EPA/600/R-92/183
September 1992
EVALUATION OF TERRESTRIAL INDICATORS FOR USE IN ECOLOGICAL
ASSESSMENTS AT HAZARDOUS WASTE SITES
prepared by
Greg Hinder
ManTech Environmental Technology, Incorporated
Elaine Ingham
Oregon State University
C. Jeffrey Brandt
Battelle Pacific Northwest Laboratory
Gray Henderson
University of Missouri-Columbia
Environmental Research Laboratory
200 S.W. 35th Street
Corvallis OR 97333
submitted to
Clarence Callahan
Project Officer
Ecological Site Assessment Program
United States Environmental Protection Agency
Environmental Research Laboratory
200 S.W. 35th Street
Corvallis OR 97333
September 1992
y&) Printed on Recycled Paper
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DISCLAIMER
The preparation of this document was funded by the United States Environmental
Protection Agency. It was subjected to the Agency's peer and administrative review,
and approved for publication as an EPA document. Mention of trade names or
commercial products does not constitute endorsement or recommendation for use.
ii
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EXECUTIVE SUMMARY
The target audience for this compendium includes regional project managers and
biological technical assistance groups. While primarily compiled to support Superfund
ecological assessment activities, the strategy and methods summarized here may be
applicable to any contaminant-related assessment problem that requires biological
evaluations of soil, sediment/ and water. The methods summarized here are grouped
according to their potential target or receptor classification— animal/ plant, or
microbial. Methods which assess soils directly are emphasized here but additional
methods applicable to wetland soils or sediments have also been included to
complement those methods readily available for aquatic and sediment toxirity
assessment. Some of the methods summarized here represent consensus standards
that have previously been identified as methods applicable to soil testing, while others
are less well characterized but remain potential candidate methods for biological
evaluation of soil contamination. Tabular guides to the selection of tests methods
applicable to various habitats and toxicity endpoints are summarized to help potential
users select the most appropriate biological assessment tool for the site under
consideration.
"Ecological indicators" for the assessment of terrestrial habitats affected by hazardous
waste sites may be numerous and highly site-specific, but in general have not been
critically evaluated. Soil toxicity is one potential ecological indicator for the assessment
of terrestrial habitats impacted by Superfund sites. Evaluation of soil contamination
also represents a potential indicator relevant to ecological assessments regardless of
site location. While numerous compilations of aquatic and sediment test methods are
available for evaluating toxicity as an ecological effect, fewer collections of test
methods are readily accessible for evaluation of soil contamination. This reference
document compiles numerous methods that are currently available for evaluating soil
toxicity, and summarizes them for review and selection by potential users. The
methods are outlined as synopses, and are not intended to provide adequate technical
support for test implementation. However, these synopses do illustrate the state of
development of assessment tools currently available for soil toxicity testing, and
references that accompany each method provide guidance toward more adequate
technical support. This compendium is intended as supplemental information to be
used in conjunction with "Ecological Assessment of Hazardous Waste Sites: A Field
and Laboratory Reference" (EPA/600/3-89/013).
For ecological assessments at hazardous waste sites, soil contamination evaluations
should include acute toxicity tests as well as short-term tests which measure biological
endpoints other than death. Consequently, the compendium includes a variety of
toxicity tests and field methods that are currently available for ecological assessment
111
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and reflects the site-specific demands made by the ecological assessment process. As
ecological indicators of soil contamination, estimates of soil toxicity may yield
information regarding acute biological responses elicited by site-samples as well as
suggest longer-term biological effects (e.g., genotoxicityor teratogenicity) potentially
associated with subacute and chronic exposures to complex chemical mixtures
characteristic of hazardous waste sites. Relative to aquatic toxicity tests methods,
however, the continuing progress in methods development currently affords fewer
well standardized tests for completing a biological assessment of soil contamination.
Owing to the designed flexibility of the test methods outlined in this collection, these
methods should be considered standard guides rather than as strictly interpreted
"cookbook recipes" that should be followed without deviation. While most of these
methods have not been subjected to inter-laboratory testing to evaluate, for example,
reference toxicants, the test methods summarized here have been developed under a
minimal set of defined standard conditions which must be fulfilled in order for the
test to be successfully completed and interpreted. In general, the soil evaluation
methods outlined in the compendium are [1] direct tests of soils or [2] indirect tests
that evaluate soil toxicity, most frequently through testing soil-derived aqueous
eluates. While eluates have been, and will continue to be, evaluated using standard
aquatic toxicity tests or some modification of these methods, supplemental aquatic
tests have also been identified in the collection, owing to their relatively recent
appearance as standard aquatic toxicity test methods. Methods that may be critical to
soil contamination evaluations, but lack sufficient site-specific applications as part of
their validation, have also been summarized here and identified as requiring additional
support when considered as biological assessment tools for any particular site. These
methods, while less well developed relative to most methods included in the
compendium, may be useful for a site-specific evaluation, if they are adequately
supported by more standardized test methods.
In addition to the selected methods catalogued in this compendium, a strategy for
using these biological assessment tools is summarized within the context of ecological
risk assessment for Superfund. From a strategy perspective, an ecological assessment
for a hazardous waste site should be considered an integration of exposure and
ecological effects information, including toxicity evaluations of various environmental
matrices, e.g., soil and water. Depending upon the site, both laboratory tests and
field methods may be required for an ecological assessment. Currently, toxicity
assessments are usually derived from laboratory-generated data, but in situ toxicity
assessments, while not as well developed as laboratory toxicity tests, are becoming
more prominent in the ecological assessment process. In situ methods potentially
offer a closer linkage between contaminant sources and effects, and help reduce the
potential problems associated with laboratory-to-field extrapolations of toxicity data.
IV
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Ecological assessments also rely upon field methods which measure ecological
endpoints, either on-site or at reference sites, and yield survey data relevant to
estimates of adverse ecological effects associated with a waste site. Ecological effects
and exposure assessments/ then/ are complex interrelated functions which yield
estimates of hazard associated with environmental contaminants in various matrices
sampled at a site. Linkages between chemical contaminants and adverse ecological
effects, however, require not only toxicity evaluations for representative species but
also chemical analyses of pertinent site samples. When toxicity assessments are
combined with (1) chemical analyses which evaluate pertinent site samples and (2)
field surveys which measure ecological endpoints, higher level biological organization
(e.g., populations and communities) may be evaluated during the site-assessment
process. Ecological data, then, as well as toxicologicaland chemical information must
be acquired within ecological assessments to assure that sound ecological management
practices are implemented within the remedial investigation/feasibility study (RI/FS)
process in Superfund.
Within ecological contexts, risk assessments must consider current and future
expressions of adverse biological and ecological effects under field conditions, and
potentially following pre- and post-remediation efforts. While not considered in this
compendium, field surveys of hazardous waste sites can identify adversely affected
communities and can provide information for assessing adverse ecological effects
potentially caused by hazardous wastes. Field surveys alone can not identify causes
of effects, however, the tests included in this compendium should be considered
complementary sources of information within the ecological assessment process. For
example, toxicity tests in conjunction with appropriate chemical data can establish
potential linkages between expression of effects and contaminants. The actual causes
mediating adverse effects observed in the field may be hazardous wastes, but these
effects could also be caused or exacerbated by habitat alteration, off-site sources of
toxic chemicals, and natural variability that spuriously suggests linkages between
contaminant and ecological effects. In general, toxicity data and field survey results
should be integrated using, for example, exploratory data analysis and similar
statistical methods, and these preliminary analyses should be considered early in the
problem formulation and scoping efforts. Ideally, the relationships between the
toxicity-derivedand field-derived data sets will be correlative and suggest cause-effect
relationships. If no correlation exists, alternative causes for the effects need to be
examined. Possible cause and effect relationships can be supported by chemical
analyses. In complex mixtures, however, it may be impossible to determine which
chemical or chemicals are causing toxicity, and ecological endpoints (e.g., altered
community structure) may express effects which reflect the composite toxicity of a
complex chemical mixture.
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CONTENTS
Page
EXECUTIVE SUMMARY iii
LIST OF FIGURES x
LIST OF TABLES xi
LIST OF ACRONYMS xii
ACKNOWLEDGEMENTS xiii
1.0 INTRODUCTION 1
2.0 APPROACH 6
3.0 AQUATIC, SEDIMENT, AND SOIL TOXICTTY TEST 8
METHODS
4.0 SOIL CLASSIFICATION AND TESTING FOR 13
EVALUATION OF SOIL MATRIX AND SITE EFFECTS
5.0 CONCLUSIONS AND RECOMMENDATIONS FOR 20
RESEARCH
6.0 BIBLIOGRAPHY 31
TA TECHNICAL APPENDICES
Description of measurement and toxicity assessment TA-1
techniques
TA1 ANIMAL TEST METHODS FOR THE ASSESSMENT OF
SOIL CONTAMINATION AT HAZARDOUS WASTE
SITES
TA1.1 Earthworm survival tests TA1.1-1
TA1.2 Sublethal effects tests using earthworms TA1.2-1
TA1.3 Other tests with soil annelids TA1.3-1
TA1.4 Tests using free-living nematodes TA1.4-1
TA1.5 Tests using soil arthropods (insects) TA1.5-1
TA1.6 Terrestrial arthropod (non-insect) and isopod tests TA1.6-1
VI
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TA1.7 Mollusk tests for evaluating terrestrial and wetland TA1.7-1
habitats
TA1.8 Amphibian test methods TA1.8-1
TA1.9 Acute, subacute or chronic tests using small TA1.9-1
mammals/fur-bearers
TA1.10 Acute and subacute avian toxicity tests TA1.10-1
TA1.11 Immunotoxicity test methods using vertebrates TA1.11-1
TA1.12 Immunotoxicity test methods using invertebrates TA1.12-1
TA2 PLANT TEST METHODS FOR THE ASSESSMENT OF
SOIL CONTAMINATION AT HAZARDOUS WASTE
SITES
TA2.1 Seed germination and root elongation tests TA2.1-1
TA2.2 Early seedling survival and vegetative vigor tests TA2.2-1
TA2.3 Laboratory evaluations with rooted aquatic plants TA2.3-1
TA2.4 Laboratory evaluations with wetland plants TA2.4-1
TA2.5 Laboratory evaluations with upland plants TA2.5-1
TA2.6 Alternative test species in seed germination, root TA2.6-1
elongation, and early seedling survival and vegetative
vigor tests
TA2.7 Short-term tests for evaluating whole plant toxicity TA2.7-1
TA2.8 Life-cycle tests using vascular plants TA2.8-1
TA2.9 Plant tissue culture tests TA2.9-1
TA2.10 Tests evaluating plant community structure TA2.10-1
TA2.11 Ambient air exposure systems and phytotoxicity testing TA2.11-1
TA2.12 Photosynthesis inhibition tests for evaluating sublethal TA2.12-1
effects in plants
TA3 SOIL BIOTA TEST METHODS FOR THE ASSESSMENT
OF SOIL CONTAMINATION AT HAZARDOUS WASTE
SITES
Methods for evaluating contaminant effects on soil biota TA3.0-1
biomass and diversity.
VII
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TA3.1 Evaluation of contaminant effects on bacterial biomass TA3.1-1
TA3.2 Evaluation of contaminant effects on fungal biomass TA3.2-1
TA3.3 Evaluation of contaminant effects on protozoan numbers TA3.3-1
and diversity
TA3.4 Evaluation of contaminant effects on nematode diversity TA3.4-1
TA3.5 Soil microbial biomass (chloroform fumigation) TA3.5-1
TA3.6 Solid-phase and aqueous-phase Microtox TA3.6-1
TA3.7 Alternative microbial toxicity tests (Toxi-and SOS- TA3.7-1
Chromotest)
TA3.8 Static, batch, and continuous-flow microbial tests TA3.8-1
TA3.9 Soil-core microcosm tests TA3.9-1
TA3.10 Microbial growth inhibition: Nitrification and soil TA3.10-1
respiration rates
TA3.11 Soil biochemistry tests: Enzymes TA3.11-1
TA3.12 Soil biochemistry tests: Opid chemistry TA3.12-1
TA3.13 Substrate uptake and decomposition of organic TA3.13-1
compounds
TA3.14 Soil nutrient dynamics tests: Nitrogen cycling TA3.14-1
TA4 FIELD METHODS FOR THE ASSESSMENT OF SOIL
CONTAMINATION AT HAZARDOUS WASTE SITES
Methods for evaluating ambient, or in situ, toxicity in TA4.0-1
terrestrial or wetiand sites
TA4.1 On-site earthworm test methods for biological evaluations TA4.1-1
of terrestrial and wetland habitats
TA4.2 In situ amphibian methods used in biological evaluations TA4.2-1
of wetlands
TA4.3 Nest-box tests using starlings to evaluate in situ toxicity .TA4.3-1
TA4.4 In situ clastogenicity tests (chromosomal aberration assay) TA4.4-1
TA4.5 Field testing with sago pondweed (Potomogeton pectinatus) TA4.5-1
TA4.6 On-site and in situ testing with terrestrial plants TA4.6-1
vxn
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TA4.7 Distribution of resistant microorganisms TA4.7-1
TA4.8 Decomposition rates TA4.8-1
TA4.9 Mycorrhizal colonization TA4.9-1
TA4.10 Lichens as biomonitors TA4.10-1
ix
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LIST OF FIGURES
Page
Figure 1. Ecological risk assessment framework. 2
Figure 2. Ecological assessment of Superfund sites: Overview. 3
Figure 3. Sources of information (toxirity, chemical, and ecological) 5
that contribute to an ecological assessment for a
hazardous waste site.
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LIST OF TABLES
Page
Table 1. Technical references for aquatic and sediment toxkity test 8
methods that may be valuable for evaluating soil
contamination at terrestrial hazardous waste sites.
Table 2. Summary of applicable methods for measurement of soil 18
properties in the field and laboratory.
Table 3. Toxicity endpoints routinely used to evaluate adverse 21
biological effects and their status as indicators of
ecological effects.
Table 4. Habitats frequently impacted by hazardous waste sites 26
and the status of test methods included in the
compendium as indicators of ecological effects.
XI
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LIST OF ACRONYMS
AOSA Association of Official Seed Analysts
ASTM American Society for Testing and Materials
BTAG Biological Technical Assistance Group
CEC Cation exchange capacity
CERCLA Comprehensive Environmental Response Compensation and Liability Act
DOC Dissolved Organic Matter
ECgo Median Effective Concentrations
FIFRA Federal Insecticide, Fungicide, and Rodenticide Act
LCsQ Median Lethal Concentrations
LOEC Lowest Observed Effects Concentration
MATC Maximum Allowable Toxicant Concentrations
NOEC No Observed Effects Concentration
OECD Organization for Economic Cooperation and Development
OM Organic Matter
RCRA Resource Conservation and Recovery Act
RI/FS Remedial investigation/Feasibilitystudy process in Superfund
SARA Superfund Amendments and Reauthorization Act
TSCA Toxic Substance Control Act
US EPA United States Environmental Protection Agency
xii
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ACKNOWLEDGEMENTS
The editors wish to acknowledge Janet Burns, Clarence Callahan, Anne Fairbrother,
Charles Hendricks, Ron Landy, Tom Pfleeger, Ron Preston, Hilman Ratsch, Gerry
Schuytema, US Environmental Protection Agency; Jerry Barker, Michael Bollman,
Rosalind James, Brad Marden, Mark Meyers, Tom Moser, Julius Nwosu, Dave
Wilborn, ManTech Environmental Technology Inc.; Nelson Beyer, Woody Hill, Jim
Fleming, and Chris Ingersoll, US Fish & Wildlife Service; Randy Wentzel, U.S. Army;
Bobby Folsom, Jr., U.S. Army Corps of Engineers; Gerry Walsh, National Park
Service; Isabel Johnson, KBN Engineering; Joe Gorsuch, Eastman Kodak; Larry
Kapustka, Ecological Planning and Toxicology; Phil Ross, The Citadel; Bob Ringer,
Michigan State University; Steve KLaine and Larry Brewer, Clemson University; Don
Miles and Bill Lower, University of Missouri; Mike Richmond, Cornell University;
John Fletcher, University of Oklahoma; Jack Bantle and Karen McBee, Oklahoma State
University; Cheryl Ingersoll and Bob Griffiths, Oregon State University; Glen Suter m
and Barbara Walton, Oak Ridge National Laboratory; and, Ron Thorn, John Thomas,
and Peter Van Voris, Battelle Pacific Northwest Laboratory for their contributions to
this report. We also thank the reviewers who provided many helpful comments on
the manuscript.
Greg Under
ManTech Environmental Technology, Inc.
Elaine Ingham
Oregon State University
Jeff Brandt
Battelle Pacific Northwest Laboratory
Gray Henderson
University of Missouri-Columbia
XI11
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1.0 INTRODUCTION.
Ecological assessments for hazardous waste sites have gained increased attention after
the passage of the Superfund Amendments and Reauthorization Act of 1986 (SARA).
As a result, US EPA has drafted numerous guidance documents which suggest
conceptual approaches to the evaluation of adverse ecological effects which may exist
at hazardous waste sites (e.g. Figure 1; US EPA 1988; US EPA 1989a; Warren-Hicks, et
al. 1989; US EPA 1991). The application of ecological assessment techniques has
become an integral part of the remedial investigation/feasibility study (RI/FS) process
and in general, ecological considerations have assumed a greater role in the RI/FS
process for Superfund sites. Alternative approaches developed under different US
EPA programs (e.g., Urban and Cook 1986) have also contributed to the strategy
developed for ecological assessments at Superfund sites.
While the concept of indicator species has not been critically evaluated, "ecological
indicators" for the assessment of terrestrial habitats impacted by Superfund Sites may
be numerous and highly site-specific (Angus 1991). As with aquatic toxicity, however,
soil toxicity evaluations represent one potential measurement endpoint that is relevant
to ecological assessments regardless of site location. For ecological assessments at
hazardous waste sites, soil contamination evaluations should include acute toxicity
tests as well as short-term tests which measure biological endpoints other than death.
The variety of toxicity tests and field methods available for the ecological assessment
reflects not only the site-specific demands made by the ecological assessment process,
but the continuing progress in methods development. As ecological indicators of soil
contamination, estimates of soil toxicity may yield information regarding acute
biological responses elicited by site-samples as well as suggest longer-term biological
effects (e.g., genotoxitity or teratogenicity) potentially associated with subacute and
chronic exposures to complex chemical mixtures characteristic of hazardous waste
sites.
1.1 Approaches to Ecological Assessment.
As ecological indicators for terrestrial hazardous waste sites, toxicity tests are but one
component of an ecological assessment for a hazardous waste site. Field methods that
evaluate contaminant effects must be given equal regard during the early phases of
site evaluation. More importantly, adequate statistical considerations must be given to
field studies in order to assure that the soil being sampled will yield information
pertinent to the site assessment. This becomes particularly important when field
sampling designs are considered, since integration of toxicity assessments (be those in
situ or laboratory-generated)and field assessments requires a well-designed sample
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plan to establish linkages among toxicity, site-sample chemistry and adverse ecological
effects, if apparent.
For an ecological assessment at a hazardous waste site, an integration of exposure and
ecological effects information yields an estimate of risk associated with any particular
waste site (Figure 2). As such, ecological assessments at Superfund sites reflect the
site-specific demands required by waste sites, and represent integrated evaluations of
ecological effects, including toxicity, and exposure. Depending upon the site, both
laboratory tests and field methods will be required for an ecological assessment.
Figure 1: Ecological risk assessment framework (after US EPA 1992)
Problem Definition/Scoping
Characterization
of
Stress
Characterization
of
Ecological
Effects
Risk Characterization
Decision-Making/
Risk Management
Verification
and
Monitoring
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Figure 2: Ecological assessment of Superfund sites: Overview (after US
1991)
PROBLEM FORMULATION
* Qualitatively evaluate contaminant release, migration, and fate
'Identify:
- Contaminants of ecological concern - Exposure pathways
- Receptors - Known effects
* Select endpoints of concern
* Specify objectives and scope
EXPOSURE ASSESSMENT
* Quantify release, migration, and fate
* Characterize receptors
* Measure or estimate
exposure point concentrations
ECOLOGICAL EFFECTS
ASSESSMENT
* Literature
* Toxicity testing
* Field studies
RISK CHARACTERIZATION
* Current adverse effects
* Future adverse effects
* Uncertainty analysis
* Ecological significance
REMEDIAL OBJECTIVES
ANALYSIS OF
REMEDIAL ALTERNATIVES
* REMEDY SELECTION
* RECORD OF DECISION
* REMEDIAL DESIGN
* REMEDIAL ACTION
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Toxicity assessments are derived from acute tests as well as subacute and chronic tests
which measure biological endpoints other than death, and are generally completed as
part of the ecological assessment. These toxicity assessments are usually derived from
laboratory-generateddata, but in situ toxicity assessments, while not as well developed
as laboratory toxicity tests, are becoming more prominent in the ecological assessment
process (Warren-Hicks, et al. 1989). In situ methods more closely infer a linkage
between toxicity and exposure functions, and reduce the problems associated with lab-
to-field extrapolations of toxicity data. Ecological assessments also rely upon field
methods which measure ecological endpoints, either on-site or at reference sites, and
yield survey data relevant to estimates of adverse ecological effects associated with a
waste site.
Depending upon the environmental matrix being tested, site-specific toxicity
assessments may be derived using various tests (e.g., Peltier and Weber 1985; Weber,
et al. 1988; US EPA 1989), and may include invertebrate, vertebrate, algal, plant, and
microbial test systems. Potential contaminant migration as well as soil attenuation of
contaminant effects may also be evaluated using these toxicity assessment tools, if soil
eluates or sediment elutriates are prepared and tested in the laboratory.
Ecological methods applicable to hazardous waste site evaluations are varied, and
reflect the diversity apparent in waste sites which occur in a variety of geographic
settings. Depending upon the environmental setting (aquatic or terrestrial) and the
site-specific questions being asked in the evaluation process, methods have been
compiled (e.g., Parkhurst, et al. 1989; LaPoint and Fairchild 1989; Kapustka 1989;
McBee 1989; Bromenshenk 1989; Simenstad, et al. 1991; Holland 1990) which provide a
source of techniques available for ecological site assessments.
1.2 Assessment strategy.
From a technical perspective, an ecological assessment may be considered an
integrated evaluation of biological effects derived through measurements of toxicity
and exposure (Warren-Hicks, et al. 1989). Within the context of Superfund ecological
assessment, chemical and biological interactions associated with contaminant
exposures in soil may be evaluated according to various assessment strategies. As
previously summarized, both chemically-based and toxicity-based approaches have
made significant contributions to ecological assessments for hazardous waste sites
(Parkhurst, et al. 1989). From an ecotoxicologicalperspective, ecological effects and
exposure assessments are complex interrelated functions which yield estimates of
hazard associated with environmental contaminants in various matrices sampled at a
site. In many respects, soil toxicity may be considered in a manner similar to
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contaminated sediments (e.g., sediment quality triad; Chapman 1986). Here, linkages
between chemical contaminants and adverse ecological effects require not only toxicity
evaluations of representative species but also chemical analyses of pertinent site
samples (Parkhurst, et al. 1989; Stevens, et al 1989). When toxicity assessments are
combined with (1) chemical analyses which evaluate pertinent site samples and (2)
field surveys which measure ecological endpoints, higher levels of biological
organization (e.g., populations and communities) may be evaluated during the site-
assessment process. Ecological data, then, as well as toxicologicaland chemical
information must be acquired within ecological assessments to assure that sound
ecological management practices are implemented (Figure 3).
Figure 3. Sources of information (toxicity, chemical, and ecological) that
contribute to an ecological assessment for a hazardous waste site.
Toxicity data Chemical data Ecological data
T
I
T
Statistical or Quantitative
Integration
I
T
Contribution to ecological
assessment
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2.0 APPROACH.
Methods development has yielded numerous candidate test methods for evaluating
contaminated soil, and those methods summarized previously (US EPA 1989; Warren-
Eficks, et al, 1989) are updated in this compendium to reflect that expanded suite of
toxicity assessment methods applicable to soil contamination evaluations. Methods
which assess soils directly are emphasized in Section 5.0 and the Technical
Appendices, but additional methods applicable to wetland soils or sediments have
been included to complement those methods readily available for aquatic and
sediment toxicity assessment (see Section 3.0). Some of the methods summarized
here represent consensus standards, for example by ASTM, that have previously been
identified as methods applicable to soil testing. Others are less well characterized but
remain potential candidate methods for biological evaluation of soil contamination.
Owing to the designed flexibility of the test methods outlined in the Technical
Appendices, these methods should be considered standard guides rather than a
strictly interpreted "cookbook recipe" that should be followed without deviation.
While most of these methods have not been subjected to inter-laboratory testing to
evaluate, for example, reference toxicants, the test methods summarized here have
been developed under a minimal set of defined standard conditions which must be
fulfilled in order for the test to be successfully completed and interpreted.
La general, the soil evaluation methods outlined in the Technical Appendices are (1)
direct tests of soils or (2) indirect tests that evaluate soil toxicity, most frequently
through testing soil-derived aqueous eluates. While eluates have been, and will
continue to be, evaluated using standard aquatic toxicity tests (e.g., Peltier and Weber
1985; Weber, et al. 1988) or some modification of these methods (e.g., US EPA 1989),
supplemental aquatic tests have been identified owing to their relatively recent
appearance as standard aquatic toxicity test methods. Methods that may be critical to
soil contamination evaluations, but lack sufficient site-specific applications as part of
their validation, have also been summarized in the Technical Appendices. These
methods, while less well developed relative to most methods described, may be useful
on a site-specific basis, but must be adequately supported by more standardized test
methods throughout the soil contamination evaluation process.
2.1 Technical Appendices
Section 5.0 presents the results of the screening and evaluation process. Tables 3 and
4 provide a key for selection of bioassays and measurements appropriate for use in
different habitats, to determine specific measurement endpoints, and to assess
particular contaminant classes. The Tables refer the reader to descriptions of these
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methods that are presented in detail in Technical Appendices. These descriptions are
grouped by animal test methods (Tl), plant test methods (T2), soil biota test methods
(T3), and field test methods (T4) for the assessment of soil contamination at hazardous
waste sites. The following general information is provided for each of the methods:
• Test method summary: A brief abstract of the test method is given
including the species and life stage used for testing, exposure regime,
and bioassay endpoints. General testing procedures for conducting the
bioassay are also provided. Where available, this information includes
ecological relevance of the bioassay, sampling requirements, sample
toxicity testing methods, and QA/QC requirements.
• Intended use: A description of the conditions or assay endpoints for
which the method was originally developed is provided. Restrictions for
use of the method or interpretation of the test results under different
conditions are given.
* Previous applications/regulatoryprecedence: A brief review of the
previous applications of the method in assessing environmental
contamination is given. If the method has been used in a regulatory
context, this is noted. Limitations apparent from previous work with the
method are also discussed.
• Requirements for development and implementation: The status of
protocol availability or development is discussed. This includes degree
of standardization, documentation, availability of test organisms and
qualified laboratories to conduct the bioassay, and ability to interpret test
results in an ecological context.
• Potential problems and limitations: Problems with the use of the
method in evaluating soil contamination are identified. Examples of
problems that could affect use include lack of sensitivity to certain
chemicals, large degree of variability in test results, restrictive
requirements when conducting the test, and difficulty in interpretation of
test results.
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3.0 AQUATIC, SEDIMENT, AND SOIL TOXICITY TEST METHODS.
Because of their ready availability, aquatic tests (marine, estuarine, and freshwater)
and sediment tests (marine, estuarine, and freshwater) have not been outlined here.
Sources of information regarding their potential in site assessment and as ecological
indicators for terrestrial hazardous waste sites have been noted, however (Table 1).
Here, only recently developed aquatic and sediment test methods that are perhaps
less widely known, have been briefly summarized along with soil tests that evaluate
terrestrial indicators and soil toxicity.
Table 1. Technical references for aquatic and sediment toxicity test methods that
may be valuable for evaluating soil contamination at terrestrial hazardous
waste sites.
TEST MATRIX
TARGET BIOTA
REFERENCE
Freshwater
Freshwater/marine/
estuarine
Freshwater
Marine
Freshwater sediments
Marine/estuarine
sediments
Vascular plants Wang, 1991
Algae and vascular plants Swanson, et al. 1991
Aquatic vertebrates and
invertebrates
Marine/estuarine
invertebrates and
vertebrates
Epifauna, infauna, and
vertebrates
Epifauna, infauna, and
vertebrates
Peltier and Weber 1985
Weber etat. 1989
ASTM 1991
Weber etal. 1988
ASTM 1991
ASTM 1991
ASTM 1991
While not as readily available as aquatic toxicity test methods (e.g., Peltier and Weber
1985; Weber, et al. 1988), methods have been identified for testing soil biota (e.g., US
EPA 1989). On a site-specific basis, the application of biological tests should provide a
comparative toxicity database upon which site-specific soil evaluations can be
completed or cleanup criteria can be established. Screening (neat soils yielding
percent effect) and definitive tests (amended soils potentially yielding median effective
concentrations) may be completed with standardized test species to evaluate toxicity
8
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within a biological assessment. Additionally, to assure adequate information for
ecological evaluations of soil contamination, species having site-specific relevance may
also be tested (Parkhurst, et al. 1989). When performed in parallel with standard test
methods/ these site-specific tests, e.g., using resident plant species) may be diagnostic
and indicate biological responses, e.g., development of metal resistance, that are
associated with soil exposures. Presently, the application of laboratory bioassays to
site assessment is increasing, particularly in developing site-specific biological data
bases which contribute to ecological assessments. To enhance the ecological relevance
of site-specific biological tests and to reduce the potential extrapolation error associated
with inter-specific comparisons, use of standard test species in ecological assessments
should be considered in soil testing.
For evaluating indicators of exposure at terrestrial hazardous waste sites, soils may be
sampled, then tested directly (i.e., soils). Alternatively, in the laboratory site-soil may
be mixed with water to yield an eluate that contains water-soluble soil constituents
that could migrate off-site or enter groundwater. The justification for eluate testing
should be influenced by site-specific considerations. As "stand-alone" methods,
however, eluate tests may not be appropriate for soil toxicity assessments unless soils
are acting as potential contaminant sources, for example, in evaluating ecological
effects associated with surface runoff (see also, Thomas, et al. 1986, Miller, et al. 1985).
In soils, toxicity tests can provide data on the acute (short-term) and chronic (long-
term) toxicity of contaminated media to aquatic and terrestrial biota. These tests are
generally conducted using standard laboratory test species, but in some cases tests on
alternative test species may be appropriate. If the test species are representative of
sensitive, resident species, the toxicity data may provide an assessment of the
potential for causing the adverse effects measured in field surveys. Toxicity endpoints
are derived from acute and chronic toxicity tests, and most frequently consider
(1) percent survival of the test organisms in 100% site sample (water, soil, or
sediment) in laboratory tests or in situ exposures, and (2) concentration-percent
survival relationships for laboratory tests run at several test concentrations of soil.
The latter test design potentially yields estimates of median lethal concentrations
(LCgoS), median effective concentrations (EC^ for growth inhibition), or maximum
allowable toxicant concentrations (MATCs).
Survival data for 100% test concentrations and the in situ exposure data provide
information on the direct toxicity of ambient concentrations of chemical contaminants
in soils. These data can be directly compared to survey data to assess probable
sources and causes of toxic effects. For example, if a 100% concentration of the test
material in a laboratory (or in situ) exposure caused mortality to earthworms in 100%
site-soil, and the soil macroinvertebrate community of the site appeared impacted,
-------
then there is a high probability that toxicity is directly or indirectly associated with
adverse effects. Similarly, concentration-percent survival relationships could be used
to extrapolate toxicity data to sites with decreasing contaminant concentrations, and if
calculated, median estimates of toxicity (e.g., LCgoS and ECgoS) or MATC estimates
could be used for toxicity comparisons among different locations within a site.
Acute tests measure lethal effects, but sublethal effects (e.g., altered reproduction)
must also be measured. Acute toxicity test results are usually expressed as LCg^ and
ECgoS for the test duration, e.g., 96-hour LC^. Other effect levels besides 50% can be
calculated, provided the test design allows those estimates, e.g., LC^s. Acute toxicity,
however, should not be the only biologicalresponse considered in the ecological
assessment. At many sites, sublethal and chronic effects associated with exposure
may be more significant in mediating ecological effects. As such, chronic tests
potentially detect both chronic lethal effects and sublethal toxicity, such as effects on
growth and reproduction. Chronic test results can be expressed in the same manner
as acute test results, but they are often expressed as estimates of safe concentrations
or toxicity threshold concentrations. MATCs are usually bounded by two test
concentrations within a site-soil amendment series. For example, the NOEC (no
observed effects concentration) is the highest test concentration that causes no
statistically significant toxic effects, and the LOEC (lowest-observed-effects-
concentration) is the lowest concentration that causes statistically significant toxic
effects. Presumedly, these two values, the NOEC and LOEC, span the toxicity
threshold for the chemical, and their geometric mean, or GMATC (geometric mean of
the MATC) would estimate the chronic toxicity threshold. Peltier and Weber (1985)
and Weber, et al. (1988) provide detailed discussions of these toxicity values and
methods for their calculation. While these references bear directly on results
generated by aquatic toxicity testing, statistically similar analyses can be applied to soil
contamination evaluations for Superfund.
3.1 Integration of Toxicity Tests with Field Surveys.
Field surveys can identify adversely affected communities and can provide information
for assessing adverse ecological effects potentially caused by hazardous wastes.
However, field surveys alone can not identify causes of effects. Toxicity tests in
conjunction with appropriate chemical data can establish potential causes. The actual
causes mediating adverse effects observed in the field may be hazardous wastes, but
these effects could also be caused or exacerbated by habitat alteration, off-site sources
of toxic chemicals, and natural variability that spuriously suggests linkages between
contaminant and ecological effects.
10
-------
In general, toxicity data and field survey results should be integrated using, for
example, exploratory data analysis and similar statistical methods (Gilbert 1987;
Gordon 1981). These preliminary analyses should be considered early in the problem
formulation and scoping efforts (see Figures 1 and 2). Ideally, the relationships
between the toxirity-derivedand field-derived data sets will be correlative and suggest
cause-effect relationships, or conversely, the lack of correlation may suggest alternative
causes for the effects observed in the field. Possible cause and effect relationships can
be supported by chemical analyses. In complex mixtures, however, it may be
impossible to determine which chemical or chemicals are causing toxicity, and
ecological endpoints (e.g., altered community structure) may be expressing effects
which reflect the composite toxicity of a complex chemical mixture. Various
fractionationand toxicity identification techniques, similar to those used in effluent
monitoring (e.g., US EPA 1991) and sediment evaluations, may also be developed to
evaluate more completely the causative toxic chemicals in complex mixtures, but again
these methods are poorly standardized when working with soils.
3.2 Toxicity tests for wetland plant-dominated subunits of terrestrial Superfund
sites
Plants associated with wetlands or other aquatic habitats have been used extensively
to assess water and sediment quality. The wide variety of tests developed have
targeted both the effects of water column and sediment-borne toxic materials. The
types of aquatic vegetation used for these purposes range from microscopic unicellular
algae to relatively large flowering plants. The three most commonly applied test
methods include chlorophyll a concentration, growth, and contaminant uptake.
Growth measurements (biomass accumulation per unit of time) have been widely
applied as an assessment method on a variety of freshwater estuarine and marine
species. Much of the testing has been conducted on sediments in the laboratory,
using unicellular phytoplanktonsuch as Sdenastrum capricornutum (freshwater) and
Skeletonema costatum (marine) (e.g. Ankley et d, 1990; Thomas et al, 1990). Use of
rooted wetland macrophyte growth has been very limited.
Growth is perhaps the least specific measurement endpoint. A response, such as
reduced growth rate, is not tied to specific sites within the plant where reactions or
processes are altered by specific chemicals. This is especially true for rooted
macrophytes. The advantage of measuring growth is that it is an integrator of all
effects of toxicants on plants, it is relatively easy to measure, there is a wide range of
past use, and it can be done with acceptable precision in both the field and laboratory.
11
-------
The physiology of chlorophyll production and maintenance is quite well known.
Chlorophyll occurs in virtually all plants, and is the primary pigment involved in the
important ecological process of photosynthesis. The correlation between chlorophyll
concentration and photosyntheticrate is commonly strong.
Chlorophyll concentration relative to contamination of water or soils has been
measured in unicellular algae, macrophytes and in periphyton communities (e.g. Bassi
et al, 1990). Chlorophyll concentration generally reflects the mass of plant material
present as well as an indication of the health of the material. Toxicants can affect the
chlorophyll molecule directly or through the process of energy transfer during
photosynthesis (e.g Judy et al. 1991).
Contaminant uptake by plants has been applied primarily to rooted macrophytes. It is
assumed that most of the uptake occurs through the roots and that the concentration
of the contaminant compounds in leaf tissues is directly related to the concentration in
the soil or sediment. Uptake has received wide application in fresh and marine
systems, and has been carried out under both laboratory and field conditions (e.g.
Kovacs, 1978; Lee etal., 1981).
Uptake of contaminants relies on several assumptions that may hinder the
interpretation of results. Chemicals may be modified to form non-toxic compounds by
the plant. Certain chemicals are not concentrated, while others are, which may bias
the interpretation of what chemicals are present in the test medium. Finally, uptake
rates may be inhibited by the toxicity of other materials in the medium and the test
organism may be inhibited in its ability to accumulate the contaminants.
While measurements of plant growth, chlorophyll content, and contaminant uptake
are the most commonly used methods, several other are in various stages of
development and implementation. These methods include measurements of
photosyntheticrate, chloroplast morphology, peroxidase activity, root growth, seed
germination, seedling growth, and reproduction (e.g. Gorsuch et al. 1990).
The strongest approach to the assessment of wetland subsystems may be to use a
combination of several methods to evaluate contamination of water and sediments.
This combination would provide indication of both ecological and physiological
responses of the plants to the media, and would increase the power of the analysis
through verification of responses using several endpoints.
12
-------
4.0 SOIL CLASSIFICATION AND TESTING FOR EVALUATION OF SOIL
MATRIX AND SITE EFFECTS
At any site the physical, chemical and biological properties of the soil partially
determine its productivity. In this document we discuss methods for evaluating the
effects of contaminants in the soil on this productivity. These contaminants primarily
cause changes in soil chemical properties but may indirectly also influence biological
and, to a lesser extent, physical soil properties. However, the "ambient" soil
properties predispose a site to a certain productivity level. It is necessary to evaluate
basic soil parameters in order to assess this ambient condition and therefore, any
deviation attributable to the contaminant. In addition, a quantitative knowledge of
certain soil properties is often necessary for calculations involving the contaminant in
question, (e.g. the conversion of contaminant concentration to loading rates on an
area! basis requires knowledge of the bulk density and depth of the contaminated soil
horizon)
Adequate description of soil properties involves both field observations and laboratory
measurements. Field observations include variables that are commonly recognized as
site characteristics while laboratory measurements are those commonly associated with
traditional soil testing. Both are important in defining soil productivity. This
discussion of soil properties is separated into a field and a laboratory section.
4.1 Field (Site) Observations
Field observations can be subdivided into those which describe general site conditions
and those which specifically describe the soil or soils that are found at a site. General
site parameters are type of overstory and understory vegetation, aspect, slope and
slope position. Aspect and slope characteristics are important parameters related to
the water relations of a site. Vegetation type and distribution is often an indicator of
spatial variability of natural soil (site) or contaminant distribution.
Specific soil parameters are those integral to a complete profile description. This
profile description should be conducted to a depth sufficient to adequately represent
the rooting volume of the soil. It should be conducted for all major soil types at a
site. The description should delineate the depth and thickness of the horizons present
in the profile, including the litter (O) horizon, if present. For each horizon, the
following information should be collected; color, structure, rock content, bulk density,
and root distribution. Homogeneous samples should be collected from each horizon
for laboratory testing (see section 4.2). In addition, it may be desirable to measure
infiltration capacities of water to the soil surface and hydraulic conductivities of water
13
-------
within the profile. These would be important if a contaminant is susceptible to being
carried by water as it moves over or through the soil profile. The utility of these types
of observations is discussed below.
Horizon description - Delineating the types and sequences of horizons present
accomplishes two major tasks. When compared to undisturbed sites a
complete description documents the absence (removal), enrichment
(augmentation or deposition), and/or mixing of soil horizons and thus indicates
the degree of disturbance. Secondly, the description forms the basis for field
observation and sampling for laboratory analyses. In this context the depth of
various horizons can be used to calculate volumetric quantities of various
elements or contaminants.
Color - A Munsell color chart (book) is used to evaluate soil color. Color is
primarily used to indicate organic matter status of a soil horizon or the water
relations in a particular horizon or profile. Dark colors indicate large contents
of organic mater while light colors generally point to lower contents. Bright
colors (generally reds and yellows) indicate soils that have greater internal
aeration; that is, they are better drained. Gleyed soils (blue-gray colors) indicate
anaerobic conditions (poor internal drainage) in which a soil is commonly
saturated for a major portion of the year. Variable coloring with interspersed
gley and bright colors (mottles) indicates a horizon or profile that is subject to a
fluctuating water table and therefore experiences alternating aerobic and
anaerobic conditions. The internal oxygen/water relationship is important
because it determines the reduction-oxidationpotential of a soil and hence
influences chemical forms and some chemical reactions that can occur in the
soil matrix.
Structure - Structure refers to the way and degree individual soil particles bind
together in larger units. The type of structure in a soil influences the paths of
water movement (in reality, is partially dependent on water movement) and
paths of root penetration and expansion. Lack of structure can be an indication
of intense physical disturbance.
Bulk density - Bulk density refers to the weight per unit volume of solid
particles in a soil. It is a measurement that must be taken in the field; it can
not be successfully reconstructed from laboratory data. Bulk density indicates
the degree of natural or induced compaction in a given horizon. It is inversely
related to porosity, or the amount of air space in a soil. Compacted (bulk
densities with greater values) Horizons are barriers to both root growth and
vertical (downward) water movement and may induce lateral pathways of
14
-------
water flow. Bulk density is also a parameter that is necessary in order to
convert concentrations of a chemical constituent (usually reported on a weight
basis) to an area or volume basis. It is used in conjunction with horizon
thickness to calculate the content of a constituent (element) in that horizon.
Rock Content - The rock or stone content of a soil horizon can be an important
consideration because of its influence on water relations, the volume of finer
soil particles important in soil chemical reactions, and/or the soil volume
available for root growth. With respect to water relations, greater stone content
reduces the soil volume active in storing soil water for subsequent plant use
(lowers soil water holding capacity). In addition the paths of water movement
within a soil are determined by the amount and arrangement of any rock that is
present. In general, the greater the rock content the greater the volume of
water interacting with the fine earth portion of a soil horizon and the greater
the potential for chemical alteration and/or leaching.
Root distribution- Root growth and the patterns of root distribution found in a
soil profile provide an integration of most of the soil properties discussed above
as well as the chemical properties discussed in Section 4.2. Lack of roots (or
reduced rooting) in a particular horizon or soil serve is an indication that a
chemical, physical or biological barrier is inhibiting root growth. However, it is
necessary to recognize there is a "natural" root distribution characterized by
decreasing root distribution with depth. This natural distribution is a function
of vegetation type and regional climate and can be characterized to provide a
standard for comparison of subject sites.
Infiltration capacity and hydraulic conductivity- Infiltration capacity is a
measurement of the ability of the soil surface to absorb water. The greater the
value the greater the rate at which water is taken into a soil and the lesser the
chance of surface runoff and thus erosion. Infiltration capacity is strongly
related to texture and organic matter content of surface horizons. Hydraulic
conductivity refers to the rates of movement within a soil profile. It is also
referred to as permeability. It is important because it characterizes the
"residence time" of water within a soil profile and hence its chances of
interacting with soil chemical constituents. Hydraulic conductivity is directly
related to infiltration capacities. The slower the hydraulic conductivity the
lesser the infiltration capacity.
15
-------
4.2 Laboratory Analysis
Various laboratory analyses are typically conducted on samples obtained from field
sites. These analyses are important for understanding the chemical, physical and
biologicalreactions that might take place in the soil. They are also important for
assessing the mobility and potential pathways of movement of a contaminant,
especially with regard to water movement and water mediated biological
decomposition. However, the results of these analyses only represent the properties
of the sample itself. Adequate interpretation must consider the entire site setting and
the horizon the sample came from in the soil profile. As such, the thoroughness of
field observations discussed hi Section 4.1 is especially critical.
Laboratory analyses suggested for routine characterization and interpretation of soils
include those for texture, organic matter, pH, cation exchange capacity, base
saturation, extractable acidity, electrical conductivity, water retention, and
concentrations of the elements nitrogen, phosphors, potassium, calcium, magnesium
and sodium. The importance of these soil properties is discussed below.
Texture - Texture refers to the proportion of particles of different size classes
(sand, silt, clay) that make up a sample. Knowledge of texture is important for
assessing water relations, erodibility, and retention of cations, anions, and
toxicants.
Organic matter content - The organic matter content of a soil is an indicator of
its fertility, ability to support microbial populations, retention of elements, and
water relations. In general, the greater the organic matter content the better the
soil serves as a medium for plant growth.
pH - The pH of a soil sample is a measure of its hydrogen ion concentration (or
more accurately, its activity) and thus its acidity or alkalinity. This is an
important property because it influences (or regulates) many chemical and
biological processes occurring in a soil, including the availability of soil
nutrients.
Cation exchange capacity - Cation exchange capacity (CEC) is a measure of the
soil's ability to retain positively charged ions. The greater the CEC the greater
the capacity of a soil to retain cations; the greater its "storage" or "buffering"
capacity. CEC is especially important for assessment of the mobility of
positively charged metallic elements.
16
-------
Base saturation and extractable acidity - These two measurements characterize
opposite ends of the spectrum relative to the composition of cations occupying
the cation exchange complex of a soil. They are strongly related to the pH of a
soil. Base saturation represents the proportion of the CEC electrically satisfied
by base cations such as calcium, magnesium, potassium and sodium whereas
extractable acidity measures the proportion satisfied by the acidity cations,
Hydrogen and Aluminum.
Electrical conductivity - Electrical conductivity is an indirect measurement of
the ionic content of a soil sample. The greater the conductivity the greater the
ion content of a soil. It is an especially important measurement in soils that
have "free" salts such as saline soils, and serves as an indicator of adverse
conditions for growth of many plants in such soils.
Water retention - Water retention refers to the amount of water that can be
retained by soil when it is subjected to different pressures. Characteristicallya
"water retention curve" is established for a soil sample by measuring its water
content (on either a volumetric or weight basis) at 1/10-, 1/3-, 1-, 5-, and 15-bar
pressures. The difference in water contents between the 1/3-bar (also termed
field capacity) and 15-bar (wilting point) pressures is termed the available water
content of a soil. It represents the amount of water storage the soil can provide
against the force of gravity; the amount available for plant uptake.
Concentrations of various elements - Knowing the concentrations of the
complete spectrum of nutrient elements (especially nutrient elements) in a soil
allows interpretation of whether a specific nutrient deficiency is possible or
whether there is an imbalance in the distribution of elements present in a soil
or soil sample. Specifically, these analyses of extractable element concentrations
provide a measure of their availability for plant or microbial uptake.
References to instructions for performing specific analyses are given in Table 2. These
references, especially the two ASA Monograph Series, also provide information
valuable in interpretation of analytical results.
17
-------
Table 2. Summary of applicable methods for measurement of soil properties in the
field and laboratory.
SOIL PROPERTY
DETERMINATION IN:
FIELD OR
LABORATORY
REPORTING UNITS
REFERENCES
Texture
Laboratory
Organic matter (carbon) Laboratory
content
pH
Cation exchange
capacity (CEC)
Base saturation
Ex tractable acidity
Phosphorous
Water retention
(Available water
capacity)
Laboratory
Laboratory
Laboratory
Extractablenutrients Laboratory
(Bases: Ca, Mg, K, Na)
Nitrogen (Various Laboratory
forms)
Laboratory
Electrical conductivity Laboratory
Laboratory
Percent by weight in
various size classes (i.e.
% sand, silt, day)
Percent of dry weight
(% C or % OM)
Laboratory and/or field pH units
cmol(+)/kg
meq/lOOg
Percent of CEC
cmol/kg
meq/lOOg
cmol/kg
meq/lOOg
Percent of dry weight or
mg/kg
mg/kg
fjzahoB
Volume percent (cm/cm)
SCS (1982) pp. 17-23;
ASA Monograph 9, Part
1, Chapter 15
SCS (1982) pp. 43-47;
ASA Monograph 9, Part
2, Chapter 29
SCS (1982) pp 87-89;
ASA Monograph 9, Part
2, Chapter 12
SCS (1982) pp. 37-40;
ASA Monograph 9, Part
2, Chapter 8
SCS (1982) p. 40;
ASA Monograph 9, Part
2, Chapter 9
SCS (1982) pp. 61-63;
ASA Monograph 9, Part
2, pp. 161-164
SCS (1982)pp. 66-72;
ASA Monograph 9, Part
2, Chapters 13-14
SCS (1982) pp. 47-48;
ASA Monograph 9, Part
2, Chapters 31-36
SCS (1982) pp. 73-75;
ASA Monograph 9, Part
2, Chapter 24
SCS (1982) pp. 93-94;
ASA Monograph 9, Part
2, Chapter 10
SCS (1982) pp. 27-30;
ASA Monograph 9, Part
1, Chapters 26,36
18
-------
SOIL PROPERTY DETERMINATION IN: REPORTING UNITS REFERENCES
FIELD OR
LABORATORY
Bulk density
Field and laboratory
g/cm3
mg/m3
SCS (1982) pp. 24-27;
ASA Monograph 9> Part
1, Chapter 13
Infiltration capacity Field cm/hour ASA Monograph 9, Part
1, Chapters 32-34
Hydraulic conductivity Field or laboratory cm/day or cm/hour ASA Monograph 9, Part
1, Chapters 28-31
19
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5.0 CONCLUSIONS AND RECOMMENDATIONS FOR RESEARCH
Within regional settings, the methods compiled in this catalogue may contribute to the
biological and ecological assessment process for Superfund. Selection of biological test
methods as components of an ecological assessment for a hazardous waste site should be
completed in consultation with regional BTAGs or similar groups, and should reflect the
regional landscape and habitat features unique to the site.
As a guide to method selection, summary tables have been prepared that use biological or
ecological endpoints (e.g., toxicity endpoints, habitats) to identify which methods may be
most critical to the ecological assessment. The selection of test methods would be most
readily accomplished if the early phases of the ecological assessment process are
completed, and site-specific characteristics dearly identified as "filters" for directing the
test method selection process. In Tables 3 and 4, each cell of the "test X endpoint" matrix
evaluates: Status of method development (Extensive data base, +++; Limited data base,
++; Needs support, +)/Status of endpoint interpretation for ecological significance
(Extensive data base, +++; Limited data base, ++; Needs support, +); "blank cell" = not
applicable; "ns" = lack sufficient data base for decision.
For example, during the problem formulation and exposure assessment phases of the
ecological assessment process suspected contaminants and potential ecological targets or
receptors should be identified. Then, the most likely candidate test methods could be
selected, based in part on their being adequate surrogate species for testing or their
potential role in ecological interpretations that may be developed as a consequence of
biological testing. These tables should be considered suggested outlines for ensuring that
the biological tests selected are the most appropriate given the existing site conditions.
Within a regional setting, however, the tables should be modified and detailed to reflect,
for example, the ecoregional characteristics that would be present across various sites that
may require an ecological assessment (e.g., Great Lakes states would probably not
consider test methods pertinent to desert habitats).
20
-------
Table 3: Toxicity endpoints routinely used to evaluate adverse biological effects and
their status as indicators of ecological effects.
Method of
evaluation (see
Technical
Appendix for
methods synopses)
TA1.1 Earthworm
survival test
TA1.2Sublethal
effects tests using
earthworms
TA1.3 Other soil
annelids
TA1.4 Tests using
free-living
nematodes
TA1.5 Tests using
soil arthropods
(insects)
TA1.6 Terrestrial
arthropod (non-
insect) and isopod
tests
TA1.7Mollusk
tests
TA1.8 Amphibian
tests
TA1.9 Small
mammals and fur-
bearers
TA1. 10 Avian tests
TA1.11 Immuno-
toxlcity tests using
vertebrates
TA1.12Immuno-
toxicity tests using
invertebrates
Survival
*
• •I /i
TTT/T
*
«*
"T**rT/T • '
++/+
++/+
+++/+
«<*
+++/+
ns
ns
Growth
X i / i i
TT/TT
+/+
++/+
ns
ns
+/+
+++/+
++/+
++/+
ns
ns
Repro-
duction
++/++
+/+
ns
ns
ns
ns
ns
++/++
•H7++
ns
ns
Develop-
mental
H*
«*
+++/+ +
ns
ns
Biochem-
ical
*
*
*
ns
+/+
*
+/+
*
«
Geno-
toxicity
**
*
«*
++/+
Immuno-
toxicity
*
+/+
«*
«*
Population
and
community
+/++
*
ns
ns
ns
ns
ns
Each cell of the "test X endpoinf matrix evaluates: Status of method development (Extensive data base, +++; Limited data base,
++; Needs support, +)/St»tus of endpoint interpretation for ecological significance (Extensive data base, +++; Limited data base, ++;
Meeds support, +); "blank cell" = not applicable; "ns" = lack sufficient data base for decision.
21
-------
Method of
evaluation (see
Technical
Appendix for
methods synopses)
TA2.1Seed
germination and
root elongation
tests
TA2.2 Early
seedling survival
and vegetative
vigor tests
TA2.3 Laboratory
evaluations with
rooted aquatic
plants
TA2.4 Laboratory
evaluations with
wetland plants
TA2.5 Laboratory
evaluations with
upland plants
TA2.6 Alternative
test species for
plant testing
TA2.7 Short-term
tests for evaluating
whole plant
tosddty
TAZ8 Life-cycle
tests using vascular
plants
TA2.9 Plant tissue
culture tests
TA2.10Tests
evaluating plant
community
structure
Survival
+-H-/+*
+++/++
+++/++
+++/++
+++/++
+/++
++/+
•H-/+
++/+
++/++
Growth
+++/+
+++/++
+++/+
+++/+
+++/+
+/+
++/+
++/+
+/+
++/+
Repro-
duction
ns
ns
ns
ns
ns
ns
ns
++/+
ns
+/+
Develop-
mental
ns
+/+
ns
Biochem-
ical
+/+
+/+
+/+
ns
+/+
ns
Geno-
toxicity
ns
Immuno-
toxicity
Population
and
community
ns
ns
+/+
Each cell of the "test X endpoinf matrix evaluates: Status of method development (Extensive data base, +++; Limited data base,
•H-; Needs support, +)/Status of endpoint interpretation for ecological significance (Extensive data base, +++; Limited data base, ++;
Needs support, +); "blank cell" = not applicable; "ns" = lack sufficient data base for decision.
22
-------
Method of
evaluation (see
Technical
Appendix for
methods synopses)
TA2.11 Ambient
air exposure
systems and
phytotoxicity
testing
TA2.12
Photosynthesis
inhibition tests
TA3.1 Evaluation
of contaminant
effects on soil
bacterial biomass
TA3.2 Evaluation
of contaminant
effects on soil
fungal biomass
TA3.3 Evaluation
of contaminant
effects on soil
protozoa
TA3.4 Evaluation
of contaminant
effects on
nematode diversity
TA3.5Soil
microbial biomass
(chloroform
fumigation)
TA3.6 Test with
solid-phase/
aqueous phase
Microtox
TA3.7 Alternative
microbial toxicity
tests
Survival
rf
*
„,„
XA/AA
TT/TT
«*.
«**
(solid)
+-H-/+
(aq)
^
Growth
«,
*
«*
XA/A
TT/T
XA/A
i T7 i
++/++
ns
ns
ns
Repro-
duction
*
ns
ns
ns
ns
ns
ns
ns
Develop-
mental
Biochem-
ical
ns
*
*•
*
*
«„
«*
«*
Geno-
toxicity
ns
ns
ns
ns
ns
ns
«*
Immuno-
toxicity
Population
and
community
*
*
*
*
*
«'»
Each cell of the "test X endpoinf matrix evaluates: Status of method development (Extensive data base, +++; Limited data base,
++; Needs support, +)/Status of endpoint interpretation for ecological significance (Extensive data base, +++; Limited data base, ++;
Needs support, +); "blank cell" = not applicable; "ns" = lack sufficient data base for decision.
23
-------
Method of
evaluation (see
Technical
Appendix for
methods synopses)
TA3.8 Static,
batch, and
continuous-flow
rniorobUl tests
TA3.9 Soil-core
microcosm tests
TA3.10 Methods to
evaluate contamin-
ant effects on
microbial growth
TA3.11 Methods to
evaluate
contaminant effects
on soU enzymes
TA3.12Methodsto
evaluate
contaminant effects
onsoillipid
chemlsby
TA3.13 Substrate
uptake «nd
decomposition of
organic
compounds
TA3.HSoil
nutrient dynamics
tests: Nitrogen
cycling
TA4.1On-site
earth-worm test
methods
TAd In situ tests
using amphibians
TA4.3 Nest-box
tests using
starlings
TA4.*ln situ
dastogenicity tests
Survival
•H-/+*
+•»•/+
+++/++
ns
ns
+++/++
++/+
++/++
++/++
++/++
ns
Growth
++/++
++/+
•H4/H-
ns
ns
++/+
++/++
++/++
++/-H-
ns
Repro-
duction
+++/++
ns
ns
ns
ns
++/++
ns
Develop-
mental
+/+
ns
+/+
Biochem-
ical
++/+
++/++
++/+
+++/++
++/++
ns
ns
ns
+/+
Geno-
toxicity
ns
ns
++/++
Immuno-
toxicity
ns
ns
Population
and
community
+/+
+/+
++/++
++/++
++/+
++/+
++/++
++/++
++/++
++/++
++/++
Each cell of the "test X endpoint" matrix evaluates: Status of method development (Extensive data base, +++; Limited data base,
•H-; Needs support, +)/SUtu* of endpoint interpretation for ecological significance (Extensive data base, +++; Limited data base, ++;
Needs support, +); "blank cell* - not applicable; "ns" = lack sufficient data base for decision.
24
-------
Method of
evaluation (see
Technical
Appendix for
methods synopses)
TA4.5 Field testing
with sago
pondweed
TA4.6 On-site and
in situ testing with
terrestrial plants
TA4.7 Distribution
of microorganisms
TA4.8
Decomposition
rates
TA4.9Mycorrhizal
colonization
TA4.10 Lichens as
biomonitors
Survival
»
++/+
++/+
+++,++
++/+
ns
Growth
++/++
++/+
++/+
+++/++
++/+
++/+
Repro-
duction
ns
ns
«*
' ++/+
Develop-
mental
++/+
Biochem-
ical
++/+
++/++
+-)•/++
ns
Geno-
toxicity
Immuno-
toxicity
Population
and
community
»,»
«»
+•»•/++
++/-H-
+++/++
++/++
Each cell of the "test X endpoint" matrix evaluates: Status of method development (Extensive data base, +++; Limited data base,
++; Needs support, +)/Statu» of endpoint interpretation for ecological significance (Extensive data base, +++; Limited data base, ++;
Needs support, +); "blank cell" = not applicable; "ns" = lack sufficient data base for decision.
25
-------
Table 4: Habitats frequently impacted by hazardous waste sites and the status of test
methods included in the compendium as indicators of ecological effects.
Method of
evaluation (see
Technical
Appendix for
methods synopses)
TA1.1 Earthworm
survival test
TAl-ZSuWethal
effects tests using
earthworms
TAI.3 Other sofl
annelids
TA1.4 Tests using
free-Hying
nematodes
TA1.5 Tests using
soil arthropods
(Injects)
TA1.6 Terrestrial
arthropod (non-
insect) and isopod
tests
TA1.7MoQusk
tests
TA1.8 Amphibian
tests
TA1.9SmaH
mammal* and fur-
bearers
TA1.10 Avian tests
TA1.11 Immuno-
toxkity tests using
vertebrates
TAl.lZImmuno-
toaddty tests using
invertebrates
Conifer
Forest
•H-/++*
+/+
ns
++/+
++/+
•H/+
+/+
ns
•K/+
++/+
ns
ns
Hard-
wood
Forest
++/++
+/+
ns
++/+
++/+
++/+
+/+
ns
++/+
++/+
ns
ns
Grassland
++/++
+/+
ns
++/+
++/+
++/+
ns
+/+
++/+
++/+
ns
ns
Desert
ns
++/+
++/+
++/+
ns
++/+
++/+
ns
Inland
Wetland
ns
++/+
++/+
++/+
++/+
++/+
++/+
ns
Estuarine
Wetland
ns
++/+
++/+
++/+
ns
+H-/+
++/+
ns
Agri-
cultural
Lands
•H-f/-H-
+/+
ns
++/+
++/+
++/+
ns
+/+
++/+
++/+
ns
ns
Tundra
•H-/+
++/+
ns
++/+
++/+
ns
Taiga
++/+
++/+
ns
•H-/+
++/+
ns
* Bach cefl of the "test X endpoinf matrix evaluates: Status of method development (Extensive data base, +++; Limited data base,
++; Needs support, +)/Stmtu» of endpoint interpretation foe ecological significance (Extensive data base, +++; Limited data base, ++;
Needs support, +); "blank ceH" » not applicable; "ns" = lack sufficient data base for decision.
26
-------
Method of
evaluation (see
Technical
Appendix for
methods synopses)
TA2.1 Seed
germination and
root elongation
tests
TA2.2 Early
seedling survival
and vegetative .
vigor tests
TA2.3 Laboratory
evaluations with
rooted aquatic
plants
TA2.4 Laboratory
evaluations with
wetland plants
TA2.5 Laboratory
evaluations with
upland plants
TA2.6 Alternative
test species for
plant testing
TA2.7 Short-term
tests for evaluating
whole plant
toxicity
TA2.8 Life-cycle
tests using vascular
plants
TA2.9 Plant tissue
culture tests
TA2.10 Tests
evaluating plant
community
structure
Conifer
Forest
++/++*
++/++
++/+
ns
+/+
ns
ns
ns
Hard-
wood
Forest
++/++
++/++
++/+
ns
+/+
ns
ns
ns
Grassland
++/++
++/++
++/+
ns
+/+
ns
ns
++/+
Desert
ns
ns
++/+
+/+
ns
ns
Inland
Wetland
ns
ns
++/+
++/+
ns
Estuarine
Wetland
ns
ns
++/+
++/+
ns
Agri-
cultural
Lands
++/++
++/++
++/+
+/+
+/+
++/+
ns
++/+
Tundra
+/+
ns
Taiga
+/+
ns
Each cell of the "test X endpoint" matrix evaluates: Status of method development (Extensive data base, +++; Limited data base,
++; Meeds support, +)/Statu» of endpoint interpretation for ecological significance (Extensive data base, +++; Limited data base, ++;
Needs support, +); "blank cell" = not applicable; "ns" = lack sufficient data base for decision.
27
-------
Method of
' evaluation (see
Technical
Appendix for
methods synopses)
TA2.11 Ambient
air exposure
systems and
phytotoxkity
testing
TA2.12
Photosynthesis
inhibition tests
TA3.1 Evaluation
of contaminant
effects on soil
bacterialbiomass
TA3.2 Evaluation
of contaminant
effects on soil
fungal bkunass
TA3.3 Evaluation
of contaminant
effects on son
protozoa
TA3.4 Evaluation
of contaminant
effects on
nematode diversity
TA3JSou
mkrobUlbiomass
(chloroform
fumigation)
TA3.6 Test with
solid-phase/
aqueous phase
Mteotox
TA3.7 Alternative
mkrobUltoxicity
tests
TA3.8 Static,
batch, and
continuous-flow
mkrobUl tests
Conifer
Forest
++/W
ns
++/+
J.J./4.
TT/T
^
«»
+/ns
+/ns
+/ns
Hard-
wood
Forest
++/++
ns
++/+
++/+
«*
«.
+/ns
+/ns
+/ns
Grassland
++/++
ns
++/++
*
++•/+
«*
j. X/4.
TT/T
+/hs
+/ns
+/ns
Desert
*
*
*
t/+.
*
+/ns
+/ns
+/ns
Inland
Wetland
*
*
«*
*
+/ns
+/ns
+/ns
Estuarine
Wetland
*
ns
-
+++/+
+/ns
+/ns
+/ns
Agri-
cultural
Lands
it /it
TT/TT
.*
«„
XJ>/4-
TT/ T
-
++/+
-
+/ns
+/ns
+/ns
Tundra
*
«*
++/+
++/+
Taiga
*
-
«*
«*
* Bach cell of the "test X endpoint" matrix evaluates: Status of method development (Extensive data base, +++; Limited data base,
++; Needs support, +)/Status of endpoint interpretation for ecological significance (Extensive data base, +++; Limited data base, ++;
Needs support, +); "Wank cell" = not applicable; "ns" = lack sufficient data base for decision.
28
-------
Method of
evaluation (see
Technical
Appendix for
methods synopses)
TA3.9 Soil-core
microcosm tests
TA3.10 Methods to
evaluate contamin-
ant effects on
microbial growth
TA3.11 Methods to
evaluate
contaminant effects
on soil enzymes
TA3.12 Methods to
evaluate
contaminant effects
on soil lipid
chemistry
TA3.13 Substrate
uptake and
decomposition of
organic
compounds
TA3.14Soil
nutrient dynamics
tests: Nitrogen
cycling
TA4.1 On-site
earth-worm test
methods
TA4.2 In situ tests
using amphibians
TA4.3 Nest-box
tests using
starlings
TA4.4In_sjh!
clastogenicity tests
TA4.5 Field testing
with sago
pondweed
Conifer
Forest
+//
++/+
+/ns
ns
++/ns
++/ns
+/+
ns
Hard-
wood
Forest
ns
+/+
+/ns
+/ns
++/ns
+/ns
+/+
ns
Grassland
++/+
+/+
++/+
+/ns
++/ns
++/+
++/+
+/+
ns
Desert
ns
+/+.
+/ns
++/ns
++/hs
ns
Inland
Wetland
++/+
+/ns
++/ns
++/+
+/+
ns
+/ns
Estuarine
Wetland
++/+
++/ns
++/+
ns
Agri-
cultural
Lands
++/+
+/+
++/+
+/ns
++/ns
++/+
++/+
+/+
ns
Tundra
H-/-H
+/ns
+/ns
•f/ns
ns
Taiga
+/+
+/ns
+/ns
+/ns
ns
Each cell of the "test X endpoinf matrix evaluates: Status of method development (Extensive data base, +++; Limited data base,
++; Needs support, +)/Statu» of endpoint interpretation for ecological significance (Extensive data base, +++; Limited data base, ++;
Needs support, +); "blank cell" = not applicable; "ns" = lack sufficient data base for decision.
29
-------
Method of
evaluation (see
Technical
Appendix for
methods synopses)
TA4.6On-slteand
in situ testing with
tenestrialplants
TA4.7 Distribution
of resistant
microorganisms
TA4.8
Decomposition
rates
TA4.9Mycorrhizal
colonization
TA4.10 Lichens as
btomonitors
Conifer
Forest
+/ns
•H-+/+
++/+
ns
Hard-
wood
Forest
ns
++/+
++/+
++/++
Grassland
+/ns*
++/ns
++/+
++/+
+/+
Desert
+/ns
+/ns
+/+
+/+
+/+
Inland
Wetland
+/ns
+/+
+/ns
Estuarine
Wetland
+/ns
+/+
+/ns
Agri-
cultural
Lands
+/+
•H-Ms
H-+/++
++/+
+/+
Tundra
+/ns
+/+
++/+
ns
Taiga
ns
+/+
+/+
.ns
Each cell of the "test X endpoinf matrix evaluates: Status of method development (Extensive data base, +++; Limited data base,
++; Needs support, +V&Utu» of endpoint interpretation for ecological tignificance (Extensive data base, +++; Limited data base, ++;
Needs support, +); "blank cell" = not applicable; "ns" = lack sufficient data base for decision.
30
-------
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on forests. In Biologic markers of air-pollution stress and damage in forests. Committee on Biologic
Markers of Air-Pollution Damage in Trees, G.M. Woodwell (Chair). National Research Council. National
Academy Press. Washington, D.C. Pp. 303-316.
Witkamp, M. and van der Drift, J. (1961) Breakdown of forest litter in relation to environmentalfactors.
Plant and Soil 15:295-311.
Wong, M. H., and J. W. C. Wong. 1986. Effects of fly ash on soil microbial activity. Environ. Pollut. Ser. A.
40:127-144.
Woods, L.E., C.V. Cole, E.T. Elliott, R.V. Anderson and D.C. Coleman. 1982. Nitrogen transformationsin
soil as affected by bacterial-microfaunafinteractions. Soil Biol. Biochem. 14:93-98.
Zilkah, S., P.P. Bocion, and J. Gressel. 1977. Cell cultures vs. whole plants for measuring phytotoxicity. n.
Correlations between phytotoxicityin seedlings and calli. Plant & Cell Physiol. 18:657-670.
Zilkah, S. and J. Gressel. 1977b. Cellculturesvs.wholeplantsformeasuringphytotoxicity.nl.
Correlations between phytotoxidtiesin cell suspension cultures, calli, and seedlings. Plant & Cell
Physiol. 18:815-820.
Zilkah, S. and J. Gressel. 1978. Correlations in phytotoxicitybetween white and green calli of Runtex
obtusifoUus,Nicotianatabacum,andLycopersiconesculentum. Pesticide Biochem. Physiol. 9:334-339.
Zilkah, S. and J. Gressel. 1977a. Cell cultures vs. whole plants for measuring phytotoxicity. I. The
establishmentand growth of callus and suspension cultures; definition of factors affecting toxicity on calli.
Plant & Cell Physiol. 18:641-655.
Zoran, M.J., T.J. Heppner and C.D. Drewes. 1986. Teratogenic effects of the fungicide benomyl on
posterior segmental regeneration in the earthworm, Eisenia foetida. Pestic. Sci. 17:641-652.
Zwick, T.C., M.F. Arthur, D.A. Tolle, and P. Van Voris. 1984. A unique laboratory method for monitoring
agro-ecosystemeffects of an industrial waste product. Plant Soil 77:395-399.
63
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TECHNICAL APPENDICES
TA1. ANIMAL TEST METHODS FOR THE ASSESSMENT OF SOIL
CONTAMINATION AT HAZARDOUS WASTE SITES
TA2. PLANT TEST METHODS FOR THE ASSESSMENT OF SOIL
CONTAMINATION AT HAZARDOUS WASTE SITES
TA3. SOIL BIOTA TEST METHODS FOR THE ASSESSMENT OF SOIL
CONTAMINATION AT HAZARDOUS WASTE SITES
TA4. FIELD METHODS FOR THE ASSESSMENT OF SOIL CONTAMINATION
AT HAZARDOUS WASTE SITES
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DESCRIPTION OF MEASUREMENT AND TOXICITY ASSESSMENT
TECHNIQUES.
While numerous methods are available for measuring the toxicity associated with soil-
derived eluates, sediment elutriates, groundwater, surface waters and waste waters
(e.g., Weber et al. 1988; see Section 3.0), few standardized methods are currently
available which measure the toxicity of soils directly. Ehition methods similar to those
applied in sediment toxicity evaluations have been applied to soils and have yielded
soil toxicity estimations derived from aquatic tests (US EPA 1989). Ecological
interpretation of these indirect tests may be difficult, however, depending upon the
exposure pathways that are considered most relevant to receptors identified as part of
the problem formulation and exposure assessment phases of the ecological assessment
(Figure 2). Consequently, methods which directly evaluate soil toxicity by testing soil
macroinvertebrates(like earthworms) and plants have been developed and applied in
site assessment (US EPA 1989). Relative to numerous methods for testing aquatic
matrices, however, fewer standardized methods exist which assess soil toxicity directly
(e.g., ASTM 1991; US EPA 1989; Hicks and Van Voris 1988). The restricted number of
test methods becomes even more critical when ecological assessment of terrestrial
hazardous waste sites are considered (US EPA 1989), since the issues related to
contaminated soil frequently reoccur during the assessment process for CERCLA or
RCRA (Resource Conservation and Recovery Act) sites.
While only a few methods which evaluate bulk soils directly are widely recognized,
there are numerous methods that should also be considered in conjunction with these
more well established test methods. The following methods represent those that may
be useful in evaluating contaminated soils at hazardous waste sites, and include
additional aquatic test methods that are well developed and potentially beneficial to
ecological effects assessments.
References.
ASTM. 1991. Annual book of ASTM standards. Volume 11.04. American Society for
Testing and Materials (ASTM). Philadelphia, PA. 19103.
Hicks, R.J., and P. Van Voris. 1988. Review and evaluation of the effects of
xenobiotic chemicals on microorganisms in soil. Report 6186. Pacific Northwest
Laboratory, U.S. Department of Energy, Battelle Memorial Institute, Richland, WA.
TA-1
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US EPA. 1989. Protocols for short term toxirity screening of hazardous waste sites.
J.C. Greene, C.L. Bartels, W.J. Warren-Hicks, B.R. Parkhurst, G.L. Linder, S. A.
Peterson, and W.E. Miller (Eds.). EPA/600/3-88/029, U.S. Environmental Protection
Agency, Environmental Research Laboratory-Corvallis,OR.
Weber, C.L, W.B. Horning, D.J. Klemm, T.W. Neiheisel, P.A. Lewis, E.L. Robinson,
J. Menkedick, and F. Kessler. 1988. Short-term methods for estimating the chronic
toxirity of effluents and receiving waters to marine and estuarine organisms.
EPA/600/4-87/028. US EPA, Environmental Monitoring and Support Laboratory,
Cincinnati, OH.
TA-2
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TA1. ANIMAL TEST METHODS FOR THE ASSESSMENT OF SOIL
CONTAMINATION AT HAZARDOUS WASTE SITES
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TAl.l. Earthworm survival tests.
As a widespread and readily available terrestrial macroinvertebrate, earthworms have
become a primary test organism for soil contamination evaluations. From an
ecological perspective, earthworms are significant in improving soil aeration, drainage,
and fertility (Edwards and Lofty 1972), although the comparative data base does not
unequivocally suggest that earthworm toxicity measurements are reflective of soil
"health/' To enhance the ecological relevance of site-specific biological tests and to
reduce the potential extrapolation error associated with interspecies comparisons,
testing with site-specific species should be considered in soil evaluations.
The earthworm bioassay most frequently used is a modification of a method described
by Goats and Edwards (1982) and Edwards (1984), and uses lumbricoid earthworms as
the test species. Eisenia foetida may be used in these tests since it is easily cultured in
the laboratory, and reaches maturity in seven to eight weeks at 25°C. £. foetida is
responsive to a wide range of toxicants, and the comparative data base suggests that
similar toxicity responses can be anticipated regardless of the subspecies being tested
(Neuhauser, et al. 1986).
Test method summary: Earthworm (Eisenia foetida) 14-day survival. In the
laboratory, the earthworm test method directly evaluates the biological effects of soils
on a representative soil macroinvertebrate. In 14-day screening tests, percent survival
is recorded at Day 7 and Day 14 for evaluations of biological effects associated with
exposures to bulk soil samples. Similarly, mortality is the most frequently measured
endpoint in definitive tests, though behavioral and pathological observations may also
be recorded (see TA1.2) In definitive tests, median lethal concentrations (LCgoS) with
their corresponding 95% confidence intervals are calculated at Day 7 and Day 14.
Positive control (e.g., 2-chloroacetamide)LC50s for both seven and 14 days using, for
example, the trimmed Spearman-Karber method should be completed for definitive
tests (Hamilton, et al. 1977).
Intended use: Earthworm tests were designed to measure the toxicity of soils that
were contaminated, or were potentially impacted, by chemicals released to the
environment, as exemplified by the regulatory concerns related to CERCLA/RCRA
(Resource Conservation and Recovery Act) and FIFRA (Federal Insecticide, Fungicide,
and Rodenticide Act)/TSCA (Toxic Substances Control Act), respectively. Earthworm
test methods have been outlined with various levels of detail by US EPA (1989), Food
and Drug Administration (FDA; 1987), and OECD (1984).
TAl.l - 1
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Previous applications/regulatoryprecedence: Of the bulk soil tests outlined in this
document, those completed with earthworms or plants represent the majority of
biological measurements used in soil contamination evaluations. As an ecoindicator
for terrestrial hazardous waste sites/ earthworm testing has been applied at sites with
a wide variety of chemical contaminants.
Requirements for development and implementation: Requirements for routinely
using earthworm tests are outlined in US EPA (1989), as well as the applied ecology
literature (e.g., Neuhauser, et al. 1986; Callahan, et al. 1985). The 14-day survival test
using Lumbricus terrestris or Eisenia foetida is commercially available, and costs for
completing the test on a per sample basis range from $1000 to $2000.
Potential problems and limitations: Unlike aquatic toxicity tests, the majority of bulk
soil tests are not commercially available. Of those tests summarized throughout this
compilation, earthworm tests are among the most readily available. From a technical
perspective, earthworm testing may also be problematic with respect to a soil
contamination evaluation, if the soil is inappropriate for the earthworm species being
tested. For example, certain soils may be incompatible with respect to minimal test
requirements (e.g., strongly acidic, wetland soils; nutrient-poor and highly alkaline
soils as seen in desert soil samples), and preclude a successful test regardless of the
contaminant status of the sample. The selection of test species and the
appropriateness of test species must be given ample consideration during the problem
formulation phases of the ecological assessment. Misinterpretation of "false negatives"
that may result from selection of an inappropriate test species can be minimized, if
adequate characterization of the soil matrix is completed within the ecological effects
assessment (see Section 4.0).
References.
Callahan, C. A., L.K. Russell and S. A. Peterson. 1985. A comparison of three
earthworm bioassay procedures for the assessment of environmental samples
containing hazardous wastes. Biol. Fert. Soils. 1:195-200.
Edwards, C.A., and J.R. Lofty. 1972. Biology of earthworms. Chapman and Hall,
Ltd., London.
TA1.1 - 2
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Edwards, C.A. 1984. Report of the second stage in development of a standardized
laboratory method for assessing the toxicity of chemical substances to earthworms.
Commission of the European Communities. EUR 9360 EN. 98pp.
FDA. 1987. Environmental assessment technical assistance document. Section 4.12.
Earthworm subacute toxicity. Food and Drug Administration, Center for Food Safety
and Applied Nutrition, Environmental Impact Section and Center for Veterinary
Medicine Environmental Staff. Washington D.C.
Goats, G., and C.A. Edwards. 1982. Testing the toxicity of industrial chemicals to
earthworms. Rothamsted Exp. Station Report, 1982. pp. 104-105.
Hamilton, M.A., R.C. Russo, and R.V. Thurston. 1977. Trimmed Spearman-Karber
method for estimating median lethal concentrations in toxicity bioassays. Environ. Sci.
Tech.
11: 714-719.
Neuhauser, E.F., P.R. Durkin, M.R. Milligan, and M. Anatra. 1986. Comparative
toxicity of ten organic chemicals to four earthworm species. Comp. Biochem.
Physiol. 83C(1):197-200.
OECD (Organisation for Economic Co-Operation and Development). 1984. OECD
guidelines for testing of chemicals. Director of Information, OECD. 2, rue Andrfc
Pascal, 75775 Paris Cedex 16, France.
US EPA. 1989. Protocols for short term toxicity screening of hazardous waste sites.
J.C. Greene, C.L. Bartels, W.J. Warren-Hicks, B.R. Parkhurst, G.L. Under, S.A.
Peterson, and W.E. Miller (Eds.). EPA/600/3-88/029, U.S. Environmental Protection
Agency, Environmental Research Laboratory-Corvallis,OR.
TA1.1 - 3
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TA1.2. Sublethal effects tests using earthworms.
While bulk soil testing with earthworms is relatively well developed, endpoints other
than death are not routinely measured. Lethality, however, may not be the only
endpoint that can be readily measured in the test, and various workers have described
alternative endpoints that yield sublethal indicators of adverse effects resulting from
soil exposures. These sublethal measurements or observations are variously described
as behavioral or gross morphological endpoints, and may be qualitative or quantitative
in character. In the laboratory, methods routinely used in diagnostic pathology may
be applied to measuring these sublethal endpoints, and rank scores, for example, may
be used to quantify sublethal responses. Concentration response relationships, then,
may be used to describe these sublethal endpoints. These endpoints may also be
used in the field to evaluate sublethal indicators of in situ toxicity.
Test method summary: Various endpoints have been described for measuring
adverse biological effects in earthworms other than death, and frequently the sublethal
endpoints being used by a variety of workers are similar across soil types and
contaminants of concern. The 14-day survival test previously outlined (TA1.1) can
easily incorporate these sublethal endpoints. At the end of the test, the surviving
earthworms should be rinsed with water, examined closely, and observations
regarding sublethal endpoints should be recorded. In supplementing the survival
data collected from the standard earthworm test, the sublethal endpoints being
evaluated should be well characterized and a scoring system, if one is used, must be
fully developed before the test is initiated and measurements taken. The toxicity
endpoints measured at test termination, either acute or subacute and sublethal, are
best described for Eisenia spp. and Lumbricus spp., where the sublethal endpoints
most frequently are:
• weight loss — change in biomass (live weight) per test container (Ma, 1982;
Wentsel and Guelta, 1987; Hartensteinef al., 1981; Bouwman and Reinecke,
1987; Stenersen, 1979; Haque and Ebing, 1983);
• segmental swelling — morphological endpoint that may be used as a
localized or systemic sublethal measurement (Stenersen, 1979; Roberts and
Dorough, 1985; Drewes et al., 1988; Roberts and Dorough, 1984; Haque and
Ebing, 1983);
• lesions, ulcers, sores, blisters, bleeding sores, epithelial sores — all of these
terms have been used to describe dennopathologic responses (Stenersen, 1979;
TA1.2 - 1
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Roberts and Dorough, 1985; Drewes et al., 1988; Roberts and Dorough, 1984;
Haque and Ebing, 1983);
• coiling — term has been consistently used to describe postural endpoints
related to coordinated muscular responsiveness (Gilman and Vardanis, 1974;
Stenersen, 1979; Drewes et al., 1988; Roberts and Dorough, 1984; Haque and
Ebing, 1983);
• rigidity and shortening, and contracted and shortening —these terms have
been used to describe postural endpoints related to muscular responsiveness
(Stenersen, 1979; Roberts and Dorough, 1985; Drewes et al., 1988; Roberts and
Dorough, 1984; Haque and Ebing, 1983; Drewes et al., 1984);
• extended and limp with lack of muscle tone, immobility —these terms are
frequently used to describe similar endpoints and may be grouped with
pathologic terms such as flaccid and/or elongated to express the similar gross
morphological condition (Roberts and Dorough, 1984; Stenersen, 1979);
• segmental constrictions and body constrictions — localized or more
systemic indicators of muscular responsiveness (Gilman and Vardanis, 1974;
Roberts and Dorough, 1984);
• self-autotomy — loss of tail region upon soil exposure (Zoran et al., 1986);
* giant fiber nerve conduction — sublethal measurement related to altered
neurophysiologicalactivity (Drewes et al., 1988; Drewes et al., 1987; Zoran et al.,
1986; Drewes and Vining, 1984; Drewes et al., 1984; Callahane* al., 1985);
• non-burrowing and rate of burrowing — the lack of burrowing in the
presence of continuous lighting; earthworms exhibit negative phototaxis under
continuous lighting (e.g., Stenersen, 1979);
• surface casts — the absence of castings on the surface of the test soil may be
used as a sign of stress to the worms, since casts are usually present in the
control (Stenersen, 1979);
• development of sexual characteristics such as clitellum and tuberculae
pubertales — growth and maturation related endpoint (Bouwman and
Reinecke, 1987; Venter and Reinecke, 1985; Lofs-Holman, 1980);
TA1.2 - 2
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• cocoon production — endpoints related to reproductive success may be
measured; however, the test should be lengthened to at least 28 days (Ma, 1984;
Neuhauser et a/., 1984; Malecki et al., 1982; Bouwman and Reinecke, 1987;
Reinecke and Venter, 1985; Haque and Ebing, 1983; Lofs-Hohnin, 1982; van
Rhee, 1977);
• number of cocoons that hatch and number of hatchlings per cocoon; size
of hatchlings; cocoon mass; shape of cocoons -another sublethal endpoint
that may be applicable for evaluating reproductive effects; however, the test
should be lengthened to at least 28 days (Reinecke and Venter, 1985; Bouwman
and Reinecke, 1987).
Other anomalies have been noted and could also be scored as sublethal endpoints,
e.g., the appearance or absence of swollen ditella. Additionally, the location of the
various dermopathologic lesions, e.g., "segmental constriction just anterior to the
clitellum" may be recorded, and the site of the lesion or pre-pathologic tissue may be
used to quantify the measurement endpoint.
Intended use: Ideally, subacute and sublethal endpoints should be recorded in any
controlled laboratory or in situ test regardless of the test species, since mortality data
may not be as valuable as subacute and sublethal data when interpretation of
ecological effects is considered beyond local extirpation. As such, the organismic-level
toxirity endpoints potentially gathered in earthworm tests should be considered as
supportive, or correlative, data that complement field survey information obtained on-
site.
Previous applications/regulatory precedence: Within regulatory contexts earthworm
testing has a relatively long history in biological assessment, and for Superfund
ecological effects assessments, earthworm testing has been consistently applied across
a range of terrestrial habitats. Integrated studies that rely on chemical as well as
laboratory and in situ toxitity evaluations (Callahan, et al. 1990) suggest that ecological
effects assessments with earthworms can help address site-specific issues related to
bioavailabilityof contaminants.
Requirements for development and implementation: While earthworm testing is
commercially available, few, if any, laboratories are presently completing the test with
readings of subacute endpoints. Growth endpoints should be relatively easy to
implement, but the behavioral and pathological endpoints require special training and
experience which must be developed on a wider scale before the subacute and
sublethal endpoints in the earthworm test are routinely available.
TA1.2 - 3
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Potential problems and limitations: The life history attributes of the routinely tested
species (Lumbricus terrestris and Eisenia foetida) allow soil testing at a variety of waste
site habitats. However, some regions of the country and some site-specific soil
conditions may preclude a successful, and interpretable, test. While numerous
laboratories provide technical services for earthworm testing, implementing this
biological assessment within an ecological effects assessment for Superfund may be
difficult owing to an absence of experienced technical support, particularly in data
interpretation. During the problem formulation phase in the ecological effects
assessment, a center question that must be addressed is the applicability of earthworm
testing, given the baseline soil conditions (e.g., acidity, alkalinity, and moisture
fraction).
References.
Bouwman, H. and A.J. Reinecke. 1987. Effects of carbofuran on the earthworm,
Eisenia foetida, using a defined medium. Bull. Environ. Contam. Toxicol. 38:171-178.
Callahan, C.A., L.K. Russell and S.A. Peterson. 1985. A comparison of three
earthworm bioassay procedures for the assessment of environmental samples
containing hazardous wastes. Biol. Pert. Soils. 1:195-200.
Callahan, C.A., C.A. Menzie, D.E. Burmaster, D.C. Wilborn, and T. Ernst. 1991.
On-site methods for assessing chemical impact on the soil environment using
earthworms: a case study at the Baird and McGuire Superfund site, Holbrook,
Massachusetts. Environ. Toxicol. Chem. 10:817-826.
Drewes, C.D., M.J. Zoran and C.A. Callahan. 1987. Sublethal neurotoxic effects of
the fungicide benomyl on earthworms (Eisenia foetida). Pestic. Sci. 19:197-208.
Drewes, C.D., E.P. Vining and C.A. Callahan. 1984. Non-invasive
electrophysiologicalmonitoring: a sensitive method for detecting sublethal
neurotoxicityin earthworms. Environ. Toxicol. Chem. 3:599-607.
Drewes, C.D., E.P. Vining and C. A. Callahan. 1988. Electrophysiologicaldetectionof
sublethal neurotoxic effects in intact earthworms. In Earthworms In Waste And
Environmental Management, C.A. Edwards and E.F. Neuhauser (eds.), SPB Academic
Publishing bv, The Hague, The Netherlands, pp. 355-366.
TA1.2 - 4
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Drewes, C.D. and E.P. Vining. 1984. In vivo neurotoxic effects of dieldrin on giant
nerve fibers and escape reflex function in the earthworm, Eisenia foetida. Pestic.
Biochem. and Physiology. 22:93-103.
Gilman, A.P. and A. Vardanis. 1974. Carbofuran. Comparative toxicity and
metabolism in the worms Lumbricus terrestris L. and Eisenia foetida S. J. Agr. Food
Chem. 22(4):625-628.
Haque, A., and W. Ebing. 1983. Toxicity determination of pesticides to earthworms
in the soil substrate. J. Plant Diseases and Protection. 90(4):395-408.
Hartenstein, R., E.F. Neuhauser and A. Narahara. 1981. Effects of heavy metal and
other elemental additives to activated sludge on growth of Eisenia foetida. J. Environ.
Qual. 10(3):372-376.
Lofs-Holmin, A. 1980. Measuring growth of earthworms as a method of testing
sublethal toxicity of pesticides. Swedish J. Agric. Res. 10:25-33.
Lofs-Holmin, A. 1982. Measuring cocoon production of the earthworm Allolobophora
caliginosa (Sav.) as a method of testing sublethal toxicity of pesticides. An experiment
with benomyl. Swedish J. Agric. Res. 12:117-119.
Ma, W. 1982. The influence of soil properties and worm-related factors on the
concentration of heavy metals in earthworms. Pedobiologia. 24:109-119.
Ma, W. 1984. Sublethal toxic effects of copper on growth, reproduction and litter
breakdown activity in the earthworm Lumbricus rubellus, with observations on the
influence of temperature and soil pH. Environ. Pollution (Series A). 33:207-219.
Malecki, M.R., E.F. Neuhauser and R.C. Loehr. 1982. The effect of metals on the
growth and reproduction of Eisenia foetida (Oligochaeta, Lumbricidae). Pedobiologia.
24:129-137.
Neuhauser, E.F., M.R. Malecki and R.C. Loehr. 1984. Growth and reproduction of
the earthworm Eisenia foetida after exposure to sublethal concentrations of metals.
Pedobiologia. 27:89-97.
Reinecke, A.J. and J.M. Venter. 1985. Influence of dieldrin on the reproduction of
the earthworm Eisenia foetida (Oligochaeta). Biol. Pert. Soils. 1:39-44.
TA1.2 - 5
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Roberts, B.L. and H.W. Dorough. 1985. Hazards of chemicals to earthworms.
Environ. Toxicol. Chem. 4:307-323.
Roberts/ B.L. and H.W. Dorough. 1984. Relative toxirities of chemicals to the
earthworm Eisenia foetida. Environ. Toxicol. Chem. 3:67-78.
Stenersen,J. 1979. Action of Pesticides on Earthworms. Parti: the toxicityof
cholinesterase-inhibitinginsecticides to earthworms as evaluated by laboratory tests.
Pestic.Sci. 10:66-74.
vanRhee, J.A. 1977. Effects of soil pollution on earthworms. Pedobiologia.
17(S):201-208.
Venter, J.M. and A.J. Reinecke. 1985. Dieldrin and growth and development of the
earthworm, Eisenia fetida (Oligochaeta). Bull. Environ. Contam. Toxicol. 35:652-659.
Wentsel, R.S. and M.A. Guelta. 1987. Toxicity of brass powder in soil to the
earthworm Lumbricus terrestris. Environ. Toxicol. Chem. 6:741-745.
Zoran, M.J., T.J. Heppner and C.D. Drewes. 1986. Teratogenic effects of the
fungicide benomyl on posterior segmental regeneration in the earthworm, Eisenia
foetida. Pestic. Sci. 17:641-652.
TA1.2 - 6
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TA1.3. Other tests with soil annelids.
While methods are relatively well developed for testing earthworms, particularly
Eisenia foetida, alternative test species should be considered when the soil matrix
precludes, or potentially confounds the interpretation of biological effects owing to the
test animal's life history attributes. For example, other soil annelids should be
considered as alternative test species when a test soil's physicochemical properties
(e.g., moisture fraction, texture) are not conducive to a successful test with E. foetida.
A soil test has been developed (RSmbke 1989) that uses Enchytraeus albidus
(Enchytraeidae, Oligochaeta) in a bulk soils test. Mortality and biomass are generally
measured during a four week exposure, and if exposure periods are sufficient (eight
weeks), reproduction can be evaluated by counting offspring. While the laboratory
test has had limited application, the method using E. albidus has proven beneficial to
evaluating contaminated soil, since complementary field survey methods have been
used previously for monitoring anthropogenic chemicals deposited in terrestrial
systems (e.g., Bengtsson and Rundgren 1982).
Test methods summary: Technically, the test with E. albidus is very similar to that
with Lumbricus spp. or Eisenia spp. Rombke (1989), for example, suggests that
contaminant evaluations completed with E. albidus could be conducted as acute tests
(mortality and biomass measured weekly for four weeks), as well as longer-term
reproductive tests (four week extended studies following initial four-week tests for
measuring offspring production). Technically, the test uses exposures similar to those
used in the standard test with E. foetida (US EPA 1989; FDA 1987).
Intended use: Certain soils may be more adequately tested with an alternative
annelid; Enchytraeus albidus and other members of the family may be considered more
ecologically relevant, and are more widely distributed in temperate zone soils than the
routine test species, Eisenia foetida. Similar to the tests using Lumbricus terrestris or E.
foetida, tests using alternative earthworm species were designed to measure the toxicity
of soils that were contaminated, or were potentially impacted, by chemicals released to
the environment, as exemplified by the regulatory concerns related to CERCLA/RCRA
(Resource Conservation and Recovery Act) and FIFRA (Federal Insecticide, Fungicide,
and Rodentitide Act)/TSCA (Toxic Substances Control Act), respectively. Earthworm
test methods with alternative test species have not previously been outlined by US
EPA (1989), but their application to soil contamination evaluations should be
considered nonetheless, particularly within a comparative toxicity assessment.
Previous applications/regulatoryprecedence: Tests with E. albidus or other
enchytraeids have not been completed within ecological effects assessments for
TA1.3 -1
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Superfund, but are potentially amenable to routine testing owing to the technical
similarities to the standard earthworm test.
Requirements for development and implementation: Tests with alternative annelids
are not available in the United States, and implementing this biological assessment
within an ecological effects assessment for Superfund may be difficult owing to an
absence of experienced testing services.
Potential problems and limitations: While the potential strengths associated with
tests completed with E. albidus are numerous (e.g., more ecologically relevant,
amenable to laboratory and field assessment), the lack of commercial availability may
limit the tesf s routine application in ecological effects assessment.
References.
Bengtsson, G., and S. Rundgren. 1982. Population density and species number of
enchytraeids in coniferous forest soils polluted by a brass mill. Pedobiologia 24:211-
218.
FDA. 1987. Environmental assessment technical assistance document. Section 4.12.
Earthworm subacute toxicity. Food and Drug Administration, Center for Food Safety
and Applied Nutrition, Environmental Impact Section and Center for Veterinary
Medicine Environmental Staff. Washington D.C.
R8mbke, J. 1989. Enchytraeus albidus (Enchytreidae, Oligochaeta) as a test organisms
in terrestrial laboratory systems. Arch. Toxicol., Suppl. 13:402-405.
US EPA. 1989. Protocols for short term toxicity screening of hazardous waste sites.
J.C. Greene, C.L. Bartels, W.J. Warren-Hicks, B.R. Parkhurst, G.L. Under, S.A.
Peterson, and W.E. Miller (Eds.). EPA/600/3-88/029, U.S. Environmental Protection
Agency, Environmental Research Laboratory-Corvallis,OR.
TA1.3 - 2
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TA1.4. Tests using free-living nematodes.
Soti-irthabitingnematodes represent one of the most readily available soil invertebrates
that should be studied during soil contamination evaluations within an ecological
effects assessment. Owing to their usually high numbers, their role in soil
decomposition processes, and their significant contribution to soil nutrient dynamics
(e.g., dispersion and grazing on microflora, potential stimulation of bacterial activity/
and promotion of nutrient mineralization)/ soil nematodes directly as well as indirectly
reflect the "health'7 of the soil. The limited consideration of soil nematodes as
ecological indicators of soil integrity stems/ in part/ from a restricted toxicologicaldata
base for these invertebrates. From a laboratory perspective/ nematodes afford many
advantages for toxicity testing. Specifically/ nematodes/ if frozen in liquid nitrogen,
and subsequently preserved (-80°C), can be rehydrated prior to testing where they
characteristically have less than a four-day day generation time. Also, for short-term
culture, nematodes may be grown in liquid culture (Vanfletren 1976). Culture and
husbandry practices for rearing laboratory nematodes are well described (Dougherty
and Calhoun 1948), and reinforce the relative ease with which nematode evaluations
could be incorporated into the Superfund ecological assessment process. While
methods and applications have been developed for aquatic and sediment toxicity
assessment/ relatively few methods have been developed for using soil-dwelling
nematodes to evaluate adverse biological effects associated with contaminated soil. At
present/ two test systems are sufficiently developed for implementation and/ if
warranted/ should be considered early in the problem formulation phase of an
ecological effects assessment.
Test methods summary: Pamgrellus redivivus has a relatively well developed literature
in aquatic toxicity testing (Samoiloff/ et al. 1980)/ and has been used for evaluating
single-chemicals and complex chemical mixtures (Samoiloff, et al. 1983). Most
frequently, P. redivivus has been used in conjunction with other biological assessments
(e.g./ Daphnia magna or Ceriodaphnia dubia testing) for evaluations of water quality/ but
the test system has also been applied to sediment toxicity testing (Samoiloff/ et al.
1983). Work with P. redivivus has been well described in the comparative toxicity
literature, but another/ more recently developed nematode test using Caenorhabditis
elegans (e.g./ van Kessel/ et al. 1989; Williams and Dusenbery 1990) may be applicable
for ecological effects assessments. P. redivivus and C. elegans tests measure acute —
lethality— and subacute or sublethal effects related to growth/ reproduction, and
mutagenicity. Both methods are short-term tests and generally require less than four
to five days for completion, although long-term tests which measure reproductive
effects (e.g., number of offspring) may require seven-day exposures.
TA1.4 - 1
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Unlike P. redivivus, C. elegans is a native soil-dwelling nematode (Briggs 1946 as cited
by van Kessel, et al. 1989), and tests with this nematode may more closely reflect soil
contaminant effects in terrestrial habitats. Williams and Dusenbery (1990) studied the
toxic effects of metals in aqueous solutions using Caenorhabditis elegans, and in their
comparative analysis, C. elegans acute toxicities (LC^) for single-compound metal
exposures complemented and were consistent with acute toxicity results from Daphnia
magna and sediment macroinvertebrates. As suggested by various authors (e.g.,
Popham and Webster, 1979; Haight, et al. 1982; Doelman, et al. 1984; van Kessel, et al.
1989), for some toxicants like heavy metals, the existing toxicity data base for
nematodes is developed, and extending these methods to soils should be considered
within ecological effects assessments. For example, while the testing with either P.
redivius or C. elegans was originally developed for testing surface water or sediment
pore waters, nematode tests are directly applicable to evaluating soil extracts or
interstitial waters.
Intended use: As complementary tests to those using earthworms Lumbricus terrestris
or Eiseniafoetida, methods using free-living nematodes have been designed to evaluate
the toxicity of soils contaminated, or potentially impacted, by chemicals released to the
environment. Soil tests with nematodes have not previously been outlined by US
EPA but their application to soil contamination evaluations should be considered
nonetheless, particularly within a comparative toxicity assessment. Standardized
sampling strategies have been established (ASTM E629 1991); while developed as part
of efficacy evaluations for nematode control agents, the methods outlined in the
standard guide could easily serve the needs of ecological effects assessments for
Superfund.
Previous applications/regulatoryprecedence: P. redivius has been proposed as a
component within a test battery for evaluating, for example, complex environmental
contaminant mixtures (Samoiloff 1987). Within a comparative context C. elegans could
also be developed for testing soils as marine nematodes Monohystera microphtalama and
Diplolaimebruciei were for testing marine and estuarine waters (Samoiloff and Bogaert
1984). For soil contamination evaluations, however, relatively little work has been
completed to address questions related to "laboratory-to-field"extrapolation errors
associated with either species, and few applied studies (e.g., Bongers 1990) have been
published regarding the effects of contaminant mixtures on soil community structure.
Requirements for development and implementation: The technique for P. redivivus
was originally developed for aquatic testing and is directly applicable to soil extracts.
Few commercially available technical support laboratories offer testing services with
either P. redivivus or C. elegans.
TA1.4-2
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Potential problems and limitations: P. redivivus is an aquatic species and may be
directly applicable to groundwater evaluations, and would be applicable to indirect
tests of soil. While a free-living soil nematode, C. elegans is not found in all temperate
soils, and may not inhabit all soil types. Neither species may be applicable to all soils,
since survival and growth may not be optimal across all potential site soil conditions.
References.
ASTM E629. 1991. Standard guide for field evaluation of nematode control agents —
determination of nematode population responses to control agents. Annual book of
ASTM standards. Volume 11.04. Pesticides; Resource Recovery; Hazardous
Substances and Oil Spill Responses; Waste Disposal; Biological Effects. American
Society for Testing and Materials (ASTM). Philadelphia, PA. 19103.
Bongers, T. 1990. The maturity index: an ecological measure of environmental
disturbance based on nematode species composition. Oecologia 83:14-19.
Doelman, P., G. Nieborer, J. Schrooten, and M. Visser. 1984. Antagonistic and
synergistic toxic effects of Pb and Cd in a simple foodchairu nematodes feeding on
bacteria and fungi. Bull. Environ. Contam. Toxicol. 32:717-723.
Dougherty, E.G. and G.H. Calhoun. 1948. Experiences in culturing Rhabditis pellio
(Schneider 1866) Butschli, 1873 (Nematoda, Rhabditidae), and related soil nematodes.
Proc. Helminthol. Soc. Wash. 15:55-68.
Haight, M., T. Mudry, and J. Pasternak. 1982. Toxicity of seven heavy metals on
Panagrellus silusiae: the efficacy of the free-living nematode as an in vivo toxicological
assay. Nematologica28:l-ll.
Popham, J.D. and J.M. Webster. 1979. Cadmium toxicity in the free-living nematode,
Caenorhabditis elegans. Environ. Res. 20:183-191.
Samoiloff, M. 1987. Nematodes as indicators of toxic environmental contaminants.
In J.A. Veech and D.W. Dickson (Eds.). Vistas on nematology. E.O. Painter Printing
Co. De Leon Springs, FL. Pp.433-439.
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Samoiloff^M., S. Schulz, Y. Jordan, K. Denich, and E. Arnott. 1980. A rapid simple
long-term toxicity assay for aquatic contaminants using the nematode Panagrellus
redivius. Can. J. Fish. Aquat. Sci. 37:1167-1174.
Samoiloff,M., J. Bell, D. Birkholz, G. Webster, E. Arnott, R. Pulak, and A. Madrid.
1983. Combined bioassay-chemicalfractionation scheme for the determination and
ranking of toxic chemicals in sediment. Environ. Sci. Technol. 17:329-333.
van Kessel, W., R. Brocades Zaalberg, and W. Seinen. 1989. Testing environmental
pollutants on soil organisms: a simple assay to investigate the toxicity of
environmental pollutants on soil organisms, using CdQ2 and nematodes. Ecotoxicol.
Environ. Safe. 18:181-190.
Vanfleteren, J.R. 1976. Large scale cultivation of a free-living nematode (Caenorhabditis
elegans). Seperatum Experientia 32:1087.
Williams, P. L. and D. B. Dusenbery. 1990. Aquatic toxicity testing using the
nematode, Caenorhabditis elegans. Environ. Toxicol. Chem. 9:1285-1290.
TA1.4-4
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TA1.5. Tests using soil arthropods (insects).
While initially driven by regulatory concerns other than Superfund, various methods
have been developed for evaluating chemical effects on terrestrial insects, especially
pesticide effects on nontarget species (e.g., US EPA 1982). As ecological indicators of
soil contamination, terrestrial insects, and soil arthropods in general, are potentially
critical targets within an ecological effects assessment at a Superfund site. Within
ecological contexts, terrestrial invertebrates play a role in communities and ecosystems
that involves integrated functions such as decomposition, grazing, predation, and
pollination (Croft 1990). While methods that evaluate adverse biological effects in
terrestrial invertebrates exposed to soil contaminants are not widely considered in the
ecological effects assessment process at present, their contributions have, and should
continue to increase in the near future. Presently, there are few standard laboratory
tests using terrestrial insects, but owing to their relative importance in evaluating
ecological effects at Superfund sites, the laboratory methods summarized here should
be considered on a site-specific basis. Through strategies similar to those used with
aquatic invertebrates (e.g., Plafkin, et al. 1989; Klemm, et al. 1990), these test methods
using terrestrial insects could then be interpreted with respect to soil contaminant
concentrations and field survey information regarding insect community structure and
population numbers. Ecological effects associated with a waste site could potentially
be described with greater certainty.
Test method summary: One standardized method for evaluating acute and subacute
chemical effects on terrestrial insects was initially developed for agrichemical
evaluations, and especially for evaluating the effects of insecticides on non-target
insects (e.g., honey bees (Apis mellifera); see US EPA 1982).
A second, well developed but infrequently applied test method, uses harvester ants
(Pogonomyrmex awyhed', Gano, et al. 1985). While currently under-used in ecological
effects assessments at waste sites, as a direct test on bulk soils, the harvester ant test
has proven sensitive to some organic contaminants in single-compound exposures
(i.e., pesticides) and complex chemical mixtures (e.g., wood preservative sludge,
drilling fluid and slop oil). Not unlike tests using other soil invertebrates, metals and
selected organic compounds (i.e., herbicides) were not consistently toxic unless
concentrations were extraordinarily elevated (Gano, et al. 1985). The method was
originally described using wild-caught animals, but culture methods for ants may
encourage wider use for toxirity evaluations. However, while culture methods may
reduce age-related differences in toxicity responses, and the heterogeneity in wild-
caught test organisms may be more relevant to ecological effects assessments than
TA1.5 - 1
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responses of a pre-selected group of animals. Survival is the test endpoint most
readily measured in the harvester ant test.
Yet a third test method may be considered when soil invertebrates are potential
receptors within an ecological effects assessment. Crickets/ much like harvester ants,
have been used only on a limited basis for ecological effects assessments at hazardous
waste sites. From a life history perspective, the orthopterans (crickets and
grasshoppers) represent terrestrial invertebrates that live in close association with
litter, grass, and soil, and owing to their omnivory, crickets and grasshoppers
potentially are targets for contaminant sources in the soil or aerially deposited
contaminants on the soil's surface. As representative soil invertebrates, crickets and
grasshoppers have been used only sparingly in ecological effects assessments for
Superfund sites. Despite their relatively limited used in Superfund at present,
methods (e.g., Walton 1980) have been developed that should be considered in
ecological assessments. For example, Walton exposed adult crickets (Acheta domesticus)
to acridine via the diet, and following 18-day exposures lethality and sublethal effects
were determined. Sublethal, but chronic effects in adults were also measured and
included weight gain, feed consumption, and feces production, while addition
measurements regarding reproductive effects was gained by measuring percent egg
hatch and fecundity. Testing was also completed with various life stages, including
eggs, nymphs, and adults. Differential toxirity was noted across these various life
stages, with the egg being the most sensitive developmental stage tested. Given the
range of soil contaminants that may be pertinent to ecological effects assessments, this
stage-dependent toxicLty should be closely considered during the problem formulation
phase of the assessment, particularly when many terrestrial insects oviposit directly in
soils and may be adversely effected by diminished soil quality.
A fourth test method that has been developed, but not used in ecological effects
assessments for Superfund, uses Drosophila melanogaster. While initially designed for
genotoxicity screening programs (e.g., OECD 1984), the work with D. melanogaster
may be applicable, if its contribution to an ecological effects assessment can be based
upon its being a representative hymenopteran.
Lastly, another source of test information for terrestrial invertebrates (insects and
mites) that may be applicable to ecological effects assessment occurs in FIFRA-driven
toxirity assessments (as amended 1982; Subsection M of Biological Assessment
Guidelines). Here, standard methods are well described (e.g., James and lighthart
1990) and could be adapted for terrestrial sites that required an ecological effects
assessment.
TA1.5 - 2
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Within an ecological effects assessment, complementary laboratory methods may also
be considered. These laboratory techniques potentially yield measurements from field-
collected soil samples that relate toxicity and exposure with ecological effects that may
be expressed by soil microarthropods. For example, to quantitatively and qualitatively
evaluate soil microarthropods, techniques are readily available to extract, enumerate
and identify these organisms in reference and impacted soil samples. Soil
microarthropods are easily extracted from the soil using Tulgren high efficiency
extractors (e.g., Anderson 1988; Seastedt and Crossley 1980). The extracted organisms
can then be counted using dissecting microscopes and identified to genus, or form-
group. Recent innovations in computer-assisted identification (hypercard) have also
reduced the time required to identify these organisms (Moldenke et a/., 1991).
Intended use: These test guidelines were all designed with agrichemicals or other
challenging agents (chemical or biological) as potential hazards. However, in
evaluating adverse biological effects, test design may be similar regardless of the
agent. And, while toxicity estimates may be derived from modifications of these
existing tests, the interpretation of the toxicity information should be weighted by site-
specific information gathered, for example, during field surveys. It is critical that
these issues regarding the interpretation of toxicity test data be addressed early in the
problem formulation phase of the ecological effects assessment.
Previous applications/regulatoryprecedence: While various methods have been
described for testing insects, few Superfund ecological effects assessments have
considered these terrestrial invertebrates beyond preliminary surveys.
Requirements for development and implementation: Few technical support
laboratories are currently providing tests with insects, particularly within the context
of Superfund, but technical support may be gained on a site-specific basis, e.g.,
through local or regional testing services available at land-grant colleges. Until
adequate technical support is available, implementing this biological assessment within
an ecological effects assessment for Superfund may be difficult.
Potential problems and limitations: Technically, efficiencies for arthropod extractions
from different soil types vary, and soil characterizations (see Section 4.0) are critical to
ecological interpretations developed from these biological data. Additionally, the data
base is relatively sparse, and reliance upon adequate reference soils for site-specific
comparisons is critical.
TA1.5 - 3
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References.
Anderson, J.M. 1988. Spatiotemporal effects of invertebrates on soil processes. Biol.
Fertil. Soils. 6:216-227.
Croft, 6. A. 1990. Arthropod biological control agents and pesticides. John Wiley &
Sons, New York, NY. 723pp.
Gano, K. A., D.W. Carlile, and L.E. Rogers. 1985. A harvester ant bioassay for
assessing hazardous chemical waste sites. PNL-5434, UC-11. Pacific Northwest
Laboratory, Richland, WA.
James, R.R. and B. lighthart. 1990. Bioassay for testing the lethal effects of bacterial
pathogens on the predatory beetle Hippodamia convergens Gue. (Coleoptera:
Coccinellidae). 600/3-90/090. U.S. Environmental Protection Agency, Environmental
Research Laboratory, Corvallis, OR.
Klemm, D.J., P. A. Lewis, F. Fulk, and J.M. Lazorchak. 1990. Macroinvertebrate field
and laboratory methods for evaluating the biological integrity of surface waters. 600/4-
90/030. U.S. Environmental Protection Agency, Environmental Monitoring Systems
Laboratory, Cincinnati, OH.
Moldenke, A., C. Shaw, and J. R. Boyle. 1991. Computer-driven image-based soil
fauna taxonomy. Agr. Ecosyst. Environ. 34:177-185.
Plafkin, J.L., M.T. Barbour, K.D. Porter, S.K. Gross, and R.M. Hughes. 1989. Rapid
bioassessment protocols for use in streams and rivers: Benthic macroinvertebratesand
fish. 440/4-89/001. U.S. Environmental Protection Agency, Assessment and
Watershed Protection Division, Washington, D.C.
OECD (Organisation for Economic Co-Operation and Development). 1984. OECD
guidelines for testing of chemicals. Director of Information, OECD. 2, rue Andre
Pascal, 75775 Paris Cedex 16, France.
Seastedt, T.R. and Crossley Jr., D.A. 1980. Effects of microarthropods on the
seasonal dynamics of nutrients in forest litter. Soil Biol. Biochem. 12:337-342.
US EPA. 1982. Pesticide assessment guidelines, Subdivision L, Hazard Evaluation:
Non-target insects. 540/9-82/019. Office of Pesticides and Toxic Substances, U.S.
Environmental Protection Agency, Washington, D.C.
TA1.5 - 4
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Walton, B.T. 1980. Differentiallife-stage susceptibility of Acheta domesticus to acridine.
Environ. Entomol. 9:18-20.
TA1.5-5
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TA1.6. Terrestrial arthropod (non-insect) and isopod tests.
Outside of North America, terrestrial arthropods other than insects have been
considered from the perspective of accidental or coincidental exposure to potentially
harmful chemicals (Croft 1990). For example, to evaluate effects of agrichemical
pesticides or biological control agents on non-target invertebrates, laboratory methods
have been standardized for evaluating chemical effects on mites (e.g., Sewell and
lighthart 1988). While these tests methods have no history in ecological effects
assessments for Superfund, their inclusion in the present compendium dearly
suggests that because of their role in the environment (Croft 1990) terrestrial isopods
should receive consideration as ecological "receptors" during the ecological assessment
within the RI/FS process. The methods developed for pesticide evaluations could be
directly applied to soil contamination evaluations. Alternatively, soil-derived eluates
could be used in the testing process, if the study design indicated that indirect routes
of exposure were likely to occur. While a variety of test species have been used in the
standard tests developed in Europe and the United States (Croft 1990; Hassan 1985;
Hassan, et al. 1987), the laboratory test methods using non-insect arthropods are
relatively straight-forward and could be easily modified to directly meet the
requirements of a soil contaminant evaluation for Superfund. Similarly, biological
assessments using terrestrial isopods have historically been considered in soil
contamination evaluations, although standardization, for example, through American
Society for Testing and Materials (ASTM) or Organization for Economic Cooperation
and Development (OECD) is lacking. As field indicators of contaminant exposure, the
isopod literature suggests that whole body, and organ-specific, contaminant
bioaccumulationmay be monitored with these animals, particularly for some
environmental chemicals, e.g., metals (Beyer, et al. 1984; Beyer and Anderson 1985;
Hopkin and Martin 1984; Hopkin 1986; Hopkin 1990).
Test methods summary: Toxicity tests completed as part of the ecological effects
assessment could use laboratory-reared species representative of the site. Currently,
test systems are well characterized although exposures are generally not associated
with soil exposure directly (Croft 1990; Hassan, et al. 1987; Hassan 1985). Within an
ecological effects assessment, however, test systems may be easily modified to assure
that exposures occur directly via site-soil (for example, contained in Petri dishes).
Alternatively, exposures could occur via glass plates or Petri dishes coated with dried
films of single-compound, defined chemical mixture, or soil eluate. Acute toxicity has
been the most easily measured endpoint following exposure periods that range from
nearly one week to four weeks. Additional endpoints should also consider
reproductive success in the test species, and is most frequently achieved from
counting the number of eggs laid during the exposure period. Testing soils may be
TA1.6 - 1
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modified depending upon test species, and at present numerous beneficial insects and
mites have been evaluated with respect to biological effects associated with chemical
exposure. While the current list of test species reflects the European origins of the
test method/ families and orders of insects and mites included in that strategy are also
representative of North American (Croft 1990).
Although the. distribution of microarthropods has not been routinely used to assess
soil toxicity, this approach could be used in Superfund site assessments. The
microarthropods are extracted from the contaminated and uncontaminated soils using
Berlese funnels or high intensity Tullgren extractors (e.g., Anderson 1988; Seastedt
and Crossley 1980). The extracted arthropods should be preserved for effects based
comparisons (e.g., total numbers) until counted and identified.
Intended use: Determination of the impacts of specific chemicals on specific
populations of microarthropod or isopods have generally been limited, although the
applied literature indicates that species-specific sensitivities may be expressed, e.g.,
copper effects on isopods. Also, a data base needs to be developed which would
allow assessment of the expected numbers and types of arthropods in any given soil.
In addition, information on specific community responses to specific chemicals and
chemical mixtures must be developed.
Previous applications/regulatoryprecedence: Moldenke and Fichter (1988) have
shown that microarthropod populations show differences even after other organisms
no longer show toxic effects. Because of the great diversity of microarthropods, it is
unlikely that any one particular population can be used as the sole indicator organisms
relative to all soil toxicants.
Requirements for development and implementation: By comparison of populations
in contaminated and uncontaminated site soils, it will be possible to demonstrate
population shifts corresponding to soil toxicity. Once these differences are defined for
a given ecosystem types, it would be possible to put together a computerized
taxonomic key which would allow relatively untrained technicians to monitor
microarthropod populations during remediation and restoration process (Moldenke et
al.f 1991). The major effort in terms of time and money would be in initial site
characterization.
Expertise in the identification of microarthropods, and determination of their numbers
exists within most land-grant universities, the USDA, and the extension service, either
at the Federal or State level. Regional centers of expertise have been suggested
although few technical support laboratories currently provide these test.
TA1.6 - 2
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Implementing this biological assessment within an ecological effects assessment for
Superfund may be restricted owing to lack of experienced testing services.
Potential problems and limitations: From a technical perspective, terrestrial
arthropods in general and non-insects in particular have a poorly established
comparative effects and toxicity database. While the potential strengths associated
with toxicity evaluations and effects measurements are numerous (e.g., more
ecologically relevant, amenable to laboratory and field assessment), the lack of
commercial availability may also limit the routine application of these methods in an
ecological effects assessment.
References.
Anderson, J.M. 1988. Spatiotemporal effects of invertebrates on soil processes. Biol.
Fertil. Soils. 6:216-227.
Beyer, W.N., G.W. Miller, and E.J. Cromartie. 1984. Contamination of the O2 soil
horizon by zinc smelting and its effect on woodlouse survival. J. Environ. Qual.
13:247-251.
Beyer, W.N. and A. Anderson. 1985. Toxicity to woodlice of zinc and lead oxides
added to soil litter. Ambio 14:173-174.
Croft, B. A. 1990. Arthropod biological control agents and pesticides. John Wiley &
Sons, New York, NY. 723pp.
Hassan, S. A. 1985. Standard methods to test the side-effects of pesticides on natural
enemies of insects and mites developed by the IOBC/WPRS work group 'Pesticides
and beneficial organisms.' Bull. OEPP/EPPO 15:214-255.
Hassan, S.A., R. Albert, F. Bigler, P. Blaisinger, H. Bogenschutz, E. Boiler, J. Brun, P.
Chiverton, P., Edwards, W.D. Engloert, P. Huang, C. Inglesfield, E. Naton, P.A.
Oomen, W.P.J. Overmeer, W. Rieckmann, L. Samsoe-Petersen, A. Staubli, J.J. Tuset,
G. Viggiani, and G. Vanwetswinkel. 1987. Results of the third joint insecticide testing
programme by the lOBC/WPRS-workinggroup "Pesticides and beneficial organisms."
J. Appl. Ent. 103:92-107.
TA1.6 - 3
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Hopkin, S.P. and M.H. Martin. 1984. Heavy metals in woodlice. Symp. zool. Soc.
Ixmd. 53:143-166.
Hopkin, S.P. 1986. The woodlouse Porcellio scaber as a 'biological indicator' of zinc,
cadmium, lead and copper pollution. Environ. Pollut. (Series B) 11:271-290.
Hopkin, S.P. 1990. Species-specific differences in the net assimilation of zinc,
cadmium, lead, copper and iron by the terrestrial isopods Oniscus asellus and Porcellio
scaber. J. Appl. Ecol. 27:460-474.
Moldenke, A., C. Shaw, and J. R. Boyle. 1991. Computer-driven image-based soil
fauna taxonomy. Agr. Ecosyst. Environ. 34:177-185.
Moldenke, A.R. and B.L Fichter. 1988. Invertebrates of the H.J. Andrews
Experimental Forest, Western Cascade mountains, Oregon: IV. The oribatid mites
(Acari't Cryptostigmata). USDA Forest Service. PNW-GTR-217.
Seastedt, T.R. and Crossley Jr., D.A. 1980. Effects of microarthropods on the
seasonal dynamics of nutrients in forest litter. Soil Biol. Biochem. 12:337-342.
TA1.6 - 4
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TA1.7. Mollusk tests for evaluating terrestrial and wetland habitats.
Wetlands are habitats that are frequently impacted by hazardous waste disposal sites,
and mollusks are often regarded as representative invertebrates characteristic of these
habitats (Pennak 1978). Coincident with these habitat-related questions, some families
of freshwater mussels (Unionidae) have been identified as critical species for ecological
risk assessments for some environmental chemicals (agrichemicals; see U.S.
Department of Interior 1989). Accordingly, methods potentially amenable to ecological
effects assessments at Superfund sites have been developed to evaluate chemical
effects and acute toxicity for sensitive life stages in various mollusk species (Johnson
1990). In contrast to concerns regarding habitat loss and contaminant effects on
freshwater mollusks, efforts to develop effective molluscides have also yielded test
methods (e.g., Crowell 1979; Getzin and Cole 1964) that may be applicable to the
ecological assessment needs within Superfund. While infrequently applied to
ecological effects assessment at present, test methods using mollusks should be
considered at Superfund sites, particularly if site-specific conditions (e.g., potential
impact on wetlands) support their use.
Test method summary: The methods summarized here should be considered
examples of terrestrial and wetland mollusks that are amenable to a toxicity
assessment included within an ecological effects assessment. Historically, marine and
estuarine mollusks have been used in toxicity and ecological effects assessments
within the Office of Pesticides Program (US EPA 1985), and these methods could be
equally applicable to contaminant-related questions for Superfund ecological risk
assessments. Analogous tests with freshwater mollusks, however, have only
relatively recently been developed and a synopsis of one representative test method is
included here. The other test method briefly summarized here comes from pest
control-related work with molluscides. These methods (e.g., Getzin and Cole 1964;
Crowell 1967) are amply suited for toxicity assessments within an ecological effects
assessment, and need only be given site-specific design considerations prior to their
use.
The Unionidae mollusks are characteristic freshwater mussels, and numerous species
could be considered within a toxicity assessment setting. In developing a freshwater
mussel test, Anodonta imbedlis was initially selected as a representative unionid
mollusk, however, the techniques described by Johnson (1990) should be applicable
for testing mussels with similar reproductive strategies. Briefly, in the aquatic toxicity
test with A. imbedlis, exposures are static or renewal, and depending upon endpoint
(e.g., survival or transformation), the exposures are 24-hr or 9 to 11 days. All tests
involve the early developmental stages of the mussel, or glochidia, and juvenile
TA1.7-1
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mussels/ depending upon endpoints being measured. Guidance for developing the
test with freshwater mussels followed ASTM E729 (1991), and while not widely used
at this time/ toxirity assessments with freshwater mussels should be considered within
an ecological effects assessment.
In contrast to the freshwater mussel test that was primarily developed in response to
ecological risk assessment questions related to agrichemical use, test methods that
evaluate terrestrial snails and slugs were developed as efficacy tests for evaluating
molluscides (e.g., Getzin and Cole 1964; Crowell 1979). In these tests, exposures occur
to solid-phase materials, most frequently tainted food or soils containing chemicals.
Technically, exposures may involve various species of snail or slug, although
laboratory-reared snails, like Deroceras reticulatum for example, may be readily
identified as a primary test species. Additionally, site-specificity may suggest species
that are representative of particular habitats impacted by hazardous waste disposal.
Exposures are generally short duration (24-hr to 48-hr), and completed with various
life-stages, although adult testing is most frequently outlined (Getzin and Cole 1964;
Crowell 1979). Contact tests with contaminated site-soil, while not directly outlined in
the efficacy testing literature, could also be considered but the method would require
modification. At present, testing with snails or slugs is relatively straight forward.
Exposures occur in test boxes or glass aquaria, and ideally these test chambers may
include refuge areas where test organisms could avoid the test matrix, be that soil or
food. Of course, the endpoint routinely measured in these efficacy tests was
mortality, but alternative endpoints could also be measured, and would depend upon
test duration, e.g., 10-day exposures could focus on survival and biomass to evaluate
less-than-acute effects.
Intended use: Neither method was designed with ecological effects assessment of
primary concern. However, from the perspective of toxicity assessment in general,
aquatic toxicity tests with freshwater mussels may be critical to a wetland evaluation, if
complex chemical mixtures characteristic of hazardous waste sites were impacting the
habitat. Similarly, the test with terrestrial snails and slugs could complement the field
survey that is routinely included in the ecological effects assessment, and while
originally designed as an efficacy test, the method directly bears upon soil
contamination evaluations and the effects associated with soil exposures.
Previous applications/regulatory precedence: Superfund ecological effects
assessments have not routinely considered terrestrial or wetland mollusks tests, but
ecological risk assessments and toxicity evaluations of agrichemicals have routinely
considered these organisms as prospective targets or receptors of environmental
chemicals, i.e., agrichemicals. Additionally, when threatened or endangered
TA1.7-2
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freshwater mussels are potential receptors at a Superfund site, these test methods
should seriously be considered.
Requirements for development and implementation: Few technical support
laboratories are currently providing tests with these organisms/ and implementing this
biological assessment within an ecological effects assessment for Superfund may be
difficult owing to an absence of experienced testing services.
Potential problems and limitations: No comparative toxirity data base has been
developed/ although the mollusk literature is widespread with an increasing amount
of work being reported that summarize ecological effects associated with exposures
involving freshwater and terrestrial mollusks.
References.
ASTM E729. 1991. Standard guide for conducting acute toxicity tests with fishes/
macroinvertebrates, and amphibians. Annual Book of Standards/ American Society
for Testing and Materials/ Philadelphia, PA.
Crowell, H. 1979. Chemical control of terrestrial slugs and snails. Station Bulletin
628. Agricultural Experiment Station, Oregon State University. Corvallis, OR.
Getzin, S. and S. Cole. 1964. Evaluation of potential molluscides for slug control.
Station Bulletin 658. Washington Agricultural Experiment Station. Pullman, WA.
Johnson, I. 1990. Proposed guide for conducting acute toxicity tests with the early
life stages of freshwater mussels. Report to Ecological Effects Branch, Office of
Pesticide Programs under EPA Contract Number 68-02-4278,87018 4-EEB-08. KBN
Engineering and Applied Sciences, Inc. Gainesville, FL.
Pennak, R. 1978. Pelecypoda. In Freshwater invertebrates of the United States.
Second Edition. John Wiley & Sons. New York, NY. Pp. 736-768.
US EPA. 1985. Hazard evaluation division, Standard evaluation procedure. Acute
toxicity tests for freshwater invertebrates. 540/9-85/005. Office of Pesticide Programs.
Washington, D.C.
TA1.7-3
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US Department of Interior. 1989. Endangered and threatened wildlife and plants. 50
CFR 17.11 and 17.12.
TA1.7-4
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TA1.8. Amphibian test methods.
Wetlands are habitats that are frequently impacted by hazardous waste sites, and
evaluating and monitoring these transition zones between upland and surface water
areas will require a variety of field and laboratory techniques (Tiner 1984; Adamus and
Brandt). Amphibians — frogs and salamanders — may be representative of the fauna
potentially critical to ecological effects assessments for wetlands. Amphibian test
systems are standardized through ASTM (American Society for Testing and Materials;
ASTM E729, E1439 1991). Early embryos of the African dawed-frog (Xenopus laevis)
are used in the standardized test; however, much work has been completed with
alternative test species and should be considered on a site-specific basis (e.g., ASTM
E1439 1991; Under, et al. 1990; Under, et al. 1991).
Test method summary: Amphibian testing using FETAX (frog embryo teratogenesis
assay: Xenopus laevis. Tests may be completed with surface waters, groundwater, or
soil/sediment-derivedeluates. In the standard test, to initiate exposures, less than
eight-hour old frog embryos are placed in Petri dishes containing aqueous test
solutions for either screening or definitive tests. In definitive tests triplicate exposure
series are set up with a maximum of five to six concentrations plus controls in each
replicate. For screening purposes, triplicate Petri dishes may contain 100% site-
samples. Once exposures have been initiated, the 96-hour static-replacement
exposures are renewed at 24-hour intervals at 22 + 2°C. At termination, data include
survivorship, growth (e.g., length), and malformations observations. Survivorship
data (LCgo or percent survival in 100% site-sample) is determined at the end of four-
day exposures. Similarly, EC^s for malformation are recorded in definitive tests, or
percent malformations is recorded in screening tests. Subacute response data will
reflect numbers of gross terata (e.g., scoliosis, lordosis, and kyphosis) developed in
exposed embryos.
Intended use: The test method was originally designed for testing surface waters and
water column exposures with sediments. At present the method is most directly
applicable to wetland evaluations that may be required as part of an ecological effects
assessment for Superfund.
Previous applications/regulatoryprecedence: The amphibian test has been used in
various regulatory and applied science contexts (e.g., ASTM E729 1991; Under, et al.
1991; Birge, et al. 1985), but is finding increased applications in a variety of chemical
hazard and risk evaluations (e.g., Stephan, et al. 1985).
TA1.8 - 1
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Requirements for development and implementation: While more laboratories are
offering testing services with amphibians, only a limited number of technical support
laboratories are currently providing tests with these organisms. Implementing this
biological assessment within an ecological effects assessment for Superfund may be
difficult owing to an absence of experienced testing services.
Potential problems and limitations: The potential problems encountered with
amphibian testing are common to the majority of toxicity test methods currently
available/ that is the ecological significance that can be interpreted from an organismic-
level test. Additionally, unless in situ methods are also included as part of the
ecological effects assessment (see TA4.2), "laboratory-to-field"extrapolation error may
confound biological assessments within an ecological risk context.
References.
ASTM E729. 1991. Standard guide for conducting acute toxicity tests with fishes,
macroinvertebrates, and amphibians. Annual Book of Standards, American Society
for Testing and Materials, Philadelphia, PA.
ASTM E1439. 1991. Standard guide for conducting the frog embryo teratogenicity
test: Xenopus. Annual Book of Standards, American Society for Testing and Materials,
Philadelphia, PA.
Adamus, P.R., and K. Brandt. 1990. Impacts on quality of inland wetlands of the
United States: a survey of indicators, techniques, and applications of community level
biomonitoringdata. (EPA/600/3-90/073). U.S. Environmental Protection Agency,
Environmental Research Laboratory, Corvallis, Oregon, 97333.
Birge, W.J., J. A. Black, and A.G. Westerman. 1985. Short-term fish and amphibian
embryo-larval tests for determining the effects of toxicant stress on early life stages
and estimating chronic values for single compounds and complex effluents. Environ.
Toxicol. Chem. 4:807-821.
Lander, Greg, Janet Barbitta, and Ty Kwaiser. 1990. Short-term amphibian toxicity
tests and paraquat toxicity assessment. In Aquatic Toxicology and Risk Assessment,
Thirteenth Symposium, ASTM STP 1096. American Society for Testing and Materials,
Philadelphia, PA. Pp. 189-198.
TA1.8 - 2
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Under, G., J. Wyant, R. Meganck, and B. Williams. 1991. Evaluating amphibian
responses in wetlands impacted by mining activities in the western United States. Ja
R.D. Comer, P.R. Davis, S.Q. Foster, C.V. Grant, S. Rush, O. Thome, and J. Todd
(Eds.). Issues and technology in the management of impacted wildlife. Thome
Ecological Institute. Boulder, CO. Pp. 17-25.
Stephan, C.E., D.I. Mount, DJ. Hansen, J.H. Gentile, G.A. Chapman, and W.A.
Brungs. 1985. Guidelines for deriving numerical national water quality criteria for the
protection of aquatic organisms and their uses. NTIS PB85-227049. U.S.
Environmental Protection Agency, Environmental Research Laboratory, Duluth, MN.
Tiner, Jr., R.W. 1984. Wetlands of the United States: Current status and recent
trends. U.S. Fish and Wildlife Service, Habitat Resources, One Gateway Center,
Newton, MA. 02158.
TA1.8 - 3
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TA1.9. Acute, subacute or chronic tests using small mammals/fur-bearers.
Acute, subacute and chronic toxicity testing procedures for wildlife species, primarily
small mammals (e.g., Cholakis, et al. 1981; McCann, et al. 1981; Under and Richmond
1990) and mammalian carnivores (mustelids like mink, Mustela vison and ferret,
Mustela putorius furo) have been developed and applied in contaminant evaluations
previously (e.g., Hornshaw, et al. 1986a,b). In many respects, these methods are
modifications of efficacy tests, or toxicity tests, that have been standardized for
evaluating vertebrate control agents such as predacides and rodentirides (ASTM 1991).
While routes of exposure have not generally considered the ingestion of contaminated
soil directly, the consumption of contaminated feed, oral intake via drinking, and
dermal uptake of toxicants has been studied with a variety of wildlife species, and
methods have been developed as modifications of these existing mammalian tests.
These methods are readily adaptable for evaluating ecological indicators that may be
pertinent to Superfund. Laboratory tests are applicable to single-compound, defined
chemical mixtures, and complex chemical mixture exposures, depending upon the
contaminant-related questions driving the toxicity evaluations within the ecological
effects assessment. These tests may be considered analogues of avian toxicity tests,
and are capable of yielding comparative toxicity information in addition to estimates
(e.g., MATCs or concentration-response curves).
Test methods summary: Any material capable of being incorporated into feed stocks
may be evaluated. Dietary exposures are frequently considered in subacute tests, and
may yield NOECs (no observable effects concentrations), LOECs (lowest observable
effects concentrations), or LC^ (lethal concentration which yields 50% mortality)
estimates from definitive feeding trials. In addition to these standard toxicity
endpoints, these test methods may also yield information pertinent to interpretation of
effects observed in the field. Necropsies (external and gross internal observations)
may be accompanied with organ weights (wet and dry weight ratios, for example) and
histopathology assessments (qualitative or quantitative) on tissue samples, for
example. These observations may be critical in fully developing the ecological effects
assessment for any particular site.
Deer mice (Peromyscus maniculatus), white-footed mice (Peromyscus leucopus, prairie
voles (Microtus ochrogaster), and meadow voles (Microtus pennsylvanicus)are suggested
candidates for test species, particularly within ecological context. House mice (Mus
musculus) may also be used in subacute testing protocols, although insectivores (e.g.,
Sorex cinerus and Blarina brevicauda) and other rodents (e.g., heteromyids) may require
that methods be modified on the basis of their relatively limited use in the testing
process. For soil contamination evaluations, test species may vary and depend upon
TA1.9 - 1
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site-specific considerations that may be reflected in a food-chain contamination study,
for example.
In considering food-chain contamination evaluations that may be included within an
ecological effects assessment, laboratory methods (e.g., Hornshaw, et al. 1986a,b)have
been developed with mink (Mustela vison) and European ferret (Mustela putorius furo)
to evaluate subacute and chronic effects associated with dietary exposures. While
infrequently applied in contaminant evaluations within an ecological assessment for
Superfund, these methods are well developed, and require adequate laboratory
facilities to conduct animal tests. For example, feeding studies with small carnivores
require environmentally-controlledexposure facilities and routinely must be conducted
over 28-day (subacute tests; LC^) to three to four months (reproduction test). At
termination various endpoints can be considered (see Hornshaw, et al. 1986a,b), but
routine data collections should include gross external and internal observations upon
necropsy, whole body and target organ weights, and food and water consumption.
In addition, for reproduction test measures of reproductive success, e.g., number of
offspring, offspring weight (total litter; individual weights) should be recorded at birth
and immediately before test termination.
Intended use: Historically, mammalian tests with wildlife species have been used for
evaluating contaminant effects — acute and chronic — for a variety of toxicants,
including agrichemicals, organics, and metals. The test systems described here are
conducive to testing various routes of exposure in laboratory settings, although
inhalation exposure systems (e.g., Driver, et al. 1991) would be costly if respirable
dusts, particulates, or volatile compounds were contaminants of concern. Acute
effects testing, for example, using surface water impacted by soil contaminants as
drinking water sources, can easily be studied using these methods. And, dietary
routes of exposure using, for example, native vegetation or soil-contaminated feed
stocks were the original considerations in designing feeding trials amenable to wildlife
hazard assessment (tinder and Richmond 1990).
Previous applications/regulatoryprecedence: From a regulatory perspective, toxicity
tests with small mammal wildlife species are not frequently completed within an
ecological effects assessment. However, given the technical support and facilities, on
a site-by-site basis these acute and subacute toxicity assessments would be relatively
easy to incorporate into the overall ecological effects assessment. For example, chronic
effects on indigenous small mammals associated with contaminated soil or
contaminated forage growing on such soil (e.g., food-chain contamination) could be
considered in relatively short-term tests (e.g., McCann, et al. 1981; Hornshaw, et al.
1986a,b;Iinder and Richmond 1990).
TA1.9 - 2
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Requirements for development and implementation: Few technical support
laboratories are currently providing tests with these organisms, and implementing this
biological assessment within an ecological effects assessment for Superfund may be
difficult owing to an absence of experienced testing services.
Potential problems and limitations: Depending upon species selection, potential site-
specific receptors may not be available for testing. The comparative data base is
currently sparse, and intersperies extrapolation errors as well as "laboratory-to-field"
extrapolation errors would have to be considered within an uncertainty analysis.
References:
ASTME552. 1991. Test method for efficacy of acute mammalian predacides. Annual
book of ASTM standards. Volume 11.04. Pesticides; Resource Recovery; Hazardous
Substances and Oil Spill Responses; Waste Disposal; Biological Effects. American
Society for Testing and Materials (ASTM). Philadelphia, PA. 19103.
ASTM E555. 1991. Practice for determining acute oral LD^ for testing vertebrate
control agents. Annual book of ASTM standards. Volume 11.04. Pesticides;
Resource Recovery; Hazardous Substances and Oil Spill Responses; Waste Disposal;
Biological Effects. American Society for Testing and Materials (ASTM). Philadelphia,
PA. 19103.
ASTME593. 1991. Test method for efficacy of a multiple-dose rodenticide under
laboratory conditions. Annual book of ASTM standards. Volume 11.04. Pesticides;
Resource Recovery; Hazardous Substances and Oil Spill Responses; Waste Disposal;
Biological Effects. American Society for Testing and Materials (ASTM). Philadelphia,
PA. 19103.
ASTM E757. 1991. Test method for efficacy of canine reproduction inhibitors.
Annual book of ASTM standards. Volume 11.04. Pesticides; Resource Recovery;
Hazardous Substances and Oil Spill Responses; Waste Disposal; Biological Effects.
American Society for Testing and Materials (ASTM). Philadelphia, PA. 19103.
ASTME758. 1991. Test method for mammalian acute percutaneous toxicity. Annual
book of ASTM standards. Volume 11.04. Pesticides; Resource Recovery; Hazardous
Substances and Oil Spill Responses; Waste Disposal; Biological Effects. American
Society for Testing and Materials (ASTM). Philadelphia, PA. 19103.
TA1.9 - 3
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ASTME1103. 1991. Test method for determining subchronic dermal toxicity. Annual
book of ASTM standards. Volume 11.04. Pesticides; Resource Recovery; Hazardous
Substances and Oil Spill Responses; Waste Disposal; Biological Effects. American
Society for Testing and Materials (ASTM). Philadelphia, PA. 19103.
ASTM E1163.' 1991. Test method for estimating acute oral toxicity in rats. Annual
book of ASTM standards. Volume 11.04. Pesticides; Resource Recovery; Hazardous
Substances and Oil Spill Responses; Waste Disposal; Biological Effects. American
Society for Testing and Materials (ASTM). Philadelphia, PA. 19103.
ASTM E1372. 1991. Test method for conducting a 90-day oral toxicity study in rats.
Annual book of ASTM standards. Volume 11.04. Pesticides; Resource Recovery;
Hazardous Substances and Oil Spill Responses; Waste Disposal; Biological Effects.
American Society for Testing and Materials (ASTM). Philadelphia, PA. 19103.
ASTM E1373. 1991. Test method for conducting a subchronic inhalation toxicity
study in rats. Annual book of ASTM standards. Volume 11.04. Pesticides; Resource
Recovery; Hazardous Substances and Oil Spill Responses; Waste Disposal; Biological
Effects. American Society for Testing and Materials (ASTM). Philadelphia, PA.
19103.
Cholakis, J.M., M.J. McKee, L.C.K. Wong, and J.D. Gile. 1981. Acute and subacute
toxicity of pesticides in microtine rodents, to D. W. Lamb and E.E. Kenaga, eds.,
Avian and Mammalian Wildlife Toxicology: Second Conference. ASTM STP 757.
American Society for Testing and Materials, Philadelphia, PA, pp. 143-154.
Driver, C.J., M.W. Ugotke, P. Van Voris, B.D. McVeety, B.J. Greenspan, and D.B.
Drown. 1991. Routes of uptake and their relative contribution to the toxicologic
response of northern bobwhite (Colinus virginianus) to an organophosphate pesticide.
Environ. Toxicol. Chem. 10:21-33.
Hornshaw, T.C., R.J. Aulerich, and R.K. Ringer. 1986b. Toxicity of o-cresol to mink
and European ferrets. Environ. Toxicol. Chem. 5:713-720.
Hornshaw, T.C., R.K. Ringer, and RJ. Aulerich. 1986a. Toxicity of sodium
monofluoroacetate (Compound 1080) to mink and European ferrets. Environ. Toxicol.
Chem. 5:213-223.
TA1.9 - 4
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Under, Greg and M.E. Richmond. 1990. Feed aversion in small mammals as a
potential source of hazard reduction for environmental chemicals: agrichemical case
studies. Environ. Toxicol. Chem. 9:95-105.
McCann, J.A., W. Teeters, and D.J. Urban, and N. Cook. 1981. A short-term dietary
toxicity test on small mammals. In D.W. Lamb and E.E. Kenaga, eds., Avian and
Mammalian Wildlife Toxicology: Second Conference. ASTM STP 757. American Society
for Testing and Materials, Philadelphia, PA, pp. 132-142.
TA1.9 - 5
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TAl.lO. Acute and subacute avian toxirity tests.
Acute and subacute exposure systems have been developed and standardized for
evaluating contaminant effects on birds. Acute methods have been generalized for
terrestrial vertebrates, and have been well developed in chemical hazard evaluations
for birds. Generally, subacute tests may be more pertinent to ecological effects
assessments and will involve feeding routes of exposure. Methods that directly
measure multiple routes of exposure (e.g., dietary, inhalation, and dermal) have also
been described, but are more costly (Driver, et d. 1990). In those methods that are
standardized (ASTM E857, E1062 1991), the toxicity endpoints are death, although
other information may be collected concurrent with mortality information.
Test method summary: The subacute test (ASTM E8571991) was designed as a five-
day test followed by a three-day post-feeding observation period, and was considered
a screening test to evaluate whether additional testing was indicated. Originally, the
test was used to evaluate single chemicals (e.g., pesticides) incorporated into an
animal's feed, but defined mixtures as well as contaminated soil could be included in
tainted feed for a laboratory study developed within an ecological effects assessment
for Superfund. While numerous species are amenable to subacute testing (northern
bobwhite (Colinus virginianus); Japanese quail (Coturnix japonica); mallard (Anas
platyrhynchos)',ring-necked pheasant (Phasianus colchkus)) using the standardized
method, alternative test species could be considered on a site-specific basis, but the
methods would have to be modified.
Given the results from a screening method like the standardized five-day dietary test,
additional testing may be indicated, and within ecological contexts, an evaluation of
chemical effects on reproduction may be in order for a particular waste site. Through
ASTM E1062 (1991), a standard practice for conducting reproductive studies with
mallard (Anas platyrhynchos)and northern bobwhite (Colinus virginianus)has been
developed. While primarily used in the agrichemical registration process, the method
outlined in ASTM E1062 (1991) may be applicable on a site-specific basis (e.g., birds
may be identified as a receptor during "Problem Formulation" [see Figure 2]), and
feeding trials with single-chemicals, defined chemical mixtures, or soil-contaminated
feed may be beneficial to the ecological effects assessment. In conjunction with field
surveys, these tests — subacute feeding and reproductive — could be critical
components in interpreting ecological effects at a site.
Intended use: Avian toxicity tests were originally designed for single-compound
testing, most commonly agrichemicals. But, these methods may be applied to various
complex chemical mixture exposure evaluations, if dietary routes of exposure are
TA1.10-1
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critical pathways that are being considered within the ecological effects assessment.
For application to Superfund ecological effects assessments/ avian tests would
generally be useful in evaluating feeding routes of exposure; other routes of exposure
(e.g./ inhalation) would be technically more difficult and costly.
Previous applications/regulatoryprecedence: Few Superfund site ecological effects
assessments have considered birds in laboratory and field studies. Avian species
frequently are identified as receptors when considered as part of the ecological effects
assessment/ but literature reviews (e.g., toxirity and bioaccumulationpotential of
target analytes) and occasional site-specific tissue analysis for target analytes are the
routine tools used in performing the ecological assessment.
Requirements for development and implementation: Because of the regulatory
requirements under FIFRA/ numerous technical support services offer both acute and
subacute avian testing; Little work/ however/ has been done under the auspices of
Superfund as far as avian toxicity assessment in the laboratory/ so testing services/
while experienced in testing various chemicals/ will be relatively unfamiliar with
Superfund regulations driving the biological assessment itself.
Potential problems and limitations: From a technical perspective/ a support service is
available for performing site-specific avian studies. As frequently noted/ while well
developed for some environmental chemicals (e.g./ agrichemicals)/ for some sites the
limitations of the comparative toxicity data base may require that surrogate test species
be carefully selected from the list of potential test species (see ASTM E857, E1062
1991).
References:
ASTM E857. 1991. Practice for conducting subacute dietary toxicity tests with avian
species. Annual book of ASTM standards. Volume 11.04. Pesticides; Resource
Recovery; Hazardous Substances and Oil Spill Responses; Waste Disposal; Biological
Effects. American Society for Testing and Materials (ASTM). Philadelphia, PA.
19103.
ASTM E1062. 1991. Practice for conducting reproductive studies with avian species.
Annual book of ASTM standards. Volume 11.04. Pesticides; Resource Recovery;
Hazardous Substances and Oil Spill Responses; Waste Disposal; Biological Effects.
American Society for Testing and Materials (ASTM). Philadelphia, PA. 19103.
TA1.10-2
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Driver, C J., M.W. Ligotke, P. Van Voris, B.D. McVeety, B.J. Greenspan, and D.B.
Drown. 1991. Routes of uptake and their relative contribution to the toxicologic
response of northern bobwhite (Colinus virginianus) to an organophosphate pesticide.
Environ. Toxicol. Chem. 10:21-33.
TA1.10-3
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TA1.11. Immunotoxicity test methods using vertebrates.
Various biomarkers have been identified recently that may be critical to weight of
evidence arguments in Superfund ecological effects assessments, and several methods
have been modified from the biomedical and veterinary sciences for evaluating
sublethal immunologic effects that may be associated with contaminant exposures in
wildlife species. Test modifications have been developed for terrestrial vertebrates,
most notably birds and mammals, and these methods may be considered within an
ecological effects assessment for Superfund.
Test methods summary: Currently, various test methods are available for evaluating
immune function in vertebrates, many being modified from biomedical or veterinary
(Rose and Friedman 1976) standard methods. In general, site-specific sampling and
analysis plans must be developed with particular methods fully detailed, but in
general tests should be selected to evaluate humoral and cellular immune function.
For example, in evaluating humoral function, plaque-forming cell assays modified
after Jerne and Nordin (1963) have been used to evaluate contaminant effects in
vertebrates, and various additional methods are potentially available to qualitatively
and quantitativelymeasure immunoglobulins (e.g., Kochwa 1976; Davis and Ho 1976).
Similarly, cellular immune function (e.g., Oppenheim and Schechter 1976) and other
immunohematologicalfunctions (e.g., complement; Gewurz and Suyehira 1976) may
be considered in vertebrate targets identified as critical to an ecological effects
assessment. Here, numerous standardized laboratory test methods would be
applicable and would include, for example, in vivo and in vitro phytohemagglutinin
tests, delayed-typehypersensitivity tests, and leukocyte migration inhibition tests.
Macrophage activation and function, as well as phagocytosis, may also be considered
with a number of standardized test methods that would probably require
modification, depending upon the site-specific receptors identified early in problem
formulation.
Intended use: At present immune function evaluations, like many biological markers,
are insufficiently developed to act as "stand alone" assessment methods. However,
when site-specific contaminant histories indicate potential immunologic effects,
multiple immunofunction tests (Exon, et al. 1986) may be considered in integrated
studies that evaluate, for example, subacute effects on avian receptors.
Immunotoxicity assessments may be beneficial to evaluating long-term ecological risks
associated with a particular site.
Previous applications/regulatoryprecedence: By design, few, if any, Superfund
ecological effects assessments have included evaluations of vertebrate immune
TA1.11 -1
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function. The comparative literature suggests, however, that the methods may be
applicable to certain contaminants that have been identified previously as
immunotoxins (e.g., selenium, Fairbrother and Fowles 1990; pentachlorophenol,
Kerkvliet, et al. 1982; polychlorinatedbiphenyls, Vos and De Roji 1972; complex
petroleum hydrocarbons mixtures, Rocke, et al. 1984).
Requirements for development and implementation: Few technical support
laboratories are currently providing tests with these organisms, and implementing this
biological assessment within an ecological effects assessment for Superfund may be
difficult owing to an absence of experienced testing services. In view of their more
general application, however, regional diagnostic veterinary laboratories may be able
to provide the technical support required, if these less-than-acute effects are developed
as part of the toxicity evaluation within an ecological effects assessment for a particular
site.
Potential problems and limitations: While the potential effects associated with altered
immune competence are significant regardless of the source of contaminants, within
Superfund ecological effects assessments the role of immunotoxicity assessment
should be secondary or supportive of other organismic or population level
information. And, while the immunotoxicity literature strongly suggests that these
endpoints may gain increased value in evaluating subacute or chronic contaminant
effects, currently they potentially contribute to weight of evidence arguments on a site-
specific basis. Depending upon the complexity and mechanisms of toxicity
characteristic of chemical mixtures at hazardous waste sites, immunotoxicity, like other
sublethal indicators of contaminant effects, will provide critical information regarding
the potential long-term health of terrestrial populations impacted by hazardous wastes.
References:
Davis, N. and M. Ho. 1976. Quantitationof immunoglobulins. In Manual of clinical
immunology, N. Rose and H. Friedman (Eds.). American Society of Microbiology.
Washington, D.C. pp. 4-16.
Exon, J.H., L.D. Koller, P.A. Talcott, C.A. O'Reilly, and G.M. Henningsen. 1986.
Immunotoxicity testing: an economical multiple-assay approach. Fund. Appl. Toxicol.
7:387-697.
TA1.11 - 2
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Fairbrother, A. and J. Fowles. 1990. Subchronic effects of sodium selenite and
selenomethionineon several immune-functions in mallards. Arch. Environ. Contam.
Toxicol. 19:836-844.
Gewurz, H. and L. Suyehira. 1976. Complement. In Manual of clinical
immunology, N. Rose and H. Friedman (Eds.). American Society of Microbiology.
Washington, D.C. pp. 36-50.
Kerkvliet, N.I., L. Baecher-Steppan, A.T. Claycomb, A.M. Craig, and C.G. Sheggeby.
1982. Immunotoxicity of technical pentachlorophenol (PCP-T): depressed humoral
immune responses to T-dependent and T-independent antigen stimulation in PCP-T
exposed mice. Fund. Appl. Toxicol. 2:90-99.
Kochwa, S. 1976. Immunoelectrophoresis. In Manual of clinical immunology, N.
Rose and H. Friedman (Eds.). American Society of Microbiology. Washington, D.C.
pp. 17-35.
Oppenheim, J. and B. Schechter. 1976. Lymphocyte transformation. In Manual of
clinical immunology, N. Rose and H. Friedman (Eds.). American Society of
Microbiology. Washington, D.C. pp. 81-94.
Rocke, T.E., T.M. Yuill, and R.D. Hinsdill. 1984. Oil and related toxicant effects on
mallard immune defenses. Environ. Res. 33:343-352.
Vos, J.G. and T.H. De Roij. 1972. Lmmunosuppressive activity of a polychlorinated
biphenyl preparation on the humoral immune response in guiena pigs. Toxicol. Appl.
Pharmacol. 21:549-555.
TA1.11 - 3
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TA1.12. Immunotoxitity test methods using invertebrates.
While immunotoxicity test methods are more widely described for vertebrates, similar
methods have been applied to terrestrial invertebrates under controlled laboratory
conditions. For example, methods have been developed for evaluating the
immunocompetence of earthworms exposed to various contaminants (Chen, et al.
1991; Mohrig, et al. 1984; Rodriguez-Grau, et al. 1989; Eyambe, et al. 1990). hi these
subacute tests, earthworms are exposed to single-chemicals or complex mixtures. The
majority of work has been completed using filter-paper contact tests, but soil exposed-
earthworms could be tested in conjunction with laboratory or in situ tests (see TA1.1
and TA4.1). Regardless the route of exposure, coelomocytes are non-invasively
harvested from exposed or control earthworms, generally Lumbricus terrestris, which
are then evaluated for their viability, ability to form secretory rosettes, total and
differential cell counts, among other potential subacute measurement endpoints (see
TA1.2). Through these and similar measurement endpoints, the health of earthworms
exposed to soil contaminants can be inferred, and on the basis of biological response
data garnered upon test termination, the results may suggest potential long-term
effects (i.e., immunocompromised populations) associated with exposure.
Test methods summary: While not presenting as long a history as immunotoxicity
evaluations in vertebrates, a few analogous test methods are currently available for
evaluating immune function in invertebrates such as earthworms (e.g., Stein and
Cooper 1988; Rodriguez-Grau, et al. 1989; Eyambe, et al. 1990). These published
methods are currently focussed on humoral immunocompetence, and in conjunction
with total and differential coelomic cell counts may be beneficial for evaluating the
potential immunotoxicity of contaminated soils.
Intended use: As a direct indication of soil quality, altered immune function in
terrestrial macroinvertebrates should be interpreted with caution when evaluating
adverse ecological effects associated with environmental contaminants. Again, as a
biomarker, altered immune function in earthworms should be considered as
supporting data in an integrated study that addressed, for example, organismic-level
measurements (e.g., standardized 14-day earthworm test; see TA1.1) and field survey
information. While not sufficiently developed at present, the future value of these
methods centers upon the comparative toxitity data base that can be developed
relative to other receptors (e.g., mammals and birds).
Previous applications/regulatoryprecedence: The method(s) introduced here are
standardized within the experimental literature, but will not be developed beyond that
context in the near future. As with many biomarkers, these methods for evaluating
TA1.12-1
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humoral immunity in invertebrates are not intended to be "stand alone" tests within
an ecological effects assessment. Rather, the strengths of these method(s) lie in their
contribution to weight of evidence arguments that are supportive of complementary
tests that dearly illustrate adverse effects associated with soil exposures, e.g. 14-day
survival earthworm tests.
Requirements for development and implementation: Few technical support
laboratories are currently providing tests with these organisms/ and implementing this
biological assessment within an ecological effects assessment for Superfund may be
difficult owing to an absence of experienced testing services.
Potential problems and limitations: At present, the interpretation of altered immune
function in soil macroinvertebrates should be guarded, particularly within the context
of an ecological effects assessment. Only through integrated studies using organismic-
level tests and field surveys will results from these invertebrate immune function tests
be ecologically relevant. However, unlike the immunotoxicity information garnered
from terrestrial vertebrates, an evaluation of the immunocompetence of soil
macroinvertebrates, such as earthworms, may directly reflect the long-term adverse
biological effects associated with soil contaminants. For terrestrial vertebrates, the
majority of exposure routes are indirect, except for direct soil ingestion, but for soil-
dwelling invertebrates the routes of exposure are more direct owing to the close
contact between receptor and contaminant source. Depending upon physicochemical
properties of the contaminant mixture at a site, cutaneous or dermal uptake of
contaminant gains nearly equal, if not greater, weight with direct soil ingestion when
exposure to contaminants is considered in earthworms.
References:
Chen, S.C., L.C. Fitzpatrick, A.J. Goven, B.J. Venables, and E.L. Cooper. 1991.
Nitroblue tetrazoliumdye reduction by earthworm (Lumbricus terrestris) coelomocytes:
an enzyme assay for nonspecific immunotoxicity of xenobiotics. Environ. Toxicol.
Chem. 10:1037-1043.
Eyambe, G.S., A.J. Goven, L.C. Fitzpatrick, B.J. Venables, and E.L. Cooper. 1990. A
non-invasive technique for sequential collection of earthworm (Lumbricus terrestris)
leukocytes during subchronic immunotoxicity studies. Lab. Animals 25:61-67.
TA1.12-2
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Mohrig, W., E. Kanschke, and M. Ehleers. 1984. Rosette formation by coelomocytes
of earthworm Lumbricus terrestris L. with sheep erythrocytes. Devel. Comp.
Immunology 8:471-476.
Rodriguez-Grau, J., B.J. Venables, L.C. Fitzpatrick, and E.L. Cooper. 1989.
Suppression of secretory rosette formation by PCBs in Lumbricus terrestris: an
earthworm assay for humoral immunotoxicity of xenobiotics. Environ. Toxicol. Chem.
8:1201-1207.
Stein, E.A. and E.L. Cooper. 1988. In vitro agglutininproduction by earthworm
leukocytes. Devel. Comp. Immunol. 12:531-547.
TA1.12-3
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TA2. PLANT TEST METHODS FOR THE ASSESSMENT OF SOIL
CONTAMINATION AT HAZARDOUS WASTE SITES
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TA2.1. Seed germination and root elongation tests.
Techniques modified from methods originally developed in the plant and weed
science disciplines have yielded short-term tests which assess toxic chemical effects on
plants. The seed germination and root elongation bioassays are laboratory toxirity
tests which directly and indirectly assess toxicity of soils, and evaluate toxirity
endpoints (seed germination and root elongation) pertinent to ecological assessments
for terrestrial and wetland habitats. Seed germination tests measure toxicity associated
with soils directly, while root elongation tests consider the indirect effects of water-
soluble constituents which may be present in site-samples.
Test method summary: 120-hour seed germination test. Numerous seed vials
containing 40 size-graded seeds (e.g. lettuce Lactuca sativa) should be collected, stored
in the refrigerator at 4°C, and be readily available for soil testing. Prior to testing,
stored seeds should be removed and allowed to warm to room temperature.
Depending on the moisture fraction of the soil sample, a standard seed germination
evaluation requires no less than 1200g (wet weight) site soil. After screening and
mixing the soil sample, the water content and water holding capacity of the sample
should be determined. Soil pH should range between pH 6 and 10, and should be
completed using a soil slurry. Following these initial pH determinations, any pH
adjustments required for test initiation can be determined.
Screening tests should be completed on uncut, homogenized soil samples. If
definitive tests follow these screening evaluations, EC^ estimates will require at least
three replicates of at least five test soil concentrations. Three replicates of negative
and positive controls must also be completed. Initial measurements of pH and
temperature should be completed prior to planting seeds. If these initial
measurements are acceptable, soils should then be loaded into petri dishes. Test soil
pH must range between pH 6 to 10 to assure germination of test seeds. When petri
dishes are readied, 40 seeds should be planted in each replicate petri dish. Forceps
may be used to space the seeds and minimize crowding. After planting, 16-mesh
cover sand should then be poured over each plate, and the petri dishes subsequently
placed into ZiplockR bags, sealed, and incubated at 24 ± 2°C for 120 hours in an
environmental chamber. An airtight seal, leaving as much air space as possible, will
assure adequate humidity for each replicate dish, and light intensity within the
environmental chamber should be 400 ± 40 foot-candles. The first 48 hours of
incubation should be occur in complete darkness, and the last 72 hours should occur
under 16:8 light:dark cycle.
TA2.1 - 1
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Following the 120-hour incubation time, germinated seeds should be counted and
germination recorded. Germinated shoot spikes must be above the soil surface to be
considered a germinated seed (Thomas and dine, 1985). Screening tests are recorded
as percent germination. If definitive tests are completed, median effective estimates
(EC^s) may be calculated. The acceptability of the test is based on percent
germination of positive and negative controls. At least 80% of the negative controls
must germinate in order for the test to be acceptable. For testing lettuce seeds, 2-
chloroacetamide should be prepared as a positive control.
Test method summary: 120-hour root elongation test. Root elongation evaluations,
as modified from Ratsch (1983), estimate the adverse biological effects of soil eluates to
lettuce seedlings (Lactuca sativa) in a 120-hour test. Screening evaluations may be
completed using uncut soil eluates; if definitive tests follow these preliminary screens,
at least three replicates must be included as part of the test design.
To initiate screening or definitive tests, Whatman No. 3 (9cm diameter) filter papers
are placed in the bottom of labelled petri dishes. Each filter paper-lined petri dish
should be charged with 4ml of eluate or dilution, then five seeds are spaced
equidistantlyper plate. Negative controls generally receive double-distilled or reverse-
osmosis grade water, and positive controls receive a standard addition of reference
toxicant, e.g., sodium fluoride solution, per replicate. Eluate pH should be noted and
recorded for each undiluted sample. If definitive tests are designed, the lowest eluate
dilution should also have pH measurements recorded. All petri dishes should be
incubated in a humidified, garbage bag-lined box which is sealed and placed in an
environmental chamber (24 + 2°C) for 120 hours.
After the 120-hour incubation, root lengths are measured from the transition point
between the hypocotyl and root to the end of the root tip. All roots are measured to
the nearest millimeter, and a minimum of 80% of the negative control seeds must
have germinated for the test to be valid. Root elongation results in screening tests are
reported as percent reduction in root lengths in treatments relative to controls; in
definitive tests, EC^ (the concentration which inhibits root elongation by 50% relative
to controls) may be calculated. Three negative and three positive controls must also
be included for definitive tests, and minimum QA/QC specifications require that
negative controls have least 80% germination, and positive controls yield growth
reductions consistent with laboratory-defined criteria.
Intended use: Both these phytotoxirity test methods were adapted from preexisting
methods drawn from the plant and weed science literature. From a phytoloxicity
testing perspective, a relatively well developed data base is available, particularly for
TA2.1 - 2
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agrichemicals (Fletcher, et al. 1988). The existing procedures, however, were primarily
designed with waste site assessment in mind. Seed germination tests should be
considered critical life stage tests suited to bulk soil testing. The germination endpoint
measured upon test completion, however, should more accurately be characterized as
a germination and emergence endpoint. The test should be considered for use in an
ecological effects assessment when field surveys suggest that plant communities have
been impacted at a site, or that future land use may require a phytotoxicity evaluation
as part of the soil contamination testing.
In contrast, the root elongation test requires that soil elutions be prepared prior to
testing. Through the root elongation test, the biological activity of water soluble soil
constituents, both contaminant and non-contaminant, is measured. The soil-derived
eluate may be directly relevant to evaluations of soil contamination and groundwater
quality relationships, or to evaluations of altered quality of surface water runoff from a
contaminated site. Also, when soil contamination directly or indirectly impacts the
plant rhizosphere, soil-derived eluates may provide information regarding interstitial
water quality that potentially influences plants inhabiting contaminated soil.
Previous applications/regulatory precedence: Vegetation tests to evaluate
phytotoxicity at hazardous waste sites have increased in the recent past (e.g., Thomas
and dine 1985; tinder, et al. 1990), and short-term tests like seed germination and
root elongationhave been codified in response to other program needs (e.g., TSCA;
CFR 1985). Technical support documents developed by various agencies have
outlined these methods (e.g., FDA 1987).
Requirements for development and implementation: The increasing demands,
primarily from Superfund and FIFRA, for evaluating phytotoxicity have resulted in a
number of laboratories offering routine testing services for seed germination and root
elongation evaluations. However, due to the relatively recent development of these
regulatory requirements, few of these laboratories have a long performance history,
and implementing these biological assessments within an ecological effects assessment
for Superfund may be difficult owing to an absence of experienced technical support.
Potential problems and limitations: The comparative toxicity data base, at least for
some species and some potential chemicals of concern (e.g., heavy metals and
agrichemicals) is relatively well developed and readily accessible (Fletcher, et al. 1988;
Fletcher, et al. 1990). However, limitations to the comparative toxicity data base
should be considered during the interpretation of phytotoxicity data within an
ecological effects assessment. For example, the current data collection is heavily
skewed toward north-temperate, agricultural species, particularly grasses and legumes
TA2.1 - 3
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(Fletcher, et al. 1988; Fletcher, et al. 1990; Gorsuch, et al. 1990), and little information is
available regarding less commercially important native plants, including woody
species.
References:
CFR [Code of Federal Regulations]. 1985. Rules and regulations; Section 797.2750,
Seed germination/root elongation toxicity test. September 27,1985. CFR 50
(188):39389-39391.
FDA [Food and Drug Administration]. 1987. Sections 4.06 [Seed germination and
root elongation];4.07 [Seedling growth]. In Environmental Assessment Technical
Assistance Handbook. NTIS, PB 87-175345. U.S. Food and Drug Administration,
Washington, D.C.
Fletcher, J.S., F.L. Johnson, and J.C. Me Farlane. 1990. Influence of greenhouse
versus field testing and taxonomic differences on plant sensitivity to chemical
treatment. Environ. Toxicol. Chem. 9:769-776.
Fletcher, J.S., F.L. Johnson, and J.C. Me Farlane. 1988. Database assessment of
phytotoxicity data published on terrestrial vascular plants. Environ. Toxicol. Chem.
7:615-622.
Gorsuch, J.W., R.O. Kringle, and K.A. Robillard. 1990. Chemical effects on the
germination and early growth of terrestrial plants. In Plants for Toxicity
Assessment, ASTM STP 1091. W. Wang, J.W. Gorsuch, and W.R. Lower, Eds.,
American Society for Testing and Materials. Philadelphia, PA. Pp. 49-58
Under, G., J.C. Greene, H. Ratsch, J. Nwosu, S. Smith, and D. Wilborn. 1990. Seed
germination and root elongation toxicity tests in hazardous waste site evaluation:
methods development and applications. In Plants for Toxicity Assessment, ASTM
STP 1091. W. Wang, J.W. Gorsuch, and W.R. Lower, Eds., American Society for
Testing and Materials. Philadelphia, PA. Pp. 177-187.
Ratsch, H. 1983. Interlaboratory root elongation testing of toxic substances on
selected plant species. NTIS, PB 83-226. U.S. Environmental Protection Agency,
Environmental Protection Agency, Washington, D.C.
TA2.1 - 4
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Thomas, J.M., and J.E. dine. 1985. Modification of the Neubauer technique to assess
toxicity of hazardous chemicals. In: soils. Environ. Toxicol. Chem. 4:201-207.
TA2.1 - 5
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TA2.2. Early seedling survival and vegetative vigor tests.
Early seedling survival and vegetative vigor tests have been briefly outlined by Hoist
and Ellwanger (1982) in response agrichemical registration requirements. The Food
and Drug Administration (FDA 1987), the Code of Federal Regulations (CFR1985) and
the Organization for Economic Cooperation and Development (OECD 1984) have also
provided guidance for conducting these tests. Early seedling survival and vegetative
vigor tests frequently require longer duration exposures that range from 20 to 90 days,
depending upon test species.
Test summary: Early seedling survival and vegetative vigor tests. Replicate
exposures should be set up using soils collected on-site from identified sampling
locations. Plants identified for testing should be selected to meet the site-specific data
needs (e.g., commercial seeds or native seeds), and should be grown under
greenhouse or environmental chamber conditions specified by their species
requirements. For example, four inch plastic pots can be used as exposure units, with
each being filled with test soil (site and reference samples; control soil). Then, these
pots could be placed in plastic growing trays previously filled with growth media
(e.g., Hoagland's). Growth media levels should be monitored daily to assure that all
soils have adequate nutrients. Plants should be grown under ambient conditions in
the greenhouse, and monitored daily under exposure termination. Supplemental
lighting may be required to ensure sufficient photosynthetically active radiation under
specified lighting regimens. Growth conditions, e.g., temperature and humidity,
should be recorded daily as well as any additional exposure conditions that are critical
to successful completion of the test.
As a minimum data collection, at test termination plant leaves and roots should be
collected from each exposure and control replicate, and total biomass should be
recorded as an endpoint for assessing plant vigor. Supplemental endpoints may also
be defined during the problem formulation phase of an ecological effects study design
(e.g., physiological and morphological indicators of plant health). In order to
adequately interpret test endpoints, soil samples should be split after being prepared
for testing and submitted for physicochemical characterization (e.g., soil moisture and
pH, textural analysis, total nitrogen and total organic matter, and cation exchange
capacity).
Intended use: Vegetative vigor and early seedling survival tests are designed to
extend the information gather using short-duration phytotoxicity tests (see TA2.1).
For example, seed germination tests and root elongation tests may not adequately
show chronic effects that are potentially associated with low concentration,
TA2.2 -1
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environmental contaminant exposures, and vegetative vigor and early seedling
survival tests/ owing to their longer duration, may yield adequate data to determine
these longer-term effects. Vegetative vigor and early seedling survival tests may also
be designed to address site-specific questions related to contaminant uptake into plant
tissues, if chemical analytical data are collected concurrent with harvest data.
Previous applications/regulatoryprecedence: Within a regulatory setting, various
agencies have outlined the requirements and specifications for vegetative vigor and
early seedling survival tests (Hoist and Ellwanger 1982; OECD 1984; FDA 1987).
Requirements for development and implementation: Despite their regulatory
requirements, few technical support laboratories are currently providing vegetative
vigor and early seedling survival tests; thus, implementing these within an ecological
effects assessment for Superfund may be difficult owing to an absence of experienced
testing services.
Potential problems and limitations: As with short-term phytotoxicity tests, the
comparative data base is sparse, and in routine practice north-temperate, agricultural
species, particularly grasses and legumes (Fletcher, et al. 1988; Fletcher, et al. 1990;
Gorsuch 1990) are tested. Similarly, little information is available for less commercially
important native plants, including woody species.
References:
CFR [Code of Federal Regulations]. 1985. Rules and regulations; Section 797.2800,
Early seedling growth toxicity test. September 27,1985. CFR 50 (188):39391-39393.
FDA [Food and Drug Administration]. 1987. Sections 4.06 [Seed germination and
root elongation]; 4.07 [Seedling growth]. In Environmental Assessment Technical
Assistance Handbook. NTIS, PB 87-175345. U.S. Food and Drug Administration,
Washington, D.C.
Fletcher, J.S., F.L. Johnson, and J.C. Me Farlane. 1988. Database assessment of
phytotoxicity data published on terrestrial vascular plants. Environ. Toxicol. Chem.
7:615-622.
TA2.2 - 2
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Fletcher, J.S., F.L. Johnson, and J.C. Me Farlane. 1990. Influence of greenhouse
versus field testing and taxonomic differences on plant sensitivity to chemical
treatment. Environ. Toxicol. Chem. 9:769-776.
Gorsuch, J.W., R.O. Kringle, and K.A. Robillard. 1990. Chemical effects on the
germination and early growth of terrestrial plants. In Plants for Toxicity
Assessment, ASTM STP 1091. W. Wang, J.W. Gorsuch, and W.R. Lower, Eds.,
American Society for Testing and Materials. Philadelphia, PA. Pp. 49-58
Hoist, R.W. and T.C. EUwanger. 1982. Pesticide assessment guidelines, Subdivision
J, Hazard evaluation: nontarget plants. 540/9-82/020. Office of Pesticides and Toxic
Substances, U.S. Environmental Protection Agency, Washington, D.C.
OECD [Organisation for Economic Co-Operation and Development]. 1984. OECD
guidelines for testing of chemicals. Director of Information, OECD. 2, rue Andre
Pascal, 75775 Paris Cedex 16, France.
TA2.2 - 3
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TA2.3. Laboratory evaluations with rooted aquatic plants.
Superfund sites situated on wetland soils frequently complicate phytotoxicity
assessments, owing to the saturated character of their soils. Wetland soils may
resemble sediments in many respects, particularly when seasonal, or ephemeral,
climatic conditions alter soil water holding capacity, which may confound
interpretations of germination and growth responses in standard plant testing species
(e.g., buttercrunch lettuce, Lactuca sativa). Standardized rooted aquatic plant toxicity
tests, however, have been developed and should be considered on a site-specific basis
for hydric soils and freshwater/estuarine sediment evaluations. The most well
developed method uses Hydrilla verticillata, which is outlined below. Additional test
methods using sago pondweed (Potamogeton pectinatus) may also be valuable in
evaluating wetland soils or sediments, but are less well developed (see TA4.5).
Test method summary: Hydrilla verticillata test. Both screening tests or definitive
tests using soil or sediment dilutions may be completed with the standardized test
system. Regardless of screening or definitive test application, each sediment or hydric
soil sample should be evaluated in triplicate with three plants per jar following a 14-
day exposure. Specifically, immediately before exposures are initiated, root and shoot
lengths should be measured and recorded for each plant being tested. After
determining initial root and shoot length, the plants should then be placed into the
test sediments and incubated. Briefly, these exposures require that sediment samples
be equally divided among six 1-liter, acid-washed Mason* jars to yield 2 cm to 4 cm of
sediment in each. To reduce sediment resuspension, each sediment surface should be
covered with a Teflon8 sheet bearing three equally spaced holes. In each MasonR jar
replicate known-aged plants have their roots inserted through each of the holes, and
500 ml of autodaved 10% Hoagland's medium amended with 200 mg/1 NaHCO3 is
gently added to each jar. After each jar is loaded with test sediment, nutrient
solution, and plants, Petri dish lids are placed over the jar openings. Test incubation
occurs in an environmental chamber under controlled temperature (25 + 1°C) with
continuous cool white fluorescent light (40 juE m"2sec"2). At a minimum, test
endpoints may include estimates of shoot and root growth, as well as biological
markers indicative of sublethal contaminant effects (Byl and Klaine 1991). For
example, after four days incubation, one plant from each jar may be removed for
chlorophyll a analysis, and after seven days a second plant from each jar may be
harvested for peroxidase activity measurements. Then at test termination (14 days),
the remaining plant should be harvested and root and shoot lengths recorded.
Test method summary: Potamogeton pectinatus test. In recognizing the potential
extrapolation error associated with ecological interpretation from a single-species test,
TA2.3 - 1
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additional aquatic plant tests that may be amenable to wetland assessments have been
developed/ and when applicable should be considered with the ecological effects
assessment. Sago pondweed (Potamogeton pectinatus) is a native north temperature
species/ and the genus is widespread and occurs in various regions of the country.
Fleming/ et al. (1991) described a laboratory test, initially working with agrichemicals,
that is technically similar to the Hydrilla vertidllata test described previously. Again/
biomass is the toxirity endpoint most readily measured/ and for the test system using
sago pondweed exposures generally occur over a four-week test period. While
various culture systems have been used in testing P. pectinatus (e.g., heterotrophic/
microcosm, and autotrophic culture systems)/ the method is not fully standardized. If
the sago pondweed test is included as a component of an ecological effects
assessment, technical details should developed from one of these culture systems in
the site-specific sampling and analysis plan. Also, one advantage of the sago
pondweed test is, that unlike H. vertidllata, members of the genus Potamogeton are
native to the United States, and parallel in situ testing could be incorporated into the
ecological effects assessment to address the "laboratory-to-field"extrapolation error
that could result from laboratory testing alone.
Intended use: Tests with emergent vegetation should be considered in wetland
evaluations that are designed within a Superfund ecological assessment. These
methods are intended to to be independent of standard short-term phytotoxicity tests
(see TA2.1) although the comparative data derived from parallel testing efforts would
benefit data interpretation within the ecological assessment process for Superfund.
Previous applications/regulatoryprecedence: Emergent plant tests have had
previous application in evaluating contaminated sediments, and similarly have a
relatively well established toxicLty data base in the literature for selected contaminants
(e.g. Byl and Klaine 1991).
Requirements for development and implementation: Few technical support
laboratories are currently providing tests with these organisms, and implementing this
biological assessment within an ecological effects assessment for Superfund may be
difficult owing to an absence of experienced testing services.
Potential problems and limitations: Technically, these tests are both relatively
straight forward, but facilities and experienced personnel will be limiting as far as
routine testing is concerned. While the number of test species is relatively limited and
the comparative toxicity data base sparse, the increasing awareness of the ecological
importance of wetland habitats should support site-specific laboratory testing with one
or both of these species.
TA2.3 - 2
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References:
Byl, T.D. and S.J. Klaine. 1991. Peroxidase activity as an indicator of sublethal stress
in the aquatic plant Hydrilla vertidllata (Royle). In Plants for Toxicity Assessment:
Second Volume. ASTM STP 1115, J.W. Gorsuch, W.R. Lower, W. Wang, and M. A.
Lewis, Eds., American Society for Testing and Materials. Philadelphia, PA. Pp. 101-
106.
Fleming, W.J., M.S. Ailstock, J.J. Momot, and C.M Norman. 1991. Response of sago
pondweed, a submerged aquatic macrophyte, to herbicides in three laboratory culture
systems, to Plants for Toxicity Assessment: Second Volume. ASTM STP 1115,
J.W. Gorsuch, W.R. Lower, W. Wang, and M. A. Lewis, Eds., American Society for
Testing and Materials. Philadelphia, PA. Pp. 267-275.
Gorsuch, J.W., R.O. Kringle, and K. A. Robillard. 1990. Chemical effects on the
germination and early growth of terrestrial plants, to Plants for Toxicity
Assessment, ASTM STP 1091. W. Wang, J.W. Gorsuch, and W.R. Lower, Eds.,
American Society for Testing and Materials. Philadelphia, PA. Pp. 49-58
TA2.3 - 3
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TA2.4. Laboratory evaluations with wetland plants.
Freshwater marsh plants may be used to evaluate sediments or hydric wetland soils as
outlined by Walsh, et al. (1991). The method described below was originally designed
to test single toxicants or defined chemical mixtures in defined media/ but it can be
modified to test field-collected sediments or wetland soils that may be appropriate to
wetland evaluations for Superfund. In general the method utilizes rooted marsh
plants and evaluates the effects of contaminated soils and sediments on early seedling
growth and survival.
Summary test method: Echinochloa crusgalli is the marsh plant specifically identified in
the test procedure, but alternative marsh plants (e.g., Spartina alterniflora) may be
identified on a site-specific basis and tested, provided the selected plants are amenable
to the test format outlined. In the test, after being collected in the field, sediments or
wetland soils are mixed, then loaded into exposure containers (e.g., Styrofoam™
cups). Seedlings of E. crusgalli, germinated from seeds four days prior to testing, are
planted in each replicate cup, and incubated under controlled environmental
conditions (e.g., 16:8 lighfcdark cycle at approximately 35/iE/m2/s). After 14 days
exposure, plants are counted and weighed to measure survival and growth endpoints.
For interpretation of ecological effects, methods are outlined for designing reference
sediments or wetland soils, if natural reference materials are not available (Walsh, et
al. 1990). Also, physicochemical characterization of the materials being tested is
outlined and described to help minimize confounding factors that may exert biological
effects independent of contaminants.
Intended use: When originally designed, marsh plant testing was primarily for
complex effluent evaluations, as well as FIFRA and TSCA related plant toxicity
assessments. In recognizing the potential routes of exposure and receptors identified
in the exposure assessment, the ecological effects assessment within Superfund
presents similar chemical mixtures that are directly or indirectly (e.g., through surface
runoff) impacting wetland plants.
Previous applications/regulatoryprecedence: Testing with marsh plants has only
recently been fully developed, and the data base as far as Superfund applications is
limited.
Requirements for development and implementation: Few technical support
laboratories are currently providing tests with the either freshwater or estuarine marsh
plants, and implementing this biological assessment within an ecological effects
TA2.4 - 1
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assessment for Superfund would require contract support or experienced testing
services, if available.
Potential problems and limitations: Technically, these tests, whether using
freshwater or estuarine wetland plants/ are both relatively straight forward, yet
ecologically relevant contaminant information can be gained in a relatively short time
period. While the number of test species is relatively limited and the comparative
toxirity data base sparse, the increasing awareness regarding the ecological significance
of wetland habitats should support a consideration of site-specific laboratory testing
with one or both of these species when sampling and analysis plans are developed.
References:
Walsh, G.E., D.E. Weber, L.K. Brashers, and T.L. Simon. 1990. Artificial sediments
for use in tests with wetland plants. Environ. Exper. Botany 30:391-396.
Walsh, G.E., D.E. Weber, T.L. Simon, and L.K. Brashers. 1991. Toxicity tests of
effluents with marsh plants in water and sediment. In Plants for Toxicity
Assessment: Second Volume. ASTM STP 1115, J.W. Gorsuch, W.R. Lower, W.
Wang, and M. A. Lewis, Eds., American Society for Testing and Materials.
Philadelphia, PA. Pp. 517-525.
TA2.4-2
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TA2.5. Laboratory evaluations with upland plants
Primarily in response to the assessment needs associated with land disposal of
dredging materials, the US Army Corps of Engineers Waterways Experiment Station
(WES) has developed a test method for evaluating phytotoxicity and bioaccumulation
potential hi a freshwater plant, the yellow nutsedge (Cyperus esculentus). The method
is applicable to Superfund site ecological effects assessments, and can be used in
either flooded wetland or upland habitats. From an ecological perspective, the test
evaluates toxicity endpoints (e.g., growth) that may directly relate to field observations
regarding plant cover or vegetative vigor.
Test method summary: The test method (WES 1989; Folsom and Price 1991) was
originally designed to evaluate field-collected dredge materials. Generally, test
preparations involve a physicochemical characterization of the sediment or wetland
soils which are mixed prior to being loaded into exposure containers (e.g., 6-8 L
greenhouse pots). For testing, tubers of C. esculentus are sprouted 5-7 days prior to
initiating the test, then planted into the greenhouse pots. Exposures can occur under
flooded (reduced) or upland (oxidized) conditions. If site-specific characteristics
warrant, exposures may be completed using both approaches. The replicate pots used
for plantings of C. esculentus are then allowed to incubate for 45 days in a greenhouse
or environmental chamber under controlled conditions (e.g., 32 + 2°C daylight
temperatures under 1,200 /iE/m2 photosyntheticallyactive radiation; 21 + 2°C night
temperatures). After 45 days the plants are harvested and total weight (i.e., growth
endpoint) determined. Plant material may also be analyzed for contaminant uptake
and incorporation into plant tissues.
Intended use: The methods described for yellow nutsedge were originally intended
for contaminant evaluations for dredge materials in either a flooded or upland
condition. Again, in view of the varied habitats potentially of interest to ecological
effects assessments in Superfund, either flooded or upland test systems would
potentially be applicable.
Previous applications/regulatoryprecedence: Yellow nutsedge testing, while well
described and applied hi the USCOE dredge materials program, is relatively
unexploited for Superfund ecological assessments. When the role of plant toxicity
assessment is considered, not only within the Superfund ecological effects assessment,
but throughout a site's life history, various remedial options such as sediment
dredging may suggest that testing with Cyperus esculentus be incorporated into site
management plans.
TA2.5 - 1
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Requirements for development and implementation: Few technical support
laboratories are currently providing tests with yellow nutsedge, and implementing this
biological assessment within an ecological effects assessment for Superfund may be
difficult owing to an absence of experienced testing services.
Potential problems and limitations: Beyond those limitations stemming from lack of
technical support/ problems in site-specific interpretation of ecological effects
associated with potential responses in the test would center upon "laboratory-to-field"
extrapolation error and interspecies variability with respect to contaminant-mediated
adverse effects.
References:
Folsom, Jr., B.L. and R. A. Price. 1991. A plant bioassay for assessing plant uptake of
contaminants from freshwater soils or dredged material. In Plants for Toxicity
Assessment: Second Volume. ASTM SIP 1115, J.W. Gorsuch, W.R. Lower, W.
Wang, and M. A. -Lewis, Eds.," American Society for Testing and Materials.
Philadelphia, PA. Pp. 172-177.
WES. (U.S. Army Corps of Engineers Waterways Experiment Station). 1989. A plant
bioassay for assessing plant uptake of heavy metals from contaminated freshwater
dredged material. Technical Note EEDP-04rll. U.S. Army Engineer Waterways
Experiment Station, Vicksburg, MS.
TA2.5 - 2
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TA2.6. Alternative test species in seed germination, root elongation, and early
seedling survival and vegetative vigor tests.
As suggested by the tests previously outlined (TA2.1-TA2.5)test endpoints frequently
are similar (e.g., growth, germination) in these procedures, but the species being
tested differ. In part these differences reflect soil matrix characteristics that might limit
the success of any given test system, and recognize that site-specific characteristics
may suggest alternative test species. For example, lettuce seed is frequently used in
seed germination tests, but some soils may not be amenable to testing with a
domesticated species selected for optimal growth in a particular soil matrix.
Contaminant effects and matrix effects may potentially be confounded when the life
history characteristics of test species preclude or potentially limit its usefulness in any
given phytotoxicity test method. Additionally, for interpretation of site-specific
ecological effects, the support of a comparative toxicity database may be insufficient
within a risk assessment context. Thus, more relevant test species may be beneficial
to an ecological effects assessment, and measurement endpoints (e.g., survival and
growth) used to evaluate relationships between ecological indicators and soil toxicity
may be considered using methods modified for tests using alternative species.
Test method summary: Methods to evaluate seed germination using a various
species of plant seeds (agricultural crops, vegetables and herbs, flowers, and trees and
shrubs) are briefly summarized by the Association of Official Seed Analysts (AOSA) in
their Rules for Testing Seeds (AOSA 1990). Here, for example, exposure conditions
specific to various species are tabulated, including suggested substrates and optimum
incubation temperatures for germination testing, as well as test duration specifications.
Furthermore, special pretreatment of native seeds, for example, prechilling or
scarification, is also specified, and methods for distinguishing between non-
germinated seeds and nonviable seeds are identified (e.g., tetrazolium and embryo
excision tests). On a site-specific basis, these alternative test species may be more
conducive to ecological interpretation, especially when soil matrix effects potentially
confound contaminant effects on seed germination and emergence.
Intended use: Testing with alternative species is intended to yield site-specific data,
or at a minimum, more closely address the problems that develop when ecological
interpretations are derived from testing with standard species (see TA2.1). While the
comparative toxicity data base is developing, test endpoints related to growth and
germination in non-standard test species may in part reflect site-specific soil matrix
characteristics that might limit the success of standard tests.
TA2.6 -1
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Previous applications/regulatoryprecedence: Alternative test species have found
limited use within plant toxirity assessments for Superfund (e.g., Linder, et al. 1990),
but the increasing awareness that tests using standard plant species may be limited
within ecological contexts (Gorsuch, et al. 1990; Fletcher, et al. 1990; Fletcher, et al.
1988; Royce, et al. 1984) should encourage use of alternative plant testing in ecological
effects assessments.
Requirements for development and implementation: Few technical support
laboratories are currently providing tests with standard test species, so the use of
alternative test species as part of a biological evaluation within an ecological effects
assessment for Superfund may be technically limited.
Potential problems and limitations: While lack of testing services may limit
alternative species testing, if adequate technical support were gained, site-specific
ecological effects assessments could potentially yield empirical data that would reduce
uncertainty associated with interspecific extrapolation errors, which presently are likely
in view of the sparse comparative toxicity database.
References:
AOSA [Associationof Official Seed Analysts]. 1990. Rules for testing seeds. J. Seed
Tech. 12:1-122.
Fletcher, J.S., F.L. Johnson, and J.C. Me Farlane. 1990. Influence of greenhouse
versus field testing and taxonomic differences on plant sensitivity to chemical
treatment. Environ. Toxicol. Chem. 9:769-776.
Fletcher, J.S., F.L. Johnson, and J.C. Me Farlane. 1988. Database assessment of
phytotoxicity data published on terrestrial vascular plants. Environ. Toxicol. Chem.
7:615-622.
Gorsuch, J.W., R.O. Kringle, and K. A. Robillard. 1990. Chemical effects on the
germination and early growth of terrestrial plants. In Plants for Toxicity
Assessment, ASTM STP 1091. W. Wang, J.W. Gorsuch, and W.R. Lower, Eds.,
American Society for Testing and Materials. Philadelphia, PA. Pp. 49-58
Under, G., J.C. Greene, H. Ratsch, J. Nwosu, S. Smith, and D. Wilborn. 1990. Seed
germination and root elongation toxicity tests in hazardous waste site evaluation:
TA2.6 - 2
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methods development and applications. IQ Plants for Toxirity Assessment, ASTM
STP 1091. W. Wang, J.W. Gorsuch, and W.R. Lower, Eds., American Society for
Testing and Materials. Philadelphia, PA. Pp. 177-187.
Royce, C.L., J.S. Fletcher, P.G. Risser, J.C. Me Farlane, and F.E. Benenati. 1984.
PHYTOTOX: a database dealing with the effect of organic chemicals on terrestrial
vascular plants. J. Chem. Inf. Comput. Sci. 24:7-10.
TA2.6 - 3
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TA2.7. Short-term tests for evaluating whole plant toxitity.
Depending upon site conditions, soil contamination evaluations could be complete
using hydroponic exposure systems/ and endpoints complementary to those measured
in the root elongation test could be determined. However, root elongation tests do
not evaluate soil contaminant effects on ecologically significant endpoints related to
photosynthesis, transpiration, flower initiation, and fruit development. A five-day,
whole plant test [TOXSCREEN] has recently been described, however, and its
potential for use in ecological effects assessment may be warranted, especially if the
soil contaminants of concern are highly mobile and water soluble. The test uses
whole plants, for example, soybean (Glycine max), barley (Hordeum vulgare) and woody
perennials, that have been grown in hydroponic culture for at least 28 days. These
28-day old plants are used in a hydroponic exposure system where the root
environment is exposed to chemicals in solution for three to five days. Following
these exposures to water-soluble contaminants, plant yields (e.g., root and shoot
growth, total biomass) are measured. Depending upon the study design, additional
endpoints could be measured, for example, contaminant uptake and transpiration
rates. The method can be used as a screening test and evaluate the indirect effects of
site-soil on the basis of eluates derived from soil samples, and if warranted, definitive
tests with site-specific chemical mixtures could be completed to help discriminate, for
example, contaminant effects from non-contaminant related soil effects.
Test method summary: As a screening test with whole plants, TOXSCREEN is
relatively short in duration, especially in relationship to the life history of test species,
since the test considered here routinely requires no less than 30 to 40 days to conduct.
Also, the method is amenable to testing various plant species, including young woody
perennials, and more definitive tests have been designed using very similar technical
approaches (Me Farlane, et al. 1990).
Briefly, species identified for testing are germinated, then grown to a specified size in
hydroponic containers (Me Farlane and Pfleeger 1987). Alternatively, known-age
plants may be purchased commercially (e.g., woody species) and adapted for
hydroponic exposures (Me Farlane, et al. 1990). Regardless of the plant source, once
plants have attained the specified size, exposures are relatively short (e.g., three to
five days), depending upon test species. Exposures occur hydroponically when
young plants are transferred from nursery containers to containers filled with
exposure media. Tests are routinely conducted under 16/8 light/dark regimen (light
intensity of 350 pmol m'V1 at top of canopy) at 25/21 ± 2°C and relative humidity
between 50 to 70%. When appropriate, solvent systems may be used as carriers, for
example, when rhizosphere exposures are designed to reflect site-specific conditions.
TA2.7-1
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Soil eluates may also be used in fining exposure containers. Toxicity endpoints
measured upon termination would routinely include survival and growth (Pfleeger, et
d. 1991), although exposure systems could be designed that allow additional
measurements for estimating sublethal effects (Me Farlane, et al. 1990).
Intended use: TOXSCREEN should be considered primarily as a screening test,
particularly if contaminants of concern are water soluble and conducive to hydroponic
exposures. Depending upon site-specific characteristics, target analytes could be used
in single compound or defined chemical mixture exposures, and if sufficient soil were
collected, eluates could be used as the exposure medium.
Previous applications/regulatoryprecedence: As with various methods summarized
in this compendium, TOXSCREEN was originally designed with regulatory
applications being the central focus (Me Farlane and Pfleeger 1985). TOXSCREEN,
then, should be considered in developing sampling and analysis plans, depending
upon site-specific contingencies.
Requirements for development and implementation: Few technical support
laboratories are currently providing tests with these organisms, and implementing this
biological assessment within an ecological effects assessment for Superfund may be
difficult owing to an absence of experienced testing services.
Potential problems and limitations: If adequate facilities and technical support are
available, the test exposure is relatively short; however, adequate technical
considerations must be made to assure that plant materials are available for testing
(e.g., hydroponic nursery facility or commercial sources). Owing to its recent
description, TOXSCREEN is not commercially available.
References:
Pfleeger,T., C. Me Farlane, R. Sherman, and G. Volk. 1991. A short-term bioassay
for whole plant toxicity. In Plants for Toxicity Assessment: Second Volume, ASTM
STP 1115. J.W. Gorsuch, W.R. Lower, W. Wang, and M.A. Lewis, Eds. American
Society for Testing and Materials, Philadelphia, PA. Pp. 355-364.
Me Farlane, J.C., T.G. Pfleeger, and J.S. Fletcher. 1990. Effect, uptake, and
disposition of nitrobenzene in several terrestrial plants. Environ. Toxicol. Chem.
9:513-520.
TA2.7-2
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Me Farlane, J.C. and T. Pfleeger. 1987. Plant exposure chambers for study of toxic
chemical-plant interactions. J. Environ. Qual. 16:361-371.
Me Farlane, J.C. and T. Pfleeger. 1985. Plant exposure laboratory and chambers.
6010/3-85/000. U.S. Environmental Protection Agency, Environmental Research
Laboratory, Corvallis, OR.
TA2.7-3
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TA2.8. life-cycle tests using vascular plants.
Two hydroponic test systems using plants are potentially applicable to ecological
effects assessments for Superfund sites. Again, as with TOXSCREEN (see TA2.7),
water-soluble constituents of waste site chemical mixtures may be evaluated with
either Arabidopsis thaliana (Ratsch, et al. 1986) or Brassica rapa (Shimabuku, et al. 1991).
Both tests measure sublethal endpoints as well as acute toxirity. For example,
ecologically significant endpoints related to reproduction (e.g., seed set) and growth
(e.g., biomass) are routinely measured in both tests. While exposure periods
(approximate seed-to-seed life cycle; 28-36 days for Arabidopsis thaliana and 36-44 days
for Brassica rapa) are longer than those used in seed germination and root elongation
testing (120 hours), the information gained regarding subtle contaminant-associated
effects may justify the longer time required for these tests.
Test methods summary: Both tests were designed for hydroponic exposure systems,
and single-chemicals or chemical mixtures to be evaluated are incorporated into the
test system's nutrient solution. Again, site-soils are not directly tested with either
plant system, but water-soluble extracts, or eluates, may be used in the test systems.
While relatively large volumes of eluate may be required for these hydroponic systems
relative to that volume used in the standard root elongation test, defined chemical
mixtures similar to those noted in site-soils could be used as an alternative exposure
system, and could be incorporated into the test system's nutrient solution.
For testing with Arabidopsis thaliana, hydroponic exposures occur in double-pot, static-
replacement systems where a venniculite-filledgrowth container is nested above a
second larger pot that serves as a nutrient solution reservoir. Nutrients and water
move from the nutrient reservoir to the vermiculite via polyester wicks that are draped
between the two pots. Seeds are uniformly planted on the surface of the vermiculite,
and greenhouse conditions or large growth chambers (e.g., minimum 16-hr
photoperiod) assure similar growing conditions for all plants. Depending upon the
exposure period and growth conditions, plants will set seeds and mature, and at test
termination various endpoints can be measured in all plants. For example, a typical
data collection from the A. thaliana test would include total biomass, as well as
vegetative and reproductive biomass (e.g., stems and leaves versus seeds and
reproductive structures). Morphological observations regarding leaf and flower
structure could also be scored as an endpoint indicative of plant health and subtle
reproductive effects associated with exposure. Similar test systems are used with
Brassica rapa. For B. rapa tests, seeds are planted in exposure pots containing growth
media (e.g., greenhouse potting soil). These exposure pots are then nested within a
nutrient solution pot, and static-replacement exposures occur via nutrient solutions
TA2.8 -1
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that potentially include single-chemicals, denned chemical mixtures, or site-soil
eluates. Measurement endpoints at test termination (usually at Day 36 to Day 42)
include morphological (e.g., biomass and foliar height) and phenological (e.g./ initial
flowering date) data. Time-to-flowering can also be a measurement endpoint/ since
subtle contaminant effects may alter the onset of flower development and
reproduction.
Intended use: Similar to TOXSCREEN, both of these full life-cycle tests are intended
to address toxicity endpoints that are inadequately considered in standardize plant
tests measuring seed germination and root elongation (see TA2.1). Again, exposures
are hydroponic, and contaminant water solubility may limit exposures for some
chemicals. Also/ if eluates are used in regard to soil contaminant as sources for
potential groundwater and rhizosphere contamination/ a direct measure of "worst
case" can be addressed using these systems.
Previous applications/regulatory precedence: Little regulatory work has been
routinely completed using either of these methods; however, chemicals of interest to
regulatory programs driven by TSCA and FIFRA have been considered in these test
systems which have proven effective (Ratsch, et al. 1986).
Requirements for development and implementation: Few technical support
laboratories are currently providing tests with these organisms, and implementing this
biological assessment within an ecological effects assessment for Superfund may be
difficult owing to an absence of experienced testing services.
Potential problems and limitations: Technically, chemicals of concern should be
water soluble, or at least reach concentrations in water that are significant within a
site-specific context.
References:
Ratsch, H.C., D.J. Johndro, and J.C. Me Farlane. 1986. Growth inhibition and
morphological effects of several chemicals in Arabidopsis thaliana (L.) Heynh. Environ.
Contam. Toxicol. 5:55-60.
Shimabuku, R.A., H.C. Ratsch, C.M. Wise, J.U. Nwosu, and L.A. Kapustka. 1991.
A new plant life-cycle bioassay for assessment of the effects of toxic chemicals using
rapid cycling Brassica. In Plants for Toxicity Assessment: Second Volume, ASTM
TA2.8 - 2
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STP 1115. J.W, Gorsuch, W.R. Lower, W. Wang, and M.A. Lewis, Eds. American
Society for Testing and Materials, Philadelphia, PA. Pp. 365-375.
TA2.8 - 3
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TA2.9. Plant tissue culture tests.
While relatively untapped as a technical information source for Superfund
assessments, various plant cell and tissue culture techniques have been developed for
evaluating ecological effects of chemical on plants (Wickloff and Fletcher 1991; Harms
and Langebartels 1986; Ebing, et al. 1984). Both callus and cell suspension cultures of
various plant species have been used in evaluating subacute chemical effects in plants,
primarily by addressing the metabolic fate of xenobiotics (e.g., herbicides) in plants;
but, little correlative work has been completed to address the ecological interpretation
of these in vitro plant cell and tissue culture methods.
Test method summary: For evaluating subacute effects, particularly chemical-related
alterations in plant metabolism, plant cell and tissue culture techniques have become
well developed over the past ten years. Frequently, these test methods involve
suspension cultures of commercially important plant species like agriculture crops,
e.g, soybean (Glydne max) and wheat (Triticum aestivum), that are exposed to
contaminants added to the culture nutrient medium. Exposures pertinent to an
ecological effects assessment for Superfund would require that eluates be prepared
from site-soil, and the indirect effects of soil contaminants could then be evaluated by
supplementing the test nutrient medium with eluate spikes. Alternatively, if the
contaminant history for the site was reliable, or if analytical information regarding soil
contaminants was available, defined chemical mixtures could be added as
supplements to the nutrient medium and subtle metabolic effects could be
determined. While cell and tissue culture test systems were not designed to evaluate
acute effects, the methods may be modified to address survival and growth in vitro.
Similar test methods have been described using callus cultures (Zilkah and Gressel
1977a,c; Zilkah, et al. 1977), and limited correlation studies have been completed that
suggest that photosyntheticallyactive callus cultures (e.g., Zilkah and Gressel 1978),
like cell suspension cultures, may be good surrogates for evaluating phytotoxicity,
especially in plants (e.g., terrestrial woody species) that are not conducive to whole
plant tests.
Intended use: The comparative toxicity data base is limited with respect to in vitro
toxicity and its interpretation to whole plant toxicity, and similar limitations are
apparent when ecological effects are being evaluated at a site. At present, whole plant
tests may be considered complementary methods that should be considered parallel to
these in vitro methods that may be beneficial to describing potential long-term effects
associated with altered plant functional capacities.
TA2.9 - 1
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Previous applications/regulatoryprecedence: The method(s) outlined and
summarized here are relatively early in the standardization process, and are not
intended to be "stand alone" tests. Rather/ the strengths of these method(s) lie in
their contribution to evaluating phytotoxicity in species which are difficult to assess
with whole plant tests. By using field surveys in conjunction with a comparative
toxicity approach, weight of evidence arguments regarding site-specific receptors may
be developed that illustrate adverse effects associated with soil exposures, e.g.
diminished vigor in woody shrubs or poor reproductive performance in forbs.
Requirements for development and implementation: Few technical support
laboratories are currently providing tests with these organisms, and implementing this
biological assessment within an ecological effects assessment for Superfund may be
difficult owing to an absence of experienced testing services.
Potential problems and limitations: The sparse comparative toxicity data base and
the potentially problematic interpretation of ecological effects based solely on in vitro
studies may restrict the methods that are available using plant cell and tissue culture.
References:
Ebing, W., A. Haque, I. Schuphan, H.Harms, C. Langebartels, D. Scheel, K.T. von
der Trenck, and H. Sanderman. 1984. Ecochemical assessment of environmental
chemicals: draft guideline of the test procedure to evaluate metabolism and
degradation of chemicals by plant cell cultures. Chemosphere 13:947-957.
Harms, H. and C. Langebartels. 1986. Standardized plant cell suspension test
systems for an ecotoxicologic evaluation of the metabolic fate of xenobiotics. Plant Sri.
45:157-165.
Widkloff, C. and J.S. Fletcher. 1991. Tissue culture as a method for evaluating the
biotransformationof xenobiotics by plants. In Plants for Toxicity Assessment:
Second Volume, ASTM STP 1115. J.W. Gorsuch, W.R. Lower, W. Wang, and M. A.
Lewis, Eds. American Society for Testing and Materials, Philadelphia, PA. Pp. 250-
257.
Zilkah, S. and J. Gressel. 1978. Correlations in phytotoxicity between white and
green calli of Rumex obtusifolius, Nicotiana tabacum, and Lycopersicon esculentum.
Pesticide Biochem. Physiol. 9:334-339.
TA2.9 - 2
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Zilkah, S. and J. Gressel. 1977a. Cell cultures vs. whole plants for measuring
phytotoxicity. I. The establishment and growth of callus and suspension cultures;
definition of factors affecting toxirity on calli. Plant & Cell Physiol. 18:641-655.
Zilkah, S. and J. Gressel. 1977b. Cell cultures vs. whole plants for measuring
phytotoxicity. EL Correlations between phytotoxidtiesin cell suspension cultures,
calli, and seedlings. Plant & Cell Physiol. 18:815-820.
Zilkah, S., P.P. Bocion, and J. Gressel. 1977. Cell cultures vs. whole plants for
measuring phytotoxicity. n. Correlations between phytotoxicity in seedlings and calli.
Plant & Cell Physiol. 18:657-670.
TA2.9 - 3
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TA2.10. Tests evaluating plant community structure.
Within an ecological effects assessment, plants may potentially be critical receptors
and indicators of adverse biological effects, including acute or chronic phytotoxicity,
associated with chemical exposures (Weinstein and Laurence 1989; Weinstein, et al.
1990). Impacts on plant communities associated with the release of chemicals from
hazardous waste sites may be assessed using defined plant communities grown in
raised beds, or by using reference soil containing a seed bank (Pfleeger 1991). With
these mesocosms, representative of larger systems/ being constructed in either
greenhouse or field settings/ percent cover by plant species can be monitored over
time using nested neighborhoods analysis based upon site-specific target species/ e.g./
native grasses/ potential ground covers. Community biomass can also be measured
following a prescribed exposure period. Exposures will vary with respect to duration
(e.g./ three to nine months)/ because plant communities will vary site-specifically
depending upon regional characteristics (e.g./ native plant species composition when
initiating test from seed bank). Additionally/ relative species abundance/ species
dominance and community structure may be regarded as measurement endpoints in
the test. Interspecific competition can be analyzed within the context of contaminant
concentration/ and temporal changes that may be apparent due to life-stage dependent
contaminant effects can be noted. If defined species mixes are used in conducting the
test/ the analysis of artificial plant communities can be a relatively simple and
economical method for observing subtle vegetative impacts associated with
contaminated soil. If seed banks from reference locations are used in the test/ these
naturally occurring species/ like the defined mixes/ will yield a test system reflecting
site-specific heterogeneity, which should be encouraged with respect to ecological
effects assessment.
Test method summary: Various methods may be applicable on a site-specific basis/
but the general format for conducting studies that address community level effects are
similar regardless the specific contaminants of concern. For example/ seed for testing
may be obtained commercially from regional native seed supply sources/ and the plant
community to be tested can be defined. Alternatively/ a seed bank may be collected
from a reference location and used in evaluations of site-soil. If seed bank sources are
used/ past land and chemical use should be documented and any confounding effects
owing to the selection of reference area seed bank should be acknowledged. Seed
banks for testing should be collected from the top 5 cm of surface soil at a reference
site/ and should be collected when vegetation cover is minimal (e.g./ seasonal
dormancy). The site-soil and reference seed bank may be sieved (6 mm) and mixed/
and on a site-specific basis/ a commercial potting soil may be used to amend test soils.
Raised beds are typically used as exposure containers and may be located in the field
TA2.10-1
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or in the greenhouse, depending upon the site-specific study design. The beds
should be sufficiently deep to minimize root interactions from adjacent vegetation, and
should be filled with test soil or reference soil to within 5 cm of the top. Defined seed
mixtures or seed bank should then be incorporated into the soil, and depending upon
the study design, irrigation and fertilization can be specified. Each site-specific study
plan may differ in their details for analysis of plant community responses to
contaminated soils, but for waste sites with similar contaminant histories and similar
habitat settings, study designs may be nearly identical.
For analysis of soil toxicity and its effects on vegetation, target plant species may be
identified for specific focus hi the study. Or, ecological endpoints may be identified
for analyzing community-level responses. For example, percent vegetative cover, total
biomass, species diversity and richness may be determined. The level of analytical
detail should be determined initially in the study design. Regardless of the study
design, identification of a reference soil is critical in the evaluation of soil
contamination and its effects on native plants.
Intended use: While concentration-response relationships may be designed as part of
the vegetation evaluation completed with a plant community study, the method may
be more valuable as a screening method complementary to more controlled plant
tests, e.g., vegetativevigor and early seedling survival. By using both an "ecotoxicity
test" like that measuring plant community responses to contaminated soil and an
organismic-leveltest like vegetativevigor and early seedling survival, uncertainty in
the risk characterization for the site may be more adequately addressed on the basis of
site-specific empirical information.
Previous applications/regulatoryprecedence: The method outlined and summarized
here is early in the standardization process, and is not intended to be a "stand alone"
test. Rather, the strengths of the method lies hi its contribution to weight-of-evidence
arguments that are supportive of those methods that clearly illustrate adverse effects
associated with soil exposures, e.g. early seedling survival and vegetativevigor
studies.
Requirements for development and implementation: Few technical support
laboratories are currently providing tests with these organisms, and implementing this
biological assessment within an ecological effects assessment for Superfund may be
difficult owing to an absence of experienced testing services.
Potential problems and limitations: As a field test, or greenhouse test, the plant
mesocosm exposure beds are relatively easy to establish, but the test is time- and
TA2.10-2
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labor-intensive, owing to the real-time growth required for biomass measurements and
data collection and reduction for evaluating community structure.
References:
Pfleeger, T. 1991. Impact of airborne pesticides on natural plant communities. In
Plant tier testing: a workshop to evaluate nontarget plant testing in Subdivision J
Pesticide Guidelines. 600/9-91/041. U.S. Environmental Protection Agency,
Environmental Research Laboratory, Corvallis, OR.
Weinstein, L.H. and J.A. Laurence. 1989. Indigenous and cultivated plants as
bioindicators. In Biologic markers of air-pollution stress and damage in forests.
Committee on Biologic Markers of Air-Pollution Damage in Trees, G.M. Woodwell
(Chair). National Research Council. National Academy Press. Washington, D.C.
Pp. 195-204.
Weinstein, L.H., J.A. Laurence, R.H. Mandl, and K. Walti. 1990. Use of native and
cultivated plants as bioindicators and biomonitors of pollution damage. In W. Wang,
J.W. Gorsuch, and W.R. Lower (Eds.). Plants for Toxicity Assessment. ASTM STP
1091. American Society for Testing and Materials, Philadelphia, PA. Pp. 117-126.
TA2.10-3
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TA2.11. Ambient air exposure systems and phytotoxicity testing.
Few, if any, Superfund ecological effects assessments have directly considered
vegetation effects associated with phytotoxicity resulting from exposure to ambient air
pollutants (e.g., to volatile organic compounds). However, to introduce plant testing
systems that would address these ambient air pathways, various methods are
available and listed here.
Test methods overview: Nearly all the exposure and toxirity test systems available
were initially developed in response to concerns directly associated with regional
ambient air quality (Hogsett, et al. 1987a). While those initial concerns regarding
regional ambient air quality regarded larger spatial scales than the majority of
Superfund sites, there may be site-specific examples when ambient air exposures, for
example^ to formaldehyde or petroleum distillates, may be important within an
ecological effects assessment for Superfund. Exposure systems outlined below would
be amenable to testing various plant species, ranging from grasses and forbs to shrubs
and trees. Detailed summaries of these methods are beyond the scope of the current
compilation, but if ambient air pathways were considered critical in the ecological
effects assessment for a contaminated soil, these exposure systems should be
considered in developing a site-specific sampling and analysis plan.
Test systems for evaluating phytotoxicity associated with ambient air exposure
pathways at Superfund sites (after Hogsett, et al. 1987a).
DRY DEPOSITION SYSTEMS
WET DEPOSITION SYSTEMS
• gaseous exposures using non-
chamber and chamber systems
• non-chamber systems — plume
exposures; air exclusion exposure
• chamber systems — outdoor
exposures; indoor exposures;
cuvette exposures
• non-gaseous or participate
exposures
• dust exposures
• rainfall deposition exposures
(indoor chambers)
• rainfall deposition exposures
(outdoor chambers)
• mist/cloud water exposures
using enclosure chambers
• aerosol exposures using
enclosure chambers
TA2.11 - 1
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While these testing systems may suggest various pathways that may be critical on a
site-specific ecological effects assessment, empirical data collection to address these
questions may not be technically feasible, even if ambient air exposures are of
concern, unless these facilities are available to the regional project manager. However,
within the context of an ecological assessment for any Superfund site, the criteria for
evaluating the contribution of ambient air pathways to exposure should be considered.
In determining whether ambient air exposures are critical to the ecological effects
assessment, various elements influencing exposure should be considered. In general
these elements may be categorized as:
contaminant physicochemical attributes in the atmosphere
soil, habitat, and atmospheric conditions that may influence exposure
and non-exposure periods
• biological attributes of receptors — plant or animal — that may be exposed
via ambient air pathways
Intended use: When site-specific conditions support evaluations of ambient air
exposure pathways, these methods should be considered within the ecological effects
assessment. None have been standardized; however, each has been developed and
applied toward larger-scale ambient air exposures and have contributed to data bases
that support ambient air quality standards.
Previous applications/regulatoryprecedence: The method(s) outlined and
summarized here are well developed but have not been standardized (e.g., Hogsett, et
(d. 1987b). The strengths of these method(s) lie in their potential contribution to
evaluating exposure pathways that generally have not been considered within an
ecological effects assessments.
Requirements for development and implementation: Few, if any, technical support
laboratories are currently providing tests to evaluate ecological effects associated with
contaminated soils and ambient air exposures, so implementing these biological
assessments within an ecological effects assessment for Superfund may be difficult
owing to an absence of experienced testing services.
Potential problems and limitations: Currently, the data base for evaluating or
interpreting ecological effects associated with ambient air exposures at Superfund sites
is poorly developed.
TA2.11 - 2
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References:
Hogsett, W.E., D. Olszyk, D.P. Ormord, G.E. Taylor, Jr., and D.T. Tingey. 1987. Air
pollution exposure systems and experimental protocols. Volume 1: A review and
evaluation of performance. 600/3-87/037a. U.S. Environmental Protection Agency,
Environmental Research Laboratory, Corvallis, OR.
Hogsett, W.E., D. Olszyk, D.P. Ormord, G.E. Taylor, Jr., and D.T. Tingey. 1987. Air
pollution exposure systems and experimental protocols. Volume 2: Description of
facilities. 600/3-87/037b. U.S. Environmental Protection Agency, Environmental
Research Laboratory, Corvallis, OR.
TA2.11 - 3
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TA2.12. Photosynthesis inhibition tests for evaluating sublethal effects in plants.
Photosynthesis in terrestrial and wetland plants is a complex and highly coordinated
physiological process (Miles 1990). Environmental stresses related to soil
contamination may impair portions of the photosynthetic system of resident plants,
and impaired photosynthetic function may be indicated by abnormal fluorescence
patterns relative to plants inhabiting unimpacted soils (e.g., Judy, et al. 1990; Judy, et
al. 1991). Measurement endpoints gained from analysis of these fluorescence patterns
or profiles are the variable and maximum fluorescence values [Fv and F^J, where Fv
is related to photosynthetic capacity and F,^ represents the maximum fluorescence
attained during any given analysis. Ratios derived from these measurements may
subsequently be scored as normal or abnormal based upon literature or reference-
specific values. In conjunction with other measurements of plant health, correlations
may be described between altered ecological endpoints (e.g., altered vegetative cover)
and soil contamination.
Test method summary: Testing may be completed on intact plants or on leaves
collected from the middle to upper regions of the plant. A transportable fluorometer
dedicated to analysis of photosynthesis may be used in the field or at a fixed
laboratory to evaluate potential vegetative impacts associated with contaminated soil.
Fluorometric analysis may suggest whether resident plants display altered
photosynthetic activity in the presence of contaminants occurring in soils or
sediments. For analysis, intact leaves or leaf segments are placed with the adaxial
surface down on the sample chamber window of a fluorometer. After dark adapting
(generally less than two minutes), fluorometric analyses are initiated. Fluorescence
profiles may be plotted on an X-Y recorder, or electronic data may be stored in a data
logger for later analysis. Variable and maximum fluorescence values [Fv and F^J are
measured from these fluorescence profiles, and plant health is in part described on the
basis of derived ratio estimators based on Fv and F,^. While the application of plant
biomarkers like fluorescence is relatively new in ecological effects assessment, if
supported by other phytotoxicityand plant effects methods, fluorometric analysis may
potentially yield a rapid and sensitive method for evaluating soil contamination effects
impacting plant photosynthesis.
Intended use: As one potential plant biomarker, evaluations of plant fluorescence
may act as exposure or effects indicators (Suter 1990). However, in considering the
multimedia exposures to complex chemical mixtures that are characteristic of
Superfund sites, the measurement of fluorescence profiles and ecological
interpretations derived from these measurements should be considered with caution.
TA2.12-1
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Previous applications/regulatoryprecedence: Biomarkers that potentially measure
functional conditions like photosynthesis have been considered previously within the
context of evaluating air-pollution stress (NRC 1989), but no regulatory applications
have directly considered measurements of plant fluorescence beyond a support role in
developing weight of evidence arguments.
Requirements for development and implementation: Few technical support
laboratories are currently providing tests with these organisms, and implementing this
biological assessment within an ecological effects assessment for Superfund may be
difficult owing to an absence of experienced testing services.
Potential problems and limitations: If considered alone, photosynthesis or
fluorescence may express sufficient variability (i.e., diurnal, seasonal, age-related, and
non-contaminant related effects) to confound interpretations regarding adverse
ecological effects (Winner 1989). However, if completed in conjunction with whole-
plant studies, altered fluorescence profiles in exposed plants may lend credence to
interpretations of adverse ecological effects on the basis of site-specific empirical data.
Also, compensatory mechanisms common to plants, e.g., reallocationof resources
between plant organs in response to contaminant stress, may make extrapolation from
leaf-level measurements to whole-plant and plant community-level interpretations
difficult.
The method(s) outlined and summarized here are early in the standardization process,
and are not intended to be "stand alone" tests. Rather, the strengths of these
method(s) lie in their contribution to weight of evidence arguments that are supportive
of those methods that clearly illustrate adverse effects associated with soil exposures,
e.g. early seedling survival and vegetative vigor studies.
References:
Judy, B.M., W.R. Lower, C.D. Miles, M.W. Thomas, and G.F. Krause. 1990.
Chlorophyll fluorescence of a higher plant as an assay for toxicity assessment of soil
and water. M W. Wang, J.W. Gorsuch, and W.R. Lower (Eds.). Plants for Toxicity
Assessment. ASTM STP 1091. American Society for Testing and Materials,
Philadelphia, PA. Pp. 308-318.
TA2.12-2
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Judy, B.M., W.R. Lower, F.A. Ireland, and G.F. Krause. 1991. A seedling
chlorophyll fluorescence toxicity assay. In Plants for Toxicity Assessment: Second
Volume, ASTM STP 1115. J.W. Gorsuch, W.R. Lower, W. Wang, and M.A. Lewis,
Eds. American Society for Testing and Materials, Philadelphia, PA. Pp. 146-158.
Miles, D. 1990. The role of chlorophyll fluorescence as a bioassay for assessment of
toxicity in plants. In W. Wang, J.W. Gorsuch, and W.R. Lower (Eds.). Plants for
Toxicity Assessment. ASTM STP 1091. American Society for Testing and Materials,
Philadelphia, PA. Pp. 297-307.
NRC [National Research Council]. 1989. Biologic markers of air-pollution stress and
damage in forests. National Academy Press, Washington, D.C.
Suter, G.W. 1990. Use of biomarkers in ecological risk assessment. In Biomarkers of
environmental contamination. J.F. McCarthy and L.R. Shugart (Eds.). Lewis
Publishers, Boca Raton, FL. Pp. 419-426.
Winner, W.E. 1989. Photosynthesis and transpiration measurements as biomarkers of
air pollution effects on forests. In Biologic markers of air-pollution stress and damage
in forests. Committee on Biologic Markers of Air-Pollution Damage in Trees, G.M.
WoodweU (Chair). National Research Council. National Academy Press.
Washington, D.C. Pp. 303-316.
TA2.12-3
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TA3. SOIL BIOTA TEST METHODS FOR THE ASSESSMENT OF SOIL
CONTAMINATION AT HAZARDOUS WASTE SITES
-------
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Methods for evaluating contaminant effects on soil biota biomass and diversity.
Nutrient cycling would not occur without organisms to perform the majority of the
critical processes. While abiotic factors influence the presence and activity of
organism, an understanding of the biological factors along with the physical
constraints allows prediction of process rates. Organisms are the filter through which
all nutrients pass; an understanding of their ability to cycle nutrients is critical.
Soil organisms perform many processes, and this redundancy is important to
resiliency of soil following disturbance (Dindal 1990). In healthy soil, there are usually
(but not always) several organism groups which perform any particular process.
Some organisms function in colder conditions, some in warmer, wetter or drier
conditions, higher or lower organic matter, high or low pH, and so forth. Thus rates
of nutrient cycling in a particular ecosystems depend on the organisms currently
performing a function.
Functional redundancy can be destroyed; such destruction has occurred in soils
undergoing desertification (Miller, 1987). The dependency of vegetation on the
presence of mycorrhizal fungi and on a functional soil-organism nutrient cycling
system needs to be quantified. Evidence is accumulating that, at least some plants,
are obligately dependent on symbiotic organisms for establishment or survival (Janos,
1987; Reeves, 1985).
The soil foodweb consists of primary producers, primary consumers, secondary
consumers, consumer generalists and higher level consumers (Coleman, 1985; Dindal,
1990; Foissner, 1986; Freckman and Baldwin, 1990; Hendrix rf al. 1986; Ingham et al,
1986a,b; Moore and De Ruiter, 1990). In soil, one can consider that the first "trophic"
step is organic matter derived from leaf litter, roots, algae, moss, small mammal and
bird fecal material, and secondary metabolites from microbial decomposition. These
materials are used by the two primary decomposer groups, saprophytic fungi and
bacteria, corresponding to the primary consumer trophic level. Nutrient and energy
from living plant material also moves into non-decomposer, root- mutualist fungi and
bacteria, such as mycorrhizae and nitrogen-fixing bacteria, and root-feeding organisms
(parasites, pathogens) such as plant root-feeding nematodes. These organisms should
be considered primary consumers as well.
The next trophic level is comprised of secondary consumers which feed on both the
decomposers and root-associates, i.e., bacterial-feeding(nematodes and protozoa), and
fungal-feeding organisms (nematodes, certain amoebae and a number of
microarthropods). Secondary consumers are fed upon by higher level predatory
TA3.0 -1
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fauna, such as mesostigmatid mites, predatory nematodes, insects, insect larvae small
mammals and birds. There are a variety of generalist consumers within the soil, such
as enchytraeids, earthworms, rotifers, and microarthropods. Thus, a typical soil
foodweb is complex, redundant, highly diverse and quite stable.
By examining foodweb diversity, both the subtle and acute, long-term and short-term
responses to any disturbance can be critically assessed. Responses of each organism
group can be assessed independently, or as part of the complete foodweb. A great
deal more information, especially supportive, explanatory information, can be obtained
if the whole foodweb is assessed. Additionally, with some knowledge of the system
being investigated, the focus can be shifted to the dominant groups of the foodweb,
reducing the workload, while maximizing information gained.
Reductions in foodweb organism numbers in treated soil as compared to control soils
indicate a negative effect, although information on specific responses to specific
toxicants have rarely been determined (Ingham et al. 1985,b). Development of a
database including the typical numbers of each foodweb organism group in different
soils from the US is being developed (Soil Ecology Society, Newsletter 4:2). Significant
alterations from typical numbers suggest a significant impact, and with time, the
ecological effect of reductions in numbers of organisms can be more clearly interpreted
with respect to whether an impact improves or degrades soil fertility and ecosystem
productivity.
References
Coleman, D.C. 1985. Through aped darkly: an ecological assessment of root-soil-
microbial-faunalinteractions. In A. H. Fitter, D. Atkinson, D.J. Read, and M.B. Usher
(Eds.), Ecological Interactions in Soil. Blackwell Scientific Publications, Cambridge,
U.K. Pp. 1-21.
Dindal, D. 1990. Soil Biology Guide. John Wiley and Sons. 1349 pp.
Foissner, W. 1986. Soil protozoa: fundamental problems, ecological significance,
adaptations, indicators of environmental quality, guide to the literature. Prog. Protist.
2:69-212.
Freckman, D.W. and J.G. Baldwin. 1990. Nematoda. pp. 155-200. IN Dindal, D.
1990. Soil Biology Guide. John Wiley and Sons. 1349 pp.
TA3.0 - 2
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Hendrix, P.F., R.W. Parmelee, D. A. Crossley, Jr., D.C. Coleman, E.P. Odum, and
P.M. Groffman. 1986. Detritus foodwebs in conventional and no-tillage
agroecosystems. Bioscience 36:374-380.
Ingham, E.R., J.A. Trofymow, R.N. Ames, H.W. Hunt, C.R. Morley, J.C. Moore, and
D.C. Coleman. 1986b. Trophic interactions and nitrogen cycling in a semiarid
grassland soil. Part n. System responses to removal of different groups of soil
microbes or fauna. J. Appl. Ecol. 23:615-630.
Ingham, E.R., J.A. Trofymow, R.N. Ames, H.W. Hunt, C.R. Morley, J.C. Moore, and
D.C. Coleman. 1986a. Trophic interactions and nitrogen cycling in a semiarid
grassland soil. Part I. Seasonal dynamics of the soil foodweb. J. Appl. Ecol. 23:608-
615.
Janos, D.P. 1987. VA mycorrhizae in hu;mid tropical ecosystems. In: Safir, G.R.
(Ed.). Ecophysiology of VA Mycorrhizal Plants. CRC Press, Inc., Boca Raton, FL.
pp. 107-134.
Miller, R.M. 1987. Mycorrhizae and succession. In: Jordan, ffl, W.R.; M.E. Gilpin
and J.D. Aber (Eds.). Restoration ecology: A synthetic approach to ecological
research. Cambridge University Press, Cambridge, U.K. PP. 205-220.
Moore, J.C. and De Ruiter, P.C. 1990. Temporal and spatial heterogeneity of trophic
interactions within belowground food webs. In D. A. Crossley, Jr. (Ed.), Modern
Techniques in Soil Ecology. Elsevier, Amsterdam. Pp. 371-398.
Reeves, F.B. 1985. Survival of VA mycorrhizal fungi: Interactions of secondary
succession, mycorrhizal dependency in plants and resource competition. N. Amer.
Conf. on Mycorrhizae 6:110-113.
TA3.0 - 3
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TA3.1 Evaluation of contaminant effects on bacterial biomass
Bacterial biomass is typically greater than fungal biomass in agricultural and grassland
soils/ and is thus likely to perform the major portion of decomposition processes in
these systems. Nutrient cycling is in large part controlled by which bacteria are
actaully present/ and the relative growth rates of the active fraction of the bacterial
biomass. Release of nitrogen immobilized in bacterial biomass by feeding of
nematodes and protozoa on bacteria can be critical for plant growth in nitrogen limited
soils.
Soil aggregation is/ in large part/ dependent on the action of bacteria in cementing
smaller particles together into micro-aggregates (Coleman/1985). Aggregate size and
pore-distributions directly influence moisture-holding capacity/ and retention of
organic matter in soils/ both factors of great importance in determining soil fertility.
Loss of significant portions of the bacterial biomass/ or loss of certain important
bacterial species/ such as N-fixers or nitrifying bacteria/ could severely impact the
productivity of a site.
Test method summaries:
Direct estimates of active bacterial numbers: A 1 gram soil sample is removed from
each sample to be assessed/ diluted in 9 ml sterile buffer (1:10 dilution)/ and shaken at
approximately 120 rpm for 5 minutes. One to five ml aliquots are removed from the
1:10 dilution of soil and stained for 3 minutes with 1 ml of a 20 ng/ml FDA (fluorescein
diacetate) solution in 0.1 M phosphate buffer/ pH 6.5 (Ingham and Klein/1984; Lodge
and Ingham/1991). One to five ml of 1.5% agar in pH 9.5 phosphate buffer is added
to the FDA-soil suspension/ mixed well/ and an aliquot placed on a slide with a well
of known depth. Using epi-fluorescent microscopy at a total magnification of 1000X or
greater (oil immersion)/ the number and diameter of all fluorescent bacteria (Zeiss epi-
fluorescent microscope)/ are counted in five fields and recorded. Background
contamination is determined by measuring FDA-stained bacteria using sterile water
samples instead of soil suspensions.
Direct estimates of total bacterial numbers: A 1:100 dilution of each soil sample is
prepared and a 1 ml aliquot stained using the FITC (fluorescein isothiocyanate)
method of Babiuk and Paul (1970). The one mL of soil suspension/ stained for three
minutes with FITC (20 Mg FITC/ml staining solution) is filtered through a 0.2 p.m pore-
size, non-fluorescent (black-stained) Nudepore polycarbonate filter, destained with
sodium carbonate and 5% pyrophosphate as described by Babiuk and Paul (1970), the
TA3.1 - 1
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filter placed on a slide and the number of fluorescent bacteria in each of ten fields on
each filter counted. Background levels of contamination for each batch of buffer
solutions will be assessed by filtering sterile buffer onto filters and counting bacteria
on the filter. These counts (which are normally zero) will be subtracted from counts
obtained on that day's samples.
Calculation of bacterial biomass: Diameters and numbers of bacteria are measured
during observation of organisms. Direct estimates of bacterial and fungal volume are
converted to biomass using hyphal density averages of 0.41 g/cm3, and bacterial
density averages of 0.33 g/cm3 (Paul and Clarke, 1988).
Bacterial community composition: The standard approach to community
composition is to spread dilutions of soil on a variety of different agar media in a
variety of different abiotic conditions/ and then isolate and identify the colonies of
bacteria which appear (see Domsch and Jangnow, 1990 for an overview). Bacterial
identity is usually determined by ability to grow and catabolize specific test nutrients
(various sugars, carbohydrates) or to produce specific enzymes (dehydrogenase,
oxidase, etc). The culture requirements for many soil bacteria are not known and the
percentage of the actual soil bacterial community which grow on the agar media
chosen cannot be determined. Molecular methods offer a bright future for solving the
problem of assessing bacterial identity, but the technology is being developed and is
not available for wide-spread use at this time.
Intended use: Any soil, sediment, litter or plant material can be tested using this
method. The specific information gained is active and total bacterial biomass.
Numbers of sensitive species (if tagged with a immuno-flourescent stain) and
ecosystem responses of bacteria can be assessed using this approach. Sensitive
species can be added and assayed as long as survival and growth requirements for the
particular species is known for the material being tested. Determination of total and
active bacterial numbers indicates effects of toxics on bacterial activity, function and
total biomass. Reductions in toxicant-impacted soil as compared to controls, or
expected levels given the soil type and organic matter level, indicate a negative effect
on bacterial activity and biomass. In combination with information on active and total
fungal biomass, protozoan numbers and community structure, and nematode
numbers and community structure, this information allows estimation of nutrient
cycling, energy flow, foodweb structure and diversity.
Previous applications/regulatoryprecedence: Several research publications have
suggested that changes in bacterial activity and biomass indicate possible changes in
decomposition rates, soil fertility, and general ecosystem function (Coleman, 1985;
TA3.1 - 2
-------
Ingham, et al. 1986a,b; Hendrix, et al. 1986; Nannipieri, et al. 1990). However, total
bacterial biomass is relatively constant over all ecosystem and soil types/ suggesting
that this metric is less sensitive to disturbance with respect to other soil foodweb
determinations. Few studies have measured active bacterial numbers using FDA.
Alternative approaches to measuring bacterial activity are given in TA 3.10 and these
have been shown to be highly sensitive to addition of toxic chemicals.
Regulatory standards exist for numbers and types of bacteria in water and wastewater.
These measurements are performed by plating on a general medium (total aerobic
bacteria) and media specific to coliforms. In soil/ there is no known group of bacteria
with a comparable indicative function such as coliforms have in water. Much more
work is needed in soils to relate species presence/ function and total numbers to a
regulatory role.
Requirements for development and implementation: Development of a database for
a variety of soil types is needed/ as is research to determine the ecological significance
to Superfund and general regulatory requirements.
Potential problems and limitations: Development is needed to indicate the
quantitative levels delineating impacts with different toxic chemicals in different soil
types.
References:
Babiuk, L.A. and E.A. Paul. 1970. The use of fluoresceinisothiocyanatein the
determination of the bacterial biomass of a grassland soil. Can. J. Microbiol. 16:57-62.
Coleman, D.C. 1985. Through a ped darkly: an ecological assessment of root-soil-
microbial-faunalinteractions. In A. H. Fitter/ D. Atkinson/ D.J. Read/ and M.B. Usher
(Eds.)/ Ecological Interactions in Soil. BlackweU Scientific Publications, Cambridge/
U.K. Pp. 1-21.
Domsch, K. H. and G. Jangnow. 1990. Soil bacteria, pg. 1-48. IN Dindal/ D. 1990.
Soil Biology Guide. John Wiley and Sons. 1349 pp.
Hendrix, P.F., R.W. Parmelee/ D.A. Crossley, Jr., D.C. Coleman, E.P. Odum, and
P.M. Groffman. 1986. Detritus foodwebs in conventional and no-tillage
agroecosystems. Bioscience 36:374-380.
TA3.1 - 3
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Ingham, E.R., J.A. Trofymow, R.N. Ames, H.W. Hunt, C.R. Morley, J.C. Moore, and
D.C. Coleman. 1986b. Trophic interactions and nitrogen cycling in a semiarid
grassland soil. Part n. System responses to removal of different groups of soil
microbes or fauna. J. Appl. Ecol. 23:615-630.
Ingham, E.R., J.A. Trofymow, R.N. Ames, H.W. Hunt, C.R. Morley, J.C. Moore, and
D.C. Coleman. 1986a. Trophic interactions and nitrogen cycling in a semiarid
grassland soil. Fart I. Seasonal dynamics of the soil foodweb. J. Appl. Ecol. 23:608-
615.
Ingham, E.R. and D. A. Klein. 1984. Soil fungi: Relationships between hyphal activity
and staining with fluoresceindiacetate. Soil Biol. Biochem. 16:273-278.
12
Lodge, D.J. and E.R. Ingham. (1991). A comparison of agar film techniques for
estimating fungal biovolumes in litter and soil. IN Crossley, D. A., Jr. Methods in
Soil Ecology. Elsevier, The Netherlands.
Nannipieri, P., S. Grego, and B. Ceccanti. 1990. Ecological significance of the
biological activity in soil. Soil Biochemistry 6:293-355.
Paul, E.A. and Clarke, F.E., 1990. Soil Microbiology and Biochemistry. Academic
Press, Inc. San Diego, CA. 272 pp.
TA3.1 - 4
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TA3.2 Evaluation of contaminant effects on fungal biomass
The functional roles of fungi are highly varied, and include saprophytes/ pathogens/
or symbionts (Foster et al. 1983). Fungi affect primary production directly via
symbiotic and pathogenic interactions and indirectly by nutrient cycling processes.
While many fungi are cosmopolitan/ the unique character of the fungal community in
specific biomes has been demonstrated/ as exemplified by the successional patterns of
fungi on decaying leaf litter material and in wood (Kendrick and Parkinson/1990).
In addition to examining fungal community structure/ assessment of fungal biomass
offers a rapid method for comparision between soils. In forest and tundra soils/ fungi
contribute the largest proportion of the total soil biomass/ and most likely contribute
to decomposition and nutrient cycling processes to a greater degree than other
foodweb components. Loss or modification of species composition would most likely
negatively impact nutrient cycling and soil fertility in these systems.
Test method summaries:
Direct estimates of active and total fungi: A 1 gram soil sample is removed from
each sample to be assessed/ diluted in 9 ml sterile buffer (1:10 dilution)/ and shaken at
approximately 120 rpm for 5 minutes. One to five ml aliquots are removed from the
1:10 dilution of soil and prepared for active (fluorescein diacetate stained) and total
(phase contrast microscopy) fungal biomass estimations (Ingham and Klein/1984) by
staining for 3 minutes with 1 mL of a 20 Mg/nU FDA (fluorescein diacetate) solution in
0.1 M phosphate buffer, pH 6.5. One ml of 1.5% agar in pH 9.5 phosphate buffer will
be added to the FDA-soil suspension/ mixed well, and an aliquot placed on a slide
with a well of known depth. The length of FDA-stained hyphae/ using epi-
fluorescent microscopy, and the length and diameter of all hyphae/ using phase
contrast microscopy (Zeiss epi-fluorescent microscope)/ will be measured and recorded
from three 18 mm length transects of each agar film. Total magnification used is 160X
and one agar film is observed per soil sample. Background contamination is
determined by measuring hyphal lengths using sterile water samples instead of soil
suspensions.
Calculation of fungal biomass: Diameters and lengths of fungi are measured during
observation of organisms. Direct estimates of fungal volume are converted to biomass
using hyphal density averages of 0.41 g/cm3 (Lodge/1987).
TA3.2 - 1
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Fungal community composition: The standard approach to community composition
is to spread dilutions of soil on a variety of different agar media in a variety of
different abiotic conditions, isolate the colonies of fungi which appear, and identify
these fungi (see Kendrick and Parkinson, 1990 for an overview). For many fungi so
isolated, the requirements for it to fruit (sexual or asexual reproductive structures) are
not known, and thus the fungus can not be identified. Additionally, the culture
requirements for many soil fungi are not known and the percentage of the actual
fungal community present in soil which grow on the agar media chosen cannot be
determined. Molecular methods offer a bright future for solving this problem of
assessing fungal identity in soil, but the technology is being developed and is not
available for for wide-spread use at this time.
Intended use: Any soil, sediment, litter or plant material can be tested using this
method. Sensitive species can be added and assayed using this approach, as long as
survival and growth requirements for the particular species is known for the material
being tested. Determination of total and active fungal length and biomass indicates
effects of toxics on fungal activity, function and total biomass. Reductions in toxicant-
impacted soil as compared to controls, or expected levels given the soil type and
organic matter level, indicate a negative effect on fungal activity and biomass.
Previous applications/regulatoryprecedence: Several research publications have
suggested that changes in fungal activity and biomass indicate possible changes in
decomposition rates, soil fertility, and general ecosystem function (Coleman, 1985;
Foster et al. 1983; Ingham, et al. 1986a,b; Hendrix, et al. 1986; Lodge, 1987; Richards
1987).
Requirements for development and implementation: Development of the database
for a variety of soil types is currently underway and research is needed to determine
the application of this test to Superfund and general regulatory settings. Soil biomass
testing services may become available at several land-grant institutions in the near
future.
Potential problems and limitations: Development is needed to indicate the
quantitative levels which delineate impacts with different toxic chemicals in different
soil types.
References:
TA3.2 - 2
-------
Coleman, D.C. 1985. Through a ped darkly: an ecological assessment of root-soil-
microbial-faunal interactions. In A. H. Fitter, D. Atkinson, D.J. Read, and M.B. Usher
(Eds.)/ Ecological Interactions in Soil. Blackwell Scientific Publications, Cambridge,
U.K. Pp. 1-21.
Foster, R.C., Rovira, A.D. and Cook, T.W., 1983. Ultrastructure of the Root-Soil
Interface. Am. Phytopath. Soc., St. Paul, MM. 157 pp.
Hendrix, P.F., R.W. Parmelee, D.A. Crossley, Jr., D.C. Coleman, E.P. Odum, and
P.M. Groffman. 1986. Detritus foodwebs in conventional and no-tillage
agroecosystems. Biosrience 36:374-380.
Ingham, E.R. and D.A. Klein. 1984. Soil fungi: Relationships between hyphal activity
and staining with fluorescein diacetate. Soil Biol. Biochem. 16:273-278.
12
Ingham, E.R., J.A. Trofymow, R.N. Ames, H.W. Hunt, C.R. Morley, J.C. Moore, and
D.C. Coleman. 1986a. Trophic interactions and nitrogen cycling in a semiarid
grassland soil. Part I. Seasonal dynamics of the soil foodweb. J. Appl. Ecol. 23:608-
615.
Ingham, E.R., J.A. Trofymow, R.N. Ames, H.W. Hunt, C.R. Morley, J.C. Moore, and
D.C. Coleman. 1986b. Trophic interactions and nitrogen cycling in a semiarid
grassland soil. Part n. System responses to removal of different groups of soil
microbes or fauna. J. Appl. Ecol. 23:615-630.
Kendrick, W. B. and D. Parkinson. 1990. Soil Fungi, pg. 49-68. IN Dindal, D. 1990.
Soil Biology Guide. John Wiley and Sons. 1349 pp.
Lodge, D.J., 1987. Nutrient concentrations, percentage moisture and density of field-
collected fungal mycelia. Soil Biol. Biochem., 19:727-733.
Richards, B.N., 1987. The Microbiology of Terrestrial Ecosystems. Longman Scientific
& Technical, UK. 399pp.
TA3.2 - 3
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TA3.3. Evaluation of contaminant effects on protozoan numbers and diversity
As indicator organisms, protozoa are perhaps unsurpassed (see the excellent review
by Foissner, 1986; also Austin et al. 1990; Domsch et al. 1983; Gupta and Germida,
1989; Lai and Saxena, 1982; Lord and Wright, 1985; Singh and Crump, 1953; Smith and
Wenzel, 1947). To improve the use of protozoa as indicators, however, a greater
understanding of: (1) their response to disturbances beyond the normal seasonal cycle,
(2) their habitat-specificity, and (3) their prey-preferences in specific habitats is needed.
Efforts should be directed towards understanding changes in protozoan community
composition in terrestrial systems.
Protozoa occur in all ecosystems of the world and several hundred species have been
described (Lee et al. 1985). Soil protozoan communities are distinctive for a wide
range of specific habitats although a considerable overlap of individual species occurs
between ecosystems (Stout et al. 1982). There are species with wide-spread
distributions and a tremendous range of adaptability to environmental fluctuations
(Corliss, 1979; Bamforth, 1985), while other species are intolerant of even small
changes in specific environmental conditions (Austin et al. 1990; Foissner, 1986;
Bodenheimer and Reich, 1933; Davis, 1981; Lousier and Bamforth, 1990; Stout, 1984).
In fact, several unique protozoan communities can occur along the length of a single
conifer root (Ingham et al. 1991b).
Protozoa can be classified to genus and often to species based on morphology alone
(Lee et al. 1985). Protozoa occur in large numbers in natural ecosystems and "capture"
of a representative picture of the entire community is not a problem, unlike mammals
or birds.
Soil protozoa are important because many of the roles they play benefit nutrient
cycling, including:
(1) mineralization of N, P and S immobilized in bacterial and fungal biomass
(Anderson et al. 1978; Bamforth, 1985; darholm, 1985; Elliott and
Coleman, 1977; Gupta and Germida, 1989; Ingham et al. 1986a, b;
Kuikman et al. 1990; Woods et al. 1982),
(2) enhanced nitrification rates (Griffiths, 1989);
(3) immobilization (sequestering) of C, N, P and other nutrients in
protozoan biomass (Hunt et al. 1977,1986; Ingham et al. 1986a, b; Ingham
and Horton, 1987),
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(4) release of C from soil via respiration (Anderson et al. 1978; Bryant et al.
1982, Foissner, 1986; Parker et al. 1984a,b; Kuikman et al. 1990),
(5) food source for predators such as nematodes and arthropods (Bryant et
al. 1982; Ihgham et al. 1985; Hunt et al. 1986),
(6) reducing bacterial numbers (Casida, 1989; Chao and Alexander, 1981;
Habte and Alexander, 1977), control of numbers and species diversity of
bacteria and fungi (Darbyshire and Greaves, 1967; Ingham et al. 1986a, b;
Kuikman et al. 1990). Bacterial survival can be improved by adding
bentonite clay which allows the formation of micro-niches in which
protozoa can not reach their bacterial prey (Heynen et al. 1988).
(7) suppression of bacterial and fungal pathogens, i.e., use as biocontrol
agents (Chakrabortyand Warcup, 1984; Chakrabortyand Old, 1982),
(8) agents of plant disease (Dollet, 1984), and
(9) indicators of disturbance, chemical impacts, addition of genetically-
altered bacteria, and soil degradation (Austin et al. 1990; Cairns et al.
1978; Pratt and Cairns, 1985; Foissner, 1986; Bamforth, 1991a,b).
Test method summary: The commonly used enumeration techniques are: (1) direct
observation, and (2) separation techniques. Significant expertise is needed to identify
protozoa using these methods. An experienced investigator is needed to reject
material in soil which appears similar in size and shape to, but are not, protozoa.
Turbidity: In aquatic systems, protozoan feeding rates are very sensitive to the
presence of toxic substances. Measurement of protozoan feeding rates on bacteria is
simple, but dependent on optimizing the concentration of both bacteria and protozoa
in the culture vessels. Known quantities of bacteria are added to aquatic samples and
using turbidiometers or spectrophotometers, density changes are determined
(Horowitz, 1991). The application to soils would likely involve a soil extract, since the
simplicity of the measurement is scoring by changes in turbidity of the cultures.
Testing needs to be performed to determine the level of sensitivity for soil extracts,
and the predictibility of soil extracts for in situ prediction of soil availability of toxic
substances.
TA3.3 - 2
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Most Probable Number (MPN) technique: Individual protozoa are diluted to the
point that there are too few to be transferred into the next dilution step (Singh, 1942,
Darbyshire et al. 1974). Ten-fold or two-fold dilution series are usually prepared, four
to ten 0.5 to 1 ml aliquots removed from each dilution and incubated to allow the
protozoa to reproduce. After 4-7 days for flagellates/ 7-10 days for naked amoebae
and dilates, or weeks for testate amoebae/ a well-mixed aliquot is removed and
scanned using a compound microscope (20 to 45 X). Presence or absence of
flagellates/ amoebae, and dilates is recorded and population density in the soil
calculated.
Tissue culture plates with either 24 (4 or 6 replicates per dilution) or 96 wells (8 or 12
replicates per dilution) per plate are convenient containers for incubation. One half
ml of soil extract agar is placed in each well and 0.5 to one ml of each dilution placed
in the well. Clarholm (1985) suggested that protozoa could be observed in the wells of
the 96-well tissue culture plates but several researchers have found significantly lower
estimates of flagellates and dilates as compared observing aliquots on a slide.
Contamination of medium is detected by growth of protozoa in all wells. Protozoan
numbers are expressed in terms of numbers and biomass, based on the following
equivalents: 1 flagellate = 10"10 g; 1 amoeba = 10^ g, and 1 dilate = 10"8 g biomass (Paul
and Clarke/1990).
Dormant protozoa excyst and become active during the incubation period. The
difference between the numbers estimated from untreated soil (trophic plus encysted)
and numbers estimated from an acid-treated (2% HQ; Cutler, 1923) subsample of the
soil (encysted number) gives the number of trophic/ or active/ protozoa present in that
soil. However/ add may destroy some cysts, may stimulate some to excyst that don't
excyst in the untreated soil/ and may remove preferred bacterial food sources/
resulting in the starvation of some excysted forms (Foissner/1986).
Bamforth (1991b) suggested examining, drop by drop, 0.2 g of soil mixed with 3 ml of
sterile soil extract solution to obtain active numbers of dilates (not appropriate for
amoebae or flagellates), followed by MPN to obtain total numbers of dilates. Total
Direct observation of watered soil suspensions (Foissner, 1986): Suspend soil in a
diluent (e.g., 2 grams in 20 ml sterile distilled water, or 0.2 g in 3 ml water) and
observe 10 to 20 drops of the suspension successively (Bamforth and Bennett, 1985).
Limitations: (1) soil obscures the protozoa and numbers may be underestimated, (2)
only small amounts of soil are observed or the method becomes extremely time-
consuming, for example up to eight hours to observe a single soil sample (Griffiths
and Ritz, 1988), (3) protozoa are nearly the same refractive index as water, and easily
TA3.3 - 3
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missed on the surface of a dark/ amorphous piece of soil material. Staining reduces
this problem.
Foissner (1986) pointed out that, using this method, recovery of dliates added to a soil
was 55-100%, 30-100% for flagellates, about 50% for diatoms, less than 10% for naked
amoebae, and 30-100% for testate amoebae. These types of addition-recovery tests of
recovery efficiency are criticized because they do not take into account that organisms
added to soil are not intimately associated .with soil fractions and are likely to be more
easily recovered.
Staining and fixation methods (Stout et id. 1982; Couteaux and Palka, 1988):
Fixatives, such as: (1) Bouin-Hollande (6.25 g copper acetate in 250 ml distilled water;
add 10 ml picric acid; filter (Whatman #1 filter paper); add 25 ml 40% formalin and 2.5
ml glacial acetic acid), (2) Schaudin's fixative (66 ml of a 6% mercuric chloride solution
in 13 ml of 95% ethanol), (3) Camay's fixative (30 ml absolute ethanol, 10 ml glacial
acetic acid), (4) Fleming's fixative (30 ml of 1% chromic acid, 8 ml of 2% osmium
tetroxide, 2 ml of glacial acetic acid), (5) glutaraldehyde fixative (place sample in 2-5%
glutaraldehyde in appropriate buffer such as phosphate, cacodylate, or water, rinse in
buffer, place in 1-4% osmium tetroxide), or (6) Hollande's fixative (4 g picric acid, 3.5 g
cupric acetate, 10 ml formalin, 5 ml glacial acetic acid, 100 ml distilled water) can be
added to give greater visual contrast between the organisms and soil particles (Stout et
al. 1982; Lee etal. 1985).
Protargol, nigrosin stains, and silver stains can be used to examine of the distribution
of cilia. Iron-hematoxylin, acid methyl-green, and Feulgen stains are used to stain
nuclei, and a variety of flagellar stains are given in Lee et al. (1985), although a good
differential interference contrast (Nomarski, Hoffman's) microscope will often obviate
the need for staining.
Protozoa can be fixed to slides to prevent cells from being washed off during fixation
steps (Lee et a/. 1985). Apply a layer to albumen to the surface of the slide, apply
fixed and dried protozoa to the albumen layer, and rehydrate the protozoa through a
isopropanol series (eg. 15, 30, 75, 85, 95% isopropanol).
Membrane filter techniques (Couteaux, 1967; Lousier and Parkinson, 1981): Dilute
soil to the appropriate degree to give maximum estimates, filter an aliquot onto a
membrane filter and observe for protozoa. Limitations: (1) the pressure involved in
filtration often destroys cells, and this unmodified filtration method is used mainly to
enumerate testate amoebae, since these forms are less sensitive to collapse (Griffiths
TA3.3 - 4
-------
and Ritz, 1988). Foissner (1986) criticized filtration, however, as destroying the very
largest size testacean species.
High resolution microscopy: Phase contrast or differential interference (Nomarski>
Hoffman) microscopy can improve the level of detail which can be observed and
resolve some of the difficulty with direct observations. Also, since the presence,
number and placement of flagella and cilia on cell surfaces, or the way in which
pseudopods are produced by amoebae, are important in identification of species,
interference contrast microscopy can be useful in observing these details. Scanning
electron microscopy (SEM) can also serve to increase resolution of these surface
characteristics of protozoa. SEM is more useful than transmission EM in this capacity,
because of visualization of surface morphological characters (Stout et al. 1982).
Density Centrifugatioru Sucrose, percoll, and other gradient materials have been
used to remove bacteria from obscuring soil particles (Bone and Balkwill, 1986; Basel et
al. 1983). Griffiths and Ritz (1988) utilized this approach for protozoa, testing a variety
of dispersing agents, mixing methods, centrifugation methods, and staining
procedures. Their procedure is: Sieve soil (5 mm mesh), add 5 g wet soil to a
mixture of 50 ml distilled water and 50mM Tris buffer, pH 7.5. Shake for 10 minutes
(wrist action shaker), settle for 60 seconds, remove a 1 ml aliquot (taken 5 cm below
the meniscus), incubate with 0.1 ml 0.4% aqueous INT (iodonitro tetrazoluim) for 4
hours at 25 C, fix with 0.1 ml of 25% glutaraldehyde, and load onto the top of 5 ml
Percoll gradient in 0.1 M Sorenson's phosphate buffer (pH 7.0) in sterile 15 ml
polycarbonate centrifuge tubes. Allow the soil-gradient to settle for 30 minutes and
centrifuge (3000 G for 2 hours). Decant the supernatant, stain with 1 ml of a 5 Mg
DAPI/ml water solution, filter onto a black 25 mm diameter 0.8 um pore size cellulose
membrane filter using suction of not more than 7 kPa. Counterstain with 10 ml of a
33 /xg acridine orange ml'1 solution. Mount filters on microscope slides and observe
using epifluorescence.
Criteria for identifying protozoa are: distinctly stained nucleus or nuclei, appropriate
size and shape, cytoplasm present, no red fluorescence indicative of chlorophyll. One
important assumption is that no shrinkage of cells occurs as a result of fixation in
glutaraldehyde, use of Tris buffer, stains, or as a result of filtration. Dispersants are
not recommended as they are toxic, distorted and reduced the volume of observed
cells.
While some protozoa are left behind in the soil, Griffiths and Ritz (1988) found that
92% of added organisms with a density of less than 1.12 g cm3 were recovered. Large
amounts of soil can be observed with this method, since the protozoa from 100 to 200
TA3.3 - 5
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grams of soil were concentrated on a single filter. Counting is still beset with
interference by soil particles and flagellates can not be differentiated from amoebae or
dilates, although some relative size class differences could be used to differentiate the
different types of protozoa. Active protozoa could be differentiated by counting those
organisms stained with INT or FDA staining and staining with acridine orange
allowed visualization of nuclei (i.e., viable cells).
Additional approaches for assessing protozoan numbers, especially of particular
protozoan groups, as available from the American Agronomy Society (Methods in Soil
Analysis, 1992).
Intended use: Soil, sediment, litter or plant material can be tested using this method.
Sensitive species can be added and assayed using this approach, as long as survival
and growth requirements for the particular species is known for the material being
tested. Determination of numbers of each protozoan group, i.e., flagellates, testate
amoebae, naked amoebae, and dilates indicates effects of toxics on protozoan function
and total biomass. Reductions in toxicant-impacted soil as compared to controls, or
expected levels given the soil type and organic matter level, indicate a negative effect
on protozoan biomass.
Previous applications/regulatoryprecedence: Several research publications have
suggested that changes in protozoan biomass indicate possible changes in soil fertility,
and general ecosystem function (Bamforth 1985; darholm 1985; Kuikman et al. 1990;
Ingham, et al. 1985a,b; Hendrixs, et al. 1986). Foissner (1986) recently published a
review of the sensitive species and their responses to specific toxic chemicals. Burton
has suggested the use of specific dilate spedes as indicators of heavy metal toxidty.
Requirements for development and implementation: Development of the database
for a variety of soil types is currently underway and research is needed to determine
the application of this test to Superfund and general regulatory settings.
While several useful keys for aquatic protozoa exist, a comprehensive taxonomic guide
to soil protozoa is needed. Sandon's monograph (1927) served as the only systematic
account for all four groups of protozoa in a single publication for many years. Kudo's
fifth edition (1966) and Jahn et al.'s second edition (1979) are useful general references,
but do not list spedes and leave out most soil spedes. Smith's key on the terrestrial
protozoa of antarctic islands (1978) furnishes a useful guide, but its scope is limited by
the restricted fauna of that harsh environment.
TA3.3 - 6
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The more recent publication by Lee et al. (1985) includes all five groups of protozoa
and describes species/ but the strong aquatic emphasis of current protozoan taxonomy
is quite apparent. Lousier and Bamforth (1990) deleted strictly aquatic species/ making
their key most useful for soil protozoology/but it differentiates only to the genus
level/ with drawings of only the common species. Foissner (1986) presents the most
complete compendium of soil tiliate species. There simply is no key which covers all
species of known/ or even common/ soil protozoa.
Potential problems and limitations: Development is needed to indicate the
quantitative levels which delineate impacts with different toxic chemicals in different
soil types.
References:
Anderson, R.V., E.T. Elliott, J.F. McClellan, D.C. Coleman, C.V. Cole and H.W.
Hunt. 1978. Trophic interactions in soils as they affect energy and nutrient dynamics.
m. Biotic interactions of bacteria/ amoebae and nematodes. Microbial Ecology 4:361-
371.
Austin, H. K., P.G. Hartel and D.C. Coleman. 1990. Effect of genetically-altered
Pseudomonas solanacearum on predatory protozoa. Soil Biol. Biochem. 22:115-117.
Bamforth, S.S. 1985. The role of protozoa in litters and soils. J. Protozool. 32:404-409.
Bamforth, S.S., and L.W. Bennet. 1985. Soil protozoa of two Utah cool deserts.
Pedobiologia 28:423-426.
Bamforth, S.S. 1991a. Implications of soil protozoan biodiversity. Soil Biodiversity
and Function: Resolving Global and Microscopic Scales. Soil Ecology Society
Meeting, April 1991, Corvallis, Oregon.
Bamforth, S.S. 1991b. Enumeration of soil ciliate active forms and cysts by a direct
count method. Agric. Ecosyst. Environ. 34:209-212.
Basel, R.M., E.R. Richter and G.J. Banwart. 1983. Monitoring microbial numbers in
food by density centrifugation. Appl. Environ. Microbiol. 45:1156-1159.
Bodenheimer, F.S. and K. Reich. 1933. Studies on soil protozoa. Soil Sci. 38:259-265.
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Bone, T.L. and D.L. Balkwill. 1986. Improved flotation technique of microscopy of in
situ soil and sediment microorganisms. Appl. Environ. Microbiol. 51:462-468.
Bryant, R.J., L.E. Woods, D.C. Coleman, B.C. Fairbanks, J.F. McClellanand C.V.
Cole. 1982. Interactions of bacterial and amoeba! populations in soil microcosms with
fluctuating moisture content. Appl. Environ. Microbiol. 43:747-752.
Cairns, A. M.E. Dutch, E.M. Guy, and J.D. Stout. 1978. Effect of irrigation with
municipal water or sewage effluent on the biology of soil cores. N.Z. J. Agric. Res.
21:1-9.
Casida, Jr, L.E. 1989. Protozoan response to the addition of bacterial predators and
other bacteria to soil. Appl. Environ. Microbiol. 55:1857-1859.
Chakrabortry, S. and J.H. Warcup. 1984. Soil amoebae and saprophytic survival of
Gaeumannomyces graminis tritici in a suppressive pasture soil. Soil Biol. Biochem.
15:181-185.
Chakraborty, S. and K.M. Old. 1982. Mycophagous soil amoeba: Interactions with
three plant pathogenic fungi. Soil Biol. Biochem. 14:247-255.
Chao, W.L. and M. Alexander. 1981. Interaction between protozoa and Rhizobium in
chemically amended soil. Soil Sri. Soc. America J. 45:48-50.
Qarholm, M. 1985. Interactions of bacteria, protozoa and plants leading to
mineralizationof soil nitrogen. Soil Biol. Biochem. 17:181-187.
Corliss, J.O. 1979. The Ciliated Protozoa. Characterization, classification and guide
to the literature. 2nd Ed. Pergamon Press, Oxford and Frankfurt. 455pp.
Couteaux, M.M. and L. Palka. 1988. A direct counting method for soil dilates. Soil
Biol. Biochem. 20:7-10.
Couteaux, M.-M. 1967. Une technique d'observationdes thecamoebiens du sol pour
1'estimationde leur densite absolute. Rev. Ecol. Biol. du Sol. 4:593-596.
Cutler, D.W. 1923. The action of protozoa on bacteria when inoculated into sterile
soil. Ann. Appl. Biol. 10:137-141.
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Darbyshire, J.F. and M.P. Greaves. 1967. Protozoa and bacteria in the rhizosphere of
Sinapis alba L., Trifolium repens L. and Lolium perene L. Can. J. Microbiol. 13:1057-1068.
Darbyshire, J.F., R.E. Wheatiey, M.P. Greaves, and R.H.E. Inkson. 1974. A rapid
micromethod for estimating bacterial and protozoan populations in soil. Ecology
61:764-771.
Davis, R.C. 1981. Structure and function of two antarctic terrestrial moss
communities. Ecol. Monogr. 51:125-143.
Dollet,W.D. 1984. Plant diseases caused by flagellate protozoa (Phytomonas). Ann.
Rev. Phytopath. 22:115-132.
Domsch, K. H., G. Jangnow, and T. H. Anderson. 1983. An ecological concept for the
assessment of side-effects of agrochemicals on soil microorganisms. Residue Rev. 86:
65-105.
Elliott, E.T. and D.C. Coleman. 1977. Soil protozoan dynamics in a shortgrass
prairie. Biochem. 9:113-118.
Foissner, W. 1986. Soil protozoa: fundamental problems, ecological significance,
adaptations, indicators of environmental quality, guide to the literature. Prog. Protist.
2:69-212.
Griffiths, B.S., and K. Ritz. 1988. A technique to extract, enumerate and measure
protozoa from mineral soils. Soil Biol. Biochem. 20:163-174.
Griffiths, B. S. 1989. Enhanced nitrification in the presence of bacteriophagous
protozoa. Soil Biol. Biochem. 21:1045-1051.
Gupta, V.V.S.R. and J.J. Germida. 1989. Influence of bacterial-amoebalinteractions
on sulfur transformations in soil. Soil Biol. Biochem. 21:921-930.
Habte, M. and M. Alexander. 1977. Further evidence for the regulation of bacterial
populations in soil by protozoa. Arch. Microbiol. 113:181-183.
Heynen, C.E., J.D. van Elsas, P.J. Kuikman, and J. A. Van Veen. 1988. Dynamics of
Rhizobium leguminosarum biovar trifolii introduced into soil: The effect of bentonite clay
on predationby protozoa. Soil Biol. Biochem. 20:483-488.
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Horowitz, A. 1991. Ecotoxirity testing using a sensitive protozoan. SETAC 12: P034.
Hunt, H. W., C.V. Cole, D. A. Klein, and D.C. Coleman. 1977. A simulation model
for the effect of predation on continuous culture. Microbial Ecology 3:259-278.
Hunt, H.W., D.C. Coleman, E.R. Ingham, R.E. Ingham, E.T. Elliott, J.C. Moore, S.L.
Rose, C.P.P. Reid and C.R. Morley. 1986. The detrital foodweb in a shortgrass
prairie. Biol. Pert. Soil 3:57-68.
Ingham, E.R., J.A. Trofymow, R.N. Ames, H.W. Hunt, C.R. Morley, J.C. Moore, and
D.C. Coleman. 1986a. Trophic interactions and nitrogen cycling in a semiarid
grassland soil. Part I. Seasonal dynamics of the soil foodweb. J. Appl. Ecol. 23:608-
615.
Ingham, E.R., J.A. Trofymow, R.N. Ames, H.W. Hunt, C.R. Morley, J.C. Moore, and
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grassland soil. Part n. System responses to removal of different groups of soil
microbes or fauna. J. Appl. Ecol. 23:615-630.
Ingham, E.R., H.8. Massicotte, and D.L. Luoma. 1991b. Protozoan communities
around conifer roots colonized by ectomycorrhizal fungi, p. 32 (abstr). IN Soil
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Ingham, R.E., J.A. Trofymow, E.R. Ingham, and D.C. Coleman. 1985. Interactions of
bacteria, fungi, and their nematode grazers: Effects on nutrient cycling and plant
growth. Ecol. Monogr. 55:119-140.
Ingham, E.R. and Horton, K.A. 1987. Bacterial, fungal and protozoan responses to
chloroform fumigation in stored prairie soil. Soil Biol. Biochem. 19:545-550.
Jahn, T.L., E.G. Bovee, and F.F. Jahn. 1979. How to Know the Protozoa 2nd. ed.
The Pictured Key Nature Series. Wm. C. Brown Co. Pub. Dubuque, Iowa.
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Kuikman, P.J., Van Elsas, J.D., Jassen, A.G., Burgers, S.L.G.E. and Van Veen, J. A.,
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Lai, R. and D.M. Saxena. 1982. Accumulation, metabolism and effects of
organochlorine insecticides in microorganisms. Microbiol. Rev. 46:95-127.
Lee, J.J., S.H. Hutner, and E.D. Bovee. 1985. An Illustrated Guide to the Protozoa.
Soc. of Protozoologists, Lawrence, Kansas. 629pp.
Lord, S. and SJ. Wright. 1985. The interactions of pesticides with free-living
protozoa. J. Protozool. (Abstr). 31:44A.
Lousier, J.D. and S.S. Bamforth. 1990. Soil Protozoa, pp.97-136. IN Dindal, D. (ed)
Soil Biology Guide. John Wiley and Sons, Inc.
Lousier, J.D. and D. Parkinson. 1981. Evaluation of a membrane filter technique to
count soil and litter testacea. Soil Biol. Biochem. 13:209-213.
Parker, L.W., D.W. Freckman, Y. Steinberger, L Driggers, and W.G. Whitford. 1984.
Effects of simulated rainfall and litter quantities on desert soil biota: Soil respiration,
microflora, and protozoa. Pedobiologia 27:185-195.
Parker, L.W., P.P. Santos, J. Phillips, and W.G. Whitford. 1984. Carbon and nitrogen
dynamics during the decomposition of litter and roots of a Chihuahuan desert annual,
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Paul, E.A. and Clarke, F.E., 1990. Soil Microbiology and Biochemistry. Academic
Press, Inc. San Diego, CA. 272 pp.
Pratt, J.R. and J. Cairns, Jr. 1985. Functional groups in the protozoa: Roles in
differing ecosystems. J. Protozool. 32:415-423
Sandon, H. 1927. The composition and distribution of the protozoan fauna of the
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Singh, B.N. and L.M. Crump. 1953. The effect of partial sterilization by steam and
formalin on the numbers of amoebae in field soil. J. Gen. Microbiol. 8:421-426.
Singh, B.N. 1942. Toxic effects of certain bacterial metabolic products on soil
protozoa. Nature (Lond.) 149:168.
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Smith, N.R. and M.E. Wenzel. 1947. Soil microorganisms are affected by some of the
new insecticides. Proc. Soil Sci. Soc. Am. 12:227-233.
Stout, J.D. 1984. The protozoan fauna of a seasonally inundated soil under
grassland. Soil Biol. Biochem. 16:121-125.
Stout, J.D. S.S. Bamforth, and J.D. Lousier. 1982. Protozoa. IN Miller, R. H. (ed)
Methods of Soil Analysis Part 2. Chemical and Microbiological Properties. 2nd ed.
Agron. Mongr. #9 ASA-SSSA, Madison, pp 1103-1120.
Woods, L.E., C.V. Cole, E.T. Elliott, R.V. Anderson and D.C. Coleman. 1982.
Nitrogen transformations in soil as affected by bacterial-microfaunalinteractions. Soil
Biol. Biochem. 14:93-98.
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TA3.4. Evaluation of contaminant effects on nematode diversity
Nematodes are important fauna in the soil foodweb (Moore and de Ruiter 1990;
Verhoef and Brussaard, 1990). Since this group includes bacterial feeders, fungal
feeders, root feeders, nematode feeders, and insect parasites, they serve ecological
roles ranging from (1) nutrient, especially nitrogen, mineralization, (2) control of
bacterial, fungal, plant, nematode and insect populations, (3) influencing soil
aggregation and thereby soil water holding capacity, and (4) organic matter
decomposition to (5) a food resource for beneficial and pest soil arthropods, insect
larvae, predatory macroarthropods, small mammals and birds (Dindal 1990; Freckman
and Baldwin, 1990).
Nematodes are exposed to contaminants by direct contact with chemicals in the soil
solution or adsorbed on surfaces of soil particles (Freckman and Baldwin, 1990).
Nematode trophic/feeding group dynamics indicate changes in their prey populations
(Moore and de Ruiter, 1990). Reductions in numbers of bacterial-feedingnematodes
indicates a loss of bacteria from the soil, reductions in fungal-feeding nematodes
indicates a reduction in active fungal biomass, and so forth, for each feeding/trophic
group (Ingham et a/., 1986a,b). This information is further useful in defining the
direction of ecosystem change and future problems with respect to nutrient cycling.
Test method summary: A variety of extraction methods are available and this
constitutes an area where work is needed (Freckman and Baldwin, 1990). Nematodes
are separated from soil by simple extraction in water (Baermann funnels; Anderson
and Coleman, 1978), by flotation in water after agitation of the soil by various
methods, followed by screening of the water extract (various methods), and by
sucrose density centrifugation. Extraction efficiency of each method varies by species
being examined and by soil type. Sucrose extraction is the least subject to these
variables, but is also the most time consuming method. A comparison of methods is
needed and one "best method" for use a Superfund sites needs to be chosen, based
on time and efficiency. Extracted nematodes are collected in 50 to 200 ml volume vials
and stored in a refrigerator until the total numbers, and identity if needed, are
determined. Extraction efficiency of each soil type should be determined by sucrose
density centrifugation of each soil before storage (Freckman and Baldwin, 1990).
The test metric is numbers of nematodes present, and the genus, or species,
composition of these nematodes. As a routine part of quality assurance, all
nematodes are identified in large samples, so the ratio of rare to common genera and
species can be assessed. Significant changes in diversity as measured by an index or
TA3.4 -1
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the disappearance of a common species in the impacted soil as compared to reference
soils, are indicators of serious .ecological impacts.
Information on damage threshold or nematode hazard index is available for some
crops. Economic threshold levels (i.e., the number of root-fcedingnematodes
necessary to result in economically significant damage to the crop) are being
developed as well. Ecologically speaking, the ratio of non-root-f ceding nematodes to
root-feeding nematodes needs to remain high in any ecosystem.
Intended use: Any soil/ sediment, litter or plant material. Addition of sensitive
species can be performed in any of these media/ except that survival and growth
requirements for the particular species must be known to exist in the material being
tested.
Previous applications/regulatoryprecedence: Nematode diversity has been used to
define the environmental hazard of only a few contaminants/ such as chloropicrin/
and several heavy metals (Freckman and Baldwin/1990; Bongers, 1989). Addition of
specific nematodes sensitive to environmental contaminants has not been often tested
(Ceanorhabditis, Panagrellus, see previous section)/ but appear potentially useful.
Requirements for development and implementation: The best extraction procedure
which will give comparable extraction of nematodes across different soil types must be
determined/ such that interpretation of numbers and genus/species composition is
reliable from site-to-site. Sensitivity of different species of nematodes to specific
contaminants must be determined. There are only a few known relationships
between contaminants and specific nematode species. As a method/ the addition of
specific/ sensitive nematodes to determine if a known contaminant is present and at
what biologically active level needs two improvements. First/ the habitat range
requirements for each sensitive species must be determined such that the soils in
which the nematode can be tested (i.e., it survives and feeds) are known. Second/ the
sensitivity of these species to a range of contaminants should be determined.
Potential problems and limitations: Assessment of the existing nematode diversity is
useful at several levels of resolution. Loss of sensitive species/ as compared to what is
present in reference soils/ or what is known to exist in that type of soil in that habitat/
is extremely useful for specific contaminant questions. Secondly/ nutrient cycling and
ecosystem level questions can be addressed by assessing loss of trophic groups or
feeding types. Once methodological questions are settled, the potential problems for
this method are extremely few.
TA3.4-2
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With respect to adding sensitive species of nematodes, the potential problems are
significant. The ability of the particular nematode to survive, grow, and feed in the
test soil, regardless of presence or absence of contaminant, must be known before
useful results can be obtained. If pH, moisture, and other abiotic conditions are
known to be acceptable to the test nematode, there should be few problems, and the
results highly useful.
Application of this concept to Superfund sites needs further assessment, with regards
to the maturity index concept suggested by Bongers (1989). This method has not been
used to define specific contaminant impacts, although by examining this range of
trophic-levels and predator-prey groups, a number of sensitive species are likely to
occur and could be monitored. Additionally, known sensitive species, depending on
the contaminant suspected, could be added to the soil and the biologically-activelevel
of contaminant determined. However, these specific associations need to be
experimentally determined.
References:
Anderson, R.V. and D.C. Coleman. 1977. The use of glass microbeads in ecological
experiments with bacteriophagic nematodes. J. Nematol. 9:319-322.
Bongers, T. 1989. Nematoden von Nederlands. The Hague, Amsterdam.
Dindal, D. 1990. Soil Biology Guide. John Wiley and Sons. 1349 pp.
Freckman, D.W. and J.G. Baldwin. 1990. Nematoda. pp. 155-200. IN Dindal, D.
1990. Soil Biology Guide. John Wiley and Sons. 1349 pp.
Ingham, E.R., J.A. Trofymow, R.N. Ames, H.W. Hunt, C.R. Morley, J.C. Moore, and
D.C. Coleman. 1986a. Trophic interactions and nitrogen cycling in a semiarid
grassland soil. Part I. Seasonal dynamics of the soil foodweb. J. Appl. Ecol. 23:608-
615.
Ingham, E.R., J.A. Trofymow, R.N. Ames, H.W. Hunt, C.R. Morley, J.C. Moore, and
D.C. Coleman. 1986b. Trophic interactions and nitrogen cycling in a semiarid
grassland soil. Part n. System responses to removal of different groups of soil
microbes or fauna. J. Appl. Ecol. 23:615-630.
TA3.4-3
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Moore, J.C. and De Ruiter, P.C., 1990. Temporal and spatial heterogeneity of trophic
interactions within belowground food webs. Iru Crossley, D. A. Jr. (Editor), Modern
Techniques in Soil Ecology. Elsevier, Amsterdam, pp. 371-398.
Verhoef, H.A. and Brussaard, L., 1990. Decomposition and nitrogen mineralization in
natural and agroecosystems: The contribution of soil animals. Biogeochemistry,
11:175-211.
TA3.4-4
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TA3.5. Soil microbial biomass (chloroform fumigation)
Chloroform fumigation, as first suggested by Jenkinson and Powelson (1976), is
intended to kill all but a small portion of the microbial population. This remnant re-
colonizes the soil, utilizing the killed microbes as substrate for growth. As a result,
the CO2 evolved during incubation is a measure of Hie killed microbial biomass
initially present. In more recent variations, the amount of N (Sparling and West,
1988), and P (Brookes et al. 1982) extracted immediately after fumigation compared to
what was present before incubation is used as a measure of the killed microbial
biomass.
Test method summary: Soil samples are removed from the site, sieved (2mm mesh)
to remove root and litter material, 50 to 100 g subsamples placed in small containers,
these containers placed in a vaccum chamber, a beaker of chloroform placed in the
chamber, and the chamber placed under pressure (Jenkinson and Powelson, 1976;
Paul and Clarke, 1988). The chloroform is allowed to "boil" and fills the chamber
atmosphere. Two or three flushes of chloroform are intially needed to circulate the
vapor within the soil pores. After the final flush, the chamber is left for 24 to 36
hours filled with chloroform vapor. The chamber is then flushed several times with
air in order to remove the chloroform. Soils are incubated in respiratory chambers
and CO2 evolved from the soils is measured over the next 10-day period. An
unfumigated control is considered necessary by some researchers to assess the normal
respiration processes that would occur in these over the incubation period.
Variations of this method have been developed based on difficulties with the original
method. For example, if a variable number of microbes remain after fumgiation in
different soils, the growth of this portion of surviving microbes may use different
carbon substrates at different rates, while the use and availablity of carbon in soil
subjected to fumigation may be different from use of C and C availability in non-
fumigated soil. Subtraction of the "control" is nonsensical and often, the unfumigated
control may evolve more carbon dioxide than the fumigated sample.
One alternative to fumigation incubation is immediate extraction of N, P and S from
fumigated soils to avoid the problem of incubating and the efficiency of growth of
different organisms on different substrates. However, differentiation of what is in the
soil before fumigation compared with after fumigation remains a problem. A further
difficulty arises with modificiationof the absorptive capacity of soil following
fumigation, such that N, P or S may be more or less retained by the soil particles and
organic matter within different soils, and before versus after fumigation.
TA3.5 - 1
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Intended use: To give a measure of microbial C, N, or P, depending on the
extraction method utilized.
Previous applications/regulatoryprecedence: This method has been used for some
time/ but there have always been questions about the pools of carbon actually utilized
by the population of bacteria/fungi that grow after fumigation, and the absorptive
capacity of the soil before versus after fumigation (Ingham and Horton, 1987).
Requirements for development and implementation: Chloroform has been listed as
a hazardous chemical and is restricted by many laboratories. An alternative is to
microwave instead of fumigate, although few comparisons between microwaving and
fumigation have been minimal (Hendricks et al. 1988). Few technical support
laboratories currently conduct these tests and implementing this biological assessment
within an ecological effects assessment for Superfund may be difficult owing to the
continuing controversy surrounding interpretation of this test.
Potential problems and limitations: Precisely what C, N, P or S pools are released
following chloroform fumigation have not been clearly determined (Ingham and
Horton, 1987). Whether the C, N, P or S released are actually a measure of microbial
biomass or labile pools in the soil has not been adequately elucidated. More work
needs to be performed to allow dear interpretation of this method.
References:
Brookes, P.C., D.S. Powelson and D.S. Jenkinson. 1982. Measurement of microbial
phosphorus in soil. Soil Biol. Biochem. 14:319-329.
Hendricks, C.W. and N. Pascoe. 1988. Soil microbial biomass: Estimates using 2450
MHz microwave irradiation. Plant and Soil 110:39-47.
Ingham, E.R. and Horton, K.A. 1987. Bacterial, fungal and protozoan responses to
chloroform fumigation in stored prairie soil. Soil Biol. Biochem. 19:545-550.
Jenkinson, D.S. and D.W. Powelson. 1976. The effects of biocidal treatments on
metabolism in soil - V. A method for measuring soil biomass. Soil Biol. Biochem.
8:209-213.
TA3.5 - 2
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Paul, E.A. and Clarke, F.E., 1990. Soil Microbiology and Biochemistry. Academic
Press, Inc. San Diego, CA. 272 pp.
Sparling, G.P. and A. W. West. 1988. Modifications to the fumigation-extraction
technique to permit simultaneous extraction of soil microbial C and N. Commun. Soil
Sci. Plant. Anal. 19:327-344.
TA3.5 - 3
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.fan
TA3.6. Solid-phase and Aqueous-phase Microtox1
While aqueous-phase testing with Microtoxto has been readily available for ten to
fifteen years, solid-phase testing has only recently been commercially available
(Microbics 1992). As previously summarized (Warren-Hicks, et al. 1989), Microtoxta
relies upon measurements of bioluminescence for an evaluation of a sample's toxicity.
The test, whether aqueous- or solid-phase, utilizes freeze-dried cultures of the marine
bacterium Photobacterium phosphoreum and is based on the inhibition of
bioluminescence by toxicants (Bulich 1979,1982,1986). The results of several studies
of pure compounds and complex chemical mixtures suggests that aqueous-phase
testing with Microtox*111 generally agrees with standard fish and invertebrate toxicity
tests (Curtis et al. 1982) Solid-phase testing with Microtox*11*, however, does not have
a comparable data base established for developing correlative statements regarding its
correspondence with standard soil tests using, for example, earthworms (see TA1.1).
Test method summary: In brief, both aqueous-phase and solid-phase tests with
Microtox*111 directly measure biological activity in water (or soil- and sediment-derived
extracts), and sediment or soil, respectively. Both test systems use luminescent
bacteria (Photobacterium phosphoreum) to measure the biological effects on culture
metabolism (inhibitory or stimulatory) that may be associated with exposure. Altered
cellular metabolism may affect the intensity of light output from the organism, and
when these changes in light output are expressed, estimates of biological effects may
be derived from screening or concentration-response curves that yield ECgoS
(concentration of sample associated with a 50% reduction in light intensity) from
plotted data. Unlike the aqueous-phase test where the sample (i.e., surface water,
groundwater, sediment pore water or soil eluate) is directly tested (samples may be
filtered), the solid-phase test requires a pre-testing extraction step. During the
extraction, a //micro-eluate//is prepared from a soil sample (ca 0.3 gram), then
incubated at 15°C. Following incubation, the soil-diluent slurry is filtered, and the
filtrate is subsequently analyzed using the Microtox*111 analyzer.
Intended use: Soil-phase Microtox*111 was developed for directly testing solid materials
like soils, sediments, and sludges. Although not a direct test of the soil where soil
particle-organism interactions are considered, by including a direct incubation step
with the well-establishedaqueous-phase test, Microtox*111 can be used to indirectly
evaluate soil contamination, particularly those adsorbed materials that are relatively
bioavailable owing to their solubility.
Previous applications/regulatoryprecedence: Aqueous-phase testing with Microtox*111
has a long history in toxicity assessment, and has been used in a variety of
TA3.6 -1
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contaminant-related studies with surface water, groundwater, aqueous and organic
extracts of solid materials, and sediment pore waters. Currently, solid-phase
Microtox*111 lack regulatory precedence, and it is sufficiently recent in development that
is does not have a large data base regarding its sensitivity in determining soil
contamination.
Requirements for development and implementation: Numerous technical support
laboratories are currently providing aqueous-phase testing with Microtoxto and the
solid-phase test should be equally accessible through these laboratories. Costs per
sample vary, but are relatively inexpensive ranging between $100 to $300, depending
upon sample numbers, turn-around times, and technical considerations (e.g., sample
manipulations to model remediation alternatives) that may be site-specific.
Potential problems and limitations: Aqueous-phase testing and solid-phase testing
with Microtox*"1 should both be performed in conjunction with other assessment
methods, for example, animal or plant tests (TA1 and TA2) as previously noted
(Warren-Hicks, et al. 1989). As a direct measure of altered soil structure and function,
and for interpretation ecological effects, the solid-phase Microtox*"11 currently requires
the support of adequately defined site-specific reference soils, as well as a comparative
data base that relates solid-phase Microtox*"1 test results with soil "health."
References:
Bulich, A. A. 1979. Use of luminescent bacteria for determining toxicity in aquatic
environments. In: Markings, L.L., and R.A. Kimerle, eds. Aquatic Toxicology.
American Society for Testing and Materials, Philadelphia, PA.
Bulich, A. A. 1982. A practical and reliable method for monitoring the toxicity of
aquatic samples. Process Biochem. 17:45-47.
Bulich, A.A. 1986. Bioluminescent assays. Pages 57-74. Iru G. Bitton and B.J.
Dutka, eds. Toxicity Testing Using Microorganisms, Vol. 1. CRC Press, Boca Raton,
FL.
Curtis, C., A. lima, S.J. Lorano, and G.D. Veith. 1982. Evaluation of a bacterial
bioluminescencebioassay as a method for predicting acute toxicity of organic
chemicals to fish. Pages 170-178. In: J.G. Pearson, R.B. Foster, and W.E. Bishop,
TA3.6 - 2
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eds. Aquatic Toxicity and Hazard Assessment, STP 766, American Society for Testing
and Materials. Philadelphia, PA.
Microbics. 1992. Microtoxtm manual. Microbics Corporation, Carlsbad, CA.
Warren-Hicks, W., B. Parkhurst, and S. Baker, Jr. (eds.). 1989. Ecological assessment
of hazardous waste sites. EPA/600/3-89/013. U.S. Environmental Protection Agency,
Environmental Research Laboratory, Corvallis, OR.
TA3.6 - 3
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TA3.7. Alternative microbial toxicity tests (Toxi- and SOS-Chromotest).
Alternative short-term microbial tests are frequently based on enzyme inhibition in
bacteria and fungi. For example, microbial test methods have been developed that
assess the toxicity of domestic and industrial effluents, discharges, and waste products
like sludge. Ecological effects assessments rely on short-term tests that measure acute
toxicity of chemicals to bacteria and other organisms. In conjunction with field
surveys to address "lab-to-field" extrapolation error, these laboratory tests may suggest
contaminant-related effects on various soil functions, e.g., representative trophic levels
that mediate the cycling of nutrients. These tests, while not standardized, may assess
a wide range of toxicants in soils and soil-derived leachates or eluates, either directly
or after extraction and concentration.
As an example, both Toxi- and SOS-Chromotest rely on measurements of enzyme
activity as a basis for toxicity testing with bacteria (Bitton and Koopman 1986;
Christensen et al. 1982). In these tests adverse biological effects associated with
contaminants are measured via de novo enzyme biosynthesis or inhibition in laboratory
microorganisms. The classic example is the indurible enzyme system beta-
galactosidase, which is controlled by the lac operon (Jacob and Monod 1961). Toxicity
tests based on the inhibition of beta-galactosidasein E. coli have also been developed
and evaluated with toxicants in single-compound and complex chemical mixture
matrices (Dutton et al. 1988). Tests based on the inhibition of beta-galactosidase
activity are only sensitive to heavy metals, but the one based on enzyme biosynthesis
apparently respond to both organic and inorganic toxicants (Dutton et al. 1988). A
modification of this test system has also been used for genotoxitity (Quillardetand
Hofhung 1985). This test is based on the induction of the gene sftA, which is
controlled by the general represser of the SOS system in E. coli. Expression of sfiA is
mediated through gene fusion with lacZ gene for beta-galactosidase.
Test method summary: Toxi-Chromotestand SOS-Chromotest (Orgenics Ltd.
1985a,b) consist of colorimetric assays of microbial enzymatic activities after incubating
various concentrations of water or sediment and soil extracts with the special test
strain E. coli (K-12 PQ37) that possesses altered cell membrane permeabilities, and
biological activity associated with soil extracts is determined measuring inducible B-
galactosidase activity. In Toxi-Chromotest, enzyme inhibition is indicative of adverse
biological effects. SOS-Chromotest differs from Toxi-Chromotest primarily in the
mode of expression of the B-galactosidase activity; SOS-Chromotest uses the altered E.
coli strain with the B-galactosidase gene linked to the SOS operator gene. In the
presence of genotoxic effects and the activation of the SOS system, B-galactosidase
activity will be induced, and the enzyme will be produced. SOS-Chromotest may also
TA3.7-1
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be completed using S-9 supplemented systems to evaluate the potential for metabolic
activation in test samples. The extent of enzyme expression can be quantified
spectrophotometrically/and the endpoint measured in either Toxi-or SOS-Chromotest
reflects toxicity (Orgenics 1985a,b).
Intended use: In recognizing the potential "lab-to-field" extrapolation error apparent
in methods that rely on laboratory test organisms such as E. coli, these alternative
tests systems relying on microbial function should be run in conjunction with other
tests and field surveys to provide information for interpretation of ecological effects.
Previous applications/regulatoryprecedence: These methods, while widely used in
applied sciences/ have not been codified within a regulatory context. Their application
to soil contamination surveys would require adequate description and support within
a site-specific sampling and analysis plan.
Requirements for development and implementation: These laboratory tests are
versatile and cost-effective assessment tools (Hicks and Van Voris 1988; Bitton and
Dutka 1986; Dutka and Bitton 1986; Liu and Dutka 1984). Because they are simple,
rapid, and relatively inexpensive procedures, they may become readily available
through contract services laboratory in the future. However, unlike tests with
Microtox*111, few technical support laboratories are currently providing tests with these
organisms, and implementing this biological assessment within an ecological effects
assessment for Superfund may be difficult owing to an absence of experienced testing
services.
Potential problems and limitations: Although enzymes are quite sensitive to heavy
metals, tests that evaluate enzyme inhibition are frequently less sensitive to organic
toxicants. Methods that evaluate enzyme induction or enzyme biosynthesis, on the
other hand, respond to both organic and inorganic toxicants (Dutton et al. 1988).
References:
Bitton, G., and B.J. Dutka, eds. 1986. Toxicity Testing Using Microorganisms, Vol. 1.
CRC Press, Boca Raton, FL.
Bitton, G., and B. Koopman. 1986. Biochemical tests for toxicity screening. Pages 27-
55. In: G. Bitton and B.J. Dutka, eds. Toxicity Testing Using Microorganisms. CRC
Press, Boca Raton, FL.
TA3.7-2
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Christensen, G.M., D. Olson, and B. Reidel. 1982. Chemical effects on the activity of
eight enzymes: A review and a discussion relevant to environmental monitoring.
Environ. Res. 29:247-255.
Dutka, B.J., and G. Bitton, eds. 1986. Toxicity Testing Using Microorganisms, Vol. 2.
CRC Press, Boca Raton, FL.
Dutton, R.J., G. Bitton, and B. Koopman. 1988. Enzyme biosynthesis versus enzyme
activity as a basis for microbial toxicity testing. Toxicity Assess. 3:245-253.
Hicks, R.J., and P. Van Voris. 1988. Review and Evaluation of the Effects of
Xenobiotic Chemicals on Microorganisms in Soil. Report 6186. Pacific Northwest
Laboratory, U.S. Department of Energy, Battelle Memorial Institute, Richland, WA.
Jacob, F., and J. Monod. 1961. Genetic regulatory mechanisms in the synthesis of
proteins. J. Mol. Biol. 3:318-356.
Liu, D., and B.J. Dutka (Eds.). 1984. Toxicity Screening Procedures Using Bacterial
Systems. Marcel Dekker, New York, N.Y.
Orgenics Ltd. 1985a. The Toxi-chromotest, Version 2 (US). Orgenics Ltd., P.O. Box
360, Yavne 70650, Israel.
Orgenics Ltd. 1985b. The SOS Chromotest Blue Kit, TwoStep Version 3. Orgenics
Ltd., P.O. Box 360, Yavne 70650, Israel.
Quillardet, P., and M. Hofnung. 1985. The SOS chromotest, a colorimetric bacterial
assay for genotoxins: Procedures. Mutation Res. 147:65-78.
TA3.7-3
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TA3.8. Static, batch, and continuous-flow microbial tests.
Static or batch methods have traditionally been used to evaluate microbial growth in
terrestrial systems, and various techniques that may be applicable to site-specific
ecological effects assessments have been described in the literature (e.g., Alexander
1986). A continuous-flow, microbial test system, however, has been developed
(Hendricks, et al. 1987) to evaluate chemical effects on soil biochemical processes such
as nitrification (Rhodes and Hendricks 1990), and should be considered during the
process of developing sampling and analysis plans focussing on soil contamination
evaluations. While initially developed for agrichemical evaluations, the method should
be considered for an evaluation of ecological effects, especially for soil contaminated
with chemical mixtures characteristic of hazardous waste sites. From a technical
perspective, the continuous-flow method for evaluating soil contamination is an
attractive waste site testing tool, since a relatively small amount of soil (15-20 g dry
weight) may be used to evaluate biological effects and toxicity associated with
contaminated soil using indigenous microbial communities.
Test method summary: Static or batch culture systems may be used to evaluate
microbial growth in terrestrial systems (Alexander 1986), and the continuous-flow
method (Hendricks, et al. 1987) that has been recently reported may also provide
information relevant to soil contamination evaluations for Superfund. For testing with
the continuous-flow method, samples may be collected from surface or subsurface
soils (e.g., at various depths along the soil profile) and prepared for testing. Soil
samples are generally moistened and incubated at room temperature five days prior to
testing. For testing, the soil samples are sieved then loaded into a test chamber (see
Hendricks, et al. 1987) and placed in-line with a continuous perfusion apparatus.
Exposures may be completed with untreated soils or reference soils perfused with
defined substrate; alternatively, if sufficientsoil-leachatepr soil-eluate is available, the
test can be completed with aqueous soil-derivatives passing through a soil to evaluate
the effects of water soluble soil constituents. Exposures in the continuous-flow test
should last three to five days, depending upon site-specific considerations in the
sampling and analysis plan.
Intended use: Ideally, the continuous-flow test using soil grab samples should be
conducted parallel with standard toxicity tests, e.g., aqueous-phase Microtox*"1 or an
alternative test for evaluating soil function (see TA3.5,3.6, 3.7). In conjunction with
field surveys, these laboratory estimates of adverse biological effects can be placed
within the context of an ecological effects assessment.
TA3.8 -1
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Previous applications/regulatoryprecedence: The methods outlined and
summarized here are early in the standardization process, and are not intended to be
"stand alone" tests. Rather, the strengths of these method lie in their contribution to
weight of evidence arguments that are supportive of those methods that dearly
illustrate adverse effects associated with soil exposures, e.g. relative decrease in soil
biomass mediated by toxic effects of soil contaminants.
Requirements for development and implementation: Few technical support
laboratories are currently providing tests with these systems, and implementing this
biological assessment within an ecological effects assessment for Superfund may be
difficult owing to an absence of experienced testing services.
Potential problems and limitations: Biological effects associated with soil processes
such as nitrogen fixation may be variously expressed in reference soils, the relative
impact of soil contaminants must be derived with respect to an adequate reference
soil. As with other methods summarized throughout this compendium,
complementary test systems and field surveys should also be completed, particularly
when ecological effects are being evaluated on the basis of laboratory tests (Blaise et al.
1985; Plotkin and Ram 1984; Burton and Stemmer 1988; Dutka and Kwan 1988; Giesy
et al. 1988).
References:
Alexander, M. 1986. Introduction to Soil Microbiology. Third Edition. Wiley, New
York, NY.
Blaise, C, N. Bermingham, and R. Van Coillie. 1985. The integrated ecotoxicological
approach to assessment of ecotoxicity. Water Qual. Bull. 10:3-10.
Burton, G.A., Jr., and B.L. Stemmer. 1988. Evaluation of surrogate tests in toxicant
impact assessments. Toxicity Assess. 3:255-269.
Dutka, B.J., and K.K. Kwan. 1988. Battery of screening tests approach applied to
sediment extracts. Toxicity Assess. 3:303-314.
Giesy, J.P., R.L. Graney, J.L. Newsted, C.J. Rosiu, A. Benda, R. J. Kreis, Jr., and F.J.
Horvath. 1988. Comparison of three sediment bioassay methods using detroit river
sediments. Environ. Toxicol. Chem. 7:483-498.
TA3.8 - 2
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Hendricks, C.W., E.A. Paul, and P.O. Brooks. 1987. Growth measurements of
terrestrial microbial species by a continuous-flow technique. Plant and Soil 101:189-
195.
Plotkin, S., and N.M. Ram. 1984. Multiple bioassays to assess the toxicity of a
sanitary landfill leachate. Arch. Environ. Contam. Toxicol. 13:197-206.
Rhodes, A.N. and C.W. Hendricks. 1990. A continuous-flow method for measuring
effects of chemicals on soil nitrification. Toxicity Assess. 5:77-89.
TA3.8 - 3
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TA3.9. Soil-core microcosm test.
The soil-core microcosm test has been standardized through ASTM E1191 (1991), and
has been validated within an ecological risk assessment context for various chemical
and biological hazards (e.g., Van Voris, et al. 1985a,b; Bolton, et al. 1991a,b). The test
is designed to evaluate the environmental fate, ecological effects, and environmental
transport of chemicals, both liquid and solid, and genetically-engineeredmicrobial
agents that may be released to terrestrial systems. For chemicals, the method can be
used to evaluate toxicity or adverse effects on growth and reproduction of native
vegetation or crops and the uptake and cycling of nutrients in a soil/plant system.
Contaminant uptake by plants can potentially be evaluated using the soil-core
microcosm as well as contaminant fate and transport, especially for describing the
relationships between groundwater and soil contaminants.
Test method summary: As a microcosm, the soil-core test potentially measures the
adverse effects, or toxicity, of chemicals in either defined or complex chemical mixture
exposures (ASTM E11911991). Briefly, the standard method follows a relatively
simple outline. Originally, the 60-cm deep by 17-cm diameter terrestrial soil-core
microcosm was designed to yield chemical effects data in soils collected from
grassland or agricultural systems, but the method may be adapted for other soil types
as necessary. The cylinder containing the intact soil core is collected from a site using
stainless steel extraction tubes (see Van Voris, et al. 1985b; Tolle, et al. 1983; Zwink, et
al. 1984), laboratory testing is completed on the intact core. Routine physicochemical
analyses are completed on the soil, e.g., percent organic material, cation exchange
capacity, and nutrient analysis, and in conjunction with field surveys, vegetation and
soil biota are characterized. Once in the laboratory, the soil core can be manipulated
following a site-specific sampling and analysis plan, but ideally exposure conditions
occur in a greenhouse or environmental chamber. The ASTM E1191 (1991) standard
guide outlines numerous exposure methods, depending upon the data needs of the
study. Generally, exposure periods are relatively long, and depending upon the site-
specific requirements, would be six to eight weeks duration. Regardless of the
exposure methods, potential endpoints measured in a soil-core microcosm study are
numerous, but ecological endpoints that are routinely considered include productivity
measurements and measurements of plant health, nutrient loss and chemical fate
testing.
Intended use: The standard guide developed by ASTM E1191 (1991) provides a
laboratory test system that evaluates environmental fate, transport, and toxicity as part
of an assessment of potential ecological impacts associated with chemicals that may be
released in the environment. Initially, the intended application was focussed on
TA3.9 -1
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agrichemical releases in agriculture settings, but the method may be modified to
evaluate other soils on a case by case basis.
Previous applications/regulatoryprecedence: While the soil-core microcosm has
been used in various hazard and risk assessment settings, no regulatory precedence
exits for routinely testing site soils using the soil-core microcosm. Within applied
contexts, the method has proven useful to evaluations of complex chemical wastes
(Van Voris, et al. 1984; Tolle, et al. 1983), hazardous wastes (Van Voris, et al. 1985b)
and agricultural chemicals (Zwink, et al. 1984; Van Voris, et al. 1983).
Requirements for development and implementation: Few technical support
laboratories are currently providing tests with these organisms, and implementing this
biological assessment within an ecological effects assessment for Superfund may be
difficult owing to an absence of experienced testing services.
Potential problems and limitations: The soil-core microcosm potentially yields data
that will be directly relevant to any soil contamination evaluation, and its limitations
are those inherent to microcosms and laboratory tests in general (Van Voris, et al.
1985a). However, if reference soils are available for concurrent testing, the soil-core
microcosm test can yield information that could be significant to an ecological effects
assessment for contaminated soils.
References:
ASTM E1191. 1991. Standard guide for conducting a terrestrial soil-core microcosm
test. Annual book of ASTM standards. Volume 11.04. Pesticides; Resource Recovery;
Hazardous Substances and Oil Spill Responses; Waste Disposal; Biological Effects.
American Society for Testing and Materials (ASTM). Philadelphia, PA. 19103.
Bolton, Jr., H., J.K. Fredrickson, S.A. Bentjen, D.J. Workman, S. W. li, and J.M.
Thomas. 1991a. Field calibration of soil-core microcosms: fate of a genetically altered
rhizobacterium. Microbial Ecol. 21:163-173.
Bolton, Jr., H., J.K. Fredrickson, J.M. Thomas, S. W. li, D.J. Workman, S. A. Bentgen,
and J.L. Smith. 1991b. Field calibration of soil-core microcosms: ecosystem structural
and functional comparisons. Microbial Ecol. 21:175-189.
TA3.9 - 2
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Tolle, D. A., M.F. Arthur, and P. Van Voris. 1983. Microcosm/field comparison of
trace element uptake in crops grown in fly ash-amended soil. Sci. Total Environ.
31:243-261.
Van Voris, P., D. Tolle, M.F. Arthur, J. Chesson. 1985a. Terrestrial microcosms:
validation, applications, and cost-benefit analysis. In Multi-species toxicity testing,
Pergamon Press, New York, NY. Pp. 117-142.
Van Voris, P., D. Tolle, and M.F. Arthur. 1985b. Experimental terrestrial soil-core
microcosm test protocol. A method for measuring the potential ecological effects, fate>
and transport fo chemicals in terrestrial ecosystems. 600/3-85/047, PNL-5450.
Environmental Research Laboratory, Corvallis, OR.
Van Voris, P., D. Tolle, M.F. Arthur, J. Chesson, and T.C. Zwick. 1984.
Development and validation of terrestrial microcosm test system for assessing
ecological effects of utility wastes. EPRI Publication N. EA-3672, Final Project Report.
Electric Power Research Institute, Palo Alto, CA.
Van Voris, P., M.F. Arthur, D. A. Tolle, J.P. Morris, and M. Larson. 1983. Use of
microcosms for monitoring nutrient cycling processes in agroecosystems. In Nutrient
cycling in agricultural ecosystems. Univ. Georgia, College of Agriculture Experimental
Station Special Publication 23. Tifton, GA. Pp. 171-182.
Zwick, T.C., M.F. Arthur, D.A. Tolle, and P. Van Voris. 1984. A unique laboratory
method for monitoring agro-ecosystem effects of an industrial waste product. Plant
Soil 77:395-399.
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TA3.10. Microbial growth inhibition: Nitrification and soil respiration rates.
Test Method Summary: Nitrification rates. Of the major nitrogen transformations
mediated by microorganisms, nitrification appears to be the most sensitive
transformation to a wide range of potential toxicants.
One approach is to compare nitrification rates in contaminated soils to rates in
uncontaminated soils. An alternative method is to add the soil to be tested to a
sensitive culture of nitrifying bacteria. There have been a number of studies in which
the nitrifying bacterium Nitrosomonas europaea has been tested in toxicity studies
(Powell and Prosser, 1986; Sato et al.r 1988). An aqueous suspension of the potential
toxic substance was added to the bacterial culture and the conversion of ammonium to
nitrite was quantified. Although there are no reports to our knowledge of this
approach being used to assess toxicity in soil samples, the results of Powell and
Prosser (1986) suggest that the method has potential usefulness.
The nitrifying organisms used hi these tests were cultures from standard sources. No
attempt was made to seek strains isolated from nonpolluted waters or soils that may
be particularly sensitive to toxicants. Because nitrification is known to be sensitive to a
wide range of toxicants, it is very likely that it should be relatively easy to select for
strains that are particularly sensitive to different groups of toxicants. Knowing the
composition of toxicants at a given site, technicians could select one or more strains
that are known to be particularly sensitive to the toxicants present on that site. The
resulting tests should be very sensitive, rapid, and easy to perform by relatively
untrained personnel.
Intended use: This approach can be used to assess the effects of chemcials on
nitrification, a critical step in the nitrogen cycle. This information correlates with soil
fertility in agricultural settings and changes hi this metric indicate effects on nutrient
cycling.
Previous applications/regulatory precedence: A broad database of information has
been published on nitrification and the effects of various toxic chemicals. A literature
search is needed to summarize this information and improve the interpretation of
hazardous chemtial impacts in these organisms and this process.
Requirements for development and implementation: Few technical support
laboratories currently provide this test and implementing this biological assessment
within an ecological effects assessment for Superfund may need development of
TA3.10-1
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standard protocols by the research laboratories before these methods will become
standard methods for application to site assessment.
Potential problems and limitations: Forest and shrub-land soils do not typically
contain significant numbers of nitrifying organisms and thus impacts of toxic
chemicals on this process in forest and shrub ecosystems can be difficult to quantify.
If large numbers of denitrifying organisms exist in the soil, the nitrate pool may
utilized by these organisms, and low rates of nitrification might be calculated when in
fact the N is being cycled further through the system. Taking soil samples may result
in the release of nutrients from the soil and cause a flush of nitrification to occur.
Test Method Summary: Soil respiration. Soil respiration is a general indicator of
microbial activity, easily measured with relatively simple tools using easy-to-follow
protocols (Dumontet and Mathur, 1989). One main advantage is that respiration can
be determined non-destructively on intact soils. The same volume of soil in the same
plot of ground can be followed over time, a distinct advantage when trying to assess
recovery in a system.
A standardized air-tight container is placed over a known volume of soil, either in the
field/ or in laboratory pots. After 1 to 24 hours, the accumulation of carbon dioxide in
the collecting vessel is determined and compared to controls. Respired gases can be
collected by trapping in alkali (KOH), by removing a known volume of gas from the
headspace of the chamber and analysing for CO2 with a gas chromatograph, for
example.
Intended use: Measurement of respiratory activity is an excellent general means of
assessing soil activity. In general, pesticides and heavy metals have significant
impacts on respiration (Nohrstedt, 1987).
Previous applications/regulatoryprecedence: Considerable data are available on the
effects of toxic chemicals on respiration rates, although this information needs to be
soil compiled into one source. In work with heavy metal contamination, respiration
has been shown to be a useful measure of impact most likely because heavy metals
have such broad effects on organisms. In a study designed to compare both short
and long-term respiration responses of the addition of Cd, Cr, Cu, Pb, Mi and Zn to
five Dutch soils (Doelman and Haanstra, 1984), respiration during the first 2 to 8
weeks was the most sensitive indicator of heavy metal toxirity. The most important
soil characteristic influencing the toxic response was the clay content for Cd, Fe
content for Cu, Pb, and Zn toxirity and pH for Ni toxicity. Inhibition was the greatest
TA3.10-2
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in sand and lowest in the day soils. This study demonstrates the potential
importance of soil characteristics on the apparent toxic effect of heavy metals.
If the toxicant is organic, there is a possibility that soil respiration may actually
increase as the result of organisms degrading the compound. In general, soil
respiration is not affected by the application of herbicides (Grossbard and Davies,
1976) or insecticides although Tu (1970) reported that oxygen consumption rates did
increase as the result of pesticide utilization by soil microorganisms. Fungicide
treatment, however, may led to an initial decrease in respiration followed by an
increase (Domsch, 1970). This is probably due to the indirect effects of killing fungi
with subsequent utilization of the fungal biomass by bacteria (Parr, 1974).
Requirements for development and implementation: Technical support laboratories
that providing biological oxygen demand or dissolved oxygen demand measurements
could provide this information. To implement this biological assessment within an
ecological effects assessment for Superfund needs development of the database for
interpreting any particular set of information.
Potential problems and limitations: As with soil enzymes, all the organisms in soil
contribute to soil respiration rates, including roots. In fact, 40-70% of the CO2
released from soils may come from root respiration. Chemcials may impact only one
component part of all the organisms present in soil and the impact on total respiration
may be small compared to the respiration of all organisms present. Pinpointing the
impacted organism is not possible with this method. Thus, the toxicant must have
broad effects in order to disrupt all soil organism components or there is little
likelihood that an effect will be seen. This metric should be used as a general
indicator of serious and far-reaching impact.
References:
Doelman, P. and L. Haanstra. 1984. Short-term and long-term effects of cadmium
chromium, copper, nickel, lead, and zinc on soil microbial respiration in relation to
abiotic soil factors. Plant Soil 79: 317-337.
Domsch, K. H. 1970. Effects of fungicides on microbial populations in soil. In:
Pesticides in Soil Ecology; Degradation and Movement Symposium. E. Lansing State
University, Mich.
TA3.10 - 3
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Dumontet, S. and S. P. Mathur. 1989. Evaluation of respiration-based methods for
measuring microbial biomass in metal-contaminated acidic mineral and organic acids.
Soil Biol. Biochem. 21:431-435.
Grossbard, F. and H. A. Davies. 1976. Specific microbial responses to herbicides.
Weed Res. 16:163-169.
Nohrstedt, H.-0.1987. A field study on forest floor respiration response to artificial
heavy metal contaminated acid rain. Scand. J. For. Res. 2:13-19.
Parr, J. F. 1974. Effects of pesticides on microorganisms in soil and water. Iru
Pesticides in Soil and Water. Guenzi, W. D., J. L. Ahlrichs, M. E. Bloodworth, G.
Chesters, and R. G. Nash (eds). Soil Sci. Soc. Amer. Inc. Madison, pp. 315-340.
Powell, S. J. and J. I. Prosser. 1986. Effect of copper on inhibition by nitrapyrin of
growth of Nitrosomonas europaea.
Sato, C., S. W. Leung, and J. L. Schnoor. 1988. Toxic response of Nitrosomonas
europaea to copper in inorganic medium and wastewater. Water Res. 22:1117-1127.
Tu, C. M. 1970. Effect of four organophosphorus insecticides on microbial activities in
soil. Appl. Microbiol. 19:479-484.
TA3.10 - 4
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TA3.11. Soil biochemistry tests: Enzymes.
Activity rates, as determined by enzyme studies, nucleic acid production and
incorporation into biomass or nucliec acids, have been used as indices of total soil
microbiological activity and linked with fertility indices in agriculture, but the
correlation of enzyme activity and fertility is rarely straight forward (Nanrdpieri et d.
1990). The approach of using a single measurement as an index of fertility fails to
appreciate the complexity of processes carried out in the soil. A better approach is to
define the aspects of activity which needs to be investigated, select the proper criteria
to measure those aspects of activity, and then utilize the appropriate methods.
A number of measures could be chosen for this work, although specific measurements
have rarely been correlated with toxicant effects. In general, several enzymes that
have been studied with respect to activity following contaminant application. In
general, oxidoreductase and hydrolase enzyme classes should be useful indicators of
environmental activity because these enzymes reflect components of major metabolic
processes common to all life forms (Burton and Lanza, 1987). Oxidoreductases
include dehydrogenase, catalase and peroxidase while hydrolases include invertase,
proteinases, phosphatase and urease. Few studies on transferases or lyases have been
performed in terrestrial systems. With respect to enzymes, each one is specific for a
particular function or group of functions and cannot reflect the soil microbial acitivty
in toto. Therefore, either a suite of enzyme assays must be performed, or some
knowledge of impact must be available in order to choose one or two indicator
enzyme assays.
Test Method Summaries: Enzyme activities depend on addition of the substrate for
the particular enzyme to utilize. Incubation times should be kept as short as possible
to prevent microbial growth and reproduction. Sorption of the substrate or products
on the surfaces of soil and clay particles needs to be prevented, limited, or measured.
Measurement of substrate disappearance, enzyme presence, or product appearance
must be kept as simple as possible, and usually is determined by a color change in the
medium (disappearance of substrate or appearance of product changes pH and a pH
sensitive sye is present in the medium), by change in turbidity, or by the production
of a precipitate or chemical whose presence can be assayed by spectrophotometry.
The dehydrogenase assay is summarized as an example of a typical enzyme assay.
Entire books have been written on the subject of soil enzymes and the reader is
referred to Ladd (1985) and Burns (1986) for specific procedures. Dehydrogenase
activity indicates relative electron transport activity (ETA), or general respiratory
activity. In general, tetrazolium salts are used as alternative electron acceptors, and
TA3.11 - 1
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are reduced to formazan, a purple-red colored precipitate, which can be quantified
spectrophotometrically. The impact of contaminanted versus non-contaminated soils
on aerobic activity of the soil microbial community can thus be measured using this
approach (Lenhard, 1968; Bitton, 1983).
ATP content and adenylate energy charge: The use of ATP (adenosine triphosphate)
content of soil as a measure of microbial biomass and possibly activity assumes (1)
ATP is a constant component of microorganisms of all types, (2) ATP does not remain
in dead cells, and (3) does not long remain adsorbed to soil particles once released
from the cell (Eiland, 1985). None of these assumptions are completely true, but the
greatest problem with measuring nucleic acids released from cells in the soil is their
strong adsorption to clays once the cells are lysed. Nucleic acids so adsorbed are
subject to decreased degradation rates as well, and as such, may be present as a pool
within soil for considerable times. A reasonable relation between ATP and soil
biomass was found if the soil was first brought to 50-60% water holding capacity, and
incubated long enough for plant debris to be decomposed (Eiland 1985). This of
course leads to questions about the representativeness of that soil to field conditions.
In any case/ as is dear from the discussion in the recent review by Nannipieri et al.
(1990), the correlation of ATP with biomass or microbial activity is highly variable,
depending on many soil and climatic factors. Of possible use in some conditions, a
great deal of investigationis needed before a standard interpretation would be
possible.
Adenylate energy charge (AEC) measures the metabolic energy stored in cells and
thus, would relate to the physiological state of the cells (Atkinson, 1977). Metabolic
state is measured as the concentrations of the adenine nucleotides adenosine
monophosphate (AMP), adenosine diphosphate (ADP) and ATP as follows: AEC =
(ATP + 0.5 ADP)/(ATP + ADP + AMP). This can vary between 1 (all ATP) and 0 (all
AMP). Even actively growing cells contain some AMP, and the highest recorded
"activity" values, from laboratory cultures, are in the range of 0.9. The intermediate
values, of cells in stationary growth, were around 0.5 to 0.75. Dead or dying cells
have values below 0.5. Spores were found to have an AEC as low as 0.08. Highly
variable levels of AEC have been found in various soils with little data to support the
reason for the wide variances. Nannipieri et al. (1990) concluded that the AEC of soil
cannot be used as an indicator to assess and predict possible damage to the
physiological functions of soil by potential toxic chemicals, and cannot be used to
assess the effect of agricultural practices or climatic conditions.
Incorporation of radiolabelled nucleic acids: Tritiated amino acids and radio-isotopes
of P and S have been used to measure incorporation into nucleic acids and indicate
TA3.11 - 2
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activity (Nannipieri et al. 1986). The accuracy of the measurement depends on (1) the
reliability of isotope dilution measurements, and on (2) the ability to assess and correct
for nonspecific labeling of cellular constituents. This becomes extremely difficult when
applied to the heterogenous soil mixture and has only been utilized in soil slurries,
never in field trials. In any case, much more research needs to be performed before
any of these methods can be interpreted with respect to contaminant impacts on
terrestrial processes..
Heat production: Calorimetric determination of heat evolved with catabolic metabolic
processes can be measured in small samples, and in soil would represent the sum of
all catabolic processes (Nannipieri et al. 1990). Microcalorimetry is conducted in sealed
ampuoles and the internal atmosphere can have a marked effect on heat production.
In addition, even in soil autodaved three times, heat may be produced from chemcial
and physical mechanisms. Sparling (1981) found that heat production from 12
different soils, with and without glucose addition, proportionately correlated with
respiration, ATP content and amylase activity, but that there was less heat produced
than expected from liquid culture studies. As with other methods, more work is
needed before this measurement could be used in contaminant studies.
Intended use: Significant reduction of activity as compared to control soils is
considered a negative impact.
Previous applications/regulatoryprecedence: There is extensive literature on the
effects of heavy metals on soil enzyme activity (Tyler, 1974,1975; Jordan and
Lechevalier, 1975; Mathur, 1981). The effects of heavy metals on microbial function
and enzyme activities are greater under acid conditions (Domsch, 1984). Arylsulfatase
and urease were more sensitive to heavy metal inhibition than phosphatase but pH
was critical to controlling the effects with pH values above 7 showing little effect
whereas at lower pHs, the effects were more pronounced (Domsch 1984). By
screening 25 microbiological responses to 71 different pesticides, Domsch (1984) found
that acid phosphatase, organic matter degradation, and nitrification were relatively
sensitive, while denitrification, urease activity and non-symbiotic nitrogen fixation
were relatively insensitive indicators of pesticide toxicity. Ammonification, CO2
release and oxygen uptake and dehydrogenase activities were intermediate in their
effect. Phosphatase activity was particularly sensitive to organic phosphorus
insecticides and nitrification was particularly sensitive to fumigants and fungicides.
Normal microbial activity returned within soils 20 to 30 days after pesticide application
(Domsch et al,, 1983), suggesting that depressed functions which could return after
TA3.11 - 3
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pesticide degradation indicated a healthy soil. In Domsch's view, effects that
continued for more than 60 days were critical and indicated real problems.
Burton et al. (1989) found that microbial community assays showed a greater range of
responses and provided for greater discrimination between test sites than any single
test. Beta galactosidase and beta glucosidase activities were the most sensitive
indicators of toxicity, but a range of enzymatic tests gave them valuable information
which lead them to conlcude that the multitrophic approach gave the clearest
indication of environmental problems within the shortest period of time.
In a similar study of two polluted Ohio streams, Burton (1989) also reported that there
were high correlations between sediment enzymatic activities and water quality
profiles. Dehydrogenase activity was positively correlated with Biological Oxygen
Demand (BOD) and oil and grease concentrations. There was a high degree of
temporal variability and no one enzyme showed consistent trends over all times and
in both streams/ again consistent with the need for a multitrophic approach.
In yet another study, Burton and Lanza (1987) concluded that multi-tiered test systems
were most suited to defining the relative toxicity of freshwater sediments.
Dehydrogenase, alkaline phosphatase, glucosidase, amylase, protease activities and
electron transport activity in freshwater sediments were assayed in both waters and
sediments. In addition, others studied in-stream fish, plankton and benthic
communities as well as 7 day Ceriodaphnia and minnow tests. Similar patterns of
toxicity were determined by all tests the degree of sensitivity was greater with the
microbial enzyme assays. However, these relatively inexpensive, rapid microbial
response measurements need to be compared with the more traditional assay methods
over the full range of test conditions.
Burton et al. (1987) also compared the effectiveness of 12 microbial enzyme assays with
7 day Ceriodaphnia survival-reproductionassays in a Montana stream polluted by
mining activity. One of the parameters that they used was resistance to metal toxicity.
Of the 12 microbial parameters that were studied, 5 were highly correlated with metal
concentration data.
Lenhard (1968) observed that in general, dehydrogenase activity is more sensitive to
metals than organics, as appears to be the case with many of the ezyme assays. To
see a depression in the activity of an indicator of general microbial activity, of which
dehydrogenase is an example, overall microbial activity must be depressed. Some
microbes might be depressed by a specific contaminant, while others are stimulated,
resulting in no net effect. This is much more likely to occur with organic compounds
TA3.11 - 4
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than heavy metals/ since organics might serve as a substrate for growth for some
microorganisms. This is particularly true if the organic is crude oil or some petroleum
product (see Burton 1989).
Requirements for development and implementation: A much more extensive
database is needed on all these potential assays. Major factors whose effects must be
considered in establishing a baseline for interpretation of these measurements are: (1)
localization of enzymes, cells/ substrates and nucliec acids in soil/ (2) standardization
of methodology, (3) sorption of substrates/ products and cells by soil clay and organic
fractions/ (4) nutrient cycling during long incubation assays/ and (5) disturbance of soil
while sampling and incubation in the laboratory gives potential rates and not in situ
rates (Nannipieri et al, 1990).
Potential problems and limitations. Activity measures two things; the response of
indigenous microbial communities and the availability of a degradable carbon source.
Heavy metals cannot be used as carbon source/ while organic pollutants might be/ if
the appropriate bacterial and fungal community is present.
There are two potential drawbacks to the enzyme assays when applied to soil/ as
exemplified by the reported difficulties with the dehydrogenase assay. For example/
in order for the tetrazolium salt to be reduced (INT or iodonitrotetrazoliumhas the
fewest direct toxic effects of any tetrazolium compound thus far tested)/ it must be
transported across the cell membrane. This assumes the presence of a utilizable
carbon substrate/ such that active metabolism can occur and INT be reduced at a
nonlimiting rate. If there is little carbon substrate present, metabolism may be too
slow to result in precipitation of the formazan within the test period. In addition/
quantitatively extracting formazan from the soil can be difficult. A number of
components in soil will retain the formazan even following vigorous extraction
procedures. Spiked controls should be used to obtain a extraction efficiency and
circumvent this problem.
General enzymes are produced by a wide variety of microorganisms/ requiring the
toxicant to affect a general reduction in the activity of soil heterotrophs before a
reduction in enzyme activity is evident. Therefore, toxicants with limited or targeted
biological activity will rarely show a general effect, e.g. non-heavy metal pollutants.
The positive aspects of assaying enzyme activity are the well established, rapidly
performed, inexpensive procedures/ which can be performed on whole soils/ as well
as soil extracts.
TA3.11 - 5
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References:
Atkinson, D.E. 1977. Cellular energy metabolism and its regulation. Academic Press,
New York.
Bitton, G. 1983. Bacterial and biochemical tests for assessing chemical toxicity in the
aquatic environment: as review. CRC Crit. Rev. Eviron. Control 13: 51-67.
Burns, R.G. 1986. Interaction of enzymes with soil mineral and organic colloids. IN
P.M. Huang and M. Schnitzer (eds) Interactions of soil minerals with natural organics
and microbes, pp. 429-451. Soil Sci. Soc. Am. Madison, WI.
r
Burton, G. A. 1989. Evaluation of seven sediment toxicity tests and their relationship
to stream parameters. Toxicity Assessment 4:149-159.
Burton, G.A., A. Drotar, J. M. Lazorchak, and L. L. Bahls. 1987. Relationship of
microbial activity and Ceridaphnia. responses to mining impacts on the Clark Fork
River, Montana. Arch. Environ. Contain. Toxicol. 16:523-430.
Burton, G. A., B. L. Stemmer, K. L. Winks, P. E. Ross, and L. C. Burnett. 1989. A
multitrophic level evaluation of sediment toxicity in Waukegan and Indiana Harbors.
Environ. Toxicol. Chem. 8:1957-1066
Burton, G. A., and G. R. Lanza. 1987. Aquatic microbial activity and macrofaunal
profiles of an Oklahoma stream. Wat. Res. 10:1173-1182.
Domsch, K. H., G. Jangnow, and T. H. Anderson. 1983. An ecological concept for the
assessment of side-effects of agrochemicals on soil microorganisms. Residue Rev. 86:
65-105.
Domsch, K. H. 1984. Effects of pesticides and heavy metals on biological processes in
soil. Plant and Soil 76:367-378.
Eiland, F. 1985. Determination of adenosine triphosphate (ATP) and adenylate
energy charge (AEC) in soil and use of adenine nucleotides as measures of soil
microbial biomass and activity. Danish J. Plant Soil Sci. 8:1777:1-193.
Jordan, M. J. and M. P. Lechevalier. 1975. Effects of zinc-smelter emissions on forest
soil microflora. Can. J. Microbiol. 21:1833-1865.
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Ladd, J.N. 1985. Soil Enzymes. IN D. Vaughan and R.E. Malcom (eds). Soil
organic matter and biological activity, pp. 175-221. Martinus Nijhoff, Dordrecht, The
Netherlands.
Lenhard, G. 1968. A standardized procedure for the determination of dehydrogenase
activity in samples from anaerobic treatment systems. Wat. Res. 2:161-167.
Mathur, S. P. 1981. The inhibitory role of copper in the enzymic degradation of
organic soils. Proceed. International Peat Sympos. Bemidji, Minn.
Nannipieri, P., S. Grego, and B. Ceccanti. 1990. Ecological significance of the
biological activity in soil. Soil Biochemistry 6:293-355.
Nannipieri, P., C. Ciardi, L. Badalucco and S. Casella. 1986. A method to determine
DNAandRNA. Soil Biol. Biochem. 18:275-281.
Sparling, G. P. 1981. Heat output of the soil biomass. Soil Biol. Biochem. 13:373-
376.
Tyler, G. 1974. Heavy metal pollution and soil enzymatic activity. Plant Soil 41: 303-
311.
Tyler, G. 1975. Heavy metal pollution and mineralization of nitrogen in forest soils.
Nature 255: 701-702.
TA3.11-7
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TA3.12. Soil biochemistry tests: Lipid chemistry
All groups of living organisms produce lipids for metabolic functions and as protective
surface materials. The chemical composition of soil lipids is complex, the direct result
of the nature and reactivity of the various compounds added to soil from plant litter,
animals, insects and microorganisms. Soil lipids generally account for 4-8% of soil
organic carbon, but can be as much as 42% (Dinel et al. 1990). Information on the
lipid composition of soil is limited as a result of the difficulty in extracting
representative lipid material from soil and the lack of instruments to characterize these
lipids. Recent advances in cross polarization magic angle spin nuclear magnetic
resonance (NMR) and mass spectrometry have opened new horizons for the
characterization of soil lipids, such that different types of carbon (aliphatic C, protein
branching patterns, long alkyl chains, carbohydrates, OH-substituted aliphatics,
aromatics, phenolic and carboxyl C) can be distinguished.
Test method summary: Two approaches hold significant promise with respect to soil
lipids. First, the identity of soil organism groups can be determined using lipid
signatures of particular groups, such as families, genera and species, lipid-structure
signatures from particular microbial groups can indicate subtle shifts in the
composition of affected soils. Second, past biodegradationprocesses, hydrophobic
properties, reactivity and soil development can be assessed by analyzing soil lipids.
To develop the lipid signature library, soil is spread on plates, colonies which grow on
the plate chosen based on morphology, the organism grown in liquid cultures, tested
for purity and a portion of that culture extracted for lipids. These lipids must then be
analyzed for the specific signature compounds, the test soils are then extracted for
lipids, the extracts analyzed, and compared to known lipid signatures.
Intended use: Assess bacterial and fungal community composition shifts and
quantify essential soil characteristics.
Previous applications/regulatoryprecedence: Vestal and White (1989) used changes
in lipid chemistry in the cell walk of microorganisms to assess changes in microbial
diversity in impacted and non-impacted soils. Nordgren et al. (1985) found that there
were changes in the patterns of fungal distributions around a brass mill where the
soils were contaminated with Cu and Zn. Multivariate analysis showed a high
correlation with heavy metal contamination but no correlation with moisture and soil
organid matter. This is one of many studies which have shown that heavy metals
and heavy metal based fungicides greatly influenced the distribution of fungi
(Nordgren et al. 1985). As might be expected, the fungal community structure may be
TA3.12-1
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a more sensitive indicator of heavy metal toxic effects than either soil respiration or
fungal biomass (Nordgren et al. 1983).
Requirements for development and implementation: As summarized by Dinel et al.
(1990):
(1) Develop better techniques and equipment to extract and characterize
chemically highly complex (especially for organisms) and polymerized
(especially for soils) lipids;
(2) Improve knowledge about mechanisms of inhibitory action of certain
lipids on microbial populations and seed germination;
(3) Assess biodegradabilityof various types of lipids in cultivated and
uncultivated soil; and
(4) Evaluate the effect of certain lipids on soil structure.
Potential problems and limitations: Reliable extraction efficiency of lipids from the
sample, whether soil or organisms in soil/ remains a problem. Characterization of
lipids is time-consuming and/ if new structures occur/ difficult. Effects of different soil
communities on lipid expression by individual organisms is a completely unknown
interaction at this time.
References:
Dinel, H., M. Schnitzer and G.R. Mehuys. 1990. Soil lipids: Origin/ nature/ content,
decomposition and effect on soil physical properties. Soil Biochem. 6:397-429.
Nordgren, A. E., E. Baath, and B. Soderstrom. 1985. Soil microflora in an area
polluted by heavy metals. Can. J. Bot. 63:448-455.
Nordgren, A. E., E. Baath, and B. Soderstrom. 1983. Microfungi and microbial activity
along a heavy metal gradient. Appl. Environ. Microbiol. 45:1829-1837.
Vestal, J.R., and D. C. White. 1989. Lipid analysis in microbial ecology. Quantitative
approaches to the study of microbial communities. Bioscience 39: 535-541.
TA3.12-2
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TA3.13. Substrate uptake and decomposition of organic compounds
Substrate uptake can be used as a means of demonstrating toxic effects by following
the fate of radiolabeled toxicant when added to a soil sample. This approach is most
appropriate with organic toxicants that are broken down by organisms with relatively
specialized function.
Test Method Summary: Labeled hydrocarbons, sugars or other substrates of interest
are added to the soil. Their utilization is assayed by determining labeled CO2
production, labeled biomass production, or disappearance of the labeled compound.
Intended use: Assess changes in bacterial and fungal utilization of substrates.
Previous applications/regulatory precedence: Radiolabeled hydrocarbon uptake and
incorporation into biomass have been used to demonstrate increased numbers of
organisms capable of degrading crude oil and petroleum products in areas
contaminated by those and related compounds. Thus, soils contaminated with
degradable organics could be assayed for effects on the ability to utilize particular
compounds by adding a particular radiolabeled compound and following if s fate.
Additionally, assaying for the enrichment of organisms capable of using the toxicant in
the impacted soil as compared to a standard soil, analogous to assaying for resistant
microorganisms, could be 'performed.
Requirements for development and implementation: A much more extensive
database is needed on all these potential assays. Major factors whose effects must be
considered in established a baseline for interpretation of these measurements are: (1)
localization of enzymes, cells, substrates and nucliec acids in soil, (2) standardization
of methodology, (3) sorption of substrates, products and cells by soil day and organic
fractions, (4) nutrient cycling during long incubation assays, and (5) disturbance of soil
while sampling and incubation in the laboratory gives potential rates and not in situ
rates (Nannipieri et al, 1990).
Potential problems and limitations: Before this approach can be highly useful, we
need to know how long after a pollutant enters the soil before enrichment of resistant,
or degradatory organisms will occur, and how long the resistant/degradatory
organisms persist in the environment after a pollutant has been degraded. This
approach is beneficial for the remediation of impacted soil however. The organisms
capable of degrading the pollutant can be isolated, high numbers grown in the
laboratory and used to inoculate the soil at the site. Since the organism was originally
from the site, novel organisms are not being placed on-site. The organisms should be
TA3.13 -1
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able to grow in the condition at the site, since they were originally isolated from the
area. Certainly testing is needed to make certain no changes in genetic capability of
the organism occurs in laboratory increases in numbers.
The main methodological drawbacks of this method are the need for relatively
expensive radiolabeled isotopes, disposal of the radiolabeled test material, the
specialized equipment needed for determining radiolabeled-compounddegradation (a
liquid scintillation counter) and the fact that no general-activity radiolabeled material is
available. If a spectrum of effects are suspected, each substrate must be tested
separately.
References:
Nannipieri, P., S. Grego, and B. Ceccanti. 1990. Ecological significance of the
biological activity in soil. SoilBiochem. 6:293-355.
TA3.13-2
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TA3.14. Soil nutrient dynamics tests - Nitrogen cycling
Of the major nitrogen transformations mediated by microorganisms, nitrogen cycling
is one of trie most important and directly related to plant productivity. In addition,
nitrification of ammonium to nitrite and then to nitrate appears to be the most
sensitive transformation to a wide range of potential toxicants.
Test method summary: One very simple method is to collect soil and assess nitrogen
pools (ammonium,nitrate, nitrite) at time zero and after incubation hi plastic "zip-lock"
bags. The difference indicates the potential nitrogen cycling rate. A second method is
to add N-15 labeled ammonium to the soil and determine the rate at which is appears
as nitrate-nitrite.
For nitrification rates, one approach is to compare nitrification rates in contaminated
soils to rates in uncontaminated soils. An alternative method is to add the soil to be
tested to a sensitive culture of nitrifying bacteria and test for continued function of the
bacterial culture.
Intended use: hi general, biogeochemical transformation of nitrate has been shown to
be highly sensitive to pesticides, herbicides/ and heavy metals (Parr/1974).
Previous applications/regulatoryprecedence: A study comparing the effects of
herbicide application on aerobic nitrogen fixation/ aerobic cellulose decomposition/
arnmonificationand nitrification showed that nitrification •was the most sensitive and
cellulose decomposition was the least affected functional processes (Helmeczi, 1977).
Fungicide and fumigant application can depress nitrification rates hi soil for extended
periods of time (Parr/1974), while denitrificationis not consistently sensitive to
pesticide or herbicide applications (Parr, 1974; Wainwright, 1978). In grassland,
agricultural, and shrubland soils, the organisms performing nitrification are several
species of autotrophic bacteria, sensitive a gamut of toxicants. In acid forest soils,
however, most nitrification is performed by heterotrophic bacteria and fungi which
don't appear as sensitive to toxicants as autotrophic nitrifiers.
There have been a number of studies in which the nitrifying bacterium Nitrosomonas
europaea has been tested in toxicity trials (Powell and Prosser, 1986; Sato et al., 1988).
An aqueous suspension of the potential toxic substance was added to the bacterial
culture and the conversion of ammonium to nitrite was quantified. Although there
are no reports of this approach being used to assess toxicity hi soil samples, the
results of Powell and Prosser (1986) suggest that it has potential usefulness.
Additionally, the nitrifying organisms used in these tests were cultures from standard
TA3.14-1
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sources. No attempt was made to seek strains isolated from nonpolluted waters or
soils that may be particularly sensitive to toxicants. Because nitrification is known to
be sensitive to a wide range of toxicants, it is very likely that it should be relatively
easy to select for strains that are particularly sensitive to different groups of toxicants.
Knowing the composition of toxicants at a given site, technicians could select one or
more strains that are known to be particularly sensitive to the toxicants present on
that site. The resulting tests should be very sensitive, rapid, and easy to perform by
relatively untrained personnel.
Requirements for development and implementation: Procedures for measurement of
soil nitrification rates have been published extensively and are considered routine.
The rate of ammonium conversion to nitrate and/or nitrite/ and the rate at which
nitrite is converted to nitrate are the endpoints measured. The concentrations of
ammonium/ nitrate and nitrite are determined colorimetrically, using either
autoanalyzersor laboratory spectrophotometers.
Potential problems and limitations: Heterotrophic bacteria must be isolated from
nonpolluted soils/ and these cultures must be determined to be sensitive to a wide
range of toxicants. The culture must be assayed for the effects of the test soil on
nitrificationrate/ or nitrogen cycling rates. The combination of substrate and bacterial
strain could be optimized for the toxicants found on a given site or a generally
applicable combination could be standardized for all Superfund sites. The advantage
of this approach is that it can be readily applied to soils, it is extremely fast/ and
sensitive.
References:
Helmeczi, B. 1977. The effect of herbicides on soil bacteria belonging to certain
physiological groups. Acta. Phytopath. Academ. Sci. Hungaricae 12: 41-49.
Parr/ J. F. 1974. Effects of pesticides on microorganisms in soil and water. In:
Pesticides in Soil and Water. Guenzi/ W. D., J. L. Ahlrichs, M. E. Bloodworth, G.
Chesters, and R. G. Nash (eds). Soil Sci. Soc. Amer. Inc. Madison, pp. 315-340.
Powell, S. J. and J. I. Prosser. 1986. Effect of copper on inhibition by nitrapyrin of
growth of Nitrosomonas europaea.
TA3.14-2
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Sato, C, S. W. Leung, and J. L. Schnoor. 1988. Toxic response of Nitrosomonas
europaea to copper in inorganic medium and wastewater. Water Res. 22:1117-1127.
Wainwright, M. 1978. A review of the effects of pesticides on microbial activity in
soils. J. Soil Sci. 29: 287-298.
TA3.14-3
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TA4. FIELD METHODS FOR THE ASSESSMENT OF SOIL CONTAMINATION
AT HAZARDOUS WASTE SITES
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Methods for evaluating ambient, or in situ, toxirity in terrestrial or wetland sites.
The field testing methods summarized here should not be considered "stand alone"
test methods; however, these field tests have had numerous applications in Superfund
and non-Superfund ecological assessments. These field methods evaluate in situ
toxicity and may contribute to an analysis of "laboratory-to-fieldextrapolation errors"
that are critical to risk characterizations required in the overall ecological effects
assessment process.
For use in the field, in situ toxicity tests are being developed and evaluated; some in
situ techniques have been applied to waste site evaluations to a limited extent (e.g.,
Rowley et al. 1983; Tice, et al. 1987; McBee, et al. 1987; Thompson, et al. 1988) but are
not toxicity tests by design. These in situ methods, however, are significant for
evaluating exposure. In situ techniques applied on a site-specific basis may help
integrate laboratory toxicity data with field-derived estimates of exposure, and
subsequently yield an estimate of the hazard associated with a particular waste site
(Figure TA4). Generally, in situ methods use resident species that naturally occur on
or near a waste site, and can be captured to evaluate toxicity or exposure. Various
levels of biological organization can be measured through in situ methods, ranging
from cellular and molecular levels through population levels of organization.
Depending upon the data quality objectives for the ecological effects assessment, the
results of these in situ toxicity tests may yield either high or low resolution
information.
Figure TA4. Relationship between laboratory and in situ toxicity assessments within
the integrated study design characteristic of an ecological assessment (see
Figures 1 and 2).
Chemical analysis
for target analyte
Physicochemical analysis
of site soils (see Sec-
tion 4.0)
Laboratory toxicity |
tests with standard i
methods and species |
Site-specific in situ
toxicity tests and
field studies
Ecological Effects
Assessment
Risk Characterization
TA4.0 - 1
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References:
McBee, K., J.W. Bickhman, K.W. Brown, and K.C. Donnelly, 1987. Chromosomal
aberrations in native small mammals (Peromyscus leucopus and Sigmodon hispidus) at a
petrochemical waste disposal site: I. Standard karyology. Arch. Environ. Contam.
Toxicol. 16:681-688.
Rowley, M.H., J.J. Christian, D.K. Basu, M.A. Pawlikowski, and C.J. Paigen. 1983.
Use of small mammals (voles) to assess a hazardous waste site at Love Canal, Niagara
Falls, New York. Arch. Environ. Contam. Toxicol. 12:383-397.
Thompson, R.A., G.D. Schroder, and T.H. Connor. 1988. Chromosomal aberrations
in the cotton rat, Sigmodon hispidus, exposed to hazardous waste. Environ. Molec.
Mutagen. 11:359-367.
Tice, R.R., B.C. Ormiston, R. Boucher, C.A. Luke, and D.E. Paquette. 1987.
Environmental biomonitoring with feral rodent species. In: S.S. Sandhu, D.M.
Demarine, M.J. Mass, M.M. Moore, and J.L. Mumford, eds.). Volume V. Short-
term bioassays in the analysis of complex environmental mixtures. Plenum Press.
New York, NY.
TA4.0 - 2
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TA4.1. On-site earthworm test methods for biological evaluations of terrestrial and
wetland habitats.
In order to address questions related to "laboratoiy-to-field"extrapolationand site-
specific differences/ for example, in contaminant bioavailability, soil contamination
evaluations completed with earthworms in the field ("on-site testing") may be
amenable to ecological effects assessments (Marquenie, et al. 1987; Callahan, et al.
1991).
Test method summary: Exposure chambers are located in presumedly contaminated
soils and in reference soils. Exposure chamber designs, though varying in physical
conformation, are screened on top and bottom to assure exposures between the test
soil and worms placed into the exposure chambers. To initiate an on-site test, all
exposure chambers should be hydrated to assure adequate soil moisture (e.g., 45%),
and ten worms (for example, Eisenia foetida or Lumbricus terrestris) should be placed
into each replicate exposure chamber. During exposures (e.g., 14-day), all chambers
should be held between 30% to 45% soil moisture, but no additional manipulations
need routinely be completed other than daily monitoring of soil temperature and
disturbance, e.g., animal damage. At test termination, each exposure chamber has its
contents emptied onto a plastic bag, and earthworms are then separated from the soil.
As in the laboratory test (see TA1.1), survivorship should be determined and sublethal
effects should be evaluated [e.g., dermopathology].
Intended use: The method(s) briefly summarized and referenced here are early in the
standardization process, and are not intended to be "stand alone" tests. Rather, the
strengths of these method(s) lie hi their contribution to weight-of-evidence arguments
that support those laboratory methods that suggest adverse effects associated with
soil exposures. In particular, on-site tests address recurring issues related to
"laboratory-to-fieldextrapolation." As with all field tests, the on-site earthworm test
was designed to address methodological biases associated with laboratory tests, and
the disadvantages implicit to field testing (e.g., less constant exposure regimen) are
off set by the potential information gained, particularly for ecological interpretations,
regarding ambient expression of soil toxicity.
Previous applications/regulatoryprecedence: On-site testing with earthworms has
not been codified, but has been completed within a regulatory context for soil
contamination evaluations (Marquenie, et al. 1987; Callahan, et al. 1991).
Requirements for development and implementation: Few technical support
laboratories are currently providing on-site or in situ tests with earthworms, and
TA4.1 - 1
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implementing this biological assessment within an ecological effects assessment for
Superfund may be difficult owing to an absence of experienced testing services.
Potential problems and limitations: The on-site test with earthworms should be
conducted in parallel with the laboratory test (see TA1.1) to gain the most information
regarding soil toxicity and its potential ecological effects. The on-site test will be
subject to various sources of relatively uncontrolled and potentially confounding
effects, but through field testing an improved assessment of ecological effects can be
developed, particularly when ambient toxicity may be confounded by site-specific
habitat limitations, e.g., physical alteration of soils.
References:
Callahan, C.A., C.A. Menzie, D.E. Burmaster, D.C. Wilborn, and T. Ernst. 1991.
On-site methods for assessing chemical impact on the soil environment using
earthworms: a case study at the Baird and McGuire Superfund site, Holbrook,
Massachusetts. Environ. Toxicol. Chem. 10:817-826.
Marquenie, J.M., J.W. Simmers, and S.H. Kay. 1987. Preliminary assessment of
bioaccumulationof metals and organic contaminants at the Times Beach confined
disposal site, Buffalo, NY. Miscellaneous Paper EL-87-6. Department of Army,
Waterways Experiment Station, Corps of Engineers. Vicksburg, MS.
TA4.1 - 2
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TA4.2. In situ amphibian methods used in biological evaluations of wetlands.
Laboratory tests with amphibians yield toxicity estimates under controlled
experimental conditions that may yield "laboratoiy-to-field"extrapolation errors,
unless field effects are accounted for during an ecological effects assessment. Test
methods applicable to field evaluations have subsequently been developed. Unlike
laboratory methods, in situ tests follow less rigid guidelines which allow for various
site-specific contingencies implicit to field testing. In situ exposures for amphibians are
designed with critical early life stages in mind and evaluate the first 4 to 10 days post-
fertilization. The measurement endpoint used to define test termination is
developmental stage rather than a strictly defined exposure period. The significance
of a reference or site-equivalent exposure cannot be ignored since stage-specific
endpoints in reference locations determine the comparative basis for evaluating site-
specific effects. Also, reference locations are critical for site evaluations, since habitat
interactions (particularly for in situ methods) and interspecies variability are
unavoidable and must be considered on a site-by-site basis.
Test method summary: To initiate in situ exposures, fertilized eggs or early embryos
from either commercial sources or pristine habitats must be available. After loading 10
- 25 early embryos into exposure chambers, they are placed into the test matrix [e.g.,
sediment and water column] at on-site locations, then secured with stainless steel
stakes or other restraints. The exposure cages track environmental conditions without
interference. Temperature should be monitored with recording or
maximum/minimum thermometers and periodic water quality measures should be
taken. Daily inspections of the exposure cages are recommended, and when possible
ancillary field work (e.g., wetland field surveys) should be completed to complement
the in situ toxicity assessment.
At termination, the minimal field data collections should yield mortality data.
Additional endpoints readily measured in mobile field-support facilities include length
and gross teratogenic endpoints (e.g., skeletal malformations). Field endpoints could
also include behavioral observations such as mobility.
Intended use: The method briefly summarized and referenced here is early in the
standardization process, and should not intended to be a "stand alone" test. Rather,
the strengths of these method(s) lie in their contribution to weight-of-evidence
arguments that support those laboratory methods that suggest adverse effects
associated with wetland soils and sediment exposures. In particular, the in situ test
with amphibians addresses recurring issues related to "laboratory-to-field
extrapolation." As with all field tests, the in situ amphibian test was designed to
TA4.2 - 1
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address methodological biases associated with laboratory tests, and the disadvantages
implicit to field testing (e.g., less constant exposure regimen) are offset by the
potential information gained, particularly for ecological interpretations, regarding
ambient expressions of toxicity.
Previous applications/regulatoryprecedence: The amphibian test method outlined
here has been applied within an ecological effects assessment at a wetland located on
a Superfund site (Under, et al. 1991), but has not been codified as a regulatory
requirement.
Requirements for development and implementation: Few technical support
laboratories are currently providing in situ tests with these organisms, and
implementing this biological assessment within an ecological effects assessment for
Superfund may be difficult owing to an absence of experienced testing services.
Potential problems and limitations: The in situ test summarized here should be
conducted in parallel with the laboratory amphibian test (see TA1.8) to gain the most
information regarding wetlands potentially impacted by contaminated soils or
sediments and the potential ecological effects associated with these matrices. The in
situ test will be subject to various sources of relatively uncontrolled and potentially
confounding effects (e.g., surface water temperature), but through field testing an
improved assessment of ecological effects can be developed, particularly when
ambient toxicity may be confounded by site-specific habitat limitations, e.g., redox
potential of sediments.
References:
Lander, G., J. Wyant, R. Meganck, and B. Williams. 1991. Evaluating amphibian
responses in wetlands impacted by mining activities in the western United States. In
R.D. Comer, P.R. Davis, S.Q. Foster, C.V. Grant, S. Rush, O. Thorne, and J. Todd
(Eds.). Issues and technology in the management of impacted wildlife. Thorne
Ecological Institute. Boulder, CO. Pp. 17-25.
TA4.2 - 2
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TA4.3. Nest-box tests using starlings to evaluate in situ toxirity.
While initially designed with agrichemical field studies as their focus (FDFRA, P.L. 92-
516), aviari field tests using nest-boxes for evaluating exposure and toxicity are
relatively well developed (Kendall, et al. 1989) and amenable to incorporation into
Superfund ecological effects assessments. As a method currently undergoing ASTM
standardization, the nest-box survey described for European starlings (Sturnus vulgaris)
is a long duration study (as long as three years, for example), but the method could
be modified as needed as a contribution to an ecological effects assessment (Grue, et
al. 1982; Grue and Hunter 1984; Stromborg, et al. 1988).
Test method summary: By design, the nest-box field test is intended to evaluate
contaminant effects on reproductive success (e.g., Robinson, et al. 1988; Schafer, et al.
1981) as measured by number of offspring fledged per sample unit. Other endpoints
are also available for measurement and include: number of fledglings per nest-box;
survivorship of fledglings; eggs per nest-box; percent hatch; and mean weight of
nestlings. Alternative test species, while not specifically addressed, are also amenable
to avian nest-box studies, and these include American kestrels (Falco sparoerius), barn
owls (Tyro alba), wood ducks (Aix sponsa), purple martins (Progne subis), great crested
flycatchers (Myiarchus crintus), great tits (Parus major), flickers (Colaptes auratus), eastern
bluebirds (Sialia sialis), tree swallows (Tachycineta tricolor), black-capped chickadees
(Parus atricapillus), and house wrens (Troglodytes aedori). Test endpoints may also be
identified from various aspects of the reproductive biology of the selected test species.
Clutch size, frequency of egg-laying and egg viability, and secondary poisoning of
offspring may all be appropriate, if developed within the site-specific sampling and
analysis plan.
Intended use: Originally designed for field studies that evaluated agrichemical effects
on birds, the methods developed are equally well suited to ecological effects
assessments for Superfund. While test duration was intended to be relatively long,
and potentially multi-year, the method may be modified to meet site-specific needs
and be designed with multiple week exposures in short-term studies. The longer-
term, multiple year designs may be quite appropriate for ecological monitoring phases
within the RI/FS process (see Figure 1).
Previous applications/regulatoryprecedence: The method(s) outlined and
summarized here are early in the standardization process, and are not intended to be
"stand alone" tests. Rather, the strengths of these method(s) lie in their contribution
to weight of evidence arguments that support laboratory methods conducted in
conjunction with these field tests and surveys. These nest box survey methods with
TA4.3 -1
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starlings were designed primarily to minimize "laboratory-to-field"extrapolation
errors, and are complementary to laboratory methods that evaluate acute and subacute
toxicity (see TA1.10).
Requirements for development and implementation: Primarily due to requirements
under FIFRA, a growing number of technical support laboratories are providing field
tests with birds, and implementing this biological assessment within an ecological
effects assessment for Superfund should be considered, if data needs require an
assessment of avian exposure and toxicity.
Potential problems and limitations: While numerous species are amenable to nest-
box studies, these methods share the same shortcomings as all field tests, e.g., lack of
strict control over study elements such as weather, that yield realistic exposure
regimens but may introduce variability implicit to field settings. However, to address
/'laboratory-to-field//extrapolation errors, integrated studies with laboratory and in situ
toxicity tests and field surveys will directly consider sources of uncertainty that should
be considered in risk characterization (see Figure 2).
References:
Grue, C.E., G.V.N. Powell, and M.J. McChesney. 1982. Care of nestlings by wild
female starlings exposed to an organophosphate pesticide. J. Appl. Ecol. 19:327-335.
Grue, C.E. and C.C. Hunter. 1984. Brain cholinesterase activity in fledgling starlings:
implications for monitoring exposure of songbirds to ChE inhibitors. Bull. Environ.
Contain. Toxicol. 32:282-289.
Kendall, R.J., L.W. Brewer, T.E. Lacher, B.T. Marden, and M.L. Whitten. 1989. The
use of starling nest boxes for field reproductive studies: provisional guidance
document and support document. 600/8-89/056. U.S. Environmental Protection
Agency, Environmental Research Laboratory, Corvallis, OR. 82pp.
Robinson, S.C., R.J. Kendall, C.J. Driver, T.E. Lacher, Jr., and R. Robinson. 1988.
Effects of agricultural spraying of methyl parathion on cholinesterase activity and
reproductive success in wild starlings. Environ. Toxicol. Chem. 7:343-349.
Schafer, Jr., E.W., R.B. Brunton, and E.G. Schafer. 1981. Test method for evaluating
the effects of chemicals on starling reproduction. In J.R. Beck (Ed.). Vertebrate Pest
TA4.3 - 2
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Control and Management Materials: Third Conference. ASTM STP , American
Society for Testing and Materials, Philadelphia, PA. Pp.52-61.
Stromborg, K.L., C.E. Grue, J.D. Nichols, G.R. Hepp, J.E. Hines, and H.C. Bourne.
1988. Postfledgling survival of European starlings exposed as nestlings to an
organophosphorus insecticide. Ecol. 69:590-601.
TA4.3 - 3
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TA4.4. In situ clastogenirity tests (chromosomal aberration assay).
The chromosome aberration assay (CA) examines mitotic cells of small mammals
arrested at metaphase for alterations and/or rearrangements in the chromosomes.
Gross microscopic aberrations observed in chromosomes (e.g., breakage,
translocations) have been strongly correlated with the presence of mutagens and may
be closely associated with carrinogenesis.
Test method summary: The assay is widely used and accepted for in vivo analysis of
dastogenic mutagens, and standardized protocols for assays conducted with
laboratory species are available from several sources including Brusick (1980) and EPA
(1985). These protocols have been adapted for in situ toxicity assessment with several
wild mammal species (Baker et al. 1982; McBee et al. 1987; Thompson et al. 1988; Tice
et al. 1987) and should be readily adaptable to other species. Although background
values for chromosome aberrations are available for only a few species of wild
mammals, it remains essential that studies be designed with concurrent chromosomal
aberration analysis at reference sites.
Intended use: By design, collections of small mammals require field surveys and
trapping sessions, but the methods of analysis are equally conducive to diagnostic
laboratory studies that may be completed concurrently with laboratory toxicity tests
(see TA1.9). Field collected small mammals from on-site and at reference locations,
when analyzed, may potentially suggest past exposure history to clastogens, provided
adequate reference locations are identified.
Previous applications/regulatory precedence: In situ clastogenirity tests have a
limited data base in the applied ecology literature, but has been used increasingly
when sites with known contaminant histories that involve potentially dastogenic
agents (e.g., McBee, et al. 1987).
Requirements for development and implementation: Few technical support
laboratories are currently providing in situ dastogenidty tests with vertebrates, and
implementing this biological assessment within an ecological effects assessment for
Superfund may be difficult owing to an absence of experienced testing services.
Potential problems and limitations: Reference locations must be adequately described
and identified to establish a chromosomal aberration baseline for comparisons to site-
collected small mammals. Parallel laboratory studies should also be considered, if
possible.
TA4.4-1
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References:
Baker, R.J., M.W. Haiduk, L.W. Robbins, A. Cadena, and B.F. Koop. 1982.
Chromosomal studies of South American bats and their systematic implications.
Special Publication. Pymatuning Laboratory. Ecol. 6:303-327.
Brusiek, D. 1980. Protocol 13: Bone marrow cytogenetic analysis in rats. In:
Principles of Genetic Toxicology. Plenum Press. New York, NY.
McBee, K., J.W. Bickhman, K.W. Brown, and K.C. Donnelly, 1987. Chromosomal
aberrations in native small mammals (Peromyscus leucopus and Sigmodon hispidus) at a
petrochemical waste disposal site: I. Standard karyology. Arch. Environ. Contam.
Toxicol. 16:681-688.
Thompson, R.A., G.D. Schroder, and T.H. Connor. 1988. Chromosomal aberrations
in the cotton rat, Sigmodon hispidus, exposed to hazardous waste. Environ. Molec.
Mutagen. 11:359-367.
Tice, R.R., B.C. Ormiston, R. Boucher, C.A. Luke, and D.E. Paquette. 1987.
Environmental biomonitoring with feral rodent species. In: Short-term bioassays in
the analysis of complex environmental mixtures. V. (Sandhu, S.S., D.M. Demarine,
M.J. Mass, M.M. Moore, and J.L. Mumford, eds.) Plenum Press. New York, NY.
US EPA. 1985. Toxic Substances Control Act Test Guidelines; Final Rules. 40 CFR
parts 796, 797, and 798.
TA4.4-2
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TA4.5. Field testing with sago pondweed (Potamogeton pectinatus).
As with most field tests, recent efforts (see Kapustka 1991; Kovacs 1978) with
emergent vegetation during wetland evaluations has yielded a native plant test that
may be applicable on a site-specific basis. Unlike laboratory methods standardization,
in situ toxirity tests and field survey methods must provide guidelines for measuring
test endpoints and be flexible in order to be amenable to site-specific requirements.
For example, as a guide, in this in situ toxicity test, sago pondweed (Potomogeton
pectinatus) winter buds may be placed in suspect wetland sediments.
Test method summary: Generally, turions may be purchased from commercial
sources or may maintained from laboratory stock, and should be sprouted in the
laboratory two weeks before initiating field testing. For transport to field locations
sprouted buds may be placed in dampened paper towels, sealed in a Ziplockta plastic
bag, and held at 4°C in a cooler during transit. All plants should be held in transit
less than 18 hours and be planted on-site the same day they were transported. To
plant the sprouted buds on-site, baseline root and shoot measurements must be taken,
then the roots of each plant should be placed into in situ exposure containers that
have had their bottom removed (e.g., 125 ml septum bottle). The young stems
emerged through the septum hole of each bottle should then be sealed with lanolin,
and the bottle placed into the sediment. To ensure adequate harvest on test
termination, at least six replicate bottles should be placed at each on-site testing
location. Test plants may be harvested after seven days in the field exposures, and
root and shoot measurements should be taken.
Intended use: The method briefly referenced here is early in the standardization
process, and should not be a "stand alone" test. The strengths of the method lies in
its contribution to weight of evidence arguments that support laboratory methods (see
TA2.4) that evaluate adverse effects associated with wetland sediment exposures; in
particular, the in situ test with emergent vegetation addresses recurring issues related
to "laboratory-to-fieldextrapolation." As with all field tests, the in situ test with sago
pondweed was designed to address methodological biases associated with laboratory
tests, and the disadvantages implicit to field testing (e.g., less constant exposure
regimen) are offset by the potential information gained, particularly for ecological
interpretations, regarding ambient expressions of toxicity.
Previous applications/regulatoryprecedence: In situ tests with sago pondweed have
been reported in the applied ecology literature, primarily in wetlands or in evaluations
of dredge materials and sediments.
TA4.5 -1
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Requirements for development and implementation: Few technical support
laboratories are currently providing these tests, and implementing this biological
assessment within an ecological effects assessment for Superfund may be difficult
owing to an absence of experienced testing services.
Potential problems and limitations: The in situ test summarized here should be
conducted in parallel with the laboratory test (see TA2.4) to gain the most information
regarding wetlands potentially impacted by contaminated soils or sediments and the
potential ecological effects associated with these matrices. The in situ test will be
subject to various sources of relatively uncontrolled and potentially confounding
effects (e.g., surface water temperature), but through field testing an improved
assessment of ecological effects can be developed, particularly when ambient toxirity
may be confounded by site-specific habitat limitations, e.g., redox potential of
sediments.
References:
Kapustka, L. 1991. Evaluating exposure and ecological effects with terrestrial plants.
Proceedings of a workshop for the US EPA Exposure Assessment Group. 108pp.
Kovacs, M. 1978. Element accumulation in submerged aquatic plant species in Lake
Balaton, Hungary. Acta. Bot. Aca. Sci. Hung. 24(3-4): 273-284.
TA4.5 - 2
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TA4.6. On-site and in situ testing with terrestrial plants.
While field trials in agricultural science have a long history in evaluating the success of
plant breeding and cultivation practices, only recently have designed field tests been
initiated within ecological effects assessments (Nwosu, et id. 1991). Here, "laboratory-
to-field" extrapolation error and reduced soil manipulations helped focus the field
design.
.2
Test method summary: In on-site tests, soil samples should be collected from 3-4 m
areas that are identified during the early phases of field study design. At each
sampling station, top soil should be sampled, then placed into test containers that are
filled to within 2-3 cm of the container's mouth. After soil samples are loaded into
test containers, no more than 20 seeds should be planted in each test container.
Seeds may be standard commercial varieties [cucumber (Cucumis sativus), lettuce
(Lactuca sativa), radish (Raphanus sativa), red clover (Trifolium pratense)], or native
species that have been selected on the basis of their site-specific relevance. After
planting, all seeds should be covered with 25g of 16-mesh cover sand. Temporary
shading or portable green houses/cloches may be beneficial to a successful completion
of on-site tests. Throughout the test, soil moisture and temperature should be
monitored in order to track environmental conditions that could potentially confound
on-site test results. For example, depending on their moisture level, soil samples
should be irrigated to assure that soil moisture is similar across "treatments" and
"species". Test soils, including the controls, should be performed in duplicate.
Quality control exposure chambers should use 20-mesh silica sand as test matrix
overlain with 16-mesh cover sand; seed performance [e.g., germination rates, viability]
should be evaluated in these controls. All test containers were covered with lids and
left under a shaded structure for at least eight days, depending upon species. Upon
termination, germination should be evaluated and any additional end points
previously identified as pertinent to the soil contamination evaluation (e.g., root length
or biomass) should be measured.
Intended use: The on-site test with terrestrial vascular plants was designed to
complement the laboratory seed germination and root elongation tests (see TA2.1).
While commercial species were used in testing, alternative seeds (AOSA1990) may be
used if possible, particularly within an ecological effects assessment.
Previous applications/regulatory precedence: Within the context of ecological effects
assessment, on-site testing with terrestrial vascular plants has a limited history at
present.
TA4.6 - 1
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Requirements for development and implementation: Few technical support
laboratories are currently providing these tests, and implementing this biological
assessment within an ecological effects assessment for Superfund may be difficult
owing to an absence of experienced testing services.
Potential problems and limitations: On-site testing with vascular plants should be
conducted in parallel with the laboratory tests (e.g., see TA2.1, TA2.2) to gain the
most information regarding soil toxicity and its potential ecological effects. The on-site
test will be subject to various sources of relatively uncontrolled and potentially
confounding effects, but through field testing an improved assessment of ecological
effects can be developed, particularly when ambient toxicity may be confounded by
site-specific habitat limitations, e.g., physical alteration of soils.
References:
AOSA [Associationof Official Seed Analysts]. 1990. Rules for testing seeds. J. Seed
Tech. 12:1-122.
Nwosu, J.U., H. Ratsch, and L.A. Kapustka. 1991. A method for on-site evaluation
of phytotoxicity at hazardous waste sites. In J.W. Gorsuch, W.R. Lower, W. Wang,
and M. A. Lewis (Eds.). Plants for toxicity assessment: Second volume, ASTM STP
1115, American Society for Testing and Materials, Philadelphia, PA. Pp.333-340.
TA4.6 - 2
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TA4.7. Distribution of resistant microorganisms.
There have been a number of studies in which increased numbers of toxicant-resistant
organisms have been shown to correlate with the presence of a toxicant. Most forms
of resistance involve some mechanism of detoxificationand the genetic information for
the resistance usually resides in bacterial plasmids which may also code for antibiotic
resistance (Gadd and Griffiths, 1978). Of the heavy metals, Hg and Cd resistances
have been most extensively studied because bacterial strains have been found which
show resistance to these metals and there is great difference between sensitive and
resistant strains. In some cases, the difference can be as high as 100 times. The
differences between sensitive and resistant strains to Pb, Ni, Co, and Ag are too slight
to be of use. Cadmium resistance has only been found in Staphylococcus aureus.
Test method summary: There are a number of ways in which resistant populations
can be estimated, mostly by plate counts or Most Probable Number (MPN)
determinations using broth cultures in test tubes. In both cases, numbers of microbes
on media with and without the toxicant present are determined. In some cases, a
range of toxicant concentrations is used. As indicated above, this approach is
probably most appropriately applied to defining the biological activity of Hg and Cd,
assuming that the higher the toxicity of these heavy metals, the higher the number of
resistant microorganisms.
Intended use: Increased numbers of toxicant resistant organisms occur in the
presence of a toxicant, indicating presence of toxicant at ecologically significant levels.
Previous applications/regulatoryprecedence: Houba and Remade (1980) conducted a
study in which there was a positive correlation between heavy metal concentrations
and increased resistance to heavy metal toxicity in bacterial populations in the
sediments of a zinc-copper factory sedimentation pond. Similar results were reported
by Horner and Hilt (1985) in their study of bacterial heavy metal resistance in Zn
polluted stream sediments. Increased resistance has also been reported in metal-
contaminated soils (Duxbury and Bicknell, 1983) and in estuarine sediments (Mills and
Colwell, 1977; Timoney et al., 1978). These studies are in contrast with a study by
Dean-Ross and Mills (1989) in which bacterial community structure and function was
analyzed along a density gradient of heavy metal concentrations in an Indiana river.
They saw no effect of the gradient on viable and total bacterial numbers, heterotrophic
activity, resistance to Pb or Cu, or species diversity. The lack of correlation was
thought to be related to the high pH of the river and the resulting reduction in heavy
metal toxicity. It would have been much more helpful if this work had been done in
conjunction with more traditional Daphnia or similar studies to determine if any other
TA4.7-1
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indications of toxirity could be found in terms of biological responses. What was
surprising in this study was that there was no indication of increased resistance to
heavy metal toxicity. This points out to the necessity of knowing the chemical form of
the metals in the environmental samples being tested.
Requirements for development and implementation: A much more extensive
database is needed. Major factors whose effects must be considered in establishing a
baseline for interpretation of these measurements are: (1) correlation of resistant
strains with occurrence and concentration of a toxicant in soil, (2) standardization of
methodology, (3) effect of toxicant sorption by clay and organic fractions, (4)
determination of the disappearance rate of resistant organisms once the toxicant is
degraded or removed from the soil.
Potential problems and limitations: At the present time, it is very difficult to define
those organisms that are resistant vs. those that are susceptible to the toxic effects of
heavy metals (Trevors, 1989). One of the most important complicating factors is the
precipitation of metals by the test organisms (Rayner and Sadler, 1990). Microbial
populations exposed to organic toxicants can use the toxin as an energy source,
reduce toxicity in its immediate environment, use it as a cometabolite, or change it's
structure (i.e., membrane composition or transport capability) or function (induced
enzymes) to cope with the presence of the toxic substance. Even with these
problems, Trevors et al. (1985) have suggested that organisms capable of metal
transformations that show a particularly high tolerance to metals might be used as
indicator organisms in defining metal polluted and unpolluted soils. The increase in a
resistant population however, does not imply that there has been a change in
microbial function unless that function is specifically impacted by the presence of the
heavy metal (Domsch, 1984).
References:
Dean-Ross, D. and A. L. Mills. 1989. Bacterial community structure and function
along a heavy metal gradient. Appl. Environ. Microbiol. 55: 2002-2009.
Domsch, K. H. 1984. Effects of pesticides and heavy metals on biological processes in
soil.
Duxbury, D. D. and 6. Bicknell. 1983. Metal-tolerant bacterial populations form natural
and metal-polluted soils. Soil Biol. Biochem. 15:243-250.
TA4.7-2
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Gadd, G. M. and A. J. Griffiths. 1978. Microorganisms and heavy metal toxieity.
Microbial Ecol. 4: 303-317.
Homer, S. G. and B. A. Hilt. 1985. Distribution of zinc-tolerant bacteria in stream
sediments. Hydrobiologia 128:155-160.
Houba, C. and J. Remade. 1980. Composition of the saprophytic bacterial
communities in freshwater systems contaminated by heavy metals. Microb. Ecol. 6: 55-
69.
Mills, A. L. and R. R. Colwell. 1977. Microbiological effects of metal ions in
Chesapeake Bay water and sediment. Bull. Environ. Contain. Toxicol. 18: 99-103.
Rayner, M. H., and P. J. Sadler. 1990. FEMS Microbiol. Let. 71: 252-258.
Timoney, J. F., J. Port, J. Giles and J. Spanier. 1978. Heavy-metal and antibiotic
resistance in the bacterial flora of sediments of New York Bight. Appl. Environ.
Microbiol. 36:465-472.
Trevors, J. T., K. M. Oddie and B. H. Belliveau. 1985. Metal resistance in bacteria.
FEMS Microbiol. Rev. 32:39-54.
Trevors, J. T. 1989. The role of microbial metal resistance and detoxification
mechanisms in environmental bioassay research. Hydrobiologia 188/189:143-147.
TA4.7-3
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TA4.8. Decomposition rates
Decomposers play important roles in agricultural systems, in accelerating
decomposition following soil tillage, increasing retention of nutrients in microbial
biomass when activity and numbers are high, and as sources of nutrients for plants
when grazer activity releases nutrients, particularly nitrogen, from the decomposer
biomass (Coleman 1985, Groffman et al. 1987, Blair and Crossley 1988, Hendrix, et al.
1986). Changes in the decomposer community following soil disturbance and residue
mixing are key components of understanding the effects of management practices in
agricultural ecosystems (Hendrix et al. 1987).
Measurement of decomposition by weight loss, CO2 released or nutrient loss per unit
time integrates a complex set of processes. These processes, in all ecosystems, result
in similar end products from litter material: CO2, mineral elements, recalcitrant organic
material, and microbial and fauna! biomass. Litter composition and the activities of
soil microflora and fauna as regulated by abiotic factors determine the rate of
decomposition processes and the end products of decomposition. Disturbances such
as drought, fire, flooding, grazing, sulfur additions and cultivation can drastically
affect the components of the decomposition process.
Test method summary: While different methods may generally measure similar
endpoints, the mechanisms measured are not the same. Thus, understanding of the
system will be improved by using a combination of methods.
Respiration Methods. A wide range of methods have been used to determine
CO2 evolution and in some cases, O2 utilization in soil (reviewed by Coleman
and Sasson, 1980 and Swift et al., 1979), but the most common is the use of
base traps to absorb CO2 over discrete time intervals. However, the
relationship between CO2 evolved per unit soil per unit time has not been
correlated quantitatively with litter decomposition rate or with total amount
decomposed (Gupta and Singh, 1981b). Respiration methods should be used as
comparative measures since they integrate a range of processes which are often
difficult to separate.
Weight Loss Measures: Confined Utter. Litter is collected, dried, weighed,
and placed into litter bags (Witkamp and van der Drift, 1961) or plastic
containers with mesh covering the open ends (Koellingand Kucera 1965,
Nyhan 1976, Vossbrink et al. 1979). When plant material is collected, the
species composition, age, substrate quality, chemical composition and structural
composition (such as ratio of leaves, stems and buds) should represent the litter
TA4.8 -1
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inputs of interest. The packing of litter should also simulate the density of litter
per unit area observed in the field. The mesh size of the litter bag or covering
must be considered carefully as well as the mesh material (nylon decomposes
in sunlight). A wide range in mesh sizes have been used in litter bag studies:
small mesh sizes (<1 mm) have been used to keep out soil animals and measure
only microbialand abiotic decomposition. However, these mesh sizes do not
restrict nematodes or protozoa from entering. Mesh sizes of 1-7 mm2 can
restrict access of the larger comminuting organisms while larger mesh sizes are
not thought to restrict soil animals. Comparison of litter bag decomposition
rates with those of unconfined litter indicates that confinement reduces rates of
weight loss (Wiegertand Evans, 1964). Litter bags or containers in the field
accumulate soil by influx with water and soil animals. Soil contamination can
be corrected by determining ash-free dry weight. Confined litter can also be
invaded by roots and tillers. Differentiationof original litter from these
materials and from root exudates or other chemical additions can be difficult. If
bags are placed on the surface of the litter or soil, the color of the bag may
induce solar heating and increase temperature in the litter bag. All these
factors must be considered when interpreting data from these methods.
Unconfined Utter (Wiegert-Evans paired plots). There are three assumptions
necessary for paired plots to measure litter decomposition accurately; 1) rates of
decomposition would have been the same in both plots if no disturbance had
occurred, 2) the species composition of living and dead material is identical in
both plots and 3) no outside material entered either plot during the time
increment. Coleman and Sasson (1980) and Swift et al. (1979) discuss the
problems of material additions not included in aboveground standing crop of
plants such as from roots, algae, mosses, and inputs from weather-mediated
events such as floods. In this technique, the rate of instantaneous
disappearance of litter in g/g'd is described as ^(Wj/WjX^-t,,), where W0 =
dead material removed from a given area at the start (t=0) and Wt = dead
material removed from a second area at t = 1 (Wiegertand Evans, 1964).
Intended use: To indicate whether decomposer-grazer community interactions are
proceeding at the same rate as the control site or if they have been altered by chemical
applications.
Previous applicatkms/regulatoryprecedence: Decomposition in different grassland
ecosystems has been measured using CO2 evolution, weight loss from litter bags and
paired plot comparisons. Soil respiration rates increase with increasing temperature
and moisture. Boreal climates have the lowest annual respiration rates because of the
TA4.8-2
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limited period in which microbial activity can occur. In addition, moisture can either
be in excess where through-flow is hampered by frozen layers below the thawed
surface, or be limiting in arid tundras (Heal et al. 1981). In both cases respiration is
reduced. Temperature and moisture effects on individual litter types from tundra
sites is extensively discussed by Flanagan and Veum (1974) and Heal and French
(1974) and respiration accounts for only 50-100% of total observed weight loss, with
leaching responsible for the rest of the loss in the first year. In both temperate and
tropical systems, CO2 evolved averages between 400-600 g C/m2/y. Reductions below
this level can often be explained by adverse abiotic factors limiting microbial activity.
For example, reduction in rates of CO2 evolution in arid grasslands such as those in
eastern Washington, Colorado, South Dakota, Chile, Australia and Africa can be
explained by high temperatures and low moisture limiting microbial activity. The
highest rates of respiration occur in the monsoon climate of India, where abiotic
factors probably are optimal throughout the year.
Litter bag decomposition rates are highly variable within climates. Annual rates
ranging from, on average, 15 to 86% can be found in both boreal and temperate
grasslands. Even in tropical systems, rates as low as 31% loss per year have been
observed. Weight loss rates of 100% per year routinely occur only in tropical systems
except in one case in a grassland in Ireland, Festuca litter completely disappeared in
less than one year. Thus, while climate has some effect on litter bag weight loss, it is
not as important as for respiration rates.
Other factors may determine weight loss from litter bags. Wiegert and Evans (1964),
Koelling and Kucera (1965), Curry (1969), Nyhan (1975), Vossbrinck et al. (1979), and
Wieder et al. (1983), explained differences in the decomposition rates of different litter
substrates by the ratio of C to N. Landscape position along a catenary sequence also
alters decomposition signficantly. Higher rates were observed in the footslope or
swale positions of a catena as opposed to the upland or shoulder positions (Wiegert
and Evans, 1964, Landerholm and Hadley, 1975). The effects of microbial and fauna!
activities also influence decomposition rates (Uvarov, 1982, Wasilewska et al. 1981,
Curry, 1969, Dickinson, 1983, Vossbrinck et al. 1979, Heal and French 1974). Seasonal
cycles and variation from year to year are responsible for many observed differences
(Wiegert and Evans, 1964). While most decomposition occurs during the growing
season, Abouguendia and Whitman (1979) observed that decomposition continued
through the winter in a midgrass prairie, albeit at lower rates. Similar results have
been obtained in shortgrass prairies and high mountain meadows under 1-5 feet of
snow in the winter (Ingham et al., 1986). Thus, it is not representative to restrict
measurement of decomposition to the growing season and decomposition must be
TA4.8 - 3
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measured for a number of years during all seasons before representative rates can be
accurately determined and valid comparisons made.
Requirements for development and implementation: The literature abounds with
decomposition studies, but needs to be summarized in some manner. Litter
decomposition methods need to be standardized (placed on a yearly basis, accounting
for seasonal variation in rates) and results assessed with respect to substrate quality
and nutrient content.
Potential problems and limitations: Decomposition is a cumulative measurement of a
number of component processes. As an integrative measure, it is indicative of a sum
of many processes, any one of which could be affected by a single chemical impact.
Often, the impact of a single chemical impact can not be detected on such a braod-
scale measurement.
References:
Abouguendia, Z.M. and W.C. Whitman. 1979. Disappearance of dead plant material
in a mixed grass prairie. Oecologia (Berl.) 42:23-29.
Blair, J.M. and D. A. Crossley, Jr. 1988. Litter decomposition, nitrogen dynamics and
litter microarthropods in a southern Appalachian hardwood forest 8 years following
cleacutting. Jr. Appl. Ecol. 25:683-698.
Coleman, D.C., 1985. Through a ped darkly: An ecological assessment of root-soil-
microbial-faunalinteractions. In: A. H. Fitter, D. Atkinson, DJ. Read, and M.B.
Usher (Editors), Ecological Interactions in Soil. Blackwell Scientific Publications,
Cambridge, U.K. pp. 1-21.
Coleman, D.C. and A. Sasson. 1980. Decomposer subsystem. In: Breymeyer, A.J.
And G.M. Van Dyne (eds.). Grasslands, Systems Analysis and Man. Cambridge
Univ. Press, p. 609-655.
Curry, J.P. 1969. The decomposition of organic matter in soil. Part 1. The role of
fauna in decaying grassland herbage. Soil Biol. Biochem. 1:253-258.
Dickenson, N.M. 1983. Decomposition of grass litter in a successional grassland.
Pedobiologia 25:117-126.
TA4.8 - 4
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Flanagan, P.P. and A.K. Veum. 1974. Relationships between respiration, weight loss,
temperature and moisture in organic residues on tundra. In: Holding, A.J., O.W.
Heal, S.F. MacLean, Jr. and P.W. Flanagan (eds.). Soil Organisms and Decomposition
in Tundra. Tundra Biome Steering Committee, Stockholm, pp. 249-277.
Groffman, P.M., P. F. Hendrix, C. Han and D.A. Crossley, Jr. 1987. Nutrient cycling
processes in a southeastern agroecosystem with winter legumes, pp. 7-8. IN J.F.
Power (ed.) The Role of Legumes in Conservation Tillage Systems. Soil Conservation
Society of America, Ankeny, IA
Gupta, S.R. and J.S. Singh. 1981. Soil respiration in a tropical grassland. Soil Biol.
Biochem. 14:261-268.
Heal, O.W. and D.D. French. 1974. Decomposition of organic matter in tundra. In:
Holding, A.J., O.W. Heal, S.F. MacLean, Jr. and P.W. Flanagan (eds.). Soil
Organisms and Decomposition in Tundra. Tundra Biome Steering Committee,
Stockholm, pp. 279-310.
Heal, O.W., P.W. Flanagan, D.D. French and S.F. MacLean, Jr. 1981.
Decomposition and accumulation of organic matter. Iru Bliss, L.C., O.W. Heal and
J.J.Moore. Tundra Ecosystems: a Comparative Analysis. Cambridge University
Press, Cambridge, England, pp. 587-633.
Hendrix, P.P., R.W. Parmelee, D.A. Crossley, Jr., D.C. Coleman, E.P. Odum, and
P.M. Groffman. 1986. Detritus foodwebs in conventional and no-tillage
agroecosystems. Bioscience 36:374-380.
Hendrix, P. F., D.A. Crossley, Jr., D.C. Coleman, R.W. Parmelee and M.H. Beare.
1987. Carbon dynamics in soil microbes and fauna in comventional and no-tillage
agroecosystems. INTECOL Bull. 15:59-63.
Ingham, E.R., J.A. Trofymow, R.N. Ames, H.W. Hunt, C.R. Morley, J.C. Moore, and
D.C. Coleman. 1986a. Trophic interactions and nitrogen cycling in a semiarid
grassland soil. Part I. Seasonal dynamics of the soil foodweb. J. Appl. Ecol. 23:608-
615.
Koelling, M.R. and C.L. Kucera. 1965. Dry matter losses and mineral leaching in
bluestem standing crop and litter. Ecology 46:529-532.
TA4.8 - 5
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Landerholm, W.A. and E.B. Hadley. 1975. Decomposition of natural and simulated
litter in OakviUe Prairie, to Wali, M.K. (ed.). Prairie: a Multiple View. University
of North Dakota Press. Grand Forks, USA. pp. 199-207.
Nyhan, J.W. 1976. Influence of soil temperature and water tension on the
decomposition rate of carbon-14 labelled herbage. SoilSci. 121:288-293.
Swift, M.J., O.W. Heal and J.M. Anderson. 1979. Decomposition in Terrestrial
Ecosystems. Univ. of Calif. Press, Berkeley.
Uvarov,A.V. 1982. Decomposition of clover green matter in an arable soil in the
Moscow region. Pedobiologia 24:9-21.
Vossbrinck, C.R., D.C. Coleman and T. A. Wooley. 1979. Abiotic and biotic factors in
litter decomposition in a semi-arid grassland. Ecology 60:265-271.
Wasilewska, L., E. Paplinska and J. Zielinski. 1981. The role of nematodes in
decomposition of plant material in a rye field. Pedobiologia21:182-191.
Wieder, R.K., J.E. Carrel, J.K. Rapp and C.L. Kucera. 1983. Decomposition of tall
fescue (Festuca ehtior var. arundinaceae) and cellulose litter on surface mines and a
taUgrass prairie in central Missouri, USA. J. Applied Ecology 20:303-321.
Wiegert, E.G. and F.C. Evans. 1964. Primary production and disappearance of dead
vegetation on an old field in southeastern Michigan. Ecology 45:49-63.
Witkamp, M. and van der Drift, J. 1961. Breakdown of forest litter in relation to
environmental factors. Plant and Soil 15:295-311.
TA4.8 - 6
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TA4.9. Mycorrhizal colonization
Colonization (or infection) of plants by vesicular-arbuscular mycorrhizal fungi (VA),
ectomycorrhizal fungi (EM), or ericoid mycorrhizal fungi changes the physiology of
the plant host of the colonizing fungi. Mycorrhizal colonization influences root
exudation, carbon allocation, and nutrient uptake by plants and affects (1) the
belowground foodweb, including bacteria, saprophytic fungi, protozoa, nematodes,
and microarthropods; (2) pathogen attack of plants; (3) litter decomposition rates; (4)
aboveground grazing by herbivores; (5) plant growth hormones; (6) soil aggregation; •
and (7) plant competition, community structure, and succession (see review by
Ingham and Molina, 1990). Certain species of plants have never been found in
association with mycorrhizal fungi, while some join in symbiosis only under certain
conditions, while still others appear to be obligately dependent on the mycorrhizal
fungi for growth (Janos, 1987; Reeves, 1985). Thus, the presence or absence of the
fungal inoculum can be of critical importance to some plant species, while being of no
consequence at all to certain other species (Klopatekef al. 1988). Facultatively-
dependent plant species may or may not require the presence of the fungus,
depending on the fertility of the soil. When fertility is high, the plant may not need
the fungus to survive or grow, but when fertility is limited, the fungus may be
required for survival.
Test method summary: Percent root colonization. Roots are obtained either by
cutting from intact plants or by sieving (suggested mesh size 2.5 mm; Mosse et al.
1981). To determine vesicular-arbuscular mycorrhizal (YAM) or ericoid mycorrhizal
colonization, roots are examined after clearing in KOH (concentration depends on
suberizationof roots) and staining with trypan-blue (Harley and Smith, 1983; Largent,
et al. 1980). Total root length and the length of the root occupied by VAM structures
(arbuscules and vesicles) are measured. Actual length of root can be measured, or
line-intersept methods can be used. VAM fungal species can be identified if spores
are present.
Ectomycorrhizal colonization is determined by counting the number of root tips
occupied by the fungus, versus the total number of root tips (France and Reid, 1983).
Morphological characteristics of the various EM fungal species can be used for
identification. The morphology of the different fungal species depends on the plant
species being utilized and only a few individuals in the U.S. are trained to identify EM
species in this manner, thus limiting identification of EM community structure. Keys
do not yet exist which show the appearance of different EM fungi on different species
of plants.
TA4.9 -1
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Bioassays. Soil is taken from the field and a two-fold dilution series prepared by
addition of sterile soil or sterile sand (i.e., free from mycorrhizal inoculum). Axenic
seedlings are planted in soil series and their growth monitored. At two to six months,
dependent on the bioassay plant used, seedlings are removed and their root systems
examined for percent colonization (see above). Using a Most Probable Number Table,
the concentration of viable inoculum present in the original soil can be determined
(Parke et al., 1984).
Intended use: To assess the mutualistic interaction between plants and these
symbiotic fungi. Reduction in colonization of dominant plant species will affect
survival, growth and reproduction of obHgately mycorrhizal plants, and will reduce
the competitive ability of facultatively mycorrhizal species in low fertility soils. Loss of
mycorrhizal inoculum will restrict plant community succession following disturbance,
especially if toxic contaminants prevent re-establishment of the fungus in the system.
Detrimental chemcial applications will remove, reduce, or slow fungal colonization,
with resultant reductions in plant growth rates.
Previous applications/regulatoryprecedence: Numerous articles have been written,
and several journals are devoted to documenting the mutualism between plants and
fungi. In mine spoil restoration, the need for mycorrhizal inoculum is well
documented and regulatory requirements for the proper storage of topsoil to insure
inoculum survival have been implemented (see Miller, 1987).
Requirements for development and implementation: Interpretation of the ecological
significance of reduction in mycorrhizal colonization of dominant or bioassay plants is
difficult and needs further research. Improved methods for identifying community
structure of mycorrhizal fungi from field soils are needed. The impact of specific
chemcials on mycorrhizal colonization should be assessed as well.
Potential problems and limitations: The ecological significance of reduction in
mycorrhizal colonization of dominant or bioassay plants needs to be quantified, and
methods for identifying mycorrhizal fungi from field soils are needed, as are methods
of predicting the impact of specific chemcials on mycorrhizal colonization.
References:
France, R.C. and Reid, C.P.P. 1983. Interactions of nitrogen and carbon in the
physiology of ectomycorrhizae. Can. J. Bot. 61:964-984.
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Harley, J.L. and Smith, S.E. 1983. Mycorrhizal Symbioses. Academic Press, London.
Ingham, E. R., and R. Molina. 1991. Interactions of mycorrhizal fungi, rhizosphere
organisms and plants. In Barbosa, P. (editor), Microorganisms, plants and herbivores.
John Wiley and Sons, New York.
Janos, D.P. 1987. VA mycorrhizas in humid tropical ecosystems. In Ecophysiology of
VA Mycorrhizal Plants. Safir, G.R. (editor), pp. 107-134. CRC Press, Inc., Boca Raton,
Florida.
KLopatek, C. C., L. F. DeBano, J. M Klopatek. 1988. Effects of simulated fire on
vesicular-arbuscular mycorrhizae in pinyon-juniper woodland soil. Plant and Soil
109:245-249.
Largent, D. L., N. Sugihara, and C. Wishner. 1980. Occurrence of mycorrhizae on
ericaceous and pyrolaceous plants in northern California. Canadian Journal of Botany
58:2274-2279.
Miller, R. M. 1987. Mycorrhizae and succession. Pp. 205-220 in Jordan, m, W.R.,
M.E. Gilpin and J.D. Aber (eds.), Restoration ecology: a synthetic approach to
ecological research. Cambridge University Press, Cambridge, U.K.
Mosse, B. Stribley, D.P. and LeTacon, F. 1981. Ecology of mycorrhizae and
mycbrrhizal fungi. Adv. Microbial Ecology 5:137-210.
Parke, J. L., R. G. Linderman, and J. M. Trappe. 1984. Inoculum potential of
ectomycorrhizal fungi in forest soil from southwest Oregon and northern California.
For Sci. 30:300-304.
Reeves, F.B. 1985. Survival of VA mycorrhizal fungi: Interactions of secondary
succession, mycorrhizal dependency in plants and resource competition. N. Amer.
Conf. on Mycorrhizae 6:110-113.
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TA4.10. Lichens as bio-monitors at Superfund sites.
Lichens are a composite of two separate and distinct types of plants, a single fungus
and one or more algae, living together in a symbiotic association. Both "components"
benefit from the relationship. Fungi are heterotrophicand are important decomposers
of carbon in terrestrial ecosystems. A lichen, however, is autotrophic because of the
photosynthetic capability of the algal component. The fungal component benefits by
having an otherwise unavailable source of food, while protecting the algal component
from insolation and desiccation.
Lichens are well know as pioneer organisms that can colonize harsh environments
where most other plants cannot survive. They are especially tolerant of low water
and high light availability. As a result, lichens are found in almost every type of
habitat, from seacoasts to mountain tops, from the equator to the polar regions, from
deserts to the tropics. They are scarce only in the moist tropical regions. Lichens are
important in soil formation and primary plant succession. They can be important in
the availability of nitrogen. They are less important in cycling of phosphorus and
potassium, and least important in cycling of calcium and magnesium. Lichens
provide an important food source for reindeer and caribou in the tundra and sub-
arctic regions.
Lichens are sensitive to very low levels of air pollution. Some lichens are killed by
pollutant levels that are too low to cause visible injury to other plants. In the late 19th
century, the Finnish lichenologistNylander (1866) was one of the first scientists to
observe that lichen populations are adversely affected by air pollution and that lichens
can serve as indicators of air quality. Lichens are now used extensively as elements in
air quality bio-monitoring networks in the United States and Europe. Some
advantages and disadvantages of using lichens as bio-monitors are described below.
Advantages and disadvantages of lichen bio-monitoring (Wilson 1991)
ADVANTAGES DISADVANTAGES
Lichens are very slow-growingand long-lived. Identification of some species can be difficult and
field work can be very labor-intensive.
Lichens lack roots and must obtain all their water Lichen species at a study site may have already
from the air. been eliminated by pollutants. (Transplants can
alleviate this problem.)
Lichens lack cuticles and stomata and are thus Lichen monitoring gives only a qualitative
unprotected against pollution. indication of air quality, so it is best used in
conjunction with standard monitoring
techniques.
TA4.10-1
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ADVANTAGES
Lichens do not shed tissue so they can
accumulate pollutants over their lives.
Lichens are more sensitive to some pollutants
than are higher plants.
Lichens are ubiquitous and are found in nearly
every habitat.
Lichens are not limited by a growing season as
are higher plants.
Good correlations exist between species
distributions and pollution concentrations.
A low level of technologyis required to collect
lichens.
Lichens indicate the history of regional air quality
when historical records of species distributions
are available.
DISADVANTAGES
Lichen species vary in sensitivity to pollutants, so
selection of a study species must account for this
variability.
Very many environmental variables affect lichen
species distribution, so they must be controlled
or accounted for when designing a study.
Test method summary: The main methods used to study the effects of air pollution
on lichens are distribution studies and floristic studies to determine presence and
abundance of species at the study site, elemental analysis to examine the amount of
pollutants absorbed by the lichens, and laboratory fumigation studies to evaluate the
responses of different lichen species to pollutants. The most thorough studies include
a combination of these techniques and may include species diversity, species density,
species frequency, and percent cover. In general, the lichen species in the study area
are identified and recorded and correlations between pollutant levels and species
present are made (see Denison and Carpenter 1973; Wilson 1991). An analysis of
elemental content of selected samples may be performed and, in conjunction with
information on environmental variables such as wind and topography, will provide
information about exposure pathways and potential sources of pollutants.
Intended use: Lichens are used as monitors primarily for ambient air pollutants such
as sulfur dioxide, hydrogen fluoride, acidic precipitation, ozone, nitrogen dioxide,
heavy metals, and radioactive compounds. The research on heavy metal pollution
(and radioactive compounds for certain federal facilities under CERCLA) is most
relevant to Superfund sites.
TA4.10-2
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Previous applications/regulatory precedence: lichens have been used extensively to
determine the presence of several different gaseous and participate air pollutants/ and
to map their spatial distribution. Superfund sites are not generally primary sources of
gaseous air pollutants usually monitored with lichens, but wind blown distribution of
metal particles and leaching and volatilizationof organic compounds may be a
problem around some sites.
Lichens appear to be very efficient collectors of airborne participate metal pollutants.
In general, lichens are very tolerant of heavy metal contamination and accumulate
metals to high concentrations. However, metal contamination does not produce
observable changes in lichen morphology or health and elemental analysis is required
to determine the extent of contamination. Lichen samples are collected, cleaned and
analyzed by a suitable method, such as atomic adsorption, spectrophotometry, X-ray
fluorescence spectrometry, or neutron activation. Elevated element levels closer to the
site and decreasing with distance from the site, are evidence that the site is the source
of contamination. Puckett (1988; as cited in Wilson 1991) gives detailed descriptions of
the techniques used for elemental analysis of the pollutants, the importance of
standardization of techniques, and a listing of studies on line sources, point sources,
remote sources, regional sources, and temporal sources of heavy metal pollutants.
Although lichens have been used infrequently as monitors for organic compounds,
the available evidence seems to indicate that they are potentially suitable bio-indicators
of atmospheric pollution by organochloride compounds (Wilson 1991).
Requirements for development and implementation: The chemicals of most concern
at Superfund sites for which lichens would be appropriate bio-monitors are heavy
metals and organic compounds. Heavy metals adhering to dust and soil particles can
be transported by the wind and deposited in the area surrounding the site. Organic
compounds that leach out of a site may be volatilized and released to the surrounding
area (Wilson 1991).
Many important aspects must be considered when planning any lichen bio-monitoring
study. Lichen populations are affected by many environmental factors, of which
pollutants are only one. Researchers must be aware that it may be difficult to separate
the effects of one or more specific pollutants from the effects of a multitude of natural
and human influences. Monitoring sites must be carefully planned and the
appropriate literature should be consulted when designing the sampling plan (Wilson
1991). Although not yet applied under Superfund, bio-monitoring techniques using
lichens are in common use for monitoring air pollutants and show great promise for
use in Superfund ecological site assessments.
TA4.10-3
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Potential problems and limitations: Generally, Superfund sites do not contribute to
ambient levels of SO2, HF, NOX/ or O3. Adverse effects on lichens observed near
Superfund sites may be attributable to these pollutants from other sources. Research
is necessary to determine the feasibility of using lichens as bio-monitors for
organochloride compounds (Wilson 1991).
References:
Denison, W.C. and S.M. Carpenter. 1973. A guide to air quality monitoring with
lichens. Lichen Technology, Inc., Corvallis, OR. 39 pp.
Nylander, W. 1866. Les lichens du jardin du Luxembourg. Bull. Roy. Bot. Soc. 13:
364-372.
Puckett, K.J. 1988. Bryophytes and lichens as monitors of metal deposition. In:
Nash HI, T.H. and V. Wirth (Eds.). lichens, bryophytes and air quality. Bibliotheca
Lichenologica 30:231-267.
Wilson, M.J. 1991. Lichens as indicators of air pollution impacts at Superfund sites.
Prepared for US EPA Exposure Assessment Group. US EPA contract 68-DO0100. 84
PP-
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