebrate assessments is economical
compared to that for chemistry or toxicological evaluations. Ohio EPA
(1987a) provided a cost of about 5700 to conduct a benthic assessment at one
sample site. However, this cost included overhead (e.g., rent, office
equipment), all travel expenses, time spent in the 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 5700 was quite economical compared to
chemical/physical testing (SI,500) or bioassay testing (53,000 to 512,000)
for each site.
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
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
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 h 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
analyze chemical variables.
3.2.3 Tendency to be Conservative--
The benthic macroinvertebrate community assessment provides a conser-
vative 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 community when an impact occurs. Because of
influences 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 information, 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 from 431 sites in Ohio, benthic macroinvertebrates
were more sensitive (conservative) indicators of water quality (Ohio EPA,
personal communication). In 35.6 percent of the sites, the benthic
assessment revealed impacts not detected by chemical analyses. In 58.1
percent of the sites, the chemical and biological assessment supported one
another. Only 6.3 percent of the sites did not have a benthic impact when
the chemistry indicated that there would be one.
3.2.4 Level of Acceptance--
Benthic macroinvertebrate community assessments of sediment/water
quality have been used in freshwater systems since the early 1900s
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
(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
establishment of numerical biological criteria in state water quality
standards.
3.2.5 Ability to be Implemented by Laboratories with Typical Equipment and
Hand!ing Facilities--
The only special pieces of equipment required are the samplers and
sieves, which are easily obtained from biological supply warehouses. Most
biological laboratories will have dissecting and compound microscopes,
chemical reagents, microscope slides and cover slips, forceps, and any other
materials needed. The laboratory's capability to identify benthic macro-
invertebrates is less common. Taxonomy is not a .widespread skill and is
more likely to be found in consulting firms than in analytical laboratories.
3.2.6 Level of Effort Required to Generate Results--
Depending upon the study objectives and level of effort needed, results
can be generated in written form in as quickly as 1 day (Plafkin et al.
1988) or in several months. For example, Ohio EPA processes over 500
individual benthic samples each year, identifies the organisms, and prepares
reports for regulatory use in less than 1 year, with fewer than three full-
time employees in their benthic macroinvertebrate unit. The critical period
is the turnaround time for the taxonomy. With artificial substrates, an
additional 6-wk colonization period is required; unless a rapid assessment
or moderate sized study is done, a written report including interpretation
of results will require between 6 mo and 1 yr.
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Freshwater Hacroinvertebrate Benthic Community
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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. Although several numerical indices that appear simple
to use are available, data interpretation relies upon all of the information
generated for a study, including chemical, physical, and toxicological
measurements, as well as indicator organisms and function measures.
3.2.8 Degree of Environmental Applicability--
Benthic macroinvertebrate community structure and function is used
extensively to evaluate sediment and water quality and characterize impacts
in lotic and lentic freshwater ecosystems.
3.2.9 Degree of Accuracy and Precision--
Since benthic macroinvertebrate's are measured directly, this method is
highly accurate for characterizing sediment/water quality effects upon
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 indicate 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 II
error is much more serious than a Type I error, which is conservative. To
reduce the possibility of a Type II error, the interpretation of the data
(including chemistry and toxicity) must be done by a trained benthic
ecologist. Resh (1988) reviewed the levels of accuracy and precision for
several of the data analysis techniques.
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Freshwater Macroirwertebrate Benthic Community
•Structure and Function
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 protocols 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 reduced when using a community-level approach rather
than a population-level approach.
4.0 STATUS
Sections 1.1 (Current Uses) and 3.0 (Usefulness) describe the status
of the discipline.
4.1 Extent of Use
This method is widely used in both regulatory and nonregulatory
sediment and 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.
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 upon other environmental data to
validate the method. The documentation provided in this paper should present
adequate documentation of the method's' validity.
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Freshwater Macroirwertebrate Benthic Community
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4.3 Reasons for Limited Use
Benthic macroinvertebrate community assessments are very common in
freshwater systems because of their relatively low cost and high information
output.
4.4 Outlook for Future Use and Amount of Development Yet Needed
The outlook for the future use of benthic macroinvertebrate community
structure and function in sediment quality assessment is very good because
of the recognition that benthic macroinvertebrates 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 whether sediment quality is
impairing the designated uses and biotic integrity. With the increasing
reliance upon numerical biocriteria, additional sediment quality problems
will be identified. The area where development is most needed iT irr
combining benthic community assessments with chemical and toxicological data
in an integrated approach for assessing sediment quality. In addition, the
functional measures, which also hold much promise for sediment assessments,
need to be validated more thoroughly.
5.0 REFERENCES
Aagaard, K. 1986. The Chironomidae fauna of north Norwegian lakes with a
discussion of community classification. Hoi. Ecol. 9:1-12.
American Public Health Association, American Water Works Association, and the
Water Pollution Control Federation. 1985. Standard methods for the
examination of water and wastewater. 16th Edition. APHA, Washington, DC.
American Society for Testing and Materials. 1988. Annual book of ASTM
standards: water and .environmental technology. Vol. 11.04. ASTM, Phila-
delphia, PA. 963 pp.
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
Armitage, P.O., and Blackburn, J.H. 1985. Chironomidae in a pennine stream
system receiving mine drainage and organic enrichment. Hydrobiologia
121:165-172.
Beck, W.M., Jr. 1977. Environmental requirements and pollution tolerance of
common freshwater Chironomidae. EPA-600/4-77/024. U.S. Environmental
Protection Agency, Office of Research and Development, Cincinnati, OH.
Beck, W.M., Jr. 1955. Suggested method for reporting biotic data. Sew.
Ind. Wastes 27:1193-1197.
Beckett, D.C., and M.C. Miller. 1982. Macroinvertebrate colonization of
multiplate samplers in the Ohio River: the effect of dams. Can. J. Fish.
Aquat. Sci. 39:1622-1627.
Bode, R.W., and M.A. Novak. 1988. Proposed biological criteria for New York
State streams, pp. 42-48. In: Proceedings of the First National Workshop
on Biological Criteria - Lincolnwood, Illinois, December 2-4, 1987. T.P.
Simon, L.L. Hoist, and L.J. Shepard (eds). EPA-905/9-39/003. U.S. EPA
Region 5 Instream Biocriteria and Ecological Assessment Committee, Chicago,
IL. 129 pp.
Bode, R.W., and K.W. Simpson. Unpublished. Communities of Chironomidae in
large lotic systems: impacted vs. unimpacted. Unpublished paper presented
at the 30th Annual Meeting of the North American Benthological Society in
Ann Arbor, MI, May 18, 1982. 15 pp.
Bray, J.R., and J.T. Curtis. 1957. An ordination of the upland forest
communities of southern Wisconsin. Ecol. Monogr. 27:325-349.
Brillouin, L. 1962. Science and information theory. Academic Press, New
York, NY. pp. 1-347.
Brinkhurst, R.O., A.L. Hamilton, and H.B. Herrington. 1968. Components of
the bottom fauna of the St. Lawrence Great Lakes. Great Lakes Inst. Report
33. University of Toronto, Toronto, Canada. 50 pp.
Brock, D.A. 1977. Comparison of community similarity indices. J. Wat.
Pollut. Control Fed. 49:2488-2494.
Cairns, J., Jr. 1981. Introduction to biological monitoring, pp. 375-409.
In: Water Quality Management: The Modern Analytical Techniques. H.B. Mark,
Jr., and J.S. Mattson (eds). -Marcel Dekker, Inc. New York, NY.
Chadwick, J.W., and S.P. Canton. 1983. Comparison of multiplate and surber
samplers in a Colorado mountain stream. J. Freshwater Ecol. 2:287-292.
7-37
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Structure and Function
Chadwick, J.W., and S.P. Canton. 1984. Inadequacy of diversity indices in
discerning metal mine drainage effects on a stream invertebrate community.
Wat. Air Soil Pollut. 22:217-223.
Chandler, J.R. 1970. A biological approach to water quality management. J.
Wat. Pollut. Control Fed. 4:415-422.
Cook, O.G., and M.G. Johnson. 1974. Benthic macroinvertebrates of the St.
Lawrence Great Lakes. J. Fish. Res. Board Can. 31:763-782.
Cooke, S.E.K. 1976. Quest for an index of community structure sensitive to
water pollution. Environ. Pollut. 11:269-288.
Courtemanch, O.L. 1987. Trophic classification of Maine lakes using benthic
Chironomidae fauna. Paper presented at the 7th International Symposium of
North American Lake Management Society, Orlando, FL. 20 pp.
Courtemanch, D.L., and S.P. Oavies. 1988. Implementation of biological
standards and criteria in Maine's water classification law. pp. 4-9. In:
Proceedings of the First National Workshop on Biological Criteria-
Lincolnwood, Illinois, December 2-4, 1987. T.P. Simon, L.L. Hoist, and L.J.
Shepard (eds). EPA-905/9-89/003. U.S. EPA Region 5 Instream Biocriteria and
Ecological Assessment Committee, Chicago, IL. 129 pp.
Grossman, J.S., and J. Cairns. 1974. A comparative study between two
different artificial substrate samplers and regular sampling techniques..
Hydrobiologia 44:517-522.
Grossman, J.S., J.R. Wright, and R.L. Kaesler. 1984. Consolidation of
baseline information, development of methodology, and investigation of
thermal impacts on freshwater shellfish, insects, and other biota. EPA-
600/7-84/042. Prepared by Tennessee Valley Authority for U.S. EPA Office of
Research and Development, Washington, DC. 159 pp.
Cummins, K.W. 1988. Rapid bioassessment using functional analysis of
running water invertebrates. pp. 49-54. In: Proceedings of the First
National Workshop on Biological Criteria - Lincolnwood, Illinois, December 2-
4, 1987. T.P. Simon, L.L. Hoist, and L.J. Shepard (eds). EPA-905/9-89/003.
U.S. EPA Region 5 Instream Biocriteria and Ecological Assessment Committee,
Chicago, IL. 129 pp.
Cummins, K.W., and M.A. Wilzbach. 1985. Field procedures for analysis of
functional feeding groups of stream macroinvertebrates. Contr. 1611 to
Appalachian Environmental Research Laboratory. University of Maryland,
Frostburg, MD. 21 pp.
Cushman, R.M. 1984. Chironomid deformities as indicators of pollution from
a synthetic, coal-derived oil. Freshwater Biology 14:179-182.
7-38
-------
u .
Freshwater Macroinvertebrate Benthic Community
Structure and Function
Cushman, R.M., and J.C. Goyert. 1984. Effects of a synthetic crude oil on
pond benthic insects. Environ. Pollut. (Ser. A) 33:163-186.
Davis, W.S., and A.L. Lubin. 1989. A statistical validation of Ohio EP'A's
invertebrate community index. Draft. Paper presented at the First Midwest
Pollution Control Biologists Meeting, U.S. EPA Region V, February 14-17,
1989, Chicago, IL. 15 pp.
Davis, W.S., and T.P. Simon. . 1988. Sampling and data evaluation require-
ments for fish and macroinvertebrate communities. pp. 89-97. In:
Proceedings of the First National Workshop on Biological Criteria-
Lincolnwood, Illinois, December 2-4, 1987. T.P. Simon, L.L. Hoist, and L.J.
Shepard (eds). EPA-90S/9-89/003. U.S. EPA Region 5 Instream Biocriteria
and Ecological Assessment Committee, Chicago, IL. 129 pp.
Oenbow, T.A., and W.S. Davis. 1986. Highway runoff water quality training
course student workbook. Chapter 7. In: Water Quality Impacts. U.S. Oept.
Transportation, Federal Highway Administration, McLean, VA.
DePauw, N., 0. Roels, and A.P. Fontoura. 1986. Use of artificial substrates
for standardized sampling of macroinvertebrates in the assessment of water
quality by the Belgian biotic index. Hydrobiologia 133:237-258.
Oupuis, T.V., P. Bertram, J. Meyer, M. Smith, N. Kobriger, and J. Kaster.
1985. Effects of highway tunoff on receiving waters. Volume II: results of
field monitoring program. Prepared by Rexnord for the Federal Highway
Administration, McLean, VA.
Fandrei, G. 1989. Personal Communication. Minnesota Pollution Control
Agency, St. Paul, MN.
Fiske, S. 1988. The use of biosurvey data in the regulation of permitted
nonpoint discharges in Vermont. pp. 67-74. In: Proceedings of the First
National Workshop on Biological Criteria - Lincolnwood, Illinois, December 2-
4, 1987. T.P. Simon, L.L. Hoist, and L.J. Shepard (eds). EPA-905/9-89/003.
U.S. EPA Region 5 Instream Biocriteria and Ecological Assessment Committee,
Chicago, IL. 129 pp.
Fitchko, J. 1986. Literature review of the effects of persistent toxic
substances on Great Lakes biota. Report of the Health of Aquatic Communities
Task Force, International Joint Commission, Windsor, Ontario. 256 pp.
Gaufin, A.R., and C.M. Tarzwell. 1952. Aquatic invertebrates as indicators
of stream pollution. Pub. Health. Report 67:57.
7-39
-------
Freshwater Macroinvertebrate Benthic Community
Structure and Function
Green, R.H. 1978. Optimal impact study design and analysis, pp. 3-28.
In: Biological Data in Water Pollution Assessment: Quantitative and
Statistical Analyses. K.L. Oickson, J. Cairns, Jr., and R.L. Livingston
(eds). ASTM STP 652. American Society for Testing and Materials, Phila-
delphia. PA.
Harris, T.L., and T.M. Lawrence. 1978. Environmental requirements and
pollution tolerance of Trichoptera. EPA-600/4-78/063. U.S. Environmental
Protection Agency, Office of Research and Development, Cincinnati, OH.
Hart, C.W., Jr., and S.L.H. Fuller (eds). 1974. Pollution ecology of
freshwater invertebrates. Academic Press, Inc. London. 389 pp.
Hawkes, H.A. 1979. Invertebrates as indicators of river water quality.
Chapter 2, pp. 1-45. In: Biological Indices of Water Quality. A. James,
and L. Evison (eds). John Wiley and Sons, New York, NY.
Hester, F.E., and J.B. Oendy. 1962. A multi-plate sampler for aquatic
macroinvertebrates. Trans. Amer. Fish. Soc. 91:420.
Hilsenhoff, W.L. 1977. Use of arthropods to evaluate water quality of
streams. Technical Bulletin No. 100. Wisconsin Department of Natural
Resources, Madison, WI. 15 pp.
Hilsenhoff, W.L. 1982. Using a biotic index to evaluate water quality in
streams. Technical Bulletin No. 132. Wisconsin Department of Natural
Resources, Madison, WI. 23 pp.
Hilsenhoff, W.L. 1987. An improved biotic index of organic stream
pollution. Great Lakes Entomologist 20:31-39.
Hilsenhoff, W.L. 1988. Rapid field assessment of organic pollution with a
family-level biotic index. J. N. Am. Benthol. Soc. 7:65-68.
Hite, R.L. 1988. Overview of stream quality assessments and stream
classification in Illinois. pp. 98-125. In: Proceedings of the First
National Workshop on Biological Criteria - Lincolnwood, Illinois, December 2-
4, 1987. T.P. Simon, L.L. Hoist, and L.J. Shepard (eds). EPA-905/9-89/003.
U.S. EPA Region 5 Instream Biocriteria and Ecological Assessment Committee,
Chicago, IL. 129 pp.
Howmiller, R.P., and M.A. Scott. . 1977. An environmental index based on
relative abundance of oligochaete species. J. Wat. Pollut. Control Fed.
49:809-815.
Hubbard, M.D., and W.L. Peters. 1978. Environmental requirements and
pollution tolerance of Ephemeroptera. EPA-600/4-78/061. U.S. Environmental
Protection Agency, Office of Research and Development, Cincinnati, OH.
7-40
-------
Freshwater Macroinvertebrate Benthic Community
Structure and Function
Hughs, B.D. 1978. The influence of factors other than pollution on the
values of Shannon's diversity index for benthic macroinvertebrates in
streams. Wat. Res. 12:359-364.
Hynes, H.B.N. 1970. The ecology of running waters. University of Toronto
Press. 555 pp.
Hynes, H.B.N. 1971. Benthos of flowing water, pp. 66-80. In: Secondary
Productivity in Freshwaters. W.T. Edmondson, and G.C. Winberg (eds). IBP
Handbook No. 17. Blackwell Scientific Pub!., Oxford, U.K.
Illinois Environmental Protection Agency. 1987. Quality assurance and
field methods nanual. Section C. Macroinvertebrate monitoring. Illinois
EPA, Division of Water Pollution Control, Springfield, IL.
Jaccard, P. 1908. Nouvelles recherches sur la distribution florale. Bull.
Soc. Vaud. Sci. Nat. XLIV(163):223-269.
Kaesler, R.L., E.E. Herricks, and J.J. Grossman. 1978. Use of indices of
diversity and hierarchial diversity in stream surveys. pp. 92-112. In:
Biological Data in Water Pollution Assessment: Quantitative and Statistical
Analyses. K.I. Dickson, J. Cairns, Jr., and R.L. Livingston (eds). ASTM
STP 652. American Society for Testing and Materials, Philadelphia, PA.
Karr, J.R., K.D., 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.
Springfield, IL. 28 pp.
Klemm, D.J. 1985. A guide to the freshwater Annelida (Polychaeta, Naidid
and Tubificid Oligochaeta, and Hirudinea) of North America. Kendall/Hunt
Publ., Dubuque, IA. 198 pp.
Krieger, K.A. 1984. Benthic macroinvertebrates as indicators of environ-
mental degradation in the southern nearshore zone of the central basin of
Lake Erie. J. Great Lakes Res. 10:197-209.
Lafont, M. 1984. Oligochaete communities as biological descriptors of
pollution in the fine sediments of rivers. Hydrobiologia 115:127-129.
Larsson, P. 1984. Transport of PCBs from aquatic to terrestrial environ-
ments by emerging chironomids. Environ. Pollut. (Ser. A) 34:283-289.
Lauritsen, D.D., S.C. Mozley, and D.S. White. 1985. Distribution of
oligochaetes in Lake Michigan and comments on their use as indices of
pollution. J. Great Lakes Res. 11:67-76.
7-41
-------
Freshwater Macroinvertebrate Benthic Community
Structure and Function
Lenat, D.R. 1988. Water quality assessment of streams using a qualitative
collection method for benthic macroinvertebrates. J. N. Am. Benthol. Soc.
7:222-233.
Maret, T. 1988. A stream inventory process to classify use support and
develop biological standards in Nebraska, pp. 55-66. In: Proceedings of
the First National Workshop on Biological Criteria - lincolnwood, Illinois,
December 2-4, 1987. T.P. Simon, L.L. Hoist, and L.J. Shepard (eds). EPA-
905/9-89/003. U.S. EPA Region 5 Instream Biocriteria and Ecological
Assessment Committee, Chicago, IL. 129 pp.
Mason, W.T., P.A., Lewis, and C.I. Weber. 1985. An evaluation of benthic
macroinvertebrate biomass methodology. Part 2. Field assessment and data
evaluation. Environ. Monitor. Assess. 5:399-422.
Merritt, R.W., and K.W. Cummins (eds). 1984. An introduction to the
aquatic insects of North America. 2nd edition. Kendall/Hunt Publ.,
Oubuque, IA. 441 pp.
Merritt, R.W., K.W. Cummins, and V.H. Resh. 1984. Collecting, sampling,
and rearing methods for aquatic insects, pp. 11-26. In: R.W. Merritt, and
K.W. Cummins (eds). An Introduction to the Aquatic Insects of North
America. 2nd edition. Kendall/Hunt Publ., Dubuque, IA.
Millard, S.P., and O.P. Lettenmaier. 1986. Optimal design of biological
sampling programs using the analysis of variance. Est. Coast. Shelf Sci.
22:637-656.
Moore, J.W., V.A. Beaubien, and D.J. Sutherland. 1979. Comparative effects
of sediment and water contamination on benthic invertebrates in four lakes.
Bull. Environ. Contam. Toxicol. 23:840-847.
Mozley, S.C. 1978. Effects of experimental oil spills on Chironomidae in
Alaska tundra ponds. Verh. Internat. Verein. Limnol. 20:1941-1945.
Mozley, S.C., and M.G. Butler. 1978. Effects of crude oil on aquatic
insects of tundra ponds. Arctic 31:229-241.
Murphy, P.M. 1978. The temporal variability in biotic indices. Environ.
Poll. 17:227-236.
Ohio Environmental Protection Agency. 1987a. Biological criteria for the
protection of aquatic life: Volume I. The role of biological data in water
quality assessment. Ohio EPA, Division of Water Quality Monitoring and
Assessment, Surface Water Section, Columbus, OH. 44 pp.
7-42
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
Ohio Environmental Protection Agency. 1987b. Biological criteria for the
protection of aquatic life: Volume II. Users manual for biological field
assessment of Ohio surface waters. Ohio EPA, Division of Water Quality
Monitoring and Assessment, Surface Water Section, Columbus, OH.
Ohio Environmental Protection Agency. 1987c. Biological criteria for the
protection of aquatic life: Volume III. Standardized biological field
sampling and laboratory methods for assessing fish and macroinvertebrate
communities. Ohio EPA, Division of Water Quality Monitoring and Assessment,
Surface Water Section, Columbus, OH.
Peckarsky, B.L. 1986. Colonization of natural substrates by stream
benthos. Can. J. Fish. Aquat. Sci. 43:700-709.
Pennack, R.W. 1978. Freshwater invertebrates of the United States. 2nd
ed. John Wiley 4 Sons, Inc., New York. 803 pp.
Penrose, O.L., and O.K. Lenat. 1982. Effects of apple orchard runoff on the
aquatic macrofauna of a mountain stream. Arch. Environ. Contam. Toxicol.
11:383-388.
Penrose, D.L., and J.R. Overton. 1988. Semiqualitative collection
techniques for benthic macroinvertebrates: uses for water pollution
assessment in North Carolina. pp. 77-88. In: Proceedings of the First
National Workshop on Biological Criteria - Lincolnwood, Illinois, December 2-
4, 1987. T.P. Simon, L.L. Hoist, and L.J. Shepard (eds). EPA-905/9-89/003.
U.S. EPA Region 5 Instream Biocriteria and Ecological Assessment Committee,
Chicago, IL. 129 pp.
Peters?"; L.B.M,. and R = C, Petersen. 1983. Anomalies in hydropsychid
capture nets from polluted streams. Freshwater Biology 13:185-191.
Plafkin, J.L., M.T. Barbour, K.D. Porter, and S.K. Gross. 1988. Rapid
bioassessment protocols for use in streams and rivers: benthic macroin-
vertebrates and fish. Draft. Prepared by EA Engineering Science and
Technology Corp. for U.S. Environmental Protection Agency, Monitoring and
Data Support Division, Washington, DC.
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.I. Kinney. 1979. Assessment of macroinvertebrate
monitoring techniques in an energy development area: a test of the
efficiency of three macrobenthic sampling methods in the White River. EPA-
600/7-79/163. U.S. Environmental Protection Agency, Office of Research and
Development, Las Vegas, NV. 26 pp.
7-43
-------
Freshwater Macroinvertebrate Benthic Community
Structure and Function
Rabeni, C.F., S.P. Davies, and K.E. Gibbs. 1985. Benthic invertebrate
response to pollution abatement: structural changes and functional implica-
tions. Wat. Res. Bull. 21:489-497.
Rae, J.G. 1989. Chironomid midges as indicators of organic pollution in
the Scioto River basin, Ohio. Ohio J. Sci. 89:5-9.
Resh, V.H. 1988. Variability, accuracy, and taxonomic costs of rapid
assessment approaches in benthic biomonitoring. Draft. Paper presented at
the 1988 North American Benthological Society Technical Information Workshop,
Tuscaloosa, AL.
Resh, V.H., and D.G. Price. 1984. Sequential sampling: a cost effective
approach for monitoring benthic macroinvertebrates in environmental impact
assessments. Environ. Manage. 8:75-80.
Richardson,
1925, Bull
R.E. 1928. The bottom fauna of the middle Illinois River, 1913-
, Illinois Natural History Survey 17:387-472.
Rooke, J.B.,
environments:
Rooke, J.B.,
environments:
1:433-442.
and G.L. Mackie. 1982a.
I. Design and testing. J.
An ecological
Freshwat. Ecol.
analysis of lotic
1:421-432.
and G.L. Mackie.
II. Comparison
1982b. An ecological analysis of lotic
to existing indices. J. Freshwat. Ecol.
Rosas, I., M. Mazari, J. Saavedra, and A.P. Baez. 1985. Benthic organisms
as iIndicators of water quality in Lake Patzcuaro, Mexico. Water Air Soil
Pollut. 25:401-414.
Rosenberg, D.M., and V.H. Resh. 1982. The use of artificial substrates in
the study of freshwater benthic macroinvertebrates. pp. 175-236. In:
Artificial Substrates. J. Cairns, Jr. (ed). Ann Arbor Science Publishers,
Ann Arbor, MI.
Rosenberg, D.M., and A.P. Wiens. 1976. Community and species responses of
Chironomidae (Diptera) to contamination of fresh waters by crude oil and
petroleum products, with special reference to the Trail River, Northwest
Territories. J. Fish. Res. Board Can. 33:1955-1963.
Saether, O.A. 1979. Chironomidae communities as indicators of water
quality. Hoi. Ecol. 2:65-74.
Shakelford, B. 1988. Rapid bioassessments of lotic macroinvertebrate
communities: biocriteria development. Arkansas Department of Pollution
Control and Ecology, Little Rock, AR. 45 pp.
7-44
-------
Freshwater Macroinvertebrate Benthic Community
Structure and Function
Shannon, C.E., and W. Weaver. 1949. The mathematical theory of communica-
tion. The University of Illinois Press, Urbana, IL. pp. 19-27, 82-83, 104-
107.
Sheldon, A.L. 1984. Cost and precision in a stream sampling program.
Hydrobiologia 111:147-152.
Shepard, R.B. 1982. Benthic insect colonization of introduced substrates in
the Sangamon River, Illinois. Trans. 111. Acad. Sci. 75:15-27.
Simpson, K.W. 1980. Abnormalities in the trachea! gills of aquatic insects
collected from streams receiving chlorinated or crude oil wastes. Freshwater
Biology 10:581-583.
Simpson, K.W. 1983. Communities of Chironomidae (Diptera) from an acid-
stressed headwater stream in the Adirondack Mountains, New York. Mem. Amer.
Entomol. Soc. 34:315-327.
Simpson, K.W., and R.W. Bode. 1980. Common larvae of Chironomidae (Diptera)
from New York State streams and rivers - with particular reference to the
fauna of artificial substrates. New York State Department of Health, New
York State Museum Bull. No. 439. Albany, NY. 105 pp.
Smith, M.E., and J.L. Kaster. 1983. Effect of rural highway runoff on
stream benthic macroinvertebrates. Environ. Pollut. Ser. A. 32:157-170.
Surber, E.W. 1937. Rainbow trout and bottom fauna production in one mile of
stream. Trans. Amer. Fish. Soc. 66:193-202.
Surber, E.W. 1970. Procedure in taking stream bottom samples with the
stream square foot bottom sampler. Proc. Conf. Southeastern Assoc. Game
Fish. Comm. 23:587-591.
Surdick, R.F., and A.R. Gaufin. 1978. Environmental requirements and
pollution tolerance of Plecoptera. EPA-600/4-78/062. U.S. Environmental
Protection Agency, Office of Research and Development, Cincinnati, OH.
U.S. Environmental Protection Agency. 1973. Biological field and laboratory
methods for measuring the quality of surface waters and effluents. C.I.
Weber (ed). EPA-670/4-73/001. U.S. Environmental Protection Agency,
National Environmental Research Center, Cincinnati, OH.
U.S. Environmental Protection Agency. 1987. A compendium of Superfund
field operations methods. Section 12, Biology/Ecology. EPA-540/P-87/001.
U.S. EPA, Office of Emergency and Remedial Response, Washington, DC.
7-45
-------
Freshwater Macroinvertebrate Benthic Community
Structure and Function
U.S. Environmental Protection Agency. 1988a. Report of the national
workshop on instream biological monitoring and criteria. U.S. EPA Region V
Instream Biological Criteria Committee, U.S. EPA Office of Water, Washington,
DC. 34 pp.
U.S. Environmental Protection Agency. 1988b. Proceedings of the first
national workshop on biological criteria - Uncolnwood, Illinois, December 2-
4, 1987. EPA-905/9-89/003. U.S. EPA Region Instream Biocriteria and
Ecological Assessment Committee, Chicago, IL. 129 pp.
U.S. Environmental Protection Agency. 1989a. Ecological assessment of
hazardous waste sites. EPA-600/3-89/013. U.S. Environmental Protection
Agency, Office of Research and Development, Corvallis, OR.
U.S. Environmental Protection Agency. 1989b. Risk assessment guidance for
Superfund - environmental evaluation manual. Interim Final. EPA-540/1-
89/001A. U.S. Environmental Protection Agency, Office of Emergency and
Remedial Response, Washington, DC.
U.S. Environmental Protection Agency. 1989c. Interim guidance for the
design and execution of sediment sampling efforts related tonavigational
maintenance dredging in Region V - May 1989. U.S. EPA Region V, Chicago, IL.
van Dyk, L.P., C.G. Greeff, and J.J. Brink. 1975. Total population
density of Crustacea and aquatic Insecta as an indicator of fenthion
pollution of river water. Bull. Environ. Contam. Toxicol. 14:426-431.
Van Horn, W.M. 1950. The biological indices of stream quality. Proc. 5th
Ind. Waste. Conf., Purdue Univ. Est. Ser. 72:215.
Warwick, W.F. 1985. Morphological abnormalities in Chironomidae (Diptera)
larvae as measures of -toxic stress in freshwater ecosystems: indexing
antennal deformities in Chironomus Meigen. Can. J. Fish. Aquat. Sci.
42:1881-1914.
Warwick, W.F., J. Fitchko, P.M. McKee, D.R. Hart, and A.J. Burt. 1987.
The incidence of deformity in Chironomus spp. from Port Hope Harbour, Lake
Ontario. J. Great Lakes Res. 13:88-92.
Washington, H.G. 1984. Diversity, biotic and similarity indices: a review
with special relevance to aquatic ecosystems. Water Res. 18:653-694.
Waterhouse, J.C., and M.P. Farrell. 1985. Identifying pollution related
changes in chironomid communities as a cunction of taxonomic rank. Can. J.
Fish. Aquat. Sci. 42:406-413.
Webb, D.W. 1980. The effects of toxaphene piscicide on benthic macroinverte-
brates. J. Kansas Entomol. Soc. 53:731-744.
7-46
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Freshwater Macroinvertebrate Benthic Community
Structure and Function
Wentsel, R., A. Mclnotsh, and V. Anderson. 1977. Sediment contamination
and benthic macroinvertebrate distribution in a metal-impacted lake.
Environ. Pollut. 14:187-193.
Wiederholm, T. 1980. Use of benthos in lake monitoring. J. Wat. Pollut.
Control Fed. 52:537-547.
Wiederholm, T. 1984a. Responses of aquatic insects to environmental
pollution. pp. 508-557. In: The Ecology of Aquatic Insects. V.H. Resh
and D.M. Rosenberg (eds). Praeger Publishers, New York,NY. 625 pp.
Wiederholm, T. 1984b. Incidence of deformed chironomid larvae (Oip-
tera:Chironomidae) in Swedish Lakes. Hydrobiologia 109:243-249.
Wihlm, J.L. 1970. Range of diversity in benthic macroinvertebrate
populations. J. Wat. Pollut. Control Fed. 42:R221-224.
Winnell, M.H., and D.S. White. 1985. Trophic status of southeastern Lake
Michigan based on the Chironomidae (Diptera). J. Great Lakes. Res. 11:540-
548.
Winner, R.W., J.S. Van Dyke, N. Caris, and M.P. Farrell. 1975. Response of
a macroinvertebrate fauna to a copper gradient in an experimentally-polluted
sStream. Verh. Internat. Verein. Limnol. 19:2121-2127.
Winner, R.W., M.W. Boesel, and M.P. Farrell. 1980. Insect community
structure as an index of heavy-metal pollution in lotic ecosystems. Can. J.
Fish. Aquat. Sci. 37:647-655.
Woodiwiss, F.S. 1964. The biological system of stream classification used
by the Trent River Board. Chem. Ind. 11:443-447.
Yasuno, M., Y. Sugaya, and T. Iwakuma. 1985. Effects of insecticides on
the benthic community in a model stream. Environ. Pollut. (Ser. A) 38:31-
43.
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Marine Benthic Community Structure
CHAPTER 8. MARINE BENTHIC COMMUNITY STRUCTURE ASSESSMENT
Betsy Day, Gary Braun, and Gordon Bilyard
Tetra Tech, Inc.
11820 Northup Way, Suite 100E
Bellevue, WA 98005
(206) 822-9596
Benthic communities are communities of organisms that live in or on the
sediment. In most benthic community structure assessments, primary emphasis
is placed on determining the species that are present and the distrioution
of individuals among those species. These community attributes are
emphasized largely for pragmatic reasons. Whereas it is relatively simple
to collect, identify, and enumerate benthic organisms, it is 'very difficult
to determine first hand the spatial distributions 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 composition and abundance
is typically used in conjunction with information in the scientific
literature to infer the distributions of species and individuals in three
dimensional space and the functional attributes 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 evaluating 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
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Marine Benthic Community Structure
in marine and estuarine systems, and coral reefs. Soft-bottom habitats
include mud and sand habitats 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 7.
1.0 SPECIFIC APPLICATIONS
Assessment of benthic community structure is an in situ method that can
be used alone, as part of other approaches [e.g., Sediment Quality Triad
(see Chapter 9) and Apparent Effects Threshold (AET) (see Chapter 10)], or in
combination with other sediment assessment techniques (e.g., sediment
toxicity bioassays). It is commonly used in three ways to assess impacts to
benthic communities 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
• 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 jn 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 through a 1.0- or 0.5-mm
sieve). Reasons for the more limited use of meiofauna are twofold.
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Marine Benthic Community Structure
• Although they may be sampled quantitatively, 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
poor.ly known, and it is therefore difficult to interpret the
importance 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 megafauna (i.e., large organisms that are typically
measured in centimeters) and the microfauna (i.e., microbes) usually
preclude them from consideration in assessments of benthic community
structure. Furthermore, although the functional importance of sediment
microbes has been studied, their structural and functional characteristics
have not been used as indicators of sediment quality.
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.
1.1.1 Organic Enrichment--
Pearson and Rosenberg (1978) performed an extensive review of benthic
community succession in relation to organic enrichment of marine and
estuarine sediments. Based on that review, they developed a generalized
model of structural community changes (i.e., numbers of species, abundances,
biomass) in relation to organic enrichment, and identified opportunistic and
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Marine Benthic Community Structure
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 benthic 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 charac-
terized 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) indicated that
these stations were under stress.
1.1.2 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 contam-
ination of the sediment. His studies showed an inverse relationship
between concentrations of metals in the sediment and the species richness
and abundance of the benthic macroinvertebrate fauna. Bryan et al. (1987)
examined population distributions of the oyster Ostrea edulis, 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 maintain their original distribution range.
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Marine Benthic Community Structure
1.1.3 Contamination Due to Toxic Organic Compounds--
Toxic organic compounds are frequently associated with municipal
discharges, industrial effluents, and storm drains. These discharges may
also result in organic enrichment and contamination by metals or metalloids.
The following benthic studies provided evaluations of sediment quality in
areas primarily affected by toxic organic compounds:
• Creosote contamination. Tagatz et al. (1983) examined the
benthic communities that colonized uncontaminated sediments
and sediments contaminated with three different concentrations
of creosote (177, 844, and 4,420 ug/g) in field and laboratory
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 reduced numbers of
individuals and species. Distribution of individuals within
species was similar for the laboratory and field assemblages
of animals.
• Oil contamination. Elmgren 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 complete recovery would require at least 5 yr.
Jackson, et al. (1989) .investigated the effects 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
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Marine Benthic Community Structure
in areas exposed to the open sea had recovered. Clifton et
al. (1984) performed field experiments in Willapa Bay, WA,
and found that oil in the sediments modified the burrowing
behavior of infaunal benthos.
1.1.4 Dredging and Construction-Related Activities--
Swartz et al. (1980) examined species richness and species abundances
just before dredging occurred in Yaquina Bay, OR, 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 appearance in a disturbed area. They suggested that
the "equilibrium community is less productive than a pioneering stage" and
suggested that productivity may be enhanced through managed disturbances.
1.1.5 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 observed 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 Biscane Bay. They documented that
recolonization occurs fairly rapidly after the defaunation period.
Oscillations in macrobenthic populations in the shallow waters of the
Peruvian coast were examined by Tarazona et al. (1988). Fluctuations in
density, biomass, species composition, and diversity were attributed to the
El Nino of 1982-1983.
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Marine Benthic Community Structure
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 9). It is .also a component of the
Apparent Effects Threshold approach (see Chapter 10).
1.2 Potential Use
To date, benthic community assessments performed to evaluate sediment
quality have focused on the relationships between community variables (e.g.,
numbers of species, total abundance, 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
indicative 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 relation to organic enrichment of the sediments
in the Southern California Bight.] A tremendous potential exists, however,
for identifying species th?t ?re 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 remedial
actions at contaminated sites. To date, it has not been possible to use
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Marine Benthic Community Structure
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 valuable byproduct of such models
would be the ability to predict the capacity of the remediated area to
support higher tropic level organisms that forage on benthic organisms,
including commercially and recreationally harvested demersal fishes.
2.0 DESCRIPTION
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;
ana-lys-es 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 interpreted in conjunction with other sediment
variables (e.g., sediment grain size, total organic carbon) that were
collected concurrently with the benthic samples.
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
environment at the sampling location
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Marine Benthic Community Structure
• Benthic community structure may be altered in a predictable
manner over time and space by chemical or physical disturb-
ances
• 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.
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
required to collect the organisms. The sorting and identification process
is labor-intensive and generally expensive. Program objectives will
determine whether the data analyses are simple or complex.
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 organisms 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 communities) 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
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Marine Benthic Community Structure
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 while
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).
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 urn 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 time-consuming when the smaller mesh is used because it retains more
abiotic materials and many smaller organisms.
2.1.2.2 Methods—Methods for collecting data on benthic community
structure are divided into three categories: program design, field methods,
and laboratory methods. Each of these categories is briefly discussed below.
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Program design includes the selection of station locations, level of
replication, type of sampler, screen size, data analysis methods (discussed
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 hydrography,
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
distributions that often result from contaminant sources.
Selection of the number of replicates is an important component of
program design because 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 accuracv and nrecision that are accentabls to
meet study objectives. A single sample may be useful for general distribu-
tional or trends analyses (Cuff and Coleman 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 (Conor and Kemp 1978).
Additional research on replication is presently being conducted by the
U.S. Environmental Protection Agency in Newport, OR under the direction of
S. Ferraro (Swartz, R.C., 15 March 1989, personal communication).
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Marine Benthic Community Structure
Power analysis can assist in determining the appropriate number of
replicates. A power analysis includes consideration of the minimum
detectable 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 probability 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), 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 example, Eleftheriou and Holme
(1984) and Swartz (1978) recommend that an area of 0.5 m^ be sampled to
assess species composition in coastal and estuarine regions. Most studies
of benthic community structure routinely involve five replicate O.l-m^ grab
samples. A single O.l-m^ grab sample may be sufficient to obtain "useful
descriptive information" for use in cluster analyses (Word 1976). However,
a single sample precludes 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
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Marine Benthic Community Structure
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 O.l-m^ sample is generally not adequate to characterize
benthic community structure and function. In general, five O.l-m^ 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. However, for monitoring programs where reoccupation 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 appropriate 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 2.1.2.1.
Field and laboratory methods must be conducted 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
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Marine Benthic Community Structure
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
stations for conventional sediment chemistry variables (e.g., sediment
organic content, sediment grain size distribution). Because organic carbon
content and sediment grain size naturally affect the composition of benthic
communities, measurement of these variables will assist in determining
whether benthic communities are affected by reduced sediment quality.
2.1.2.3 Types of Data Required—The two primary structural attributes
of any benthic community are the distributions 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 functional attributes of the benthic
community, including trophic relationships, primary and secondary produc-
tivity, and interactions between the resident biota and.the abiotic habitat.
The data required for analysis of the structural and functional attributes
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 imple-
mentation of data analyses are equally important, and are discussed in the
remainder of this section. The data analyses presented in this section
address primarily structural components of benthic communities. However,
functional attributes can be inferred from many of those structural
attributes.
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Marine Benthic Community Structure
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 collected 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, descriptive indices such as diversity [the distribution of
individuals among .species; see Washington (1984) for additional definitions
of diversity], evenness (the evenness with which individuals are distributed
among taxa), and dominance (the degree to which one or a few species
dominate the community) are usually calculated.
Most programs evaluate the temporal or spatial differences in benthic
community structure. Typically, comparisons of one or more indices are made
at the same station over time and compared to a baseline value, or compari-
.sons are made between stations in a study area and stations in a rsfarsncs
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) statistical analyses,
multivariate (i.e., multiple variables) 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 (i.e.,
grouping similar stations into clusters) and ordination [i.e., representing
sample or species relationships as faithfully as possible in a low-dimen-
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Marine Benthic Community Structure
sional (two-four dimensions) space] methods (see Gauch 1982 for an overview
of multivariate methods). Multivariate techniques group data and display
them on a two-dimensional plot or dendrogram so that stations exhibiting
similar communities are located closer to one another than to stations with
dissimilar communities. The numerical and graphical results can then be
compared with physical and chemical data collected concurrently to determine
whether those variables correlate with trends in benthic communities. A
commonly used classification technique involves first computing a matrix of
similarity indices that represent the degree of similarity in species
composition between two stations. Commonly used similarity indices include
Bray-Curtis, Canberra metric, and Euclidian distance indices. The similarity
matrix is then entered into a clustering algorithm (e.g., pair-wise
averaging, flexible sorting) to produce a dendrogram depicting similarities
among stations. Commonly used ordination 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 multivariate analyses may involve
data transformations. Transformations of the original data may be necessary
for one or more of the following reesons:
• Benthic data sets are usually characterized by large
abundances of a few species and small abundances of many
species
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Marine Benthic Community Structure
• The distribution of individuals among species tends to be log-
normal
• 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). Standardizations are alterations that
depend on some property of the data under consideration. A common standardi-
zation is the conversion of values to percentages.
Benthic data are transformed to better meet the assumptions of
parametric tests (e.g., normality, homogeneity of variances). In multi-
variate analyses, data are often transformed using logarithms [e.g., log
(x+1)] because of the presence of zero scores. This transformation is also
applied 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 compared with chemical and
physical data that are collected concurrently. These comparisons may take
the form of simple linear correlations, correlations with cluster groups, or
correlations using multivariate techniques such as discriminant analyses.
Multiple discriminant analysis attempts to isolate groups of similar
stations so that variables 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 environment, or changes in sediment
quality due to toxic substances or organic materials.
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Marine Benthic Community Structure
The use of different methods and analyses may result in different inter-
pretations of the same data. For example, use of the same data with
different standardization methods in a classification analysis can yield
very different results (Austin and Grieg-Smith 1968). Generally, the more
analyses that are conducted on the-data, the higher the probability of
interpreting the data accurately.
2.1.2.4 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,
appropriate sampler, sieves, sample storage containers, buffered fixative);
and standard biological laboratory equipment (e.g., microscopes, sieves,
hydrometers or pipets, and a balance). More specialized equipment includes
a muffle furnace for determining total volatile solids concentrations, a
taxonomic reference collection, and a taxonomic reference library.
Computer equipment and appropriate software are required to make studies
cost-effective. A microcomputer is sufficient for most analyses, but some
complicated multivariate analyses may require the use of a mini- or
mainframe computer.
Trained benthic taxonomists are required to ensure accurate identifica-
tions. Some computer programming and some level of data management are
usually required. A trained benthic 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 interpretation of
descriptive indices.
2.1.3 Adequacy of Documentation--
Many different approaches and methods are used to analyze benthic data,
some of which have their origins in classical terrestrial community
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Marine Benthic Community Structure
ecology. Because analysis of benthic community structure is a relatively old
assessment tool, literally thousands of papers have been written about the
method. Several books and protocols have also been developed to describe
field and laboratory 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 describes standardized
procedures for analyzing and interpreting benthic community data is lacking.
The most commonly used interpretive approaches include measures of
diversity and classification. 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; PxaffaGlli
1987), lognormal distribution (Gray and Mirza 1979), Index 5 (Bernstein and
Smith 1986)]. Each of these methods has shown promise in some situations,
but more testing and validation 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 ordinations run on various hierarchical
levels of taxonomic 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
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Marine Senthic Community Structure
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 taxonomic level than
natural environmental variables, which influence the fauna more by species
replacement." 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 community
structure between impacted areas and reference 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 U. S.
2.2 ADD!icabilitv of Method to Human Health. Aquatic Life, or Wildlife
Protection
The assessment of benthic community structure is directly applicable to
the protection of aquatic life. Because benthic organisms are aquatic,
assessments of benthic community structure provide a direct measure of the
condition of aquatic life. Furthermore, because benthic organisms 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
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Marine Benthic Community Structure
of benthic infaunal communities (e.g., see Kleppel et al. 1980), as has the
distribution of the starfish Astropecten verilTi (Strip!in 1987).
Assessment of benthic community structure may be directly or indirectly
applied to the protection of human health. When changes in community
structure are caused by the presence of toxic contaminants, then the
bioaccumulation of those contaminants in more tolerant species may sometimes
be postulated. Those contaminated benthic infauna 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).
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 chemicals, 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 chemical-specific effects.
3.0 USEFULNESS
Assessment of benthic community structure has become a valued tool for
determining sediment 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.
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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 contamination.
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 occupying different types of sediment should not be compared.
Considerable 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 particulate
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 dif-
ferences in community structure between nonimpacted and potentially impacted
stations, the reference and test stations should exhibit, to the extent
possible, similar sediment characteristics (as v.ell as similar water depths
because benthic communities naturally vary by depth). However, it is not
always oossible for the reference and test stations to have sediment that
has a similar composition (e.g., dredged material at a dump site may have
different characteristi.es than native sediment surrounding the dump.site).
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Marine Benthic Community Structure
If the experimental design is based upon 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 assessment 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 community structure between stations with similar
grain sizes and water depths.
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,b, 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 loading
(Pearson and Rosenberg 1978). In an area receiving both organic enrichment
and toxic contaminants, it may be difficult to distinguish the effects of
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Marine Benthic Community Structure
these forms of pollution from each other. Additional research is greatly
needed to help separate the effects of multiple sources of contaminants.
3.1.3 Suitability for Predicting Effects on Different Organisms--
Changes in benthic communities that result from the presence of organic
enrichment or chemical 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
important components of this research.
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 Chapters 9, and 10, and 11) in which benthic community
structure is the only in situ biological measure.
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, benthic
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Marine Benthic Community Structure
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 chemicals or classes of chemicals present in the
sediment, they are of limited 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 indicator 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 uncontaminated particulates dilute the contaminated sediment. Further-
more, this assessment technique would be useful only in areas where other
uncontrolled sources would not obscure sediment recovery 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 monitoring regional sediment recovery. The concentration and type of the
contaminants, and hydrodynamics of the study area will govern the length of
time over which recovery will occur (Perez, K., 1 May 1989, personal
communication).
3.1.6 Suitability for Disposal Applications--
Regulations concerning biological testing of sediment that is dredged
under Sections 401 and 404 of the Clean Water Act do not include assessments
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 bioassays, and bioaccumulation
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Marine Benthic Community Structure
studies can, however, be used to assess sediment quality throughout the
dredging depth. Section 102 of the Marine Protection Research and Sanctuary
Act does call for monitoring of benthic community structure in areas where
dredged material is disposed.
The International Joint Commission (IJC) recommends use of benthic
communities to determine whether areas of concern exist in sediments that
require dredging (IJC 1988a,b). However, they do not discuss whether benthic
community structure would be used to determine the suitability of dredged
material for open-water disposal.
Analysis of benthic community structure is appropriate for post-disposal
monitoring of confined and unconfined disposal sites and for monitoring
recovery of areas that were dredged. As part of the Puget Sound Dredged
Disposal Analysis (PSDDA) post-disposal monitoring program, benthic community
structure will be used to monitor the potential transport of disposed
material away from the disposal site (Tilley et al. 1988). The purpose of
this aspect of the monitoring 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 integrity at the Commencement Bay
Nearshore/Tideflats Superfund site in Tacoma, WA (Tetra Tech 1988). As
described earlier, Swartz et al. (1980) documented recovery in Yaquina Bay,
OR following dredging. Rhoads et al. (1978) suggested that periodic
disturbance such as dredging and disposal may enhance benthic productivity.
3.2 General Advantages apj Limitations
General advantages of using benthic community 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
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Marine Benthic Community Structure
theoretical approach. However, as with most assessment techniques involving
field studies, the evaluation of benthic communities is costly and time-
consuming. 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 technique has been incorporated into other predictive
techniques (see Chapters 9, 10, and 11) that provide information more easily
used by resource managers.
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 established
criteria. However, the use of benthic community structure as a sediment
quality assessment tool is widely accepted, and trained benthic ecologists
are available throughout the country. By using highly experienced in-
dividuals to conduct the field, laboratory, and data analysis work,
potential problems (such as generating "noisy" data that obscure real
trends, or arriving at different interpretations using the same data) should
not occur.
3.2.2 Relative Cost--
The relative cost of conducting an assessment of benthic communities
is less than the cost to 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, then the relative cost of conducting a benthic survey is greater
than the cost of analyzing sediment for contaminant concentrations and
comparing those data to the values to determine sediment quality. Sediment
toxicity bioassays are generally less costly than analysis of replicated
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Marine Benthic Community Structure
benthic samples. Because the Triad approach requires synoptic analyses of
sediment chemistry, sediment toxicity, and benthic communities, 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 assessment 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., S400-S1,000),
but varies greatly depending on the size of the area sampled, the screen
size, the level of the taxonomic identifications, and the environment
sampled.
3.2.3 Tendency to be Conservative--
Benthic community structure is a moderately conservative measure of
sediment quality. Because benthic community structure reflects the
collective response of all species, responses of individual species that are
susceptible—to degradation 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 environmental 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.
3.2.4 Level of Acceptance--
Benthic community assessments have been used as a sediment quality
assessment tool for several decades in North America, Europe, and Australia,
as well as in South Africa, China, and Japan. The method has gained
widespread acceptance because of its inherent capability to assess sediment
quality at the community level, thereby documenting ecological response to
sediment perturbations.
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Marine Benthic Community Structure
Many methods may be used to analyze benthic 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 log-normal distribu-
tion (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.
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 oversee 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
component of taxonomic laboratories.
3.2.6 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 station, taxonomic level of the identifi-
cations, 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
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Marine Benthic Community Structure
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 require 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.
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 of the fauna and the statistical techniques
that are routinely used to analyze the data. Interpretation of the many
data points generated by this approach may require many weeks before
meaningful trends are recognized. The inherent variability of benthic
communities has so far prevented the development of specific benthic
criteria for use in assessing pollutant-related trends in sediment quality.
3.2.8 Degree of Environmental Applicability--
The assessment of benthic community structure is a direct measure of the
environmental effects of pollutants and, as such, is highly applicable as a
method to assess sediment quality. Its applicability lies in its ability
to provide information on the effects of pollutants on ecological processes
within the sedimentary environment.
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 (defined as the consistency and reliability of the
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Marine Benthic Community Structure
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 community 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 designing 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
placement. 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.
4.0 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.
4.1 Extent of Use
Assessment of benthic community structure has been a valued tool in
marine, estuarine, and freshwater environments for several decades. Many
of the early programs examined benthic communities from an academic
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Marine Benthic Community Structure
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,
eutrophication, organic enrichment, oil spills, and mine tailings disposal
(see Section 1.1). It has also been used to determine the suitability of
sediments for dredged material disposal, to monitor dredged material
disposal sites, and to monitor recovery of impacted areas following the
cessation of contaminant loading.
4.2 Extent to Which Approach Has Been Field-Validated
Because benthic community structure is an in situ sediment quality
assessment tool, it does not require additional field validation.
4.3 Reasons for Limited Use
Although conducting studies of benthic community structure is a common
practice, the cost and amount of time required to generate usable results may
prevent the method from being implemented 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 probability
of producing valuable data and results.
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 assessment tool is particularly bright because of the
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Marine Benthic Community Structure
continuing development of new data analysis methods by researchers in North
America and Europe. The objective of these methods is generally to reduce
cost or variability within the data by relating aspects of the distributions
of organisms or organism biomass to specific kinds of environmental perturba-
tions. Gray and Mirza (1979) determined that the log-normal 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 abundance/biomass comparison (ABC)
method developed by Warwick (1986). This method proposes that the relation-
ship 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.
Rhoads and Germano (1986) developed the REMOTS™ (remote ecological mapping
of the seafloor) system. They use a vessel-deployed sediment-profile camera
to photograph vertical sections of the sediment. Although REMOTS™ cannot
determine the species composition of the benthic community, it can document
relationships between organisms and sediment. Rhoads and Germano (1986)
characterized the successional stages of benthic communities, and suggest
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Marine Benthic Community Structure
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.
The sediment profile camera has been used for a variety of other
purposes including assessing the relationships between sediment quality and
eutrophication (Day et al. 1987; Revelas et al. 1987; Rhoads, D.C., 1 May
1989, personal communication), monitoring the perimeter of dredged material
disposal sites (Rhoads, O.C., 1 May 1989, personal communication; Diaz,
R.J., 1 May 1989, 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.
5.0 REFERENCES
Amjad, S., and J.S. Gray. 1983. Use of the nematode/copepod ratio as an
index of organic pollution. Mar. Poll. Bull. 14:178-181.
Austin, M.P., and P. Grieg-Smith. 1968. The application of quantitative
methods to vegetation survey. II. Some methodological problems of data
from rain forest. J. Ecol. 56:327-844.
Beukema, J.J. 1988. An evaluation of the ABC method (abundance-biomass
comparison) as applied to macrozoobenthic communities living on tidal flats
in the Dutch Wadden Sea. Mar. Biol. 99:425-433.
Bernstein, B.B., and R.W. Smith. 1986. Community Approaches to Monitoring.
IEEE Oceans '86 Conference Proceedings, Washington, DC, September 23-25,
1986. pp. 934-939.
Bilyard, G.R. 1987. The value of benthic infauna in marine pollution
monitoring studies. Mar. Poll. Bull. 18:581-585.
Bishop, J.D.D., and J.P. Hartley. 1986. A comparison of the fauna retained
on 0.5 mm and 1.0 mm meshes from benthic samples taken in the Beatrice
Oilfield, Moray Firth, Scotland. Proc. Royal Soc. Edinburgh. 918:247-262.
8-34
-------
Marine Benthic Community Structure
Boesch, D.F. 1977. Application of numerical classification in ecological
investigations of water pollution. EPA 600/3-77-033. U.S. Environmental
Protection Agency, Corvallis, OR. 115 pp.
Bros, W.E., and B.C. Cowell. 1987. A technique for optimizing sample size
(replication). J. Exp. Mar. Biol. Ecol. 114:63-71.
Bryan, G.W., P.E. Gibbs, L.G. Hummerstone, G.R. Burt. 1987. Copper, zinc,
and organotin as long-term factors governing the distribution of organisms
in the Fal Estuary in Southwest England. Estuaries 10:208-219.
Clifford, H.T., and W. Stephenson. 1975. An introduction to numerical
classification. Academic Press, San Francisco, CA. 229 pp.
Clifton, H.E., K.A. Kvenvolden, and J.P. Rapp. 1984. Spilled oil and
infaunal activity-modification of burrowing behavior and redistribution of
oil. Mar. Environ. Res. 11:111-136.
Cohen, J. 1977. Statistical power analysis for the behavioral sciences.
Academic Press, New York, NY.
Cuff, W., and N. Coleman. 1979. Optional survey design: lessons from a
stratified random sample of macrobenthos. J. Fish. Res. Bd. Can.
36:351-361.
Oauer, D.M., and W.G. Conner. 1980. Effects of moderate sewage input on
benth-fc-polychaete populations. Est. Mar. Sci. 10:335-346.
Day, B., L.C. Schaffner, R.J. Diaz, and J. Ryther, Jr. 1987. Long Island
Sound sediment quality survey and analyses. Prepared for National Oceanic
and Atmospheric Administration, Rockville, MO. Evans-Hanmilton, Inc.,
Seattle, WA. 113 pp. + appendices.
Day, J.H., J.G. Field, and M.P. Montgomery. 1971. The use of numerical
methods to determine the distribution of the benthic fauna across the
continental shelf of North Carolina. J. Anim. Ecol. 40:93-125.
Diaz, R.J. 1 May 1989. Personal Communication (phone by Ms. Betsy Day,
Tetra Tech, Inc., Bellevue, WA regarding uses of the sediment profile camera
system). Virginia Institute of Marine Science, Gloucester Point, VA.
Dobbs, F.L., and J.M. Vozarik.- 1983. Immediate effects of a storm on
coastal infauna. Mar. Ecol. Prog. Ser. 11:273-279.
Eleftheriou, A., and N.A. Holme. 1984. Macrofauna techniques, pp. 140-
216. In: Methods for the Study of Marine Benthos. N.A. Holme and A.O.
Mclntyre (eds). Blackwell Scientific Publications, Oxford, U.K.
8-35
-------
Marine Benthic Community Structure
Elliott, J.M. 1977. Some methods for the statistical analysis of samples
of benthic invertebrates. 2nd ed. Freshwater Biological Association.
Titus Wilson & Son Ltd., Kendal, U.K. 156 pp.
Elmgren, R., S. Hansson, U. Larsson, B. Sundelin, and P.O. Boehm. 1983.
The "Tsesis" oil spill: acute and long-term impact on the benthos. Mar.
Biol. 73:51-65.
Fabrikant, R. 1984. The effect of sewage effluent on the population
density and size of the clam Parvilucina tenuisculpta. Mar. Poll. Bull.
15:249-253.
Gauch, H.G. 1982. Multivariate analysis in community ecology. Cambridge
University Press, New York, NY. 298 pp.
Conor, J.J., and P.F. Kemp. 1978. Procedures for quantitative ecological
assessments in intertidal environments. EPA 600/3-78-078. U.S. Environ-
mental Protection Agency, Corvallis, OR. 104 pp.
Gray, J.S., and F.B. Mirza. 1979. A possible method for the detection of
pollution-induced disturbance on marine benthic communities. Mar. Po-11.
Bull. 10:142-146.
Green, R.H., and G.L. Vascotto. 1978. A method for analysis of environ-
mental factors controlling patterns of species composition in aquatic
communities. Water Res. 12:583-590.
Grizzle, R.E. 1984. Pollution indicator species of macrobenthos in a
coastal lagoon. Mar. Ecol. Prog. Ser. 18:191-200.
Holme, N.A., and A.O. Mclntyre (eds). 1984. Methods for the study of
marine benthos. Blackwell Scientific Publications, Oxford, U.K. 387 pp.
Ibanez, F., and J. Oauvin. 1988. Long-term changes (1977-1987) in a muddy
fine sand Abra alba-Melinna palmata community from the western English
Channel: multivariate time-series analysis. Mar. Ecol. Prog. Ser. 49:65-
81.
International Joint Commission. 19S8a. Procedures for the assessment of
contaminated sediment problems in the Great Lakes. IJC, Windsor, Ontario,
Canada. 140 pp.
International Joint Commission. 1988b. Options for the remediation of
contaminated sediments in the Great Lakes. IJC, Windsor, Ontario, Canada.
78 pp.
8-36
-------
Marine Benthic Community Structure
Jackson, J.B.C., J.D. Cubit, B.D. Keller, V. Batista, K. Burns, H.M. Caffey,
R.L. Caldwell, S.D. Garrity, C.D. Getter, C. Gonzales, H.M. Guzman, K.W.
Kaufman, A.M. Knap, S.C. Levings, M.J. Marshall, R. Steger, R.C. Thompson,
and E. Weil. 1989. Ecological effects of a major oil spill on Panamanian
coastal marine communities. Sci. 243:37-44.
Kleppel, G.S., J.Q. Word, and J. Roney. 1980. Demersal fish feeding in
Santa Monica Bay and off Palos Verdes. pp. 309-318. In: Coastal Water
Research Project Biennial Report 1979-1980. Southern California Coastal
Water Research Project, El Segundo, CA.
Kronberg, I. 1987. Accuracy of species and abundance minimal areas
determined by similarity area curves. Mar. Biol. 96:555-561.
Lambshead, P.J.D. 1984. The nematode/copepod ratio, some anomalous results
from the Firth of Clyde. Mar. Poll. Bull. 15:256-259.
Long, B., and J.B. Lewis. 1987. Distribution and community structure of
the benthic fauna of the north shore of the Gulf of St. Lawrence described
by numerical methods of classification and ordination. Mar. Biol. 95:93-101.
Lunz, J.D., and D.R. Kendall. 1982. Benthic resource analysis technique, a
method for quantifying the effects of benthic community changes on fish
resources, pp. 1021-1027. In: Conference Proceedings on Marine Pollution,
Oceans 1982. National Oceanic and Atmospheric Administration, Office of
Marine Pollution Assessment, Rockville, MO.
Mirza, F.B., and J.S. Gray. 1981. The fauna of benthic sediments from the
organically enriched Oslofjord, Norway. J. Exp. Mar. Biol. Ecol. 54:181-207.
Moore* S.F.j and D.B. McLai!nh!in. 1978. Design of field experiments to
determine the ecological effects of petroleum in intertidal ecosystems.
Water Res. 12:1091-1099.
Nalepa, T.F., M.A. Quigley, and R.W. Ziegler. 1988. Sampling efficiency of
the ponar grab in two different benthic environments. J. Great Lakes
Research 14:89-93.
Nemec, A.F.L., and R.O. Brinkhurst. 1988a. Using the bootstrap to assess
statistical significance in the cluster analysis of species abundance data.
Can. J. Fish. Aquatic Sci. 45:965-970.
Nemec, A.F.L., and R.O. Brinkhurst. 1988b. The Fowlkes-Mailows statistic
and the comparison of two independently determined dendrograms. Can. J.
Fish. Aquatic Sci. 45:971-975.
Parker, H.R. 1975. The study of benthic communities. A model and review.
Elsevier Oceanography Series 9. Elsevier, Amsterdam.
8-37
-------
Marine Benthic Community Structure
Pearson, T.H., and R. Rosenberg. 1978. Macrobenthic succession in relation
to organic enrichment and pollution of the marine environment. Oceanogr.
Mar. Biol. Annu. Rev. 16:229-311.
Perez, K. 1 May 1989. Personal communication (phone by Ms. Betsy Day,
Tetra Tech, Inc., Bellevue, WA regarding mesocosm experiments to determine
rates of bentnic recovery). U.S. Environmental Protection Agency, Environ-
mental Research Laboratory, Narragansett, RI.
Preston, F.W. 1948. The commonness, and rarity, of species. Ecology
29:254-283.
PTI and Tetra Tech. 1988. Elliott Bay Action Program: Analysis of toxic
problem areas. Draft Report. Prepared for the U.S. Environmental Protection
Agency, Region X, Office of Puget Sound. Tetra Tech, Inc., Bellevue, WA.
Raf f aell i , 0. 1987. The behavior of the nematode/copepod ratio in organic
pollution studies. Mar. Environ. Res. 23:135-152.
Revelas, E.G., O.C. Rhoads, and J.O. Germano. 1987. San Francisco Bay
sediment quality survey and analyses. Prepared for National Oceanic and
Atmospheric Administration, Rockville, MO. Science Applications Inter-
national Corporation, Newport, RI. 127 pp. + appendices.
Rhoads, D.C. 1 May 1989. Personal Communication (phone by Ms. Betsy Day,
Tetra Tech, Inc., Bellevue, WA regarding uses of the REMOTS™ sediment
profile camera system). Science Applications International Corporation,
Woods Hole, MA.
Rhoads, O.C., and L.F. Boyer. 1982. The effects of marine benthos on
physical properties of sediments: a successional perspective. pp. 3-52.
In: Animal -Sediment Relations. P.L. McCall and M.J.S. Trevesz (eds).
Plenum Press.
Rhoads, D.C., and J.D. Germano. 1986. Interpreting long-term changes in
benthic community structure: a new protocol. Hydrobiologia.
Rhoads, D.C., and O.K. Young. 1970. The influence of deposit-feeding
organisms on sediment stability and community trophic structure. J. Mar.
Res. 28:150-178.
Rhoads, O.C., P.L. McCall, and J.Y. Yingst. 1978. Disturbance and
production on the estuarine seafloor. Amer. Sci. 66:577-586.
B. 1985a. Distribution of species along pollution-induced diversity
gradients in benthic communities in Norwegian fjords. Mar. Poll. Bull.
12:469-474.
8-38
-------
Marine Benthic Community Structure
Rygg, B. 1985b. Effect of sediment copper on benthic infauna. Mar. Ecol.
Prog. Ser. 25:83-89.
Rygg, B. 1986. Heavy-metal pollution and log-normal distribution of
individuals among species in benthic communities. Mar. Poll. Bull. 17:31-
36.
Saila, S.B., R.A. Pikanowski, and O.S. Vaughan. 1976. Optimum allocation
strategies for sampling benthos in the New York Bight. Est. Coast. Mar.
Sci. 4:119-128.
Sanders, H.L. 1968. Marine benthic diversity: a comparative study. Amer.
Nat. 102:243-282.
Santos, S.L., and J.L. Simon. 1980. Response of soft-bottom benthos to
annual catastrophic disturbance in a south Florida estuary. Mar. Ecol.
Prog. Ser. 3:347-355.
Schaffner, L.C., and R.J. Diaz. 1988. Distribution and abundance of
overwintering blue crab Callinectes sapidus in the lower Chesapeake Bay.
Estuaries 11:68-72.
Self, S.G., and R.H. Mauritsen. 1988. Power/sample size calculations for
generalized linear models. Biometrics 44:79-86.
Shiells, G.M., and K.J. Anderson. 1985. Pollution monitoring using the
nematode/copepod ratio, a practical application. Mar. Poll. Bull. 16:62-
68.
Shin, P.K.S. 1982. Multiple discriminant analysis of macrobenthic infaunal
assemblages. J. Exp. Mar. Biol. Ecol. 59:39-50.
Sokal, R.R., and F.J. Rohlf. 1981. Biometry. 2nd ed. W.H. Freeman and
Company, San Francisco, CA, 859 pp.
Stephenson, M., and G.L. Mackie. 1988. Multivariate analysis of cor-
relations between environmental parameters and cadmium concentrations in
HyalTella azteca (Crustacea: Amphipoda) from central Ontario lakes. Can. J.
Fish. Aquatic Sci. 45:1705-1710.
Stephenson, W., W.T. Williams, and G.W. Lance. 1970. The macrobenthos of
Moreton Bay. Ecol. Managr. 40:459-494.
Stephenson, W., W.T. Williams, and S.D. Cook. 1972. Computer analyses of
Petersen's original data on bottom communities. Ecol. Monogr. 42:387-415.
Stephenson, W., W.T. Williams, and S.D. Cook. 1974. The benthic fauna of
soft bottoms, Southern Moreton Bay. Mem. Qd. Mus. 17:73-123.
8-39
-------
Marine Benthic Community Structure
Striplin, P.L. 1987. Resource utilization by Astropecten verrilli along
gradients of organic enrichment. M. Sc. Thesis. California State University
at Long Beach, Long Beach, CA. 108 pp. •»• appendices.
Swartz, R.C. 1978. Techniques for sampling and analyzing the marine
macrobenthos. EPA 600/3-78-030. U.S. Environmental Protection Agency,
Corvallis, OR. 27 pp.
Swartz, R.C., W.A. DeBen, F.A. Cole, and L.C. Bentsen. 1980. Recovery of
the macrobenthos at a dredge site in Yaquina Bay, Oregon, pp. 391-408. In:
Contaminants and Sediments, Vol. 2. R. Baker (ed). Ann Arbor Science, Ann
Arbor, MI.
Swartz, R.C. 15 March 1989. Personal communication (phone by Ms. Betsy
Day, Tetra Tech, Inc., Bellevue, WA regarding status of replication study
using samples collected during the Everett Harbor Action Program survey).
U.S. Environmental Protection Agency, Newport, OR.
Tagatz, M.E., G.R. Plaia, C.H. Deans, and E.M. Lores. 1983. Toxicity of
creosote-contaminated sediment to field-and laboratory-colonized estuarine
benthic communities. Environ. Tox. Chem. 2:441-450.
Tarazona, J., H. Salzwedel, and W. Arntz. 1988. Oscillations of macro-
benthos in shallow waters of the Peruvian central coast induced by El Nino
1982-83. J. Mar. Res. 46:593-611.
Tetra Tech. 1986a. Quality assurance/quality control JQA/QC) for 301(h)
•onitoring programs: guidance on field and laboratory methods. Prepared
for the U.S. Environmental Protection Agency, Office of Marine and Estuarine
Protection, Marine Operations Division, Washington, DC. Tetra Tech, Inc.,
Bellevue, WA.
Tetra Tech. 1986b. Recommended protocols for measuring selected environ-
mental variables in Puget Sound. Prepared for the Puget Sound Estuary
Program, U.S. Environmental Protection Agency, Region X, Seattle, WA. Tetra
Tech, Inc., Bellevue, WA.
Tetra Tech. 1987. Technical support document for ODES statistical power
analysis. Prepared for Marine Operations Division, Office of Marine and
Estuarine Division, Office of Marine and Estuarine Protection, U.S. Environ-
mental Protection Agency. Tetra Tech, Inc., Bellevue, WA. 34 pp. + ap-
pendices.
Tetra Tech. 1988. Commencement Bay nearshore/tideflats feasibility study.
Prepared for Washington Department of Ecology and U.S. Environmental
Protection Agency. Tetra Tech, Inc., Bellevue, WA.
8-40
-------
Marine Benthic Community Structure
Tilley, S., D. Jamison, J. Thornton, B. Parker, and J. Malek. 1988.
Management plans technical appendix. Prepared for Puget Sound Dredged
Disposal Analysis. U.S. Army Corps of Engineers, Seattle, WA.
Vezina, A.F. 1988. Sampling variance and the design of quantitative surveys
of the marine benthos. Mar. Biol. 97:151-155.
Vidakovic, J. 1983. The influence of raw domestic sewage on density and
distribution of meiofauna. Mar. Poll. Bull. 14:84-88.
Warwick, R.M. 1986. A new method for detecting pollution effects on
marine macrobenthic communities. Mar. Biol. 92:557-562.
Warwick, R.M. 1988. The level of taxonomic discrimination required to
detect pollution effects on marine benthic communities. Mar. Poll. Bull.
19:259-268.
Warwick, R.M., T.H. Pearson, and Ruswahyuni. 1987. Detection of pollution
effects on marine macrobenthos: further evaluation of the species abun-
dance/biomass method. Mar. Biol. 95:-193-200.
Washington, H.G. 1984. Diversity, biotic, and similarity indices. A
review with special relevance to aquatic ecosystems. Water Res. 18:653-694.
Winer, B.J. 1971. Statistical principles in experimental design. McGraw-
Hill Book Company, New York, NY.
Word, J.Q. 1976. Biological comparison of—grab" sampling devices.
pp. 189-194. In: Coastal Water Research Project Annual Report. Southern
California Coastal Water Research Project, El Segundo, CA.
Word, J.Q. 1373. The infaunal trophic index. pp. 19-39. In: Coastal
Water Research Project Annual Report for 1978. Southern California Coastal
Water Research Project, El Segundo, CA.
Word, J.Q. 1980. Classification of benthic invertebrates into infaunal
trophic index feeding groups. pp. 103-121. In: Coastal Water Research
Project. Biennial Report of the years 1979-1980. W. Bascom (ed). Southern
California Coastal Water Research Project, Long Beach, CA.
Word, J.Q., B.L. Myers, and A.J. Mearns. 1977. Animals that are indicators
of marine pollution. pp. 199-206. In: Coastal Water Research Project
Annual Report. Southern California Coastal Water Research Project, El
Segundo, CA.
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Sediment Quality Triad
CHAPTER 9. SEDIMENT QUALITY TRIAD APPROACH
Peter M. Chapman
E.V.S. Consultants Ltd.
195 Pemberton Avenue
North Vancouver, BC
Canada V7P 2R4
(604) 986-4331
The Sediment Quality Triad (Triad) approach is an effects-based
approach to describing sediment quality. It typically incorporates measures
of sediment chemistry, sediment toxicity, and benthic infauna communities,
although other variables can be used. This combination method is both
descriptive 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, in press-a).
1.0 SPECIFIC APPLICATIONS
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 et al. 1986, 1987a, 1988, in preparation; Chapman in press-
fa). It can also be used to determine numerical sediment quality criteria
directly (e.g., Chapman 1986, in press-a) and, through manipulations, to
determine Apparent Effects Threshold (AET) values (see Chapter 10). Triad
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, and in freshwater environments including the Great Lakes
(Long and Chapman 1985; Chapman et al . 1986, 1987a, 1988, in press-a, in
9-1
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Sediment Quality Triad
preparation, unpublished). Current uses of the Triad approach are summarized
in Table 9-1 and discussed in Section 3.1 (Environmental Applicability).
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
• To predict where such degradation will occur based on levels
of contamination and toxicity.
It can be used in any number of situations and is not restricted to aquatic-
sediments. For example, Triad can be used in water column work with phyto-
plankton and in terrestrial hazardous waste dump studies with other
organisms of concern. Other uses are described in Section 3.1.
2.0 DESCRIPTION
2.1 Description of Method
The Triad approach consists of three components (Figure 9-1):
• Sediment chemistry, to measure chemical contamination
• Sediment bioassays, to measure toxicity
9-2
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BULK
SEDIMENT CHEMISTRY
Rafervnca: Adapted from Chapman (1966).
Figure 9-1. Conceptual modal ol tfw Sediment Quality Tnad. wmcn combines data from chemistry, toxiaty bioassays.
and m situ studies. Chemistry and bioassay asomates are based on laooratory measurements with Meld
collected sediments, in vtu studies generally include, but are not limited to. measures of bentnic
community structure. Areas where the tnree facets ot the tnad show the greatest overlap (m terms of
eiffier posiove or negaove results) provide me strongest data for determining sediment quality cntena.
9-3
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TABLE 9-1. CURRENT USES OF THE SEDIMENT QUALITY TRIAD APPROACH
Use
Comment
General Locations
Where Implemented3
Prioritize areas for
remedial actions
Determine size of areas
for remedial actions
Verify quality of
reference areas
Determine contaminant
concentrations always
associated with effects
Describe ecological
relationships between
sediment properties and
biota at risk
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
environmental protection
PS, GM, SF,
VH, FW
PS
PS
PS
PS, VH, FW
PS
GM
SF
VH
FW
Puget Sound, various locations (Long and Chapman 1985).
Gulf of Mexico, oil platform (Chapman et al. 1988, in preparation).
San Francisco Bay, various locations (Chapman et al. 1986, 1987a).
Vancouver Harbor, Canada, various locations (Chapman et al. 1989).
Various freshwater environments (Chapman unpublished data; Rogers
Texas State, unpublished data).
North
9-4
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Sediment Quality Triad
• 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 ottv-
components. For instance, sediment chemistry 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 that 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 dredging) 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 (e.g., grain size, organic content, pH, Eh, chemical form,
presence of other chemicals), the importance of any particular concentration
of a chemical or suite of chemicals in sediments cannot be determined solely
from chemical measurements.
The three components of the Triad approach integrate chemical and
biological response data. They also provide the strong 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
nondegraded conditions. Conclusions that can be drawn from intermediate
responses are listed in Table 9-2.
9-5
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TABLE 9-2. POSSIBLE CONCLUSIONS PROVIDED BY
USING THE SEDIMENT QUALITY TRIAD APPROACH
Possible
Outcome Contamination Toxicity Alteration Possible Conclusions
1. + + + Strong evidence for pollution-
induced degradation
2. - - Strong evidence for absence of
pollution-induced degradation
3. * - Contaminants are not bioavailable
4. - + Unmeasured chemicals or conditions
exist that have the potential to
cause degradation
5. - - + Alteration is probably not due to
toxic chemical contamination
6. + •»• - Toxic chemicals are stressing the
system
7. - •»• + Unmeasured toxic chemicals are
causing degradation
8. -i- - •»• Chemicals are not bioavailable or
alteration is not due to toxic
chemicals
a -t- » Measured difference between test and control or reference conditions.
- - No measurable difference between test and control or reference conditions.
9-6
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Sediment Quality Triad
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 interactions between contami-
nants in complex sediment mixtures (e.g., additivity,
antagonism, synergism), 2) the actions of unidentified toxic
chemicals, and 3) the effect of environmental factors that
influence biological responses (including toxicant concen-
trations)
• Selected chemical contaminant concentrations are appropriate
indicators of overall chemical contamination
• Bioassay test results and values of selected benthic community
structure variables are appropriate indicators of biological
effects.
These components are presently treated in an additive manner, with each
having equal weight because there is insufficient information available to
assign weightings.
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 detailed by Chapman (1988), and benthic infauna
samples should be collected at the same sampling location. Chemistry and
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Sediment Quality Triad
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
separate sampling containers. In general, chemistry and bioassay samples
include laboratory rather than field 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 7, Section 2.1.2.2
herein). Coincident rather than synoptic sampling is possible (e.g., Long
and Chapman 1985), but data interpretation is complicated by spatial
heterogeneity in sediment contamination and toxicity (cf. Swartz et al.
1982).
Adequate quality assurance/quality control (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 1986b).
The first component of Triad involves identification and quantifi-
cation of inorganic and organic contaminants present in the sediments.
Chemical analytes measured are generally restricted by equipment, technology,
and the availability of funds and facilities. Local concerns and existing
data also affect target analytes measured. Fiscal conservatism, if a
factor, must be balanced against the need for a analytical database suffi-
ciently large to allow determination of the presence (or absence) of known
toxicants of concern.
An example of the types and classes of compounds required to provide a
reasonable characterization of chemical contamination is shown in Table 9-3.
Total organic carbon and grain size are measured to provide a basis for
normalizing the data to different types of sediments. Coprostanol, an
9-8
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TABLE 9-3. EXAMPLE ANALYTES AND DETECTION LIMITS
FOR USE IN THE CHEMISTRY COMPONENT OF TRIAD
Detection
Analyte Limit
Analyte
Detection
Limit
Conventional? (ma/ka. dry}
Grain size
TOCa
Sul fides
Inoraanics (ma/ka
Arsenic
Iron
Chromium
Copper
Cadmium
Lead
Mercury
Nickel
Si Iver
Selenium
Zinc
Oraanics fua/ka.
LPAH°
Benzo(a)pyrene
8enz(e)pyrene
n/a
n/a
0.5
. drv)
0.05
2.5
1.0
0.5
0.05
0.05
0.01
1.0
0.05
0.05
0.5
dry)
5
10
10
Bipnenyl
Perylene
Coprostanol
op '-ODD
op '-ODE
op '-DOT
pp'-OOO
pp'-ODE
pp'-ODT
Oieldrin
Heptachlor
Hexach 1 orobenzene
Lindane
Mi rex
PCBsc
PCPd
TCPe
5
5
10
0.15
0.25
0.15
0.15
0.10
0.10
0.10
0.10
0.10
0.15
0.10
2.5
1.0
1.0
Benz(a) anthracene 10
Chrysene
Dibenzanthracene
Fluoranthene
Pyrene
10
16
5
5
The detection limits are the instrumental estimates. Actual detection
limits may be higher because of matrix effects.
a TOC = total organic carbon.
b LPAH = low molecular weight polycyclicaromatic hydrocarbons (includes
acenapthene, anthracene, naphthalene and methylated naphthalenes, fluorene,
phenanthrene, and methylated phenanthrenes).
c PCBs = polychlorinated biphenyls.
d PCP = pentachlorophenol.
e TCP = tetrachlorophenol.
9-9
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Sediment Quality Triad
indicator of human waste, is measured to differentiate sewage inputs from
industrial inputs.
The second Triad component involves identification and quantification
of toxicity based on laboratory tests using field-collected sediments.
Ideally, one would test the toxicity of the sediments to all ecologically
and commercially important 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 character-
ization of the degree of toxicity are shown in Table 9-4. Sediment
bioassays with estuarine waters have been developed but are not yet as well
accepted as those for fresh and marine waters. Obvious omissions 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 in press).
The final Triad component involves the evaluation of in situ biological
alteration. Generally this component is provided by benthic infauna
community data because benthic organisms are relatively sessile and
location-specific. Histopathology of bottom fish has also been used for
this Triad component (Chapman 1986), but for area-wide rather than site-
specific studies, as 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, total abundance,
and percentage composition of major taxonomic groups. In the marine
environment, this last category includes any or all of polychaetes,
amphipods, molluscs, and echinoderms. In the freshwater environment,
oligochaetes, chironomids, and other major insect groups would fit into the
last category.
9-10
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TABLE 9-4. POSSIBLE STATIC SEDIMENT BIOASSAYS
Bioassay
Duration
Endpoint
Amount of
Sediment
Required (L)
Marine Waters
Rhepoxynius
(adult amphi
abronius
pod)
Bivalve larvae
development
Neanthes sp.
10
48
20
days
h
days
Survival ,
Survival ,
Survival ,
avoidance
development
growth
1
0
2
.5
.5
.0
(juvenile polychaetes)
Fresh Waters
Hyalella azteca 10 days
(adult amphipod)
Qaphnia magna 10 days
(water flea)
Chironomus tentans 25 days
(juvenile insect)
Estuarine Waters
Eohauston'us estuarinus 10 days
(adult amphipod)
Survival, avoidance
Survival, growth
Survival, avoidance
1.5
Survival, reproduction 0.5
1.5
1.5
9-11
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Sediment Quality Triad
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 (Figure 9-1). All data are compared on a quantitative basis,
and are normalized to reference site values by converting them to ratio-to-
reference (RTR) values as described by Chapman et al. (1986, 1987a) and
Chapman (in press-b). 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 concentration 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 dif-
ferences 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
stations for which data are averaged.
The RTR criterion is based but not dependent on the assumption that
the reference site concentrations are, in fact, 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 is calculated for each
station by separately determining 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. An index of
toxicity is calculated by combining bioassay RTR values as a single mean
value. An index of biological alteration is calculated in the same manner
as is toxicity, using benthic community structure data. The indices of
contamination are used, to rank stations. These summary ranks are also
compared with the ranks generated using the sediment bioassay and infaunal
data.
9-12
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Sediment Quality Triad
The composite RTR values for each Triad component can also provide
useful visual indices. 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 9-2,
for the eight possible situations detailed in Table 9-2. These plots also
provide a visual guide to the characteristics of "background" or reference
stations. Since reference data usually involve a site containing more than
one reference station, RTR comparisons should also be made against individual
reference stations.
2.1.2.1 Type of Sampling Required — Synoptic sampling is preferred for
all three Triad components, as described above. Any reasonable sampling
procedure can be used if it provides suitable sediment samples for quanti-
fying sediment contamination, toxicity, and biological alteration. Studies
to date have used remote samplers such as a O.l-m^ van Veen grab operated
from a vessel.
2.1.2.2 Methods—Typical variables included in the chemical analyses
and sediment bioassays are listed in Tables 9-3 and 9-4, respectively.
Details for benthic infauna analyses are provided in Chapter 7 herein.
Although unit costs vary, costs are typically on the order of $1,500 for
three separate replicated (n-5) sediment bioassays, $1,500 for unreplicated
chemical analyses, and $2,000 for replicated (n-5) benthos.
2.1.2.3 Types of Data Required —Standard measurements of chemistry,
toxicity, and biological alteration are required. These measurements are
then combined, as described above.
2.1.2.4 Necessary Hardware and Ski! 1 s—Appropriate sampling equipment
and trained field and laboratory personnel are required for chemical
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Toxicmr
• I
x.
t . i CONTAMINATION
t .1
ALTERATION
Toxicmr
1 .1 CONTAMINATION
< . i
ALTERATION
Toxicmr i. >
1 .1 CONTAMINATION
ALTERATION
Reference: Chapman (in pnws-b).
"jre 9-2. The Sediment Quality Triad determined, m 9ie example situation, for each of the eight possible outcomes
described in Table 9-2. Toxictty. contammaoon. and altaraoon are shown normalized to Raoo-to-Aeferences
values as described by Chapman et al. (1986. I987a). 1 0 • reference conditions. Note tnat ffie exact symmeffy in
these examples would not be reuonely expected in actual studies.
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Sediment Quality Triad
analyses, toxicity testing, and benthic infaunal analyses. Although the
equipment required can be both costly and sophisticated, it is in common use
for investigations related to sediment contamination. The necessary
equipment, facilities, and expertise are generally available through a wide
variety of government, university, commercial, and private facilities.
2.1.3 Adequacy of Documentation--
Documentation for use of this method is provided by Long and Chapman
(1985), Chapman (1986, in press-a, in press-b), and Chapman et al. (1986,
1987a, 1988, in preparation). Using many of the references cited immediately
above, other investigators have successfully applied this method (e.g.,
Wiederholm et al. in press).
2.2 Apolicabilitv 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
components are redirected towards issues such as bacterial contamination and
toxic contaminant bioaccumulation. For instance, Triad could be used to
address bacterial problems by 1) measuring bacterial contamination in water
or sediment, 2) measuring bacterial diseases or concentrations in tissues,
and 3) performing laboratory tests to quantify relationships between
sediment/water concentrations and effects. Toxic contaminant bioaccumu-
lation could be addressed by 1) measuring toxic contaminant concentrations
in water' or sediment, 2) measuring bioconcentration/biomagnification in
tissues, and 3) performing laboratory tests to determine effects related to
bioconcentration and.biomagnification.
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Sediment Quality Triad
2.3 Ability of Method to Generate Numerical Criteria for Specific Chemicals
Triad 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 areas and concentrations that were
associated with no or minimal biological effects. The criteria fall within
a factor of 2 to 10 of values generated for these contaminants by the
screening level concentration (see Chapter 10, Section 1.1.2), the AET
approach (see Chapter 10), and laboratory toxicity methods (Chapman et al.
1987b). As detailed by Chapman (in press-a), the Triad approach is similar
to the AET approach except that the former combines all bioassay and in situ
biological effects data to provide a single value, while the latter provides
criteria for benthic infauna and each bioassay conducted. However, there
has been little work since Chapman (1986) on development of the Triad
approach for the production of numerical sediment quality criteria.
3.0 USEFULNESS
3.1 Environmental Applicability
Although the Triad approach is both labor-intensive and expensive, its
strengths render it extremely cost-effective for the level of information
provided. First, it provides empirical evidence of sediment quality (i.e.,
based on observation, not theory). Second, it allows ecological interpre-
tation 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 indices.
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Sediment Quality Triad
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 biological ' effects occur?).
Standards in the form of sediment quality criteria (Chapman 1986, in press-a;
PTI 1988a,b) can be set from the contaminant concentrations that are always
associated with effects. 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 unpub-
lished). 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 to develop RTR values. This
comparison helps investigators determine whether dredging will affect useful
habitat or result in material that is unacceptable for ocean disposal.
Similarly, 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 due to a natural event (e.g., storm, oxygen depletion) or physical
anthropogenic impact (e.g., recent dredging, boat scour), the other two
Triad components (i.e., sediment chemistry and toxicity) provide information
where conventional univariate approaches would prove deficient. Such cases
enforce the need to use knowledge of an area in making any type of environ-
mental 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
de1:~eate areas that are excessively contaminated with toxic chemicals
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Sediment Quality Triad
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 et al. 1986, 1937a) and in
freshwater environments in Europe (e.g., Wiederholm 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, and TOC) that
may be important in interpreting the Triad compounds. For example, 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 because of lack of oxygen, making it difficult to
distinguish that mortality from toxicity due to high concentrations of
contaminants.
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
different chemicals. For this reason, a battery of bioassay tests is
recommended. Caution must be used when testing sediment extracts that may
be specific to certain chemical classes. Interpretation of the results must
be restricted to only those chemicals.
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 that are used in the bioassay tests. In other words, all biological
effects data are based on a single species. Ln such cases, independence of
the infaunal community analyses and bioassay test results cannot be assumed.
'9-18
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Sediment Quality Triad
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.
3.1.4 Suitability for In-Place Pollutant Control--
The Triad approach provides a comprehensive approach to in-place
pollutant control as it allows for assessment of all potential interactions
between chemical mixtures and the environment. The comprehensiveness of
this method results from the fact that it includes measurements of multiple
chemicals as well as potential toxic effects of both measured and unmeasured
chemicals.
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 effluents and other discharges.
3.1.5 Suitability for Disposal Applications--
The Triad approach has been used for disposal applications, including
Navy Homeporting work in San Francisco Bay. In that study, it 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.
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Sediment Quality Triad
General Advantaggs
There are several major advantages to the Triad approach:
• The combination of the three separate components .of Triad
provides a preponderance-of-evidence approach
• This approach does not require a priori assumptions concerning
the specific mechanisms of interaction between organisms and
toxic contaminants
• This method can be used to develop sediment quality values
(including criteria) for any measured contaminant or a
combination of contaminants, including both acute and chronic
effects
• It provides empirical evidence of sediment quality
• It can be used for any sediment type
• It allows ecological interpretation of both physical -chemical
and biological properties
• Follow-up is usually not necessary when a complete Triad
study is conducted.
There are also several major limitations to the Triad approach:
• Statistical criteria have not been developed for use with the
Triad approach
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Sediment Quality Triad
• Rigorous criteria for calculating single indices from each of
the sediment chemistry, bioassay, and in situ biological
effects data sets have not been developed
, A large database is required
• If this method is used to determine single-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 interpretation is labor-
intensive and expensive
• The choice of a reference site is often made without
.adequate information regarding how degraded that site may be.
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.
3.2.2 Relative Cost--
Relative cost can be evaluated in terms of either dollars or environ-
mental damage. The Triad approach will not prevent environmental damage.
but it can be used to identify contaminated areas for future remediation.
In terms of dollars, the Triad approach requires substantial resources to :e
9-21
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Sediment Quality Triad
implemented 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 is not
expensive.
3.2.3 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.
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; Wiederholm et
al. in press). In addition, Canada will be conducting Triad studies in
Vancouver and Halifax to determine the suitability of this approach for
implementation of the new Canadian Environmental Protection Act.
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 reasonably well equipped.
The major requirements are adequate QA/QC procedures for chemical measure-
ments; appropriate detection limits; and, for biological analyses, taxonomic
experts and a taxonomic reference library or museum.
3.2.6 Level of Effort Required to Generate Results--
Different levels of effort will generate different levels of results.
For instance, results can be generated by simply measuring one or two
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Sediment Quality Triad
chemicals, determining the number of infauna present, and conducting a single
sediment toxidty bioassay. However, the applicability of these results may
be severely limited. Consequently, multiple chemicals including inorganic
and organic compounds should be measured, and multiple measures of in situ
biological alteration and sediment toxicity should be made. Although it is
possible to use previously collected nonsynoptic data to derive results in a
"paper" study (e.g., Long and Chapman 1985), fieldwork and synoptic sampling
generate the most useful results.
3.2.7 Degree to Which Results Lend Themselves to Interpretation--
Beyond the general conclusions noted in Table 9-2, expert judgment is
required to implement and interpret the Triad approach. In particular, the
definition of "minimal'1 and "severe" biological effects is required to
establish chemical-specific criteria. The Triad approach reflects the
complexity of the issues that must be addressed to assess environmental
quality.
3.2.8 Degree of Environmental Applicability--
The Triad approach has an extremely high degree of environmental
applicability, as detailed above in Section 3.1.
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.
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Sediment Quality Triad
4.0 STATUS
4.1 Extent of Use
Development of the formalized Triad concept has occurred relatively
recently (Long and Chapman 1985; Chapman 1986; Chapman et al. 1986, 1987a,
1988, in press-b). The Triad approach has been used directly to establish
sediment quality criteria (Chapman 1986) and, through data manipulations,
to determine AET values for sediment quality criteria (Tetra Tech 1986a; PTI
1988a,b).
Triad has been used to identify spatial and temporal trends of
pollution-induced degradation. Indices developed using the Triad approach
can be numeric (i.e., numeric sediment quality criteria) 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.
4.2 Extent to Which Approach Has Been Field-Validated
The Triad approach includes field measurements of in situ biological
alteration. As such, it can be considered that field validation is an
integral part of each and every complete Triad investigation.
4.3 Reasons for Limited Use
As previously described, the Triad approach is being used in the U.S.,
Canada, and Europe for marine, estuarine, and freshwater areas. It is not
being used in small projects due to cost and expertise required for full
implementation.
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Sediment Quality Triad
4.4 Outlook for Future USP and Amount *f Development Y»t Maa^Q^
The following areas of the Triad approach require development:
• Determining the appropriateness of different endpoints of
different bioassays, selected chemical contaminants, selected
measures of benthic community structure, and other potential
measures of in situ biological alteration
• Determining the appropriateness of an additive treatment of
the data (e.g., summing bioassay responses to provide a single
index for toxicity).
• Development of statistical criteria.
• Development of rigorous criteria for determining single
indices for each of the three Triad components.
m~ Methods standardization for sediment toxicity bioassays.
However, even without development of the above, 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 alone,
it is expected that the Triad approach will be much more widely used in
future.
5.0 REFERENCES
Chapman, P.M. 1986. Sediment quality criteria from the Sediment Quality
Triad • an example. Environ. Toxicol. Chem. 5: 957-964.
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Sediment Quality Triad
Chapman, P.M. 1988. Marine sediment toxicity tests. pp. 391-402. In:
Chemical and Biological Characterization of Sludges, Sediments, Dredge
Spoils, and Drilling Muds. J.J. lichtenberg, F.A. Winter, C.I. Weber, and
I. Fradkin (eds). ASTM STP 976. American Society for Testing and Materials,
Philadelphia, PA.
Chapman, P.M. (In press-a). A critical review of current approaches to
developing sediment quality criteria. Environ. Toxicol. Chem. 8.
Chapman, P.M. (In press-b). The Sediment Quality Triad approach to
determining pollution-induced degradation. Sci. Total Environ.
Chapman, P.M., R.N. Dexter, S.F. Cross, and O.G. Mitchell. 1986. A field
trial of the Sediment Quality Triad in San Francisco Bay. NCAA Tech. Memo.
NOS OMA 25. National Oceanic and Atmospheric Administration. 127 pp.
Chapman, P.M., R.N. Dexter, and E.R. Long. 1987a. Synoptic measures of
sediment contamination, 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. Environ. Toxicol. Chem. 6:723-725.
Chapman, P.M., R.N. Dexter, H.A. Andersen, and 8.A. Power. 1988. Testing of
field collected sediments and evaluation of the Sediment Quality Triad
concept. Unpublished report prepared for the American Petroleum Institute.
E.V.S. Consultants, Seattle, Washington.
Chapman, P.M., C.A. McPherson, and K.R. Munkittrick. 1989. An assessment
of the Ocean Dumping Tiered Testing approach using the Sediment Quality
Triad. Unpublished report prepared for Environmental Protection Canada.
E.V.S. Consultants, North Vancouver, BC., Canada.
Chapman, P.M., R.N. Dexter, H.A. Andersen, and 8.A. Power. (In preparation).
Evaluation of effects associated with an oil platform, using the Sediment
Quality Triad. Manuscript for submittal to Mar. Poll. Bull.
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 Gewassensedimente - Allgemeine Problematik und
internationaler stand der Diskussion. Wasser-Abwasser-Forsch 20:54-59.
Keith, L.H., W. Crummett, J. Deegan, Jr., R.A. Libby, J.K. Taylor, and
G. Wentler. 1983. Principles of environmental analysis. Anal. Chem.
55:2210-2213.
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Sediment Quality Triad
Long, E.R., and M.F. Buchman. (In press). An evaluation of candidate
measures of biological effects for the National Status and Trends Program.
NCAA Tech. Memo. NOS OMA. National Oceanic and Atmospheric Administration.
Long, E.R., and P.M. Chapman. 1985. A sediment quality triad: measures of
sediment contamination, toxicity and infaunal community composition in
Puget Sound. Mar. Poll. Bull. 16:405-415.
PTI Environmental Services, Inc. 198Sa. Sediment 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 Program,
EPA Contract No. 68-02-43441. PTI Environmental Services, Inc., Bellevue.
WA.
PTI Environmental Services, Inc. 1938b. 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. OeBen, K.A. 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. 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.
Tetra Tech. 1986b. Recommended protocols for measuring selected environ-
mental variables in Puget Sound. Prepared for the Puget Sound Estuary
Program, U.S. Environmental Protection Agency, Region X, Seattle, Washington.
Tstra Teen, Inc.* S?ll§vui; WA.
Wiederholm, T., A.-M. Wiederholm, and G. Mi.lbrink. (In press). Field
validation of T. tubifex bioassays with lake sediments. Water Air Soil
Pollut.
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AET
10.0 APPARENT EFFECTS THRESHOLD APPROACH
Catherine Krueger
Office of Puget Sound
U.S. Environmental Protection Agency Region X
1200 Sixth Avenue
Seattle, WA 98101
(206) 442-1287
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 expected. 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.
r.O ' SPECIFIC APPLICATIONS
1.I Currant Use
At present, the AET approach is being used by several programs to
develop guidelines for the protection of aquatic life in Puget Sound. These
guidelines 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 biological effects. An underlying strategy
10-1
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AET
in many of these programs was to develop two sets of sediment 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 concen-
trations above which multiple biological effects are expected.
The programs incorporate direct biological testing in concentration ranges
between these two extremes 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.
1.1.1 Commencement Bay Nearshore/Tideflats Super'fund Investigation--
Commencement Bay is a heavily industrialized 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 contamination, 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 and define volumes of contaminated sediment for remediation.
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AET
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.
1.1.2 Puget Sound Dredged Disposal Analysis Program--
In 1985, the Puget Sound Dredged Disposal Analysis (PSDOA) program was
initiated to develop environmentally safe and publicly acceptable options
for unconfined, open-water disposal of dredged material. PSDOA 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 Department of Natural
Resources (WONR). AET values were used to develop chemical-specif ic
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
concentration or SIC) biological testing is required to determine the
suitability of dredged material for unconfined, open-water disposal. Based
primarily on AET values for multiple biological indicators, i 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).
1.1.3 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
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AET
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 (Authority) and other state agencies and local govern-
ments. Major funding and overall guidance for the program is provided by
U.S. EPA Office of Marine and Estuarine Protection.
In the PSEP urban bay program, AET values are used in conjunction with
site-specific biological tests during the assessment of sediment contamina-
tion to define and rank problem areas. Source control actions are well
underway, but sediment remediation has not yet begun at any of the sites
(PTI 1988).
1.2 Potential Use
The AET approach to determining sediment quality can also be used as
follows:
• To determine the spatial extent and relative 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
• With appropriate safety factors or other modifications, to
screen sediments in regulatory programs that involve
extensive biological testing.
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AET
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 sediment 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 1.1 and involves active participation by Ecology, U.S.
EPA. the Authority, WONR, the Corps (Seattle District), and various public
interest groups. The draft regulation currently under development affects
only sediments in Puget Sound. As additional data become available from
other locations, the adopted regulation will eventually be broadened and
modified to include the entire state.
2.0 DESCRIPTION
2.1 Description of Method
AET values are derived using a straightforward algorithm that relates
biological and chemical 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
(P<0.05) relative to appropriate reference conditions. The calculation of
AET for each chemical and biological indicator is conducted as follows:
1. Collect "matched" chemical and biological effects data--
Conduct chemical and biological effects testing on subsamples
of the same field sample. (To avoid unaccountable losses of
benthic organisms, benthic infaunal and chemical analyses are
conducted on separate samples collected concurrently at the
same location.)
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AET
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 concentrations of any
toxic chemicals, an absence of other adverse effects, and
physical characteristics that are directly comparable 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
indicator as the highest detected concentration among
sediment samples that do not exhibit statistically significant
effects. (If the chemical is undetected in all nonimpacted
samples, then no AET can be established for that chemical
and biological indicator.)
4. Check for preliminary AET--Verify thatstatistically
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 1-4 for each biological indicator.
The AET approach for a group of field-collected sediment samples is
shown in Figure 10-1. The samples were collected at various locations and
were analyzed for 1) toxicity in a laboratory bioassay and 2) the concen-
trations 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:
iO-6
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Lead
SP-14
IMPACTED
660 ppm
RS-18
*
N ON IMPACTED
AET
10 100 1000 10000
INCREASING CONCENTRATION —
OH- c -CHS 4-Methylphenol
100000
IMPACTED
3600 QQb SP-U
I
£ IMIII.I '.'• iBwn-n •••lit; rrr. IIH - ~~ Q
NONIMPACTED
AET
10 100 1000 10000 100000 100000
INCREASING CONCENTRATION »
Figure 10-1. The AET approach applied to sediments tested for lead and
4-methylphenol concentrations-and toxicity response during
bioassays.
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AET
• Sediments that did not exhibit statistically significant
(P>0.05) toxicity relative to reference conditions ("non-
impacted" stations)
• Sediments that exhibited statistically significant (P<0.05)
toxicity in bioassays relative to reference conditions
("impacted" stations).
Over the observed range of concentrations for these sediment samples
(horizontal axis in Figure 10-1), the sediments fall into two groups for each
chemical :
• At low to moderate concentrations, significant sediment
toxicity occurred in some samples, but not in others
• At concentrations above an apparent threshold value,
significant sediment toxicity occurred in all samples.
The AET value is defined for each chemical as the highest concentration
of that chemical in the sediments that did not exhibit sediment toxicity.
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
biological effect.
For each chemical, additional AET values could be defined for other
biological indicators that were tested (e.g., other bioassay responses or
depressions in the abundances of certain indigenous benthic infauna).
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AET
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 .interrelationships in the environment
without relying upon * 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 assumptions 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)
• When biological impacts are observed at concentrations below
an AET value for a given chemical, it is assumed that the
imparts m.av be related to another chemical, chemical
interactive effects, or other environmental factors (e.g.,
sediment anoxia)
• The AET concept is consistent with a relationship between
increasing concentrations of toxic chemicals and increasing
biological effects (as observed in laboratory exposure
studies).
The assumptions in interpreting environmental data are demonstrated
below with actual field data. Using Figure 10-1 as an example, sediment from
Station SP-14 exhibited severe toxicity, potentially related to a great:;/
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AET
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 10-1). 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 associated 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-methylphenol 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 concentrations at other sites without
associated biological effects). Based on the results for lead and 4-methyl-
phenol, bioassay toxicity at five of the impacted sites shown in the figure
may be associated with elevated concentrations of 4-methylphenol, 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 protective against biological effects,
only that single chemicals may not account for all stations with biological
effects. By developing AETs for multiple chemicals, a high percentage of all
stations with biological effects are accounted for with the AET approach
(see Section 3.2.9 and U.S. EPA 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
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AET
gaps between stations and/or do not cover a wide range of concentrations
must be scrutinized carefully (e.g., to discern whether chemical concentra-
tions in the data set exceed reference concentrations) to determine whether
AET generation is appropriate.
2.1.2. Level of Effort--
%.\.2.l Type of Sampling Required—Collection of field data for initial
generation of AET 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 collected, the diversity of
biological indicators that are tested, and the range of chemicals measured).
One means of minimizing these costs is to compile existing data that meet
appropriate quality assurance criteria. There are no definitive requirements
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 recommendations for data collection:
• Collect or compile chemical and biological 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
(preferably over at least 1-2 orders of magnitude).
• 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.
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AET
• Ensure that detection limits of <1QQ 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 moderate concentrations; as noted
previously, only detected data are used in AET calculations.
Metals are naturally occurring substances and most metals
concentrations 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
described at the beginning of Section 2.1). Matched data sets should be
used to reduce the possibility that uneven (spatially variable) sediment
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.
2.1.2.2 Methods—Methodological details for the generation of AET
values are described at the beginning of Section 2.1.
2.1.2.3 Types of Data Required—There are two fundamental kinds of data
analysis required for AET generation:
• Statistical analysis of the significance of biological effects
relative to reference conditions (i.e., classification of
stations as impacted or nonimpacted for each biological
indicator)
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AET
• Generation of an AET value for each chemical and biological
indicator (essentially 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 chemical data, and
evaluation of the appropriateness of reference area stations. These topics
have been described elsewhere (e.g., Seller et al. 1986; Barrick et al.
1988).
The AET method does not intrinsically require a specific method of
statistical analysis for determination of significance of biological effects
relative to reference conditions. Existing Puget Sound AET have relied
largely on pair-wise t-tests; details of statistical analyses performed for
the generation of Puget Sound AET have been described elsewhere (U.S. EPA
1988; Barrick et al. 1988; Seller et al. 1986). For example, the following
steps were used to determine the statistical 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 wsrs calculated
• Results from each potentially impacted site were then compared
statistically with the reference conditions using pair-wise
analysis
• Tne Fmax test (Sokal and Rohlf 1969) was used to test for
homogeneity of variances between each pair of mean values
• If variances were homogenous, then a t-test was used to
compare the two means
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AET
• 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 pair-wise error
rate of 0.05 to ensure consistency among studies of differing
sample sizes.
Data analyses that have been applied to other biological indicators are
described elsewhere (Seller et al. 1986; Barrick et al. 1988). Notably,
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 collected 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 impacted 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 o*
benthic 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 approach, but rather
are options to address potentially unusual matrices or biological conditions.
2.1.2.4 Necessary Hardware and Skills — The primary skills required for
AET generation are related to the development of the biological/chemical
database. Expertise in environmental chemistry is required to evaluate
chemical data quality, and the need for normalization of chemical data and
related factors. Biological and statistical expertise are required for the
determination of statistical significance. For benthic data in particular.
evaluation of appropriate reference conditions and knowledge of benthic
taxonomy and ecology are necessary.
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AET
Computers are recommended for the efficient generation of AET values.
A menu-driven database (SEDQUAL) has been developed for U.S. EPA Region X
that is capable of a number of data manipulation tasks, including the
following: 1) storing chemical and biological data, 2) calculating 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
users 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) and is available at no cost from U.S. EPA Region X.
2.1.3 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 (U.S.
EPA 1988)
• Policy implications of effects-based marine sediment criteria
(PTI 1987).
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AET
^.2 Applicability of Mo»Hnrt 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 association between chemical
contamination and adverse biological effects, predictions can be made
regarding the levels of contamination that are always associated with
adverse effects (i.e., the AET values). These critical levels of contamina-
tion 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 AET
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 AET developed on the basis of
acute biological responses.
2.3 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 AET, 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
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AET
detection frequencies (e.g., greater than 10 detections). A partial list of
chemicals for which AET have been developed is presented in Table 10-1.
3.0 USEFULNESS
3,1 Environmental Aoolicability
3.1.1 Suitability for Different Sediment Types--
The AET approach can be applied to any sediment type in saltwater or
freshwater environments for which biological tests can be conducted. By
normalizing chemical concentrations to appropriate sediment variables (e.g.,
percent organic carbon), differences between different sediment types can be
minimized in the generation of AET. In practice, identification of unique or
atypical sediment matrices is important in determining the general appli-
cability of AET values generated from a specific set of data.
Differences in physical characteristics (e.g., grain size, habitat
exposure) is one major factor that may account for stations that do not
meet predictions based on existing AET values. In Puget Sound studies, for
example, fine-grained sediments dominated stations that had significant
amohipod mortality but wars not predicted to be so, and coarse-grained
sediments dominated stations that had significant depressions in benthic
infauna but were not predicted as impacted by benthic AET (Barrick et al.
1988).
3.1.2 Suitability for Different Chemicals or Classes of Chemicals--
There are no constraints on the types of chemicals for which AET can be
developed. An AET can be developed for any measured chemical (organic or
inorganic) that spans a wide concentration range in the data set used to
generate A£Ts. The availability of a wide diversity of chemical data
increases the probability that toxic agents (or chemicals that covary in the
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TABLE 10-1. SELECTED CHEMICALS FOR WHICH AET
HAVE BEEN DEVELOPED IN PUGET SOUND
Metals
Antimony
Arsenic
Cadmium
Chromium
Copper
Lead
Mercury
Nickel
Silver
Zinc
Organic Compounds
Low molecular weight PAH
Naphthalene
Acenaphthylene
Acenaphthene
Fluorene
Phenanthrene
Anthracene
2-Methylnaphthalene
Chlorinated benzenes
1,3-Oichlorobenzene
1,4-Oichlorobenzene
l.,2-0ichlorobenzene
1,2,4-Trichlorobenzene
Hexachlorobenzene (HC8)
Total PCBs
Pesticides
p.p'-OOE
p,p*-OOD
p,p'-OOT
Miscellaneous Extractables
Benzyl alcohol
Benzoic acid
Oibenzofuran
Hexachlorobutadiene
N-Ni trosodiphenylamine
High molecular weight PAH
Fluoranthene
Pyrene
Benz(a)anthracene
Chrysene
Benzofluoranthenes
Benzo(a)pyrene
Indeno(1,2,3-c,d)pyrene
Oibenzo(a,h)anthracene
Benzo(g,h,i)perylene
Phthalates
Dimethyl phthalate
Diethyl phthalate
Oi-n-butyl phthalate
Butyl benzyl phthalate
Bis(2ethylhexyl)phthalate
Oi-n-octyl phthalate
Phenols
Phenol
2-Methylphenol
4-Methylphenol
2,4-Oimethylphenol
Pentachlorophenol
Volatile Organics
Tetrachloroethene
Ethylbenzene
Total xylenes
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AET
environment with toxic agents) can be included in interpreting observed
biological impacts.
To date, AET have been developed for over 60 chemicals frequently
detected in the environment, including 16 polycyclic aromatic hydrocarbons
(PAH); several alkylated PAH and related nitrogen-, sulfur-, and oxygen-
containing heterocycles; 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; and 10
metals and metalloids; 3 volatile organic compounds; and 5 miscellaneous
extractable 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-um particulate material,
in dry weight of sediment). These latter AET values did not appear to offer
advantages in predictive reliability over the more commonly used dry weight
and TOC normalizations (Seller et al. 1986).
3.1.3 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 relies on empirical
information that measures the chemical concentrations associated with
samples exhibiting adverse effects, the results are directly applicable to
predicting effects on the organisms used to generate the A£T. The results
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AET
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 sensitive 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 macroihvertebrates.
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
condi tions
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AST
• Provide an environmental basis for triggering sediment
remedial action
• 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.
3.1.5 Suitability for Source Control--
The AET approach is well suited for identifying 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 environmental samples). The link between
problem areas and potential sources of contamination is established 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 approach 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 too 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 AET is
predicted to result in mitigation of the adverse biological effects. It is
possible that such source controls may be effective because of the con-
comitant removal of an unmeasured contaminant.
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AET
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 preponderance-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 or will not occur in
dredged material after disposal at aquatic sites.
3.2 General Advantages and Limitations
3.2.1 Ease of Use--
In this section, "use" is treated as both generation 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 expenditure of effort,
as would the statistical analysis of a large number of samples. However, if
data are available and statistical analyses have been performed, the
generation of AET values is very easy with the SEDQUAL database (described
in Section 2.1.2.4). The menu-driven . system allows for a considerable
amount of flexibility in choosing stations and biological indicators to be
included in AET generation. Aoolication of AET (i.e., comparison of A£T
values to chemical concentrations in field samples) is also very easy *hen
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AET
using SEDQUAL, provided that the field data have been computerized.
Application of AET values to chemical data presented in existing literature
is also straightforward.
3.2.2 Relative Cost--
The cost of developing AET values can span a wide range, depending on
the stage of database development and the numbers and kinds of chemicals 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 if existing AET values are applicable would still be
useful.) The historical database could be based on the pooled results from
various studies conducted in a region, providing 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 AET 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, inexpensive
biological test. It is recommended that the approach be based on a
relatively wide range of chemicals, and if possible, several kinds of
biological 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
AET varied considerably among the four indicators. For example, laboratory
costs for the least expensive indicator (i.e., Microtox bioassay) were
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AET
approximately $200 per station, whereas costs for the most expensive
indicator (i.e., abundances of benthic macroinvertebrates) were as high as
SI,800 per station. Therefore, 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.
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
underprotective of the environment. The occurrence of biologically impacted
stations at concentrations below the AET of a given chemical (see Figure 10-
1) may appear to be underprotective. However, the occurrence of impacted
stations at concentrations below the AET of a single chemical does not imply
that AETs in general are not protective against biological effects, only that
single chemicals may not account for all stations with biological 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 bioassays of
mortality in adult amphipods, developmental abnormality in larval bivalves,
and metabolic alterations in bacteria. All of these organism/endpoint
combinations are considered to be sensitive to chemical toxicity. AETs have
also been generated for in situ reductions in the abundances of benthic
macroinvertebrates. Because these reductions incorporate chronic (i.e.,
long-term) exposure to contaminants, they can also be considered as sensitive
measures of the effects of chemical toxicity. However, a more protective
acproach would be to use the lowest of the four kinds of AET for each
chemical as the concentration upon which predictions are made. A-itar-
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AET
natively, the protect!veness of any kind of AET could be modified by
developing sediment quality values based on "safety factors" applied to
existing AETs.
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 contaminated
sediments in urban bays. PSOOA has used AET values as a tool to develop
chemical guidelines for determining whether biological testing is necessary
for dredged sediments proposed for unconfined, open-water disposal. Ecology
has used AET to guide several stages of remedial action and to draft
sediment standards for classifying sediments according to their potential
for causing adverse biological effects. In several of these applications,
AET have been modified by "safety factors" to enhance their protectiveness.
Several major characteristics influence the acceptability of the AET
approach. The most attractive characteristic of the approach is probably the
reliance on empirical information based on field-collected sediments or
indigenous organisms, and exposure of laboratory test organisms to environ-
mental 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 observed. This characteristic provides
consistency that, with a representative database used to generate 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).
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AET
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 chem'ical data to address the
bioavailability of contaminants to organisms. Organic carbon-normalization
may be most appropriate for nonpolar organic contaminants based on the-
oretical considerations. In addition, nonprotective 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
guidelines for chemical and biological data are discussed by Barrick et al.
(1988). The AET approach is currently under review by the U.S. EPA Science
Advisory Board.
3.2.5 Ability to be Implemented by Laboratories with Typical Equipment and
Handling Faci1ities--
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 necessary 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.
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AET
3.2.6 Level of Effort Required to Generate Results--
As noted in Section 3.2.1, the SEDQUAL database facilitates AET
generation and application. After field data have been collected, the most
time-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
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 requires
that the numerical criteria be entered in the database.
3.2.7 Degree to Which Results Lend Themselves to Interpretation--
The manner in which the AET approach can be used to interpret matched
biological and chemical data from field-collected sediments is described in
Section 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 relationships 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. Although the AET concept is simple, the generation of AET .alues
based on environmental data incorporates many .complex biological-chemical
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AET
interrelationships. For example, the AET approach incorporates the net
effects of the following factors that may be important in field-collected
sediments:
• Interactive effects of chemicals (e.g., synergism, antagonism,
and additivity)
• Unmeasured chemicals and other unmeasured, potentially adverse
variables
• Matrix effects and bioavailabil ity (i.e., phase associations
between contaminants and sediments that affect bioavailabil ity
of the contaminants, such as the incorporation of PAH in soot
particles).
The AET approach cannot quantify the individual contributions of
interactive effects, unmeasured chemicals, or matrix effects in environmental
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 predictions are
made. Alternatively, isolated occurrences of such environmental factors in
a data set used to generate AETs may limit the predictive reliability of
those AET values. If confounding environmental factors render the AET
approach unreliable, then this should be evident from validation tests in
which biological effects are predicted in actual environmental samples.
A more detailed discussion of the interpretation of AETs and the
confounding effects of environmental factors is presented in U.S. EPA (1988).
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AET
3.2.8 Degree of Environmental Appl icabil ity-
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 information 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 (U.S.
EPA 1988). The overall reliability of the four tests ranged from 35 to 96
percent, indicating that all four kinds of AET were relatively accurate at
predicting the presence or absence of effects for samples from the existing
database. This high level of reliability suggests that AET have a relatively
high degree of environmental applicability in Puget Sound, and has been a
primary factor for 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 criteria to provide an initial estimate of
the magnitude of potential problem areas in coastal regions of the U.S. for
the U.S. EPA Office of Policy Analysis (PTI 1987).
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:
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AET
• Sensitivity in detecting environmental problems (i.e., are all
biologically impacted sediments identified by the predictions
of the chemical sediment criteria?)
• Efficiency in screening environmental problems (i.e., are only
biologically impacted sediments identified by the predictions
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 inefficient. 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 10-2.
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:
All stations correctly predicted as impacted
Overall reliability - All stations correctly predicted as nonimpacted
Total number of stations evaluated
High reliability results from correct prediction of a large percentage of the
impacted stations (i.e., high sensitivity, few false negatives) and correct
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Q
IMPACTED
/ I
( )
( )
\
PI1EOICIEO
o
t
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 UlOlOGlCAl INOICAIOH
A All ItlAIIONS PREDICTED TO BE IMPACIEO
B Alt aiAUONS KNOWN IO ME IUPACTEO
C ALL &IAIIONS CORHECILY fHCDlCUU IO UE IMPACIEO
Figure 10-2. Measures of reliability (sensitivity and efficiency).
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AET
prediction of a large percentage of the nonimpacted stations (i.e., high
efficiency, few false positives).
An assessment of AET reliability was recently conducted using a large
database comprising 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 AET generated from a 188-sample data set were evaluated with
an independent 146-sample 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 amphipod 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 conditions) and nonparametric methods (e.g., ranking of stations
by concentration), and thus are not amenable to the routine definition of
confidence intervals. 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
• Classification error associated with the statistical
significance of biological indicator results (i.e., whether a
station is properly classified as impacted or nonimpacted, as
related to Type I and Type II statistical error)
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• The weight of evidence or number of observations supporting a
given AET value
• The analytical error associated with quantification of
chemical results.
Detailed discussion of these factors is provided in Seller 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 concentration 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
stations used to establish an AET would be expected to have a marked effect
on AET uncertainty because small data sets would tend to have less continuous
distributions "of-chemical concentrations than large data sets. Based on
analyses conducted with Puget Sound data, the magnitude of the AET uncer-
tainty for 10 chemicals or chemical groups that are commonly detected is
typically less than one=third to one-half of the value of the AET itself
(considering both amphipod 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.
4.0 STATUS
4.1 Extent of Use
The AET approach has been used by several agencies in the Puget Sound
region to provide guidelines for regulatory decisions. The U.S. EPA has
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AET
used AET to develop sediment quality values with which to evaluate the
potential toxicity of contaminated sediments in urban bays. PSOOA has used
the AET approach as a tool for developing guidelines to determine whether
biological testing is necessary for dredged sediments proposed for uncon-
fined, open-water disposal. Ecology has used AET values to establish draft
sediment standards for classifying sediments according to their potential
to cause adverse biological effects. Ecology and U.S. EPA have also used
AET values to identify problem chemicals, link contaminated sediments to
potential sources, and provide reference points for the establishment of
sediment cleanup goals in the Commencement Bay RI/FS.
Several strategies have been developed for using the AET approach for
different regulatory purposes in Puget Sound. In the Superfund program
locally, the lowest AET (termed LAST) for the four kinds of AETs used in
Puget Sound have been used to establish goals for sediment remedial actions.
In dredged material assessment, sediment quality values have been developed
for use as protective screening chemical levels by applying "safety factors"
to the AET. Because biological effects are rarely expected to occur when
chemical concentrations are below these screening levels, additional testing
of sediments usually is not required. The AET approach also been used to
develop maximum chemical levels, above which adverse effects are predicted
for all of the biological tests used to generate AETs in Puget Sound. These
maximum levels have been set by the highest AET (termed HAET) for the four
biological indicators evaluated in Puget Sound.
Outside the Puget Sound region, chemical and biological data from San
Francisco Bay, San Diego Bay, and the Southern California Bight are currently
being evaluated for use in developing region-specific AETs for the California
State Water Resources Control Board. These California AETs will be compared
with Puget Sound AET to evaluate similarities and differences between the two
kinds of information.
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AET
A.? Extent to Which Aoornarh u.s Rppn Field-Validated
As described in U.S. EPA (1988), the reliability of AETs generated from
Puget Sound data was evaluated with tests of sensitivity and efficiency
(defined in Section 3.2.9). Tests of the sensitivity and efficiency of the
AET approach were carried out in several steps, as described below:
• 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 chemistry stations with associated benthic infaunal data
were grouped together (201 stations), all chemistry stations
with associated oyster larvae bioassay data were grouped
together (56 stations), and all chemistry stations with
associated Microtox 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»,,v, and t-tests at alpha • 0.05).
• • nia« r I
m 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 sen-
sitivity and efficiency tests were performed against the
same database for each biological indicator.
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AET
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., ionizable organic
compounds, metals, and metalloids), and (b) using TOC-
normalized data for all chemicals. Test 2 allowed for a
posteriori evaluation of the relative success of dry
weight and TOC normalization 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 stations. 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
completely 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
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of Puget Sound AET values using matched biological/chemical data from other
geographic areas is desirable before recommending direct application of the
Puget Sound values in other geographic regions.
4.3 Reasons for Limited Use
The AET approach was developed in the Puget Sound region, and has been
used to provide agencies with guidelines for evaluating and managing
contaminated sediments. The approach is not yet commonly used outside of
Puget Sound. Because the approach is based on empirical data, region-
specific values should be evaluated thoroughly by experts before application
in other regions. For example, because regional reference areas are used to
determine the significance of adverse biological effects in the approach,
there may be concern that AET developed for one region may be overprotective
or underprotective of other areas.
Development of site-specific AET for other geographic areas may require
additional sampling. Because many past studies were not multidisciplinary,
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 inte-
gration or comparison of AET data sets among different regions can also be
restricted because appropriate biological indicators for generating AETs may
vary among regions.
4.4 Outlook for Future Use and Amount of Development Yet Needed
The following two approaches to AET development could be particularly
beneficial in expanding the use of this approach:
• Use of laboratory cause-effect (spiking) studies to eva-luate
AET predictions on a chemical-specifie basis
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AET
• 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 method is currently under technical review by the U.S. EPA Science
Advisory Board. Based on this review, additional development of the method
may be recommended.
5.0 REFERENCES
Barrick, R.C., S. Becker, I. Brown, H. Seller, 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.
Environmental- Protection Agency Region 10, Office of Puget Sound. PTI
Environmental Services, Bellevue, WA. 74 pp. + appendices.
Becker, O.S., R.P. Pastorok, R.C. Barrick, P.N. Booth, and I.A. Jacobs.
1989. Contaminated sediments criteria report. Prepared for the Washington
Department of Ecology, Sediment Management Unit. PTI Environmental
Services, Bellevue, WA. 99 pp. * appendices.
Bellan-Santini, 0. 1980. Relationship between populations of amphipods and
pollution. Mar. Poll. Bull. 11:224-227.
Seller, H.R., R.C. Barrick, and O.S. Becker. 1986. Development of sediment
quality values for Puget Sound. Prepared for Resource Planning 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, 0. 1988. SEDQUAL users manual. Prepared for Tetra Tech, Inc. and
U.S. Environmental Protection Agency Region 10, Office of Puget Sound. PTI
Environmental Services, Bellevue, WA.
Pearson, T.H., and R. Rosenberg. 1978. Macrobenthic succession in relation
to organic enrichment and pollution of the marine environment. Oceanogr.
Mar. Biol. Annu. Rev. 16:229-311.
Phillips, <.. 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, '«A.
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AET
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 American Management Systems and U.S. Environmental Protection
Agency, Office of Policy Analysis. PTI Environmental Services, Bellevue, 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.
U.S. Environmental Protection Agency. 1988. Briefing report to the EPA
Science Advisory Board. Prepared for Battelle and U.S. Environmental
Protection Agency Region 10, Of- ce of Puget Sound. PTI Environmental
Services, Bellevue, WA. 57 pp.
Sokal, R.R., and F.J. Rohlf. 1969. Biometry. W.H. Freeman and Company, San
Francisco, CA. 859 pp.
Swartz, R.C., W.A. OeBen, J.K. Phillips, J.O. Lamberson, and F.A. Cole.
1985. Phoxocephalid amphipod bioassay for marine sediment toxicity. pp.
284-307. In: Aquatic Toxicology and Hazard Assessment: Proceedings :f the
Seventh Annual Symposium. R.O. 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 environmen-
tal variables in Puget Sound. Final Report. Prepared for the U.S. EPA,
Region X, Office of Puget Sound, Seattle, WA. Tetra Tech, Inc., Bellevue,
WA.
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IJC
CHAPTER 11. A SUMMARY OF THE SEDIMENT ASSESSMENT STRATEGY
RECOMMENDED BY THE INTERNATIONAL JOINT COMMISSION
Philippe Ross
Illinois Natural History Survey
607 East Peabody Drive
Champaign, IL 61820-6970
(217) 244-5054 or (312) 353-0117
The International Joint Commission (IJC) Sediment Subcommittee has
published a document entitled Procedures for the Assessment of Contaminated
Sediment Problems in the Great Lakes (IJC.1988a). An overview of the IJC
(1938a) 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 (19S8a) document itself for an explanation of details
that are not provided herein.
1.0 SPECIFIC APPLICATIONS
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 the
work plan for sediment investigations at Great Lakes areas of concern (AOCs,
as identified by the IJC). 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 demonstration program called
ARCS (Assessment and Remediation of Contaminated Sediments). The strategy
recommended by IJC (19S8a) will be app' ed through a series of activities
involving physical mapping and characterization, sampling, chenvcai
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analyses, toxicity testing, and in situ community analysis. The assessment
will begin in 1989, with completion scheduled for 1991.
\.% 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.
2.0 DESCRIPTION
2.1 Description of Method
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 assessments of areas (e.g.,
bays, harbors, rivers, other depositional zones) where sediment
contamination and the need for remedial action are suspected. 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 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-making activities, such as sample-by-sample disposal of
dre^-ed material from navigation channels. Nevertheless, some of the
component methods described could be useful and cost-effective in this
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regard. The first major assumption, therefore, 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 focuses on whether sediment contaminants are exerting biological
stress or are being bioaccumulated. 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 Chapter 9). The proposed strategy has the following
objectives:
• To provide accurate assessments of specific problems by using
a modified 'triad' approach, which integrates chemical,
physical, and biological information
• To perform tasks in a sequence so that the results from each
technique can be used to reduce subsequent sampling require-
ments and costs
• To provide adequate proof of linkage between the contamination
and the observed biological impact
• To quantify problem severity, thereby enabling inter-
comparisons between and within areas of investigation (thus
allowing a priority list for remedial actions to be developed
and the objective selection of appropriate remedial options)
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• 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 IJC 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 assessment portion of the document identifies a
set of the best currently available assessment tools (see Section 2.1.2.2),
it is assumed that decisions will be made based on the circumstances unique
to each AOC. There is no substitute for experience (expert judgment), and
.it is also assumed that appropriate expertise will be assembled before the
assessment study plan is formulated.
2.1.2 Level of Effort--
2.1.2.1 Type of Sampling Required — The IJC (1983a) approach involves
two stages. Stage I, the initial assessment, is used for areas where an
inadequate or outdated database exists. 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 identifications),
fish contaminant body burdens (one important species, selected by expert
judgment), and external abnormalities on collected specimens. Any one of
the following criteria provide 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
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IJC
• Concentrations of hazardous persistent organic compounds above
detection levels in fish or benthos
• The absence of a healthy benthic community (e.g., absence of
clean water organisms such as amphipods or mayflies, presence
of a community dominated by oligochaetes, the complete
absence of invertebrates)
• 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 and 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 assessment or 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, but rather logical groupings of work units. The expert investi-
gator should be responsible for the final study design.
In Phase I of Stage II, extensive information on the physical
composition of the sediments is collected. These data are used to define
areas or zones of homogeneity within a study area. Knowledge of these zones
allows sampling requirements for Phase II to be estimated.
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In Phase II, the benthic community structure is examined to the lowest
possible taxonomic 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 data to reduce the sampling effort in the next phase.
In Phase III, 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 II sample set. Since
fresh sediment must be collected for this phase, precision position-finding
equipment is required to relocate previously sampled sites. Phase III costs
can be reduced by performing acute lethality bioassays on a sediment sample
before proceeding to tests that measure chronic or sublethal effects. Also
in Phase III, sediment cores are collected, dated, and sectioned for strati-
fied chemical analyses and bioassays. Finally, adult fish are examined
histopathologically for internal (e.g., liver) tumors. In relatively
confined geographical areas, Phases II and III may be combined, as further
sampling may be more costly than conducting additional bioassays, ana
relocating Phase II sets for Phase III sampling may be difficult. In .this
case. Phase II sampling will include extra material for Phase III.
In the fourth and final phase, sediment dynamics (e.g., accumulation,
resuspension, movement) and factors affecting them are quantified. 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 in
the selection and cost evaluation of remedial options.
Criteria that clearly indicate when some form of remedial action n.ust
be considered (based on the results of Stage II) are essential. Due to the
absence of definitive sediment action criteria at time of writing, the
criteria proposed by the IJC (198Sa) are highly conservative, following ire
language of the 1978 Great Lakes Water Quality Agreement as revised in 1937
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(especially Annexes 1 and 12), in order to promote maximum protection and
effective restoration of the Great Lakes ecosystem. The IJC (1988a) urges
that these criteria be reviewed regularly to ensure that they continue to
fulfill their intended purpose.
g. 1.2.2 HethodS'-During Stage I, the minimum amount of information
necessary to assess potential problem sediments is collected. A variety of
physical, chemical, and biological measurements are recommended, as outlined
below:
• A geographical description of the area and its bathymetry
are required.
• Sediment grain size - Size analysis techniques based on
settling velocity (American Society for Testing and Materials
1964; Duncan and LaHaie 1979) are recommended. The sand
fraction is removed by a 62-um sieve and analyzed separately
from the fine-grained material.
• Sediment water content - The water content can be determined
during sample preparation 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 recommended by IJC (1988a)].
• Organic carbon - It is recommended that total sediment
organic carbon be measured as described by Plumb (1981).
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Phosphorus - TWO measurements are suggested: total phosphorus
as extracted from sediment by sodium carbonate fusion or by
perchloric acid digestion, and bioavailable phosphorus as
estimated by NaOH extractable 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 AOCs. 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) protocols for broad scans and analyses of
individual compounds. When the strategy was written, no
standardized chemical protocols for estimating bioavailabi1ity
of trace organic compounds were identified.
External abnormalities in fish - The presence of one or more
external abnormalities is often indicative of anthropo-
genically 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 recommended that locally occurring catfish
(particularly /. nebu/osus) be examined for tumors, melanoma,
blindness, and barbel abnormalities during a Stage I assess-
ment.
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• 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 carpi a)
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 contaminated 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
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 inexpensive 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 2.1.2.1 above).
The detailed assessment of Stage II consists of more focused analyses
to supplement or complement information obtained in Stage I. Phase I :•*
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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, area! 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 II of the detailed assessment focuses on surficial sediment
chemistry and benthic community structure. Based on the previous mapping of
homogeneous zones (Phase I), effort in Phase II can be expended in deposi-
tional areas and in those areas with fine-grained sediments. Surficial
chemistry sampling should be coincident with the sampling for detailsa
benthic community structure analysis. Total organic carbon, redox
potential, pH, metals, and persistent organics should be measured.
Investigators 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 II community 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 analyses. To help investigators assess
community impact, IJC (198Sa) provides a partial list of literat-ir-
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descriptions of normal nearshore communities in habitats that most closely
approximate Great Lakes AOCs. A detailed discussion of statistical methods
is also included.
Phase III of the detailed assessment consists of obtaining additional
information concerning sediment toxicity (i.e., bioassays and fish histopath-
ology) and stratigraphic characterization of sediment cores. A suite of
bioassays is proposed for toxicological evaluation of sediments:
• Microtox - an acute, liquid-phase (elutriate or pore-water)
test with luminescent bacteria (Bulich 1984)
• Algal photosynthesis - an acute, liquid-phase test using
natural communities [algal fractionation bioassay (Munawar and
Munawar 1987)] or the laboratory species Selenastrum capri-
cornutum (Ross et al. 1988)
• Zooplankton life-cycle tests (Daphnia magna liquid and solid
phases) monitoring growth and reproduction (Nebeker et al.
1984; LeBlanc and Surprenant 1985)
• Chronic, solid-phase tests using the benthic invertebrates
Chironomus tentans (Nebeker et al. 1984), Hyalella azteca
(Nebeker et al. 1984), or Hexagenia limbata (Malueg et al.
1983)
• A solid-phase fish bioaccumulation test with the fathead
minnow Pimephales promelas (Mac et al. 1984)
• The liquid-phase (extract) Ames Sa7/none/?a/microsome assay, a
bacterial mutagenicity test (Tennant et al. 1987).
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In addition to bioassays, histopathological examinations of indigenous
adult fish (especially Ictalurus nebulosus), focusing on preneoplastic and
neoplastic liver lesions (Couch and Harshbarger 1985), are recommended.
Also included in "Phase III work are chemical analyses and dating of
sediment cores. Isotopic (14C, 210Pb, 55Fe, 137Cs) and biostratigraphic
[i.e., ragweed (Ambrosia) pollen] methods are both recommended for dating
sediment cores. This dating is necessary to establish the three-dimensional
configuration of the contaminated sediment mass and to assign a date to the
sediment depositional unit.
In Phase IV of the detailed assessment, studies on sediment dynamics are
necessary to determine the following:
• Potential water column impacts through resuspension
• Movement of contaminated sediment out of the AOC
• The quality and rate of new sediment accumulation
• 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.
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 taxonomic identification); metals concentrations acove
cackground levels in the surficial sediments (ascertained from dating);
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hazardous persistent 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 sediments. The
biological impacts should be assessed 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 requirements to be reduced. When Phases I and II
have revealed homogeneous zones of sediment type and similar community
structure, the number of Phase III samples can be appropriately scaled down.
Impairment due to sediment contamination and the probable need for
remediation are established when the biomonitoring results from the
detailed assessment demonstrate significant departures from controls.
Each section of IJC (1988a) contains a detailed discussion of the
statistical procedures required, with references and examples. The
preferred method of interpretation is left to the expert investigator in
many cases.
2.1.2.4 Necessary Hardware and Skills—The initial assessment, and to
an even greater degree the detailed assessment, require a large array of
field and laboratory equipment. Although none of the items recommended are
unusual or inordinately sophisticated, one laboratory or field unit is
unlikely to have all the required apparatus. Specific suggestions for
hardware and skills are provided by IJC (19S8a). Because this approach is
intended for major sediment assessment efforts, several groups would
probably have to be mobilized to contribute to the effort.
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2.1.3 Adequacy of Documentation--
Each component method described in IJC (1988a) is fully referenced in
the text and accompanied 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.
2.2 Aoolicabilitv 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
IJC (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.
2.3 Ability of Method to Generate Numerical Criteria for Specific Chemicals
The document was designed to provide guidance to assessment programs.
Nevertheless, since chemical, toxicological, and infaunal data are collected
in the Stage II assessment, it is possible that these data could be used to
develop chemical-specifie 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 bi'ota in locations other than r.ne
ACC under consideration.
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3.0 USEFULNESS
7 | Environmental Analirabilitv
3.1.1 Suitability for Different Sediment Types--
The approach recommended in IJC (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 sediment 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.
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 IJC (19S8a-) are effective for most contaminants
found in Great Lakes sediments. The broad-based nature of the approach
contains sufficient flexibility to deal with anomalous situations.
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,
laboratory results (i.e., chemistry, toxicity) which can be compared to
standard conditions and literature values may be placed in the context of
empirically derived effects data from the site under investigation.
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3.1.4 Suitability for In-Place Pollutant Control--
The guidance document was developed specifically 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
upon which to base such decisions. A companion document (IJC 1988b) provides
guidance in the selection of courses of remediation.
3.1.5 Suitability for Source Control--
The detailed assessment provides an adequate framework for identifying
hot spots, and for establishing significant differences from background
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 IJC (1988a) strategy.
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.
3.2 General Advantages and Limitations
3.2.1 Ease of Use--
The proposed strategy is designed to be applicable to the AOC under
investigation. It is intended to flexible, relying on the judgment anc
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experience of those who apply u. A detailed assessment would only be
practical in cases where a major remedial effort is contemplated.
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
compared to the potential remedial costs, the assessment costs are relatively
small. The sequential approach is designed to reduce sampling, analysis,
and expense 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 II assessment. Alterna-
tively, if the Stage I assessment produces no results of concern, then Stage
II need not be undertaken. The cost of a detailed assessment, although
relatively high, is controlled somewhat by the sequential approach to data
collection. No firm cost figures are currently available, but assessments
planned for priority AOCs under Section 118(c)(3) 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.
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 investi-
gator, the degree of conservatism in decision-making will be variable.
3.2.4 Level of Acceptance--
The guidance document (IJC 19S3a) does not describe a new method, but
rather a combination of several types of .methods, each widely accepted in
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its own sphere. The strategy as a whole is being used for the first time in
1989.
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 upon a variety of expertise in a given geographical
area.
3.2.6 Level of Effort Required to Generate Results--
The total level of effort for a detailed assessment will be relatively
high in most cases. This strategy is most suitable for major evaluation
projects.
3.2.7 Degree to which Results Lend Themselves to Interpretation--
The actual statistical analysis and interpretation 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 ultimate decision is left to the investigator.
The inclusion of chemical, toxicological, and infaunal information in the
database allows the investigator to compare different types of indicators
before making decisions.
3.2.8 Degree of Environmental Applicability--
One of the strengths of a strategy that includes in situ community
analysis is that effects data have a high degree of environmental relevance.
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Site-relevant species can even be substituted in the bioassay battery if
necessary, and the body burden and community structure data are always site-
specific.
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. Nevertheless, 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 strategy is difficult to assess, but the methods recommended by
the IJC (1988a) were chosen for their relevance to the Great Lakes ecosystem.
4.0 STATUS
4.1 Extent of Use
IJC's (1988a) document was published in December 1988, and distributed
in early 1989. The strategy is intended for the Great Lakes, and will be
used for the first time in 1989. Most of the individual methods recommended
are widely used and accepted.
4.2 Extent to Which the Approach Has Been Field-Validated
The first extensive field validation of the approach will take place in
1989-1991 as part of the ARCS program under Section HS(c)(3) of the Water
Quality Act of 1987.
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uc
4.3. Reasons for Limited Use
Most component protocols are in wide use. Because the IJC (1988a)
document has only recently appeared, it has not yet been applied.
4.4 Outlook for Future Use and Development
With the backing of both signatories to the Great Lakes Water Quality
Agreement, the document seems destined for widespread use in the Great
Lakes basin. As methods will progress, the document will be updated in each
of its sections.
5.0 REFERENCES
Adams, D.O., D.A. Darby, and R.J. Young. 1980. Selected analytical
techniques for characterizing the metal chemistry and geology of fine-
grained sediments and interstitial water. In: Contaminants and 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. Philadelphia, PA. 535 pp.
Bulich, A.A. 1984. Microtox - a bacterial toxicity test with general
environmental applications, pp. 55-64. In: Toxicity Screening Procedures
Using Bacterial Systems. 0. Lin and B.S. Outka (eds). Marcel Oekker, New
York, NY.
Cairns, J., Jr. 1986. The myth of the most sensitive species. BioScience
36:670-672.
Chapman, P.M., and E.R. Long. 1983. The use of bioassays as part of a
comprehensive approach to marine pollution assessment. Mar. Pollut. Bull.
14:81-84.
Couch, J.A., 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
se.diaientology laboratory, NWRI. Env. Can. NWRI contribution. 23 pp.
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Elliott, J.M. 1977. Some methods for the statistical analysis of samples
of benthic invertebrates. Scientific Publication No. 25. Freshwater
Biological Association. 160 pp.
Green, R.H. 1979. Sampling design and statistical methods for environ-
mental biologists. J. Wiley and Sons, New York, NY. 257 pp.
International Joint Commission. 1988a. Procedures for the assessment of
contaminated sediment problems in the Great Lakes. IJC, Windsor, Ontario,
Canada. 140 pp.
International Joint Commission. 1988b. Options for the remediation of
contaminated sediments in the Great Lakes. IJC, Windsor, Ontario, Canada.
73 pp.
LeBlanc. G.A., and D.J. Surprenant. 1985. A method for assessing the
toxicity of contaminated freshwater sediments, pp. 269-283. In: Aquatic
Toxicology and Hazard Assessment, Seventh Symposium. R.D. Cardwell, R.
Purdy, and R.C. Bahner (eds). ASTM STP 854. American Society for Testing
and Materials, Philadelphia, PA.
Mac, M.J., C.C. Edsall, R.J. Hesselberg, and R.E. Sayers, Jr. 1984. Flow-
through bioassay for measuring bioaccumulation of toxic substances from
sediment. EPA OW-930095-01-0. U.S. Environmental Protection Agency,
Chicago, IL. 26 pp.
Malueg, K.W., G.S. Schuytema, J.H. Gakstatter, and O.F. Krawczyk. 1983.
Effect of Hexagenia on Oaphnia response in sediment toxicity tests. Env.
Toxicol. Chem. 2:73-82.
Munawar, M., and I.F. Munawar. 1987. Phytoplankton bioassays for evaluating
toxicity of in situ sediment contaminants. Hydrobiologia 149:87-105.
Nebeker, A.V., M.A. Cairns, J.H. Gakstatter, K.W. Malueg, and G.S. Schuytema.
1984. Biological methods for determining toxicity of contaminated freshwater
sediments to invertebrates. 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-31-1. U.S. Environ-
mental Protection Agency/U.S. Army Corps of Engineers Technical Committee on
Criteria for Dredged and Fill Material, U.S. Army Waterways Experiment
Station, Vicksburg, MS. 471 pp.
Poulton, D.J. 1987. Trace contaminant status of Hamilton Harbor. J.
Great Lakes Res. 13:193-201.
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Ross, P.E., V. Jarry, and H. Sloterdijk. 1988. A rapid bioassay using the
green alga Selenastrum capricornutum to screen for toxicity in St. Lawrence
River sediments. American Society for Testing and Materials. STP 988:68-
73.
Schmitt, C.J., M.A. Ribick, J.L. Ludke, and T.W. May. 1983. National
pesticide monitoring program: organochlorine residues in freshwater fish,
1976-79. Fish and Wildlife Service Res. Publ. No. 152. U.S. Oept. 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, Hamilton, Ontario. May 17-20,
1988.
Sokal , R.R., and F.J. Rohlf. 1969. Biometry. W.H. Freeman and Co., San
Francisco, CA.
Tennant, R.W., 8.H. Margolin, 0.0. Shelby, £. Zeiger, O.K. Haseman, J.
Spalding, W. Caspary, M. Resnick, S. Stasiewicz, B. Anderson, and R. Minor.
1987. Prediction of chemical carcinogenicity in rodents from ;n situ
genetic toxicity assays. Science 236:933-941.
Terzaghi, <., and R.B. Peck. 1967. Soil mechanics in engineering practice.
John Wiley and Sons, New York, NY. 729 pp.
Tsai. C.-H., and W. Lick. 1986. A portable device for measuring sediment
resuspension. J. Great Lakes Res. 12:314-321.
U.S. Environmental Protection Agency. 1984. Guidelines establishing test
procedures for the analysis of pollutants under the Clean Water Act; final
rule and interim final rule and proposed rule. U.S. EPA, Washington, DC.
Federal Register Vol. 49, No. 209, Part VIII. pp. 1-210.
Williams, J.O.H., H. Shear, and R.L. Thomas. 1980. Availability to
Scenedesmus quadricauda of different forms of phosphorus in sedimentary
materials in the Great Lakes. Limnol. Oceanogr. 25:1-11.
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