United States
Environmental Protection
Agency
Office of Research and
Development
Washington DC 20460
EPA/600/3-90/073
August 1990
vvEPA
Impacts on Quality of
Inland Wetlands of the
United States:
A Survey of Indicators,
Techniques, and
Applications of Community
Level Biomonitoring Data
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EPA/600/3-90/073
August 1990
IMPACTS ON QUALITY OF
INLAND WETLANDS OF THE UNITED STATES:
A SURVEY OF INDICATORS, TECHNIQUES, AND APPLICATIONS OF
COMMUNITY-LEVEL BIOMONITORING DATA
by:
Paul R. Adamus
NSI Technology Services Corporation
US EPA Environmental Research Laboratory
200 SW 35th St.
Corvallis, OR 97333
and
Karla Brandt
Center for Wetlands
University of Florida
Gainesville, FL 32611
u.s. Agency
ward, 12th Floor
Chicago, "
Eric M. Preston, Project Officer
USEPA Environmental Research Laboratory
200 SW 35th St.
Corvallis, Oregon 97333
Printed on Recycled Paper
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DISCLAIMER/CREDITS ON CONTRACTS
This project has been funded by the United States Environmental Protection Agency (EPA) and conducted
through contract 68-C8-006 to NSI Technology Services Corporation. It has been subjected to the Agency's
peer review. The opinions expressed herein are those of the authors and do not necessarily reflect those
of EPA The official endorsement of the Agency should not be inferred. Mention of trade names of
commercial products does not constitute endorsement or recommendation for use.
11
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CONTENTS
SECTION PAGE
LIST OF TABLES v
SUMMARY vi
ACKNOWLEDGEMENTS vii
1.0 INTRODUCTION 1
1.1 SCOPE OF COVERAGE and ORGANIZATION 1
1.2 HOW THE REPORT WAS PREPARED 11
2.0 APPROACHES FOR PROTECTING WETLAND QUALITY 14
2.1 REGULATORY BACKGROUND 14
2.2 USE DESIGNATION AND CLASSIFICATION 14
2.3 DEVELOPMENT OF PRIORITIZED USE SUB-CATEGORIES 17
2.4 NARRATIVE CRITERIA TO PROTECT WETLAND DESIGNATED USES 20
2.5 NUMERIC CRITERIA TO PROTECT WETLAND DESIGNATED USES 21
3.0 GENERAL GUIDELINES FOR WETLANDS BIOLOGICAL CHARACTERIZATION 24
3.1 WHAT TO MONITOR 24
3.2 TYPES OF MONITORING 27
3.3 STUDY DESIGN 30
3.4 DATA ANALYSIS AND INTERPRETATION 31
4.0 WETLAND MICROBIAL COMMUNITIES 35
4.1 USE AS INDICATORS 35
4.2 SAMPLING METHODS AND EQUIPMENT 36
4.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS 37
5.0 WETLAND ALGAE 39
5.1 USE AS INDICATORS 39
5.2 SAMPLING EQUIPMENT AND METHODS 41
5.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS 42
6.0 NON-WOODY (HERBACEOUS) VEGETATION 43
6.1 USE AS INDICATORS 43
6.2 SAMPLING METHODS AND EQUIPMENT 58
6.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS 60
7.0 WOODED WETLAND VEGETATION 62
7.1 USE AS INDICATORS 62
7.2 SAMPLING METHODS AND EQUIPMENT 65
7.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS 66
8.0 WETLAND INVERTEBRATE COMMUNITIES 67
8.1 USE AS INDICATORS 67
8.2 SAMPLING METHODS AND EQUIPMENT 76
8.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS 79
ill
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9.0 WETLAND FISH COMMUNITIES 82
9.1 USE AS INDICATORS 82
9.2 SAMPLING METHODS AND EQUIPMENT 86
9.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS 88
10.0 WETLAND AMPHIBIANS AND REPTILES 90
10.1 USE AS INDICATORS 90
10.2 SAMPLING METHODS AND EQUIPMENT 93
10.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS 95
11.0 WETLAND BIRD COMMUNITIES 96
11.1 USE AS INDICATORS 96
11.2 SAMPLING METHODS AND EQUIPMENT 101
11.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS 102
12.0 WETLAND MAMMAL COMMUNITIES 129
12.1 USE AS INDICATORS 129
12.2 SAMPLING METHODS AND EQUIPMENT 131
12.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS 133
13.0 BIOLOGICAL PROCESS MEASUREMENTS IN WETLANDS 134
13.1 USE AS INDICATORS 134
13.2 SAMPLING METHODS AND EQUIPMENT 136
13.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS 137
14.0 LITERATURE CITED 139
APPENDIX A. Summary of Advantages and Disadvantages of Use of Major Taxa in Monitoring
Wetland Ecological Condition 192
APPENDIX B. Wetland Biomonitoring Sites, Referenced and Mapped by State 199
APPENDIX C. Inland Wetland Community Profile Reports of the U.S. Fish and Wildlife
Service 461
IV
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LIST OF TABLES
Table 1. Examples of Major Federal Laws, Directives, and Regulations for the Management and
Protection of Wetlands 2
Table 2. Potential Metrics for Wetland Biomonitoring 3
Table 3. Examples of Biological Metrics Describing Wetland Community Structure and Function. . . 4
Table 4. Examples of Analytical Metrics, Indices, and Procedures Used for Wetland Community
Studies 5
Table 5. Stressors Addressed in this Report 9
Table 6. Wetland Monitoring Indicators Suggested by Various Scientists 28
Table 7. Examples of Aquatic Macrophytes Tolerant of Saline Conditions in Inland Wetlands 49
Table 8. Examples of Aquatic Plants That May Indicate Reduced Light Penetration Due to Greater
Turbidity or Shade 54
Table 9. Examples of Aquatic Invertebrates That May Indicate Eutrophic Conditions in
Wetlands 68
Table 10. Examples of Aquatic Invertebrates That Tolerate Low-Oxygen Conditions in Wetlands. . . 70
Table 11. Examples of Invertebrates That May Tolerate or Prefer Acidic Conditions in Wetlands. . . 73
Table 12. Examples of Invertebrate Density and Biomass Estimates from Wetlands 81
Table 13. Examples of Wetland Fish Species That Tolerate Low Dissolved Oxygen 83
Table 14. Examples of Wetland Birds Categorized by Major Food Source 98
Table 15. Within-Year Variability of Breeding Bird Richness and Density, Among Wetlands, by
State 108
Table 16. Breeding Bird Richness and Density, by Wetland Type and State 115
v
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SUMMARY
This report describes what is known about ecological community response to anthropogenic stressors in
inland wetlands. Because wetlands are shallow, located in a topographically low position in the landscape,
and have low hydraulic exchange rates, they are particularly sensitive to accumulation of pollutants and
changes in water tables. Despite this situation, and the fact that unimpacted wetlands support exceptional
biological production, government biomonitoring programs to date have focused mainly on rivers and lakes
to the exclusion of wetlands. Monitoring of wetlands has focused mainly on extent of the resource, rather
than changes in wetland quality (e.g., ecological structure and function, condition).
Based on a synopsis of the literature, the report describes the potential effects upon wetland community
structure of the following stressors: eutrophication, organic loading, contaminant toxicity, acidification,
salinization, sedimentation, turbidity/shade, vegetation removal, thermal alteration, dehydration, inundation,
and fragmentation of habitat. The incidence and geographic extent of these stressors in wetlands is currently
unknown. Information is provided concerning the effect of each stressor on potential indicators of wetland
condition-wetland microbes, algae, vascular plants, invertebrates, fish, amphibians, reptiles, birds, mammals,
and selected biological processes in wetlands.
The report describes options for using potential indicators to (a) develop and incorporate biocriteria for the
protection of sustainable ecological conditions, and (b) help identify and prioritize degraded wetlands that
may be candidates for restoration. Because of the lack of appropriate comparative studies of wetlands, the
report does not provide biocriteria for wetlands, evaluate or prioritize potential indicators of wetland
condition, nor endorse specific techniques for wetland biomonitoring and data analysis. Its intended use is
mainly as a technical source document for future design, testing, and reporting of indicators.
The focus is primarily on community-level (as opposed to individual-organism) responses to the stressors.
Techniques for sampling each of the taxonomic groups in wetlands are described generally. To the extent
allowed by published data, the range of density, richness, and diversity within some taxonomic groups is
reported, and most-sensitive species are noted. To facilitate regionalization of future efforts and to further
cooperation among researchers and use/analysis of extant data, the locations of a large portion of published
wetland community studies are depicted on state maps, referenced to state bibliographies. Important
elements in future use and regionalization of this report's information should be continued reviews by other
scientists of literature published after 1989, and expanded compilations of existing data on responses of
individual species.
Copies of this report are available from:
US EPA Center for Environmental Research Information
26 Martin Luther King Drive
Cincinnati, OH 45268
VI
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ACKNOWLEDGEMENTS
We acknowledge the support jointly provided to this effort by EPA's Office of Policy, Planning, and
Evaluation (OPPE) and EPA's Environmental Monitoring and Assessment Program (EMAP). We also
acknowledge with appreciation the contribution made by many individuals during the preparation of this
report. John Maxted, Doreen Robb, and Diane Fish at the EPA Office of Wetlands Protection and F. Kim
Devonald at the EPA Office of Policy, Planning, and Evaluation were instrumental in initiating the effort.
Dr. Mark Brown at the University of Florida served as project officer of the Cooperative Agreement which
provided assistance on the effort.
At the EPA Environmental Research Laboratory in Corvallis, Oregon, Dr. Eric M. Preston, the EPA Project
Officer and Manager of EPA's Wetland Research Program, facilitated external communications necessary
to the project's success. Barbara Hagler was instrumental in locating hundreds of journal articles for
subsequent review. The tasks of data plotting and literature database construction and retrieval were capably
handled by Robin Renteria, Eric Schneidermann, and Jeff Irish, with assistance from Scott Leibowitz and
Donna Frostholm. Jo Ellen Honea and Kristina Heike assisted with formatting and layout.
Within EPA, the final draft was reviewed, in part or in toto, by Doreen Robb, Diane Fish, and Martha
Stout (Office of Wetlands Protection), William Shippen (Office of Water Regulation and Standards), Ruth
Miller (Office of Policy, Planning, and Evaluation), Wayne S. Davis (Region 5), and Louisa Squires (NSI,
Corvallis Environmental Research Laboratory).
From other agencies, reviews (in part or in toto) were provided by Carl Armour, Greg Auble, R. Bruce
Bury, Richard Schroeder, and Michael Scott of the U.S. Fish and Wildlife Service; Barbara Kleiss, K. Jack
Kilgore, Charles Klimas, Thomas Roberts, William Taylor, and James Wakeley of the U.S. Army Corps of
Engineers Waterways Experiment Station; and Dr. James LaBaugh of the U.S. Geological Survey.
From the scientific community, reviews of the final draft were provided by Drs. Robert Brooks, Joan
Ehrenfeld, Jerry Longcore, William Niering, and Fredrick Reid. Early drafts of the report were reviewed
externally by Drs. Robert Brooks, David Cooper, James Karr, and R. Wayne Nelson.
The contributions of the many wetland specialists who responded to our written inquiries are particularly
appreciated. Access to the Wetland Values Database was kindly expedited by Craig Johnson of the U.S.
Fish and Wildlife Service, Division of Endangered Species and Habitat Conservation. Dr. Ronald Hellenthal
of Notre Dame University graciously accessed bioindicator data in the ERAPT database. Bird data collected
by thousands of volunteers and compiled by the Cornell Laboratory of Ornithology and the U.S. Fish and
Wildlife Service was provided by these institutions.
VII
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1.0 INTRODUCTION
The widespread physical loss of North American wetlands has been generally documented (e.g., Tiner 1984).
However, uncertainty exists regarding the ecological condition of the wetlands that remain. Although
wetlands passively provide for many public uses-e.g., water purification, flood control, aquatic life and
wildlife support—the extent to which these functions are being impaired in the remaining wetland resources
is unclear. The Environmental Protection Agency (EPA) has the responsibility under several legal mandates
(Table 1) for determining this.
Wetland ecological quality assumes special significance because of current State and Federal interest in
adopting a policy of "no net loss" of the nation's wetlands. As expressed by EPA's Wetlands Action Plan,
this implies no net loss of either acreage or function. To determine whether particular functions or uses,
such as support of aquatic life, are being impaired in wetlands, "indicators" of these functions must be
identified and protocols articulated for their measurement and interpretation.
1.1 SCOPE OF COVERAGE and ORGANIZATION
This report focuses on inland wetlands of the conterminous United States. Except for those bordering the
Great Lakes, these are not subject to significant tidal fluctuations. They are generally fresh water wetlands,
except for saline wetlands in some mid-continent and western regions. Other tidal, tundra, and tropical
wetlands were not included because their consideration would have involved a greatly expanded scope of
work. Protocols for biological sampling of tidal wetlands have been presented by Simenstad et al. (1989)
and others. For purposes of this report, "wetlands" are considered to be vegetated areas transitional between
uplands and open water.
A principal goal of this report is to encourage each state to track their progress in protecting wetland
ecological condition. As one of many components needed to achieve this, this report identifies data gaps
and provides guidance that describes (a) how existing resource data might be applied in the designation of
"uses" for wetlands, (b) ambient biological criteria for wetlands might be developed or modified, and (c) how
wetlands might be periodically sampled (and data interpreted) to estimate their relative ecological condition,
compliance with biological criteria, or need for restoration. Publication of this report is not intended to
imply that sufficient knowledge exists to develop community-based biocriteria for all wetlands at the present
time.
This report emphasizes the biological functions of wetlands-habitat for fish, wildlife, and related organisms
and the processes that support biological functions. Its purpose is to provide State and Federal water
quality and wetland managers with a synopsis of selected literature describing the community-level response
of wetlands and similar aquatic systems to particular stressors. In most cases, this document does not
synthesize the literature into statements applicable to all wetlands, or to all wetlands, taxa, or stressors
of a certain type. Such a synthesis was generally avoided because the technical literature lacks a sufficient
number of studies that demonstrate causal relationships (as opposed to correlation) or that allow statistical
extrapolation (i.e., synthesis) to entire taxa, stressor types, or wetland types, regions, or states.
Biological sampling can be carried out at several ecological levels-the organism, the population, the
community, or the ecosystem (Table 2). This report focuses on measurements of biological communities,
that is, associations of interacting populations, usually delimited by their interactions or by spatial
occurrence. Tables 3 and 4 show specific metrics (that is, characteristics or indices) used to describe the
communities. This report also discusses, to a more limited extent, the measurement and use of biological
processes as indicators of anthropogenic stress.
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Table 1. Examples of Major Federal Laws, Directives, and Regulations for the Management and Protection
of Wetlands.
Directive
Executive Order 11990
Date
May 1977
Responsible Agency
All agencies
Protection of Wetlands
Executive Order 11988
Floodplain Management
Federal Water Pollution
Control Act (PL 92-500)
as Amended
Section 401- Water
Quality Certification
Section 404- Dredge and
Fill Permit Program
reporting requirements for
Section 305(b)
National Environmental
Policy Act
Coastal Zone Management Act
May 1977
1972, 1977
All agencies
1975
1972
EPA, States
EPA, Corps of Engineers
States
All agencies
Office of Coastal Zone Management
This report's focus on biological communities does not mean other measurements are less important or
useful. Indeed, there are numerous situations where alternative indicators~in particular, wetland flooding
regime, bioaccumulation of contaminants, sedimentation rate, population demographics, and habitat structure
-can more cost-effectively reflect the ecological condition, impact causes, and sustainability of a wetland than
can community-level biological methods. Quantitative literature on the community ecology of wetlands has
been singled out for focus, largely because of current EPA interest in applying this approach when assisting
States with the development of community-based biocriteria for surface waters (Plafkin et al. 1989, USEPA
1987, 1990).
This focus on community-level measurements coincides with a growing body of literature which suggests that,
at least for many applications in flowing waters, monitoring of biological community structure provides cost-
effective information about ecological condition or as some have termed it, "health" (Krueger et al. 1988).
Biological monitoring directly addresses the result of pollution, not its possible cause. Measurements of
community structure can integrate intermittent stressor conditions. They can also detect impacts from many
sources for which chemical criteria are poorly suited to detect (e.g., alteration of hydrologic regimes,
synergistic pollutant effects, nonpoint runoff). If community-level measurements suggest that a stress is
occurring, traditional methods (e.g., direct hydrologic monitoring, tissue analysis, chemical sampling) can be
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Table 2. Potential Metrics for Wetland Biomonitoring.
Organismal Level
Altered Behavioral Responses
o foraging/feeding effectiveness
o response to odors, pheromones, temperature, chemicals
o reproductive behavior (courtship, mating, maternal/paternal)
o predator avoidance (reaction time, evasiveness)
o migratory/dispersal behavior
o social interactions/territoriality
Altered Metabolism/Homeostasis
o thermo/osmo/hydro regulation
o oxygen consumption, photosynthesis
o nutrient uptake and translocation, food conversion efficiency
o enzyme/protein activation/inhibition (e.g., cholinesterase)
o hormone balances
Altered Reproductive Success
o seed set, tillering, flowering, vegetative (clonal) growth
o sexual maturity, conception/implantation, parturition
Altered Growth and Development
o growth rate (e.g., tree ring analysis)
o size at age, morphological abnormalities
Decreased Disease Resistance
Direct Tissue/Organ Damage (e.g., lesions, tumors)
Changes in Stamina (e.g., plant vigor)
Bioaccumulation
Population Level
o survival/mortality
o sex ratio, fecundity
o population abundance, biomass, density
o age structure and recruitment
o gene pool
o intraspecific competition
o population behavior, migration, dispersal
o susceptibility to predation
o population rate of decline or increase
Community Level (see Table 3 for details)
Structure (taxonomic and functional)
Function (process)
Ecosystem Level
Mass Balance of Nutrients
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Table 3. Examples of Biological Metrics Describing Wetland Community Structure and Function.
Community Structure
Abundance. The number of individuals of an organism or organisms. As an analytic metric, tends to
exaggerate the importance of small, abundant species.
Biomass. The weight of living material in all or part of a community. For this report, it includes
measurements of chlorophyll or caloric content as well. As an analytic metric, tends to exaggerate the
importance of large, uncommon species.
Density. The number of individuals of an organism or organisms, per unit area or per unit volume.
Richness. The number of species, size classes, or other functional groups, per unit area or volume, or per
number of individuals.
Diversity. The variety (richness) of species, life forms (physiognomy), genetic material, or functional groups,
taking into account the relative abundance (evenness and dominance) of each species or group.
Community Composition. Qualitative descriptions of the members of a community (e.g., species lists),
perhaps describing as well their relative abundance and grouped by their attributes (e.g., exotic vs. native,
migrant vs. resident, response guild).
Community Attributes
Colonization rates
Stability
o resistance, assimilation capacity
o resilience, recovery rate
Successional relationships
Food web structure, trophic interactions
Competition among species
Predator/prey relationships
Grazing/herbivory relationships
Parasite/host relationships, symbiosis
Community Function (Process)
Decomposition/leaching
Productivity, Photosynthesis, Respiration
Denitrification, Nitrogen Fixation
Other Biogeochemical Functions (e.g., methanogenesis)
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Table 4. Examples of Analytical Metrics, Indices, and Procedures Used for Wetland Community Studies.
Similarity (Comparative) Indices. Metrics that reflect the number of species or functional groups in
common between multiple wetlands or time periods. May be weighted by relative abundance, biomass,
taxonomic dissimilarity, or caloric content of the component species. Includes Jaccard coefficient, Bray-
Curtis coefficient, rank coefficients, overlap indices, the "community degradation index" (Ramm 1988), and
others.
Cluster Analysis and Ordination. Procedures that detect statistical patterns and associations in community
data. Can be used to hypothesize relationships to a stressor. Includes principal components analysis,
reciprocal averaging, detrended correspondence analysis, TWINSPAN, canonical correlation, and others. Can
be used to identify guilds (see below). A useful reference is Pielou (1984), and a cautionary note is
expressed by Reals (1973).
Food Web Analysis. Procedures that measure length of food chains, number of trophic levels, ratio of
number of trophic species to trophic links, and similar measures (e.g., Patten et al. 1989, Turner 1988). As
yet, they have seldom been tested in stressed wetlands.
Tolerance Indices. Metrics that reflect proportionate composition of tolerant vs. intolerant taxa. Includes
saprobic indices, macroinvertebrate EPT index, Hilsenhoff index, and others detailed and compared in
Hellawell (1984) and Washington (1984). "Tolerance" usually means tolerance to organic pollution; tolerance
to many toxicants and physical habitat alterations may not be well-reflected by available indices.
Guild Analysis. Procedures in which individual species are assigned to functional groups (species
assemblages) based on similar facets of their:
o life history
o habitat preference
o trophic level, assumed niche breadth
o size, biomass, caloric content
o toxicological sensitivity
o behavioral characteristics
o phenological characteristics
o sensitivity to human presence
o status as an exotic or indigenous species
o resident vs. migrant status
o harvested vs. protected status
o other factors
Indices of Biotic Integrity. Indices that are a composite of weighted metrics describing richness, pollution-
tolerance, trophic levels, abundance, hybridization, and deformities. Widely used in stream fish studies (see
Karr 1981).
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used to help determine cause. Moreover, ambient biological criteria can directly provide realistic evaluations
of whether specific areas designated for protection of aquatic life are meeting this objective, or require
restoration.
In most cases, if biological community monitoring data are to be correctly interpreted, they should be
collected over time periods spanning several years, and should be accompanied with hydrologic and water
quality measurements. Hydrologic measurements typically describe the variability, temporal pattern, extent,
frequency, depth, and duration of surface waters and/or saturated condition (e.g., Gunderson 1989, Poff and
Ward 1989). They may be expressed, for example, as water residence time distribution, water yield (net
water balance), and stage or flow exceedence curves (i.e., percentage of time a particular water level or
discharge is exceeded). Typical equipment for measuring these includes precipitation gauges, flourescent
dyes, stage-discharge recorders, piezometers, redox probes, and sediment traps. For further information on
the use of hydrologic and sediment measurements in wetland monitoring, readers may find the following
references particularly useful:
Faulkner et al. 1989, Gunderson 1989, Heliotis and DeWitt 1987, Kadlec 1984, Kadlec 1988,
LaBaugh 1986, Rosenberry 1990, USEPA 1985, Van Haveren 1986, Welcomme 1979, and
Zimmerman 1988.
Monitoring protocols for estimating bioaccumulation in wetlands will be published in a manual by the U.S.
Fish and Wildlife Service in 1990. General summaries of aquatic bioaccumulation processes and effects are
contained in Biddinger and Gloss 1984, Fagerstrom 1979, Phillips 1980, Robinson-Wilson 1981, and
Sonstegard 1977. Examples of bioaccumulation studies in wetlands include:
Anthony and Kozlowski 1982, Aulio 1980, Behan et al. 1979, Lambing et al. 1988, Larsen and
Schierup 1981, Mclntosh et al. 1978, Metcalf et al. 1984, Mouvet 1985, Niethammer et al. 1985,
Schierup and Larsen 1981, Stephenson and Mackie 1988, Taylor and Crowder 1983, and others.
It is assumed that each state will determine how best to sample wetlands, incorporate wetland biological
criteria into its water quality management programs, and establish restoration priorities. For this reason,
much of the information contained in this report is presented as "could's" or "might's," and details regarding
"how" the many technical statements should be interpreted and implemented are left to other agencies and
institutions which have diverse goals and which encounter a wide variety of political and environmental
conditions. To date, only a single state (Florida) has drafted survey-based biological criteria for some of its
wetland resource (described by Schwarz 1987).
This report is also intended to serve as once source of technical support for the EPA's National Guidance
on Water Quality Standards for Wetlands, prepared jointly by the Office of Wetlands Protection and the
Office of Water Regulations and Standards. This report pursues this goal partly by providing just one
input-a literature review-for identifying and interpreting biological indicators of wetland ecological
condition.
Many factors other than technical data must be considered in developing biological criteria and setting
restoration priorities. Decisions concerning selection of which resources, uses, or functions to protect or
enhance are inevitably complex, since the criteria for protecting one resource or use may be counter to
protecting another (Duinker and Beanlands 1986, Graul and Miller 1984, Smith and Theberge 1987). A
generalized list of wetland functions or uses that might be the focus of protection or restoration is contained
in Section 404 of the Clean Water Act. These are as follows (from 33 CFR 320 (b)(2)):
a. Food Chain Production (i)
b. General Habitat (i)
c. Research, Education, and Refuges (ii)
d. Hydrologic Modification (iii)
e. Sediment Modification (iii)
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f. Wave Buffering and Erosion Control (iv)
g. Flood Storage (v)
h. Ground Water Recharge or Discharge (vi)
i. Water Purification (vii)
j. Uniqueness/Scarcity (viii)
Any such list could include many additional or more specific values of wetlands, e.g., maintenance of
biodiversity, landscape value as corridors or habitat islands, role in global climate warming, timber harvest.
This report begins, in Section 2, with a description of possible technical approaches that state and local
agencies might use in designating "uses" for wetlands and eventually, perhaps, developing community-level
biological criteria. Section 3 then describes general considerations in the design of wetland biomonitoring
studies. Remaining sections of the report are delimited by major taxonomic groups (e.g., birds, fish). Each
of these taxonomic sections is divided according to the following themes:
Use as Indicators
Sampling Protocols and Equipment
Spatial and Temporal Variability, Data Gaps
Originally, our intent was to organize the discussions by wetland type. This is because wetland types are
generally believed to differ in their community-level responses to particular stressors. Thus, wetland "type"
may be an important qualifier of any biocriteria that might be developed in the future. However, studies
of specific anthropogenic stressors within individual types of wetlands were often so few that attempts to
organize sections by wetland type proved futile. Nonetheless, within the discussions of particular taxa and
stressors, statements about indicator metrics and taxa have been couched whenever possible in terms
descriptive of wetland type/region. Also, attempts were made to organize the descriptions of sampling
techniques according to wetland type. Although sampling protocols and appropriate equipment differ
between flowing-water wetlands, wetlands with standing surface water, and wetlands without surface water,
a finer classification of types is difficult to specify without knowledge of study objectives. Usually, having
a clear definition of the objectives of a particular study is more important to study design than is knowledge
the particular types of wetlands that happen to be included in a study.
The subsection discussions of Use as Indicators attempt to document community-level shifts that occur as
a result of particular anthropogenic stressors. Stressors considered in this report are listed and defined in
Table 5. Their effects on biota are often cumulative and interactive, thus complicating the use of biota as
indicators of any individual stressor. Although several previous documents have summarized impacts to
wetland biota (e.g., Brennen 1985, Brown et al. 1989, Darnell et al. 1976, Davis and Brinson 1980, USEPA
1983, USEPA and USFWS 1984), not all taxa, wetland, and stressor types have been covered and inferences
have commonly been drawn from non-wetland aquatic environments.
It is important to understand that statements made in this report reflect strongly the particular wetland
locations and types that were studied, and considerable uncertainty exists regarding whether such
conclusions (e.g., about the value of specific taxa as indicators) can be transferred to other wetland types
and regions. Many cited studies reflect one-visit or one-season data collections from a single wetland type,
rather than recurrent monitoring. There are very few statistically-valid studies that adequately quantify the
exposures of wetland organisms to stressors using factorial designs, e.g., studying areas both with/without
treatment and staggered before/after measurements (Stewart-Oaten et al. 1986, Walters et al. 1988).
Although it may appear, from the quantity of studies contained in the maps, in their bibliographies, and in
the extensive literature citations in the text that inland wetlands in some regions have been extensively
monitored, in truth relatively little is known about wetland biological response to anthropogenic stressors.
Compared to monitoring of streams and lakes, sampling of wetlands on a recurrent or comparative regional
basis has been almost non-existent, partly due to lack of government sponsorship of wetland biomonitoring
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programs.
Also, the response of a wetland community to anthropogenic stress depends not only on the taxa present
and the severity of the stressor, but also on the geomorphic, physical, and chemical environment of the
wetlands (Adamus et al. 1987, Adamus and Stockwell 1983). For example:
o Wetland biological communities most vulnerable to sedimentation effects might be those located in
shallow basins without outlets, so sediment quickly accumulates;
o Wetland biological communities most vulnerable to eutrophication and contaminant effects might
be those in wetlands that get most of their water directly from precipitation (e.g., ombrotrophic
bogs), which have low alkalinity, and/or which have types of sediments that adsorb (but do not
render biologically unavailable or harmless) the nutrients and contaminants during the short time
that runoff passes through the wetland, e.g. Goldsborough and Beck (1989).
o Wetland biological communities most vulnerable to effects of many anthropogenic changes might
be those that:
(a) have no prior exposure to similar levels or types of stress; and/or
(b) exist in wetland types or regions that are characteristically stable (relatively speaking) over time;
and/or
(c) are physically isolated from sources of colonizers, so that recovery occurs slowly; and/or
(d) are located in regions that have experienced especially rapid losses of wetlands of a similar type.
Considerably more investigation may be required before candidate indicators of wetland ecological condition
can be fairly rated relative to one another, and exact numerical criteria specified. Thus, users of the report
are urged to obtain, whenever possible, assistance from local wetland scientists when attempting to apply
the information reported herein.
In this report, the representativeness, replication, and field and data analysis techniques used by cited studies
were not evaluated; the overwhelming majority of citations are peer-reviewed papers from professional
journals. Also, no attempt is made to give equal coverage to all topics within the general theme, because
availability of data varies greatly among topics.
In the "Use as Indicators" subsections, discussions focus on the community metrics that are defined in Table
4. An important metric that is frequently discussed is "richness." References in this report to the response
of richness to stressors should be assumed, unless otherwise noted, to refer to changes in taxonomic richness
within a wetland. However, readers should be aware that some stressors may increase richness of a major
taxonomic group within a wetland (alpha diversity) while decreasing richness on a regional level (beta and
gamma diversity). This may occur as the result of a net increase in species within the wetland, but an
increase in which regionally rare species originally inhabiting the wetland are replaced by a larger number
of regionally common and widespread species. Thus, no value judgement should necessarily be attached
to statements that richness increases in response to a stressor. Moreover, design of future studies evaluating
changes in community richness should in many cases include information on the regional rarity of species
that may be displaced.
The subsection discussions of Sampling Protocols and Equipment focus on techniques for sampling each
taxonomic group, e.g., how, where, when, and how often sampling has been done. However, this report is
not intended to be a prescriptive manual. Rather, the intent is to present the user with choices. Choices
are provided by summarizing the types of equipment, protocols, and community metrics that have been used
previously to monitor wetland communities. Choices, rather than prescriptions, are given because rigid
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Table 5. Stressors Addressed in This Report
Enrichment/Eutrophication. Increases in concentration or availability of nitrogen and phosphorus. Typically
associated with fertilizer application, cattle, ineffective wastewater treatment systems, fossil fuel combustion,
urban runoff, and other sources.
Organic Loading and Reduced DO. Increases in carbon, to the point where an increased biological oxygen
demand reduces dissolve oxygen in sediments and the water column and increases toxic gases (e.g., hydrogen
sulfide, ammonia). Typically associated with ineffective wastewater treatment systems.
Contaminant Toxicity. Increases in concentration, availability, and/or tenacity of metals and synthetic organic
substances. Typically associated with agriculture (pesticide applications), aquatic weed control, mining, urban
runoff, landfills, hazardous waste sites, fossil fuel combustion, wastewater treatment systems, and other
sources.
Acidification. Increases in acidity (decreases in pH). Typically associated with mining and fossil fuel
combustion.
Salinization. Increases in dissolved salts, particularly chloride, and related parameters such as conductivity
and alkalinity. Typically associated with road salt used for winter ice control, irrigation return waters,
seawater intrusion (e.g., due to land loss or aquifer exploitation), and domestic/industrial wastes.
Sedimentation/Burial. Increases in deposited sediments, resulting in partial or complete burial of organisms
and alteration of substrate. Typically associated with agriculture, disturbance of stream flow regimes, urban
runoff, ineffective wastewater treatment plants, deposition of dredged or other fill material, and erosion
from mining and construction sites.
Turbidity/Shade. Reductions in solar penetration of waters as a result of blockage by suspended sediments
and/or overstory vegetation or other physical obstructions. Typically associated with agriculture, disturbance
of stream flow regimes, urban runoff, ineffective wastewater treatment plants, and erosion from mining and
construction sites, as well as from natural succession, placement of bridges and other structures, and
^suspension by fish (e.g., common carp) and wind.
Vegetation Removal. Defoliation and possibly reduction of vegetation through physical removal, with
concomitant increases in solar radiation. Typically associated with aquatic weed control, agricultural and
sOvjcultural activities, channelization, bank stabilization, urban development, defoliation from airborne
contaminants and other stressors included in this report, grazing/herbivory (e.g., from muskrat, grass carp,
geese, crayfish, insects), disease, and fire.
Thermal Alteration. Long-term changes (especially increases) in temperature of water or sediment
Typically associated with power plants, other industrial facilities, and global climate warming.
Dehydration. Reductions in wetland water levels and/or increased frequency, duration, or extent of
desiccation of-wetland sediments. Typically associated with ditching, channelization of nearby streams,
invasion of wetlands by highly transpirative plant species, outlet widening, subsurface drainage, global climate
change, and ground or surface water withdrawals for agricultural, industrial, or residential use.
Inundation. Increases in wetland water levels and/or increase in the frequency, duration, or extent of
saturation of wetland sediments. Typically associated with impoundment (e.g., for cranberry or rice
cultivation, flood control, water supply, waterfowl management) or changes in watershed land use that result
in more runoff being provided to wetlands.
Fragmentation of Habitat Increases in the distance between, and reduction in sizes of, patches of suitable
habitat
Other Human Presence. Increases in noise, predation from pets, disturbance from visitation, invasion by
aggressive species capable of outcompeting species that normally characterize intact communities;
electromagnetic, ultraviolet (UV-B), and other radiation, and other factors not addressed above.
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standardization of wetland monitoring techniques may not be desirable or feasible given the current lack of
comparative studies. Exceptions may include situations where litigation is probable or efficacy of a
regulatory program must be determined. Also, the need for diverse and adaptive sampling stategies is
suggested by the extreme temporal and spatial variability within and among wetlands, and the variety of
purposes for which wetlands are monitored.
The Sampling Protocols subsection also notes where data are indicate that one protocol, type of sampler,
or metric is better than another. However, we have not evaluated these ourselves, except to note situations
where the use of a particular sampler, protocol, or community metric seems clearly inappropriate.
Finally, the subsection discussions of Spatial and Temporal Variability-Data Gaps summarize numerical
data, both temporal and spatial, on wetland community ecology. The range in values of, say,
macroinvertebrate density, is noted for wetland types for which such data are available.
Data on variability is potentially useful for helping develop wetland biocriteria. For example, taxa whose
community structure naturally varies the least with time and space tend to be most practical for use as
indicators of anthropogenic influences. Also, the spatial variability in community composition among
wetlands may be less in disturbed landscapes than in natural landscapes, if inferences from other ecosystem
types are applicable (Sheehan 1984). Such information is useful in design of regional monitoring programs.
If data that describe variability were drawn from a sufficient number of wetlands to represent the wetland
resource of a region, and with a sufficient frequency to capture the range of changing conditions, then such
data might be used as one basis for establishing numeric criteria for protection of wetland aquatic life. They
might also be used to target gaps and reduce costs in the statistical design of more rigorous biomonitoring
efforts. Such an approach has been proposed for use in EPA's new Environmental Monitoring and
Assessment Program (EMAP), and has been applied successfully to stream ecosystems in Ohio and Arkansas
(e.g., Giese et al. 1987).
However, existing data, such as those presented, are of uncertain statistical representativeness. They were
compiled from all relevant, published studies. As such, they may represent only a first-guess or "default"
estimate of expected or baseline levels of community-level metrics, relevant only when local data are lacking.
As noted earlier, conclusions drawn from these data cannot be extrapolated to other wetlands with known
certainty.
Areas of missing biological information ("data gaps") are also noted. As appropriate, gaps are identified by
geographic region, by wetland type, and by type of stressor. Emphasis is on geographic gaps, rather than
on thematic gaps (thematic gaps have been identified in Adamus 1989, USEPA 1988, and in many other
documents). Information on gaps was gained partly by plotting all relevant studies on state maps (Appendix
B).
1.2 HOW THE REPORT WAS PREPARED
In August 1988, EPA's Wetlands Research Program sponsored a workshop in Easton, Maryland, a part of
which focused on identifying organisms and metrics that might be useful for indicating wetland ecological
condition. Findings were summarized in an EPA report , "Wetlands and Water Quality: EPA's Research
and Monitoring Implementation Plan for the Years 1989 - 1994" (Adamus 1989). That report noted a need
for synthesizing existing regional literature in ways that would allow candidate bioindicators to be identified
and available data to be numerically compiled. Potential categories of indicators applicable to surface waters
(in general) were targeted in EPA contracted reports (AMS 1987, Mittleman et al. 1987,) and by another
EPA workshop held in early 1989 (Temple, Barker, & Sloane 1989).
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A preliminary synthesis of wetland indicator literature was completed in August 1989 (Brown et al,
unpublished) as part of EPA's planning efforts for the new Environmental Monitoring and Assessment
Program (EMAP). That effort included a review of abiotic as well as biotic indicators, but did not attempt
a comprehensive review of technical literature. At the same time, EPA's Office of Policy, Planning, and
Evaluation (OPPE), acting on a request from EPA's Office of Wetlands Protection, asked EPA's Corvallis
Environmental Research Laboratory to modify and expand the scope of the similar, unpublished EMAP
report. Representatives of EPA's Office of Water Regulations and Standards (OWRS) were also involved
in early discussions of the scope of the effort. It was agreed that the modified report would focus more
strongly than did the EMAP effort on compiling quantitative measurements of wetland ecological
communities. In particular, it would attempt to describe the variability in community responses by region,
stressor type, and taxon. Additional support from EMAP would complement OPPE's support. This report
represents that effort.
Literature review began with an automated bibliographic search of the Wetland Values Database of the U.S.
Fish and Wildlife (Ruta Stuber 1986). Other bibliographic databases were also searched using terms wholly
or partly synonymous with wetlands, e.g.:
alluvial; aquatic moss; aquic; aquod; backwater; bayou; benthic/aquatic/submersed/submerged
plant/vegetation; black(-)water; bog; bosque; brown(-)water; depression; ditch; dystroph-; fen;
floodplain; fluventic; fluvisol; histic; histosol; hydrophy-; intermit- stream; inundated soil;
lagoon; lentic; littoral; lowland; macrophy-; marsh; mire; muck; muskeg; oxbow; playa;
pluvial; pocosin; pond; poorly drained; pothole; riparian; saprophilic; seep; shallow lake;
shoal; sphag-; stockpond; stream corridor; swamp; vernal pool; wash; water log-; wet land;
wet meadow; wet prairie
Literature was included if it met the following criteria:
o quantitative biological measurements were described (i.e., not just species lists or faunistic surveys);
o inland nontidal wet areas were covered;
o oriented towards the community level of ecological structure (i.e., transects or point data in which
a full range of vascular plant, fish, bird, amphibian, or mammal species was measured, not just single
species);
o if not community-oriented, then focused on the sensitivity of ecosystem process (e.g., productivity,
decomposition) to environmental stressors, or on the relative usefulness of particular species as
"bioindicators."
References resulting from the preliminary literature review were compiled by state and circulated, with the
criteria, for comment to persons from the following groups:
o wetland coordinators from the EPA Regions and wetland biologists from other EPA Labs
o selected offices of the Corps of Engineers in each region
o a majority of members of the Society of Wetland Scientists
o wetland coordinators for all state highway departments
o state biologists of the Soil Conservation Service
o refuge managers of the National Wildlife Refuges
o attendees from the Easton workshop
o other persons selected from Wetland Research Program mailing lists
In addition to soliciting comments on published literature, we asked these persons to suggest data meeting
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our criteria that could be found in the following types of less-available literature:
o student theses
o biomonitored mitigation sites
o impact statements or permit applications
o Superfund site assessments
o water quality bioassessment reports
o utility siting plans
o fish/wildlife agency studies
o forest management monitoring plans
o grazing management monitoring plans
o aquatic weed control impact studies
A large number of responses were received, and along with secondary citations discovered in literature we
collected, resulted in significant expansion of our bibliography. Some unpublished and ongoing data sets
recommended by respondees were included as well. Despite the considerable effort, some experts were
undoubtedly not contacted and it is likely that some number of references meeting our criteria were not
discovered.
Subsequently, all literature contained in the bibliography but not presently in the EPA - Corvallis wetlands
library was obtained. Study locations were plotted on state maps (Appendix B) using a geographic
information system at the Lab (Arclnfo GIS), quantitative data were extracted and compiled for chapter
tables, the "Methods" sections of papers were reviewed, and the narrative descriptions presented in the
following chapters were prepared. Quality of individual data sets or their locations on the maps could not
be checked or assured.
In addition, various national databases exist that frequently contain wetland community data. Data from
sites associated with these databases were obtained and/or the site locations were plotted on the digital
maps. These include:
o LTER network (all areas plotted; Long Term Environmental Research sites sponsored by the
National Science Foundation);
o Christmas Bird Count database (all areas plotted); from Cornell Laboratory of Ornithology;
o Breeding Bird Survey database (all areas plotted); from U.S. Fish and Wildlife Service;
o Breeding Bird Census database (only wetland areas plotted); from Cornell Laboratory of
Ornithology;
o Waterfowl Surveys (Migrating, Wintering, Spring Waterfowl Surveys, Summer Brood Count/
Breeding Ground Surveys); from Waterfowl Flyway Technical Representatives in each state;
o International Shorebird Survey (all inland wetlands); from Manomet Bird Observatory, Manomet,
Massachusetts.
In addition, several data sets exist that may include relevant wetlands biological data, but with the limited
effort of this project, such data could not be easily compiled or separated from non-wetland data. Examples
of these include:
o wetland boundary determinations by consultants and agencies (a vast source of botanical data);
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o measured data collected in support of HEP analyses by numerous consultants and agencies;
o river basin reports of government water quality monitoring programs (a source of fish and
invertebrate data, if "wetland" stations could be separated from others);
o monthly bird counts of the National Wildlife Refuges;
o data from the Nest Card and Colonial Waterbird databases of the Cornel] Laboratory of
Ornithology;
o private notes of birders, botanists, and other naturalists.
Wetland data not included because of their failure to meet one or more of our criteria included the
following:
o National Contaminant Biomonitoring Program data of the U.S. Fish and Wildlife Service (focuses
on bioaccumulation and generally does not include measurement of community-level variables);
o Inventories of wetland threatened/ endangered species (not measurement of community-level
variables);
o Inventories of wetland acreage and distribution (not measurement of community-level variables).
o Databases of The Nature Conservancy and state heritage programs (field data often not quantitative)
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2.0 APPROACHES FOR PROTECTING WETLAND QUALITY
2.1 REGULATORY BACKGROUND
Statutes to reduce the impacts of the disposal of dredged and fill material in wetlands (e.g., Section 404 of
the Clean Water Act) do not directly address impacts to wetlands from drainage, vegetation removal, and
nonpoint-source discharges. However, other provisions of the Clean Water Act, if applied more vigorously
to wetlands, have the potential to significantly reduce these impacts (USEPA 19895). Moreover, interest
in restoring degraded lands, including wetlands, appears to be growing, and because not all degraded areas
can be restored immediately, priorities based partly on the existing degree of degradation must be developed.
Questions arise, then, as to how best to measure, protect, and restore the quality of wetlands.
Many activities and discharges of pollutants into lakes and streams are regulated by State and Federal
agencies. For example, under the State water quality certification authority of Section 401 of the Clean
Water Act, States may grant, deny, or condition Federal permits or licenses that authorize a wetland
alteration within that state. States are also mandated to develop and adopt water quality standards, as
provided in Section 303 of the Clean Water Act, and all have done so. These standards must be applied
to all waters of a State. The standards are the basis upon which States review permits to determine whether
a proposed activity will meet a "use" that has been designated by the State for a particular water. Federal
agencies reviewing applications for wetland alteration must comply with State decisions rendered under
Section 401.
EPA, in its Water Quality Standards Program, requires State programs to include five components, two of
which are the focus of this chapter:
o Designating Uses
o Applying Water Quality Criteria
The following sections of this report define these and" describe optional approaches for States to consider
as they address the future application of water quality standards to wetlands.
2.2 USE DESIGNATION AND CLASSIFICATION
"Designated uses" are uses or goals—such as public water supply, propagation of fish and wildlife, and
recreation—that may be specified for each water body or wetland, whether or not they are currently being
attained. Because of the high biological productivity of many wetlands, fish and wildlife uses are often
emphasized, and the designated use category of "fishable/swimmable" that already covers other surface waters
in most State programs can, as a first step, be administratively extended to cover wetlands. Use-designations
may reflect either an acceptable current use of a wetland, or particularly in the case of restoration programs,
a desired or attainable future use. They may be described in either general (e.g., "wetlands in Basin A
should sustain commercial fishery production") or specific terms (e.g., "wetlands in Basin A should support
a Fish Index of at least 3.5"). Multiple uses may occur or be designated within a single wetland or wetland
type, and criteria appropriate for protecting one use may differ from criteria appropriate for protecting
another.
Wetlands also typically have "uses" not commonly designated in State programs concerned with other surface
waters. Examples include floodwater storage, groundwater recharge, and shoreline stabilization. These uses
(commonly termed "functions" by wetland managers) have been recognized, for example, in Section 404(b)(l)
Guidelines of the Clean Water Act. Currently, few State water quality or wetland management programs
have promulgated explicit procedures or standards for designating and protecting these uses.
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Specifying a goal or "use" is not the only way for a State to protect the quality of its wetlands. Under the
Antidegradation provisions of the Clean Water Act (40 CFR 131.12(a)(3)), States can simply declare all
wetlands, all wetlands of a certain type, or particular wetlands to be "outstanding national resource waters."
If the States' antidegradation policies are at least as protective as EPA's, this generally affords these wetlands
the highest degree of protection because no degradation is allowed, except for "short term changes that have
no long-term consequences" (USEPA 1989b). Other approaches, both regulatory and non-regulatory, might
also be used for protecting wetland quality, including (but certainly not limited to) EPA's Advance
Identification initiatives, nonpoint source management plans, water allocation negotiations, State Wetland
Conservation Plans, State Conservation of Outdoor Recreation Plans (SCORP's), emission control programs,
and others.
An obvious first step for any approach to regulating wetland water quality is to determine the general
distribution and location of wetlands. The most comprehensive source of such information is the series of
quadrangle-based wetland maps available for most of the United States from the U.S. Fish and Wildlife
Service (obtained by phoning 1-800-USA-MAPS). These maps classify wetlands into a number of categories
not necessarily related to their functions or uses. Although field-checking is often required to determine
if wetlands on these maps (and possibly some that are not) are subject to regulation under Section 404, such
intensive verification is not required for States to include wetlands under their definition of State waters.
As noted previously, existing use designations for State waters may be extended to wetlands. However, the
inherent diversity of wetlands is best protected by developing different sub-categories of a use to different
wetland types, at which point wetlands would be classified further. Designated uses, and criteria for
protecting these uses, can be assigned as described below to each wetland, to landscape units of wetlands,
or to each wetland type. To achieve this, two major options are presented: (a) Strategic Setting, and (b)
Probable Functions.
The Strategic Setting option involves assigning designated uses to wetlands based on their landscape position,
relative to connected waters or adjoining lands which potentially benefit from uses or services the wetlands
provide. A rudimentary application of this approach would involve assigning to wetlands the same "use"
currently designated for all waters into which they flow. Some technical consideration should be given
regarding the nature of the hydrologic connection that exists between the wetlands and the receiving waters;
even wetlands that lack surface water outlets •are sometimes intimately connected to other waters via
subsurface flow. A distance criterion may also be appropriate, because beyond some distance, the
contribution of wetlands to certain functions of receiving water uses may, even on a cumulative basis, be
indetectable. To reduce subjectivity, some simple models (e.g., Phillips 1989) might be used to assign
technically-derived distance coefficients for use in different river basin types.
A conceptually similar but somewhat more descriptive option for assigning designated uses to wetlands can
be used to supplement the above, and to prioritize wetlands for more detailed scrutiny in the use-
designation process. Under this option, the designation of uses for a particular wetland is based on the
presence of nearby cultural features that would be expected to benefit from functions typically performed
by most wetlands. For example, a wetland might be assigned a use-designation of "flow regime maintenance"
if it is located a reasonable distance upstream from a floodplain area that contains many dwellings
susceptible to costly flooding. Other cultural features that may benefit from typical functions of upstream
wetlands might include the following (these are only a few examples):
o sole source aquifers
o waters with known fish kill or eutrophication problems
o other waters believed to be in violation of water quality criteria
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o waters in periodic need of dredging
o areas intensively managed or protected for their ecological resources
o floodplains containing economically significant development
In operation, this approach would begin by reviewing and, if appropriate, expanding upon the above list.
Such a list might then be circulated for public input and perhaps narrowed to include just those for which
geographic data are available. Then, these cultural features are mapped and their drainage areas or other
functionally connected areas are delineated. Finally, a distance criterion is specified, wherein all wetlands
containing, located upstream from, or otherwise functionally influenced by the listed feature (and within the
specified distance) are considered to be strategically situated. That is, they are positioned so as to
individually or collectively deliver or support the particular designated use. If no cultural features of
concern are positioned within an appropriate distance from a wetland, the wetland may be assigned a general
designated use that has been assumed for all wetlands statewide (perhaps "aquatic life*) unless a use
attainability analysis provides evidence to the contrary.
Once the existing geographic data have been assembled, the Strategic Setting option can be applied rapidly
to large areas (river basins or entire states), because it mainly designates uses according to watersheds or
drainage areas, rather than requiring wetlands to be evaluated individually. However, it makes no evaluation
as to whether uses that are reputed to be provided by wetlands generally are actually provided by a
particular wetland; it only evaluates whether a wetland is positioned so that such uses, if performed, will
have an important recipient or user.
The approach does not require that all wetlands be mapped. Once the strategic watersheds have been
identified, the responsibility for determining the locations of wetlands that are affected might be assigned
to permit applicants. Managers must also keep in mind that downstream cultural features, and thus the
strategic status of a wetland, can change with time. Accordingly, an overall designation of attainable uses
should always be provided.
A second option, the Probable Function option, generally involves designating a use or uses to individual
wetlands, wetland landscapes, or wetland types based on (a) direct measurement of the use, i.e.-, wetland
function, or, (b) structural indicators of the use. Ideally, the uses or functions are measured directly in each
wetland and their verification becomes the basis for establishing that designated use in each wetland.
However, such individual verification of uses is seldom feasible without significant time and funds. Even
then, uses verified to exist in wetlands under one set of annual climatic conditions may not exist in
subsequent years under different climatic conditions.
An alternative is to employ structural indicators of wetland function-such as degree of channel meandering,
watershed position, and connectedness-to modify more general descriptors provided by existing classification
schemes such as the Cowardin et al. (1979) classification scheme. The literature on such structural indicators
of wetland function has been summarized by Adamus and Stockwell (1983) and Adamus et al. (1990).
Cursory evaluation of these structural indicators can be accomplished largely by reviewing airphotos and
topographic maps depicting the wetland and its associated landscape, with perhaps a single brief site visit.
At the simplest level, the classification scheme of the U.S. Fish and Wildlife Service (Cowardin et al. 1979)
might be used as the sole structural indicator of wetland function. Uses that are expected to be typically
attainable for each wetland map category (e.g., riverine emergent, palustrine emergent, palustrine forested)
are described, without regard for where on the landscape the wetland exists. Such an approach can be
implemented quickly due to the availability of wetland maps, but some uncertainty exists as to whether map-
based classification schemes designed for more general purposes are sufficiently sensitive to the wide
variability in degree of function among wetland types.
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2.3 DEVELOPING PRIORITIZED USE SUB-CATEGORIES
The description of the above two options (Strategic Setting and Probable Function) has focused on ways
these options might be employed for designating uses of wetlands. Additionally, in some cases they might
be used for (a) establishing sub-categories of use, and/or (b) establishing criteria that describe the conditions
necessary to protect particular uses. This section focuses on (a), and the following section (2.4) addresses
(b). Although most States establish standards for water bodies simply according to the uses they are
designated to support, other States have established a hierarchy of uses with the higher uses denoting higher
water quality. The former situation was described above, and the latter is described in this section.
Applying the Strategic Setting option to establish a sub-category of use might involve designating not only
a general use such as "drinking water" to wetlands upstream of a drinking water intake, but also assigning
a sub-category of use entitled "high quality for drinking water protection."
In another case, if the Probable Function option is used, wetlands which appear (for example) to support
aquatic life might be assigned to a sub-category of use entitled "high quality wetland for shorebirds" based
on direct functional measurement, Cowardin class, or structural indicators. Direct functional measurement
could involve regional biosurveys and use of community-level indices to establish sub-categories descriptive
of wetland quality, as some States have used for non-wetland surface waters. If, instead, structural indicators
were used as the basis for establishing higher sub-categories within the "aquatic life" use, this could involve
defining the "best" wetlands for this use in terms of their seasonal hydrology, vegetation, soils, landscape
position, and other factors.
As part of EPA's "Advance Identification" program, some EPA regional offices, localities, and states are
identifying or rating functions and uses of hundreds of individual wetlands using structural indicators or,
rarely, direct functional measurement. Some States (e.g., Florida, Swihart et al. 1986) have similarly ranked
wetlands as part of efforts to (a) designate "Outstanding Natural Resource Waters" under state water quality
laws, or (b) designate wetlands of exceptional importance to waterfowl under State wetland
conservation/recreation plans and the North American Waterfowl Management Plan. In these cases, existing
evaluations might be used as one data source for considering appropriate sub-categories of designated use.
Regardless of whether use sub-categories are identified by direct measurement (e.g., biosurveys) or through
use of structural indicators, considerable effort and time is required. Three options for making the task
more tractable are available and are described as follows.
One option is to measure the functions (uses) in a limited set of "reference wetlands," which might be either
a randomly selected set of regional wetlands, or a set of wetlands selected because they are believed to
represent the least disturbed conditions. Once the reference wetlands are chosen, measurements of their
structure and/or function (e.g., diversity, biogeochemical cycling rates, hydrologic transfer rates, community
composition) might be used for defining the highest sub-category of protection. Such regional efforts would
involve four steps:
1. "Reference wetlands" are chosen.
2. Functional (use) data are collected from these and compiled.
3. Spatial and temporal variability in the data, within and among reference wetlands, is compiled.
4. Use-subcategories are developed or modified.
If reference wetlands are selected using the random approach (e.g., Abbruzzese et al. 1988), measurements
of these may not reflect "best attainable" levels of function, because in some regions, a majority of wetlands
may not be functioning at desired levels, due to landscape-scale impacts. Conversely, if the "least disturbed"
selection approach is used, some of the selected wetlands may not be providing some functions at levels
desired by some segments of the public; for example, greater benefits to some components of aquatic life
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may be provided by wetlands that are actively managed rather than undisturbed. Agencies might desire that
certain species or processes be the focus of protection in a particular wetland or watershed, and these might
be dependent on continuation of existing management practices. Thus, if definition of "best attainable"
functional condition is the desired objective, data from reference wetlands selected by either approach would
serve only as a starting point. The reference wetland data would eventually need to be evaluated from the
perspective of resource goals (as described above under Use Designation). Attainable condition could also
be evaluated using data from from similar wetlands in less-disturbed adjoining regions, and/or from analysis
of pre-settlement conditions.
To begin the selection of reference wetlands using the "least-disturbed" approach, attention would be given
to identifying wetlands having superficial characteristics such as the following:
o wetland arose naturally and at a considerable time in the past, rather than being recently
constructed;
o surrounding watershed, particularly within 500 feet of the wetland transition with upland, is largely
undeveloped;
o water levels fluctuate naturally, not being affected by diversions, dams, or nearby wells;
o wetland has not been recently used for silviculture, grazing, or other human uses that potentially
impact vegetation and/or water quality and quantity.
Frequently, when defining the "least disturbed" condition (either for an individual wetland or the regional
wetland resource), the objective is to maintain the use within an "envelope" of expected temporal variability
at a site or within a region. Although quantifying this can be a challenge, in some wetland types, recent
developments in methods such as seed bank analysis (e.g., Poiani and Johnson 1989), tree ring analysis
(Bowers et al. 1985, Hupp and Morris 1990, Sigafoos 1964), as well as sediment core analysis of pollen
(palynological analysis)(Agbeti and Dickman 1989, Battarbee and Charles 1987) and sediment deposition
(with measurement of lead 210 and/or cesium 137, see Bloesch and Evans 1982, Ritchie and McHenry 1985)
can be used to identify the extent of hydrologic and botanical variability that has existed in both recent and
distant historic times.
To select reference wetlands using the "random selection" approach, a statistical sample of all wetlands in
each general category (e.g., riverine emergent, palustrine emergent, palustrine forested) is visited and
probable functions of each wetland are assessed using structural indicators (as organized in any of several
rapid methods for wetland evaluation, see Kusler and Riexinger (1986) for examples) or direct measurement
of function (e.g., biological surveys). As enough wetlands are sampled to overcome variability within a broad
class, generalities about the class in a particular region may begin to emerge. This approach was
demonstrated by Schiefele and Mulamoottil (1988), who used structural indicators, and by Ohio EPA (1987),
who measured functions (aquatic life values) directly, in non-wetland surface waters. These characterizations
of the functions (uses) of wetlands of a general type then could be used for assigning distinctive levels or
types of protection to specific wetland types. This approach also has the advantage that potential biases in
selecting "least-disturbed" wetlands are avoided. By incorporating statistically representative data that will
be collected beginning in the mid-1990's by EPA's Environmental Monitoring and Assessment Program
(EMAP), the attractiveness of this approach may grow in future years.
Experiences of the EPA Wetlands Research Team suggest that, if a subset of a population of wetlands is
to be visited and evaluated, the subset should contain 10 to 20 times as many wetlands than will actually
be sampled, because denial of access is a common problem. Finding this amount of suitable wetlands when
wetlands normally comprise less than 10 percent of the landscape can be a daunting task. In states with
digitized (GIS-based) wetland maps, however, it can be much simpler.
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If a decision is made to stratify the set of sampled wetlands, the degree of detail associated with the
description may be important. For example, a stratification based only on broad general map categories
such as "palustrine" and "lacustrine" wetlands would result in much higher variability than one in which the
definition of "types" is based on structural features that are expected to best relate to the intended use or
function. More samples of reference wetlands would be required to adequately define "typical" uses and
functions. If, instead, types of wetlands were defined by a larger set of structural indicators (e.g., landscape
position, channel meandering, soil type), variability among wetlands within types would be less and less
monitoring would be required to characterize the typical uses of each type. For some purposes, the host
of diverse structural indicators could be synthesized into fewer "types" by recognizing categories defined by
expected biogeochemical forcing functions, e.g., wetlands dominated by flowing water vs. wetlands dominated
by wave energy vs. wetlands dominated by ground water influx. However, without costly field-checking,
wetlands cannot be reliably classified across entire regions according to such a scheme. Its application poses
several operational problems with definitions as well.
A third option for making the Probable Function component more tractable is to focus evaluation at a
landscape level, assessing function directly or measuring its structural indicators at a coarse scale rather than
wetland-by-wetland. Regional information on structural indicators of wetland function used in such an
approach might include soil types, runoff, and landform, as depicted on existing maps and geographic
databases which summarize these. Although such data are seldom collected and compiled at a similar scale
and level of resolution, for planning-level estimates they might be combined to yield qualitative estimates
of the cumulative contribution of wetlands to landscape function in a region (Abbruzzese et al. In Press).
Once the relevent data layers have been identified, acquired, and assembled, the categorization of similar
landscape units and designation of their uses may proceed quite rapidly.
Regardless of which option is used for reducing the data collection effort of the Probable Function
component, data interpretation will remain an important concern. Specifically, considerable subjectivity may
surround decisions as to what numeric thresholds in the data should define particular sub-categories of use.
Accordingly, the definition of use sub-categories should involve both public and technical inputs. Systematic
procedures are available for helping reduce complexity and subjectivity in data interpretation (e.g., Krebs
1989). These include statistical approaches, e.g., systematically clustering data in groups, identifying quartiles
or distributional nodes in the data, "break points" in cumulative frequency curves, and similar procedures.
The eventual result is a sub-categorization of uses according to the degree to which they should be satisfied.
For example, a State might wish to define "Class A" wetlands as:
o All bogs that contain greater than 80 % of the bog-dependent amphibians found in bogs of the
same size in the region, OR,
o All wetlands that contain greater than 4 vegetation strata," OR,
o All herbaceous floodplains with a net annual productivity of greater than 2000 grams of
carbon per square meter per year."
For a less pristine (e.g., "Class B") sub-category, the above ligures might be relaxed to 70%, 3, and 1000 per
square meter, respectively. Again, these specific levels are only illustrative, and would need to be derived
from a biosurvey of wetlands in each region.
For greatest replicability, sub-categories of use describe the desired or actual condition of biota that can be
directly measured, rather than only describing the sub-category as "degraded", "pristine", and similar
qualitative terms. Biological descriptions contained in the descriptions are related directly to management
goals. Descriptions of use sub-categories might include lists or ratios of organisms that characteristically
dominate altered and unaltered wetlands, or organisms physiologically tolerant or intolerant of a particular
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type of stressor, if such information is regionally available for the particular wetland type. Descriptions of
use sub-categories that include multiple trophic levels (e.g., algae, fish, birds) are more difficult to develop
but may improve reliability and provide flexibility in applying assessment techniques. Descriptions of use
sub-categories should also footnote particular assessment protocols (perhaps including methods for
determining the requisite number and distribution of samples) used to develop the criteria and to be used
to document compliance or non-compliance or to refine use sub-category descriptions.
2.4 NARRATIVE CRITERIA TO PROTECT WETLAND DESIGNATED USES
Once uses are designated, narrative or numeric criteria must be established to protect each use. EPA has
indicated that, by September 30, 1993, States and qualified Indian Tribes shall apply standards to wetlands
that incorporate, among others, designated uses, aesthetic narrative criteria (e.g., "free from..."), and narrative
biological criteria, as well as appropriate numeric criteria. As States desired to become more protective of
wetlands, these requirements are to be based on existing information. However, new information will be
needed to eventually refine the standards.
At the simplest level, a State might define its narrative criteria is to state that no activity be permitted that
results in a net loss of wetland acreage. This is because wetland extent is the most fundamental measure
of wetland function. Wetland extent is best evaluated on a landscape or regional scale, as it is at this level
that wetlands may provide the most significant benefits to aquatic life and wildlife. For example, the
cumulative acreage of wetlands in a region, and the specific combinations and juxtapositioning of wetland
types, can mean more to highly mobile waterfowl than does the contamination status of a particular wetland.
This is because the daily and annual movements of many animals encompass several wetlands.
If wetland extent is to be used to define desirable sub-categories of use, it may be necessary to initially
define "reference conditions" at a landscape level. For wetland extent, this may mean determining, from
existing wetland maps and airphotos, mean densities of wetlands (acres per square mile), perhaps of various
types and in various landscape contexts within a region. This metric could be determined just for "least-
disturbed" landscapes, or for all landscape units in a region. Alternatively, an appropriate series of archival
airphotos could be interpreted to yield information on historical wetland density (e.g., acreage of wetlands
per square mile).
At a somewhat more detailed level, narrative criteria could specify that wetlands shall not be changed from
one Cowardin type to another. This would be based on an assumption that maintaining a particular
Cowardin type maintains the designated uses. However, wetland science indicates that considerable
degradation of a wetland's functions (uses) can occur without being manifested as a change in Cowardin
type.
In contrast, if structural indicators are used (as described above in descriptions of the Probable Function
option), uses might be somewhat better-protected. For example, narrative criteria might specify that no
activity be permitted that decreases a wetland's probability rating for "Aquatic Diversity," as indicated by an
accepted, structure-based wetland evaluation procedure. However, wetland science also indicates that
structural indicators of wetland function are not always reliable; wetland function can be considerably
degraded without obvious signs of structural change. Also, wetland evaluation methods have not been
designed to distinguish which of the structural features they employ are determinants of wetland function
(and thus useful in criteria development) vs. indicators (mere correlates) of wetland function. In either case,
these structural factors rrfay not currently be subject to legal regulation and/or may not normally be altered
by development. Also, protection of the structural integrity of wetlands in one county or state may or may
not guarantee against degradation of the associated use that results from stresses that occur beyond
jurisdictional boundaries.
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2.5 NUMERIC CRITERIA TO PROTECT WETLAND DESIGNATED USES
Where the predominant stress to a wetland is one that is commonly regulated in other surface waters (i.e.,
Contaminant Toxicity, Enrichment/Eutrophication, Organic Loading/Dissolved Oxygen, Acidification,
Salinization, Turbidity, Thermal Alteration), the existing aquatic life criteria (USEPA 1986) that are
commonly applied may also be applied to wetlands. In addition, it can be assumed that all human health
criteria for other surface waters apply to wetlands. In the longer term, regionally-based numeric criteria
might be developed for wetlands by measuring functions/uses directly, as opposed to narrative criteria which
would rely mostly on Cowardin type or structural indicators.
In the case of "aquatic life" uses of wetlands, this could involve eventually developing biocriteria for
wetlands. If regional biosurveys are not immediately feasible as a basis for developing biocriteria, existing
field data sets interpreted with great caution might sometimes be used. For example, existing data may be
sufficient to indicate the approximate number and proportions of particular species expected to occur in a
particular wetland type, and this knowledge could be used to help establish biocriteria protective of that
general use. However, any such criteria derived from the literature must consider the likely non-
representativeness of the data and potential biases arising from species-area effects and variable levels of
effort.
One concern that arises is that existing narrative and numeric criteria are inadequate to address some
impacts that critically affect wetland function and use, such as hydrologic and physical alteration. Of the
stressors listed in Table 5, the physical impacts of Dehydration, Inundation, Vegetation Removal,
Sedimentation, Shade, and Habitat Fragmentation in particular are currently seldom addressed by water
quality protection programs. The impacts to wetland uses from these stressors are likely to often exceed
impairments to use resulting from chemical contamination.
Existing information is sufficient to develop narrative criteria to address these stressors. However, there are
probably insufficient data to promulgate numeric criteria at present, except perhaps site-specifically for a few
well-studied wetlands. Future, long-term development of numeric criteria for protecting wetland uses against
physical alteration may require new research protocols, such as expanded use of field mesocosm and whole-
wetland or whole-watershed manipulations. Laboratory testing-the typical approach used for contaminants-
-would be less appropriate due to the scales involved and the complexity of interactions. Also, if numeric
criteria were to be developed for physical stressors, the criteria would need to be keyed in to specific uses,
because (for example) the amount of sedimentation that is detrimental in a wetland to some uses might be
beneficial to others.
In the long term, however, some chemical criteria may need to be re-evaluated site-specifically because of
the unusual conditions encountered in some wetland types. Undisturbed wetlands sometimes have lower pH
and dissolved oxygen; higher organic carbon, humic acids, temperature, ammonia, and sulfide; extreme
reducing conditions; more potential for photodegradation, biodegradation, chelation and organic
complexation than do surface waters generally and laboratory waters specifically. Under certain
circumstances and for some contaminants, these conditions can profoundly affect the rate and direction of
contaminant mobility in wetlands, as well as the bioavailability and toxicity of these contaminants (e.g.,
Winner 1984). Moreover, the spatial and temporal variability of these conditions is believed to be much
greater in wetlands than in non-wetland waters, due to their shallow depths, prevalence of vegetation, and
closer dependence on hydrologic forces.
In applying existing numeric criteria to wetlands, careful consideration is most appropriate when (a) the
laboratory water used to develop existing criteria differs significantly from extremes found in a particular
wetland type, and/or (b) the types of organisms in the region differ significantly in their physiologic
responses or propensity for bioaccumulation from those used in testing. Thus, in cases where verification
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of the applicability of existing numeric criteria to a particular wetland or wetland type is essential, a four-
phase procedure might be used, involving (a) review of data describing laboratory water chemistry and
indicator species used previously to develop the criteria, (b) analysis of water samples or sediments from the
subject wetland or wetland type to determine if they are significantly dissimilar from laboratory water used
in testing, (c) biological survey of species inhabiting the subject wetland or wetland type to determine if they
are significantly dissimilar (in terms of physiologic responses or bioaccumulation potential) from species used
in the laboratory bioassays, and (d) rerunning the toxicity tests, if necessary as indicated from the results of
(a)-(c).
If re-running of toxicity tests is desired and feasible, ideally, a tiered testing scheme is used (Kimerle et al.
1978, Ongley et al. 1988). Traditionally, risk is assessed and numeric criteria are developed by incorporating
a hierarchy of toxicity tests of increasing complexity, chemical data, and expected exposure regimes (La Point
and Perry 1989). This involves use of a combination of single-species laboratory assessments of acute
toxicity, field microcosms, field mesocosms, and modeling of toxicity, transport, and fate based on chemical
structure and other factors (Matthews et al. 1982). This could be done using bioassays featuring indigenous
wetland organisms (e.g., Fremling and Mauk 1980, Lee et al. 1987), and/or by manipulating experimentally-
confined wetlands to determine biotic responses (e.g., Carpenter and Chancy 1983, Hurlburt et al. 1972,
Richardson et al. 1983). EPA protocols specify that establishment of an acute value for a freshwater
criterion be based on a minimum of eight different taxonomic families, including a freshwater alga or
vascular plant, a planktonic crustacean, a benthic crustacean, an insect, a nonarthropod nonchordate, another
insect or a new phylum, a salmonid, another fish, and another chordate. Partly because such testing has been
carried out chemical-by-chemical, the development of criteria has typically been a lengthy process and efforts
have only recently begun to better define chemical interactions, through use of whole-effluent toxicity testing
and other approaches. However, the criteria which result are regarded as simple to apply and interpret, thus
allowing regulation of an effluent to be undertaken incrementally through licences and permits.
Consideration of the need for re-testing might focus initially on substances whose criteria were based on
testing of the fewest species, because the probability is less that these would include a sufficient number of
wetland species to fulfill the EPA requirement for bioassay of at least eight taxonomic families.
Given the large number of wetlands potentially exposed to contaminants, the costs associated with such site-
specific testing might be justified only where existing biochemical data had indicated that a particular
wetland was significantly different from biochemical test conditions. Because there is seldom enough
available data on background biochemical conditions of large numbers of wetlands to indicate that they
differ significantly from test conditions, it may be necessary to either (a) use exposure indicators (e.g.,
proximity to hazardous waste sites) or administrative needs (e.g., permit applications) to select wetlands
for site-specific testing or modification of criteria, or (b) statistically select a representative series of
wetlands, stratified by their probable, naturally-occurring biochemical type, that are suspected to deviate the
most from test conditions, and then confirm their biochemical categorization with field measurements and
re-test their biota. Any resulting modifications to existing criteria would be applied to the entire regional
population of wetlands of that biochemical type.
Future applications of numeric criteria to wetlands could include performance standards, impact standards,
or both (Courtemanch et al. 1989). Performance standards are characterized by a focus on each pollutant,
and are commonly expressed as "end-of-the-pipe" or "receiving water" desired concentrations or loadings.
These are often specified in terms of allowable magnitude, duration of exposure, and frequency, and are
designed to protect aquatic life both from risks due to bioaccumulation and from acute and chronic toxicity.
In contrast, impact standards require that a certain result be achieved. They are typically specified in terms
of biocriteria for ambient waters or sediments, e.g., desired species composition and richness, as described
in the earlier.
As States begin to protect and restore wetlands through a biocriteria approach, a question arises as to which
features, processes, or organisms best indicate the ecological "health" of the wetland resource, or are
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desirable due to convenience of monitoring or other reasons. Because EPA's national goal for wetland
protection is "no net loss in acreage or function." it may be desirable to additionally examine the community
structure and processes within wetlands, to establish criteria for biological function and to monitor
attainment of the functional quality goal. This can be done using the approach described above, i.e.,
identifying reference conditions, compiling data, analyzing variability, and ultimately establishing use-
designation criteria or setting restoration priorities-cither through field surveys or professional consensus.
However, in doing so, one faces the questions:
o "What are the best indicators of wetland biological function?"
o "How to monitor wetland biological function?"
These are the subjects of the next chapter.
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3.0 GENERAL GUIDELINES FOR WETLANDS BIOLOGICAL CHARACTERIZATION
Wetlands pose unusual challenges for monitoring programs. Because wetlands, as transitional environments,
are located between uplands and deepwater areas, their biota exhibits extreme spatial variability, triggered
by very slight changes in elevation. Temporal variability is also great, because the shallowness of any
surface water results in its being highly influenced by slight, fleeting changes in precipitation, evaporation,
or infiltration. Only a minority of all wetlands in the United States have permanent surface water (Shaw
and Fredine 1956), so sampling techniques developed for other surface waters are not always applicable.
The extreme spatial and temporal variability often requires that large numbers of samples be collected if the
wetland community is to be properly characterized.
Such extensive sampling is made difficult, however, by potentially severe problems of access. Physically,
access to many wetlands is hindered by water too shallow for rapid boat access, soil too fluid for rapid foot
or vehicular access, and vegetation canopies too dense for easy aerial or airboat access. Access to many
wetlands is also seriously hindered by the widespread (and sometimes misguided) public perception that
wetlands, in contrast to other waters regulated by the Clean Water Act, are exclusively private land.
Landowner awareness of the potential for regulation has led to commonplace denial of requests for access
to wetlands during other EPA projects. Proportionally few wetlands are publicly owned, and these are not
necessarily representative of the total wetlands population. These factors all combine to potentially increase
the costs of an effective wetland monitoring program, and pose significant demands for study design and
logistical planning.
Despite these difficulties, the need for more vigorous wetland sampling efforts is compelling. Because most
wetlands are located in a topographically low, depositional environment and have long hydraulic detention
times, they accumulate contaminants from a wide area. At the same time, undisturbed wetlands are
characterized by exceptional biological productivity, suggesting a greater need for more extensive monitoring
of wetlands. However, wetlands seldom are monitored, so much remains to be learned about the extent to
which contamination and other stressors have altered their condition.
3.1 WHAT TO MONITOR
Monitoring of multiple indicators-having both short and long lifespans, and both localized and broad home
ranges-is preferable to monitoring a few because indicators differ in their sensitivity to different types of
stress in different types of wetlands, and in their temporal and spatial occurrence. By monitoring both
short- and long-lived taxa, the effects both of stressors that occur briefly (e.g., herbicides) and of those that
occur over longer time periods (e.g., bioaccumulation of metals) can be detected. By monitoring both
resident and wide-ranging/ migrant species, the cumulative landscape-level impacts that may not be
detectable on a local scale may become apparent. Ideally, monitoring of a wetland should encompass as
long a time period, as many indicators, and as many microhabitats within the wetland as possible, given
available resources. However, the need to make choices is inevitable.
Another choice concerns the which level of ecological hierarchy should be measured-e.g., physiology of
individuals, demographics of a population, structure of a community, or processes of an entire ecosystem.
Conclusions from one level cannot necessarily be extrapolated to another. As noted in Chapter 1, the scope
of this report is limited primarily to the community level. A good discussion of factors affecting the choice
of an appropriate hierarchical level in wetlands is presented by Farmer and Adams (1989).
Sometimes, the analysis of initial data collections can be used to target particularly sensitive groups or
processes and identify optimal numbers of samples. Also, if life histories and ecological relationships are
sufficiently well-understood in a particular area, monitoring could be limited to a few taxa known for their
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sensitivity to a particular stressor or their role as ecological "keystones." Keystone species include those
which physically alter the landscape so profoundly that they create or destroy habitat for a much larger
group of species over a wide area.
Examples of taxa that are considered to be keystones in particular regions or wetland types include:
o woodpeckers, which excavate cavities required by dozens of species;
o bees and other pollinating or seed-dispersing organisms, which control habitat structure through
their major collective effects on vegetation;
o gopher tortoises and other burrowing species that create shelter critical to survival of many other
animals;
o beaver, which create wetlands and temporarily destroy forest;
o muskrats, alligators, and some herbivorous birds, which through grazing and physical movement
cause locally major increases in open water patchiness of wetlands.
Caution is necessary because it is seldom possible to validly infer trends in all species by monitoring only
one or a very few "keystone" or "indicator" species. Thus, changes in community-level metrics usually give
a clearer indication of "abnormal" biological stress than does the presence or absence of a single indicator
species, regardless of its reputation as a keystone (Browder 1988, Cairns 1974, Couch 1982, Grigal 1972,
Hellawell 1984, Karr 1987, Kelly and Harwell 1989, Landres et al. 1988).
In other aquatic systems, stable isotope techniques have been used to help identify keystone species,
ecosystem components, or processes. In the case of vascular plants, attempts to identify the most sensitive
species have also been made by measuring exposure of a host of species to a particular substance (e.g., a
nutrient) and then monitoring the varying degrees to which the substance accumulates in tissue (e.g.,
Canfield et al. 1983), or alters germination and other physiological processes. Species which accumulate the
substance and/or show the greatest physiological response the most are presumed to be likely to be affected
if the substance increases.
To identify the most sensitive indicators, greater efforts could be made to comb the literature on
experimental toxicology. However, although use of standardized conditions in most toxicity testing allows
some degree of comparison among taxa regarding their relative sensitivities, the usefulness of laboratory
toxicological data can be limited by the dissimilarity of test conditions and typical wetland conditions (e.g.,
altered toxicant mobility and toxicity due to increased organic carbon; interactions between hydroperiod
effects and chemical toxicity--see Chapter 2.0).
Conceptual models (e.g., Patterson and Whillans 1984) or simulation models (e.g., Summers and McKellar
1981) of wetland ecosystems also could be applied to identify impact networks and thus, taxa that are likely
to be most vulnerable to a particular stressor, and/or are potential keystones in ecosystem energy flow
(Levins 1973). However, modeling approaches are also limited by lack of data on many wetland species and
stressors (e.g., tolerance of wetland organisms to desiccation, burial).
Inevitably, the choice of what to monitor is governed by both policy and scientific considerations. The
following criteria (derived from AMS 1987, Hellawell 1984, Kelly and Harwell 1989, Landres et al. 1988,
Schaeffer et al. 1988, and Temple, Barker, & Sloane 1989), may apply:
Decision factors related to policy implications:
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o Unambiguous - The indicator is socially relevant and easily understood as an indicator of ecological
integrity and/or health;
o Evaluative - The indicator is capable of evaluating the effectiveness of regulations, control, or
management strategies;
o Cost-effective - The indicator is capable of giving a maximum amount of information for a minimum
cost, and thus fiscally attractive;
o Accessible - The indicator is capable of being generated from accessible data sources;
o Anticipatory - The indicator is capable of providing a warning in time to avoid widespread or
irreversible damage.
Decision factors related to scientific implications:
o Sensitivity - The indicator is responsive to the range of conditions likely to be encountered;
o Common - The indicator is sufficiently present in wetlands to be captured by reasonable sampling
effort;
o Integrative - The indicator is capable of integrating effects over time and space;
o Standardized - The indicator is either broadly used and possessing standard methods, or capable of
development of standard methods;
o Reliable - The indicator provides comparable results over a wide range of conditions;
o Predictive - The indicator provides a predictable response to a given stressor or set of stressors;
o Rigorous - The indicator is scientifically accurate, precise, explicit and capable of standard
measurement and reporting protocols that are congruent with the data quality objectives.
The relative weights given each of these evaluation factors will vary depending on the programmatic context,
i.e., for which of the following potential purposes the indicator is being used:
o Determining simply whether a wetland is changing, and in what direction;
o Assessing how aberrant is the community structure of a particular wetland, e.g., to set priorities for
restoration or strategies for mitigation;
o Evaluating the success of management of a wetland, e.g., compliance with permits and mitigation
plans;
o Pinpointing the source of degradation of a wetland;
o Evaluating overall program success of wetland quality protection efforts;
o Priority ranking of wetlands;
o Gaining an understanding of fundamental wetland processes and advancing the science.
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As we examined the technical literature on the most commonly monitored taxonomic groups, we applied
the unweighted criteria to the indicators in a non-systematic, qualitative manner. A resulting summary of
the advantages and disadvantages of each taxonomic group is presented as Appendix A. As data become
available, a more thorough analysis would consider, more specifically, the differences among taxa with regard
to particular stressors in particular wetland types.
To date, there appears to be only one field study (Brooks et al. 1990) that has attempted to compare the
relative sensitivity of major phyla (at the level of community structure) for indicating anthropogenic stress
in inland wetlands. Experimental studies making such comparisons are also virtually non-existent. Future
efforts to develop and compare indicators could focus on studies that circumstantially span a gradient of
disturbed and undisturbed (but otherwise as similar as possible) wetlands of all types. They could compare
all taxa, metrics and data reduction techniques, which, from a theoretical perspective and studies to date,
show promise for use (e.g., Do vegetation similarity measures respond more sensitively to heavy metal
pollution than does wetland invertebrate biomass?). Such future efforts to develop and compare metrics
could emphasize comparisons under different types of temporal and spatial variability.
Given this situation, an alternative approach is to query wetland experts regarding their personal opinions
of taxa and metrics that might be most useful for a given purpose. Some of these opinions have been
published (Table 6). However, recommendations can be unintentionally colored by the expert's degree of
experience with a particular taxon.
As resources allow, rigorous approaches to indicator evaluation might involve integrated laboratory and field
dosing experiments, conducted in parallel with empirical field studies of a series of wetlands that are as
similar as possible but are situationally exposed to various levels (i.e., a gradient) of the same stressor.
This is proposed in EPA's implementation plan for wetland - water quality research (Adamus 1989).
3.2 TYPES OF MONITORING
Monitoring methods might be classified as qualitative and quantitative. Qualitative methods are generally
faster, based largely on visual observation, require little or no sampling equipment, and are usually applied
just along the edges of wetlands. Compared to measurement-based quantitative methods, qualitative
methods are often less replicable and accurate.
One type of qualitative method used in wetland biological monitoring involves use of ground-level (or low-
level) photography. This typically consists of establishing fixed stations at several points around or within
a wetland and taking photographs at specified times. Stations may be surveyed in to known benchmarks
to assure that they may be subsequently located with accuracy, or objects expected to be immobile over time
(e.g., heavy metal stakes) may be included in each picture. Range poles can also be included in pictures to
document scale. Photographs are often pieced together to form a panorama, and video cameras are being
used increasingly to comprehensively document conditions. Photographs can subsequently be evaluated
visually, primarily for major changes in woody vegetation. Time-lapse photography can be used in some
settings to monitor wildlife use. Cameras tethered to balloons have also been used in emergent wetlands
to record interspersion of open water areas with vegetation, and distribution of submersed macrophytes (e.g.,
Edwards and Brown 1960).
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Table 6. Wetland Monitoring Indicators Suggested by Various Scientists.
Aust et al. (1988):
These authors studied silvicultural impacts to wetlands, and found that the most efficient indices of
changes in ecological function (from helicopter logging, skidding, and herbiciding) were soil acidity,
redox potential, oxygen concentration, temperature, soil mechanical resistance, sedimentation, and
vegetation cover. These require short sampling periods, a minimum of laboratory work, and easily
operable and maintainable equipment. Less complex to interpret were sedimentation, net primary
productivity, plant N and P uptake, cellulose decomposition, and bird richness, diversity, and
abundance. Most responsive to disturbance (i.e., showing significant differences across gradients or
between treatments) were total N and P concentrations in soil water, soil acidity, rcdox potential,
saturated hydraulic conductivity, temperature, soil mechanical resistance, sedimentation, net primary
productivity, plant N and P uptake, and cellulose decomposition. Most Inlegratlve of ecological
processes were soil redox potential, net primary productivity, plant N and P uptake, and ccllullose
decomposition rates.
Brooks et al. (1989) and Brooks and Hughes (1988):
For general monitoring of Inland wetlands, the following monitoring parameters were suggested:
hydrology, water quality, hydric soils, vegetation (richness, density, productivity, vertical stand
structure, horizontal patchlness), macroinvertebrates, fish, amphibians, birds, mammals.
Brown et al. (1989):
They proposed the following (in approximate priority order) be monitored for EMAP (EPA's
proposed Environmental Monitoring and Assessment Program for wetlands, in which a probability
sample of 3000 wetlands (50-100 of each of about 13 types) nationwide would be visited once every
3-4 years, with perhaps more-frequent airphoto coverage):
1. Regional changes in the acreage, type diversity, and spatial patterns of wetlands.
2. Nutrients
3. Other pollutants in sediments
4. Hydro period
5. Vegetation (patterns, abundance, richness, composition)
6. Sediment and organic matter accretion
7. Waterbird abundance and species composition
8. Bioaccumulation
9. Macroinvertebrates (abundance, biomass, composition)
10. Leaf area, percent light transmittance, greenness
11. Microbial community structure
12. Bioassays and biomarker measurement
Florida DER (Schwarz et al. 1987):
For state-required monitoring of wooded and cattail-dominated wetlands receiving treated
wastewater, the following parameters are measured: water quality, detention time, vegetation
("importance value' of dominant species), macroinvertebrates (Shannon diversity index), and fish
(biomass ratio of rough fish to sport and forage fish).
Kadlec (1988):
Chemical inputs and outputs normalized to flow, vegetation biomass, sediment and organic matter
accretion.
USEPA (1983):
For monitoring of wetlands receiving wastewater, the following parameters were listed: hydrology,
nutrients, other dissolved substances, trace metals, refractory chemicals, sedimentation, vegetation
(species composition, areal distribution, biomass, growth, production), detrital cycling (organic matter
accretion), bioaccumulation, macroinvertebrates, fish (productivity, biomass, spawning success,
bloassays, incidence of disease), wildlife communities (habitat structure, species richness, density,
indicator species, incidence of disease).
USEPA (Sherman et al. 1989):
For comparison of multiple sets of constructed wetlands with reference wetlands in Florida, New
England, and the Pacific Northwest, the following were measured: water depth, depth to water table,
ambient nutrient concentrations, sediment chemistry, soil oxidation, morphometry and bank slope,
vegetation (species composition, cover, natives vs. exotics).
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Qualitative methods such as these can be used to develop maps of vegetation within wetlands, e.g., Farney
and Bookhout (1982), Meeks (1969), and Morgan and Philipp (1986). Use of cover maps, aerial photos,
and ground photos can be used to identify broad changes in plant composition, as well as providing
permanent records. Suitable, existing, low-altitude color photographs often can be obtained from state
offices of the Agricultural Stabilization and Conservation Service, from the U.S. Forest Service (forest pest
management monitoring programs), and (near roads) from state highway departments, as well as other
sources. Remote sensing has also been used under ideal circumstances to estimate soil saturation, primary
productivity, and sedimentation (Heilman 1982).
A second type of qualitative monitoring involves making visual, ground-level estimates simply of
presence/absence of indicator species and physical conditions (e.g., Terrell and Perfetti 1989) and, in the case
of vegetation, of percent cover. Vast numbers of such unpublished "species lists" are available from
university botanical visits to wetlands, consultant reports, and other sources. While preferable to no data
at all, these represent the "data rich - information poor" syndrome. However, some investigators go beyond
a simple listing of species and visually estimate abundance in relative terms, e.g., rare, common, and this
allows improved interpretation of data. Examples are reports by Dunn and Sharitz (1987), Ehrenfeld 1983,
Kadlec and Hammer (1980), Nilsson and Keddy (1988), Taylor and Erman 1979, Wilcox (1986).
A third type of qualitative monitoring approach involves the use of "wetland evaluation" methods. Many
such methods are available (e.g., see reviews by Adamus 1989, Kusler and Riexinger 1986, Lonard et al.
1981), but differ little in terms of their time requirements. Perhaps the most widely used are:
o Habitat Evaluation Procedures (HEP) of the U.S. Fish and Wildlife Service.
o Wetland Evaluation Technique (WET) developed by EPA and the Corps of Engineers (Adamus et
al. 1987, Adamus et al. 1990).
Although these methods may benefit from or require a limited number of field measurements, they are
predominantly qualitative. They do not directly measure biological communities, but rather, assume
biological community structure or wetland function using information on habitat structure (Schroeder 1987).
Most are applicable at the individual-site level (e.g., WET), while others (e.g., the "Synoptic Approach"-
Abbruzzese et al. in press) operate at regional scales and require more cursory data inputs.
Quantitative methods are the focus of this report. Although many reports and books describe protocols for
biological sampling of lakes and flowing waters, few have attempted to comprehensively describe or evaluate
sampling modifications appropriate for the highly variable, transitional environments of wetlands. Some
relevant information can be found in the following:
Brooks 1989, Brooks and Hughes 1988, Erwin 1989, Fredrickson and Reid 1988a,b, Harris et al.
1984, Homer and Raedeke 1989, Murkin and Murkin 1989, Platts et al. 1987, USEPA 1986,
Welcomme 1979, Woods 1985.
Other parts of EPA's Wetlands Research Program have developed protocols for wetland sampling. For
example, at the EPA-Corvallis Laboratory, the Wetlands Team has developed protocols for monitoring
created or restored (mitigation) wetlands (Sherman et al. 1989). The EPA-Duluth Laboratory has developed
protocols for biological field-sampling of wetlands impacted by a variety of stressors. EPA is refining these
and developing other protocols for support of its nationwide Environmental Monitoring and Assessment
Program (EMAP). This report is not intended to substitute for these other protocols, but rather, includes
them in discussions of a full range of methods available.
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3.3 STUDY DESIGN
For detecting wetland ecological change and estimating its causes, a statistically powerful approach would
involve sampling both before and after the expected change, in both exposed wetlands and in similar,
unexposed wetlands (or in a large, random sample of similar wetlands with unknown exposure history).
Selection of unexposed, or "reference" wetlands is discussed briefly in section 2.1. Statistical issues associated
with wetland biomonitoring are discussed in greater detail by Simenstad et al. (1989).
Because of the high variability of wetland environments, sample collections should be replicated, both within
and among wetlands, and within and among sampling times. One simple option for estimating the minimum
effective number of samples or hours of effort involves plotting a curve. The "x" axis of the curve would
describe the number of samples collected and the "y" axis would describe the community metric being
measured (e.g., cumulative number of species), or its cumulative percent error or variance. Assuming a
reasonably large number of samples have been initially collected, the point where the curve levels off might
be considered to represent the minimum effective sampling effort. Statistical protocols are also available
for estimating requisite number of samples in wetlands, given a desired detection level and initial
information on sample variability (e.g., Downing and Anderson 1985, Eberhardt 1978, Jackson and Resh
1988, Resh and Price 1984).
There are several options for placement of sampling stations. In previous wetland studies, stations most
often have been situated in one of the following ways:
o randomly;
o along transects (usually perpendicular to wetland gradient or flow and extending to the deepest part
of the wetland, and sometimes intentionally aligned to intersect all habitat or topographic "types"
within the wetland);
o at ecotones (spatial boundaries between major vegetation types, and open water and vegetation);
o in proportion to occurrence of habitat types (or hydroperiod classes) present within the wetland;
o at locations subjectively felt by the investigator to "represent" the wetland.
Seasonal timing of sampling is also important, and can be scheduled to coincide with (a) times at which
organisms of concern are most likely to be at maximum numbers, (b) times when these organisms are most
physiologically sensitive to a particular stressor, and (c) times at which concentration of, or organism
exposure to, the stressor is greatest. From this, it is apparent that cost-effective wetland biomonitoring
requires knowledge of (a) life history aspects of wetland organisms, (b) physiology and relative sensitivities
to stressors of the component organisms, and (c) dynamics of physical and chemical factors that largely
determine stressor availability. Most biological surveys of wetlands have been conducted during the growing
season, and relatively little is known of exposure or community structure and function during stressful
conditions of ice cover, severe anoxia, or drought. Time-of-day is also an important consideration,
particularly when monitoring vertebrates. Unless diurnal behavior patterns are well-understood, or there
is sufficient labor available to sample wetlands simultaneously, it may be desirable to alternate the order
in which wetlands are visited, to avoid temporal bias.
The optimal seasonal timing from a biological perspective may not coincide with the best timing from a
perspective of physical human access. Physical access into wetlands is notoriously difficult, and the more
accessible edges of a wetland do not represent the biological conditions in a wetland generally. Although
interior parts of wetlands may be more accessible during ice cover or drought, seldom are these the most
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biologically appropriate times for sampling. Previous investigators have used hip boots, canoes, inflatable
rafts, airboats, helicopters, snowshoes (in summer, for distributing weight in peat bogs to prevent sinking),
and scuba gear for dealing with problems posed by the semi-fluid substrate of many wetlands. For
vegetation, remote sensors can be used for general coverage estimates. Low-altitude video can provide
digital data directly, facilitating spatial analysis (pers. comm., M. Scott, U.S. Fish and Wildlife Service, Fort
Collins, CO). Biomass of submersed aquatic macrophytes was measured electronically by Canfield et al.
1983, Duarte 1987, and Thomas et al. 1990.
3.4 DATA ANALYSIS AND INTERPRETATION
After addressing the question, "What should we measure?" the next logical question is "How do we express
the data?" Thus, in developing and applying wetland biocriteria, the selection and interpretation of
appropriate metrics is at least as important as the selection of appropriate taxa and sampling techniques.
Questions related to wetland metric selection, such as the following, must inevitably be addressed if "data"
are to be converted to "information:"
o Is abundance, biomass, or species richness a more sensitive indicator of wetland biological change?
o When are "guilds" an appropriate way to compile data?
o Do similarity indices and ordination procedures indicate stress from contaminants better than they
show stress from hydroperiod alteration?
o When metrics describing ecosystem structure (such as the above) show that a wetland has changed,
what can be inferred about the wetland's change in function?
Providing a detailed description of all possible techniques for analysis and interpretation of wetland data was
considered beyond the scope of this report. Similarly, the validity and sensitivity of various metrics and
procedures, as applied to the specific taxa and stressors described in later chapters, is not evaluated by this
report. Such an evaluation, perhaps using the evaluation factors listed in section 3.1, would be extremely
important in developing biocriteria for wetlands, but is not currently feasible due to lack of sufficient
comparative data. Some of the more commonly used metrics and analysis procedures are shown in Tables
2, 3, and 4. Review and comparisons of performance of various indices in non-wetland ecosystems are given,
for example, in Green and Vascotto 1978, Huhta 1979, Krebs 1989, Magurran 1988, Matthews et al. 1982,
Polovino et al. 1983, Wolda 1981, Washington 1984, and others. For further information on statistical
analysis of wetland community data the following references (among hundreds) might be consulted: Gauch
1982, Green and Vascotto 1978, Hill 1979, Isom 1986, Jongman et al. 1987, Ludwig and Reynolds 1988,
Pielou 1984, and Wiegleb 1981.
Community-level metrics can also vary greatly in their sensitivity for detecting environmental stress. To
optimize detection of ecologically degraded condition, it is usually best to use several metrics or procedures
in combination (Schindler 1987), as is done by the "Index of Biotic Integrity" that was developed for other
surface waters (Karr 1981).
For other surface waters, information compiled by Sheehan (1984) and others suggests that the approximate
statistical sensitivity of community-level metrics/procedures to pollution has often been:
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cluster/ordination > similarity > richness per unit area or effort > biomass/abundance
procedures indices and diversity indices
However, generalizations such as this contain a high degree of uncertainty. This is because of biases
potentially arising from unknown (and perhaps inconsistent) dependence on a metric's or procedure's
sensitivity to (a) statistical properties of the data set, (b) the particular combination of taxa contained in the
data set (and associated life histories varying from sample to sample), (c) taxonomic level-of-identification,
(d) wetland or community type, (e) spatial scale of measurement, (f) temporal scale of measurement (e.g.,
frequency of sampling, time elapsed since the stressor was maximal), (g) sampling equipment, level-of-
effort, and techniques used. Thus, when only a few metrics and statistical procedures can be applied, results
may be difficult to interpret. Unfortunately, few wetland studies have examined these potential biases. Also,
of particular interest would be (a) the correlation of responses of metrics at several ecological levels, e.g.,
do metrics based on response at the organism level show the same response as those based on data from
the population, community, and ecosystem levels? and (b) the correlation of responses of metrics to
responses in ecosystem function (processes).
A single number from a metric, if used alone, sometimes provides little useful information. Often more
instructive is the particular taxonomic composition that led to a particular summary metric value. Thus,
where data on sensitivities and life histories of organisms are available, aggregating species-level monitoring
data by functional groups ("guilds", see Table 4) of species can provide for more meaningful data
interpretations. It can also reduce the statistical variability in data sets, thus reducing the number of
requisite samples (pers. comm., Dr. James Karr, University of Virginia).
Moreover, shifts in taxonomic composition in response to contaminants frequently are likelier to occur than
changes in total number of species or biomass (e.g., Ferrington and Crisp 1989). However, predicting which
species will become dominant following a wetland disturbance is generally more difficult than predicting
that species composition, overall richness, or biomass-abundance will change (Nilsson and Keddy 1988). In
wetland macrophyte communities, richness is frequently correlated with biomass (Nilsson and Keddy 1988).
This is not true in some communities of wetland fish (Tonn 1985).
All of the above metrics/procedures, except biomass/abundance, commonly employ species-level data. Such
data are easily determined for taxa such as birds, but are much more difficult to acquire for microbial
communities, which have large numbers of species, and for which comprehensive regional references on
taxonomy are virtually non-existent. The need for species-level identifications for the determination of
anthropogenic effects has been asserted by some studies and disputed by others; the need may depend on
the factors listed above that pertain to metric biases, as well as on costs of making more-detailed
identifications vs. costs of collecting a larger number of samples that are only identified at a general
taxonomic level.
The utility of some metrics and procedures, as well as their sensitivity, may vary by wetland type. For
example, metrics and procedures that depend on species-level data (richness, ordination, similarity indices)
may be ineffective in describing the ecological condition of wetlands that characteristically have low species
richness (e.g., breeding bird richness in salt marshes, fish richness in montane wetlands).
The metrics and procedures listed in this report represent only our current abilities to quantify wetland
community structure. From the emerging discipline of "stress ecology" (e.g., Lugo 1978, Odum 1979, 1985),
there may be additional theoretical properties of wetland community structure-such as inertia, elasticity,
amplitude, resilience, hysteresis, malleability, and persistence (to use the terms of Sheehan 1984 and
Westman 1978)~that have potential for quantification and testing. However, only a very few experimental
studies (e.g., Meffe and Sheldon 1990) have quantitatively examined some of these in a regional set of inland
wetlands. If conceptual and operational problems associated with these metrics can be overcome, they may
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hold potential for more sensitively measuring impacts.
After addressing the question of "How do we express the data?" the next logical question is "What represents
normal (or desirable) conditions?" Data interpretation is critical to every monitoring program, and (as
discussed in Chapter 2) "normal" can be defined either in terms of (a) the condition of a reference wetland,
(b) average regional conditions, or (c) ecological conditions necessary for sustaining the ecosystem type
and/or a dynamic balance of its important species. In deference to the vital processes of natural succession
that prevail in many wetland types, a definition of "normal condition" should encompass not only a mean
condition, but the naturally-occurring extremes in structure and function that may be expected over decades
of time (i.e., temporal and spatial variability). This report has not sought to go beyond this general
consideration and attempt to define nominal (normal) and subnominal (abnormal) wetland conditions. Such
an exercise would require an understanding of specific resource management objectives, considerably more
data, and significant public involvement.
Finally, if it has been determined that a wetland is "abnormal," it may sometimes be necessary to conclusively
determine causality. This typically involves laboratory and field bioassay work, a discussion of which is
beyond the scope of this report.
Regardless of which approach is used, caution must be exercised in interpreting community-level data as a
potential indication of anthropogenic stress. Absence of a species may be due merely to random events
(e.g., Grigal 1985). Sampling metrics, particularly species richness, are often very sensitive to the intensity
of sampling, i.e., number of samples, level of effort, size and natural heterogeneity of the wetland sampled.
Also, genetic mutation, natural selection, and/or adaptation can result in evolution of tolerant "ecotypes"-
-local forms of a species that have become tolerant even of normally toxic contaminants. This can alter
competitive relationships and ultimately, community structure. Although it is uncertain as to how
widespread this phenomenon may be, it can be locally important and has been documented to occur in
communities of microbes (Baath 1989), macrophytes (e.g., Christy and Sharitz 1980, McNaughton et al.
1974), aquatic invertebrates (e.g., Krantzberg and Stokes 1989, Kraus and Kraus 1986), and amphibians (e.g.,
Karns 1984).
Also, the possibility that mobile fish or wildlife are avoiding contaminated areas (even temporarily) should
be considered when evaluating community-level vertebrate data. Conversely, wide-ranging biological
indicators may not occur even in the "healthiest" wetlands if most other surrounding wetlands have been
contaminated or altered.
Finally, wetland function cannot always be assumed to change whenever the structure of the biological
community changes. Changes in community composition may be compensatory, such that new species
replace the function of original species and overall community biomass and perhaps richness does not
change (Cairns and Pratt 1986, Herricks and Cairns 1982). An example of this specifically from wetlands
is provided by Cattaneo and Kalff (1986), who conducted invertebrate exclusion experiments in an aquatic
bed wetland. For this reason, it may be advisable to develop and employ, whenever possible, measurements
of both structure and function.
The following sections of this report summarize information relevant to monitoring specific taxonomic
groups, wetland types, and stressors. Again, the purpose is not to be prescriptive, but rather to partially
survey techniques used by other investigators and summarize conclusions that are relevant to future
monitoring. It is expected that these descriptions will be refined and evolve during the review process and
as more data are collected from wetlands. The order of these sections does not necessarily reflect priorities,
but rather is based on phylogeny (taxonomic relationships). Despite the manner of organization, by major
taxa, it is important to recognize that a massive array of interactions can occur in any wetland among the
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separate taxonomic groups, and such competitive interactions, as noted in a few cases in the following text,
can temper the response of an individual taxon to a particular stressor.
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4.0 WETLAND MICROBIAL COMMUNITIES
4.1 USE AS INDICATORS
As used here, "microbes" includes bacteria, viruses, yeasts, and microscopic fungi. In wetlands, these have
most often been measured indirectly, in the pursuit of estimates of microbe-related processes relevant to
element cycling, such as decomposition and denitrification. Although microbial responses to contaminants
have been summarized for other surface waters (e.g., Cairns et al. 1972) and upland soils (Baath 1989), few
studies have looked at microbial community structure specifically in wetlands, or identified particular
microbes as indicators of wetland ecological condition.
Following are discussions of community responses to various stressors. Although we have included some
discussion of decomposition rates (an indirect measure of microbial biomass), that process is mainly
discussed in Chapter 13.
Enrichment/eutrophication. Microbial abundance and community structure are profoundly affected by
trophic status. Enrichment typically results in major increases in microbial abundance (e.g., Tate and Terry
1980) and sometimes richness (Pratt et al. 1989). Enrichment with nitrogen in particular may affect
microbial communities, at least in riverine detritus-based systems. Adding nitrogen to streams increased leaf
decomposition, microbial biomass, and microbial activity; added phosphorus alone had no effect (Fairchild
et al. 1984). Photosynthetic protozoans appear to respond most immediately to nutrient additions (Pratt
and Cairns 1985a). However, effects on species richness and community structure have not been extensively
studied in most wetland types, and little is known of "indicator taxa" whose use might be most appropriate
for signaling enrichment in wetlands.
Microbial colonization rates in a series of shallow Florida ponds was used by Henebry and Cairns (1984)
to indicate trophic status. In pond systems (Schmider and Ottow 1985), enrichment increased microbial
population densities and number of facultative-anaerobic bacteria (e.g., Streptococci, Enterobacteriaceae and
aerobic spore forms, e.g., Bacillus spp., Pseudomonas alcaligenes. and Aeromonas spp.). Mesotrophic ponds
had highest numbers of fluorescent pseudomonads. Oligotrophic water had more denitrifiers (Pseudomonas
fluorescens and Vibrio spp.).
Organic loading/reduced DO. Given the naturally large organic concentrations in wetlands, it is probable
that unique or adapted microbial communities are sometimes present (Felton et al. 1966). Indeed, microbial
communities respond strongly to organic additions (Tate and Terry 1980). However, few studies have
investigated the effects of increased organic loading and decreased dissolved oxygen on wetland microbial
community structure. Low dissolved oxygen (DO) is tolerated or preferred by some taxa, so changes in DO
probably trigger significant shifts in community composition.
In cypress domes that received wastewater, Dierberg and Ewel (1984) found faster rates of leaf
decomposition (a mainly microbial process). In other surface waters, considerable attention has been
focused on coliform bacteria and nuisance growths of Sphaerotilus spp. Large populations of these microbes
characteristically develop where sewage has been introduced.
Contaminant Toxicity. The literature concerning response of microbial community structure to heavy metals
is summarized by Baath (1989), who includes one study from presumably wetland (organic) soils. That study
found a reduction in bacterial abundance at copper concentrations exceeding 275 ug/g. Evidence
summarized from non-wetland soils indicates that heavy metal contamination reduces taxonomic richness of
the microbial community and causes distinct shifts in taxa; some taxa with potential indicator value are
identified by Baath (1989). A shift toward more fungal and gram-negative (vs. gram-positive) taxa may
occur, but there is apparently little change in the overall ratio of mycorrhizal to decomposer fungi.
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Addition of oils and synthetic organics may result in increased abundance of microbes, particularly species
known to degrade and be sustained by petroleum (Walker and Colwell 1977). Microbes, particularly those
associated with wetland plants, can be largely responsible for detoxifying some synthetic organic compounds
(Hodson 1980) such as pentachlorophenol (Pignatello et al. 1985), and the herbicide glyphosate (brand
name, Roundup or Rodeo) (Goldsborough and Beck 1989) as well as detergents (Federle and Schwab 1989).
Thermal Alteration. Although microbial communities are highly sensitive to temperature, few studies have
directly examined the effects of thermal stress on community structure in wetlands. In other surface waters,
Thermus aquaticus has been found only where heated effluents were introduced (Brock and Yoder 1971).
Acidification. Bogs and other acidic wetlands in many cases contain relatively low richness of microbial taxa
(Stout and Heal 1967) and secondary production of microbes can be reduced under such conditions (Benner
et al. 1985). However, naturally acidic bogs can have well-adapted, moderately diverse microbial
communities (Henebry et al. 1981). Zooflagellate microbes and the ratio of bacteria to fungi can decline
with acidification (Leivestad et al. 1976).
Fragmentation of Habitat. We found no explicit information on microbial community response to
fragmentation of regional wetland resources. A study of microbial colonization at various distances from
an intermittently flooded Virginia wetland (McCormick et al. 1987) found that fewer species colonized
introduced substrates that were located a far distance from the wetland; similarity of microbial communities
also decreased with increasing distance. One can surmise that as the distance between wetlands with
potential microbial colonizers becomes greater, microbial taxa with narrow environmental tolerances and
which do not disperse easily might disappear first.
Salinization; Sedimentation/Burial; Turbidity/Shade; Vegetation Removal; Dehydration; Inundation. We
found no explicit information on microbial indicators or community response to these stressors in wetlands.
From knowledge of microbial responses in other surface waters, it appears likely that microbes in wetlands
could respond dramatically to many of these stressors. Undoubtedly data are available from non-wetland
surface waters that identify indicator assemblages and document microbial community response to many of
these stressors (e.g., Krueger et al. 1988). However, reviewing these was beyond the scope of the present
effort, and the transferability of these data to wetlands remains uncertain.
4.2 SAMPLING METHODS AND EQUIPMENT
It is particularly important when using microbes as indicators of anthropogenic disturbance that the
comparison wetlands are of about equal age and have similar sedimentary regimes and vegetation densities.
This is because microbial communities respond strongly to changes in sediment organic matter, which usually
accumulates with wetland age. For example, recently disturbed ponds were found to have fewer microbe
species than did natural and older reclaimed ponds on a surface-mined site (Pratt and Cairns 1985b).
However, microbial communities in ponds more than two years old were indistinguishable from those in
older reclaimed, unreclaimed, and natural ponds despite differences in water quality. Other factors that
could be important to standardize among collections of microbial communities include:
light penetration (water depth, turbidity, shade), temperature, sediment oxygen, baseline
chemistry of waters (particularly pH and conductivity), detention time, current velocity,
vegetation density, dominant vegetation species, and moisture (e.g., time elapsed since last
runoff, inundation, or desiccation event).
Replication requirements for microbial collections are usually significant, due to extremely great spatial and
temporal variability of microbial density and diversity. Protozoan "blooms" are more likely to occur in
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wetlands than in rapidly flowing surface waters, and entirely different communities may exist without
apparent cause within millimeters of each other (Carlough 1989). Sampling can occur at any season, but
microbial biomass is often greatest in late summer (e.g., Murray and Hodson 1985) and autumn. Standard
protocols are available; one is the manual by Britton and Greeson (1988).
Bacterial and fungal abundance are usually estimated as colony forming units (CPU) using plate count
techniques. However, concerns have been raised about the validity of this technique for monitoring fungi;
use of low-nutrient culture media (rather than the typical enriched media) are also recommended (Baath
1989).
Microbial communities in wetlands are generally collected from sediment samples, water column samples,
artificial substrates, or natural organic substrates (e.g., leaf packs). These are described as follows.
Sediment sampling. Sediment sampling of microbial communities can be conducted in all types of wetlands.
Dierberg and Brezonik (1982), working in Florida cypress swamps, sampled microbial communities of surface
sediments using a sterile piston corer and a plastic syringe with an attached tube.
Water column sampling. Any wetland types that have surface water permanently or seasonally can be
sampled using sterile, volumetric containers.
Artificial substrates. Plexiglass plates, acrylic rods, polyurethane foam, or similar inert, sterile surfaces can
be placed in any wetlands that have surface water permanently or seasonally, and allowed to be colonized
by microbes over a period of several weeks (e.g., Goldsborough and Robinson, 1983, Pratt et al. 1985, Pratt
and Cairns 1985b). Substrates are then retrieved and community structure is analyzed. The use of artificial
substrates may be a more practical method of sampling protozoa in wetlands than is direct collecting,
because of the diversity of microhabitats in wetlands (Henebry and Cairns 1984).
Natural substrates. Natural organic substrates typically contain great numbers of microbes. Consequently,
microbial communities have often been collected directly from detrital material, or have been indirectly
monitored through measurement of leaf litter decomposition rates. Microbial biomass can also be indirectly
monitored by analyzing relative levels of adenosine triphosphate (ATP), e.g., the ratio nM ATP/g ash-free
dry weight (Meyer and Johnson 1983). Activity of certain microbial communities was estimated by
measuring relative rates of lipid biosynthesis (Fairchild et al. 1984). Adenylate (ATP, ADP, AMP) energy
charge ratios in microbes also have been suggested as metrics of ecosystem stress (Witzel 1979).
4.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS
In no region of the country, and in no wetland type, have data on microbial community structure been
uniformly collected from a series of statistically representative wetlands. Thus, it is currently impossible to
state what are "normal" levels for parameters such as seasonal density, species richness, and their temporal
and spatial variability. Even qualitatively, lists of "expected" wetland microbial taxa have not been compiled
for any region or wetland type.
Limited data suggest that among-wetland variability in microbial community structure is less than variability
in vascular plant community structure, but that clear differences exist in microbial communities of marshes,
fens, and bogs (Henebry et al. 1981). Microbial density, species richness, and/or colonization rates can be
higher in some wetlands than in other surface waters (Duarte et al. 1988, Henebry et al. 1981).
Studies that have compared microbial communities among wetlands (spatial variation) apparently include
only Henebry et al. (1981, 1984) and Pratt et al. (1989). The former study, covering 13 Michigan wetlands
over a 5-year period, found a range of 93 to 365 protozoan species; Sorenson's similarity index ranged from
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0 to 40, with a mean of 21. The latter study, covering 28 Florida ponds, found a range of 112 to 410
species, with a mean of 338 species in non-artificial ponds. Functional group structure of the resident
microbial fauna changed slightly from year to year, but wetlands in the same geographic region and
experiencing the similar climatic patterns had similar proportions of species in each functional group (Pratt
et al. 1989). Microbial densities can vary by 2 to 5 orders of magnitude between sediments, aquatic plants,
and the water column (Kusnetsov 1970).
Another study, which examined only one wetland complex (Okefenokee Swamp, Georgia) reported that
microbial biomass in sediment ranged from 1 to 28 micrograms gram (dry weight) (Murray and Hodson
1984). A third study, Felton et al.(1967), from Louisiana, reported microbial densities of up to 10*.
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5.0 WETLAND ALGAE
This discussion concerns wetland communities containing phytoplankton, metaphyton, benthic algae,
periphyton, and epiphytic algae. Wetlands may contain algal communities that differ from other surface
waters, or that indirectly influence community composition of algae in receiving waters. For example, acidic
wetland waters commonly are rich in desmid species and acid-tolerant diatoms, such as Eunotia. Frustulia.
and Pinnularia (Flensburg and Spalding 1973, Graffius 1958, Patrick 1977). Marshes may become dominated
by Nostoc pruniforme. Microcoleus paludosus. Vaucheria sessilis. and sometimes Aphanothece stagnina
(Prescott 1968). In a study of the effect on periphyton in a river above and below a marsh, Perdue et al.
(1981) found some species of Navicula were common upriver of a marsh but almost non-existent below the
marsh; several Nitzschia spp. and Fragilaria spp. were common below but rare above the marsh. Fragilaria
construens was abundant in both areas.
5.1 USE AS INDICATORS
As with microbial communities, algal communities in wetlands have most often been measured indirectly,
in the pursuit of estimates of photosynthesis, respiration, and productivity. Few studies have quantified algal
community structure in wetlands, or identified particular wetland algal species as indicators of wetland
ecological condition. However, paleoecological studies of several peatlands have been undertaken. These
use diatoms and pollen from peat cores as indicators of ancient environmental conditions (e.g., Agbeti and
Dickman 1989, Battarbee and Charles 1987).
Following are discussions of algal community responses to various stressors.
Enrichment/eutrophication and Organic loading . Algal blooms are synonymous with eutrophication, so
algae (particularly blue-green forms) are obvious indicators of trophic state, at least in lakes (Hecky and
Kilham 1988). As concentrations of phosphorus in flowing water begin to exceed 0.020 mg/L, or 0.015 mg/L
(and frequently less) in standing water, significant changes in algal communities can begin to occur (e.g.,
Traaen 1978), particularly if flow-adjusted loads are greater than 0.22 g/m3 (Craig and Day 1977). Florida
regulations for discharge of treated wastewater to forested wetlands specify that, on an annual average basis,
waters entering the wetland contain less than 3 mg/L nitrogen and less than 1 mg/L phosphorus; the
monthly average for total ammonia must be less than 2.0 mg/L.
Enriched conditions can be associated with either increased (e.g., Morgan 1987) or decreased (e.g., Hooper
1982, Schindler and Turner 1982) species richness of algal communities, depending on whether algae are
mostly epiphytic or benthic, the pH, water regime, original state of the system, and other factors. Few
studies have used algal community composition to classify the trophic state of wetlands. In other shallow
surface waters, taxa such as the following (for example) have become dominant in response to fertilization
(Mulligan et al. 1976, Patrick 1977, Prescott 1968):
Anabaena Oscillatoria
Aphanizomenon Pandorina
Closterium Pediastrum
Cosmarium Scenedesmus
Dinobrvon Staurastrum
Micrasterias Schroederia
Microcystis
In New Jersey streams exposed to residential and agricultural runoff, Morgan (1987) reported a shift from
species characteristic of the region to species that had been geographically peripheral to the region. Algal
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community structure in some cases might be capable of reflecting the form of enrichment; based on
experiments in a Michigan bog, chlorophytean species responded particularly to ammonium, whereas blue-
green (cyanobacteria) species dominated when phosphate was added (Hooper 1982). Euglenophytes (one-
celled, mobile algae) in particular respond to increases in ammonium and Kjeldahl nitrogen (rather than to
nitrate alone), as well as to other substances associated with decomposing organic matter (Hutchinson 1975).
Near a wastewater-disposal pipeline in a Michigan bog, several algal species bloomed-Cladophera glomerata.
Microspora. Euglena. and Spirogvra (Richardson and Schwegler 1986); algal growth rates were faster at the
outfall site than at the control and at various distances away from the outfall.
Contaminant Toxicity. Numerous studies have demonstrated adverse effects of heavy metals (Whitton 1971),
herbicides, synthetic organics, oil, and/or heavy metals on freshwater algae. Most such studies have been
conducted in laboratories or non-wetland mesocosms, and/or have generally not examined community
structure. Several (e.g., Hurlbert et al. 1972) report major algal blooms occurring after insecticide
application due to temporary suppression of grazing by aquatic invertebrates. Herbicides have been shown
to cause a shift in community composition from large filamentous chlorophyes (green algae) to smaller
diatom species and blue-green algal species, particularly those of the order Chaemaesiphonales
(Goldsborough and Robinson 1986, Gurney and Robinson 1989, Hamilton et al. 1987, Herman et al. 1986).
Following application of phenol to a shallow pond mesocosm, Giddings et al. (1984, 1985) found and
indirectly-caused increase in the dominance of the taxa Euglena, Phacus. Gonium. Coleochaeta. and
Scenedesmus. Oil was predicted by Werner et al. (1985) to shift community composition from algae to
heterotrophic microbes. In other studies, tolerance to high arsenic levels was demonstrated by Chlorella
vulgaris (Maeda et al. 1983) and in a lake contaminated with copper, lead, and zinc, Rhizosoenia eriensis
bloomed while other species declined (Deniseger et al. 1990). Algal assays using highway runoff have
demonstrated chronic toxicity in several cases, probably due to combined effects of heavy metals, road salt,
and sediment (FHWA 1988).
Acidification. Algal responses to acidification in lakes are summarized by Stokes (1981, 1984). Algal
species richness can decline in acidified lakes, particularly in the presence of heavy metals (Dillon et al.
1979). Filamentous algae typically show a proportionate increase, and the genus Mougeotia has been
reported to be a useful indicator of acidification. Nonetheless, algal production can be relatively high in
some naturally acidic wetlands (e.g., Bricker and Gannon 1976).
Thermal Alteration. From knowledge of algal responses in other surface waters (e.g., Squires et al. 1979),
it appears likely that algae in wetlands would respond dramatically to thermal effluents, and that suitable
assemblages of "most-sensitive species" could eventually be identified.
Dehydration/Inundation. Drawdown of wetland water levels often concentrates nutrients and mobilizes
nutrients locked up in exposed peat. This can cause algal blooms in remaining surface water (Schlosser and
Karr 1981, Schoenberg and Oliver 1988). Inundation may have the opposite effect, diluting nutrients,
reducing nutrient mobilization via oxidation, increasing algal competition with vascular plants, and thus
reducing biomass of some algal taxa. However, inundation typically increases the leaf surface area available
for colonization by algae, and provides increased opportunities for dispersal of some algal taxa into and out
of a wetland. In some Prairie pothole wetlands, metaphyton (unattached, filamentous algae that float in a
visible mat) and periphyton (attached algae) increase, while phytoplankton decreases, as higher water levels
reduce the density of vascular plants and increase light penetration (Hosseini 1986).
Other Human Disturbance. In other surface waters, species suggestive of "clean" water include Melosira
islandica and Cyclotella ocellata. Algal or microbial species that can indicate "contaminated" water include
Chlamvdomonas. Euglena viridis. Nitzschia palea. Microcvstis aeruginos. Oscillatoria tenuis. O. limosa.
Stigeocloneum tenue. and Aphanizomenon flos-aquae (Prescott 1968, APHA 1980).
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Salinization; Sedimentation/Burial; Vegetation Removal; Fragmentation of Habitat. We found no explicit
information on algal indicators or algal community response to these stressors in wetlands. From knowledge
of algal responses in other surface waters (e.g., Dickman and Gochnauer 1978), it appears likely that algae
in wetlands would respond dramatically to many of these stressors, and that suitable assemblages of "most-
sensitive species" could be identified.
5.2 SAMPLING EQUIPMENT AND METHODS
Factors that could be important to standardize (if possible) among collections of algal communities include:
age of wetland (successional status), light penetration (water depth, turbidity, shade),
hydraulic residence time, temperature, conductivity and baseline chemistry of waters, current
velocity, leaf surface area and stand density of associated vascular plants, density of grazing
aquatic invertebrates, typical duration and frequency of wetland inundataion, and time
elapsed since last runoff or inundation event.
Standard protocols for algal monitoring are available, although uncertainty exists concerning their
applicability to wetlands. One is presented by the manual of Britton and Greeson (1988).
Replication requirements in wetland algal studies are significant, due to large spatial and temporal
variability. Some investigators have recommended that samples that will be assumed to come from the
same time period should be sampled within a time period less than the hydraulic residence time of the
wetland. Rapid succession in dominant flagellate species was typical of shallow, eutrophic ponds where
conditions fluctuate quickly (Estep and Remsen 1985).
Sampling can occur at any season, but algal biomass is often greatest during the mid to late growing season
(e.g., Crumpton 1989, Hooper 1978, Hooper-Reid and Robinson 1978a, b). In deeper waters, it may be
advisable to sample phytoplankton at mid-day, due to vertical movements at other times (Estep and Remsen
1985). The pigment, chlorophyll-a is sometimes sampled from the water column as an indicator of algal
biomass, but yields little information on community structure. Rabe and Gibson (1984) found greater
phytoplankton density in a shallow vegetated pond than at nonvegetated sites, but species composition was
similar. In contrast, Seelbach and McDiffett (1983) found that a pond with submerged vegetation had more
taxa but lower population density than an open-water pond.
Algal communities in wetlands are generally collected from sediment samples, water column samples,
artificial substrates, or natural organic substrates. Methods are described as follows.
Sediment sampling. Algae can be sampled from sediment surfaces in all types of wetlands. Piston corers,
plastic syringes, or other suction devices are typically used.
Water column sampling. Any wetland types that have surface water permanently or seasonally can be
sampled. Samples from surface waters commonly involve use of volumetric containers or fine-mesh nets.
Vertically-integrating, automated samplers can be used (e.g., Schoenberg and Oliver 1988). Surface
microlayers (top 250-440 micrometers) can be sampled using fine nets or screens mounted on a frame (e.g.,
Estep and Remsen 1985). In flowing-water wetlands, fine nets can be mounted to intercept algae carried
by currents.
Artificial substrates. Artificial substrates (initially sterile materials placed in a wetland and subjected to
natural colonization) may integrate algal assemblages from a large variety of microhabitats. As with
microbial communities, algal communities can be monitored by installing plexiglass plates or similar inert,
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sterile surfaces in any wetlands that have surface water permanently or seasonally, and allowing them to be
colonized by attached algae over a period of several weeks. Substrates are then retrieved and community
structure is analyzed (e.g., Hooper-Reid 1978).
Natural substrates. Natural organic substrates, particularly those in shallow water, may contain a great
biomass of algae. Epiphytic and epibenthic algae are often sampled using a quadrat approach, in which a
frame is placed over a standard-sized area of bottom or a standard volume of the water column is enclosed.
Frame sizes of 10 x 10 cm (Atchue et al. 1983) and 1-2 m2 (Schoenberg and Oliver 1988) have been used.
If algal density is to be estimated accurately, the surface area of substrate must be quantified. This can be
a daunting task in the case of epiphytic algae, where plant surface areas need to be measured. Some
investigators have approached this by measuring surface areas of a random sample of plants, sometimes with
the use of a digital scanner, then measuring their volumes (by displacement) or dry weights and developing
area-volume or area-weight calibration curves. The curves can be used to estimate plant surface area from
future, simpler measurements of the volume or weight of other plants of the same species.
5.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS
In no region of the country, and in no wetland type, have data on algal community structure been uniformly
collected from a series of statistically representative wetlands. Thus, it is currently impossible to state what
are "normal" levels for parameters such as seasonal density, species richness, and their temporal and spatial
variability.
Studies that have compared algal community structure among wetlands (spatial variation) apparently include
only Hern et al. (1978) who studied the Atchafalaya system in Louisiana, and Sykora(1984), who reported
a range of 9 to 21 phytoplankton taxa per ml (mean=9, S.D.=2.3) from a series of six West Virginia
wetlands. Phytoplankton density (cells per ml) ranged from 19 to 2581 (mean=203, S.D.=126). Atchue et
al. (1982) found 56 taxa of phytoplankton in 8 springtime collections from a one-hectare temporary swamp
pool in Virginia. We encountered no journal papers that quantified measurement errors or year-to-year
variation in microbial community structure in U.S. inland wetlands.
Even qualitatively, lists of "expected" wetland algal taxa appear not to have been compiled for any region
or wetland type. Limited qualitative information may be available by wetland type from the "community
profile" publication series of the U.S. Fish and Wildlife Service (USFWS)(Appendix C).
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6.0 NON-WOODY (HERBACEOUS) VEGETATION
This discussion concerns communities dominated by mosses, lichens, liverworts, ferns, sedges, etc., and
includes emergent, floating-leaved, and submersed forms. These taxa probably have been studied more than
any other wetland taxa.
6.1 USE AS INDICATORS
Following are discussions of the community-level responses of herbaceous vegetation to various stressors.
Enrichment/eutrophication. Species richness of herbaceous plants, particularly emergent species, can
increase with moderate enrichment (Graneli and Solander 1988). However, severe enrichment drastically
shifts community structure, and can decrease species richness (e.g., Lachavanne 1985, Lind and Cottam 1969,
Tilman 1987, Hough et al. 1989, Toivonen and Back 1989). This might be particularly true of macrophyte
communities in flowing water wetlands (e.g., Pip 1987), where nutrients otherwise tend to be less limiting
than in most standing water (basin) wetlands. Duarte and Kalff (1988), studying lacustrine macrophytes,
similarly found that the effect of fertilization was influenced by hydrologic energy (e.g., wave action).
The greatest richness of emergent plants has been reported to occur when standing biomass of the
community is less than 1000 g/m^ in British wetlands, 400-500 g/m^ in Netherlands wetlands (Vermeer and
Berendse 1983), and 60-500 g/m^ in Ontario wetlands. If a goal is to maintain within-wetland species
richness, the particular nutrient loadings that result in a desired biomass might be calculated from empirical
data (e.g., Duarte and Kalff 1986, Duarte et al. 1986) to derive very approximate criteria for nutrient
loadings, and perhaps, with further testing, for other factors that can increase plant biomass (e.g., thermal
warming, hydrologic regime). However, the numeric ranges just given are probably less valid for wetlands
that are grazed or subject to other significant vegetation removal processes.
Changes in composition and growth of herbaceous communities as a probable result of increased nutrients
have been reported by many ecologists, including:
Guntenspergen 1984, Haslam 1982, Kullberg 1974, Jensen 1979, Kadlec and Hammer 1980,
Klowsowski 1985, Kohler et al, Mahoney 1977, Pringle and van Ryswyck 1968, Schwartz and
Gruendling 1985, Seddon 1972.
An important regional impact of excessive enrichment is that small, regionally rare plant species (that often
characterize infertile wetlands or wetlands whose chemistry reflects weak buffering) are often out-competed
by large, regionally common species (Day et al. 1988, Moore et al. 1989). Insectivorous plants, quillworts,
many species that typify fen wetlands, and some orchids and mosses that typify oligotrophic wetlands, are
particularly sensitive to enrichment, either airborne or waterborne (Moore et al. 1989, Roelofs 1983, 1986,
Schuurkes et al. 1986). Percent cover of the dominant peat-forming mosses of bogs can probably be reduced
by atmospheric nitrogen deposition rates of 4.3 g/m^/yr, but not 2.0 g/m^/yr (Ferguson et al. 1984).
However, because great variability exists among tolerances of moss species, a limit of 2.0 and possibly (in
oligotrophic wetlands) 1.0 g/m^/yr has been suggested (Schuurkes et al. 1987, Liljelund and Torstensson
1988).
Submersed and floating-leaved or mat-forming species usually respond more strongly to enrichment than do
emergents (e.g., Ozimek 1978, Shimoda 1984), because the former obtain nutrients directly from the water
column, whereas the latter obtain them from sediments. In many regions, vascular floating-leaved plants
such as pondweed (Nurrhar), duckweed (Lernna), and water-meal (Wolffia) become more prevalent with
increasing enrichment (e.g., Bevis and Kadlec 1990, Burk et al. 1976, Ewel 1979, Kadlec et al. 1980), and
in severely eutrophic lakes, emergent species may survive as floating mats (Graneli and Solander 1989).
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Species shifts may be less immediate or noticeable when moderate amounts of nutrients are added to
wetlands that are already eutrophic, because high microbial populations that characterize such environments
can be highly effective at first in competing for the nutrients (e.g., Richardson and Marshall 1986). With
extreme enrichment, submersed macrophytes can eventually decline, probably as a result of being shaded out
by algae (e.g., Mulligan et al. 1976, Phillips et al. 1978), and emergent species may increase.
Among emergent species, extreme eutrophication causes decreased species richness because (a) turbidity of
phytoplankton blooms shades out many submersed species, and (b) as phytoplankton decays, resultant oxygen
deficits in bottom sediments probably stress the most sensitive rooted species (e.g., Hartog et al, 1989). Of
the emergent species, cat-tail (Tvpha) and common reed (Phragmites) often dominate enriched wetlands and
may be the least sensitive to the initial stages of eutrophication (e.g., Kadlec 1979, Hartland-Rowe and
Wright 1975, Kadlec 1990, Kadlec and Bevis 1990, Moore et al. 1989). Cat-tail biomass and production
respond to annual fluctuations in nitrate, making cat-tail a successful opportunist capable of dominating
wetlands that have erratic inputs of nutrients (Davis 1989). Although Phragmites can exist without any
obvious sign of harm in wetlands with at least 6 mg/L phosphorus and 10 mg/L nitrogen (Ostendorp 1989),
massive die-offs of this species in European wetlands have been attributed by some to excessive enrichment
(Hartog et al. 1989, Ostendorp 1989). Another emergent plant-manna grass (Glvceria grandisV-increased
in dominance in an Ontario wetland subjected to treated effluent (Mudroch and Capobianco 1979).
In Michigan, moderately eutrophic lakes were dominated by Ccratophvllum demersum. Utricularia vulgaris.
and Cladophora fracta (Hough et al. 1989). In England, Potamogeton pectinatus. Mvriophvllum spicatum.
and Hippuris vulgaris dominate in highly eutrophic waters (Butcher 1946, Seddon 1971). However, Kadlec
et al. (1980) and Mulligan et al. (1974) found Mvriophvllum declined under increasing fertilization, along
with Ceratophvllum demersum. Polygonum and Utricularia. Many wetland plant species are categorized by
their nutrient-level preferences, and thus as their potential as indicators of eutrophication, in reports by
Ellenberg Jeglum 1971, Moyle 1945, Pip 1979, Stewart and Kantrud 1972, Swindal and Curtis 1957, and
Zoltai and Johnson 1988.
Even the submersed types of herbaceous vegetation appear a poorer indicator of eutrophication than are
algal communities, which respond more quickly (Crumpton 1989). Neither macrophyte nor algal taxa are
reliable indicators of moderate enrichment in naturally enriched waters, e.g., minerotrophic fens, wetlands
in karst limestone regions (Hellawell 1984, Strange 1976).
Organic loading/reduced DO. Existing literature often does not adequately distinguish the effects on
herbaceous plants of organic loading/reduced DO, from the effects of nutrients (discussed above) or
inundation (discussed below).
At least in the short term, the biomass of herbaceous plants generally increases with moderate additions of
wastewater. In acidic, oligotrophic wetlands (e.g., bogs), species richness may increase (e.g., Guntenspergen
1984). Community components with short turnover times, such as aboveground biomass and leaf area of
annual plants, can respond most sensitively (e.g., Brown 1981, Odum et al. 1984).
Aggressive, introduced annuals sometimes replace native perennial species (e.g., Finlayson et al. 1986).
While the occurrence of rarer, perennial species is often correlated with specific chemical conditions, the
occurrence of aggressive, common species often is not (Pip 1979). Populations of such species tend to be
more plastic in their response to wastewater enrichment (e.g., Guntenspergen 1984).
Over longer periods of time and/or excessive loading, wastewater additions may result in stress from low
dissolved oxygen, increased hydrogen sulfide, and excessive accumulation of sediment organic matter. These
conditions can selectively inhibit certain plant taxa (Barko and Smart 1983), particularly those that are
unable to translocate oxygen to their roots (Brennan 1985). While cattail (Tvpha) require only trace
amounts of dissolved oxygen for germination (Leek and Graveline 1979), bud development is more successful
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in reeds (Phragmites) if flooded soils are aerated (Haslam 1973), as is sprouting of purple nutsedge ((
rotundus) (Al-Ali et al. 1978).
Morgan and Philipp (1986) surveyed a host of New Jersey streams and listed 22 species found only in
streams that, based on their location and limited chemical sampling, were assumed to be "polluted." The
researchers found 18 only in "unpolluted" streams, and 21 in both types. Callitirche heterophylla. Ludwigia
palustris. Polvgonum punctatum. Potamogeton epihydrus. and Sparganium americanum were locally dominant
only in polluted streams, and Sagittaria englemanniana. Scirpus subterminalis. and Vaccinium macrocarpon
were dominant only in unpolluted streams. Polluted sites, with high nitrate and pH, had a higher
percentage of non-indigenous species, vines, and herbaceous (vs. woody) plants. Vines and other low-
growing species also were found by Nilsson and Grelsson (1990) to dominate riverine sites with intermediate
accumulations of organic matter (i.e., 100-200 g/m^ leaf litter), whereas sites with very low or very high
accumulations of organic matter were dominated by stemmed species. Emergent plant species richness also
showed such a quadratic correlation with accumulated organic matter.
Contaminant Toxicity. Some herbaceous plants are quite sensitive to heavy metals and other contaminants,
and as a result, contamination can alter species composition, and decrease species richness, canopy coverage,
and net annual productivity of wetland communities (e.g., Cooper and Emerick 1989, Olson 1979). Based
on studies of eight Colorado wetlands exposed to varying degrees of heavy metal-contaminated runoff,
Cooper and Emerick (1989) noted:
"Subalpine fen wetlands in the Colorado Front Range that have less than 3 vascular plant
species growing in the main part of the wetland (not the edges) and have less than 50
percent total canopy coverage and less than 100 g/rn^ total annual primary production, are
likely to indicate impact from heavy metal toxicity. An exception is areas that are flooded
or have ponded water for much of the growing season."
Forbs (herbaceous dicots in that study) seemed particularly uncommon in polluted wetlands. The authors
noted no species that occurred only at contaminated sites, but found that the sedges, Carex aquatilis. C.
utricularia. and/or C. scopulorum, predominated in these areas. Species absent from areas contaminated by
large concentrations of heavy metals included the following:
Swertia perennis Cardamine cordifolia
Caltha leptosepala Epilobium lactiflorum
Geum macrophvllum Galium trifidum
Sedum (Clementsia) rhodantha Juncus albescens
Bistorta bistortoides B. vivipara
Polvgonum (Bistorta) bistortoides and vivipara
Duckweed (Lernna) is particularly sensitive to the heavy metals cadmium and nickel, and chromium
concentrations of 10 mg/L are inhibitory (Huffman and Allaway 1973). Cattail (Tvpha latifolia) can tolerate
lead, copper, and chromium accumulations of at least 10 micrograms/g dry weight of aboveground biomass;
zinc accumulations in cattail may reach 25 micrograms/g dry weight without apparent ill effects (Mudroch
and Capobianco 1979). The common reed (Phragmites) can tolerate industrial wastewater with high levels
of heavy metals (e.g., up to 250 micrograms/g sediment copper concentrations), as do bulrushes (e.g., Seidel
1966).
Heavy metals and other toxicants borne in air currents and precipitation have widely been reported to alter
community composition of mosses and lichens (e.g., Lee et al. 1987, Sigal and Nash 1980). Species of
mosses and lichens differ considerably in their sensitivity to metals, and are prevalent in many wetland
types. Thus, they may have considerable potential for use as indicators of this type of pollution.
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A decline in Asclepias syriaca (milkweed) and an overall increase in species richness and equitability may
have been related to contaminants associated with incinerator residue deposited in an emergent marsh in
Massachusetts (Mika et al. 1985). In a major Ohio river, Stuckey and Wentz (1969) reported the following
species to be rare or absent from waters contaminated by industrial effluents, but common in analogous
uncontaminated habitats:
Justicia americana Lippia lanceolata
Saururus cernuus Helenium autumnale
Phytostegia virginiana Eclipta alba
Rumex verticillatus Scirpus americanus
Samolus parviflorus Amaranthus tuberculatus
Carex frankii Hibiscus militaris
Lvcopus rubellus Strophostyles helvola
Also, these investigators found the following plants to be common in industrially contaminated waters:
Polygonum hvdropiper Echinochloa pungens
P. persicaria Leersia oryzoides
P. pensylvanicum Ambrosia trifida
P. coccineum Urtica dioica
P. lapathifolium Arctium minus
P. punctatum Bidens frondosa
Sagittaria latifolia
In an Ontario river, submersed species (Elodea, Ceratophyllum. and Mvriophyllum') appeared to be less
tolerant of industrial wastes than floating-leaved and short, rooted aquatic plants (Potamogeton. Nuphar.
and Nvmphaeal. which were in turn less tolerant than cattail (Typha) and common reed
(Phragmites)(Dickman et al. 1980, 1983, Dickman 1988).
Floating-leaved herbaceous plants are sensitive to the physical effects of oil, and growth of the duckweed
Spirodela oligorhiza is affected by PCB concentrations of 5 mg/L (Mahanty 1975). Cattail can tolerate
petroleum oil concentrations of 1 g/L (Merezhko 1973) and, along with common reed (Phragmites),
appeared to be the most tolerant macrophyte downstream from an industrial effluent source in Ontario
(Dickman 1988). The response of wetland species to an oil spill in a Massachusetts inland wetland (Burk
1977) was as follows (* = annual species):
Species not recorded after oil spill:
Bidens cernua* H. virginicum
B. connata* Iris versicolor
B. frondosa* Lvcopus uniflorus
Echinochloa waited* Mimulus ringens
Eleocharis obtusa Polygonum punctatum*
Galium tinctorium* P. sagittatum*
Hvpericum mutilum Sparganium americanum
Spirodela ployrhiza Vallisneria americana
Verbena hastata
Species reduced after oil spill:
Cephalanthus occidentalis Najas flexilis*
Eleocharis acicularis Onoclea sinsibilis
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Galium trifidum Pilea fontana*
Leersia oryzoides Pontederia cordata
Lindernia dubia* Scirpus pedicellatus
Ludwigia palustris Zizania aquatica
Species apparently unaffected or increasing after oil spill:
Alisma subcordatum Polygonum coccineum
Carex lurida Potamogcton crispus
Ceratophvllum demersum P. epihydrus
Dulichium arundinaceum Sagittaria graminea
Eleocharis palustris S. latifolia
Elodea nuttallii Salix nigra
Equisetum fluviatile Scirpus cyperinus
Lemna minor S. validus
Lvsimachia terrestris Scutellaria leteriflora
Nuphar variegatum Sium suave
Veronica scutellataq Vitis labrusca
Bulrushes are killed by phenol concentrations of 100 mg/L and abnormalities occur at large phenol
concentrations, but new shoots form quickly (Seidel 1966). Herbicides have often been used to control some
herbaceous species, notably purple loosestrife (Lythrum) and common reed (Phragrnites), and undoubtedly
affect some non-target species as well. However, herbicide effects can be species-specific, with the result
being that some applications result in overall increase in algae and plant richness (although perhaps lower
overall productivity), as monotypic or dominant stands are opened for invasion by less aggressive species
(e.g., Murphy et al. 1981). Detergent concentrations of 15 mg/L can damage wetland macrophytes (Agami
et al. 1976).
Additional toxicological information may be available through EPA's PHYTOTOX (Royce et al. 1984) and
AQUIRE databases.
Acidification. Ambient pH is one of the most important factors affecting community composition of
emergent and aquatic bed wetlands bordering northern lakes (Hultberg and Grahn 1975), as well as
peatlands (e.g., Anderson 1986, Jeglum 1971) and perhaps other low-alkalinity, standing water wetlands. It
can be a stronger influence in these systems than nutrient status or water transparency (e.g., Jackson and
Charles 1988). However, its effect on overall species richness is unclear (Eilers et al. 1984, Jackson and
Charles 1988, Yan et al. 1985,). Usually, fewer species of macrophytes are found in acidic lakes than in
circumneutral lakes (e.g., Friday 1987, Hunter et al. 1986, Hutchinson et al. 1985), but these are often
species that are regionally rare (Moore et al. 1989).
The study of Adirondack (New York) lacustrine wetlands by Jackson and Charles (1988) reported the
following taxa to be relatively intolerant of acidification: Najas flexilis. Nitella flexilis. Potamogeton pusillus.
P. natans. and P. amplifolius. Submersed and floating-leaved species present at pH lower than 5.5 but not
in less acidic conditions included Potamogeton confervoides and Sparganium angustifolium: species present
in both acidic (pH < 5.0) and circumneutral wetlands included Nuphar. Juncus pelocarpus. Drepanocladus
fluitans. Utricularia vulgaris. Isoetes muricata, Eriocaulon septangulare. Sagittaria graminea. and
Mvriophyllum tenellum (Jackson and Charles 1988). Emergent species present in both acidic (pH < 5.0)
and circumneutral wetlands included Calla pallustris. Juncus brevicaudatus. Dulichium arundinaceum.
Lvsimachia terrestris, and Juncus pelocarpus (Jackson and Charles 1988). Wolffia. Lemna. and Spirodela
have optimal pH's of 5.0, 6.2, and 7.0 respectively, whereas their tolerated ranges are (respectively) 4 - 10,
4 - 10, and 3 -10 (McClay 1976).
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In some northern wetlands, especially those that are heavily shaded, acidification can result in increased
presence of mat-forming mosses of the genus Sphagnum (e.g., Gignac 1987, Grahn 1976, Roberts et al.
1985), and these mosses can further lower the pH (e.g., Glime et al. 1982). However, under severe
acidification and accompanying deposition of industrial pollutants, Sphagnum can decline and in some cases,
be replaced by cottongrass (Erior2horurn)(e.g., Gorham et al. 1987, Lee et al. 1987a).
Cattail, rushes, and sedges occur in sediments with a pH of at least 4.7 (Dykyjova and Ulehlova 1978), while
common reed and nutsedge can tolerate a pH as low as 2.0 (Al-Ali et al. 1978, Dykyjova and Ulehlova
1978). Natural stands of sedge (Carex) have a pH range from 4.9 to 7.4 (Baker 1971), while the range for
reed canary-grass (Phalaris) is 6.1 to 7.7 (Gross and Jung 1978, Dean and Clark 1972, Niehaus 1971,
Allinson 1972). Many regional botanical texts describe approximate pH ranges of individual wetland species
(e.g., Crow and Hellquist 1981), as does some literature not excerpted here (e.g., Jeglum 1971, Swanson
1988).
Reductions in plant species diversity, decreased productivity, and life cycle disruptions were among the
effects attributed to high pH values downslope from a Massachusetts hazardous waste lagoon (USEPA
1989a).
Salinization. Saline inland wetlands commonly have fewer species of macrophytes (Pip 1979, Reynolds and
Reynolds 1975), and may be particularly deficient in species that typically form floating mats (Lieffers 1984).
Most freshwater macrophytes cannot tolerate more than 10 ppt dissolved salts (Reimold and Queen 1974).
Inland wetland plants that reportedly tolerate specific conductivity of greater than about 5 mS/cm are shown
in Table 7, from Kantrud et al. (1989). Other data on salinity tolerances of inland wetland plants are
provided by Reimold and Queen (1974) and others listed in Table 7.
Contamination of a northern Indiana bog with road salt resulted in almost complete elimination of endemic
species and replacement by non-bog species, dominated by Tvpha angustifolia (Wilcox 1987), which can
sometimes tolerate salinities of up to 25.5 ppt (Philipp and Brown 1965, Shekov 1974), at least for short
periods. As salt concentrations declined in the four years of the study, endemic plants began to recolonize
the affected area; biomass and growth of Sphagnum fimbriatum was significantly reduced at NaCl
concentrations greater than 900 mg/L Cl- (Wilcox 1987). The common reed (Phragmites communis')
tolerates salinities of up to 45 ppt, although seedlings may be killed by salinities of 10 ppt. Duckweed
(Lemna minor) has reduced growth at salinities above 7 ppt (Haller et al. 1974, Stanley and Madewell
1976). For many species, these values vary by genetic population, life stage, duration of exposure,
temperature, and other factors. The freshwater cattail, Tvpha latifolia. as expected, is less salinity-tolerant
than the estuarine cattail, Tvpha angustifolia, mentioned above (McNaughton 1966). However, a presumed
hybrid, Tvpha gauca. appeared resistent to road salt runoff (Bayly and O'Neil 1972). Even Tvpha latifolia
seeds appeared more tolerant of road salt in snowmelt than germinating wool-grass CScirpus cvperinus) and
three-way sedge (Dulichium arundinaceum); purple loosestrife seeds (Lythrum salicarial were similarly
tolerant (Isabelle et al. 1987). The rush, Scirpus acutus. appears more salt-tolerant than its many of its
congeners (Smith 1983).
48
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Table 7. Examples of Aquatic Macrophytes Tolerant of Saline Conditions in Inland Wetlands.
These lists are reproduced from Kantrud et al. (1989), and deal primarily with prairie pothole wetlands;
applicability to other wetland types is unknown. Additional salt-tolerant species may be found in lists of
Haller et al. 1974, Kauskik 1963, Lieffers 1984, Mall 1969, McKee and Mendelssohn 1989, Millar 1976,
Millar 1978, Moyle 1945, Nelson 1954, Pip 1979, 1987a,b, Reynolds and Reynolds 1975, and Stewart and
Kantrud 1972.
Specific conductivity (mS/cm)*
Species Mean Min. Max.
Vernonia fasciculata
Agrostis stolonifera
Lycopus americanus 0. 3
Potentilla rivalis 0. 3
Carex stipata 0.4
Equisetum arvense 0. 4
Juncus interior 0. 4
Aster sagittifolius
Plantago manor
Potentilla norvegica
Juncus dudlevi 0.4 0.3
Carex buxbaumii 1.2 1.0 1.4
Lvsiraachia hybrida 0.1 0.1 1.6
Carex vulpinoidea 1.0 0.1 1.7
Ranunculus macounii 1.1 0.1 2.1
Rumex mexicanus
Juncus bufonius
Cirsium arvense 2.5
Bidens cernua 1.5 0.7 2.5
Helenium autumnale 1.5 0.5 2.5
Carex praegracilis 0.3 0.1 3.0
Echinochloa crusgalli 1.3 0.5 3.2
Carex laeviconica 1.5 0.1 3.2
Rorippa islandica 1.7 0.1 3.2
Poa palustris 1.4 Tr.° 3.4
Stachvs palustris 1.8 0.1 3.6
Calamaarostis canadensis 1.4 '0.4 3.8
Carex sartwellii 1.5 0.4 3.8
Lycopus asper 1.9 0.4 4.4
Epilobium alandulosum 1.5 0.5 4.7
Mentha arvensis 1.6 0.1 4.9
Apocvnum sibiricum 1.8 0.4 5.0
Eleocharis compressa 2.0 0.7 5.0
Carex tetanica 2.0 0.9 5.5
Potentilla anserina 1.6 0.1 6.0
Boltonia asteroides 1.4 0.1 6.8
Carex lanuginosa 2.0 0.1 9.1
Teucrium occidentale 3.1 0.2 9.1
Aster hesperius 2.4 0.4 9.8
Juncus torreyi 1.7 0.2 10.0
Aster simplex 1.8 0.1 16.1
Ca1amaarostis inexpansa 2.6 Tr. 17.6
Juncus balticus 3.3 0.1 20.1
49
-------
Table 7 continued
Specific conductivity fmS/cm)*
Species Mean Min. Max.
Spartina crracilis
Plantacro eriopoda
Sonchus arvensis
Spartina pectinata
Muhlenbercria asperifolia
Hordeum iubatum
Trialochin maritima
Distichlis spicata
Atriplex patula
9.0
9.8
5.2
3.0
11.0
7.8
12.5
17.0
23.0
0.7
1.0
0.5
Tr^
0.7
Tr.
0.7
0.5
6.9
20.1
20.1
20.8
33.5
38.5
48.6
50.9
76.4
76.4
"Underlined means (Disrud 1968; Kantrud et al. 1989) indi-
cate surface water measurements in wetlands where the
species reached peak abundance; underlined ranges (ibid)
are for instances where the species occurred in waters of
greater or lesser salinity than that recorded by Smeins
(1967).
Indicates measurements <0.05 mS/cm.
50
-------
Tab] t? 7 continued
Species
Hater regime
Specific conductivity fas/cm^*
Mean Min. Max.
E^iSStUB fluviatile
Galiua trifidua
Scutellaria qalericulata
Jmpatiens biflora
Mimulus ringens
EVPflt°riu.ni aaculatua
Saqittarifl cuneata
Glvceria striata
Ranunculus gaelini
Asclepias incarnata
Pamassia alauca
Glvceria boreal is
Salix interior
Carex lacustris
Solidago oraainifolia
Polyqonua aaphibiua
Scirpus atrovirens
Cicuta maculata
Eriphorua anqustifoliua
Carex rostrata
Polyqonua coccineua
Phalaris arundinacea
Carex aouatilis
Lvsiaachia thrvsi flora
Glvceria qrandis
Slum suave
Scirpus heterochaetus
Alopecurus aegualis
Sparganiua eurvcarpua
Eleocharis acicularis
Scirpus validus
TYPha X qlauca
Saqittaria cuneata
Scirpus fluviatilis
Alisaa qranineua
Carex atherodes
Ranunculus sceleratus
Bectaaannia syziqachne
Alisaa plantaqo-aquatica
Ranunculus cymbalaria
Scolochloa festucacea
Phraqaites austral is
Tvpha latifolia
ZypJlA anqustifolia
Eleocharl.s palustris
Scirpus nevadcnsis
Scirpus acutus
Suaeda depressa
Scirpus aaericanus
Scirpus maritiaus
Puccinell,ia nuttalliana
SE
SA
SA
SA
SA
SA
SE
SA
SA
SA
SA
SE
SA
SA
SA
SE
SA
SA
SA
SA
SE
SE
SA
SA
SE
SE
SP
SE
SE
SE
SP,SA
SP
SE
SP
SE
SE
SE
SE
SE
SE
SE
SA
SP,SA
SP
SE
SE
SP
SE
SE
SP
SE
SL2.
0.3
SLA
SLA
0.7
O*l
SLS.
SLS.
SLS.
0.9
1*O
SLA
1.4
Qt8
0.6
1.0
1.4
1.8
1.1
1.3
1.6
1.6
1.7
0.7
1.8
1.4
1.1
1.8
1.5
1.8
SLS.
1.8
1.9
2.0
2.0
3.6
1.5
1.8
3.5
3.4
3.5
2.1
3.4
2.7
15.7
4.3
24.0
4.9
10.3
20.0
—
—
—
—
— —
—
—
—
SLS.
—
—
0.3
0.9
0.1
0.1
SLS
SLS
0.5
Tr .
p. i
0.3
SLS.
Tr .
p. i
0*1
0 , i
IE*.
0*1
SL2.
0. 1
0. l
0.3
0.3
0 . i
Tr.
lEx
0.6
0,1
0*1
SLA
0*1
12.0
SL2.
5_*£
SLS.
1*1
—
— —
—
—
— —
—
—
—
SL3.
—
—
1.7.
1.7
1*1
2 , 2
2*2.
2.2
2.2
2.6
3.4
3.8
3.8
3.8
4.0
1*£
4.2
4.5
4.6
5.8
6.2
6.6
6.7
£*1
6.7
8.5
9.5
9.5
9.5
12.1
12.4
13.6
13.6
14.5
20. S
24.0
66.0
70.0
76.4
76.4
"Underlined Beans (Disrud 1968; Kantrud et al. 1989) indicate
surface water aeasurements in wetlands where the species reached
peak abundance; underlined ranges (ibid) are for instances where
the species occurred in waters of greater or lesser salinity than
that recorded by Sacins (1967).
Indicates neasureaents <0.05 aS/ca.
SA=semiannual, SE=seasonal, SP=semipermanent
Si
-------
Table 7 continued
Species
Specific conductivity fmS/cm)*
Mean or single
measurement Min. Max.
Myriophyllum pinnatum
Nuphar yariegatum
Najas f lexilis
Elodea canadensis
Potamoqeton friesii
MyriophyHum verticillatum
Potamoqeton gramineus
Callitriche palustris
Hippuris vulgaris
Callitriche hermaphroditica
Ranunculus flabellaris
Potamoqeton zosteriformis
Spirodela polyrhiza
Ricciocarpus natans
Drepanocladus spp.
Potamoqeton vaginatus
Potamoqeton richardsonii
Ranunculus subrigidus
Riccia f luitans
Ceratophvllum demersum
Potamoqeton pusillus
Myriophvllum spicatum
Utricularia vulgaris
Lerona minor
Lerona trisulca
Ruppia maritima var . occidental
Zannichellia palustris
Chara spp.
Potamoqeton pectinatus
Ruppia maritima var. rostrata
0.4
0.6
0.8
Oil
1.2
2.0
1.0
1.5
2.2
1.2
3.3
1.7
1.4
2.1
2.1
2.1
2.2
2.7
3.1
3.2
is 4.1
4.8
2.2
6.5
36.1
0.3
on
oTT
0.6
0.6
0.3
1.0
1.2
2.5
2.5
2.8
3.0
3.2
4.0
4.5
4.7
5.1
8.1
10.9
13.9
14.2
25.0
42.0
60.0
66.0
"Underlined means (Disrud 1968; Kantrud et al. 1989) indicate
surface water measurements in wetlands where the species reached
peak abundance; underlined ranges (ibid) are for instances where
the species occurred in waters of greater or lesser salinity than
that recorded by Smeins (1967).
52
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Sedimentation/Burial. We found little explicit information on overall macrophyte community response to
burial or sedimentation. In some cases, sedimentation creates shoals in rivers or lakes, which provide
sufficient substrate within the euphotic zone for herbaceous wetlands to become established or expand, at
least until a major scouring flood re-occurs (e.g., Burton and King 1983). Where sedimentation is severe,
water may become too shallow for some submersed species and a shift to emergent species may occur
(Edwards 1969).
Differences probably exist among herbaceous plant species with regard to their intrinsic tolerance and
adaptability to excessive sedimentation. Species often noted as occurring in disturbed, sediment-laden
wetlands include the common reed (Phragmites). reed canary-grass (Phalaris)(Reed et al. 1977), and other
large and robust taxa. Repeated burial by as little as 5 cm of sediment per year can be detrimental to some
emergent species (van der Valk et al. 1981).
Shading/turbidity. Increased shade or turbidity (whether from suspended sediment, phytoplankton, natural
staining, or other sources) generally results in a shift in community structure from submersed species to
floating-leaved or emergent species (Hough and Forwall 1988). Turbidity increases and decreases in bank
stability may also favor an increase in the proportion of invasive, dominating species to the exclusion of less
aggressive native macrophytes (Morin et al. 1989).
A 25 NTU (nephelometric turbidity units, or about 100 mg/L suspended solids) increase in turbidity in a
shallow riverine wetland can reduce production of algae and submerged aquatics by 50 percent (Lloyd et al.
1987) and a mere 5 NTU increase (about 20 mg/L) has been shown to reduce the productive area of a lake
by about 80 percent (Lloyd et al. 1987). The sensitivity of submersed plants to turbidity can be expressed
by the ratio of the depth maxima of species to the Secchi transparency depth, i.e., the "turbidity tolerance
index" (Davis and Brinson 1980). Data on depth maxima and ranges for many submersed species are
compiled in Davis and Brinson (1980). The more shade-tolerant non-emergent herbaceous species are listed
in Table 8.
Vegetation removal. Harvesting of "aquatic weeds" comprises a direct impact on submersed vegetation, and
can shift the community composition at least temporarily. Species richness can either increase or decrease,
depending on the initial state and species that are harvested (Sheldon 1986). At the deepwater edge of
lacustrine wetlands with submersed plants, milfoil (Mvriophvllum spicatum) frequently becomes dominant
following the catastrophic alteration of more diverse communities by dredging, herbicides, disease, storms,
herbivory, or other factors (Nichols 1984).
Removal of woody overstory generally increases herbaceous vegetation biomass and diversity (Madsen and
Adams 1989). In the Prairie pothole region, specific information on shifts in community composition as a
result of vegetation removal from grazing, haying, and cultivation, is reported by Kantrud et al. (1989:
Appendix B). Annual burning, at least of emergent wetlands of the mesic pine-wiregrass savannas of North
Carolina, can increase species richness (Walker 1985).
Thermal Alteration. Changes in wetland thermal regime can cause changes in production and shifts in
species composition of the herbaceous plant community (Allen and Gorham 1973, Haag and Gorham 1977).
An eventual shift from perennial and woody species to annual and herbaceous species may also occur in
wetlands exposed to intermediate degrees of thermal warming (Dunn and Scott 1987, Sharitz et al. 1974).
Changes are due both to physiological factors and (in northern wetlands) to changes in ice cover (Geis
1984) and growing season length. Most aquatic plants are killed by temperatures warmer than 45°C for 10
minutes, and by somewhat cooler temperatures for longer periods (Christy and Sharitz 1980). Despite this
fact, and the fact that macrophyte species richness may be positively correlated with temperature across
broad geographic regions, temperature in itself is probably not a major factor governing the distribution of
herbaceous wetland plants (Pip 1989).
53
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Changes in community composition as a result of thermal alteration begin with changes in the germinations,
growth, and survival of individual species. For example, the introduction over one year of continuously
discharged heated water into a Wisconsin marsh resulted in failed shoot emergence, spring emergence
instead of fall emergence, fewer number of shoots, and greater height of shoots in the sedge, Carex lacustris
(Bedford 1977). Seedling survivorship of one common floodplain species, Ludwigia leptocarpa, was reduced
at 42°C. (Christy and Sharitz 1980). Seedling germination of this species did not vary significantly over the
range 22-42°C. Cattail, Tvpha latifolia. was killed as the probable result of heat-induced depletion of non-
structural carbohydrates in its underground storage organs. This cattail may grow best at a water
temperature of 30°C, but survival is poor at 35°C and seed germination requires temperatures of 13-24°C
(Jones et al. 1979). The common reed (Phragmites communis) may grow best when temperatures fluctuate
within the 20-30°C range (Haslam 1973), and reed canarygrass (Phalaris) may grow best at about 25°C
(McWilliam et al. 1969). For most species, these values vary by genetic population, life stage, duration of
exposure, day length, light intensity, and other factors.
Table 8. Examples of Aquatic Plants That May Indicate Reduced Light Penetration Due to Greater
Turbidity or Shade.
From Davis and Brinson (1980) and other sources. Note that these species may occur as well in wetlands
that are NOT turbid, although usually in smaller proportion relative to other species.
Alisma plantago-aquatica
Ceratophvllum demersum
Eichhornia crassipes
Elodea canadensis
Heteranthera dubia
Hvdrilla verticillata
Lemna minor
Myriophyllum spicatum
Najas flcxilis
Najas guadalupensis
Najas minor
Nuphar lutes
Potamogeton crispus
Potamogeton pectinatus
Potamogeton perfoliatus var. bupleuroides
Potamogeton pusillus
Potamogeton richardsoni
Riccia fluitans
Ricciocarpus natans
Spirodela polvrhiza
Vallisneria americana
Zannichellia palustris
Dehydration. Deviations of seasonal and annual hydrologic cycles from their "normal" regime (including
stabilization of usually fluctuating regimes) can profoundly affect structure of herbaceous wetland plant
communities, perhaps even more so than the actual magnitude of the deviation (Hartog et al. 1989,
Zimmerman 1988). In some cases, community changes reflect the "intermediate disturbance" hypothesis,
wherein "moderate" deviations from "normal" conditions increase community diversity. For example, in
54
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Okefenokee Swamp in Georgia, Greening and Gerritsen (1987) found greater species diversity and variation
in biomass at a site where drawdown was occasional and less predictable than at more predictable sites.
Many herbaceous plant communities, particularly those with rigid stems (e.g., cat-tail, common reed) can
endure (and may even require) periods of a few hours or days of occasional dehydration without changing.
Even a few non-rigid species can survive two or more weeks of exposure, e.g., water milfoil (Myriophvllum
spicaturn). bladderwort (Utricularia gibba), duckweed (Lcmna minor), pondweed (Potamogeton pectinatus).
and Ceratophyllum demersum (e.g., Cooke 1980).
However, if dehydrated shorelines subside (collapse) or complete water level drawdown is sustained over
many days (particularly if it occurs during the growing season and results in desaturation of sediments)
dehydration can trigger significant changes in wetland community structure. This is largely due to the
increased availability of nutrients as sediments become desaturated and oxidized, and partly due to enhanced
germination of seeds of wetland plants that have lain dormant for years in sediments.
In wetlands that are strongly influenced by ground water discharge, erect vegetation may be less vulnerable
to effects of drawdown, because sediments are less likely to become totally dewatered during intentional
drawdown (Cooke 1980). Effects are likely to be most severe when drawdown occurs during extremes of
heat or cold.
In the short-term, complete drawdown often shifts the balance of community structure in favor of emergent
and woody species, and away from submersed species. In the Southeast, aggressive aquatic plants such as
alligatorweed (Alternanthcra philoxeroides) and naiad (Najas flexilis) can increase following partial
drawdown, while muskgrass (Chara vulgaris), water lily (Nuphar spp.), and water hyacinth (Eichhornia
crassipes) can decrease (Holcomb and Wegener 1971, Lantz et al. 1964). In prairie potholes, complete
water loss year after year results in reduced richness even of herbaceous plants, with Carex and Polygonum
generally becoming dominant (Driver 1977). However, partial drawdown, particularly if it occurs for short
periods, may greatly increase macrophyte biomass and growth, due to enhanced nutrient and light availability
that otherwise limit submersed species (Wegener et al. 1974).. In Minnesota peatlands, artificial drainage
resulted in increased dominance of the sedge Carex lasiocarpa (Glaser et al. 1981).
In Indiana, woolgrass (Scirpus cvperinus) was believed to indicate dessication and related disturbance of
former wetlands (Wilcox et al. 1985), as was the sedge, Carex antherodes. in central Canada (Millar 1973).
Woolgrass, along with reed canary-grass (Phalaris) tolerated severe water level drawdown in a New York
reservoir (Burt 1988). In temporarily drained wetlands, Mallik and Wein (1986) found that Tvpha (cat-
tail), Calamagrostis canadensis and Brachvthecium salebrosum had highest cover values. Cover and stem
density of Tvpha increased after draining, while plant height and stem diameter decreased, compared to a
flooded area. Tvpha may not be a good indicator of wetland dehydration, however, as the same study
showed that on the flooded area, Tvpha. Sphagnum squarrosum (a moss) and Pellia epiphvlla had the
highest cover values.
In a literature review on the effects of lake drawdown for control of macrophytes in Wisconsin eutrophic
lakes, Cooke (1980) surmised that only three species-Brasenia schreberi (a water shield), Hvdrochloa
carolinensis. and Potamogeton robbinsii (a pondweed) always decline following temporary drawdown, and
Nuphar spp. and Mvriophvllum spp. often decline following drawdown. Species that appear always to
increase following temporary drawdown include Alternanthera philoxeroides (alligator weed), Lemna minor
(duckweed), Leersia oxvzoides (cutgrass), and Najas flexilis.
In southern Florida, drainage of wet prairies and cypress domes results in increased sawgrass, broomsedge
(Andropogon). common reed (Phragmites communist maidencane (Amphicarpum). chainfern (Woodwardia),
and many graminoids and shrubs; in deeper waters of wetlands, cattail (Tvpha) may increase (Alexander and
Crook 1974, Rochow 1983, Rochow and Lopez 1984, Worth 1983, Atkins 1981).
55
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A wealth of qualitative information about hydrologic tolerances of plants has been compiled for the U.S.
Fish and Wildlife Service's "National List of Plant Species that Occur in Wetlands" (Reed 1988). This
publically-available database classifies all U.S. wetland plants according to their fidelity to wet environments,
i.e., obligate (nearly always in wetlands) or facultative (usually or sometimes in wetlands). As one might
imagine, the obligate taxa in general tend to be less tolerant of desiccation than the facultative taxa listed
in that database. Information on hydric preferences of species might be numerically summarized using the
index of Michener (1983).
The U.S. Fish and Wildlife Service, through the National Ecology Research Center in Fort Collins,
Colorado, has also compiled information on "moist soil management techniques." Use of proposed models
will allow users to predict the effect of water level changes on herbaceous wetland plants, or perhaps
conversely, what the presence of particular plants suggest about prior hydrologic regimes.
Drawdown of wetland water levels in some regions results in increased susceptibility to fires, which in turn
can trigger significant changes in wetland chemistry and vegetation.
Inundation/impoundment. Effects of inundation on emergent herbaceous species are extensively compiled
in Fredrickson and Taylor (1982), Knighton (1985), and Whitlow and Harris (1979). Herbaceous species
of the prairie pothole region are classified according to 12 life history types, related to flooding regime, by
van der Valk (1981). A few additional studies that have examined inundation effects on herbaceous plants
are summarized here.
Increased water levels in aquatic bed (submersed and floating-leaved plant) wetlands appear to have little
effect in some instances (Davis and Brinson 1980). However, water level increases in other instances may
result in increased wave action and initially greater turbidity, which is detrimental to many aquatic plants.
Addition of permanent open water to a non-permanently flooded, emergent wetland increases the
opportunity for invasion by many submersed and floating-leaved species, and generally results in an increase
in on-site species richness. Normally aggressive, perennial emergents such as purple loosestrife, cat-tail,
common reed, and water hyacinth may be reduced or eliminated along with less aggressove species as
flooding increases.
Although species richness of an entire non-permanently flooded wetland sometimes declines for a few years
after flooding becomes permanent (Sjoberg and Danell 1983), overall community richness may change only
slightly in the long term, and only the position of the submersed, emergent, and meadow zones may shift
(Harris and Marshall 1963, van der Valk and Davis 1976). Such zonal shifts serve as indications of long-
term water level change within a wetland (Bolts and Cowell 1988). They occur as dormant seeds of wetland
plants, which require specific water depths for germination (Moore and Keddy 1988), germinate along the
upland boundary as a result of the new flooding. At the same time, down-gradient species subjected to
inundation may be lost as a result of suffocation, build-up of compounds toxic to roots, and alteration of
physical conditions, e.g., erosion and scour. Floating-mat vegetation may survive.
The effects of flooding also will depend on flooding depth, frequency, duration, dominant plant species,
sediment type, water velocity, and other factors. Short periods of flooding (days or weeks) were reported
in Wisconsin to have no effect on wetland community composition (e.g., Nichols et al. 1989). Assemblages
of herbaceous wetland plants can be a more sensitive indicator of water level change than assemblages of
woody wetland plants, which respond more slowly (Paratley and Fahey 1986).
Deviations of seasonal and annual hydrologic cycles from their "normal" regime in wetlands, and particularly,
the elimination of seasonal fluctuations, may reduce overall plant species richness. Native and perennial
species, particularly grasses and sedges, may be replaced by taxa that are more aggressive, exotic, clonal,
56
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and/or annuals. Depending on the initial water level, these commonly include cat-tails, bulrush,
pickerelweed (Sagittaria) and pondweed (Pontederia)(e.g., Bolts and Cowell 1988, Mclntyre et al. 1988). In
some cases, community changes reflect the "intermediate disturbance" hypothesis, wherein "moderate"
deviations from "normal" annual hydrologic conditions increase community diversity.
In wetlands along Lake Erie in Ohio, diking of a marsh increased the dominance of liverwort (Riccia spp.),
duckweeds (Lemna minor. Spirodela polvrhizat. coontail (Ceratophvllum demersumX water milfoil
(Mvriophvllum speciatum'). pondweeds, and bladderwort (Utricularia vulgaris^ (Farney and Bookhout 1982).
Permanent inundation of other marshes has decreased the number of plant species and the dominance of
Carex spp. (Farney and Bookhout 1982, Sjoberg and Danell 1983).
In Colorado, subalpine wetlands flooded for longer than about 30-45 days during the growing season had
fewer emergent plant species than those inundated for shorter periods (Cooper and Emerick 1989). In
Sweden, lakeshore wetlands with less than 40-60 days of flooding during the growing season had maximum
richness and cover of macrophytes, in contrast to those flooded for longer periods (Nilsson and Keddy 1988).
Cattails generally tolerate deeper water than most rushes (Lathwell et al. 1973), which tolerate deeper water
than most sedges (van der Valk and Davis 1976). Cattail (Tvpha) can dominate wetlands with water depths
generally greater than 15 cm for 6 to 12 months (Mall 1969). The common reed (Phragmites) typically
occurs in water depths of 0 to 1.5 meters (Haslam 1970, Spence 1982). For most species, these values vary
by genetic population, life stage, duration of exposure, water chemistry, and other factors. The horse-tail,
Equisetum fluviatile. appeared to be the most tolerant of several emergent species to modest increases in
water depth (Sjoberg and Danell 1983). Depth-to-water-table preferences of many peatland species are given
by Jeglum (1971).
As noted above, a wealth of qualitative information about hydrologic tolerances of plants is represented by
the U.S. Fish and Wildlife Service's "National List of Plant Species that Occur in Wetlands" (Reed 1988),
and in their models currently being developed for moist-soil management and in-stream flow management.
Fragmentation of Habitat. We found no explicit information on macrophyte community response to
fragmentation of regional wetland resources. Biomass and cover of submersed wetland plants generally
decreases with increasing lake size, while the converse is true for emergent species (Duarte et al. 1986). In
England, Helliwell (1983) found greater macrophyte species richness in larger wetlands, but a large amount
of the variation could be attributed to other factors. Larger wetlands also may have greater macrophyte
richness because they tend to be visited more often than smaller wetlands by birds and other animals capable
of introducing new plants (Pip 1987). However, regionally rarer species often occur in small wetlands with
unique physical and chemical environments (Moore et al. 1989). One can surmise that species with broad
environmental tolerances and that disperse easily (e.g., Godwin 1923) might be least affected as wetlands
become more isolated from one another.
Other human presence. The bottoms of Sierra lakes in California with higher levels of human visitation
had more coverage with rooted macrophytes (Isoetes, Anacharis, Nitella) and bottom algae (Rhizoclonium)
than those less frequently visited; this phenomenon was evident even in lakes where use had been restricted
for 10 to 20 years (Taylor and Erman 1979). Trampling and other impacts on riparian wetland vegetation
are documented in Cole and Marion (1988) and in studies of wetland buffer zones in New Jersey.
In wetlands of "developed" watersheds in New Jersey, uncharacteristic herbaceous species replaced endemic
ones, and herbs and vines were more prevalent than in "undeveloped" watersheds (Ehrenfeld 1983, Schneider
and Ehrenfeld 1987). Common reed (Phragmites) typically characterizes many disturbed, nutrient-poor
wetlands (Haslam 1971), as do woolgrass and soft rush (Juncus effusus^) in Pennsylvania (Hepp 1987). In
Ohio, increased eutrophication, warming, and turbidity were implicated in the decline of Najas gracillima
and N. flexilis. and an increase in N. marina. N. minor, and N. guadalupensis over a 70-year period (Wentz
and Stuckey 1971).
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Channels with heavy shipping traffic connecting the Great Lakes had less dense beds of submersed
macrophytes than in channels with less ship traffic. Dominance shifted from Myriophyllum spicatum. Elodea
canadensis and Heteranthera dubia in the relatively undisturbed channel to Characeae, Potamogeton
richardsonii. and Najas flexilis in the disturbed channel (Schloesser and Manny 1989). Similar reductions
of macrophyte biomass from recreational boating have been found in Europe, as compiled by Liddle and
Scorgie (1980) and Murphy and Eaton (1983).
6.2 SAMPLING METHODS AND EQUIPMENT
Factors that could be important to standardize (if possible) among wetlands when monitoring community
structure of macrophytes include:
age of wetland (successional status), light penetration (particularly for submersed species),
water or saturation depth, conductivity and baseline chemistry of waters and sediments,
current velocity, abundance of herbivores (particularly muskrat, geese, grazing cattle,
crayfish), stream order or ratio of discharge to watershed size (riverine wetlands), sediment
type, existence of any prior planting programs, and the duration, frequency, and seasonal
timing of regular inundation, as well as time elapsed since the last severe inundation,
drought, or fire.
References that provide more detailed guidance on sampling herbaceous wetland vegetation include
Fredrickson and Reid 1988a, Moore and Chapman 1986, Mueller-Dombois and Ellenberg 1974, Murkin
and Murkin 1989, Phillips 1959, Schwoerbel 1970, and Woods 1975. Guidelines for collecting specimens of
aquatic plants for preservation are given by Britton and Greeson (1988), and Haynes (1984). References
useful in data analysis are listed in Chapter 3.
If wetlands can be sampled only once, mid-growing season is usually the recommended time. However,
many plants are apparent and/or identifiable only for a few weeks of the growing season. Thus, if the aim
is to quantify community composition accurately, repetitive visits that account for the diverse phenologies
of wetland species should be implemented. Ideally, annual visits could be timed to coincide with year-
specific weather conditions, rather than calendar date. For example, Grigal (1985), who sampled vegetation
over three years, did field work at slightly different times each year. This increased the chances of finding
species in flower, making identification easier. In northern bogs, early fall may be a desirable sampling time,
e.g., Wilcox (1986) sampled vegetation in a bog in the first week of September because the maximum
number of species was identifiable at that time in Indiana. Optimal sampling times vary geographically.
Whenever possible, plants should be identified in the field rather than collected. Trampling of herbaceous
vegetation and compaction of saturated soils during even a single site visit can induce community changes
detectable in subsequent visits. Thus, field crews should be as small as possible and follow the same path
in and out of a wetland. In riverine and lacustrine wetlands, underwater SCUBA transects can be run
(Schmid 1965).
Equipment commonly used to destructively sample herbaceous wetland vegetation (especially submersed
species) includes dredges, oyster tongs, plant grappling hooks, steel garden rakes, and similar devices (Britton
and Greeson 1988). Equipment designed specifically for sampling herbaceous macrophytes is described by
Dromgoole and Brown 1976, Macan 1949, Satake (1987), 1977, Wood 1975, and others.
Types of commonly-collected data on herbaceous wetland communities include species per plot and percent
cover. Less often, total stem count per m^, and stem count per species per plot are determined. Stem
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counts are usually made only of species perceived to be dominant, and may possibly include a few
subdominants.
Herbaceous plant community composition is typically quantified using belt transects or replicate quadrats.
Transects and quadrats can be used in all wetland types, but may give less reliable data where vegetation
is submerged or otherwise difficult to access. Sampling schemes involving transects or quadrats can yield
data that is particularly amenable to statistical analysis. The number, size, and spacing of transects and
quadrats in a wetland depends primarily on wetland size, shape, internal heterogeneity (e.g., as perceived
during an initial reconnaissance visit and/or from aerial photographs), and the statistical power one wishes
to have in detecting spatial change in various community metrics. Larger wetlands require more transects
or quadrats, usually spaced farther apart, to accurately characterize overall community composition. More
linear wetlands (e.g., narrow fringe marshes along lakes) may require more tightly spaced sampling points,
as may ecotone areas along transects. Sampling stations along transects are usually situated at even
intervals, and quadrats can be placed evenly (e.g., in a grid), randomly, or clustered. Random placement of
plots for the purpose of statistically characterizing a wetland is usually prohibitively expensive, due to the
extreme spatial variability of most wetlands (Durham et al. 1985). Plots or transect lines are often marked
for future relocation.
In most studies of herbaceous wetlands, investigators have located transects or quadrats in a manner that
parallels or spans a likely stressor gradient (e.g., parallel to basin gradient, perpendicular to flow, or parallel
to flow path of discharge from a chemical outfall). If the stressor is a point source, the transect should
be long enough to allow complete definition of gradients in response to the stressor. Thirty meters was not
far enough to show distance effects of wastewater disposal in a bog/marsh system studied by Kadlec and
Hammer (1980). To avoid problems of treatment effects spilling over into control plots, Loveland and
Ungar (1983) used a randomized block design of 0.25-m^ plots in each of three vegetation zones. Each
zone contained five replications of each block; for controls, five plots were randomly spaced in each zone.
In another study of artificial enrichment (Duarte and Kalff 1988), plots were not isolated; fertilized plots
were 9 m apart and control plots were 3 m from the corresponding treated plots.
Where multiple, non-overlapping gradients are perceived, transects may be located perpendicular to, or at
other appropriate angles to, each other. The number of transects and quadrats in particular cover types
within the wetland may also be designed to be proportional to the overall coverage of these cover types.
Transects used for herbaceous community monitoring have ranged upwards from about 100 meters in length
(depending on wetland size and shape); quadrats have ranged upwards from 0.05 m^, and may be
rectangular, square, or circular. The minimum effective size can be determined statistically or by plotting
of initial data, as described in section 3.3. Based on statistical analysis of dozens of published studies of
submersed vegetation, Downing and Anderson (1985) suggested it is better to use small quadrats with great
replication than large quadrats with little replication, especially where vegetation stands are not dense.
However, they suggested cautious interpretation of this recommendation if small quadrats are being placed
in dense macrophyte beds. From a study of 18 Canadian lakes, France (1988) determined that at least 21
replicate samples are required to achieve estimates within 20 percent of the mean biomass, using a sampler
with an area of 45.6 cm^. A different number of replicates would probably be required if determination of
richness, rather than biomass, was the objective.
Variable-sized plots also can be used, where plot size depends on life form of vegetation present in
proximity to each particular point in the wetland (e.g., Mader et al. 1988). Nested frequency quadrats, in
which only the number of times a species is present is recorded-have also been used (e.g., Frenkel and
Franklin 1987). These have the advantage of easily data collection, objectivity, and no need to relocate
plots, but interpretation depends on plot size and shape and spatial distribution of species, and this
approach cannot easily be used to quantify spatial patterns, cover, or biomass.
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6.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS
In general, the parameters most often measured in studies of herbaceous wetlands are "percent cover" and
"biomass (standing crop)." Measures of community structure of submersed or emergent aquatic communities
have not been uniformly collected from a series of statistically representative wetlands in any region of the
country. Thus, it is currently impossible to state what are "normal" levels for descriptors of community
structure such as seasonal plant density or species richness, and their temporal and spatial variability, in any
type of herbaceous wetland.
Perhaps the closest approximation of a broad-scale effort is that of Duarte et al. (1986). They looked at
just one parameter-lacustrine macrophyte biomass--and examined causes of local and regional variability.
From their resultant equations, expected ("nominal") levels of biomass of both emergent and submersed
macrophytes in lakes might be estimated. Approximate data describing lake area, depth, slope, and a few
other simple parameters are needed to run the calculations.
A few, usually localized, studies of inland wetlands have published Shannon diversity index values.
example:
For
State type
SC Pfo
MA Lab
GA P
WA P
N
? 3.58
120 1.47
? 0.83
>300 1.48
min.value
max, value
3.78
3.71
1.54
2.65
citation
Sharitz et al. 1974
Burk 1977
Greening & Gerritsen 1987
Meehan-Martin & Swanson
1988,1989
Species richness has been reported by many studies, but is not always standardized per unit effort or per
unit area as it should. Examples include:
State type
SC Pfo
MA Lab
IA Pern
NY Pern
NJ P
N
650
120
28
90
min.value
1.2/0.25m2
SE=0.4
2.3/0.25m2
SE=0.21
3.2/m2
SD=0.5
7.7/m2
18 26.8/600m
SE=2.7
max. value
9.6/0.25m2
SE=0.8
8.3/0.25m2
SE=0.56
8.3/m2
SD=2.0
12.2/m2
41.4/600m
SE=5.7
citation
Dunn and
Sharitz 1987
Burk 1977
van der Valk
& Davis 1976
Paratley &
Fahey 1986
Schneider
& Ehrenfeld 1987
In addition, Ehrenfeld (1983) summarized her data as follows:
Mean species richness per 600m2 (N=16):
Disturbed sites = 33.9 +_ 2.17; range 17-47
Pristine sites = 27.8 +. 2.24; range 13-44
A similar study by Morgan and Philipp (1986) reported the following values for coefficient of similarity
(based on 12 plots per stream, each plot 600m2 in area):
between polluted and unpolluted streams = 16%
among polluted streams = 28%
among unpolluted streams = 26%
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Many other studies, although not publishing or summarizing in a useful form their statistics on community
structure, have compared herbaceous vegetation among wetlands in a region (i.e., spatial variation). Some
of the more systematic or extensive quantitative comparisons include:
Albert et al. 1987, Bunfield and Evans 1982, Canfield et al. 1983, Canfield and Duarte 1988, Duarte
et al. 1986, Ehrenfeld 1986, Henebry et al. 1981, Pip 1979, 1987a,b, Sheath et al. 1986, Stewart and
Kantrud 1972, and Terry and Tanner 1984.
One of the more geographically extensive ongoing studies is a survey of vegetation in a large number of
Great Lakes wetlands in Michigan (Albert et al. 1987). Survey locations are shown in Appendix B.
Another extensive and long-term survey of wetland vegetation is being conducted as part of monitoring
studies for the ELF military radiocommunications facility (Blake et al. 1987).
A significant number of studies have compared long-term change (but seldom year-to-year variation) in plant
community structure in wetlands, in some cases by use of paleoecological techniques. Changes in most cases
have not been quantitatively linked with particular stressors. These chronological studies include:
Baumann et al. 1974, Brown 1987, Bumby 1977, Burk 1977, Burton and King 1983, Harris et al.
1981, Hale and Miller 1978, Kadlec 1979, Niemier and Hubert 1986, Schwintzer and Williams 1974,
Southwick and Pine 1975, Stuckey 1971, van der Valk and Davis 1979, and Wentz and Stuckey 1971.
Qualitative data on community structure of inland herbaceous wetlands appears to be most available for
Florida, Minnesota-Michigan-Wisconsin, Louisiana, New York, and North Dakota. Apparently the least
amounts of such data are for playa wetlands, and for herbaceous wetlands in the Appalachians, southern
Great Plains, and Southwest. Information is most available on impacts of hydrologic alteration and
nutrients, and least on impacts of partial burial, contaminant toxicity, and habitat fragmentation.
Reasonably complete, qualitative lists of "expected" wetland herbaceous plants are available for most regions
through the USFWS's "National List of Wetland Plants" database, and databases of The Nature Conservancy.
Quantitative data are generally most available for vascular emergent species, and less common for submersed
plants and mosses. Limited qualitative information may also be available by wetland type from the
"community profile" publication series of the USFWS (Appendix C).
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7.0 WOODED WETLAND VEGETATION
Discussions in this section focus on trees and shrubs that normally characterize wetlands. In many cases,
community structure of wood vegetation is less effective as an indicator of short-term anthropogenic stress
than is structure of herbaceous plant communities. This is because species composition of wooded wetlands
responds slowly to stress, and is suitable mainly as an integrator of conditions occurring over many months
and years.
7.1 USE AS INDICATORS
Enrichment/eutrophication. Changes in community composition of wooded wetlands were attributed to
increased nutrients in a Michigan wetland exposed to wastewater, by Kadlec and Hammer (1980), but most
studies have been too short to detect significant change. Moreover, changes in nutrient concentration are
often associated with changes in hydroperiod, and distinguishing the effects of the two can be difficult.
Effects of nutrient increases have more often been detected at the level of the individual plant (e.g., growth,
foliage and root nutrient concentrations) than at the community level. However, effects at the individual-
plant level can often be eventually translated into effects on community composition. Community-level
measurements of woody vegetation may be a poorer indicator of eutrophication than are algal or herbaceous
plant communities, which respond more quickly.
Organic loading/reduced DO. Existing literature often does not adequately distinguish the effects on woody
plants of organic loading/reduced DO, from the effects of nutrients (discussed above) or inundation
(discussed below). For example, in one Minnesota wetland experimentally exposed to wastewater, tree
mortality could have resulted from hydrologic changes or methodological variation, and so was not attributed
specifically to the effluent (Schimpf 1989).
Feedlot effluent entering an Illinois swamp caused increases in species richness, invasion by new species, and
changes in species dominance (Pinkowski et al. 1985). Straub (1984) looked for changes in growth rates in
isolated swamps with added fertilizer or wastewater, but after five years found none. Increased tree growth
has been noted in Florida wetlands exposed to secondarily treated effluent, but untreated effluent appears
to be detrimental (Brown and van Peer 1989, Lemlich and Ewel 1984). Florida regulations for treated
wastewater discharges to wetlands specify that "the importance value of any of the dominant plant species
(excluding some exotics) occupying the canopy or subcanopy shall not be reduced by more than 50 percent
at any monitoring station, or 25 percent overall in the wetland." Exceptions may be allowed if changes can
be attributed to catastrophic natural events such as hurricanes or fire. Dominant plant species are defined
as those that have a total relative importance value of at least 90 percent during the baseline monitoring
period (Schwartz 1987).
Contaminant Toxicity. Few if any studies of have been conducted of community-level response of woody
vegetation to contaminants in wetlands. Shallow-rooted species are generally believed to be more sensitive
to contaminants than deep-rooted species, due to their greater exposure to waterborne contaminants
(Sheehan 1984). A four-year study of the response of wetland species to an oil spill in a Massachusetts
inland wetland (Burk 1977) reported post-spill absence of red maple (Acer rubrumX and no effect or
increase in sugar maple fAcer saccharinum) and wild grape (Vitis labrusca). Additional toxicological
information may be available through EPA's PHYTOTOX database (Royce et al. 1984).
Acidification. There are apparently no community-level studies of effects on wooded vegetation specifically
in wetlands.
Salinization. In general, woody plants are more sensitive than herbaceous species because they are usually
unable to release salts back into the soil, and must therefore rid themselves of it through leaf loss or dying
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branches. However, we found very little explicit information on woody wetland community response to
salinization. An experimental study in Florida demonstrated stress to individual trees from chloride-
enriched water (Richardson et al. 1983). Adverse impacts of road salt on forest communities (probably
including some wetland species) have been frequently demonstrated. Some tolerance data also may be
available from studies of freshwater tidal wetlands. In North Carolina, Brinson et al. (1985) reported
reduced tree basal area and density, and greater litterfall, in forested wetlands temporarily exposed to waters
of higher salinity.
Sedimentation/burial. Trees (and especially seedlings) are killed when trunks or stems are partially buried
or sediment deposition is sufficient to cut off root oxygen exchange (e.g., Eichholz et al. 1979, Kennedy
1970, Harms et al. 1980, Maki et al. 1980). Floodplain trees in Florida were killed by 0.8 m or more of fill,
and tree vigor was reduced by only 0.04 to 0.12 m of fill (Clewell and McAninch 1977). Also, where
sedimentation is severe, the frequency and duration of inundation may change, causing shifts in community
structure (see below). Siltation can also reduce stem height and diameter growth (Kennedy 1970), thus
altering competition and ultimately, community structure. Relatively sediment-tolerant species include
eastern cottonwood, baldcypress, water tupelo and black willow (Broadfoot 1973).
Where sedimentation creates shoals in rivers or lakes, these sometimes provide additional substrate for
establishment or expansion of wooded wetlands. Moderate amounts of sediment may also have a fertilizing
effect.
Turbidity/shade; Vegetation removal. Alteration of the canopy within wooded wetlands may be expected to
trigger long-term shifts in community composition of woody species, particularly shrub species. Shade
tolerances of most woody species are relatively well-known. Logging has an obvious immediate impact on
forested wetlands, but the long-term effects on community structure are poorly known and probably
dependent upon initial state and the specific silvicultural procedures used. Repeated "high-grading" (i.e.,
removal of largest trees of the most valuable species) results in high-density, low-biomass stands of shade-
tolerant species such as elm, maple, and willow. Grazing also affects community composition, and woody
plants differ in their palatability and thus sensitivity to grazing. Usually, evergreens (particularly cedar) and
thorny species are less grazed than other deciduous species, but utilization also depends on local availability.
Thermal alteration. Decreases in plant species richness, basal area, and stem density have occurred in South
Carolina wooded wetlands as a result of warmed waters (Scott et al. 1985).
Dehydration. Many woody plant communities in wetlands require the absence of surface water, while others
require its presence. In the latter case, many of the species comprising such communities can endure brief
periods (e.g., a few hours) of occasional drawdown without changing, so long as sediments remain saturated,
and many such communities change little despite weeks, months, or years without surface water (e.g., Parker
and Schneider 1975). However, in other wetland communities adapted to flooding, if drawdown is sustained
over many days (particularly if it occurs during the growing season and results in desaturation of sediments),
dehydration can trigger significant changes in soil chemistry and in wooded wetland community structure.
This is largely due to the increased availability of nutrients as sediments become desaturated and oxidized,
and partly due to enhanced germination of seeds of woody plants that have lain dormant for years in
sediments.
In the short-term, complete drawdown of flood-adapted communities often shifts the balance of community
structure in favor of woody species, and away from submersed and emergent species. For example, drainage
and groundwater withdrawals near wet prairie and cypress wetlands in Florida have resulted in invasion of
these areas by willows (especially in burned and logged areas), maidencane, Brazilian pepper, wax myrtle,
dahoon holly, gallberry, saltbush, buttonbush, slash pine, red bay, water oak, cabbage palm, and red maple
(Alexander and Crook 1974, Carlson 1982, Duever et al. 1979, Lowe et al. 1984, Richardson 1977). In
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southwestern riparian wetlands, flow regime alteration by dams, and its effect on sedimentation, has resulted
in replacement of native riparian woodlands with non-native salt cedar (Tamarix spp.) (Brady 1985, Stevens
1989). The exact successional pattern will depend on initial state and other factors.
Density and species richness of woody species may increase following drainage and/or across a spatial
gradient of decreasing inundation duration (e.g., Thibodeau and Nickerson 1985, Maki et al. 1980). From
their limited data, Taylor and Davilla (1986) concluded that the effects of river flow diversion (dehydration)
on California riparian communities were more distinguishable in the smallest and largest streams (orders
1 and 4) than in streams of intermediate size (orders 2 and 3).
The presence of seasonally elevated water levels can sometimes be inferred by water marks and drift lines
on vegetation, presence of adventitious root "knees", signs of current scouring, subsidence and bank collapse,
and other secondary features. If these are found in a wetland whose water levels currently remain low
throughout the year, then some evidence is provided that dehydration (e.g., by flow diversion) has occurred.
Also, several investigators have sought to compile hydrologic tolerance or preference data into quantitative
metrics. For example, the Corps of Engineers has quantified much of the tolerance data for woody plants
in its "Flood Tolerance Index" (FTI), which is based on a weighting of cover estimates according to flood
tolerance of the species (Theriot and Sanders 1986). A conceptually similar index is described by
Wentworth et al. (1988) and tested by Carter et al. (1988). Either index might be tested to determine its
potential for use as an indicator of persistant dehydration, i.e., based on the proportion of facultative species
found in an area and reflected by the index value.
Inundation/impoundment. Although the existence of many woody wetland communities is absolutely
dependent upon inundation, deviations of seasonal and annual hydrologic cycles from their "normal" regime
(including stabilization of usually fluctuating regimes) can profoundly affect structure of woody plant
communities in wetlands.
Presence of surface water is generally much more detrimental to woody seedling survival than is simple soil
saturation (Hosner 1960). Flooding later in the growing season, when seedlings have leafed out, has the
potential for greater impacts than earlier floods (e.g., Scott et al. 1985). Also, stagnant, deepwater flooded
conditions may be more detrimental than aerated conditions, e.g., where water is shallow and flowing,
organic loading is light, and water levels fluctuate according to a natural seasonal pattern (Teskey and
Hinckley 1977).
Species richness of woody wetland plants generally decreases with increasing flood duration (Brown and
Giese 1988, Klimas et al. 1981). Several instances have been reported where frequent flooding has
selectively removed smaller trees and shrubs (e.g., Ehrenfeld 1986, Maki et al. 1980, Noble and Murphy
1975) and may favor emergent vascular plants and mosses (Jeglum 1975). In western riparian areas, shallow-
rooted woody species may be more sensitive to flooding than species with deep tap roots (Stevens and
Waring 1985). In an eastern floodplain swamp, mortality was lowest in trees greater than 38 cm dbh
(diameter) and greatest in trees less than 13 cm dbh (Harms et al. 1980).
As little as 3 days of flooding during the growing season can result in loss of some woody vegetation
through suffocation and compounds toxic to roots (Boelter and Close 1974, Harms et al. 1980, Stoeckel
1967, Jeglum 1975, Davis and Humphrys 1977, Keddy 1989, Maki et al. 1980, Southern Forest Experiment
Station 1958), or through alteration of physical conditions, e.g., erosion and scour. Although some species
survive at least 3 years of continuous flooding (Green 1947), most cannot survive growing-season inundation
for more than a year or two (Broadfoot and Williston 1973).
A wealth of other information about hydrologic tolerances of woody plants has been compiled in several
reports, e.g.:
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Burton 1984, Whitlow and Harris 1979, Hook 1984, Teskey and Hinckley 1977a,b, 1978a,b,c,d, 1980,
Walters et al. 1980.
The U.S. Fish and Wildlife Service's "National List of Wetland Plants" and its FORFLO model (Brody and
Pendleton 1987) also compiled substantial databases on hydrologic tolerances of plants in the course of their
development. The FORFLO model quantitatively predicts wooded wetland community change, given data
on expected hydrologic change. The USFWS and others (e.g., Harris et al. 1985, Kondolf et al. 1987) are
currently developing methods for relating hydrologic tolerances of woody plants to instream flows. Intensive,
site-specific procedures for quantifying the tolerated days, depths, and seasons of flooding in forested
wetlands are demonstrated by Grondin and Couillard (1988).
Individual tree growth may also be affected by inundation. Deviations from normal flooding cycles can
reduce tree growth (Malecki et al. 1983). However, temporary flooding by rivers may fertilize floodplain
trees, increasing growth (Mitsch et al. 1979). In some cases the basal increment can be larger in the
remaining trees as compared to unflooded areas. It should not be assumed that basal growth is a good
indicator of flooding stress or survival (Franklin and Frenkel 1987).
Fragmentation of habitat. We found no explicit information on forested wetland community response to
fragmentation of regional wetland resources. One can surmise that as the distance between wetlands with
seed sources becomes greater and dispersal corridors become hydrologically disrupted, species with narrow
environmental tolerances and which do not disperse easily might be most affected. This assumption was
used by Hanson et al. (1990), who developed a model which predicted that fragmentation will lead to lower
woody plant diversity in riparian wetlands. Those authors classified several woody species according to their
seed dispersal ability.
Other human presence. In "developed" watersheds, the frequency of characteristic wetland shrub species was
reported to be less than in wetlands in "undeveloped" watersheds (Ehrenfeld 1983).
7.2 SAMPLING METHODS AND EQUIPMENT
Natural factors that could be important to standardize (if possible) among wetlands when monitoring
anthropogenic effects on community structure of woody plant communities include:
age of wetland (successional status), water or saturation depth, sediment type, conductivity
and baseline chemistry of waters and sediments, current velocity, abundance of herbivores
(particularly beaver, grazing cattle), stream order or ratio of discharge to watershed size
(riverine wetlands), and the duration, frequency, and seasonal timing of regular inundation,
as well as time elapsed (years) since the last severe inundation, drought, windstorm, or fire.
Seasonal timing of woody plant sampling is less critical than is seasonal timing for sampling of herbaceous
plants, because most woody plants are present and identifiable throughout the year. Mid-growing season
is usually the recommended time, because of the visibility of seedlings and the relative ease in identifying
species then. However, access to woody plants in wetlands may be best in winter if ice is present.
The reference texts and choices for protocols described Section 6 for herbaceous plants generally apply to
wooded wetlands as well. However, quadrats are usually larger (at least 1 mr, and often over 10 m^) and
transects may be longer. Percent cover is less often determined where woody plant canopies are larger than
1 m^ in diameter. Belt transects and line-intercept methods (Canfield 1941, Mueller-Dombois and Ellenberg
1974) are more frequently employed, and dbh (diameter at breast height) of dominant and subdominant
stems is commonly measured. Working in large tracts of bottomland hardwood wetland, Durham et al.
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(1985) recommended 0.1 ha fixed-area plots for overstory and saplings, with 0.025 ha subplots for sampling
shrubs. The State of Florida's regulations for monitoring of discharge of treated wastewater into wooded
wetlands specify that quadrat size shall be at least 100 tar for canopy vegetation and 50 m^ for subcanopy
vegetation, and that the number of quadrats shall be that number needed to provide 90% certainty of being
within 15% of the mean number of species of the population. Additional guidance for sampling streamside
wetlands is given by Ohmart and Anderson (1986).
7.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS
In general, quantitative community-level data on wooded wetland vegetation have not been uniformly
collected from a series of statistically representative wetlands in any region of the country. Thus, it is
currently impossible to state what are "normal" levels for parameters such as seasonal plant density, species
richness, biomass, or productivity, and their temporal and spatial variability, in any type of wooded wetlands.
Perhaps the closest approximation of such a data set base is the U.S. Forest Service's Forest Inventory and
Assessment database (FIA), and Continuous Forest Inventories (CFI). At least in theory, mean density and
species richness could be calculated by state for each of the forest types that characteristically occur in
wetlands. A large data set describing these metrics also was collected by the U.S. Army Corps of Engineers,
along the lower Mississippi River (Klimas 1988), and another was collected by Jensen et al. (1989) in
western riparian systems.
Data on another community metric-importance value-were presented in some of the soil-vegetation
correlation studies sponsored by the U.S. Fish and Wildlife Service (e.g., Dick-Peddie et al. 1987, Erickson
and Leslie 1987, Hubbard et al. 1988, Nachlinger 1988). In addition, examples of studies of multiple
forested wetlands across a region, that quantify woody biomass, stem density, or basal area, include the
following:
Dale 1984, Ehrenfeld 1986, Faulkner and Patrick n.d., Jones 1981, Klimas 1988, Osterkamp
and Hupp 1984, Reiners 1972, Robertson et al. 1984.
A few published studies have quantified long-term successional changes in community structure of riparian
or other wooded wetland communities, sometimes in the engineering context of reconstructing past flood
histories. Examples include Malecki et al. 1983, Schwintzer and Williams 1974, and studies cited in Hupp
(1988).
Quantitative data on community composition of wooded wetlands appears to be most available for
California, the lower Mississippi basin, and Minnesota-Michigan-Wisconsin; and least for New England
wooded swamps, Pacific Northwest swamps, and Midwestern riparian systems. Information is most available
on impacts of hydrologic alteration, and least on impacts of partial burial, contaminant toxicity, salinization,
and habitat fragmentation.
Reasonably complete, qualitative lists of "expected" wetland woody plants are available for most regions
through the USFWS's "National List of Plant Species that Occur in Wetlands" (Reed 1988) and databases
of The Nature Conservancy. Also, qualitative information may be available by wetland type from the
"community profile" publication series of the USFWS (Appendix C).
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8.0 WETLAND INVERTEBRATE COMMUNITIES
8.1 USE AS INDICATORS
Discussions under the heading "Invertebrates" here include aquatic insects, freshwater crustaceans (e.g.,
amphipods, crayfish), aquatic annelids (e.g., worms), zooplankton, and terrestrial insects (e.g., the butterfly,
bog elfin, and others listed by Niering 1985) that are found predominantly in wetlands.
Enrichment/eutrophication. Wetland invertebrates respond strongly to trophic condition. Abundance
generally increases with increased nutrient concentrations (e.g., Cyr and Downing 1988, Tucker 1958) and
species richness may decrease (Wiederholm and Eriksson 1979) or increase (Tucker 1958). Particular species
assemblages of invertebrates have commonly been reported to be useful indicators of lake trophic state
(Table 9) and may find similar usage in wetlands. These include:
o aquatic worms (Oligochaeta) (Gatter 1986, Milbrink 1978, Lafont 1984, Lauritsen et al. 1985);
o midges (Chironomidae) (Rae 1989, Wiederholm and Eriksson 1979, Winnell and White 1985);
o snails (Gastropoda)(Clarke 1979a); and
o clams (Sphaeriidae)(Clarke 1979b, Klimowicz 1959).
In particular, the ratios of (a) tubificid worms to aquatic insects, (b) the chironomid subfamilies Tanypodinae
and/or Chironomini to the subfamily Orthocladiinae), and/or (c) cladocerans to rotifers, have been reported
to increase with increasingly eutrophic conditions (Ferrington and Crisp 1989, Gatter 1986, Radwan and
Popiolek 1989, Rosenberg et al. 1984). As species shifts occur with increasing eutrophication, chironomid
species richness may decline; however, chironomid biomass and/or abundance increase (Ferrington and Crisp
1989, Johnson and McNeil 1988). Indeed, chironomid emergence was recommended as an efficient indicator
of secondary production in lakes by Welch et al. (1988).
Organic loading/reduced DO. Excessive organic loading of surface waters, including wetlands, is known to
alter community composition (CH2M Hill 1989), usually reduces invertebrate diversity and evenness (e.g.,
Sedana 1987), and sometimes reduces density and biomass (e.g., Hartland-Rowe and Wright 1975, Pezeshik
1987, Schwartz and Gruendling 1985, USEPA 1983). However, density and biomass of benthic invertebrates
in a southern Quebec wastewater wetland was significantly greater than in unexposed wetlands (Belanger and
Couture 1988). Density (Sedana 1987) and richness of invertebrates also increased in an Alabama pond
after a single episodic addition of manure, but after four weeks richness declined to less than in a control
pond (Deutsch 1988). Florida regulations for treated wastewater discharges to wetlands specify that "the
Shannon-Weaver diversity index of benthic macroinvertebrates cannot be reduced below 50 percent of
background levels as measured using standard techniques."
Under moderate loading, attached algae may increase and consequently, herbivorous mayflies and midges
may dominate the community (Jones and Clark 1987). However, if turbidity and hydroperiod conditions
allow submersed or floating-leaved aquatic plants (e.g., Lemna) to out-compete algae, other aquatic
invertebrates may become dominant.
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Table 9. Examples of Aquatic Invertebrates That May Indicate Eutrophic Conditions in Wetlands.
Only Chironomidae are listed. Compiled from the ERAPT database (Dawson and Hellenthal 1986) and the
following references: Harvey and McArdle 1986, Rosenberg et al. (1984), Saether (1975), Strange 1976, and
Walker et al. 1985, Wiederholm and Eriksson (1979). Note that these species may occur as well in wetlands
that are NOT eutrophic, although usually in smaller proportion relative to other species.
Qiironomus atlenuatus
Chironomus crassicaudatus
Chironomus riparius
Chironomus stiematenis
Crvptochironomus blarina
Crvplotendipcs casuarius
Crvptotendipes emorsus
Dicrotendipes incurvus
Dicrotendipcs modestus
Einfeldia natchitocheae
Gryptotendipes barbipes
Glvptotendipes meridionalis
Goeldichironomus holoprasinus
Harnischia bovdi
Harnischia galealor
Kiefferulus dux
Leplochironomus nierovittatus
Paeastiella orophila
Parachironomus directus
Parachironomus monochromus
Parachironomus scheideri
Paralauterborniella elachista
Paralauterborniella subcincta
Pedionomus beckae
Polvpedilum dieitifer
Polvpedilum illinoense
Polvpedilum trieonum
Tribelos quadripunctatus
Pseudochironomus richardsoni
Calopsectra dendvi
Calopsectra xantha
Micropsectra dubia
Tanvtarsus bucklevi
Tanvtarsus recens
Cricotopus bicinctus
Cricotopus svlvestris
Psectrocladius dvari
Coelotanvpus concinnus
Coelotanvpus tricolor
Ablabesmvia aequifasciata
Ablabesmvia annulata
Ablabesmvia basalis
Ablabesmvia hauberi
Ablabesmvia mallochi
Ablabesmvia ornata
Ablabesmvia peleensis
Ablabesmvia rhamphe
Guttipelopia currant
Labrundinia johannseni
Labrundinia pilosella
Monopelopia boliekae
Procladius bellus
Procladius denticulatus
Tanvpus carinatus
Tanvpus punctipennis
Chironomus carus
Chironomus plumosus
Chironomus staeeeri
Chironomus tentans
CTVptochironomus fulvus
Crvptotendipes darbvi
Diaotendipes californicus
Dicroiendipcs leucoscelis
Dicrotendipes nervosus
Endochironomus niericans
Glvptotendipes lobiferus
Glvploiendipes paripes
Harnischia amachaerus
Harnischia edwardsi
Harnischia viridulus
Lauterbomiella varipcnnis
Ornisus pica
Parachironomus carinatus
Parachironomus hinalatus
Parachironomus pectinatellae
Parachironomus tenuicaudatus
Paralauterborniella nigronalteralis
Paratendipes subaequalis
Phaenopsectra profusa
Polvpedilum halterale
Polvpedilum simulans
Stenochironomus hilaris
Pseudochironomus fulvrventris
Calopsectra confusa
Calopsectra neoflavella
qadotanvtarsus viridiventris
Micropsectra nigripila
Tanvtarsus quadratus
Parametriocnemus lundbeckii
Cricotopus remus
Nanocladius altemanthera
Clinotanvpus pinguis
Coelotanvpus scapularis
Psectrotanvpus vernalis
Ablabesmvia americana
Ablabesmvia aspera
Ablabesmvia cinctipes
Ablabesmvia illinoensis
Ablabesmvia monilis
Ablabesmvia parajanta
Ablabesmvia philosphagnos
Ablabesmvia tarella
Labrundinia floridana
Labrundinia neopilosella
Labrundinia virescens
Procladius adumbratus
Procladius culiciformis
Procladius riparius
Tanvpus grodhausi
Tanvpus stellatus
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In an isolated Florida cypress swamp dosed with treated wastewater, the following taxa were dominant:
Nais obtusa
Psvchoda albicans. altcrnata
Chironomus riparius
Polvpedilum convictus. flavus
Invertebrates that were absent (but present in untreated swamps nearby) included the following (Brightman
1976):
Lioplax subcarinata
Glvptotendipes lobiferous
Goeldichironomus sp.
Tanvpus stellatus
Anomalagrion hastatum
Orthemis ferraginea
In another Florida wetland, McMahan and Davis (1978) detected no impact on terrestrial invertebrate
diversity from wastewater additions, despite eutrophicated conditions that resulted. After addition of
manure, some Alabama ponds previously dominated by Cladotanvtarsus. Clinotanypus. and Procladius
became dominated by Dero. Stylaria. and Physa (Deutsch 1988).
In a Vermont wastewater-impacted wetland, caddisflies, clams, snails, water spiders, Crustacea, and all aquatic
insects except midges were significantly impacted (Schwartz and Gruendling 1985). The impact was due
largely to the shading out of submersed plant substrates by algal blooms. Consequently, herbivorous
mayflies and midges can begin to dominate such communities (e.g., Jones and Clark 1987). However, if
turbidity and hydroperiod conditions allow submersed or floating-leaved aquatic plants (e.g., Lemna] to
out-compete algae, other aquatic invertebrates may become dominant.
Even in the absence of human-related wastewater influences, invertebrate communities in wetlands that
naturally have low dissolved oxygen are sometimes depauperate compared to those naturally having greater
oxygen (e.g., White 1985). Wetland invertebrates that appear to tolerate low oxygen levels (which typify
even some undisturbed wetlands) are listed in Table 10. Ratios of tolerant to intolerant species have often
been used to indicate ecological status of surface waters, and could be similarly tested for use in wetlands.
Contaminant Toxicity. The availability of vegetation may be particularly important to invertebrates in
wetlands having contaminated, persistently anoxic, or highly saline sediments. In such situations, vegetation
provides an colonization surface isolated from sediments, where contaminants often are concentrated;
richness and abundance of epiphytic and nektonic invertebrate groups may thus remain high in well-
vegetated wetlands (McLachlan 1975).
Under more severe exposure to contaminants (e.g., large ambient concentrations of dissolved metals), aquatic
invertebrate species richness and density both decline, at least in shallower wetlands (Ferrington et al. 1988,
Krueger et al. 1988, Winner et al. 1975). Richness and density can decline even with levels (of phenols and
oil-water ratios) not known to be toxic in laboratory studies (Cushman and Goyert 1984).
Shifts in community composition occur as well. Specifically, shifts in structure away from aquatic insects
and toward a community dominated by certain oligochaetes (aquatic worms) have been noted in sediments
severely contaminated by heavy metals (e.g., Wentsel et al. 1978, Howmiller and Scott 1977, Winner et al.
1980). Areas that are at least moderately contaminated often are dominated by chironomid midges (Winner
et al. 1980, Cushman and Goyert 1984, Rosas et al. 1985, Waterhouse and Farrell 1985) and other aquatic
invertebrate species whose adults have wings and short life cycles, e.g., water bugs and water beetles
(Borthwick 1988, Courtemanch and Gibbs 1979, Gibbs et al. 1981). However, responses to low levels of
copper seem to be family- or genus- specific, rather than occuring at the "order" level of taxonomic
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Table 10. Examples of Aquatic Invertebrates That Tolerate Low-Oxygen Conditions in Wetlands.
Compiled from the ERAPT database (Dawson and Hellenthal 1986) and its supporting documents (Beck
1977b, Harris et al. 1978). Note that these species may occur as well in wetlands that are NOT anoxic,
although usually in smaller proportion relative to other species.
EPHEMEROPTERA (mayflies):
Callibaetis Floridanus
Hexaeenia limbata
Caen is diminuta
CHIRONOMIDAE (midges):
Chironomus attenuatus
Chironomus crassicaudaius
Chironomus riparius
Chironomus tentans
Crvpiochironomus oirtilamellatus
Crvptoiendipcs darbvi
Dicrotendipes fumidus
Dicrotendipes neomodestus
Endochironomus niericans
Glvptoiendipcs paripes
Pscctrocladius dvari
Goeldichironomus holoprasinus
Harnischia viridulus
Parachironomus monochromus
Paralauterborniella elachista
Paraiendipes albimanus
Polvpedilum aviceps
Polypedilum halterale
Polvpedilum nieritum
Polvpedilum Ontario
Polvpedilum tritum
Pseudochironomus chen
Qadotanvtarsus viridiventris
Tanvtarsus bucklcvi
Cricotopus belkini
Cricotopus politus
Cricotopus svlvestris
Coelotanvpus concinnus
Ablabesmvia aequifasciata
Ablabesmvia mallochi
Ablabesmvia rhamphe
Procladius bellus
Tanvpus carinatus-
Tanvpus neopunctipennis su
Tanvpus stellatus
Chironomus cams
Chironomus riparius
Chironomus tentans
Endochironomus nign'cans
Glvptotendipes lobiferus
Goeldichironomus holoprasinus
Cricotopus remus
Tanvpus carinatus
Tanvpus stellatus
Chironomus chelonia
Chironomus plumosus
Chironomus stigmaterus
Crvptochironomus blarina
Crvptochironomus fulvus
Dicrotendipes californicus
Dicrotendipes modestus
Dicrotendipes nervosus
Glvptotendipes barbipes
Glvptotendipes lobiferus
Glvptotendipes meridionalis
Harnischia galeator
Kiefferulus dux
Parachironomus tenuicaudatus
Paraiauterborniella subcincta
Phaenopsectra profusa
Polvpedilum digitifer
Polvpedilum illinoense
Polvpedilum obtusum
Polvpedilum scalaenum
Pseudochironomus aix
Pseudochironomus richardsoni
Micropsectra nigripila
Brillia flaviftons
Cricotopus bicinctus
Cricotopus remus
Psectrocladius dvari
Coelotanvpus tricolor
Ablabesmvia aspera
Ablabesmvia monilis
Larsia decolorata
Procladius culiciformis
Tanvpus grodhausi
Tanvpus punctipennis
Chironomus attenuatus
Chironomus plumosus
Chironomus stigmaterus
Crvptochironomus fulvus
Glvptotendipes barbipes
Glvptotendipes meridionalis
Micropsectra nigripila
Procladius culiciformis
Tanvpus punctipennis
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classification (Leland et al. 1989). Additions of heavy metals to aquatic ecosystems may increase the ratio
of predators to herbivores and detritivores, at least initially (Leland et al. 1989). Nematodes may be
particularly sensitive indicators of contaminant toxicity in wetlands that lack surface water; those of the
subclass Adenophorea tend to be more sensitive than those of the subclass Secernentea (Bongers 1990, Platt
et al. 1984, Zullini and Peretti 1986).
The commonly used herbicide, Atrazine, has been shown to cause shifts in community composition and
emergence times of aquatic insects at a concentration of 2 mg/L (Dewey 1986). Other herbicides used in
wetlands have been shown to increase the dominance of invertebrates tolerant of low dissolved oxygen, a
result related to the large oxygen deficit commonly caused by decay of massive amounts of plants (Scorgie
1980). Also, oil and associated phenols reduced richness, diversity, and total abundance of aquatic insects
in one set of wetland experiments (Cushman and Goyert 1984). The midge Cricotopus bicinctus and the
aquatic worm Limnodrilus hoffmeistcri were more prevalent downstream of than upstream from an oil spill
(Penrose 1989).
However, some midges (e.g., Nilotanypus fimbriatus) are reportedly very sensitive to oil (Rosenberg and
Wiens 1976) and pesticides (Hanson 1952). Mayflies (except burrowing species) are particularly sensitive
to metals (Leland et al. 1989, Wagerman et al. 1978), oil (Giddings et al. 1984, Cushman and Goyert 1984),
and pesticides (Hurlbert et al. 1972, AH and Stanley 1982, Van Dyk et al. 1975). Amphipods, at least the
genera Gammarus and Hyallela. and the clam shrimp (Lynceus brachvurus) appear to be very sensitive to
certain pesticides. As indicators of contamination, these freshwater shrimp have the added benefit of being
relatively stationary (i.e., because they do not emerge and fly away like aquatic insects, their presence may
be more indicative of the longer-term conditions of a wetland). Dosed populations have taken up to a year
to recover. They occur in most wetlands with standing water, and their response to pesticides has been
documented in prairie pothole wetlands (Borthwick 1988) and Maine bog ponds (Gibbs et al. 1981,
Courtemanch and Gibbs 1979). They also have been reported as absent from stormwater treatment wetlands
while present in nearby unexposed wetlands (Homer 1988).
It is conceivable that other Crustacea, such as crayfish, respond similarly. However, few community-level data
are available. Crayfish are damaged by copper levels of greater than 0.5 mg/L (Hobbs and Hall 1974),
cadmium levels greater than 10 mg/L (Fennikoh et al. 1978), and mercury levels greater than about 2 mg/L
(Doyle et al. 1978).
In wetlands that lack permanent standing water (e.g., bogs, floodplains), data on heavy metal toxicity from
terrestrial invertebrate studies may be pertinent. A summary of such studies by Bengtsson and Tranvik
(1989) reports the following:
o Species richness and, less often, total abundance of terrestrial invertebrates declines with increasing
metal concentration;
o Rare species appear more sensitive than common, widespread species;
o Least sensitive groups include soft-bodied invertebrates such as earthworms, terrestrial herbivores
such as ants and weevils, and invertebrates that inhabit the upper soil layers;
o Oribatid mites, the nematode suborder Dorylaimina, and many ground beetles (Carabidae) are highly
sensitive, whereas springtails (Collembola) as a whole are less so.
The authors suggest maximum allowable concentrations for lead of less than 100-200 mg/kg; less than 100
mg/kg for copper; less than 500 mg/kg for zinc, and less than 10-50 mg/kg for cadmium.
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Other thresholds of invertebrate toxicity for metals and/or synthetic organics are given by Johnson and
Finley (1980), USEPA (1986), EPA's "AQUIRE" database and the US Fish and Wildlife Service's
"Contaminant Hazard Reviews" series that summarizes data on arsenic, cadmium, chromium, lead, mercury,
selenium, mirex, carbofuran, taxaphene, PCBs, and chlorpyrifos.
Although not directly manifested in changes in community structure, physical deformities of individuals often
accompany severe pollution. Midges with deformed mouth parts were noted in areas of synthetic-coal-
derived oil pollution (Cushman and Goyert 1984).
Acidification. Knowledge of acidification effects on wetland invertebrate communities comes mainly from
studies in acidified lakes and streams exposed to mine drainage. As compared to circumneutral or slightly
alkaline waters, acidic waters (natural or recently induced acidity) generally have less invertebrate biomass
and/or species richness, lower ratio of consumers to producers, and fewer clearly dominant taxa (e.g., Friday
1987, Hall and Likens 1980, Harvey and McArdle 1986, Letterman and Mitsch 1978, Parsons 1968, Smock
et al. 1981, 1985, Thorp et al. 1985, Walker et al. 1985a, Warner 1971). However, several studies, e.g.,
those of some acidic "blackwater" streams, have detected no significant differences in lake or stream
invertebrate numbers or richness attributable to pH differences (e.g., Bradt et al. 1986, Bradt and Bert 1987,
Collins et al. 1981, Crisman et al. 1980, Kelso et al. 1982, Winterbourn and Collier 1987). The effects of
acidification may interact with and possibly be overshadowed by trophic conditions of wetlands (Brett 1989,
Kerekes et al. 1984, Schell and Kerekes 1989).
Shifts in community composition are probably the most frequently measured effect of acidification.
Particularly acid-sensitive are species of gastropods (snails), pelecypods (clams and mussels), daphnids,
ephemeropterans (mayflies), amphipods (freshwater shrimp), and some midges (particularly the subfamilies
Chironominae and Orthocladinae) (Allard and Moreau 1987, Bell 1971, Friday 1987, Hall et al. 1980, Harvey
and McArdle 1986). Some of the first species to be affected by acidification are Crustacea- the predaceous
copepod, Epischura lacustris (Sprules 1975), and the freshwater shrimp, Hyalella azteca (Zischke et al. 1983)
and Gammarus lacustris. Taxa reported to be more prevalent under acidic conditions include oligochaetes,
acarids (water mites), the phantom midge, Chaoborus. and midges of the subfamilies Tanypodinae and
possibly Chironomini (Allard and Moreau 1987, Bradt and Bert 1987). A few caddisflies, freshwater
sponges, dragonflies, water bugs (Corixidae), water beetles (Dytiscidae), and Tanytarsini midges tolerate
weakly acid conditions (Fowler et al. 1985, Walker et al. 1985a). Species of midges and caddisflies known
to occur under acidic conditions are listed in Table 11, based on data compiled by Beck (1977b) and others.
Salinization. Naturally saline, nontidal wetlands typically have low diversity of aquatic invertebrates
(Kantrud 1989) and are dominated by brine shrimp (Arternia), brine flies (Ephydra). and a few species of
midges and aquatic worms. Severe increases in salinity of freshwater habitats also can diminish invertebrate
community biomass and species richness. However, rather few data have been collected specifically from
inland brackish wetlands, so relative tolerances of species to increased salinity are poorly known.
Other taxa known to be relatively tolerant include certain species of midges, mosquitoes, aquatic worms,
dragonflies, water bugs, and water beetles (Kreis and Johnson 1968). Crayfish generally require salinities
less than 15 ppt (Loyacano 1967). Former salt marshes that were converted to freshwater wetlands were
found to have fewer midges of the subfamily Orthocladiinae than expected (Walker et al. 1985a).
If more information were available on tolerances, such data might be used (e.g., as a ratio of salt-tolerant
to salt-intolerant species) in conjunction with background chemical data to indicate stress to wetlands from
irrigation runoff water, cultivation of saline soils, coastal saltwater intrusion, or other salinity sources.
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Table 11. Examples of Invertebrates That May Tolerate or Prefer Acidic Conditions in Wetlands.
Compiled from the ERAPT database (Dawson and Hellenthal 1986) and the following references: Beck
1977b, Kimerle and Enns 1968, Smock et al. 1981, Walker et al 1985a. Note that these species may occur
as well in wetlands that are NOT acidic, although usually in smaller proportion relative to other species.
Ablabesmvia americana. A. aspera. A. basilis. A. hauberi. and A. parajanta. A. peleensis. A. philosphagnos
Chironomus (some species)
Cladopclma (some species)
Cladotanytarsus (some species)
Corynoneura taris
Crvptotendipes casuarius
Dicrotendipes incurvus. D. leucoscelis
Guttipelopia currani
Harnischia amachaerus. H. bovdi
Krcnosmittia (some species)
Labrundinia floridana. L. johannseni. L. neopilosella, L. virescens
Lauterborniella varipennis
Metriocnemus abdomino-flavatus. M. hamatus. M. knabi
Monopelopia tillandsia
Monopsectrocladius (some species)
Nilotanvpus americanus
Nimbocera (some species)
Omisus pica
Orthocladius annectens
Pagastiella orophila
Parachironomus alatus. P. scheideri
Paramerina anomala
Polvpedilum braseniae. P. nvmphaeorum. P. obtusum
Procladius bellus
Tanvpus neopunctipennis
Tanvtarsus (some species)
Thienemannimvia senata
Tribelos quadripunctatus
Trissocladius (some species)
Dugesia tigrina
Nais (some species)
Limnodrilus hoffmeisteri
Aulodrilus piqueti
Crangonvx (some species)
Hvdracarina
Callibaetis diminuta
Caenis diminuta
Oxvethira (some species)
Palpomvia (some species)
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Burial/sedimentation. High rates of sedimentation (7 cm/yr) resulted in lower diversity, richness, and total
community biomass in a southern river system (Cooper 1987). Fine-particle sediments, particularly if anoxic,
support reduced diversity and richness of invertebrates (Wilbur 1974). Species of mayflies and chironomids
Species of mayflies and chironomids that feed mainly on algae are particularly affected, while burrowing
invertebrates might be expected to be least-affected. In Lake Erie, the abundance of tubificid worms was
correlated with the sediment accumulation rate and organic carbon flux (but not to organic carbon)
(Robbins et al. 1989). Excessive sedimentation may be indicated by absence of the freshwater bryozoans,
e.g., Pectinella magnifica. and the fingernail clam Sphaerium rhomboideum (Cooper 1987, Cooper and
Burris 1984).
Turbidity/Shade; Vegetation Removal. Removal of aquatic bed vegetation can increase algae in wetlands,
thus increasing the ratio of herbivorous species (e.g., certain mayflies) to detritivorous species (e.g., certain
midges and worms). Submersed plants and logs have among the highest densities and species richness of
any aquatic substrate, e.g.:
Armstrong and Nudds 1985, Boerger et al. 1982, Chubb and Listen 1986, Crowder and Cooper 1982,
Durocher et al. 1984, Dvorak and Best 1982, Floyd et al. 1984, Gilinsky 1984, Hall and Werner
1977, Kallemeyn and Novotny 1977; Kimble and Wesche 196, Krecker 1939, Krull 1970, Menzie
1980, Minkley 1963, Miller et al. 1989, Mittelbach 1981, Poe et al. 1986; Scheffer et al. 1984,
Schramm et al. 1987, Teels et al. 1978, Voigts 1976, Ware and Gasaway 1978, Wetzel 1975.
Indeed, equations for predicting the density of aquatic invertebrates in submersed vegetation (lacustrine
aquatic bed) have been developed by Cyr and Downing (1988), using data on biomass of individual
macrophyte species and season. Thus, removal or loss of aquatic vegetation due to shading/turbidity can be
expected to profoundly affect the invertebrate resource (e.g., Bettoli 1987, Vander Zouwen 1983). On the
other hand, selective removal of dense macrophyte stands can increase density, biomass, and/or richness of
remaining invertebrate communities (e.g., Beck et al. 1987, Broschart and Linder 1986, Kaminski and Prince
1981, Kenow and Rusch 1989, Murkin and Kadlec 1986). Removal of the canopy of one forested floodplain
wetland had little effect on aquatic invertebrate richness and density (Boschung and O'Neil 1981). The
degree to which vegetation removal has a neutral or beneficial effect on macroinvertebrates may depend
partly on the type of removal procedure (e.g., mechanical thinning, ditching, burning, herbicides, crayfish
introduction) and the spatial patterns created (Nelson and Kadlec 1984).
Non-aquatic invertebrates may also respond to removal of woody and emergent vegetation. For example,
a decline of wetland spider richness accompanied peat harvesting in a bog (Koponen 1979).
Thermal Alteration. Heated effluents generally reduce the richness of invertebrate communities in wetlands
and may either increase or decrease their density and productivity (Gibbons and Sharitz 1974, McKnaught
and Fenlon 1972, Nichols 1981, Poff and Matthews 1986, Whitehouse 1971, Wiederholm 1971). Increases
in secondary productivity are the result of higher primary productivity associated with warmer temperatures
and longer growing seasons. Crayfish generally cannot tolerate temperatures greater than about 30 C
(Becker et al. 1975). Backswimmers (Corixidae) and midges appear to tolerate moderately warmed surface
waters (Gibbons and Sharitz 1974). Temperatures of over 40 C apparently do not significantly affect the
life cycle of the midges Chironomus sp., Tanvpus neopunctipennis. or Tanvtarsus sp.; where deep, soft
substrates are available as refugia for burrowing species, damage from thermal increases may be lessened
(Coler and Kondratieff 1989). The ratio of burrowing oligochaetes, nematodes, gastropods, chironomid
midges, and nektonic invertebrates to other aquatic invertebrates might thus be tested as one indicator of
thermal disturbance.
Dehydration, Inundation. Water levels profoundly affect the abundance and community composition of
invertebrates (Reid 1985, Wiggins et al. 1980). Addition of permanent open water to a non-permanently
flooded wetland increases the opportunity for invasion by many submersed and floating-leaved species that
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provide complex substrates for aquatic invertebrates. This consequently can result in an increase in on-
site species richness, and perhaps increased density, of wetland invertebrates. For example, inundation of
emergent wetlands was noted to increase the density, biomass, and richness of invertebrates (Huener 1984),
and cause a shift in community composition toward herbivores and detritivores (Murkin and Kadlec 1986).
For Mississippi River borrow pit wetlands, "days flooded" was the most significant factor explaining
invertebrate density in a multivariate regression; flooding in the sampled wetlands ranged from 24 to 115
days annually, with a mean of 81 (Cobb et al. 1984).
However, if inundation in some wetlands is prolonged (throughout the growing season) and deep, the
resulting oxygen and light deficits may result in diminished richness and density of aquatic plants (Ebert and
Balko 1987). Prolonged growing-season flooding, when it occurs in wetlands that have no prior history of
such flooding, results in diminished invertebrate density and richness (Driver 1977, Hynes and Yadev 1985,
Neckles et al. 1990). In forested floodplain wetlands, invertebrate species richness and abundance decrease
with increasing soil moisture and flood frequency (e.g., Uetz et al. 1979) and with disruption of normal
sequencing of flooding (Sklar and Conner 1979).
When wetlands that normally contain standing water are almost totally dehydrated for short periods (i.e.,
"drawdown"), the result is usually a major increase in nutrients, algae, and invertebrate density (Benson and
Hudson 1975, Reid 1985, Wegener et al. 1974). This effect may be less pronounced if a dense canopy
prevents sufficient light for algal growth, and exchange rates of wetland water with adjacent waters are
minimal. Also, less mobile taxa, such as freshwater clams, may be particularly sensitive to drawdown. They
can become stranded and perish during rapid drawdown unless underlying sediments remain saturated and
soft so individuals can burrow down into the saturated zone (Jiffry 1984).
Although invertebrate density may increase following reflooding of dehydrated wetlands, invertebrate richness
may not, particularly if sediments have become heavily oxidized and hardened during exposure (Hunt and
Jones 1972). If wetlands are dehydrated irregularly and rapidly (e.g., by frequent passage of large ships)
or for long periods (e.g., reservoir fluctuations), both abundance and richness of invertebrates can decline
(Hale and Baynes 1983, Smith et al. 1987).
Invertebrate taxa can be classified into groups (response guilds) related to their life cycles and preference
for particular wetland hydroperiods. Conceivably, ratios of these groups (e.g., density-weighted ratio of
short-lived/mobile species to longer-lived/immobile species) could be tested as an indicator of wetland
hydrologic status, as has been done with midges (Driver 1977) and water beetles (Hanson and Swanson
1989). In prairie pothole wetlands, chironomid diversity was also found to increase with permanency of the
hydroperiod (Driver 1977), although contrary evidence is presented by Neckles et al. (1990). Individual taxa
might be assigned to the following response groups (Delucchi 1987, Jeffries 1989, McLachlan 1970, 1975,
1985, Wiggins et al. 1980):
o Overwintering Residents: disperse passively; include many snails, mollusks, amphipods, worms,
leeches, crayfish.
o Overwintering Spring Recruits: reproduction depends on water availability; include most midges,
some beetles.
o Overwintering Summer Recruits: reproduce independent of surface water availability, requiring only
saturated sediment; include dragonflies, mosquitoes, phantom midges.
o Non-wintering Spring Migrants: mostly require surface water for overwintering, adults leave
temporary water before it disappears in spring or summer; includes most water bugs, some water
beetles.
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Thus, changes in density-weighted ratios of response groups, monitored from a large regional set of wetlands,
might be used to indicate changing hydrologic conditions over time. However, additional research may be
needed because some recent evidence suggests that certain taxa (species of Dytiscidae, Corixidae,
Ceratopogonidae, Ephydridae, and even Chironomidae) may be unaffected by water regime in some
situations (Neckles et al. 1990).
Fragmentation of Habitat. We found no explicit information on wetland invertebrate community response
to fragmentation of regional wetland resources. A study of prairie potholes indicated increased diversity
with increased wetland size, and the author suggested that might be due to the increased distance of smaller
areas from larger and more stable wetlands (Driver 1977). Increased richness and interspersion of plant
forms within a wetland can result in increased macroinvertebrate richness and numbers (Voigts 1976).
One can surmise that as the distance between wetlands with colonizers becomes greater, species with narrow
environmental tolerances and which do not disperse easily might be most affected. Indeed, in a study of
essentially identical wetlands, Jeffries (1989) found that statistical clusters of invertebrate taxa were defined
by the distance and surface water connection of their associated wetland from a much larger regional water
body. However, even apparently "immobile" species such as amphipods and clams have some capability for
dispersal (Swanson 1984).
Landscapes where wetlands are interspersed with uplands can have almost 70 percent more invertebrate
species than those containing only uplands (Coulson and Butterfield 1985). In lakes, the species richness
of mollusks (Aho 1978, Lassen 1975), midges (Driver 1977), and crustaceans (Fryer 1985) increases with
increasing lake area.
8.2 SAMPLING METHODS AND EQUIPMENT
Natural factors that could be important to measure and (if possible) standardize among wetlands when
monitoring anthropogenic effects on community structure of invertebrates include:
age of wetland (successional status), water or saturation depth, conductivity and baseline chemistry
of waters and sediments (especially pH, alkalinity or calcium, and organic carbon), sediment type,
current velocity, presence of fish, stream order or ratio of discharge to watershed size (in riverine
wetlands), density, type, and form of vegetation and woody debris (particularly, total surface area),
ratio of open water to vegetated wetland, and the duration, frequency, and seasonal timing of regular
inundation, as well as time elapsed since the last severe inundation or drought.
Sampling methods for wetland or lake littoral invertebrates are described in Downing and Rigler 1984,
Edmondson and Winberg 1971, Fredrickson and Reid 1988b, Isom 1986, Murkin and Murkin 1989, Witter
and Croson 1976, and others. Although addressing streams, the book by Elliott (1971) is an important
reference for sampling program design and data analysis.
Larval aquatic invertebrates can be found in wetlands throughout the year. If wetlands can be sampled only
once, then the late wet season or beginning of the dry season, if they coincide with the growing season, are
usually the recommended time, as density and richness tend to be greatest then (Marchant 1982).
Alternatively, if conditions among a series of years are to be compared and the primary desire is to minimize
variability, then dry-season measurements made just before the onset of flooding may be best (McElravy et
al. 1989). However, the chronology of density peaks can vary even among wetlands in close proximity,
possibly due in some cases to differences in predation (Campbell 1983).
In either case, and particularly in disturbed and intermittently flooded wetlands, caution is needed to
schedule sampling to coincide with phenologies of particular taxa (Sklar 1985). For example, one might
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want to avoid sampling immediately after a synchronous emergence of the usually dominant species.
Maximum information is often obtained when most invertebrates are within a size range (later instars)
retained by nets used to sample them, and can be identified with greatest confidence. For biomass
estimates, Hanson et al. (1989) reported that samples collected at 4- and 6- week intervals were very similar
to those based on 9 biweekly collections. For a bog stream monitored over 23 months, Boerger et al. (1982)
reported a 17-fold variation in midge densities, and even greater variation was reported by Gatter (1986).
The choice of equipment depends largely on the wetland microhabitat to be sampled. Different assemblages
of wetland invertebrates inhabit sediments (benthos), rooted plants or algae (phytomacrofauna), open water
(nekton), and the surface film (neuston). Subsequent data analysis can use groupings based on ecological
niches associated with each taxon (e.g., Cummins and Wilzbach 1985).
A significant problem in analyzing wetland invertebrate data arises from difficulties in determining the
spatial dimensions of the area from which a sample was drawn. Accurate estimates of density (individuals
per unit area) are difficult to achieve due to difficulties in accurately measuring the complex wetland
substrate (submerged plants, tree trunks, emergent plant stems, logs, etc.). To address this, some
investigators have removed the substrate along with the collected sample, weighted both, and reported
density as weight or number of organisms per unit weight of substrate. In some cases regression coefficients
have been calculated to convert plant weights to plant area, which may be further converted to invertebrate
density (Downing 1986). Another approach has been to base comparisons among similar wetland habitats
on similarity indices and richness (per number of individuals), rather than on density and biomass.
If the objective is to sample invertebrate communities attached to wetland plants (e.g., snails, many mayflies)
and the water column, sweep nets (dip nets) are commonly used. These are the familiar long-handled insect
nets. They may be used in water or air, so long as vegetation is not dense. Usually, they are either swept
through a standard length of vegetation, or placed on the bottom and hauled vertically through the water
column in a rapid stroke. They are convenient to use, and are particularly suited for capturing large (e.g.,
crayfish) or quick-moving species not collected by other methods, such as adult dragonflies and water
striders. Disadvantages include user variability and the fact that their samples are not strictly quantitative,
since the unit of area swept is difficult to accurately determine (Adamus 1984, Plafkin et al. 1989).
Trials by Furse et al. (1981) and Friday (1987) indicate that at least 80 percent of the species found to be
present in a particular aquatic plant bed using 5 to 10 sweeps can be captured in half that number. In trial
comparisons against a modified Gerking sampler (see below), Kaminski and Murkin (1981) found sweep nets
to be just as effective in sampling water-column taxa, although Gillespie and Brown (1966) had come to the
opposite conclusion. In wetland studies, sweep nets have been used by Borthwick (1988), Courtemanch
and Gibbs (1979), Smith et al. 1987, Voigts 1976, White 1985, and others.
Another option for sampling plant-dwelling invertebrates in wetlands involves directly clipping the vegetation
and returning it in an enclosed box to the lab. This can be used for both submersed and emergent plants,
and provides more precise quantification than does use of sweep nets. Vacuum suction can also be used
to remove small invertebrates from foliage in the field (Southwood 1981). Downing and Cyr (1985) found
the most cost-effective quadrat size for clipping to be 500 cm^. Plants were enclosed in a 6-liter plastic box.
Clipping aquatic macrophtyes in quadrats of varying sizes yielded five times higher populations than did
sampling with Gerking, Macan, Minto, or KUG samplers. Gates et al. (1987) described a sampler useful
for taking simultaneous samples of sediment invertebrates and plant-dwelling invertebrates. They found this
to give results for plant invertebrates at least as precise and sometimes more accurate than obtained by
clipping macrophytes.
A third option for sampling invertebrates of wetland plants involves use of artificial substrates. Plants are
not sampled directly, but rather, plastic plants or other sterile surfaces (e.g., Hester-Dendy plate samplers)
are totally submersed in the wetland water column and allowed to be colonized over a period of at least a
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month (Macan and Kitching 1972). Because they standardize surface area and texture, collections from
substrate samplers are highly comparable to each other, making them attractive for use in monitoring of
water column water quality. They also are lightweight, can be used in areas difficult to sample by other
means (e.g., deep rivers), and sample processing is relatively clean. However, disadvantages include the fact
that a return trip to the wetland is required, vandalism may be a problem, their use is limited to wetlands
with surface water, they sample only epiphytic species, and representativeness can be questioned (Adamus
1984).
In an aquatic bed wetland, Gerrish and Bristow (1979) used plastic mimics of the pondweed, Potamogeton
richardsonii. interspersed among live experimental plants. Although this yielded no significantly different
numbers of invertebrates or species per unit of surface area than were found on real plants, aquatic worms
were significantly more common on the artificial substrates, and the substrates did not accurately reflect the
densities of invertebrates on the nearby Myriophyllum or Vallisneria plants.
Natural substrates initially devoid of organisms can also be used as colonization substrates. For example,
plant litter was placed in boxes made of hardware cloth by Batema et al. (1985) and White (1985), for
sampling macroinvertebrates in eastern floodplain forests. Artificial substrates are often ineffective for
collecting large crustaceans (e.g., crayfish) and mollusks.
If the objective is to sample invertebrate communities inhabiting wetland sediments, then dredges-- also
called grab samplers (Ekman, Ponar, etc.)~are often used. They essentially consist of a box with jaws that
is lowered onto the sediment. The jaws enclose a specified area of bottom, and retrieve sediments and
associated organisms to a sediment depth of about 5 cm. Dredges are used only where surface waters of
at least 0.5 m in depth are present, and are not effective where there are rocks, aquatic plants, or logs to
jam the jaws. They have been used in wetlands by Bradt and Bert (1987), Driver (1977), and Krull (1970).
Estimates of density are only crudely quantitative because jaws seldom close tightly, allowing organisms to
escape. Large organisms (e.g., crayfish), water column organisms, and fast-moving species are poorly
sampled.
Another option for sampling sediments is to use core samplers. Unlike grabs, corers do not have jaws, and
instead rely on compactive force or suction to retrieve sediments. They suffer the same disadvantages as
dredges. Samples may be more precisely quantitative, but the mean size of organisms effectively captured
may be smaller, due to the narrowness of corers. Core samplers may be the only option for quantitatively
sampling sediment organisms in wetlands that lack surface water, and a variety of designs are available (e.g.,
Bay and Caton 1969, Coler and Haynes 1966). Core samplers are widely used in paleoecological studies of
wetlands. Florida regulations for monitoring treated wastewater discharges to forested wetlands specify that,
if a core sampler is used, devices with minimum sampling area of 45 crn^ be used, and that the number of
samples at a given station within the wetland be that number needed to be 90% certain of being within 15%
of the mean diversity of the population. Where aquatic plants interfere, some investigators have suggested
a saw blade might be welded to the leading edge of the corer, for clipping heavy roots and stems (e.g.,
Murkin and Kadlec 1986). Where sediments are frozen, metal ice spades have been used to collect samples
(e.g., Jacobi 1978).
Where sediments or soils are not covered by water (e.g., in peat bogs), pitfall traps and soil extraction
techniques can also be used to augment vacuum sampling and sweep-net sampling, and may produce the
highest densities and species richness (Coulson and Butterfield 1985).
If the objective is to sample invertebrates that inhabit the water column, tubular samplers (e.g., "Gerking
samplers", "stovepipes", "Hess samplers", "box samplers") can be used. These are wide cylinders that enclose
a standard area of bottom and usually are not designed for effectively penetrating the sediment. In some,
the bottom can be sealed off with a sliding door, plug, or similar feature once the sampler is in place.
Some have been fitted with a reinforced cutting edge on the bottom. Designs are described by Freeman et
al. (1984), Gerking (1957), Korinkova (1971), Hiley et al. 1981, Legner et al. 1975, Mackay and Quadri
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(1971), Martin and Shireman (1976), Minto (1977), and Swanson 1978. They are not effective for catching
quick-moving organisms, burrowing species, very large taxa, many epiphytic species, or for use in flowing
water.
Emergence traps and funnel traps consist of nets or funnels anchored at and just above the water surface.
They passively collect aquatic insects as they pass into their winged adult stage and emerge from the water
column. Funnel traps are used to collect swimming, air-breathing insects as well as emerging species (e.g.,
Greenstone 1979, Henrickson and Oscarson 1978, Kaminski and Murkin 1981). Traps-either submerged,
at the water surface, or above it- can be fitted with lights to increase their attraction to some adult insects,
for example:
Aiken 1979, Apperson and Yows 1976, Carlson 1971, Carlson 1972, Espinosa and Clark 1972,
Hungerford et al. 1955, Husbands 1967).
A variety of designs for emergence and funnel traps have been tried, for example:
Corbet 1965, Daniel et al. 1985, Deonier 1972, Lammers 1977, Lemke and Mattson 1969, McCauley
1976a,b, Pritchard and Scholefield 1980, Rosenberg et al. 1980, Voigts 1973, and Washino and
Hokama 1968, and some were evaluated by Kimerle and Anderson (1967).
Use of funnel and emergence traps is limited to wetlands containing open patches of surface water during
the growing season, when most insects emerge. They can be used in both still-water and slow-flowing
wetlands, particularly those difficult to sample by other means, and samples are relatively debris-free and
easy to process. Because they are left in place (sometimes for many weeks), they avoid the problem
encountered by other samplers of missing key species due to inappropriate time of visit.
Because emerging insects come from a variety of microenvironments, emergence traps can integrate well the
extreme spatial heterogeneity within many wetlands. On the other hand, this makes it impossible to
standardize or determine the unit of area measured. Thus, they would not be suitable for tracing the
leading edge of an effluent plume within a small wetland. Samplers designed for passively collecting
terrestrial insects (e.g., light traps, pitfall traps, malaise traps) encounter the same problem. Also, many of
the wetland invertebrates most sensitive to pollution do not emerge (e.g., amphipods, aquatic worms, snails),
so are not collected by emergence traps. Initial purchase of traps can be costly, and vandalism may be
problematic.
In prairie potholes, conical emergence traps were situated at 3 m intervals perpendicular to shore (Driver
1977). To detect immediate effects of pesticide application, Gibbs et al. (1981) emptied emergence traps
every two hours, from 6 AM to 8 PM. Normally, traps are left in place for many days or weeks. Welch
et al. (1988) used submerged funnel traps to catch emerging midges in a lake. They found no difference
in total catch between 0.142-m^ and 0.283-m^ trap sizes. Traps with inverted funnels inserted in the jar
necks caught more pupae than traps without funnels, and total catch in the traps without jars was 58
percent of the catch in traps with funnels. Rosenberg et al. (1984) submerged their funnel traps, situating
them at depths of 1, 2, 3.5, and 4.5 m.
8.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS
In general, quantitative data on wetland macroinvertebrates has not been uniformly collected from a series
of statistically representative wetlands in any region of the country. Thus, it is currently impossible to state
what are "normal" levels for parameters such as seasonal invertebrate density, species richness, or biomass,
and their temporal and spatial variability, in any type of wetland.
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Perhaps the closest approximation of such a data set is that of Giese et al. (1987). These invertebrate data
were collected in part from streams flowing through relatively pristine floodplain wetlands, and thus help
serve as a regional baseline for bottomland hardwood wetlands. Collecting a single, timed (30-minute),
series of dip-net samples from each site, the investigators found an average of 50-60 invertebrate taxa,
containing an average total of about 800 individuals. The Shannon diversity index averaged 4.17 to 4.67 in
these relatively pristine lowland streams.
Also, the U.S. Geological Survey is presently initiating a program (NAWQA) to monitor stream invertebrate
communities inhabiting a carefully selected sample of watersheds throughout the United States. Although
wetlands will not be a specific target of the monitoring, the spatial and regional variability of invertebrates
may become better known from this probability sampling approach. Another regionally extensive project
was undertaken by Corkum (1989) in the Pacific Northwest/Alaska, and resulted in an ecologically-based
classification of stream types for that region. Other data on wetland macroinvertebrates is selectively
summarized in Table 12.
Coefficients of variability for invertebrates in streams range from about 0.2 to 0.8 (Eberhardt 1978). Few
such values apparently have been published for wetlands. Variation in invertebrate density among habitats
within wetlands has been documented in some cases, for example:
Beck 1977a, Erman and Erman 1975, Gatter 1986, Kansas Biological Survey 1987, Neuswanger et
al. 1982, Thorp et al. 1985.
However, only a few studies in the U.S. have quantified invertebrate community differences among a series
of wetlands in a region, and have mostly focused on lacustrine or riverine wetlands. These include:
Bradt et al. 1986, Campbell 1983, Cobb et al. 1984, Erman and Erman 1975, Ferren and
Pritchett 1988, Garono and MacLean 1988, Haack et al. 1989, Hepp 1987, Kallemeyn and
Novotny 1977, Krull 1970, Lowery et al. 1987, Mathis et al. 1981, McAuley and Longcore
1988, Stoddard 1987, Teels et al. 1978.
Although data exist that quantify year-to-year variation in invertebrate community structure in other surface
waters (e.g., McElravy et al. 1989), such studies have apparently not been published for wetlands.
Conceivably such unpublished data may be available from sites of the U.S. Department of Energy's National
Environmental Research Park system, as well as the following sites of the National Science Foundation's
Long Term Ecological Research (LTER) program (that contain studied wetlands): Illinois Pool 19 site,
Illinois-Mississippi Rivers sites, New Hampshire Hubbard Brook riparian forest, Oregon Andrews
Experimental Forest riparian forest, and Michigan Kellogg Biological Station site.
Quantitative published data on composition of aquatic invertebrate communities appears to be most
available for submersed vegetation (aquatic bed wetlands), particularly in the Southeast and Prairie pothole
region. Apparently such data are most limited for wetlands that are saturated but mostly lack standing
water (e.g., bogs), as well as for playas and non-Southeastern riparian wetlands. Information is most
available on impacts of hydrologic alteration, acidification, and nutrients, and least on impacts of
salinization, sedimentation, thermal alteration, and habitat fragmentation.
Even qualitative lists of "expected" aquatic invertebrates in wetlands of various types do not appear to have
been compiled, either nationally or by individual states. The USFWS has begun to compile such lists (pers.
comm., Buck Reed, USFWS, St. Petersburg, FL) and some publications in the "community profile" series
of the USFWS (Appendix C) mention particular taxa known to occur in wetlands.
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Table 12. Examples of Invertebrate Density and Biomass Estimates from Wetlands.
BIOMASS (g/m2)
State type* N
AR** L
AR** Pfo
LA L
LA Pab 48
LA Pab ?
MS Pfo ?
WA Pfo 18
WI Pem ?
SD Pem 220
CA P(fen) 0.9
min.value
max/value
15.0
mean=8.50
1.3
0.5 15.7
4.2 6.8
3.2
2.5 5.7
0.6 1706
1.3 8.5
8.5
DENSITY (number/m2)
State type N min.value
AR* L
AR*** Pfo
CA Pem 230 6952
FL Pfo
FL Pfo
IA Pem
IA L
KS Pem ? 508
LA Pab ?
LA Pfo
LA Pfo
LA Pfo ?
LA Pfo 70
Pem 13
MI Lab
MO Pfo
MS Pfo ? 1675
GA Pab ?
MO Pfo 4
IL-MO L 33 247
KY Pem 84 739
NJ Pem ? 196
WI Pem ?
MI Lab ? 7665
SD Pem 175 584
SD Pem 220 3533
OR R 64 33
OR Pem ? 11
max/value
> 10,000
mean=2967
23,857
1,102
2.5
>20,000
4,108
18,676
76,990
16,198
12.5
16,000
mean=95/grab
mean=2900/grab
> 10,000
>9,000
9248
mean=12,093/m2
5045
4321
5143
335,547
35,730
13,243
5929
15,193
15,700
1745
citation
Lowery et al. 1987
Cobb et al. 1984
Tebo 1955
Sklar 1985
Sklar & Conner 1983
Baker et al. 1988
Meehan-Martin & Swanson 1988
Schmal & Sanders 1978
Broschart & Linder 1986
Erman & Erman 1975
citation
Lowry et al. 1987
Cobb et al. 1984
Erman and Erman 1975
Brightman 1984
Brightman 1984
Voights 1976
Tebo 1955
Kansas Biol. Surv. 1987
Sklar and Conner 1979
Sklar 1985
Sklar 1985
Sklar & Conner 1979
Beck 1977a
Beck 1977a
Lowery et al. 1987
Batema et al.
Baker et al. 1988
Smock & Stoneburner 1980
Batema et al. 1985
Jones et al. 1985
Bosserman & Hill 1985
Gatter 1986
Schmal et al. 1978
Hiltunen & Manny 1982
Benson & Hudson 1975
Broschart and Linder 1986
Kreis & Johnson 1968
Fishman 1989
* wetland codes (Cowardin et al. 1979): Pab=palustrine aquatic bed, Pem=palustrine emergent,
Pfo=palustrine forested, L=lacustrine, R=riverine
** includes TN and MS
** includes LA, MS, and MO
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9.0 WETLAND FISH COMMUNITIES
This discussion includes both adult and larval fish, both game and nongame species. Few freshwater fish
spend their entire life in wetlands, and wetlands that seldom contain surface water (e.g., raised bogs) do not
usually have fish. Although fish community structure has been widely described in lakes and rivers, and
"indices of ecological integrity" which integrate community data have been developed and tested (Karr 1981),
such efforts have not yet been transferred to wetlands. Advantages and disadvantages of use of fish as
indicators are shown in Appendix A. The paper by Munkittrick and Dixon (1989) provides further
discussion of the value of fish as indicators of ecosystem condition. They assert that fish populations, in
general, respond to reduced food resources initially by a decline in fecundity, followed by reduced condition
factor, an increase in mean age, and finally a drop in population level. They suggest that these
characteristics might be used to indicate the "health" of a particular population, and in some cases, the types
of stress that are impairing population health.
9.1 USE AS INDICATORS
Enrichment/Eutrophication. Nutrient enrichment can result in increased fish biomass (Colby et al. 1972,
Gascon and Leggett 1977) and altered species diversity (Nakashima et al. 1977) in lakeshore wetlands.
Increased biomass may result from increased biomass of invertebrate fish foods, these having increased as
a result of increased attachment surfaces and detritus provided by nutrient-induced expansion of submersed
wetland plant beds (Pardue 1973). If fish food is already abundant, eutrophication may result in population
increases in addition to biomass increases (Nakashima and Leggett 1975). Omnivorous species may benefit
the most from the increase in submersed plants (Camp, Dresser and McKee 1989). Walleye fStizostedion
vitreum) and Mosquitofish (Garnbusia) are two of dozens of wetland species that tolerate eutrophic
conditions (Dawson and Hellenthal 1986), but few species occur exclusively in eutrophic waters.
Organic Loading/Reduced DO. Among northern lacustrine wetlands, Rahel (1984) reported that the ratio
of cyprinid to centrarchid fish was greater where winter anoxia occurred. Rivers downstream from sewage
and industrial waste outfalls showed a decline in fish community richness in Illinois (Lewis et al. 1981) and
in Louisiana (Gunning and Suttkus 1984). In the latter study, two species of darter, Ammocrypta vivax and
Etheostoma histrio, were particularly intolerant of the effluents. A southern wetland exposed to treated
wastewater experienced increased fish productivity and decreased fish species richness (Camp, Dresser, and
McKee 1989). Fish habitat in another wetland, a cypress pond in Florida, was degraded by wastewater
effluent (letter and Harris 1976).
The State of Florida's regulations for discharge of treated wastewater into wooded wetlands specify that the
biomass of sport-commercial or forage fish shall not be allowed to decline by more than 10%; exceptions
may be allowed if such declines can be attributed, through analysis of covariance, to other factors. The
State also specifies that the biomass of rough fish shall not increase more than 25% unless the ratio of sport
and commercial fish to rough fish is maintained; sampling protocols are specified. Florida regulations
consider the following fish taxa to be most tolerant of treated wastewater: suckers (all Catostomidae), tilapia
(all Chichlidae), gar (Lepisosteidae), bowfin (Amia calva). grass carp (Ctenopharyngodon idella). common
carp (Cyprinus carpioX and gizzard shad (Dorosoma cepedianum). Table 13 includes some other species
that tolerate relatively low levels of dissolved oxygen.
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Table 13. Examples of Wetland Fish Species That Tolerate Low Dissolved Oxygen.
Compiled from the ERAPT database (Dawson and Hellenthal 1986). Note that these species may occur as
well in wetlands that are NOT anoxic, although usually in smaller proportion relative to other species.
Amia calva Bowfin
Cvprinus carpio Common Carp
Eriomvzon sucetta Lake Chubsucker
Etheostoma nigrum Johnny Darter
Ictalurus melas Black Bullhead
Ictalurus natalis Yellow Bullhead
Ictalurus nebulosus Brown Bullhead
Moxostoma carinatum River Redhorse
Notemigonus crvsoleucas Golden Shiner
Notropis buchanani Ghost Shiner
Notropis heterodon Blackjaw Shiner
Notropis heterolepis Blacknose Shiner
Noturus gyrnus Tadpole Madtom
Umbra limi Central Mudminnow
Contaminant Toxicity. Declines in species richness and density of fish as a result of contaminants (oil, heavy
metals, pesticides, etc.) have been widely documented in lakes and streams, but less often in wetlands. There
exists a wealth of lexicological data from laboratory bioassays and tissue analyses. These include Johnson
and Finley (1980), USEPA (1986), USEPA's "AQUIRE" database, and the US Fish and Wildlife Service's
"Contaminant Hazard Reviews" series that summarizes data on arsenic, cadmium, chromium, lead, mercury,
selenium, mirex, carbofuran, toxaphene, PCBs, and chlorpyrifos. However, relatively few field data are
available for judging which wetland species are most sensitive.
Acidification. Acidity clearly affects fish species richness in lacustrine wetlands (Jackson and Harvey 1989,
Rahel and Magnuson 1983, Tonn and Magnuson 1982). Fish species richness declined among a series of
lacustrine wetlands with progressively more acidic conditions (Rahel 1984,1986). Various reviews (e.g., Ford
1989, Hastings 1984, Wiener et al. 1983) indicate that, in northern lakes and streams, species most
susceptible to the effects of acidification include lake trout, brook trout, Atlantic salmon, smallmouth bass,
walleye, burbot, and common shiner and various other species of minnows. Data on acidification effects in
other regions and wetland types are limited.
Salinization. No quantitative, published information was found concerning the effects of salinization of
wetlands on community structure of indigenous fishes.
Sedimentation/Burial. No quantitative information was found concerning the effects of sedimentation in
wetlands on community structure of indigenous fishes. It is widely documented that one common wetland
fish-carp, Cvprinus carpio-resuspends deposited sediments and in doing so, may alter community structure
of wetland plants and invertebrates, as well as fish. Since the feeding and reproductive habits of most fish
are well-documented, it might be possible to detect gross sedimentation by the density-weighted ratio of
sediment-feeding/breeding species to intolerant species.
Vegetation removal. Removal of canopy of forested wetlands generally results in increased algal production
and possible increases in herbaceous wetland plants. Removal of submersed macrophytes (e.g., "aquatic weed
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control") may similarly increase algae. As vegetation is thinned, herbivorous fish species can increase and
those that depend on macrophytes for cover can decrease disproportionately (e.g., Homer and Williams 1986,
Wiley et al. 1984). However, total abundance and biomass may change little (e.g., Boschung and O'Neil,
Mikol 1985, Wile 1978), and with the removal of vegetation the juveniles of some species may become
more vulnerable to predation (Peterson 1982).
In contrast, if submersed vegetation becomes too dense, species richness can decline. For example, Lyons
(1989) presents data supporting the theory that extirpation of many shiner, darter, and minnow species was
caused by invasion and excessive growth (as a consequence of of the exotic milfoil, Mvriophvllum spicatum.
in Lake Mendota, Wisconsin. Some experimental studies of macrophyte removal have shown declines in
total forage fish standing crop, but increases in growth rates, at least initially, of predatory (piscivore) fish;
density of six of eight sunfish species declined while density of two cyprinids increased (Bettoli 1987).
Physical alteration of channel structure within wetlands can reduce fish biomass and total production, both
within riverine wetlands (e.g., Arner et al. 1976, Portt et al. 1986) and within lacustrine wetlands (e.g., Eadie
and Keast 1984). Species assemblages also shift. Poe et al. (1986) suggested that percid-cyprinid-
cyprinodontid assemblages had a stronger need for diverse habitats and a lower tolerance for habitat
alteration than did assemblages of centrarchids. These investigators found that the percid-cyprinid-
cyprinodontid assemblage dominated an area with an undisturbed littoral zone, high water quality and high
species richness of aquatic macrophytes. A nearby altered site with bulkheaded shoreline, dredged area,
degraded water quality, and low species richness of aquatic macrophytes was dominated by a centrarchid
assemblage. Moring et al. (1985) found brook trout to be particularly sensitive to canopy removal in
western floodplain wetlands. Brook trout were replaced by a greater dominance of white sucker, northern
redbelly dace, blacknose dace, creek chub, and common shiner.
Abundance of fish larvae in a southeastern floodplain swamp stream was found to be 16 times higher in
macrophyte beds than in open channels during the daylight hours (Paller 1987). Durocher et al. (1984)
found a highly significant positive relationship (P<0.01) between percent submerged vegetation and
largemouth bass (Micropterus salmoides). Any reduction below 20 percent of the total lake coverage of
vegetation caused a decrease in recruitment and standing stock of bass.
Turbidity/Shade. Increased turbidity, especially when it occurs over extended periods, generally decreases
fish species richness and alters species composition (Menzel et al. 1984). Slight or moderate, seasonal
increases in turbidity may or may not change fish density and biomass. Species commonly associated with
elevated turbidity include carp, carpsuckers, black bullhead, green sunfish, and others (Menzel et al. 1984).
Species apparently intolerant of elevated turbidity include fantail darter, smallmouth bass, northern
hogsucker, rosyface shiner, hornyhead chub, southern redbelly dace, black redhorse, brook stickleback
(Menzel et al. 1984) and many others listed in Plafkin et al. (1989). Also see above discussions of
sedimentation/turbidity and vegetation removal.
Thermal Alteration. No quantitative information was found concerning the effects of thermal alteration on
community structure of fishes specifically in wetlands. A shift toward warmer-water assemblages, e.g., carp,
downstream from heated discharges seems inevitable.
Inundation/Dehydration. Virtually all fish depend on shallow-water habitats (i.e., generally wetlands) at
some point in their life history. Some species depend more strongly than others on shallow areas and
floodplain wetlands for feeding and reproduction. The proportion of highly-dependent species could
theoretically be used as one indicator of hydrologic alteration of a wetland system.
Inundation alters the spatial and temporal distribution of suitable habitat, with unpredictable effects on
floodplain-dependent species. Effects depend in part on habitat structure and soil chemistry of the areas
being flooded, and whether inundation increases the exposure of isolated populations to predators or
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aggressive competitors. In southeastern floodplain wetlands many fish species benefit if water levels remain
stable during the spawning period following seasonal inundation (e.g., Liston and Chubb 1984, Miranda et
al. 1984). In the Florida Everglades, stable water levels resulted in increased fish community richness,
diversity, biomass, average size of fish, and proportion of carnivorous species; however, fish density decreased
(Kushlan 1976). In Mississippi River floodplain ponds, "days flooded" was the most significant factor in a
multivariate regression for explaining total community biomass and biomass of catastomids, clupeids,
crappies, cyprinids, and ictalurids; flooding in the sampled wetlands ranged from 24 to 115 days annually,
with a mean of 81 (Cobb et al. 1984).
Dehydration reduces wetland fish diversity if it results in (for example) stranding of fish, anoxic conditions,
cutting off of access, increased vulnerability to terrestrial predators, reduced area of productive periodically
flooded areas, or altered food supply. However, periods of higher precipitation that follow droughts (or
periods of inundation following partial drawdown) can result in increased fish production in wetlands; this
could be due to increased nutrient availability or temporary elimination by drought of large competing or
predatory invertebrates such as dragonfly larvae (Freeman 1989).
Where hydrologic alterations occur, the seasonality of their effects is critical in determining the effect they
will have on fish community structure. Species considered by Mundy and Boschung (1981) to be most likely
to decline with impoundment in Alabama floodplain wetlands were as follows: Bluehead Chub, Striped
Shiner, Creek Chub, Creek Chubsucker, Frecklebelly Madtom, Crystal Darter, Scaly Sand Darter, and Redfin
Darter.
Species that are most dependent on wetland portions of larger water bodies might be identified from existing
regional literature (e.g., Crance 1988, Giese et al. 1987, Kwak 1988, Liston and Chubb 1984, Ross and Baker
1983, Tarplee 1975, Walker et al. 1985b) as well as from results of several ongoing studies of floodplain fish
communities, e.g., studies being conducted by the Cooperative Fisheries Research Unit at Auburn University;
the Corps of Engineers Waterways Experiment Station in Vicksburg, Mississippi; the U.S. Geological Survey
in Tallahassee, Florida, and others.
Wetlands that normally contain surface waters but then are briefly dehydrated can, upon reflooding, support
exceptionally high productivity and biomass of fish (Wegener et al. 1974, Welcomme 1979). However, this
assumes fish have access into and out of the wetland as water levels change, and that sediments do not
contain significant levels of oxidizable contaminants. Severely fluctuating water levels (i.e., causing repeated
exposure of sediments every few hours or days) associated with hydropower generation or boat wakes can
kill fish larvae (Holland 1987).
Fragmentation of Habitat. We found no explicit information on wetland fish community response to
fragmentation of regional wetland resources. One can surmise that as the distance between wetlands
containing fish becomes greater, and/or hydrologic connections become severed by dehydration or dams,
species most dependent on floodplain habitats and/or which do not disperse easily might be most affected.
The magnitude of the effect may depend on the size and intrinsic habitat heterogeneity of the wetlands that
are being fragmented.
Availability of patches of relatively unaltered habitat with natural flow regimes, such as may occur in lower-
order tributaries, can help sustain mainstem fish populations even when mainstem habitats are periodically
subjected to pollution or extreme hydrologic alteration (e.g., Gammon and Reidy 1981). The distances
between such patches may be important. In streams, individual non-anadromous fish over the course of a
year seldom disperse more than a kilometer (Hill and Grossman 1987); however, substantially greater
mobility (frequent movements of up to 12.7 km) was reported for fish inhabiting North Carolina floodplain
wetlands (Whitehurst 1981).
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In lakes, fish species diversity increases with increasing surface area and length of shoreline (Barbour and
Brown 1974, Moyle and Cech 1982, Tonn and Magnuson 1982), probably as a result of increased habitat
heterogeneity and thermal stratification (Eadie and Keast 1984).
Other Human Presence. Sport and commercial fishing comprise an obvious impact to certain wetland fish
species, in some cases at the population level.
9.2 SAMPLING METHODS AND EQUIPMENT
Some factors that could be important to measure and (if possible) standardize among wetlands when
monitoring anthropogenic effects on community structure of fishes include:
hydrologic access, water depth, winter ice cover, conductivity and baseline chemistry of
waters and sediments (especially pH and dissolved oxygen), sediment type, current velocity,
fishing pressure (harvest), stream order or ratio of discharge to watershed size (riverine
wetlands), shade, amount and distribution of cover (logs, undercut banks, etc.), ratio of open
water to vegetated wetland, and the duration, frequency, and seasonal timing of regular
inundation, as well as time elapsed since the last severe inundation or drought.
Methods for sampling fish communities are described in Kushlan 1974b, Nielsen and Johnson 1985, Plafkin
et al. 1989, Welcomme 1979, and many others.
Often, fish can be found in wetlands only during certain seasons of the year. If wetlands can be sampled
only once, then the period just after seasonal rise in water levels, if it coincides with favorable temperatures,
is usually recommended. In most regions, numbers of easily identifiable fish will be greatest late in the
season due to annual recruitment of juveniles. However, caution is needed to time sampling to coincide
with phenologies of particular taxa. Significant, regular events of fish life histories include migration,
dispersal, territory establishment, spawning, and development (Brooks 1989).
Larval fish sampling is best accomplished at night to minimize sample bias due to fish avoiding the sampling
gear (Chubb and Liston 1986). Schramm and Pennington (1981) also suggested nighttime sampling and
showed a maximum of larvae at dusk, high diversity at night and dawn, and low diversity in the daytime.
Nighttime samples were particularly important for collection of hiodontids, ictalurids, and percichthyids.
Equipment used in wetlands for fish sampling potentially includes seines, nets, trawls, electrofishing,
ichthyocides, and various types of pot gear (Hocutt 1978, Nielsen and Johnson 1985, Plafkin et al. 1989).
For sampling larval and egg stages, push-nets and modified plankton nets are often used (e.g., Meador and
Bulak 1987), while in dense vegetation, suction pumps and light traps are often used. A study by Pardue
and Huish (1981) evaluated techniques for collecting adult fish in forested wetland streams, and found that
no single technique collected all species. Thus, they are best used in combination. Scientific collecting
permits, available from state fish agencies, are generally required.
Electrofishing. temporarily stuns fish and thus allows them to be scooped into a bucket, identified and
measured, and quickly released. Electrofishing equipment is commercially available, and permits for
scientific collecting are typically required from state agencies. If the sampled wetland has clearly defined
inlets and outlets, these may be blocked with nets to prevent fish from escaping ahead of the electrical field.
Repeated passes are typically made. Electrical currents are not always used to stun fish; they may also be
used to guide fish into nets or block their escape from a seining area (Nielsen and Johnson 1985).
Electrofishing can quickly obtain fish from many wetland habitats that are difficult to sample with nets, e.g.,
undercut banks, submersed plant beds. For quantification, data are best expressed as number of fish per
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unit area shocked. However, quantitative accuracy is good only for narrow, non-turbid channels. Morgan
et al. (1988) reported that the effectiveness of electrofishing generally decreased as plant density or turbidity
increased, due to the difficulty in locating and retrieving stunned fishes. Backpack shockers may be too
bulky and unsafe for use in wetlands with extensive debris, very soft substrate, and/or ice. Boat-mounted
shockers are limited by shallow water depths, debris, and ice.
It is often difficult in wetlands to confine the area being sampled, so fish may flee the advancing electrical
field. Also, fish stunned by shockers are not necessarily representative of the general fish community.
Collections tend to be biased toward larger individuals and species. Larvae are not captured. Some studies
suggest that catchability of fish declines with successive passes through a wetland, with the effects lasting up
to 24 hours.
Sampling efficiency can also be influenced by water quality. Pulsating, direct current (DC) units are effective
in perhaps the widest range of conditions, but in the "soft" waters of many wetlands (particularly bog
streams), AC units with outputs exceeding 500 volts might work just as well. Some investigators in small,
confined soft-water wetlands have increased shocker effectiveness by placing salt blocks in the water, which
increases conductivity. Extremely high conductivity can reduce effectiveness as well. Bosserman and Hill
(1985) found that shockers were not effective in waters made highly conductive by acid mine drainage.
Seines, are robust nets, several meters long and with a width usually equal or greater than water depth, that
are pulled by people or boats through shallow areas to confine and capture fish (often by herding them
toward shore). When aquatic plants and debris interfere, seines can instead be placed in adjoining open
areas and fish herded into the seines for capture (Nielsen and Johnson 1985). Seines are too ineffective for
accurately estimating fish densities, but may allow a fair estimation of species richness and of relative
dominance of species. Leidy and Fiedler (1985) used a 3 m long seine of 6 mm mesh to sample shallow
streams. Ohio streams were sampled using a 4 ft x 8 ft "Common Sense" minnow sein with 1/8 inch mesh;
about 30 seine hauls were required for thorough sampling (Tramer and Rogers 1973). For wetlands, Hocutt
(1978) recommended the 5 ft x 10 ft "Common Sense" seine with 1/8 inch mesh. A mesh size of 1/8 inch
mesh was recommended to capture smaller species and/or life stages. In situations where larger fish may
outswim smaller seines, monofilament gill nets can be used for seining.
Sweep nets, (dip nets) can be used to capture fish as well as invertebrates. They can be effective for
qualitative sampling in very confined, shallow, clearwater pools. Walker et al. (1985b) had limited success
when dip-netting floodplain fish immobilized with spotlights at night. Studies using sweep nets include those
by Leidy and Fiedler (1985) and Chubb and Listen (1986).
Other types of nets, are used to catch wetland fish, in a passive manner. All nets tend to be selective due
to their design and thus usually provide the best catch results when used in combination. Fyke nets have
been used to sample fish in wetlands (Nielsen and Johnson 1985, Swales 1982, Tonn 1985). Wetland
vegetation is sometimes removed in a small area to make room for the net. Gill nets can be used to take
a variety of fish and can be adapted to different depths (Hocutt 1978). These nets are very effective in
wetlands, and are highly selective for particular size classes and species (Pennington et al. 1981). Gill nets
of five mesh sizes between 2.54 and 12.7 cm were placed near shore in Ohio riverine areas by Hassel et al.
(1988) and checked after 24 hours. Gill net selectivity produced catches dominated by relatively large species
such as longnose gar and channel catfish. Trammel nets purportedly are less size selective than gill nets
(Pennington et al. 1981), but select for fish species with rough surfaces and protrusions (Nielsen and
Johnson 1985). Although traditional trawl nets are not effectively used in wetlands, Herke (1969) described
a boat-mounted push-trawl useful for sampling marshes.
Lift nets, constructed with rectangular frames, hoops, or spreaders are set on the bottom below the water
surface, then lifted to capture small schooling fish (Nielsen and Johnson 1985). Camp, Dresser and McKee
(1989) reported on a lift net specifically designed for use in forested wetland systems. The 1 meter square
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net was made of two weighted PVC loops (a top and bottom) with netting of black fiberglass screen. When
fully extended, the bottomless net measured 39.4 inches x 39.4 inches x 36 inches with a 6 inch flap along
the base. The bottom frame was attached to the substrate and the top frame was connected to a rope and
pulley system to allow the trap to be sprung (lifted) from a remote location without frightening any fish
within the net. Small portable drop nets were used by Freeman et al. (1984) to sample fish in a heavily
vegetated freshwater wetland. These collected significantly more fish per unit area than did seining. Large
drop nets suffer problems of mobility, and when designed to be portable, create disturbance by movement
and shadows (Freeman et al. 1984).
Pot gear, (fish traps) of wood, wire mesh, and/or acrylic plastic have been routinely used by several
experimenters. Traps can be used in a variety of areas of moderate depth and/or heavy cover, and when
baited, are strongly selective for particular species and size classes (Pennington et al. 1981). Studies that
used fish traps in wetlands include Finger and Stewart (1988), Tonn and Magnuson (1982), Walker et al.
1985b.
Ichthyocides are poisons (preferably biodegradable) that can be used to destructively sample the entire fish
population of a wetland. They are undoubtably the most efficient tools for obtaining both quantitative and
qualitative fish samples. However, when used by inexperienced collectors, problems may outweigh benefits
(Hocutt 1978). Examples of use in wetlands include studies by Durocher et al. (1984) and Walker et al.
(1985b).
Also, radiotelemetric methods can be used to track individuals (e.g., Savitz et al. 1983) and estimate
potential wetland dependency.
9.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS
In general, quantitative data on wetland fish community structure has not been uniformly collected from a
series of statistically representative wetlands in any region of the country. Thus, it is currently impossible
to state what are "normal" levels for parameters such as fish density, species richness, biomass, Index of
Biotic Integrity (IBI, Karr 1981) and their temporal and spatial variability, in any type of wetland.
A data set that is perhaps the closest to meeting this objective was collected from a series of relatively
pristine Arkansas rivers that are mostly bordered by wetlands (Giese et al. 1987). These fish data were
collected in part from streams flowing through relatively pristine floodplain wetlands, and thus help serve
as a regional baseline for bottomland hardwood wetlands. Although data on fish density were not
developed, up to 36 species per stream were found and community structure of relatively pristine streams
was defined. In nearby Kentucky, a riverine slough wetland supported at least 12 species (Bosserman and
Hill 1985).
In submersed wetland plant beds, up to 255 fish per 10m2 may be present (Morgan et al. 1988). On the
floodplain of the Kankakee River in Illinois, 481 fish were captured during 4800 hours of trapping (Kwak
1988). In one of the few studies of larval fish communities, Chubb and Liston (1986) reported densities of
up to 32.2 larvae per m3 from Great Lakes emergent wetlands.
For stream fish studies, coefficients of spatial variation have ranged from about 50 to 150 percent
(Eberhardt 1978). In submersed vegetation, this coefficient may range from 9 to 80 percent (Morgan et al.
1988). Studies that have compared fish communities among wetlands (spatial variation) have largely been
conducted along the lower Mississippi River, and include:
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Baker et al. 1988, Cobb and Clark 1981, Cobb et al. 1984, Conner et al. 1983, Felley and
Hill 1983, Hall 1979, Lowery et al. 1987, Mathis et al. 1981, Pennington et al. 1980, and
others.
Only a few studies (Clady 1976, Freeman 1989, Kushlan 1976, Lyons 1989) have quantified year-to-year or
long-term variation in fish community structure in wetlands, but conceivably unpublished data may be
available from sites of the U.S. Department of Energy's National Environmental Research Park system, as
well as the following sites of the National Science Foundation's Long Term Ecological Research (LTER)
program (that contain studied wetlands): Illinois Pool 19 site, Illinois-Mississippi Rivers sites, New
Hampshire Hubbard Brook riparian forest, Oregon Andrews Experimental Forest riparian forest, and
Michigan Kellogg Biological Station site. Temporal (year-to-year) variation in western riparian fish
communities was quantified by Platts and Nelson (1988). Although state fishery agencies undoubtedly have
long-term data on average biomass or length of captured game fish, these data may not have been
systematically collected from wetland sites, and do not include all wetland fish species.
Quantitative data on community composition of wetland fish appears to be most available for lacustrine
aquatic bed (herbaceous) wetlands, western riparian wetlands, and southeastern bottomland hardwood
systems. Apparently such data are least available for riverine herbaceous wetlands and for riparian wetlands
in other regions.
Even qualitative lists of "expected" fish in wetlands of various types do not appear to have been compiled,
although regional distribution of fish is relatively well-documented (e.g., Hocutt and Wiley 1988; Lee et al.
1980). Some publications in the "community profile" series of the USFWS (Appendix C) mention particular
taxa known to occur in wetlands, and wetland fish are listed in the ERAPT database (Dawson and
Hellenthal 1986), in Niering (1985), and in the "Vertebrate Characterization Abstracts" database managed
by The Nature Conservancy and various state Natural Heritage Programs. Quantitative data are generally
most available for harvested species, and less available for non-game species.
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10.0 WETLAND AMPHIBIANS AND REPTILES
10.1 USE AS INDICATORS
This discussion addresses the monitoring of "herptiles"--turtles, frogs, salamanders, snakes, crocodilians, and
lizards that occur in wetlands. The life histories and requirements of amphibians differ greatly from those
of reptiles, and species within each group also differ significantly. Most amphibian species and many reptiles
spend all or critical parts of their life in wetlands. However, with only a few exceptions (Brooks and
Croonquist 1990, Corn and Bury 1989), their responses to anthropogenic stressors in wetlands have barely
been studied in the United States at the community level. Most recent ecological research on herptiles can
be characterized as assessments of the occurrence and abundance of particular species in specific micro-
habitats. Advantages and disadvantages of use of herptiles as indicators are shown in Appendix A. A
possible approach for using assemblages of anuran amphibian species (frogs and toads) as indicators of
wetland condition is described by Beiswenger (1988).
Enrichment/Eutrophication. The effects of enrichment on overall community structure of herptiles
apparently have not been documented in wetlands, and indicator assemblages of "most sensitive species"
remain speculative for this stressor. In southern England, Beebee (1987) found that the bullfrog, Bufo
calamita. consistently selected the more eutrophic wetlands.
Organic Loading/Reduced DO. The effects of severe organic loading, e.g., from wastewater outfalls, on
overall community structure of herptiles apparently have not been documented in wetlands, and indicator
assemblages of "most sensitive species" remain undefined for this stressor. Toxicological data were reviewed
by Birge et al. (1980). Anderson (1965) noted that a moderate amount of sanitary sewage pollution
seemingly increased the dominance of soft-shelled and snapping turtles in parts of the Missouri and
Mississippi Rivers, but heavy industrial waste nearly eradicated turtles for miles downstream, especially the
Ouachita map turtle, in part a mollusk-eater.
Contaminant Toxicity. The effects of heavy metals, pesticides, oil, and other contaminants on the overall
community structure of herptiles apparently have seldom been documented in wetlands, and indicator
assemblages of "most sensitive species" remain speculative for such stressors. Speculation about causes of
regionwide or even global declines in several wetland amphibians (e.g., northern leopard frog, boreal toad,
spotted frog, tiger salamander in the Rocky Mountains) has often focused on either (a) effects of airborne
contaminants on growth and development of tadpoles (Phillips 1990), or (b) effects of increased ultraviolet-
B radiation as a result of trophospheric ozone depletion, since such declines have been noted in otherwise
seemingly pristine wetlands.
Some laboratory based toxicological data for individual species may be found in USEPA (1986), EPA's
"AQUIRE" database, and the U.S. Fish and Wildlife Service's "Contaminant Hazard Reviews" series that
summarizes data on arsenic, cadmium, chromium, lead, mercury, selenium, mirex, carbofuran, toxaphene,
PCBs, and chlorpyrifos. However, relatively few field data are available for judging which wetland species
are most sensitive.
Acidification. Larval stages of amphibians have been suspected of being highly sensitive to acidification
effects. Although impacts at the species level have most often been reported (e.g., Clark 1986a,b, Corn and
Fogelman 1984), acidification impacts on the overall community structure of herptiles have been documented
in wetlands only recently (e.g., Corn et al. 1989, Leuven et al. 1986). Turner and Fowler (1981) found
significantly fewer species in wetlands with pH of less than 5.5.
A few species, e.g., wood frog (Rana svlvatica). are known to be particularly tolerant of acidic pH in bogs
(Karns 1984). However, most amphibians require a pH of higher than 4.5 to 5.0 for embryo survival and
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metamorphosis (Corn et al. 1989, Freda 1986, Freda and Dunson 1986, Gosner and Black 1957, Karns 1984,
Pierce 1985).
Acidic conditions in surface-mine (constructed) wetlands were implicated as a reason for reduced amphibian
use by Hepp (1987). Based on a single pH measurement from each surface-mine pond, the mean pH at
which various species occurred was given by Turner and Fowler (1981) as follows:
pH # of ponds
Northern Spring Peeper 5.2 16
Pickerel Frog 5.42 11
Red-spotted Newt 5.80 8
Gray Tree Frog 5.% 9
Bullfrog 5.91 6
American/Fowler's Toad 5.97 7
Northern Cricket Frog 6.00 1
Wood Frog 6.25 2
Green Frog 6.26 8
Spotted Salamander 6.32 8
Upland Chorus Frog 6.33 8
Similar types of data are presented by Clark (1986b) for Ontario wetlands.
Most of the true frogs are thought to be especially sensitive to acidic precipitation because they respire
through their skin. During foggy periods such respiration may occur while they are out of the water. At
such times, they may be directly exposed to airborne contaminants.
Salinization. The effects of salinization, e.g., from irrigation return water and oil drilling wastes, on overall
community, structure of herptiles apparently have not been documented in wetlands, and indicator
assemblages of "most sensitive species" remain undefined for monitoring salinization. In softwater lakes
and streams, moderate increases in water hardness and alkalinity can result in increased amphibian densities
(Hepp 1987).
Sedimentation/Burial. Moderate increases in soft bottom sediments can increase habitat for overwintering
turtles. However, excessive sedimentation can smother eggs of many amphibians and alter food sources.
The North American dusky salamander (Desmognathus fuscus) and the spring salamander (Gvrinophilus
porphvriticus) are reportedly very sensitive to effects of bank erosion, sedimentation, and turbidity (Campbell
1974, Orser and Shure 1972). However, the effects of sedimentation/ burial (e.g., of amphibian eggs) on
overall community structure of herptiles apparently have not been documented in wetlands, and indicator
assemblages of "most sensitive species" remain speculative for this stressor.
Turbidity/Shade, Vegetation Removal. Many herptiles are sensitive to the presence and type of vegetation
and its juxtapositioning with open water, particularly in arid regions. In Colorado River riparian zones,
lizards were most abundant in shoreline habitats, moderately dense in riparian habitats, and least dense in
non-riparian or upland habitats; densities depended on insects inhabiting herbaceous debris heaps and litter
piles washed up by the river (Jones and Glinski 1985, Warren and Schwalbe 1985). In more humid Oregon
watersheds, amphibian richness, density, and biomass were less in logged watersheds than in unlogged
watersheds, particularly when vegetation removal occurred primarily in headwater areas (Corn and Bury
1989). Herptile richness was also less in Pennsyvania watersheds with disturbed stream corridors than in
those with intact riparian vegetation (Croonquist 1990). Use of riverine wetlands by herpetofaunas has been
positively related to number of cover types, sinuosity, circumneutral pH, and gradual shoreline slopes (Hill
1986). Richness of breeding frogs may be related also to the variety of herbaceous plant forms in a wetland
(e.g., Diaz-Paniagua 1987).
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The community composition of Minnesota amphibians was found to be correlated with wetland vegetation
form. The leopard frog (Rana pipiens) was found most frequently in sedge mat and less commonly in the
very wet tamarack zone. The wood frog (Rana svlvatica) was found primarily in the fir-ash zone with lesser
numbers in the spruce and tamarack. Spring peeper (Hvla cruciferl and swamp frog (Pseudacris nigrita')
were found in the two zones most distant from the pond, spruce and fir-ash (Marshall and Buell 1955).
Despite these initial efforts, indicator assemblages of "most sensitive species" of herptiles remain speculative
in most of the U.S. for monitoring effects of vegetation removal, and the effects of vegetation removal on
overall community structure of herptiles apparently have not been documented in wetlands.
Thermal alteration. Herptiles, as ectotherms, are particularly sensitive to thermal alteration of wetlands.
Although a vast literature exists describing thermal preferenda of individual species, the effects of thermal
alteration on overall community structure of herptiles apparently have seldom been documented in wetlands.
Lack of comparative studies has resulted in a lack of information on most-sensitive indicator assemblages.
Dehydration/Inundation. Changes in wetland water level alter the quantity and quality of herptile habitat,
and may trigger immigration, emmigration, and breeding of particular species and their predators (Pechmann
et al. 1988). The effects of dehydration may be particularly severe if dehydration occurs during herptile
hibernation, due to the effects of exposure and increased predation of eggs (Campbell 1974).
Impoundment has been reported to increase the regional populations of toads and turtles (Anderson 1965),
or at least causes a shift in spatial distribution of habitat. However, inundation can reduce and alter the
seasonal timing of flooding of downstream habitats. The resultant changes in vegetation and floodplain leaf
litter accumulation can reduce both abundance and diversity of reptiles, as reported by Jones (1988) for
Arizona riparian systems. Also, if inundation causes temporarily flooded wetlands to become connected to
permanent waters, predatory fishes can gain access to the temporary wetlands, perhaps resulting in
reductions in some amphibians (e.g., Dodd and Charest 1988). Temporary dehydration of wetlands may have
the opposite effect, benefitting amphibians by reducing fish predation. The ratio of non-predatory to
predatory salamanders can increase in wetlands following dry springtime conditions (Cortwright 1987).
Herptile taxa that characterize seasonally flooded wetlands or have terrestrial phases appear to resist effects
of urbanization more than those that characterize permanently flooded wetlands or which spend their entire
life cycle in wetlands (Minton 1968). In San Francisco, Banta and Morafka (1966) attributed the decline
of the native California red-legged frog (Rana aurora dravtoni) and the introduced leopard frog (Rana
pipiens) to dehydration and filling of wetlands. Leopard frogs also declined in Colorado as a probable result
of drying up of breeding ponds during a drought (Corn and Fogleman 1984). Vickers et al. (1985), studying
aquatic and semi-aquatic amphibians in northern Florida cypress wetlands, found no change in mean
numbers, numbers of species, or species diversity in ditched versus unditched wetlands. However, species
richness declined and terrestrial species became more abundant with ditching.
Fragmentation of Habitat. We found no explicit documentation of herptile community response to
fragmentation of regional wetland resources, although the presence of some individual species, e.g., spotted
salamander, is known to sometimes depend on proximity to source ponds (Cortwright 1987). One can
surmise that as the distance between wetlands containing herptiles becomes greater, and/or hydrologic
connections become severed by dehydrated channels, dams, or (particularly) roads, species most dependent
on wetlands and/or which do not disperse easily might be most affected (Campbell 1974, Croonquist 1990).
In Oregon, Corn and Bury (1989) found that logging upstream from unlogged habitats had no effect on the
presence, density, or biomass of any species inhabiting the unlogged habitat.
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Other Human Presence. The introduction by humans of non-indigenous aggressive predators (e.g., bullfrog,
snapping turtle, and several predatory fishes) into particular water systems has sometimes led to a decline
in richness of indigenous frog communities (e.g., Hammerson 1982, Hayes and Jennings 1986, Moyle 1973).
10.2 SAMPLING METHODS AND EQUIPMENT
Some factors that could be important to measure and (if possible) standardize among wetlands when
monitoring anthropogenic effects on community structure of herptiles include:
water depth, temperature (site elevation, aspect), conductivity and baseline chemistry of
waters and sediments (especially pH, DO, and suspended sediment), current velocity, stream
order or ratio of discharge to watershed size (riverine wetlands), shade, amount and
distribution of cover (logs, crevices, etc.), ratio of open water to vegetated wetland, extent
of plant litter and rotting logs, vegetation type, and the duration, frequency, and seasonal
timing of regular inundation, as well as time elapsed since the last severe inundation or
drought.
Sampling methods for wetland herptile communities are described in Bury and Raphael 1988, Corn and Bury
1990, Halvorson 1984, Jones 1986, Scott 1982, Vogt and Hine 1982, and others.
Because amphibian distribution and abundance has strong ties to seasonal hydrologic phenomena and the
capability of particular species for dispersal, the temporal and spatial variability in amphibian community
structure strongly reflects these factors. As is the case with sampling macroinvertebrates whose communities
are similarly dependent on ephemeral hydrologic events, sampling amphibian communities can require
several repeated visits to a wetland to fully describe community composition. Nonetheless, Corn and Bury
(1989) assert that, at least for riparian communities of the Pacific Northwest, amphibian population densities
are usually stable in undisturbed habitat and serve as better indicators of habitat quality than do similar
densities of birds or mammals.
Herptiles can be sampled during the mid- to late growing season when maximum numbers of developing
juveniles are present. However, many species are easy to find only after the first few days of rain following
a drought, during late-summer thunderstorms, during the first spring thaw in northern areas, during mid-
day basking hours, or at night (Kaplan 1981). Occasionally, traditional winter hibernation areas can be
located and used to count individuals representing a larger (but undefinable) area. For Arizona, Jones
(1986) noted trapping was most effective in riparian habitats between May and July.
Methods used in wetlands for herptile sampling potentially include pitfall traps (often with drift fences and
baited), visual belt transects, direct capture methods, and vocalization recording.
Pitfall traps and funnels are perhaps the most widely used mechanism for capturing herptiles (Jones 1986).
Animals enter and cannot find the opening to escape. They are subsequently identified, counted, measured,
and released. To reduce loss of trapped animals to predation, traps and funnels can be checked regularly
(at least every other day) and can be shaded, and/or filled with sufficient moist plant litter to minimize
physiologic stress to animals. Pitfall traps are impractical in many wetlands where the water table is so close
to the land surface that pits fill rapidly with water.
Pitfall traps and funnels often produce more species per sampling effort than direct capture methods (Jones
1986). The size of the trap, baits used, and trap placement can affect the species that are caught. Trap and
funnel methods can provide relatively quantitative data, when arranged systematically and level-of-effort (e.g.,
"trap-hours") is standardized.
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They involve emplanting a container in the soil, either on the periphery of the wetland or within it (if
surface water is absent), with the lip of the container placed flush with the ground surface. Herptiles
stumble in and cannot climb the steep sides to escape. Because some species can drown if the container
fills with rainwater, Jones (1986) recommends placing floatable material (e.g., styrofoam) in the container
to reduce mortality.
Funnel openings are usually oriented toward land for greatest effectiveness. Hoop funnel traps are generally
used for turtles, and other funnel traps are used (particularly in deeper wetlands) for catching salamanders,
frogs, and occasionally snakes. A special kind of floating pitfall trap can be used to sample basking turtles
(see Jones 1986 for description). Aquatic turtles in a Missouri marsh were captured using hoop and net
traps. Traps were baited with sardines, local fish, tadpoles, frog, crayfish, dragonfly larvae, snails, and clams
(Kofron and Schreiber 1987).
The efficiency of traps and funnels can be increased by channeling movements of herptiles in the direction
of the trap or funnel. This is commonly done with "drift fences" (Gibbons and Bennett 1974). These are
fences constructed of wire screen or polyethylene plastic, with lengths upwards of 15 m. Traps are placed
at both ends of the drift fence, along the fence at various points, or at the junction of several intersecting
fences. The bottom edge of the fence is emplanted in the ground, or at least no space is provided for
herptiles to crawl under the fence.
Drift fences and pit traps can be more effective and less biased than log-turning, walking transects,
electroshocking in streams, or searching and digging through litter. However, they are expensive; time and
cost estimates for drift fence trapping are provided by Gibbons and Semlitsch (1982). Jones (1986)
comments that, for quantifying herptile communities, drift fence/pitfall trap methods are less effective for
frogs, toads, large snakes, terrestrial turtles, and salamanders than for small snakes and lizards.
Sizes and shapes of containers and associated drift fences and their configurations vary greatly, depending
partly on target species and wetland type. Vickers et al. (1985) sampled in and around cypress ponds using
arrays of four 7.6 m lengths of 0.75 m high, 6 mm polyethylene drift fences arranged perpendicularly and
attached at the center. The fences were held upright with wooden laths and buried to 10 cm depth to avoid
animals passing underneath. Two aluminum screen wire funnel traps, 75 cm long with 20 cm entrance
funnels were placed beside each drift fence. To insure that a trap would always fall on the ecotone
regrardless of pond size, distance between arrays was standardized at one-half the pond radius. Working
in peatland vegetation in Maine, Stockwell (1985) censused herptofauna in eight vegetation types -lagg,
forested bog, wooded heath, shrub heath, moss lawns, pools, streamside meadow, and shrub thicket. Drift
fences of free standing aluminum flashing were used as well as those of lath supported polyethylene. Pitfall
traps were made of two #10 tin cans joined with tape and silicone. Funnels made from margarine tubs were
used in the top of each trap to prevent escape and 2-3 cm of water placed in the bottom of each trap to
prevent desiccation of captives. Similar traps were made by Jones and Glinski (1985) using double-deep 3
Ib. coffee cans with a lid placed 15 cm over the top to prevent desiccation.
The above methods require multiple visits to a wetland, first to set up and later to check traps. Herptiles
can also be monitored directly, that is, during a single visit, or without having to wait for traps to catch
individuals. However, direct methods usually do not provide accurate quantitative data on abundance.
Unless frequent visits are made and the correct microhabitats are searched at the proper times of year,
direct methods are also unlikely to yield good estimates of species dominance or richness. However, they
can provide a useful complement to trap methods, locating species that are not easily trapped.
The simplest type of direct search involves scanning a wetland with binoculars to observe the more obviously
visible species such as basking turtles, frogs, and alligators. In some cases, floating egg masses of amphibians
can also be detected visually and identified to species. Observational methods can be done formally, along
defined transects. Searches on foot, perhaps employing many people shoulder-to-shoulder (e.g., Marshall
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and Buell 1955) have been used, but could be impractical and destructive of habitat in many wetlands.
Long-handled nets can be used to surround logs and rocks as they are lifted to search for herptiles, so as
to catch individual herptiles as they flee. In riverine wetlands, fine-mesh seines (see Fish section above) can
be used for similar purposes.
To enhance opportunities for encountering herptiles during direct searches, electrofishing and identification
of vocalizations and tracks can be used. Electrofishing methods (described in section 9, used in conjunction
with sweep nets or seines, are particularly effective for retrieving larger salamanders and frogs. Because
some species leave distinctive tracks, travel corridors can be searched periodically for tracks. Frogs can
sometimes be located more easily at night, as their eyes reflect in the beam of a flashlight. Vocalizations
of many frogs and toads are easily identified (commercially-available recordings are available to learn these)
and can be used to augment observations. Frogs and toads can sometimes be induced to vocalize by
introducing sharp, loud sounds or played-back tape recordings of vocalizations. Low-altitude overflights or
aerial photography under favorable conditions can be used to identify alligator holes and paths.
10.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS
In general, quantitative data on structure of the entire herptile community of wetlands has not been
uniformly collected from a sufficiently large, statistically-drawn sample of wetlands in any region of the
country. Thus, it is currently impossible to state what are "normal" levels for parameters such as herptile
density, species richness, or biomass, and their temporal and spatial variability, in any type of wetland.
We found only a few published studies that quantified the entire herptile community (or a large proportion
of it) across a region and/or among a set of wetlands:
Brooks et al. 1985, 1987, 1989, Clark 1986b, Congdon et al. 1986, Corn and Bury 1989, Corn et al.
1989, Fowler et al. 1985, Gibbons and Semlitsch 1982, Jackie and Gatz 1985, Karns 1984. Hepp
1987, Jackie and Gatz 1985, Stockwell 1985, Turner and Fowler 1981, Ward 1988.
We found no journal articles that quantified year-to-year variation in the entire community structure of
herptiles in wetlands, but conceivably such unpublished data may be available from sites of the U.S.
Department of Energy's National Environmental Research Park system, and sites of the National Science
Foundation's Long Term Ecological Research (LTER) program. Some studies (e.g., Corn et al. 1989) have
featured qualitative re-checking of wetlands known in previous decades to have particular species, but
probably could not be termed "long-term monitoring."
Quantitative data on community composition of wetland herptiles is virtually lacking from all regions except
the Southeast, Southwestern riparian areas, and parts of the Northeast and Pacific Northwest. Information
on impacts is limited mostly to studies of hydrologic effects and vegetation removal; especially little is
known of impacts from contaminants, salinization, sedimentation, and habitat fragmentation.
Qualitative lists of "expected" herptiles have been compiled by statewide herptile atlas projects in Illinois,
Kansas, Massachusetts, Maine, and perhaps other states, as well as by less comprehensive surveys in various
smaller areas. Species that show highest affinity for wetlands of various types might be identified by
consulting with local herpetologists, Niering (1985), and the "Vertebrate Characterization Abstracts" database
managed by The Nature Conservancy and various state Natural Heritage Programs. Limited qualitative
information may be available by wetland type from some of the "community profile" publication series of
the USFWS (Appendix C).
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11.0 WETLAND BIRD COMMUNITIES
11.1 USE AS INDICATORS
This discussion addresses the monitoring of wetland birds, e.g., waterfowl, shorebirds, wading birds, and
wetland-dependent songbirds. The use of birds as environmental indicators is discussed by Morrison (1986),
Reichholf (1976), and particularly, by Temple and Wiens (1989). Statistical aspects of regional bird trend
analysis are discussed in Sauer and Droege (1990). Advantages and disadvantages of using birds as
indicators are summarized in Appendix A.
Because most vertebrates use wetlands at some time during the year, defining what truly constitutes a
"wetland-dependent" bird species is difficult. One could argue dependency based on diet,
energetics/metabolism, requirement for a particular structural component found only in wetlands, or duration
of seasonal use. As with some wetland plant groups, many degrees of dependency occur, from species that
spend their entire life in wetlands to species that use wetlands opportunistically and/or for only brief periods.
Species that may be casual users of wetlands of a particular type in one region may be obligate users of a
different type in the same region, or of the same type in a different region. Dependency in highly altered
landscapes may be less related to the intrinsic characteristics of wetlands than to the fact that little other
undeveloped habitat remains, forcing species that normally occur less frequently in wetlands to use what
remains, regardless of its condition. In such cases, bird density may be a poorer indicator of habitat quality
(the ability of the habitat to sustain successfully reproducing individuals over the long-term) than
measurements of population demographics or measurements made at the organism level (Van Home 1983).
An empirical approach for testing wetland-dependence of birds is demonstrated by Finch (1990).
Monitoring of wetland birds, particularly waterfowl, has been extensive in many regions. Wetland birds can
be categorized as (a) those most strongly dependent on larval insects, non-insect aquatic invertebrates,
amphibians, fish, and submersed plants, and (b) those most strongly dependent on adult (terrestrial)
invertebrates, emergent plants, and rodents. In general, the former group-which includes waterfowl and
wading birds-tends to respond more immediately to contamination and water level changes than does the
latter group-which usually includes marsh wrens, certain warblers, red-winged blackbirds, and swallows.
Diets (and thus, guild assignments) of particular species can be confirmed through stomach content analysis
or, less destructively, through close-range, automated photography of nest visits. In general, though, habitat
requirements, life histories, and species assemblages of wetland birds are relatively well-known. Still,
information on community-level response to particular stressors has been difficult to collect, in part because
most bird species--as very mobile organisms—may be better at integrating overall landscape conditions than
they are at indicating the conditions in a particular wetland.
Enrichment/Eutrophication. The effects of enrichment on overall community structure of birds are poorly
documented in wetlands, and indicator assemblages of "most sensitive species" remain mostly speculative for
this stressor. Weller and Spatcher (1965) defined a species assemblage that inhabits a "late marsh"
successional stage, and species that inhabit the upland transitional zones of wetlands are well-known.
However, the dominance of these species assemblages may be related as much to physical factors
(geomorphology, fire, extreme climate events) as to nutrient enrichment. For Great Lakes wetlands,
Crowder and Bristow (1988) hypothesized the following series of events that might lead to a waterfowl
decline as a result of eutrophication:
"For the waterfowl, the effect of inshore eutrophication is thus an initial increase in food
plants, a gradual replacement of favorite species by less desirable plants, and finally a total
loss of submersed and floating-leaved plants coincident with an extension of cattail marsh.
The extended marsh in turn declines, having been exposed to wave erosion through loss of
the deeper zones of vegetation."
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However, not all aquatic plants that increase with eutrophication are poor waterfowl foods. For example,
in a study at Lake Okeechobee, Florida, Johnson and Montalbano (1984) found that waterfowl diversity in
Hydrilla beds (a widespread, exotic species) was significantly greater than in several indigenous wetland plant
communities (bulrush, cat-tail, pondweed, spikerush, and others).
Organic Loading/Reduced DO. The effects of severe organic loading, e.g., from wastewater outfalls, on
overall community structure of wetland birds have been investigated in a few cases. Generally, abundance
and/or on-site diversity of songbirds (Brightman 1976, Hanowski and Niemi 1989) and sometimes waterfowl
(Belanger and Couture 1988, Piest and Sowls 1985) have tended to increase with increased abundance of
aquatic invertebrates. The effect may depend on the type and configuration of the particular wastewater
treatment system (Fuller and Glue 1980). Other bird groups have responded more to water levels (and
associated effects on vegetation and invertebrates) than to contamination status (e.g., Ramsay 1978). In the
Houghton Lake, Michigan, wetland that was exposed to treated wastewater, Kadlec (1979) reported no major
shifts over a 3-year period in bird abundance or species composition.
Where introduction of organic wastes results in anoxic conditions lethal to fish and some amphibian larvae,
community composition may shift from fish-eating species (e.g., herons, loons, grebes) to invertebrate-
eating species and opportunists (e.g., shorebirds, songbirds, gulls, terns). Indeed, migrant shorebirds and
gulls often appear to concentrate at sewage lagoons, turf farms, and wetlands mildly polluted with organic
wastes (e.g., Campbell 1984, Fuller and Glue 1980).
Contaminant Toxicity. The effects of bioaccumulation of contaminants in wetland bird tissues have been
widely measured, and the disasterous effects of naturally-occurring toxicants on community structure of
wetlands have occasionally been documented (see discussion of Salinization below). Species assemblages for
indicating the physical effects of oil spills can be easily identified based on characteristic behaviors of some
wetland birds. However, the effects of pesticides, heavy metals, and other contaminants on overall structure
of wetland bird communities are poorly documented in wetlands, and indicator assemblages of "most
sensitive species" remain mostly speculative for these stressors (Grue et al. 1986).
Bird reproductive failure in wetlands from effects of heavy metal contamination (e.g., Scheuhammer 1987,
Kraus 1989) and pesticides have been documented, but only for a few species. Lethal thresholds for metals
and synthetic organics are reported in Hudson et al. (1984), EPA's "AQUIRE" database, and the US Fish
and Wildlife Service's "Contaminant Hazard Reviews" series that summarizes data on arsenic, cadmium,
chromium, lead, mercury, selenium, mirex, carbofuran, toxaphene, PCBs, and chlorpyrifos. However,
relatively few field data are available for judging which wetland species are most sensitive. Additional testing
of chemical toxicity to wildlife is currently being sponsored by EPA
Numerous anecdotal reports exist describing relatively stable bird assemblages in traditionally-used wetlands
even after years of progressive contamination. This might be attributed to the loss of nearby wetlands that
otherwise would have been preferred, to behavioral avoidance of contaminated microenvironments and foods,
and/or to replacement of contaminated individuals by immigrants.
Acidification. Naturally acidic wetlands sometimes have lower densities and species richness of birds,
particularly in winter, than do non-acidic wetlands (Brewer 1967, Ewert 1982). Bird use of acid mine
drainage wetlands in Pennsylvania was found to be less than use of natural wetlands, probably because of
physical degradation of habitat rather than inferior water quality alone (Hill 1986). Acidification has also
been demonstrated to reduce reproductive success and juvenile survival of some species in wetlands (e.g.,
tree swallows-Blancher and McNichol 1988, black ducks-DesGranges and Hunter 1987, ring-necked ducks-
-McAuley and Longcore 1988). Bird responses to anthropogenic acidification, summarized by McNichol et
al. (1987), are felt indirectly as a result of alteration in the dominance of various food sources and possibly,
changes in physical habitat (e.g., composition and distribution of submersed macrophytes). Shifts from fish
to aquatic insects in lakes and streams can cause a corresponding shift from fish-eating species to those that
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critically depend on aquatic invertebrates, to those that feed on aquatic invertebrates opportunistically
(assuming other habitat features remain relatively constant). Wetland bird groups in each category are
listed in Table 14. Strong presence of a particular feeding group relative to others might be used to suggest
acidification effects, if the role of other stressors (such as others listed in this section) can be ruled out.
Table 14. Examples of Wetland Birds Categorized by Major Food Source.
Predominantly feeding on fish or amphibians (at some season or life stage, in some regions):
loons, grebes, cormorants, anhinga, some herons and egrets, terns, bald eagle, osprey, kingfishers
Aquatic invertebrate obligates (at some season or life stage, in some regions):
some herons and egrets, diving ducks, some dabbling ducks, bitterns, rails, shorebirds,
yellow-headed blackbird
Aquatic invertebrate facultatives:
most dabbling ducks, swallows, marsh wrens, common yellowthroat, red-winged blackbird, many other
songbirds (see Adamus 1987 for list for the Northeast)
Salinization. Breeding waterfowl in hypersaline wetlands reportedly prefer fresher portions of these
wetlands, and inland wetlands that are naturally saline generally have fewer nesting waterfowl (Kantrud and
Stewart 1977). However, high densities of a few species, e.g., Northern Phalarope, can occur during
migration in some naturally saline wetlands. The effects of salinization on structure of wetland bird
communities have not been widely studied, despite publicity given to events such as the catastrophic
mortality at Kesterson National Wildlife Refuge. Assemblages of species that might be used as indicators
of salinization remain speculative.
Sedimentation/Burial; Turbidity/Shade. Wetland bird species that prefer soft-bottomed wetlands can be
defined, but probably with insufficient precision to warrant their use as indicators of excessive sedimentation.
Sedimentation affects community structure of wetland bird communities primarily by altering the type and
distribution of submersed plants, and perhaps also by affecting invertebrate food sources and interfering with
feeding of birds that rely on visual cues.
Vegetation Removal. Effects of vegetation removal associated with grazing and/or fire are described by
Fritzell 1975, Landin 1985, Schultz 1987, and others summarized by Kantrud (1986). "Moderate" levels of
grazing and/or mowing, if occurring at a time in the season when nests are not disturbed, can increase
wetland bird species richness in floodplain ponds (Landin 1985) and emergent wetlands (Nelson and Kadlec
1984). However, severe grazing, mowing, or fire at inappropriate times is detrimental (Duebbert and Frank
1984, Kantrud and Stewart 1984), and total removal of woody riparian vegetation dramatically alters species
composition, density, and richness of the mammalian community (Cross 1985, Malecki and Sullivan 1987,
Possardt and Dodge 1978).
Many species benefit from increased openings in dense stands of vegetation and from reduced floodplain
ground cover, while others, including ground-nesting species such as Northern Harrier and Short-eared Owl
(USDA Soil Conservation Service 1985), do not. As patches of open water are created in formerly
continuous stands of emergent vegetation, the diversity of species using a wetland typically increases (Harris
et al. 1983, Kaminski and Prince 1981). These may be species that are generally widespread in the region,
so the contribution of vegetation removal to overall regional diversity of birds may be slight. Species
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assemblages associated with vegetation structural changes can be defined by region and wetland type. Brown
et al. (1989), and Durham et al. (1985) have done so for vertebrates in bottomland hardwood wetlands, and
Short (1983, 1989) for midwestern and Arizona wetlands.
Effects of silvicultural activities in forested wetlands have received only limited study. Birds in forested
wetlands respond very strongly to changes in vertical and horizontal vegetation structure (Finch 1990, Rice
et al. 1980). Because the habitat structural needs of most forested wetland birds are relatively well-known,
at least qualitatively (e.g., see Durham et al. 1985, Swift et al. 1984), indicator associations could probably
be easily developed that reflect bird response to different levels and types of silvicultural practices in forested
wetlands. An old-growth forested floodplain wetland in South Carolina was compared by Hamel (1989) to
clearcut and selectively cut portions of the same area. More species (and particularly cavity-nesters)
achieved their highest densities in the old-growth habitat than in the disturbed forested wetland, and those
species that did achieve higher density in the disturbed forested wetlands were widespread throughout the
region. In a southwestern riparian wetland, Carothers et al. (1974) reported 46 percent fewer breeding birds
where vegetation had been thinned to 25 trees per acre, as compared to a similar reference wetland with
116 trees per acre.
Thermal alteration. The effects of thermal alteration on overall community structure of birds are poorly
documented in wetlands, and indicator assemblages of "most sensitive species" remain mostly speculative for
this stressor. Effects of heated wastewater are mostly indirect, affecting habitat and bird distribution by
prolonging ice-free conditions in northern wetlands, altering vegetation type and structure, and affecting the
type and seasonal availability of food sources (e.g., Haymes and Sheehan 1982). On a regional level, species
most sensitive to changes in temperature are often those occurring at the periphery of their geographic
ranges. These are easily defined by local ornithologists.
Inundation/Dehydration. The response of bird community structure to water level alteration has been the
subject of dozens of studies, many conducted to improve the management of waterfowl habitat. Water level
alterations (either increases or decreases) can increase or decrease overall bird abundance and richness,
depending on their duration and many other factors.
Both sustained increases and sustained decreases in water levels directly affect habitat availability and
dramatically shift community composition. For example, construction of dams on the lower Colorado River
produced a relatively stable environment that favored high invertebrate densities and consequently increases
in diving ducks, but diminished numbers of riparian species (Anderson and Ohmart 1988). Alteration of
the flooding regime of a southern forested wetland from seasonal flooding to permanent flooding (for a
greentree reservoir) had little overall effect on bird diversity; waterfowl and common grackles increased while
white-throated sparrow and a few other species decreased (Newling 1981).
Water level changes of short durations (weeks or months), while having less affect on habitat availability,
have the potential for long-term impacts to habitat quality by altering water chemistry, invertebrate
populations, and seed germination. For example, dam-induced alterations in hydrologic regime have
decreased bird richness partly by encouraging the spread of non-native salt cedar (Tamarix spp.)(Ohmart et
al. 1977, Hunter et al. 1985).
Addition of permanent open water to a non-permanently flooded wetland usually increases the opportunity
for use by waterfowl and fish-eating birds. Moreover, the typical increase in submersed and floating-leaved
plants that accompanies creation of a permanent pool within a wetland provides for a more diversified plant
and invertebrate food source. This consequently can result in an increase in on-site species richness of birds.
Many studies have found that productivity and diversity of waterbirds are greatest within basins having a
permanently flooded pool or channel that is surrounded by shallowly flooded (<10 inches depth) wetlands
that are gradually dehydrated at regular seasonal or frequent (3-5 year) annual intervals (Fredrickson and
Taylor 1982, Reid 1985). Among a series of Massachusetts forested wetlands, Swift et al. (1984) found that
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the driest wetlands supported the lowest abundance and richness of birds, even though in some regions such
wetlands have the greatest diversity of vertical habitat structure and plant species richness.
Wetland bird species vary in their water depth requirements and sensitivity to water level change. Much
information on depth requirements is summarized in Fredrickson and Taylor (1982) and Fredrickson and
Reid (1986). This information could be used to define hydroperiod "response guilds" of birds. The most
sensitive species may be those which (a) nest along water edges (e.g., Western Grebe, Redhead-Wolf (1955),
or (b) feed on mudflats (e.g., shorebirds), or (c) require a particular combination of wetland hydroperiod
types in a region (e.g., Kantrud and Stewart 1984). In contrast, species with nests typically well-above water
level (e.g., marsh wren, prothonotary warbler) may be less vulnerable. For arid, deep-water marshes in
eastern Oregon, Littlefield and Thompson (1989) suggested that presence of yellow-headed blackbird might
be a good indicator of ecologically "healthy" conditions.
Fragmentation of habitat Only a single study (Brown and Dinsmore 1986, 1988) has looked directly at the
effects of fragmenting regional wetland resources. Others had previously noted the effects of fragmentation,
using knowledge of species-specific life histories or data from non-wetland forest systems. Essentially, as
the distance between wetlands containing certain species becomes greater, and/or hydrologic connections and
vegetation corridors become severed by dehydrated channels, bank-clearing, or (particularly) roads, species
most dependent on wetlands and/or which do not disperse easily could be most affected. Moreover, some
species require not just a particular density of wetlands, but a particular combination of wetland types (or
wetland types and other land cover types) at a particular density on the landscape or in close proximity to
each other (Cowardin 1969, Kantrud and Stewart 1984, Ohmart et al. 1985, Weller 1979, Rake 1979,
Patterson 1976). Although individual birds, being highly mobile, can disperse to new areas having the
proper combination of types at a sufficient density, this can cause diminished reproductive success and thus,
non-sustainable populations.
Territorial size requirements of wetland birds are highly variable, but can be used (with empirical
observations of presence/absence in wetlands of various sizes and degrees of isolation) to define assemblages
of species that are likely to be most sensitive to habitat fragmentation (Brown et al. 1989). Such studies
must employ a standard level of effort (e.g., censusing time) per unit area if results are to be comparable.
Radiotelemetric methods can be used to track individuals and determine home range sizes under various
combinations of landscape cover patterns (Hegdal and Colvin 1986 describe techniques).
Other Human Presence. Several studies (e.g., Brooks et al. 1990, Robertson and Flood 1980, Todt 1989)
have reported changes in species composition of wetland bird communities in response to general watershed
"development," reduction in natural land cover types surrounding the wetlands, and increased visitation of
wetlands by humans. Developed areas are characterized by a typical suite of species that include European
Starling, Rock Dove, American Crow, House Sparrow, American Robin, and perhaps a few others (Graber
and Graber 1976).
Human disturbance can discourage use by wildlife (Pomerantz et al. 1988), especially (a) hunting (Conroy
et al. 1987, Gordan et al. 1987) and people traveling on foot (Burger 1981), and (b) during the breeding
season or under harsh weather conditions. Effects of noise disturbances on wildlife are summarized by
Gladwin et al. (1988). The most sensitive species appear to be ducks, geese, and other long-distance
migrants which feed in large flocks at the ground or water level (Burger 1981), as well as colonially-nesting
species (e.g., Markham and Brechtel 1979, Tremblay and Ellison 1979) and large species (e.g., Stalmaster
and Newman 1978). Sensitivity to human disturbance may also be species-specific. Reduced use of human-
visited wetlands by waterfowl or nongame waterbirds has been demonstrated by Hoy (1987), Josselyn et al.
(1989), and Kaiser and Fritzell (1984). To some extent, presence of screening vegetation can permit closer
approach to waterbirds by humans (e.g., Milligan 1985).
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Many waterbirds take flight when humans approach within 75 to 175 feet (e.g., Josselyn et al. 1989).
Wintering bald eagles may take flight when approached from a distance of 800-1,000 feet (Knight and
Knight 1984; Stalmaster and Newman 1978). Motorboat activities can disturb waterfowl up to 1,000 m away
(Hoy 1987). This results in more time being spent in energetically costly behaviors. Disturbance can also
increase the food consumption needs of waterbirds. Korschgen et al. (1985) found that only 5 boating
disturbances per day increased the energy requirements of canvasbacks by 20 percent, requiring consumption
of an additional 23 g of food daily.
Other direct human influences on wetland birds include mortality from collisions with vehicles and
powerlines, and predation by hunters and housecats. Hunting comprises an obvious impact to certain
wetland bird species, in some cases resulting in changes at the population level.
11.2 SAMPLING METHODS AND EQUIPMENT
Some factors that could be important to measure and (if possible) standardize among wetlands when
monitoring anthropogenic effects on community structure of birds include:
distribution of water depth classes, vegetation (type, and vertical and horizontal diversity and
arrangement), conductivity and baseline chemistry of waters and sediments (especially
conductivity), current velocity, distance and connectedness to other wetlands of similar or
different type, surrounding land cover (particularly within 500 feet of wetland perimeter),
shoreline slope, wetland size, ratio of open water to vegetated wetland and its spatial
interspersion, and the duration, frequency, and seasonal timing of regular inundation, as
well as time elapsed since the last severe inundation or drought.
Methods for surveying bird communities are described by Burnham et al. 1980, Halvorson 1984, Ralph and
Scott 1981, Verner 1985, Verner and Ritter 1985, and others. Censusing marsh and shorebirds specifically
is discussed in detail by Connors (1986) and Weller (1986); censusing of waterfowl by Eng (1986) and Kirby
(1980); censusing of colonial waterbirds by Speich (1986); and censusing of birds in bottomland hardwood
wetlands, by Durham et al. (1985). An effort to refine techniques for monitoring wetland birds is presently
being sponsored by the Maine Department of Inland Fisheries and Wildlife.
Even when apparently similar wetlands are censused, it is sometimes impossible to attribute changes in
wetland bird communities to human activities within the wetland being sampled, because birds move widely
across regions and continents. However, by calculating density-weighted ratios of declining resident to
declining non-resident species (with similar habitat requirements), the possible role of this factor might be
estimated.
Birds are present in wetlands throughout the year, but densities of birds vary greatly by season, depending
on region. As with many other taxa, if only a single annual visit can be made, it should be timed to account
for major life history events, such as nesting, molting, dispersal, migration, or wintering. The most severe
reductions in bird density and richness occur in winter in northern emergent wetlands and bogs that
completely freeze over. In southern wetlands, density and diversity are generally greatest in winter, while
in northern wetlands, density and diversity are usually greatest in summer (Harris and Vickers 1984).
During spring and fall, large numbers and high diversity may be present in either northern or southern
wetlands. The species richness of wetlands in arid regions often increases the greatest during spring and fall,
as many species seek temporary refuge during migration (e.g., songbirds in riparian oases, shorebirds in
flooded fields).
A survey covering several wetlands should occur simultaneously or within consecutive days, unless severe
weather conditions intervene. If the objective is to compare between-year trends in a species, total species,
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or species richness, then simple count methods (e.g., transects) are probably appropriate. However, if the
objective is to rank wetland types or relative abundance of species, more time-consuming censusing to
develop estimates of density are required (Steele et al. 1984). Determination of indices of relative annual
abundance, rather than exhaustive population censusing, is suitable for most purposes (Emlen 1981).
For reasonably accurate estimates of breeding bird richness in a wetland, three visits spread over the
breeding season may be desirable (Brooks et al. 1989, Weller 1986). Sampling non-wetland environments,
Steele et al. (1984) reported that three repetitions of a 2 km transect were adequate to estimate bird
abundance and richness of a habitat. In an inventory of birds in 87 Maine wetlands, Longcore et al.
(pers.comm.) counted birds from an overview for two hours at dawn and two hours at dusk on at least two
dates; as many observation points as necessary to view the entire wetland were used. The actual number
and duration of visits required in a particular instance will depend on the size of the wetland, its habitat
heterogeneity, visibility, and other factors.
If not only richness, but density, must be determined, then at least eight visits may be needed (Ralph and
Scott 1981). Although most common songbirds will not be disturbed by frequent visits by monitoring
personnel, raptors, waterfowl, other large or colonial species, and ground-nesting species may be susceptible.
Wetland songbird surveys are commonly conducted in during May - July, when breeding birds are most
detectable by song.
Species detection (especially of most songbirds) is greatest during early morning hours. However, thrushes
and a few other species are more detectable in the evening, and in winter, some species may be most active
at mid-day. Night-time coverage may be warranted, not only for typically nocturnal species such as owls,
but also for waterfowl and wading birds which use different wetland types for roosting and feeding.
Secretive species (e.g., rails, some passerines) have sometimes been surveyed more effectively by playing back
of tape recorded calls, use of predator decoys, use of dogs, and by dragging ropes or chains through wetlands
(e.g., Glahn 1974, Ralph and Scott 1981).
Surveys may be conducted from ground level, from elevated observation posts, or aerially. In the case of
species that nest or roost colonially and in exposed locations, photography may be used to assist counting
of individuals. Ground-level, visual techniques cannot be used effectively in wetlands with tall vegetation
(mid-season emergent marshes, forested wetlands). Boats are typically used for surveys of wetlands wider
than about 100 meters, as visibility from shore, even using a spotting scope, becomes restricted.
Many methods have been developed for monitoring wetland bird communities using visual, auditory, and
capture techniques. These include point counts, line transects, nest counts, mist netting, and regional
surveys (Brooks et al. 1989). Methods differ mainly in the degree of quantification they provide, the level-
of-effort required, and the taxa they are most effective in censusing. These methods can be used in virtually
all types of wetlands.
11.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS
Quantitative community-level data on birds have not been uniformly collected from a series of statistically
representative wetlands in any region of the country. Thus, it is currently not possible to state what are
"normal" levels in wetlands of various types for parameters such as bird density, species richness, biomass,
or productivity. Data on temporal and spatial variability of wetland birds among wetlands and years has
been systematically collected in only a few instances. These few data sets are available largely because of
the existence of two important national data collection networks, which are described as follows.
The Breeding Bird Survey (BBS) database has existed since 1966, and includes all 50 states and some
Canadian provinces. Data on bird relative abundance have been collected, usually recurrently, from about
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2500 transects ("routes"), each 25 miles in length and containing 50 evenly-spaced data collection points.
Density of coverage varies from 1 to 16 routes per degree (latitude-longitude) block. The survey routes are
not located to intentionally intersect wetlands, so wetlands are included only randomly. Routes are run only
once annually, so many species may be missed. Also, some routes are conducted later in the season than
is optimal for detecting some wetland species. Because routes follow roads and rely largely on auditory
detection more suitable for forest birds, they may further underestimate wetland species. Nevertherless, the
BBS database, by its sheer quantity of spatial and temporal coverage, represents a valuable resource for
helping define "average" bird densities (in relative terms) and for aiding detection of regional trends in
wetland birds. Locations of routes are included on the state maps in Appendix B.
The Breeding Bird Census (BBC) database also provides useful information. This database is a compilation
of individual censuses conducted by volunteers throughout the United States. Compared to methods used
by the BBS, the BBC protocols are more intensive, but coverage is not nearly as extensive. Whereas the
BBS measures only relative abundance using a single annual visit to an area, the BBC attempts to measure
population density using repeated visits. The BBC also differs from the BBS in that some habitat data are
collected, but habitat heterogeneity within census plots is not quantified, the acreage of censused plots is
not consistent among censuses, and only a small portion of the plots are revisited annually. In most cases,
census plots are too small and heterogeneous to adequately census species with large home ranges (Terborgh
1989), as is typical in wetlands. A few of the BBC's have focused exclusively on wetlands, but these wetlands
have not been chosen randomly or systematically. These are included on the state maps in Appendix B.
Selected data are presented in Tables 15 and 16, located at the end of this chapter. These are based on
data compilation conducted by the Cornell Laboratory of Ornithology and sponsored by the EPA Wetlands
Research Program. These tables are summarized in the following paragraphs.
Median number of breeding species ranged from 3.5 for all censused Florida wetlands to 51 for all censused
Montana wetlands, where the national maximum of 68 species was found in one censused wetland (a
bulrush-cattail marsh). As expected, salt marshes at all locations had the lowest number of breeding species.
The greatest variability in species richness occurred among a set of 21 Wisconsin wetlands, a set of five
Kansas wetlands, and a set of seven Florida wetlands. Most repetitively-censused wetland types (NUM>1)
had less than 15 percent variation in species richness among years, and less than 10 percent variation in pair
density among years.
Median density of breeding birds (i.e., pairs per square kilometer) ranged from 138 in Alaskan wetlands to
1857 in North Carolina wetlands. The two highest densities of all counts were from riparian willow
woodlands in California. One, with 4547 pairs and 35 species, was dominated by Chesnut-backed Chickadee,
Bewick's Wren, Song Sparrow, and Yellow Warbler. The investigator attributed the high density to extreme
density of vegetation and abundant food, despite low plant diversity. The other remarkable California
riparian count, 3208 pairs per krn^ and 13 species, was dominated by Mourning Dove, Lazuli Bunting,
Bewick's Wren, and Wilson's Warbler. Other high densities were in a California lacustrine marsh (3684
pairs, mainly Tricolored Blackbird), and in a cattail bulrush wetland in North Dakota (3418 pairs, mainly
Yellow-headed Blackbird). The greatest variability of pair density among censused wetland types occurred
among a set of four Nebraska wetlands (114 percent).
Table 15 summarizes the same parameters for each state/province, but does so by individual years of census.
With regard to number of species, during a given year most states had less than 38 percent variability among
their wetlands. Within any single year, the greatest variability in species richness among censused wetland
types occurred between 2 Florida wetlands in 1983, which differed by 110 percent. With regard to pair
density, during a given year most states had less than 54 percent variability among their wetlands. The
greatest variability of pair density among censused wetland types occurred among 3 Colorado wetlands in
1973, which differed by 144 percent.
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In general, analysis of these 478 census plots showed the following statistically significant (p<0.05), linear
relationships, based on log-transformed data:
o the median number of species was correlated with pair density and number of repeat censuses
(years) on a plot;
o variability in number of species was inversely correlated with number of species;
o the median pair density was not correlated with number of repeat censuses (years) on a plot;
o variability in pair density was correlated with pair density and number of repeat censuses (years)
conducted on a plot;
o variability in pair density was correlated with variability in number of species.
Despite their statistical significance, there was considerable scatter in all of these relationships, and the
correlation coefficients (r) never exceeded 0.5.
Published studies (other than from the national databases described above) that have compared year-to-
year or long-term variation in bird community structure in wetlands include Bellrose 1979, Blake et al. 1987,
Harris et al. 1983, Hanowski and Niemi 1987, and Rice et al. 1980. Conceivably some unpublished data on
annual variation in wetland bird communities may be available from sites of the U.S. Fish and Wildlife
Service's Northern Prairie Research Station, the U.S. Department of Energy's National Environmental
Research Park system, and the Illinois Pool 19 and Illinois-Mississippi Rivers sites of the National Science
Foundation's Long Term Ecological Research (LTER) program.
Other national bird databases exist, and new ones are being developed, for example:
o International Shorebird Survey
o Christmas Bird Count database
o Colonial Wading Bird database
o Monitoring Avian Productivity (MAP) database
o Winter Bird-population Censuses
o Migratory Waterfowl Surveys
o Mid-winter Waterfowl Survey
o breeding bird atlases in dozens of states
None of these pertain exclusively to wetlands, and it is not always possible to separate the portion of the
data that includes wetlands. Still, on a collective basis, these databases could be analyzed to yield more
information on community structure in different regions and occasionally, in different wetland types.
Overviews of some are provided by Muir and Davis (1989) and Terborgh (1989).
Lists of breeding wetland birds have been compiled by "block" (a unit generally smaller than about 50 sq.
mi.) by statewide atlas projects in many states, and along with data from Christmas Bird Counts, other
104
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databases listed above, and records kept by thousands of volunteers, these can be used to define "expected"
species in wetlands. Species that show highest affinity for wetlands of various types might be identified in
discussions with local birders and by accessing the "Vertebrate Characterization Abstracts" database managed
by The Nature Conservancy and various state Natural Heritage Programs. Limited qualitative information
may be available by wetland type from the "community profile" publication series of the USFWS (Appendix
C).
Quantitative data are most available for harvested groups, like waterfowl, and least available for the majority
of wetland species, which are not harvested. In a survey of waterfowl migration/ wintering habitat in the
United States, Bellrose and Trudeau (1988) reported the following to represent at least "moderate" densities
of waterfowl (number of birds per acre per day):
Dabbling Bay
Ducks Divers Geese
Atlantic Flyway 0.17 0.36 0.26
Mississippi Flyway 0.44 0.06 0.13
Central Flyway 0.73 0.09 0.34
Pacific Flyway 2.87 0.21 0.41
Of studies that have compared bird community structure among many wetlands in a region (spatial
variation), perhaps the most notable for their large sample sizes are those of bottomland hardwoods by
Durham et al. 1985, and prairie potholes (Kantrud and Stewart 1984, Stewart and Kantrud 1973). The latter
study--of 1321 wetlands-reported the following mean densities:
Density Density
Wetland (pairs/km2) (pairs/
class wetland^ N
Ephemeral 200 1 4
Temporary 633.1 .76 190
Seasonal 431.8 3.52 808
Semipermanent 723.8 39.92 168
Permanent 38.6 30.80 14
Alkali 52.1 33.59 8
Fen 673.5 37.12 11
Undifferentiated
tillage 89.3 0.09 118
Other quantitative studies of multiple wetlands include:
Anderson and Ohmart 1985, Blake et al. 1987, Brewster et al. 1976, Briggs 1982, Brooks
et al. 1987, 1989, Brown and Dinsmore 1986, DesGranges and Darveau 1985, Evans and
Kerbs 1977, Flake et al. 1977, Hardin 1975, Harris and Vickers 1984, Heitmeyer and Vohs
1984, Hepp 1987, Hill 1986, Hudson 1983, Hunter et al. 1985, Klett et al. 1988, Knopf 1985,
Landin 1985, Lawrence et al. 1985, Mack and Flake 1980, Maki et al. 1980, Menzel et al.,
Milligan 1985, Ohmart et al. 1985, Rector et al. 1979, Rice et al. 1980, Smith 1953, Stauffer
and Best 1980, Swift et al. 1984, Wheeler and Marsh 1979, and others.
105
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In summary, quantitative data on community composition of wetland birds is most available for breeding
populations and least for wintering and migrating populations. Perhaps least-studied are montane wetlands;
Northwestern wetlands; southeastern and southwestern herbaceous wetlands; and southern Great Plains
wetlands. Information on impacts is most available for hydrologic alteration, vegetation removal, and
acidification. Apparently the least information is available on impacts to community structure from
eutrophication, sedimentation, contamination, and habitat fragmentation.
106
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Tables 15 and 16 (Introduction).
These data are presented to quantitatively illustrate the variability that exists within a particular resource
(wetlands). Data in this table might be used, with caution paid to the several limitations described below,
to place data from a newly studied wetland into a context of other wetland studies described here. The
summary metrics used here-richness and density-are only two of many community metrics that might be
used for such purposes.
Explanation:
Data shown in these tables were collected by volunteers with diverse capabilities and without use of a strictly
standardized protocol. Lack of standardization of study area size, and inclusion of species whose home
ranges are often larger than the 10-20 mean size of most of these wetlands, introduces a significant bias into
the data set. Each record below consists of a single breeding bird census, involving multiple visits during
the breeding season of a single year, sometime during the period 1937-1988. The number of visits per
season and the size of the censused areas varies greatly among these reported data. In Table 15, records
where "NUM" >1 are sites that were visited during multiple, usually contiguous, years ("NUM" is the
number of years visited). In Table 16, records where "NUM">1 are years where more than one wetland
subtype in a state was visited.
These particular records were selected by the Cornell Laboratory of Ornithology as being ones most likely
to include wetlands and riparian areas. This table does not contain ALL breeding bird censuses conducted
in U.S. wetlands. Conversely, some records in this table may be from predominantly non-wetland habitats,
or from nest colonies where densities may be atypical of overall wetland habitat. Phrases (in the
•SUBTYPES' column) used to describe the sites were assigned by individual volunteers familiar with the
sites, and no standardized wetland classification scheme was used. Detailed information on vegetative
composition and bird species composition of most sites is available in the journal American Birds. Other
columns of the tables are defined as follows:
Table 15:
MED_SPP: The median number of species, for all years when the same site was censused; when the site
was censused only one year (NUM=1), the median is the cumulative total of all species
found that year.
MIN_SPP: The minimum number of species in any year; when the site was censused only one year
(NUM= 1), the minimum is the cumulative total species found from all visits that year.
MAX_SPP: The maximum number of species in any year, when the site was censused only one year
(NUM = 1), the maximum is the cumulative total species found from all visits that year.
CV_SPP: The among-year coefficient of variation for all years when the same site was censused; when
the site was censused only one year (NUM=1), no CV was calculated.
MED_DEN, MIN_DEN, MAX_DEN, CV_DEN: Same as above, but applying to density (number of
breeding pairs per km-).
Table 16:
MED_SPP: The median number of species, for all wetlands in the state (NUM) that were censused
during the named year, years in which only one site was censused were excluded.
MIN_SPP: The minimum number of species; for all wetlands in the state that were censused during
the named year.
MAX_SPP: The maximum number of species; for all wetlands in the state that were censused during
the named year.
CV_SPP: The among-wetland coefficient of variation for all wetlands when several were censused the
same year.
MED_DEN, MIN_DEN, MAX_DEN, CV_DEN: Same as above, but applying to density (number of
breeding pairs per square kilometer).
107
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12.0 WETLAND MAMMAL COMMUNITIES
12.1 USE AS INDICATORS
In general, wetlands are permanently inhabited by fewer mammal species than are upland ecosystems.
However, the association of some mammals with wetlands is very strong. These include river otter, muskrat,
nutria, beaver, mink, raccoon, swamp rabbit, marsh rice rat, and others. In contrast to most wetland birds,
many wetland mammals are herbivores or omnivores, i.e., they consume wetland plants directly or have a
mixed animal-plant diet. Muskrat in particular can have major impacts on wetland herbaceous plants (e.g.,
McCabe 1982). Advantages and disadvantages of using mammals as indicators are summarized in Appendix
A.
As with birds, because a majority of mammals use wetlands at least briefly at some time during the year,
defining what truly constitutes "wetland-dependent" is difficult. For example, individual bobcats and black
and grizzly bears use wetlands extensively in some regions (e.g., Helgren and Vaughn 1989), but it is
sometimes unclear whether this is the general preference of the species, and if so, whether alternative
habitats infrequently visited by humans are suitable substitutes. Some species of mammals have been
categorized according to wetland dependency by Brooks and Croonquist 1990, Durham et al. 1985, and
Fritzell 1988.
In one comparison of existing data, prairie pothole wetlands were reported to support fewer species of
mammals than either northern bogs/fens, or southern bottomland hardwoods (Fritzell 1988). Response to
particular stressors is described below.
Enrichment/Eutrophication. The effects of enrichment on overall community structure of wetland mammals
has not been documented, and indicator assemblages of species "most sensitive" to eutrophication remain
speculative.
Organic Loading/Reduced DO. Attempts have been made in a few instances to measure the effects of
severe organic loading, e.g., from wastewater outfalls, on overall community structure of wetland mammals.
However, results generally have been equivocal and indicator assemblages of species "most sensitive" to
organic loading remain speculative.
It can be hypothesized that, where introduction of organic wastes results in anoxic conditions lethal to
mammal foods (e.g., fish and some amphibians), community composition may shift from fish-eating species
(e.g., otter, mink) to vegetarian or invertebrate-eating species and opportunists (e.g., muskrat, opossum).
Contaminant Toxicity. The effects of bioaccumulation of contaminants in wetland mammal tissues have
sometimes been measured. Species assemblages for indicating the physical effects of oil spills can be easily
identified based on characteristic behaviors of some wetland mammals. However, the effects of pesticides,
heavy metals, and other contaminants on overall structure of wetland mammal communities are poorly
documented in wetlands, and indicator assemblages of "most sensitive species" remain mostly speculative for
these stressors.
Acidification. Effects of acidification on the overall community structure of wetland mammals apparently
have not been documented and indicator assemblages of "most sensitive" species remain speculative. It can
be hypothesized that, where acidification becomes severe, community composition may shift from fish-
eating species (e.g., otter, mink) to vegetarian or invertebrate-eating species and opportunists (e.g., muskrat,
opossum).
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Salinization. The effects of salinization, e.g., from irrigation return water and oil drilling wastes, on overall
community structure of mammals has not been documented in wetlands, and indicator assemblages of "most
sensitive species" remain speculative.
Sedimentation/Burial. Excessive sedimentation can alter food sources of wetland mammal communities.
However, the effects of sedimentation/ burial on overall community structure of wetland mammals has not
been documented, and indicator assemblages of "most sensitive" species remain speculative.
Vegetation Removal. Many mammals are sensitive to the presence and type of vegetation and its
juxtapositioning with open water. Species richness of small mammals in wetlands has been correlated with
complexity of vegetation structure (Arner et al. 1976, Landin 1985, Maki et al. 1980, Nordquist and Birney
1980, Stockwell 1985, Searls 1974, Simons 1985). Vegetation removal and associated long-term destruction
of den sites in both wooded and emergent wetlands has resulted in changes in furbearer populations and
small mammal communities (Krapu et al. 1970, Malecki and Sullivan 1987, Possardt and Dodge 1978), while
restoration of riparian vegetation has led to increases in use by mink (Burgess and Bider 1980). However,
many small mammals are more abundant in the denser herbaceous ground cover that results from overstory
removal, as shown in a Texas riparian system by Dickson and Williamson (988). Grazing at levels
recommended by the Soil Conservation Service had no significant effect on abundance or distribution pattern
of small mammals in a Colorado cottonwood floodplain (Samson et al. 1988).
Species in Iowa considered by Geier and Best (1980) to be least tolerant of vegetation change include
Microtus pennsylvanicus. Spermophilus tridecemlineatus. Reithrodontomvs megalotis. Peromvscus
maniculatus. and Mus musculus. Species considered "moderately tolerant" included Sorex cinereus and
Blarina brevicauda. The Eastern chipmunk (Tamias striatus) and white-footed mouse (Peromvscus leucopus)
were considered the most tolerant in Iowa, and this was also found to be true in the Vermont study of
Dodge et al. (1976) and Possardt and Dodge (1978). Species considered most sensitive to riparian
vegetation removal in Vermont were jumping mice (Zapus hudsonicus and Napeozapus insignis) and shrews
(Blarina brevicauda. Sorex cinereus).
Geier and Best (1980) predicted that a reduction in shrub cover would reduce populations of T. striatus and
S. cinereus. T. striatus would be especially affected by the selective removal of eastern red cedar.
Populations of T striatus, Peromvscus leucopus, and the two shrew species would suffer from the loss of
woody plant debris (logs, brushpiles, and stumps).
Despite these initial efforts, indicator assemblages of mammals "most sensitive" to vegetation removal remain
speculative in most of the U.S., and the effects of vegetation removal on overall community structure of
mammals have not been well-documented in wetlands.
Thermal alteration. The effects of thermal alteration on overall community structure of mammals
apparently have not been documented in wetlands, and "most-sensitive" indicator assemblages remain
speculative.
Dehydration/Inundation. Changes in wetland water level and soil moisture alter the quantity and quality
of mammal habitat, and may trigger immigration and emmigration of particular species. The effects of
dehydration may be particularly severe if they occur during hibernation, due to the effects of exposure. In
northern wetlands, muskrats, for example, require deep water in winter for successful hibernation (Bellrose
and Low 1943). Although muskrats and minks appeared to tolerate temporary flooding in an Illinois
forested floodplain, opossums, red foxes, gray foxes, striped skunks, and woodchucks were evicted by flood
conditions (Yaeger 1949).
In northern Florida cypress ponds, Harris and Vickers (1984) found an increase in relative abundance of rice
rats and a decrease in cotton rats with any addition of water. In a series of Maine bogs, species richness
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of small mammals was highest in the driest part of the bog, near the upland edge (Stockwell 1985). In
prairie pothole wetlands, small mammals select habitats based on soil moisture levels (Pendleton 1984). In
Colorado (Olson and Knopf 1988), mammal species richness, relative diversity, and faunal similarity were
greater in upland communities than in riparian wetlands. Richness was also less in Washington riparian
areas than in adjoining uplands, although presence of water-placed woody material within the wetter areas
mediated this effect to some degree (Mason 1989).
Fossorial mammals (e.g., moles and shrews) that inhabitat subsurface areas may be particularly sensitive to
moisture level changes. However, local changes in moisture regimes and other aspects of wetland habitat
quality are frequently not reflected by indicator species of mammals because of the ability of mammals to
move freely, in and out of impacted areas.
Despite these initial efforts, indicator assemblages of mammals "most sensitive" to habitat dehydration or
inundation remain speculative in most of the U.S., and the effects of these stressors on overall community
structure of mammals have not been well-documented in wetlands.
Fragmentation/Isolation of Habitat Although habitat fragmentation has been widely implicated in the
decline of some large mammals, we found little explicit documentation of overall mammal community
response to fragmentation of regional wetland resources. One can surmise that as the distance between
wetlands containing wetland-dependent mammals becomes greater, and/or hydrologic connections and
vegetated corridors become severed by dehydrated channels, bank-clearing, or (particularly) roads, the more
sensitive mammals or those which do not disperse easily might be most affected. Although individual
mammals, being highly mobile, can disperse to new areas having the proper combination of wetland types
at a sufficient density, they probably do so at risk of greater predation and energetic loss.
Sensitive species can be grouped into "guilds" that exhibit similar responses to fragmentation. For example,
Brooks et al. (1989, 1990) found significant differences in mammal communities in disturbed vs. undisturbed
watersheds, and recommended that stream corridors be at least 100 m in width. Home range sizes of
wetland mammals have also been used for defining wildlife guilds and required buffer strip sizes (Brown et
al. 1989). However, home range sizes can vary greatly by season and habitat type. They can be determined
from observations of presence/absence in wetland patches of various sizes and degrees of isolation, or by
using radiotelemetry (Hegdal and Colvin 1986 describe techniques).
12.2 SAMPLING METHODS AND EQUIPMENT
Some factors that could be important to measure and (if possible) standardize among wetlands when
monitoring anthropogenic effects on community structure of mammals include:
distribution of water depth classes, vegetation and woody debris (type, and vertical and
horizontal diversity and arrangement), current velocity, distance and connectedness to other
wetlands of similar or different type, surrounding land cover (particularly within 500 feet
of wetland perimeter), wetland size, ratio of open water to vegetated wetland and its spatial
interspersion, and the duration, frequency, and seasonal timing of regular inundation, as well
as time elapsed since the last severe inundation or drought.
Methods for surveying mammal communities are described in Cooperrider et al. (1986), Halvorson (1984),
and others.
Mammals occur in wetlands throughout the year. Mammal density and richness may be reduced during and
immediately after floods in riverine wetlands. Surveys covering several wetlands, if not conducted
simultaneously, should occur within consecutive days, unless severe weather conditions intervene. For
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efficient censusing, advantage can be taken of species that congregate seasonally in wetlands (e.g., white-
tailed deer in northern cedar swamps). Diurnally, detection of most species is greatest at night. Visual
surveys of larger, day-active species can be conducted from ground level, from elevated observation posts,
or aerially. Low-altitude overflights or aerial photography can be used to identify some beaver dams and
beaver and muskrat lodges, and to census moose and large mammals in open country. Ground-level, direct-
observation techniques cannot be used effectively in wetlands with tall vegetation (mid-season emergent
marshes, forested wetlands).
Many methods have been developed for monitoring wetland mammal communities, and generally rely on
various types of traps. Tracks, scat, den trees, burrows, vocalizations, eyeshine, and other sign may also be
counted using point counts, line transects, or similar methods. Some species can be attracted to scent
stations or salt blocks. Most non-capture methods can be used in virtually all types of wetlands. Methods
differ mainly in the degree of quantification they provide, the level-of-effort required, and the taxa they are
most effective in censusing. Thus, whenever possible a variety of methods should be used.
Spring-loaded snap traps, live (cage) traps, pitfall traps, and funnel traps are widely used for capturing
mammals. Animals are attracted by bait or, in the case of pitfall traps, stumble into a confining pit and
usually cannot escape. They are subsequently identified, counted, measured, and released. To reduce loss
of trapped animals to predation, traps and funnels are checked regularly (at least every other day) and can
be shaded, and/or filled with sufficient moist plant litter to minimize physiologic stress to animals.
The efficiency of traps and funnels can be increased by channeling small animal movements in the direction
of the trap or funnel. This is commonly done with "drift fences" (Gibbons and Bennett 1974). These are
fences constructed of wire screen or polyethylene plastic, with lengths of at least 5-15 m. Lengths less than
2.5 m are not very effective (Bury and Corn 1987). Traps are placed at both ends of the drift fence, along
the fence at various points, or at the junction of several intersecting fences. The bottom edge of the fence
is emplanted in the ground, or at least no space is provided for non-burrowing animals to crawl under the
fence. Sizes and shapes of containers and associated drift fences and their configurations vary greatly,
depending partly on target species and wetland type. Trap and funnel methods can provide relatively
quantitative data, when arranged systematically and level-of-effort (e.g., "trap-hours") is standardized.
The size of the trap, baits used, and trap placement can affect the species that are caught. Thus, a variety
of methods should be used if possible (Szaro et al. 1988). Snap traps are effective for cricetids and many
other small rodents (e.g., meadow vole, short-tail shrew, house mouse, western harvest mouse, masked
shrew)(Geier and Best 1980), whereas pitfall traps are more effective for rodents that are primarily
insectivorous and/or fossorial (moles and shrews)(Szaro 1988). Funnel traps are ineffective in capturing
many forest mammals (Bury and Corn 1987). If only a single type of capture method can be used and the
aim is to capture the widest variety of small mammals, then in Pacific Northwest forests, Bury and Corn
(1987) recommend use of pit traps over a continuous 60-day period; a list of the most common species
could be compiled by using pitfall traps only for a typical 10-day trapping period. However, the high water
table in many wetlands can render pitfall traps impractical due to flooding. In these situations, spring-
loaded traps mounted on floating platforms are effective for detecting some species (pers. comm., T.
Roberts, Waterways Experiment Station, Vicksburg, MS).
Examples of community-level mammal studies in wetlands include, for example:
Cross 1985 (Oregon), Geier and Best 1980 (Iowa), Landin 1985 (Mississippi), McConnell and
Samuel 1985 (West Virginia), Olson and Knopf 1988 (Colorado), Scelsi (n.d.)(New Jersey), and
Urbanek and Klimstra 1986 (Illinois).
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12.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS
In general, quantitative data on structure of the entire mammalian community of wetlands has not been
uniformly collected from a series of statistically representative wetlands in any region of the country. Thus,
it is currently impossible to state what are "normal" levels for parameters such as mammal density, species
richness, or biomass, and their temporal and spatial variability, in any type of wetland.
We found only a few published studies that quantified the entire mammalian community (or a large
proportion of it) among a set of wetlands: Brooks et al. 1985, 1987, 1989, Geier and Best 1980, Landin
1985, Nordquist and Birney 1980, Pardue et al. 1975, Stockwell 1985, and Urbanek and Klimstra 1986
(Illinois).
We found no journal articles that quantified year-to-year or long-term variation in mammalian community
structure in wetlands, but conceivably such unpublished data may be available from sites of the U.S.
Department of Energy's National Environmental Research Park system, sites of the National Science
Foundation's Long Term Ecological Research (LTER) program, and regional studies of the ELF military
communications facility (Blake et al. 1987).
Quantitative data on composition of wetland mammalian communities is virtually lacking from all regions
except parts of the Northeast and some riparian systems. Information on impacts is limited mostly to
studies of hydrologic effects and vegetation removal; especially little is known of impacts to community
structure from contaminants, salinization, sedimentation, and habitat fragmentation.
Qualitative lists of "expected" mammals in wetlands can be easily developed in most regions from Niering
(1985), Fritzell (1988), and the "Vertebrate Characterization Abstracts" database managed by The Nature
Conservancy and various state Natural Heritage Programs. Limited qualitative information may be available
by wetland type from some of the "community profile" publications of the USFWS (Appendix C).
However, fine gradations in degree of dependency of individual species upon wetlands have not been defined.
Quantitative data are most available for harvested species, while the majority of wetland mammals, which
are not harvested, are seldom studied quantitatively in wetlands.
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13.0 BIOLOGICAL PROCESS MEASUREMENTS IN WETLANDS
13.1 USE AS INDICATORS
This discussion addresses biological processes that are commonly monitored in inland wetlands. "Processes"
here are considered to be synonymous with wetland "functions." Included are litterfall and decomposition,
nutrient translocation, growth and production, and respiration. We have limited consideration mainly to
studies where these processes have been monitored for an entire wetland, not just a dominant species or
community within the wetland. Relatively few studies have monitored wetland biological processes in the
context of evaluating a specific anthropogenic stressor, as has Mader et al. (1988).
Although an understanding of wetland processes and their vulnerability to anthropogenic stressors is
fundamental for predicting future impacts, the limited evidence to date suggests that biological processes
usually respond only weakly and slowly to stressors in wetlands. This may be because biological processes
represent the net result of many potentially compensating mechanisms within biological communities
(Schaeffer et al. 1988, Schindler 1987). In contrast to changes in community structure which tend to occur
gradually, changes in processes, when they ultimately occur, may occur suddenly and catastrophically.
Perhaps with further testing and development of new ways to measure and quantify biological processes,
their utility to regulatory monitoring programs will increase. Advantages and disadvantages of use of
wetland ecosystem processes as indicators of ecological condition are shown in Appendix A
Enrichment/Eutrophication. The effects of enrichment on annual productivity, decomposition and
denitrification have been studied primarily in cypress dome and northern bog wetlands. Responses are
generally typical of what has been found in other aquatic systems-increased productivity with "moderate"
enrichment and a decline in productivity with "severe" enrichment.
Effects of enrichment on decomposition rates are highly variable, with both increased decomposition and
no effect reported (Almazan and Boyd 1978, Andersen 1979, Chamie 1976, Fairchild et al. 1984, Farrish and
Grigal 1988, Meyer and Johnson 1983, Richardson et al. 1976). Differing conclusions may be due to
differences in current velocity, leaf type, temperature, fertilizer type, ambient water quality, and other factors.
Enrichment of wetlands with nitrogen-rich runoff may lead to an increased proportion of nitrous oxide
release (vs. N2 release), which is of potential concern because even small changes in the production of
nitrous oxide are potentially significant considering the role of this gas in destroying stratospheric ozone
(Hahn and Crutzen 1982).
Enrichment commonly increases secondary production. For example, aquatic invertebrate production was
correlated with enrichment (total phosphorus concentration) in Plante and Downing's (1989) analysis of
aquatic bed community data from 51 lakes (164 samples) from temperate regions of the world.
Organic Loading/Reduced DO. The effects of severe organic loading, e.g., from wastewater outfalls, on
annual productivity have been studied primarily in cypress dome and northern bog wetlands, and results were
similar to the above. With regard to decomposition, Brinson et al. (1981) reviewed the available literature
and concluded that decomposition in wetlands should occur most rapidly with aerobic conditions under some
optimum regime of wetting and drying; alternating conditions of aerobic and anaerobic result in slower
decomposition.
Contaminant Toxicity. The literature summary by Baath (1989) reports heavy metal-induced impairment
of several microbial processes, such as respiration, phosphatase enzyme activity, denitrification (Grant and
Payne 1982) and decomposition of leaf litter (Jackson and Watson 1977), in wetland soils. In one case
enrichment has been demonstrated to mitigate toxicity effects (Fairchild et al. 1984). In general, the relative
toxicity of metals to microbial processes decreases in the order Cd> Cu> Zn> Pb (Baath 1989). Cadmium
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in shrub wetlands can interfere with nitrogen fixation (Wickliff et al. 1980). The effects of metals on
primary and secondary production, and the effects of other contaminants on other processes, have not been
widely studied in wetlands.
Acidification. The effects of acidification on biological processes have generally not been studied in inland
wetlands. Long-term decomposition rates, particularly of the most refractile litter components, are generally
slower in acidic water bodies (Friberg et al. 1980), and few kinds of decomposer bacteria operate effectively
below pH 4 (Doetsch and Cook 1973). Artificial acidification has been shown to decrease the
decomposition rate of litter from an herbaceous wetland plant (Leuven and Wolfs 1988), but the degree of
inhibition may depend on the buffering capacity of the litter (Gallagher et al. 1987). Increasing the pH by
adding lime can speed decomposition in acidic wetlands (Ivarson 1977); Acidification can also affect
nitrification rates in wetlands (Dierberg and Brezonik 1982), and secondary production. Aquatic invertebrate
production was correlated inversely to pH in Plante and Downing's (1989) analysis of aquatic bed community
data from 51 lakes (164 samples) from temperate regions of the world.
Salinization. The effects of salinization, e.g., from irrigation return water and oil drilling wastes, on
biological processes have generally not been studied in inland wetlands.
Sedimentation/Burial. The effects of excessive sedimentation on biological processes have generally not been
studied in inland wetlands. Based on studies in other surface waters, respiration is likely to increase initially
and decomposition rates may decrease.
Turbidity/Shade, Vegetation Removal. The impacts of increased turbidity on biological processes have
generally not been studied in inland wetlands. Based on studies in other surface waters, primary production
increases with increased solar energy, and secondary production may increase as well, depending on habitat
availability and other factors. Decomposition in a southern forested wetland, as measured by the "cotton
rate of rotting (CRR)" was 50 percent greater after removal of vegetation by herbicide than in an
undisturbed forest (Mader et al. 1988).
Thermal Alteration. Decomposition may be enhanced by moderate temperature increases, but thermal
effects are more likely to be overshadowed by effects of litter type, depth, consumer invertebrate density, and
canopy cover (Hauer et al. 1986). Primary and secondary production generally increase with increasing
temperature, but thresholds beyond which these processes start to decline are not known for any wetland
type, and thermal loading may decrease the primary productivity of specific taxa and communities (e.g., Scott
et al. 1985). Aquatic invertebrate production was correlated with water temperature in Plante and
Downing's (1989) analysis of aquatic bed community data from 51 lakes (164 samples) from temperate
regions of the world.
Dehydration/Inundation. In southern floodplains, production of woody vegetation was greater in forested
wetlands that are flooded during some portion of the year but are well-drained (except for small,
intermittent storms) during the growing season (Birch and Cooley 1983). Decomposition rates are generally
slower in wetlands with longer duration flooding, anoxia, and greater water depths (Brinson 1981, Day et
al. 1988), but dehydrated wetlands may experience considerable accretion of organic matter (Burton 1984,
Elder and Cairns 1982). Effects of various inundation regimes on vegetation biomass have been reported
by Knighton (1985), Fredrickson and Taylor (1982), Robel 1962, and others.
Fragmentation of Habitat. We found no studies that attributed a decline in individual wetland annual
productivity, decomposition or denitrification rates to the regional declines in wetlands that have occurred.
One can surmise that as the distance between wetlands becomes greater, and/or hydrologic connections
become severed by dehydrated channels or dams, the simplified community structure of the remaining
wetlands would support lower biological rates. However, this has not been tested.
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13.2 SAMPLING METHODS AND EQUIPMENT
Some factors that could be important to measure and (if possible) standardize among wetlands when
monitoring anthropogenic effects on the processes of annual productivity, decomposition and denitrification
include:
age of wetland (successional status), water depth, temperature (site elevation, aspect),
hydraulic residence time, conductivity and baseline chemistry of waters and sediments
(especially pH, DO, organic carbon, and suspended sediment), current velocity, sediment
type, stream order or ratio of discharge to watershed size (riverine wetlands), shade, ratio
of open water to vegetated wetland, vegetation type, and the duration, frequency, and
seasonal timing of regular inundation, as well as time elapsed since the last severe
inundation or drought.
Methods for measuring productivity and other biological processes in aquatic environments are described
in Edmondson and Winberg 1971, Kibby et al. 1980, Murkin and Murkin 1989, Smith and Kadlec 1985,
Symbula and Day 1988, and others. A method for measuring whole-wetland respiration is described by
Madenjian et al. (1990).
Because all biological processes are expressed as rates, they require data from at least two points in time.
To measure annual productivity in wetlands, measurements of plant biomass are made at the onset of the
growing season and at the time of peak live biomass. Measurements of decomposition are generally initiated
during the mid to late growing season.
Methods that have been used in wetlands are described only briefly below. Measurement of biological
processes in wetlands has generally been done with great innovation and adaptation, with few studies
employing exactly the same procedures. Thus, only two measurements are described below-decomposition
and tree growth. For other processes and parameters, methods used in other surface waters, e.g., for
measurement of invertebrate production, might sometimes be applicable to wetlands.
Decomposition methods. Typically, several packs of biodegradable material are placed in surface water and
subsets are removed over periods ranging from weeks (usually) to years. Decomposition is inferred by
difference in weight over a specified period of time.
Organic matter decomposition rate in one wetland study was measured by as the tensile strength losses of
soil burial cloth (93 percent cellulose) after 9 days (Mader et al. 1988). In another study, cypress leaves
in mesh fiberglass screen bags were placed in the deepest spots (Dierberg and Ewel 1984); these authors
cited the finding of Deghi et al. (1980) that there is no significant difference in decomposition rates between
center and edges of cypress swamps. Five litter bags were collected at 15, 29, 58, 114, 205, 390, and 570
days. In a third study (in a stream), bags of air-dried leaves collected just before leaf-fall were placed in
riffles in a control and a treatment stream. Bags were collected at 10, 30, 58, 87, and 115 days after
placement (Meyer and Johnson 1983).
The litter decomposition rate can integrate short-term indexes of microbial activity (such as ATP, CO2
evolution, and microfaunal counts) over periods of several years (Edmonds 1987). Tree leaf and grass litter
is collected and air dried; litter bags are set out and collected at 1, 2, 5, and 7 years.
Decomposition of different sections of three plant species was studied by Hill (1985), who collected
Nelumbo leaf laminae and petioles, Typha leaves, and whole Ludwigia plants in the fall when the leaves
were beginning to turn yellow. The litter was air-dried and cut into 10-cm pieces, and 2-5 g samples were
put into nylon mesh leaf bags (15 cm^, 3-mm octagonal openings). Three to five replicates of each type
were put between wire mesh to hold them on the sediment below the water level of a reservoir. Samples
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were collected from an inundated site at 2, 4, 7, 14, 21, 28, 63, 91, 119, and 154 days and from a drawdown
site at 2, 4, 7, 14, 35, and 63 days. Macroinvertebrates were removed from the samples before air-drying.
Tree Growth. Increment cores can be used to estimate tree ages and growth rates, as well as for shrubs
(Ehrenfeld 1986). Data from ring counts can be checked against aerial photographs (Klimas 1987). Lemlich
and Ewel (1984) took cores of pondcypress (Taxodium distichum var. nutans), a difficult species to age
because of the presence of false rings. They identified false rings by their gradual change in cell size, as
contrasted with true rings, in which small latewood cells are readily distinguishable from large earlywood
cells.
Leavitt and Long (1989), working with southwestern conifers, described a method of using tree ring analysis
to reconstruct historic precipitation and drought patterns. Their method is based on ratios of ^C to ^C,
using the principle that, under drought conditions when stomates are closed, the tree will use a greater
proportion of carbon-13 in photosynthesis.
Repeated measurement of tree diameter also can be used to gauge growth. It is important to define
precisely where on the trunk the measurement is to be taken. In Franklin and Frenkel's study (1987), tree
data could not be compared between years because the heights on the boles at which diameters were
measured were not standardized. Straub (1984) took diameters of cypress trees at 1.37 m above ground or
above buttresses, if present. Small nails were hammered into the trunks so remeasurement would be done
at the same point on the tree. Aluminum vernier tree bands calibrated to one-hundredth of an inch are
also used to measure tree growth (Sklar and Conner 1983).
More detailed measurements of diameter were used by Scott et al. (1985) in a South Carolina floodplain
swamp. Tree biomass was measured by taking five diameters at 5-cm intervals above and below breast
height. A nail was driven into the stem at the topmost measuring point to facilitate subsequent
measurements; a chain with measurement intervals marked on it can be hung from the nail.
13.3 SPATIAL AND TEMPORAL VARIABILITY, DATA GAPS
In general, data on community-level biological processes have not been uniformly collected from a series of
statistically representative wetlands in any region of the country. Thus, it is currently impossible to state,
for any wetland type, what are "normal" rates for processes such as annual productivity, decomposition and
denitrification.
Only a few studies have compared biological processes among wetlands or aquatic environments in a region
or among regions. These include Brinson et al. 1981, Gushing et al. 1983, and Plante and Downing (1989).
Apparently few studies have compared year-to-year or long-term variation in biological processes in wetlands.
Such unpublished data may be available from sites of the U.S. Department of Energy's National
Environmental Research Park system, and sites of the National Science Foundation's Long Term Ecological
Research (LTER) program.
Existing data on wetland plant productivity, collected by a wide variety of methods, was reported by Adamus
1983, Kibby et al. 1980, and (for Carex wetlands only) by Bernard et al. 1988. Net annual primary
productivity of some inland wetland emergent species can exceed 6000 g/m^/yr, but usually is less than about
2000 g/m^/yr. Biomass of submersed macrophytes spans four orders of magnitude (Moeller 1975).
Decomposition of emergent macrophytes in lacustrine wetlands may take from about 200 to 1000 days for
90 percent weight loss (Hill 1985). Breakdown rates (per day) range from 0.0008 for woody plants in bogs
to 0.0190 for non-woody plants in riparian wetlands (Webster and Benfield 1986).
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Secondary production in wetlands has been measured much less often than primary production. For
invertebrates, Smock et al. (1985) reported 3.09 g/m-2 annual production from an acidic South Carolina
forested wetland; Plante and Downing (1989) compile estimates of invertebrate production from lacustrine
wetlands. Fish production in a 4-year study of the Okefenokee Swamp in Georgia ranged from 43 to 187
kg wet mass/ha (Freeman 1989).
Limited quantitative data on other biological processes is available by wetland type in some of the
"community profile" publications of the USFWS (Appendix C).
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191
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APPENDIX A. Summary of Advantages and Disadvantages of Use of Major Taxa in Monitoring Wetland
Ecological Condition.
Microbial Communities
ADVANTAGES
o tight linkage to fundamental processes (e.g., decomposition, denitrification, respiration)
o samples easily collected, transported, and analyzed
o some taxa linked to animal welfare (e.g., streptococci)
o EPA protocols available
o immediate response to contamination
o measurable in wetlands which lack surface water
o sensitive to presence of some contaminants (e.g., Ames test, Microtox test)
o "indicator taxa" relatively well-known (especially protozoans)
o some culture bioassay data are available
DISADVANTAGES
o response is often not identifiably stressor-specific
o laborious and slow (plate culture) identification; process measurements difficult to interpret with
regard to ecological significance
o general absence of existing regional field databases
o rapid turnover requires frequent sampling; do not integrate conditions over time very well
o naturally great micro-spatial variation, especially in tidal wetlands
o drifting cells in riverine wetlands complicate interpretation
o low social recognition of their importance
o bioaccumulation is irrelevant and impractical to detect
ADVANTAGES
o tight linkage to fundamental processes (e.g., photosynthesis, respiration)
o pivotal relationships in food webs
o EPA protocols available (may need modification to wetlands)
o measurable in some wetlands which lack surface water
o tolerances and indicator value are relatively well-known, particularly to nutrients, and most are very
sensitive to herbicides
o simple collection procedures with minimal wetland impact
o response to stressors is usually immediate
o generally immobile and thus reflective of site conditions, useful for in situ exposure assessments and
whole-effluent bioassays
DISADVANTAGES
o response is often not identifiably stressor-specific
o laborious identification
o some regional field databases exist, but not for wetlands
o rapid turnover requires frequent sampling
o cannot be effectively sampled during dormant season
o low social recognition of their importance
o bioaccumulation is unmeasurable
o drifting cells of unattached species complicate interpretation
•
192
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o most relatively insensitive to heavy metals and pesticides (Hellawell 1986)
Mosses. Liverworts. Ferns
ADVANTAGES
o a few taxa are reputed indicator species for physicochemical contaminants
o perhaps the most sensitive indicator of hydric regimes
o the only integrator of the long-term geologic record (ie, peat core analyses for metals accumulation,
land cover change, ground water flow reversals)
o immobile and thus reflective of site conditions, useful for in situ exposure assessments
DISADVANTAGES
o response is often not identifiably stressor-specific
o laborious sampling and identification
o low social recognition of their importance
o few regional field databases exist
Submersed Aquatic Vascular Plants
ADVANTAGES
o extremely sensitive to turbidity, eutrophication, hydroperiod
o sensitivities of several indicator species are well known
o relatively important in food webs (e.g., waterfowl)
o immobile and thus reflective of site conditions, useful for in situ exposure assessments
o patterns interpretable using remote sensing
DISADVANTAGES
o difficult to sample systematically throughout a wetland
o cannot be effectively sampled during dormant season
o absent from wetlands that lack standing water (e.g., bogs)
o tolerant of intermittent pollution
o laborious identification
o low social recognition of their importance
o few regional field databases exist
Non-rooted Aquatic Vascular Plants
ADVANTAGES
o extremely sensitive to nutrient additions
o sensitivities of some indicator species (e.g., Lemna) are well known
o important in food webs (e.g., waterfowl)
o mostly immobile and thus reflective of site conditions, useful for in situ exposure assessments
o patterns sometimes interpretable using remote sensing
DISADVANTAGES
o difficult to sample systematically throughout a wetland
o limited bioaccumulation due to short lifespan
o absent from wetlands that lack standing water (e.g., bogs)
o laborious identification
193
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o low social recognition of their importance
o few regional field databases exist
o cannot be effectively sampled during dormant season
Emergent (Herbaceous) Vascular Plants
ADVANTAGES
o occur in virtually all wetlands
o sensitivities of some indicator species (e.g., Typha. Phragmites. Phalaris) to nutrients/sediment are
well known
o immobile and thus reflective of site conditions, useful for in situ exposure assessments
o bioaccumulate to a moderate degree
o patterns interpretable using remote sensing
o sampling techniques and community metrics well-developed
o moderately sensitive to nutrients and hydroperiod alteration
o some regional field databases exist
DISADVANTAGES
o not highly sensitive to contaminants and sedimentation
o lagged response to stressors (episodic contamination may not be reflected)
o low social recognition of importance
o sampling and identification is laborious
o community cannot be completely characterized during the dormant season
o dispersal, herbivory, soil type and other factors often overshadow contaminant effects
Forested/Shrub (Woody) Vascular Plants
ADVANTAGES
o occur widely
o sensitivities of many species to hydroperiod change are relatively well known
o immobile and thus reflective of site conditions
o bioaccumulate to a moderate degree
o patterns interpretable using remote sensing
o sampling techniques and community metrics well-developed
o some regional field databases exist
o trends can be inferred (with care) using tree ring analyses
o signs of stress (e.g., die-offs) are socially recognized
o sampling and identification are fairly easy
o community can be characterized even in the dormant season
DISADVANTAGES
o not highly reflective of contaminants and sedimentation
o long lagged response to stressors (episodic contamination may not be reflected); in situ
experimentation is impractical
o response difficult to interpret where past management (e.g., silviculture) has been practiced
Aquatic Insects (e.g.. dragonflies. midges)
ADVANTAGES
194
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o occur in all wetland types, even those lacking surface water
o community metrics/indices well-developed (e.g., Index of Biotic Integrity) but may need adaptation
for wetlands
o intermediate lifespans reflect episodic events without requiring extremely frequent sampling
o bioaccumulate to a moderate degree
o can be caged for whole-effluent bioassays or in situ assessments
o relatively important in food webs
o community can usually be sampled year-round
o some regional field databases exist, though few for wetlands
o show characteristic response to all major wetland stressors (hydroperiod, sediment, nutrients,
contaminants)
o some taxa linked to human welfare (e.g., mosquitoes)
o EPA sampling protocols available, but need modification for wetlands
o contaminants may induce identifiable deformities
DISADVANTAGES
o occurrence in isolated wetlands may be strongly tied to sources of colonizers and their dispersal
mechanisms
o sampling difficult and true densities very difficult to determine in wetlands with herbaceous
vegetation
o laborious identification
o low social recognition of their importance
o naturally great micro-spatial variation
o community composition potentially affected by selective predation (e.g., by fish, waterfowl)
Benthic/Epiphytic Macro-crustaceans (e.g., amphipods, crayfish, oligochaetes, isopods)
ADVANTAGES
o less subject to dispersal than aquatic insects (and thus more reflective of conditions in a particular
wetland)
o may be more sensitive than aquatic insects to contaminants
o fairly simple sampling and identification
o social recognition of some species (e.g., crayfish, sandworms)
o other advantages — similar to Aquatic Insects, above
DISADVANTAGES
o mostly absent from wetlands which lack standing water
o naturally great micro-spatial variation
o community composition potentially affected by selective predation (e.g., by fish, waterfowl)
Mollusks
ADVANTAGES
o highly immobile and thus most reflective of site conditions, useful for in situ exposure assessments
o highly bioaccumulative (e.g., clams, mussels)
o depuration procedures can indicate potential contaminant uptake rates
o bioassay data fairly extensive
o contaminants may induce identifiable deformities
o can be sampled year-round
o historic recreation of growth is possible (with care)
195
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o presumptive indicator of hydroperiod (complete, sustained wetland drawdown)
o EPA protocols available
o high social importance of coastal species (shellfish)
DISADVANTAGES
o very localized occurrence, related largely to dissolved solids rather than contaminants
o laborious sampling and (in freshwater) identification
Fish
ADVANTAGES
o community metrics well-developed (Index of Biotic Integrity), though not for wetlands; many
reputed indicators (e.g., carp)
o most comprehensive set of bioassay data
o can be caged for whole effluent bioassay and in situ studies, or avoidance measured using
radiotelemetry
o moderately bioaccumulative
o fairly simple identification (except larval stages)
o population characteristics, growth fairly easy to discern
o contaminants may induce identifiable deformities
o can be sampled year-round
o presumptive indicator of hydroperiod (absent from isolated wetlands with complete, sustained
drawdown)
o EPA protocols available
o integrate broad, longer-term, landscape-level impacts because of their mobility, high trophic position,
and longer life span
o high social importance of most species; existing water quality standards for aquatic life focus on fish
DISADVANTAGES
o mobility makes it difficult to locate specific contaminant sources
o absent (or present for only brief periods) in most wetlands
o laborious sampling
o early life stages and non-game species may be difficult to identify
Amphibians and Reptiles
ADVANTAGES
o small home range relative to larger vertebrates
o highly (e.g., snapping turtle, alligator) to moderately bioaccumulative; can be caged for in situ
assessments
o some social recognition
o fairly simple identification
o fairly well-established sampling protocols
o sensitive to hydroperiod alteration
o present in most inland wetland types
DISADVANTAGES
o sampling limited to certain seasons in some regions
o mostly absent from tidal wetlands
o sampling can be laborious
196
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presence can be strongly influenced by natural dispersal conditions
Birds
ADVANTAGES
o high social recognition, particularly waterfowl
o have the only relatively extensive nationwide databases on trends, habitat needs, distribution
o moderately extensive bioassay data
o some species (e.g., wading birds, harrier) are highly bioaccumulative
o avoidance is measurable using radiotelemetry, and in situ assessments are possible (caged or clipped
individuals)
o simple sampling and identification
o present in all wetland types
o established sampling protocols are available
o the only suitable indicator of degradation occurring at the landscape scale
DISADVANTAGES
o in general, community structure is highly controlled by physical habitat, and perhaps hunting
mortality, rather than contaminants
o mobility makes it difficult to locate specific causes of mortality sources (could be thousands of miles
away)
o essentially absent from some wetlands in winter
Mammals
ADVANTAGES
o many (e.g., otter) are highly bioaccumulative
o high social recognition and value (e.g., muskrat)
o avoidance is measurable using radiotelemetry, and in situ assessments are possible (caged individuals)
o fairly simple sampling and identification
o some sign (e.g., beaver dams) can be remotely sensed
o present in all wetland types
o established sampling protocols are available
o an extensive database of acute toxicity data for mice/rats may be partially transferable
DISADVANTAGES
o great temporal and spatial variation (many species are cyclic) makes data interpretation difficult
o in general, community structure is highly controlled by physical habitat, and perhaps trapping
mortality, rather than contaminants
o mobility (and frequent use of non-wetland habitat) makes it difficult to locate specific causes of
mortality sources
Biological Processes (Functions)
Definition: Whole-wetland measurement of photosynthesis, primary productivity, respiration, denitrification,
nitrogen fixation, decomposition, leaching, and/or similar processes
ADVANTAGES
o most important indicators of wetland sustainability and life support function
197
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DISADVANTAGES
o not as sensitive to contamination as is community structure or tissue analysis (Schindler 1987)
o measurement is laborious, time-consuming (e.g., isotopes)
o social recognition of importance is weak
o extreme spatial and temporal variation
o measured values may reflect natural successional stage rather than human-induced stress
198
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APPENDIX B. Wetland Biomonitoring Sites, Referenced and Mapped by State.
The following maps are provided (a) to facilitate regionalization of future efforts, (b) to further cooperation
among researchers, and (c) to encourage use/analysis of extant data. These maps and their associated
bibliographies DO NOT depict ALL wetland research sites. Nonetheless, a systematic, extensive process was
used to develop them, as described in section 1.2. Criteria for including studies in this listing were described
in section 1.2, but a small portion of sites may not fully meet these criteria. The quality of individual
studies or descriptions of their locations have not be verified or assured. Digital versions of all maps and
their bibliographies currently reside with the Wetlands Team at EPA's Environmental Research Laboratory
in Corvallis, Oregon.
The numbers on the maps are keyed to citations listed on pages that follow each state map. The
abbreviation on the line above each citation (e.g., MOBBC21) includes the state code, number reference
to map, and (in some cases) the following additionally identifying abbreviations:
BBC = wetland breeding bird census plot, from Cornell database
BBS = wetland breeding bird survey route, from USFWS database
BSB = inland shorebird migration site, monitored by the International Shorebird Survey
BW = waterfowl survey site (breeding, mid-winter, or other) monitored by state and/or federal agencies
LTR = long-term environmental research site, usually funded in part by the National Science Foundation
EPA = reference wetland studied by USEPA Wetlands Research Team and its contractors
Abbreviations at the end of citations refer to topical coverage, as follows:
A = algae
AI= aquatic invertebrates
B = birds
BA= bioaccumulation
D = decomposition
F = fish
H = herptiles (amphibians and reptiles)
I = impacts of human activities
MA= mammals
MI= microbial communities
P = plants (generally), PB= bog plants, PE= emergent plants,
PM= submersed macrophytes, PW= woody plants
R = regional studies (> 5 wetlands simultaneously sampled)
RS= remote sensing
S = spatial distribution of wetlands
SO= sediment/organic matter accumulation
TS= time series measurements (> 3 years)
199
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Inland Wetlands Having Biological
Community Measurements
A I abama
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE * or - I 81111
R«3»°rch Study Si to
Migratory Shor«b,rd Survey CBSB) site
Br«Bd(ng Bird Census (BBC) site that includes uetland
Annual Chr ittna* Bird Count area ( IS-*i le diameter)
Th i • nap does NOT portray ALL wetland sampling si
E«pha«i« » on «ite» where common i ty- I eve I data w
collected Se« chapter t for inclusion crit*ria
Breeding Bird Survey Starting points fo
ma i r. \ y non-wetland habitat
transects
Site* ore referenced by cod* nunb«r to tK« accompanying
•tat* bibliography
SITE LOCATED IN COUNTY, SPECIFIC LOCATION. Oregon
Data Compilation Paul Adamu* and Robin Renteria Cortogfaphy Jeff Iri«h
200
-------
ALABAMA
Mapped
All
Pennington, C.H., J.A. Baker, F.G. Howell, and C.L. Bond. 1981. Study of Cutoff Bendways on the Tombigbee
River. Tech. Rep. E-81-14, U.S. Army Engr. Uaterw. Expt.Stn.. Vicksburg, MS. f
AL3
Nader, S.F., U.M. Aust, and R. Lea. 1988. Changes in Net Primary Productivity and Cellulose Decomposition
Rates in a Water Tupelo-Bald Cypress Swamp Following Timber Harvest. Fifth Biennial S. Silvicul. Res. Confer.,
Memphis, TN. 5 pp. PU D I
AL4
Teels, B.M., G. Anding, D.H. Arner, E.D. Norwood and N.E. Wesley. 1978. Aquatic plant-invertebrate and
waterfowl associations in Mississippi. Proc. Southeast. Assoc. Fish Wildl. Agenc. 30:610-616.
AL5
Aust, W.M., S.F. Mader, and R. Lea. 1988. Abiotic changes of a tupeIo-cypress swamp following helicopter and
rubber-tired skidder timber harvest. Fifth Southern Silviculture Res. Conf., Memphis, TN. PW 1
AL7
Hall, T.F. and W.T. Penfound. 1943. Cypress-gum communities in the Blue Girth Swamp near Selma, Alabama. Ecol.
24(2):208-217. PW
AL7
Hall, T.F., W.T. Penfound, and A.D. Hess. 1946. Water level relationships of plants in the Tennessee Valley
with particular reference to malaria control. J. Tenn. Acad. Sci. 21:18-59.
AL8
Auburn University. 1982. Fisheries studies on Gainesville and Aliceville Lakes on the upper Tombigbee River
system, Alabama-Mississippi.Environ. Studies Inland Waterway Sys., CESAM-PD-EI. F
AL9-12
U.S. Fish & Wildl. Service. 1985. Survey of waterfowl utilization activities, Tennessee-Tombigbee Waterway,
Alabama and Mississippi. USFWS Division of Ecological Services, Daphine, AL. B
AL9-12
U.S. Fish & Wildl. Service. 1986. Tennessee-Tombigbee Waterway, waterfowl survey. 1985-1986 annual Report.
Division of Ecological Services, Daphine, AL. B
201
-------
ALABAMA (continued)
AL9-12
U.S. Fish & Wildl. Service. 1987. Tennessee-Tombigbee Waterway, waterfowl survey. 1986-1987 annual Report.
Division of Ecological Services, Oaphine, AL. B
AL9-12
U.S. Fish & Wildl. Service. 1986. A study of cutoff bendways on the Tennessee-Tombigbee Waterway. 1986 Annual
Report. Division of Ecological Services, Daphine, AL. AI F
AL13
Pardue, W.J. and D.H. Webb. 1985. A comparison of aquatic macroinvertebrates occurring in association with
Eurasian Watermilfoil with those found in the open littoral zone. J. Freshw. Ecol. 3(1):69-79.
ALK
Tomljanovich, D.A., G.A. Brodie, and D.A. Hammer. 1988. Constructed Wetlands for Treating Acid Drainage at
TVA Facilities. Off. of Nat. Res. and Economic Dev., TVA/ONRED/WRF-88/2. Knoxville, TN.
AL15-16
James, W.K., D.R. Lowery, D.H. Webb, and W.B. Wrenn. 1989. Supplement to White Amur Project Report. Tennessee
Valley Authority. Resource development. River Basin Operations, Water Resources. TVA/WR/AB--89/1. Muscle
Shoals, AL.
AL18
Peltier, W.H. and E.B. Welch. 1969. Factors affecting growth of rooted aquatic plants in a river. Weed Sci.
17:412-416. P PM
AL18
Peltier, W.H. and E.B. Welch. 1970. Factors affecting growth of rooted aquatic plants in a reservoir. Weed
Sci. 18:7-9. PM I
ALBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
Ithaca, NY. B
ALBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
ALBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, MA. B
ALBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
202
-------
ALABAMA (continued)
ALCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
Not Happed
Deutsch, U.G. 1988. Community structure and production of benthic macroinvertebrates in 0.04-hectare ponds
with and without organic loading. Ph.D. Diss., Auburn Univ., Auburn, AL. 163 pp.
Johnson, R.C, J.U. Preacher, J.R. Gwaltney, Jr., and J.E. Kennamer. 1975. Evaluation of habitat manipulation
for ducks in an Alabama beaver pond complex. Southeastern Assoc. Game Fish Comm. Proc. 29: 512-518.
Miranda, L.E., W.I. Shelton, and T.D. Bryce. 1984. Effects of water level manipulation on abundance,
mortality, and growth of young-of-the-year largemouth bass in West Point Reservoir, Alabama - Georgia. N. Amer.
J. Fish. Manage. 4:314-320. F
Murad, H.A. 1987. Acidification as environmental pollution: effects on fishpond ecology. Ph.D. Diss., Auburn
Univ., Auburn, AL. 101 pp.
Sedana, I.P. 1987. Development of benthos and its relationship to fish production in ponds with organic
loading. Ph.D. Diss., Auburn Univ., Auburn, AL. 118 pp.
Smith, B.M. and E.P. Hill. 1979. The potential of coal strip mine in Alabama as waterfowl habitat. Proc. Ann.
Conf. S.E. Assoc. Fish Wildl. Agencies: 33:1-10. B
Speake, D.W. 1955. Waterfowl use of creeks, beaver swamps, and small impoundments in Lee County, Alabama.
pp. 178-185 In: Proc. SE Assoc. Game Fish Comm. B
Webb, D.H. and A.L. Bates. 1989. The aquatic vascular flora and plant communities along rivers and reservoirs
of the Tennessee River system. J. Tennessee Acad. of Sci. 64(3):197-203. PM
203
-------
Inland Wetlands Having Biological
Community Measurements
Arkansas
Th i • nap does NOT portray ALL wetland sonpttng site*
Enphaa i • is on *>t«« wh«r« conmun i ty- I *v«1 data w«r*
coM^ctco* S«* chapter I for inclusion crtt*ria
5it«« or« r«f«r*nced by cod« nu«b«r to
•tat* bibliography
accompanying
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE » or - I ani
9 R.s.arch Study Site
| Migratory Sr-orebird Survey CBSB? site
Q Breeding Bird Census (BBC) site that includes wetland
O Annual Christmas Bird Count area (!5-mi!e diameter)
Mo»l cover mainly non-weiland habitat
+ Breeding Bird Survey Starting points for 25»i transects
SITE LOCATED IN COUNTY. SPECIFIC LOCATION(S) NOT PLOTTED
* State/Federal waterfoul survey
USEPA Environment*! R«».«rch Lebor.tory. Cor»elll».
Data Compilation Pout Adanui and Robin Renter 10
Cartography Jeff Ir,«h
204
-------
ARKANSAS
Happed
AR1-3
Landin, H.C. 1985. Bird and Mammal Use of Selected Lower Mississippi River Borrow Pits. Ph.D. Diss.,
Mississippi State Univ., 405 pp. B MA
AR4-7
Cobb, S.P., C.H. Pennington, J.A. Baker, and J.E. Scott. 1984. Fishery and ecological investigations of main
stem levee borrow pits along the lower Mississippi River. Mississippi R. Conn., Vicksburg, MS. 120 pp. F
AR6
Rainwater, W.C. and A. Houser. 1982. Species composition and biomass of fish in selected coves in Beaver Lake,
Arkansas, during the first 18 years of impoundment, 1963-1980. N. Amer. J. Fish. Manage. 2(4):316-325.
AR8-20
Dale, E.E.,Jr. 1984. Wetland forest communities as indicators of flooding potential in backwater areas of
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AR21
Harris, J.L., F.L. Burnside, B.L. Richardson, and W.K. Welch. 1984. Methods for analysis of highway
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AR21
Harris, J.L., F.L. Burnside, B.L. Richardson, and W.K. Welch. In Prep. Eight-year biomonitoring of Oats Creek
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AR22
Bryant, C.T., C.T. Bryant, Jr., and J.D. Rickett. 1988. Use Attainability Analysis of Rosenbaum Lake Pulaski
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AR23
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AR25-34
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Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
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U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
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International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
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U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
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205
-------
ARKANSAS (continued)
Not Mapped
Bedinger, M.S. 1971. Forest Species as Indicators of Flooding in the Lower White River Valley, Arkansas.
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206
-------
Inland Wetlands Having Biological
Community Measurements
Ar i zona
ACCURACY OF SITE LOCATION? ESTIMATED TO BE » or - 10m,
4 Research Study Si te
I Migratory Shc-r*bird Survey (BSB) site
Q Breeding Bird Census (BBC) site that includes wetland
O Annual Christmas Bird Count area CIS-nil* diameter)
Most cover mainly non-wetland habitat
. Th i • nap doe* NOT portray ALL w*tlar>d campling site*
+ Breeding Bird Survey Starting poinls for 2S«i transects
AND points uhere transects enter nes county Host cover E»pho« i « i* on *it*« wh«re commun i t y- I eve I data were
mainly non-wetland habitat collected Se* cKapt*r t for inclusion criteria
SITE LOCATED IN COUNTY, SPECIFIC LOCATION'S) NOT PLOTTED
* State/Federal waterfowl survey
Site* are referenced by code number to the accompanying
•tot* bibliography
USEPA En«lron««nt«l R«»«erch Llborttory. Cor««lll>, Or«gon
Data Compilation Paul Adamu* and Robin R*nt*ria
Cartography J*ff Iri»n
208
-------
ARIZONA
Happed
A21
Heede, B.H. 1985. Interactions Between Streamside Vegetation
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AZ4
Ohmart, R.D., B.W. Anderson, and U.C. Hunter. 1988. The Ecology of the Lower Colorado River from Davis Dam
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AZ5-10
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AZ8
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AZ10
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AZ10-12
Anderson, B.W. and R.D. Ohmart. 1988. Structure of the winter duck community of the lower Colorado River:
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AZ11
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Piest, L.A. and L.K. Sowls. 1985. Breeding duck use of a sewage marsh in Arizona. J. Wildl. Manage. 49:580-
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AZ21
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-------
ARIZONA (continued)
AZ23
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AZ25
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AZ25
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AZ30
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International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
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210
-------
ARIZONA (continued)
AZBW1-
U.S. Fish & Uildl. Service. Unpub. Waterfowl Survey Data. B
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ARIZONA (continued)
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212
-------
Inland Wetlands Having Biological
Community Measurements
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE » or - 18m,
e Research Study Site
f Migratory Shorebird Survey CSSB) site
Q Breeding Bird Census CB6C) s.te that includes wetland
O Annual Christmas Bird Count ar»a CIS-mile diameter)
+ Breeding .Bird Survey Storting points for 25mi transects
AND points where transects enter new county Most cover
SITE LOCATED IN COUNTY. SPECIFIC LOCATIONCS) NOT PLOTTED
» State/Federal waterfowl survey
CaIifornia
Th i s map do*« NOT por troy AUL w«t I and *oi»p 1 tnQ * 11*«
Emphasis is on sites where community~I eve I data w«re
col I »c ted $•* chapter 1 for i ne 1
-------
CALIFORNIA
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CA1
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CALIFORNIA (continued)
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CA59
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CA60
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CA61
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CA62
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CA65
Cal Expo Floodplain Manage. Plan Task Force. 1987. Cal Expo Floodplain Management Plan. Sacramento County
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CA67
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217
-------
CALIFORNIA (continued)
CA69
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CA73
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Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
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U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
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U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
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Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
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218
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219
-------
Inland Wetlands Having Biological
Community Measurements
Co I or ado
•i#V
• —'^^y\
<
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE * or - I 0m,
+ Research Study Site
£ Migratory Shorebird Survey CBSB) site
Q Breeding Bird Census (BBC) site that includes wetland
O Annual Christmas Bird Cour.1 area OS-mil* diameter)
~t~ Breeding Bird Survey Starting point* for 25mi transects
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONCS) NOT PLOTTED
+ State/Federal waterfowl survev
This mop do»s NOT portray ALL wetland sampling sites Sites are referenced by cod* number to the accompanying
E rfiphasis is on 3itea where comnunity-I eve I dato w«re state bibI iogrophy
coI Iected S*« chapter t for incI us > on crit er i a
USEPA Environ««nt•I R*»««rch Laboratory* Corvalli*. Ora^on
Campi I ation PauI Adamus and Robin Rent*
Car tography Jeff Ir ish
July 1998
220
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C01
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C02
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COLORADO (continued)
C016
Cooper, D.J. 1987, 1988. Monitoring of a created wetland. Colorado School of Mines, Golden, CO.
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C021
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COBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
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U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
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COLTR
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222
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COLORADO (continued)
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223
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D
O
O>
O
O
00
O>
C
O
X
T>
C
O
-------
CONNECTICUT
Happed
CT1
Anderson, P.M., M.U. Leforland, and W.C. Kennard. 1980. Forested wetlands in eastern Connecticut: Their
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CTl
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CT2
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CT3
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CTBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
CTCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christinas Bird Count Data. Cornell University,
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Not Happed
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Conard, W.H. 1961. A floristic study of Beckley Bog. M.S. Thesis, Yale Univ., New Haven, CT.
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Niering, W.A. and R.H. Goodwin. 1962. Ecological studies in the Connecticut Arboretum Natural Area. I.
Introduction and a survey of vegetation types. Ecol. 43:41-54.
225
-------
Inland Wetlands Having Biological
Community Measurements
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE - or- - 10m,
• Research Study Site
| Migratory SKorebird Survey CBSB> site
Q Breeding Bird Census CBBO site that includes wetland
O Annual Christmas Bird Count area CIS-mile diameter)
Most cover mainly non-wet'and habitat
-t- Breeding Bird Survey Starting points for 2Smi transects
AND points where transects enter new county Host cover
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONS) NOT PLOTTED
^ State/Federal waterfowl survey
De;
Th i * nap doe* NOT por troy ALL wetland sanplmg • i iee
EnphciB 19 19 on sites where commun i ty - I eve I data were
coI Iected See chapter I for incIue
-------
DELAWARE
Happed
DEBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
Ithaca, NY. B
DEBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
DEBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, MA. B
DEBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
DECBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christinas Bird Count Data. Cornell University,
Ithaca, NY. B
227
-------
Inland Wetlands Having Biological
Community Measurements
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE + or - 10™.
0 Research Study Site
( Migratory Shorebird Survey (BSB) site
p Breeding Bird Census C8BC) *tt« that , nc 1 ud*s wetland
O Annual Christmas Bird Count area CIS-mil* diameter^
Most cover ma inly non-wetland habttat
~t~ Breeding Bird Sur v*y Storting poirite for 2Si» i transects
AND points where transects enter new county Most cover
noinIy non-weI 1 and hobttat
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONCS) NOT PLOTTED
^ State/FederaI waterfowl survey
Thi• map doe* NOT por tray ALL wet I and samp Iing site*
Edphas is is on s i tes where commun itylevet data ware
coMecl»d See chapter t for inclusion criteria
Site* are referenced by code number to the accompany \ r>g
state bibllography
US£PA Environmental Research Laboratory* Corvalli*. Oregon
Compi \aI> on Paul Adonu* and Rob > n Renter i a Car tography Jeff Irish
228
-------
FLORIDA
Mapped
FL4, 17-20
Canfield, D.E. and J.R. Jones. 1984. Assessing the trophic status of lakes with aquatic macrophytes. Lake
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FL6
Knight, R.L., B.H. Winchester, and J.C. Higroan. 1985. Ecology, hydrology, and advanced wastewater treatment
potential of an artificial wetland in north-central, Florida. Wetlands 5:167-180. WQ P I
FL7
Duever, M.J. and J. Mucollom. 1987. Plant community boundaries and water levels at Lake Hatchineha, Florida.
In: Proc. of the Nat. Wetlands Symposium, Wetland Hydrology. ASWM Tech. Rep. 6, Berne, NY. P
FL8
Winchester, B.H., J.S. Bays, and J.C. Higman. 1987. Inundation characteristics of wet prairie and marsh
wetlands in southwestern Florida, pp. 243-252 In: Proc. Nat. Wetland Symposium, Wetland Hydrology. ASWM Tech,
Rep. 6, Berne, NY.
FL10
Straub, P.A. 1984. Effects of wastewater and inorganic fertilizer on growth rates and nutrient concentrations
in dominant tree species in cypress domes, pp. 127-140 In: K.C. Ewel and H.T. Odum (eds.). Cypress Swamps.
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FL11
Dierberg, F.E. and P.L. Brezonik. 1982. Nitrifying population densities and inhibition of ammonium oxidation
in natural and sewage-enriched cypress swamps. Water Res. 16:123-126. MI I
FL12
Dierberg, F.E., P.A. Straub, and C.D. Hendry. 1986. Leaf-to-twig transfer conserves nitrogen and phosphorus
in nutrient poor and enriched cypress swamps. For. Sci. 32(4):900-913. PW I
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Inland Wetlands Having Biologica
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236
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GEORGIA
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Moran, M.A. 1987. Microbial community dynamics and transformations of vascular plant detritus in two wetland
ecosystems. Ph.D. Diss., Univ. Georgia, Athens. 159 pp.
GA7
Murray, R.E. and R.E. Hodson. 1985. Annual cycle of bacterial secondary production in five aquatic habitats
of the Okefenokee Swamp ecosystem. Appl. Environ. Microbiol. 49(3):650-655. MI
237
-------
GEORGIA (continued)
GA7
Murray, R.E. and R.E. Hodson. 1984. Microbial biomass and utilization of dissolved organic matter in the
Okefenokee Swamp ecosystem. Appl. Environ. Microbial. 47(4):685-692. MI
GA7
Murray, R.E. and R.E. Hodson. 1986. Influence of macrophyte decomposition on growth rate and community
structure of Okefenokee Swamp bacterioplanckton. Appl. Environ. Microbiol. 51(2):293-301. MI D
GA7
Oliver, J.D. 1987. Effects of biogenic and simulated nutrient enrichment on fish and other components of
Okefenokee Swamp marshes. Ph.D. Diss., Univ. Georgia, Athens. 179 pp.
GA7
Oliver, J.D. and S.A. Schoenberg. 1989. Residual influence of macronutrient enrichment on the aquatic food
web of an Okefenokee Swamp abandoned bird rookery. Oikos 55:175-182. P
GA7
Schlesinger, W.H. 1978. Community structure, dynamics and nutrient cycling in the Okefenokee Cypress
Swamp-Forest. Ecol. Monogr. 48:43-65. P
GA7
Schoenberg, S.A. and J.D. Oliver. 1988. Temporal dynamics and spatial variation of algae in relation to
hydrology and sediment characteristics in the Okefenokee Swamp, Georgia. Hydrobiologia 162:123-133. A
GA 7
Stinner, E.H. 1983. Colonial wading birds and nutrient cycling in the Okefenokee Swamp. Ph.D. Diss., Univ.
Georgia, Athens, GA. 143 pp.
GA11-13
Environmental Protection Division, Georgia Dept. Nat. Res. 1985. Water Quality Investigation of Falling Creek
Jasper and Jones Counties, Georgia, Ocmulgee River Basin. Atlanta, GA.
GA14
Smock, L.A. and D.L. Stoneburner. 1980. The response of macroinvertebrates to aquatic macrophyte
decomposition. Oikos. 35:397-403. AI
GA15
Hale, M.M. and D.R. Bayne. 1983. Effects of water level fluctuations on the littoral macroinvertebrates of
West Point Reservoir. Proc. Ann. Conf. S.E. Assoc. Fish & Wildl. Agencies. 34:175-180.
GABBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
Ithaca, NY. B
GABBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
GABSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, MA. B
238
-------
GEORGIA (continued)
GABU1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
GACBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
Not Happed
Allred, P.M. 1981. Leaf litter decomposition studies in a blackwater stream. Ph.D. Diss., Emory Univ.,
Atlanta, GA. 229 pp.
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subtropical blackwater river: the importance of habitat and life history. Ecol. Monogr. 54:25-63. AI
Birch, J.B. and J.L. Cooley. 1983. Effect of Hydroperiod on Floodplain Forest Production. Georgia Water
Resour. Res. Center, Atlanta, Tech. Completion Rep., 98 pp. PW
Boyd, H.E. 1976. Biological productivity in two Georgia Swamps. Ph.D. Diss., Univ. of Tennessee, Knoxville,
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Carlough, L.A. 1989. Fluctuations in the community composition of water-column protozoa in two southeastern
blackwater rivers (Georgia, USA). Hydrobiologia 185:55-62.
Cuffney, T.F. 1984. Characteristics of riparian flooding and its impact upon the processing and exchange of
organic matter in coastal plain streams of Georgia. Ph.D. Diss., Univ. Georgia, Athens. 181 pp.
Fail, J.L. 1983. Structure, biomass, production, and element accumulation in riparian forests of an
agricultural watershed. Ph.D. Diss., Univ. Georgia, Athens, GA. 310 pp.
Greear, P.F-C. 1967. Composition, diversity, and structure of the vegetation of some natural ponds in
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Lochmiller, R.L. 1979. Use of beaver ponds by southeastern woodpeckers in winter. J. Uildl. Manage.
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Smock, L.A. and C.M. MacGregor. 1988. Impact of the American Chestnut Blight on aquatic shredding
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Strange, J.R. 1976. Effects of high levels of inorganic phosphate on aquatic organisms in phosphate-rich
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239
-------
GEORGIA (continued)
Thorp, J.H., E.M. McEwan, M.F. Flynn and F.R. Hauer. 1985. Invertebrate colonization of submerged wood in a
cypress-tupeIo swamp and blackwater stream. Amer. Midi. Nat. 113(1):56-68.
Walther, P.B. 1983. Decomposition processes across a flooding gradient, with special reference to earthworm
populations. Ph.D. Diss., Univ. Georgia, Athens, GA. 148 pp.
240
-------
Inland Wetlands Having Biologica
Community Measurements
lowc
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE » or - 10mi
9 Research Sludy S.te
| Migratory Shor.bird Survey CBSB) s.te
Q Breeding Bird Census CBBC) srte that includes wetland
O Annual Christmas Bird Count area (15-mile diameter)
"t" Breeding Bird Survey Starting points for 25m i transects
AND points where transects enter new county Host cover
SITE LOCATED IN COUNTY. SPECIFIC LOCATIONS) NOT PLOTTED
» State/Federal uaterfoul survey
Th<« map does NOT portray ALL wetland sampling •ite«
Emphasis is on sites whero conmunity-1 eve I data were
collected See chapter t for inclusion criteria
Sites are referenced by cods number to the accompanying
state bibliography
USEPA En» iroti»sn(«l R«5«irch Laboratory. Cory«lli». Oregon
Data Compilatiori Paul Addnus and Robin Renter la Cartography Jsff Irish
242
-------
IOWA
Happed
IA1
Niemeier, P.E. and U.A. Hubert. 1986. The 85-year history of the aquatic macrophyte species composition in
a eutrophic prairie lake (United States). Aquatic Bot. 25:83-89. TS P
IA3&4
Geier, A.R. and L.B. Best. 1980. Habitat selection by small mammals of riparian communities: evaluating
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IA5-6
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of habitat alterations. J. Wildl. Manage. 44:1-15. B I
IA7
van der Valk, A.G. and C.B. Davis. 1976. Changes in the composition, structure, and production of plant
communities along a perturbed wetland coenocline. Vegetatio 32(2):87-96. P
IA8
Krapu, G.L., D.R. Parsons, and M.W. Welter. 1970. Waterfowl in relation to land use and water levels on the
spring run area. Iowa State J. Sci. 44(4):437-452. B I
IA9
Voigts, O.K. 1975. Aquatic invertebrate abundance in relation to changing marsh vegetation. Amer. Midi. Nat.
95(2):319-322. AI
IA11
U.S. Fish & Wildl. Serv. (In Process). Upper Mississippi Biological Monitoring Program.
IA13
Brown, M. and J.J. Dinsmore. 1986. Implications of marsh siie and isolation for marsh bird management. J.
Wildl. Manage. 50(3)-.392-397. S B
IA13
Brown, M. and J.J. Dinsmore. 1988. Habitat islands and the equilibrium theory of island biogeography: testing
some predictions. Oecologia 75:426-429. S B
IA14
Menzel, B.W., J.B. Barnum, and L.M. Antosch. 1984. Ecological alterations of Iowa prairie-agricultural
streams. Iowa State J. of Research 59(1).
IA15
Davis, C. B., A.G. van der Valk, and J. L. Baker. 1983. The role of four macrophyte species in the removal
of nitrogen and phosphorus from nutrient-rich water in a prairie marsh, Iowa. Madrono 30(3):133-142.
IA16
van der Valk, A.G. and C.B. Davis. 1980. The impact of a natural drawdown on the growth of four emergent
species in a prairie glacial marsh. Aquat. Bot. 9:301-322.
IA17
Weinhold, C.E. and A.G. van der Valk. 1988. The impact of duration of drainage on the seed banks of northern
prairie wetlands. Can. J. Bot. 67:1878-1884.
IA18
van der Valk, A.G. 1976. Zonation, Competitive Displacement and Standing Crop of Northwest Iowa Fen
Communities. Proc. Iowa Acad. Sci. 83(2):50-53.
IA19
van der Valk, A.G. and C.B. Davis. 1978. The role of seed banks in the vegetation dynamics of prairie glacial
marshes. Ecol. 59(2):322-335.
243
-------
IOWA (continued)
IA20
Eckblad, J.U., N.L. Peterson, and K. Ostlie. 1977. The morphometry, benthos and sedimentation rates of a
floodplain lake in Pool 9 of the upper Mississippi River. Amer. Midi. Nat. 97(2)-.433-443.
IA21
Provost, M.W. 1947. Nesting of birds in the marshes of northwest Iowa. Amer. Midi. Nat. 38:485-503.
IA22
van der Valk, A.G. and C.B. Davis. 1979. A reconstruction of the vegetational history of a prairie marsh.
Eagle Lake, Iowa, from its seed bank. Aquat. Bot. 6-.Z9-51. P
IA23
Hansen, O.K. 1971. Effects of Stream Channelization on Fishes and Bottom Fauna in the Little Sioux River,
Iowa. State Water Resour. Res. Inst., Ames. ISWRRI-38 W71-10751. I AI F
IABBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
Ithaca, NY. B
IABBS1-
U.S. Fish & Uildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
IABSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, MA. B
IABW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
IACBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
Not Mapped
Begres, F.M. 1971. The diatoms of Clear Lake and Ventura Marsh, Iowa. Ph.D. Diss., Iowa St. Univ., Ames.
202 pp.
Best, L.B. and D.F. Stauffer. 1980. Factors affecting nesting success in riparian bird communities. Condor
82:149-158. B
Best, L.B., D.F. Stauffer, A.R. Geier, and K.L. Varland. 1982. Effects of habitat alterations on riparian
plant and animal communities in Iowa. U.S. Fish & Wildl. Serv., Washington, DC. FWS/OBS-81/26. 55 pp. I PW
B
Betancourt, C. 1981. Aquatic hyphomycetes of central and northeast Iowa. Ph.D. Diss., Iowa St. Univ., Ames,
IA. 125 pp.
Bishop, R.A., R.D. Andrews, and R.J. Bridges. 1979. Marsh management and its relationship to vegetation,
waterfowl and muskrats. Proc. Iowa Acad. Sci. 86(2)-.50-56. B MA
Clambey, G.K. 1975. A survey of wetland vegetation in north-central Iowa. Ph.D. Diss., Iowa St. Univ., Ames.
Crum, G.H. and R.W. Bauhmann. 1973. Submersed aquatic plants of the Iowa Great Lakes region. Iowa State J.
Res. 48:147-173. PM
Holte, K.E. 1966. A floristic and ecological analysis of the Excelsio fen complex in northwest Iowa. Ph.D.
Diss., Univ. Iowa, Iowa City. 306 pp.
Hosseini, S.Y. 1986. The effects of water level fluctuations on algal communities of freshwater marshes.
Ph.D. Diss., Iowa State Univ., Ames, IA. A
244
-------
IOWA (continued)
Kallemeyn, L.S. and J.F. Novotny. 1977. Fish and fish food organisms in various habitats of the Missouri
River in South Dakota, Nebraska and Iowa. U.S. Fish & Wildl. Serv. FWS/OBS-77/25.IX + 100 pp. AI F
Mrachek, R.J. 1966. Macroscopic invertebrates on the higher plants at Clear Lake, Iowa. Iowa Acad. Sci.
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Poiani, K.A. and U.C. Johnson. 1989. Effect of hydroperiod on seed-bank composition in semi-permanent prairie
wetlands. Can. J. Bot. 67:856-864.
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Iowa. Ph.D. Diss., Iowa St. Univ., Ames.
Roosa, D.M. 1981. Marsh vegetation dynamics at Goose Lake, Hamilton County, Iowa: the role of historical,
cyclical, and annual events. Ph.D. Diss., Iowa St. Univ., Ames, IA. 205 pp.
Ruhr, C.E. 1951. Fish populations of a mining pit lake, Marion County, Iowa. M.S. Thesis, Iowa State Univ.
77pp. F
Smith, P.E. 1962. An ecological analysis of a northern Iowa Sphagnum bog and adjoining pond. Ph.D. Diss.,
Univ. Iowa, Iowa City. 156 pp.
Strohmeyer D.L. and L.H. Fredrickson. 1967. An evaluation of dynamited potholes in northwest Iowa. J. Uildl.
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Tebo, L.B. 1955. Bottom fauna of a shallow eutrophic lake. Lizard Lake, Pocohontas County. Amer. Midi. Nat.
54:89-94.
Thompson, J.D. 1973. Feeding ecology of diving ducks on Keokuk Pool, Mississippi River. J. Wildl. Manage.
37:367-381. B
van der Valk, A.G. and C.B. Davis. 1979. A reconstruction of the recent vegetalionaI history of a prairie
marsh. Eagle Lake, Iowa, from its seed bank. Aquat. Bot. 6:29-51. TS P
Van Dyke, G.D. 1972. Aspects relating to emergent vegetation dynamics in a deep marsh, northcentral Iowa.
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Voigts, D.K. 1973. An odonate emergence trap for use in marshes. Proc. Iowa Acad. Sci. AI T
Weller, M.U. 1975. Studies of cattail in relation to management for marsh wildlife Iowa State. J. Sci.
49:383-412.
Weller, M.W. and L.H. Fredrickson. 1973. Avian ecology of a managed marsh. Living Bird 12:269-291.
Weller, M.W. and C.S. Spatcher. 1965. Role of habitat in the distribution and abundance of marsh birds.
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following a drought. Proc. Iowa Acad. Sci. 90(2): 50-54.
245
-------
Inland Wetlands Having Biologica
Community Measurements
ACCURACY OF SITE LOCATIONS E5TIHATED TO BE » or - I 0m,
6 Research Study Site
g Migratory Shorebird Survey CBSB) site
Q Breed,n3 Bird Census site that includes uetland
O Annual Christmas Bird Count area CIS-mile diameter)
+ Breeding Bird Surrey Stortina points for 2Smi transects
SITE LOCATED IN COUNTY, SPECIFIC LOCATION(S) NOT PLOTTED
State/Federal waterfoul survey
Idaho
Th i * map do*s NOT por tf ay ALL w«t I and sanpl i rig • i
coI I*c t«d S*« chapter I for iocIu«ion crit«r i a
Sit«» ar* r«f»r«nc«d by cod« numb«r lo lh« accompanyt
stata bibJ togrophy
USEPA Env l ronatn t*l R**««reh L*bor»tor/* Cor»«)l
Data Conp i I at t on Pau I Adanus and Rob i n R«nt«r i a Cor logrophy J»f f Ir t »h
Oregon
246
-------
IDAHO
Mapped
ID1
Wolf, K. 1955. Some effects of fluctuating and falling water levels on waterfowl production. J. Uildl.
Manage. 19(1):13-23. B I
ID2
Halford, O.K., O.D. Markham, and R.L. Dickson. 1982. Radiation doses to waterfowl using a liquid radioactive
waste disposal area. J. Uildl. Manage. 46(4):905-913. B
ID3
Gregg, W.W. and F.L. Rose. 1982. The effects of aquatic macrophytes on the stream microenvironment. Aquat.
Bot. 14:309-324. PM
ID4-5
Jensen, S., R. Ryel and W.S. Platts. 1989. Classification of riverine/riparian habitat and assessment of
nonpoint source impacts. North Fork Humboldt River, Nevada. USDA Forest Service, Intermountain Res. Stn.,
Boise Fisheries Unit.
ID6
Army Corps, of Engineers. 1983. Interim Feasibility Report. Clear Lakes Hydropower, Snake River, ID.
ID7
Rumely, J.H. 1956. Plant ecology of a bog in northern Idaho. Ph.D. Dissertation, Washington St. Univ.,
Pullman. 85 pp.
IDS
Miller, T.B. 1976. Ecology of riparian communities dominated by white alder in western Idaho. M.S. Thesis,
Univ. Idaho, Moscow. 154 pp.
ID9
Manuel-Faler, C.Y. 1981. Production and fate of aquatic macrophytes in Deep Creek, Idaho. Ph.D. Dissertation,
Idaho St. Univ., Pocatello. 127 pp.
ID10
Tuhy, J.S. 1981. Stream bottom community classification for the Sawtooth Valley, Idaho. M.S. Thesis, Univ.
Idaho, Moscow. 175 pp.
ID11
Tuhy, J.S. 1982. Riparian Classification for the Upper Salmon/Middle Fork Salmon River Drainages, Idaho.
Smithfield Associates, Smithfield, UT. 153 pp.
ID12-13
Mutz, K. and J. Queiroz. 1983. Riparian Community Classification for the Centennial Mountains and South Fork
Salmon River, Idaho. Meiji Resource Consultants, Layton, UT. 96 pp.
ID14
Youngblood, A.P., W.G. Padgett, and A.M. Winward. 1985. Riparian Community Classification of Eastern Idaho-
Western Wyoming. Res. Rep. R4-Ecol-85-01. USDA Forest Serv., Ogden, UT. 78 pp.
IDBBS1-
U.S. Fish & Uildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
IDBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
IDCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, MY. B
247
-------
IDAHO (continued)
Not Mapped
Asherin, D.A. and J.J. Claar. 1976. Inventory of riparian habitats and associated wildlife along the
Columbia and Snake Rivers. Idaho Coop. Uildl. Res. Unit, Univ. Idaho, Vol. 3A and 3b.
Breckenridge, P.p., L.R. Wheeler, and J.F. Ginsburg. 1983. Biomass production and chemical cycling in a man-
made geothermal wetland. Wetlands 3:26-43.
Falter, C.H., J. Leonard, R. Naskali, F. Rube, and H. Bobisud. 1974. Aquatic macrophytes of the Columbia and
Snake River Drainage. College For. and Dept. Biol. Sci., Univ. Idaho, Moscow, ID. PM
Huschle, G. 1975. Analysis of the vegetation along the middle and lower Snake River. Master Thesis, Univ.
Idaho, Moscow, ID. P
Minshall, G.W. 1981. Structure and temporal variations of the benthic macroinvertebrate community inhabiting
Mink Creek Idaho, USA, a third order rocky mountain stream. J. Freshw. Ecol. 1:13-26.
Oring,L.W. 1964. Behavior and ecology of certain ducks during the postbreeding period. J. Wildl. Manage.
28:223-233. B
Rumely, J.H. 1956. Plant ecology of a bog in northern Idaho. Ph.D. Diss., Washington St. Univ., Pullman, WA.
93 pp.
Steete P.E., P.O. Dalke, and E.G. Bizeau. 1956. Duck production at Gray's Lake, Idaho, 1949-1951. J. Wildl.
Manage. 20(3):279-285.
Workman, G.W. 1963. An ecological study of the Bear Lake littoral zone, Utah-Idaho. Ph.D. Diss., Utah St.
Univ., Logan, UT. 104 pp.
248
-------
Inland Wetlands Having BioIogica
Community Measurements
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE + o- - ' ft- VVT / ^i-' / - ' I HO 1 S
9 Research Study S,t*
B Migratory Shorebird Survey CBSB) site
Q Breeding Btrd Census site that includes wetland
O Annual Christmas Bird Count area {15-mile diameter)
. . Thi» map doe» NOT portray ALL w«lland sonip I i ny *it*»
+ Breeding Bird Survey Starting points for 25mi transects
AND po.nts where transects enter new county Most cover Empha«.« •• on «tt«* wh»r» commun.Iy-1 eve I data w«r«
mainly non-w»ttand habitat collected $•• chapter 1 for inclusion crit*ria
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONCS) NOT PLOTTED
+ State/Federal waterfowl survey
USCPA £nviron*tntat Rataarch
Site* are referenced by code number to the accompanying
mtot* bib)logrophy
oratory. CorvaI I i»• Or•aon
Data Compilation Paul Adomu* and Robin Renteria
Car tography Jeff Iri»h
250
-------
ILLINOIS
Happed
IL2
Clark, U.D. and J.R. Karr. 1979. Effects of highways on Red-Winged Blackbirds and Horned Lark populations.
Wilson Bull. 91(1):143-145. B I
IL3
Wiley, M.J., R.W. Gorden. S.W. Waite, and T. Powers. 1984. The relationship between aquatic macrophytes and
sport fish production in Illinois ponds: A simple model. N. Amer. J. Fish. Manage. 4:111-119. F
IL4
Brown, S. and L. Giese. 1988. Tree Growth Rates and Regeneration of Buttonland Swamp, Southern Illinois.
Final Report to Illinois Dept. of Conserv., Cache River Basin Study. Dept. of Forestry, Univ. of Illinois,
Urbana, IL. 67 pp. PW
IL5
Yeager, L.E. 1949. Effect of Permanent Flooding on a River Bottom Timber Area. Bull. III. Nat. History
Surv., Urbana, IL 25(2):33-65. PW
IL6-8
Jones, D.W., M.J. McEUigott, and R.H. Mannz. 1985. Summary of Biological, Chemical, and Morphological
Characterizations of 33 Surface-Mine Lakes in Illinois and Missouri, pp. 211-238 In: R.P. Brooks, D.E. Samuel,
and J.B. Hill (eds.). Wetlands and Water Management on Mined Lands. Penn. St. Univ., University Park, PA.
R
IL9-12
Lawrence, J.S., W.D. Kilmstra, W.G. O'Leary, and G.A. Perkins. 1985. Contribution of surface-mined wetlands
to selected avifauna in Illinois, pp. 317-326 Penn. St. Univ., University Park, PA. In: R.P. Brooks, D.E.
Samuel, and J.B. Hill (eds.). Wetlands and Water Management on Mined Lands. Penn. St. Univ., University
Park, PA. B
IL13-14
Paller, M.H., R.C. Heidinger, and W.M. Lewis. 1988. Impact of nonchlorinated secondary and tertiary effluents
on warm water fish communities. Water Res. Bull. 24(1):65-76. F I
IL15-16
Robertson, P.A., M.D. Mackenzie, and L.F. Elliott. 1984. Gradient analysis and classification of the woody
vegetation for four sites in southern Illinois and adjacent Missouri. Vegetalio 58:87-104. PW
IL17
Urbanek, R.P. and W.D. Klimstra. 1986. Vertebrates and vegetation on a surface-mined area in southern
Illinois. Trans. Illinois Acad. Sci. 79(3):175-187. B P
IL18
Mitsch, W.J. and W.G. Rust. 1984. Tree growth responses to flooding in a bottomland forest in northeastern
Illinois.For. Sci. 30:499-510. PW
IL18
Mitsch, W.J., C.L. Dorge, and Wiemhoff, Jr. 1979. Ecosystem dynamics and a phosphorus budget of an alluvial
cypress swamp in southern Illinois. Ecol. 60(6):1116-1124. WQ
IL19
Pinkowski, R.H., G.L. Rolfe, and L.E. Arnold. 1985. Effect of feedlot runoff on a southern Illinois forested
watershed. J. Environ. Qual. 14(1):47-54. I P
IL20
Grubaugh, J.W., R.V. Anderson, D.M. Day, K.S. Lubinski, and R.E. Sparks. 1986. Production and fate of organic
material from Sagittaria latifolla and Nelumbo lutea on Pool 19, Mississippi River. J. Freshw. Ecol.
3(4):477-484. SO P
IL21
U.S. Fish & Wildl. Service. 1989. Long Term Resource Monitoring Program for the Upper Mississippi River
System. First Annual Report. Environ. Manage. Tech. Center, Onalaska, WI. 85pp + Apps. B TS
251
-------
ILLINOIS (continued)
IL22
Bel I rose, F.C., F.L. Paveglio, Jr., and D.W. Steffeck. 1979. Waterfowl populations and the changing
environment of the Illinois River Valley. Illinois Nat. Hist. Surv. Bull. 32(1):54. B I TS
IL23
Kwak, Thomas J. 1988. Lateral movement and use of floodplain habitat by fishes of the Kankakee River,
Illinois. Amer. Midi. Nat. 120(2):241-149.
IL24
Uetz, G.W., K.L. Van der Laan, G.F. Summers, P.A. Gibson, and L.L. Getz. 1979. The effects of flooding on
floodplain arthropod distribution, abundance and community structure. Amer. Midi. Nat. 101(2):286-299. AI
IL25
Greenfield, D.W. and J.D. Rogner. 1984. An assessment of the fish fauna of Lake Calumet and its adjacent
wetlands, Chicago, IL: Past, Present and Future. Trans. Illinois Acad. Sci. 77(1):77-93. F TS
IL26
Peterson, Mark J. 1988. The vascular flora, fishes, and aquatic macroinvertebrates of Lovets Pond, Jackson
County, Illinois. M.S. Thesis, Biol. Sci. Prog. Grad. School, Southern Illinois Univ., Carbondale, IL. 74 pp.
1L27
Cole, C.A. 1988. Wetland ecosystem development on a reclaimed surface coal mine in southern Illinois.Ph.D.
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ILBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
Ithaca, NY. B
ILBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
ILBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, MA. B
ILBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
ILCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
ILLTR
Anderson, R.V. et al. In Process. Long Term Environmental Research Wetland Site: Illinois (Pool 19) LTER
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ILLTR
Sparkes, R.E. et al. In Process. Long Term Environmental Research Wetland Site: Illinois and Mississippi
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Baker, F.C. 1910. The ecology of the Skokie Marsh area, with special reference to the Mollusca. Bull. Illinois
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Bell, D.T. 1974. Tree stratum composition and distribution in the streamside forest. Amer. Midi. Natur.
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252
-------
ILLINOIS (continued)
Bell. D.T. and R. del Moral. 1977. Vegetation gradients in the streamside forest of Hickory Creek, Wil County,
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Cole, C.A. 1988. Wetland ecosystem development on a reclaimed surface coal mine in southern Illinois. Ph.D.
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Coss, R.D. 1981. Wildlife habitats provided by aquatic plant communities of surface mine lakes. M.S. Thesis,
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Fausch, K.D., J.R. Karr, and P.R. Yant. 1984. Regional application of an index of biotic integrity based on
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Goff, C.C. 1952. Flood-plain animal communities. Amer. Midi. Natur. 47(2):478-486.
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Harper, M. 1938. The ecological distribution of earthworms as found in developmental stages of the
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Henebry, M.S. and R.W. Gordon. 1989. Microbial populations of a managed wetland, pp. 1019-1028 In: R.R.
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Hosner, J.F., and L.S. Minckler. 1960. Hardwood reproduction in the river bottoms of southern Illinois.
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Hosner, J.F., and L.S. Minckler. 1963. Bottomland hardwood forests of southern Illinois-regeneration and
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Jackson, H.O. and W.C. Starrett. 1959. Turbidity and sedimentation at Lake Chautaqua, Illinois. J. Wildl.
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Johnson, F.L. and D.T. Bell. 1976. Plant biomass and net primary production along a flood-frequency gradient
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Krull, J.N. and W.A. Hubert. 1973. Seasonal abundance and diversity of benthos in a southern Illinois swamp.
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Larimore, R.W. and P.W. Smith. 1963. The fishes of Champaign County, Illinois, as affected by 60 years of
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253
-------
ILLINOIS (continued)
O'Leary, W.G. 1984. Uaterfoul habitats provided by surface mine wetlands in southwestern Illinois. MA Thesis,
Dept of Zoology, Southern Illinois Univ., Carbondale, IL. 141 pp. B P
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Perkins, G.A. and J.S. Lawrence. 1981. Bird use of wetlands created by surface mining. Trans. Illinois Acad.
Sci. 78. M B
Peterson, O.L. and G.L. Rolfe. 1982. Nutrient dynamics of herbaceous vegetation in upland and floodplain
forest communities. Amer. Midi. Nat. 107:325-339. PE
Schlosser, I.J. 1985. Flow regime, juvenile abundance, and the assemblage structure of stream fishes. Ecol.
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Smith, J.R. 1986. Ecological relations of fauna and flora on a pre-law coal surface-mined area in Perry Co.,
Illinois. Ph.D. Diss., Dept. of Zoology, Southern Illinois Univ., Carbondale, IL. 265 pp. B P
Smith, P.W. 1971. Illinois Streams: A Classification Based on Their Fishes and an Analysis of Factors
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Sponsler, M.L. 1982. A regional comparison of avian populations on unmined and reclaimed lands in Illinois.
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Urbanek, R.P. 1976. Vertebrate and floral diversity on strip-mined land in Williamson and Saline counties,
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Verts, B.J. 1956. An evaluation of wildlife and recreational values of a strip-mined area. M.S. Thesis,
Southern Illinois Univ., Carbondale. 60 pp.
254
-------
Inland Wetlands Having Biological
Communi ty Measurements
Indi
TKi* *op do«» NOT portray ALL w«tland •anpling sit«»
coI I•et*d S*« chapter t for incIu*ion cr r t*r ia
Sit«» ar* r«f«r«nc*d by cod* number to th« accompanying
•lat« bibltography
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE + or - 10mi
0 Research Study Sits
f Migratory Shorebird Surrey CBSB) s.te
Q Breedmg Bird Census CFBC) s,ie thai .nclodes wetland
O Annual Christmas B(rd Count area CI5-m i ie diameter)
Most cover mainly non-w«l!and Kabilot
H- Breeding Bird Survey Starting points lor 25mi transects
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONCS) NOT PLOTTED
* State/Federal waterfowl survey
USEPA Env tr on»*nt«l Research Laboratory, CorvalM** Or«0«n
Data Comp > I ation PauI Adamu* and Robin Renteria Car togr aphy J*ff Iri»h
256
-------
INDIANA
Happed
IN1
Uitcox, D.A., N.B. Pavlovic, andM.L. Mueggler. 1985. Selected ecological characteristics of Scirpus cyperinus
and its role as an invader of disturbed wetlands. Wetlands 5:87-97. PE
IN2
Uilcox, D.A., R.J. Shedlock, and W.H. Hendrickson. 1986. Hydrology, water chemistry and ecological relations
in the raised mound of Cowles Bog. J. of Ecol. 74:1103-1117. P
IN2
Uilcox, D.A., S.I. Apfelbaum, and R.D. Hiebert. 1984. Cattail invasion of sedge meadows following hydrologic
disturbance in the Cowles Bog Wetland Complex, Indiana Dunes National Lakeshore. Wetlands 4:115-128. P
IN2
R.D. Hiebert, D.A. Wilcox, and N.B. Pavlovic. 1986. Vegetation patterns in and among pannes (Calcareous
Intradunal Ponds) at the Indiana Dunes National Lakeshore, Indiana. Amer. Midi. Nat. 116(2):276-281.
IN2
Wilcox, D.A. and H.A. Simonin. 1987. A chronosequence of aquatic macrophyte communities in dune ponds.
Aquatic Bot. 28:227-242. P
IN2
Jackson, S.T., R.P. Futyma, and D.A. Wilcox. 1988. A paIeoecological test of a classical hydrosere in the
Lake Michigan Dunes. Ecol. 69(4):928-936. P
IN2.5
Wilcox, D.A., S.I. Apfelbaum, and R.D. Hiebert. 1985. Cattail invasion of sedge meadows following hydrologic
disturbance in the Cowles Bog Wetland Complex, Indiana Dunes National Lakeshore. Wetlands 4:115-128. PE I
IN3
Wilcox, D.A. 1986. The effects of deicing salts on vegetation in Pinhook Bog, Indiana. Can. J. Bot.
64:865-874. P I
IM
Wilcox, D.A. and R.E. Andrus. 1987. The role of Sphagnum fimbriatum of secondary succession in a road salt
impacted bog. Can. J. Bot. 65:2270-2275. PB I
INBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
Ithaca, NY. B
INBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
INBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, MA. B
INBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
INCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
Not Mapped
Clifford, H.F. 1966. The ecology of invertebrates in an intermittent stream. Invest. Indiana Lakes & Streams
7:57-98. AI
257
-------
INDIANA (continued)
Cortwright, S.A. 1987. Impacts of species interactions and geographical-historical factors on larval amphibian
community structure. Ph.D. Diss., Indiana Univ., Bloomington, IN. 237 pp.
Hendrickson, W.H. and D.A. Uilcox. 1979. Relationship between some physical properties and the vegetation
found in Cowles Bog National Natural Landmark, Indiana. Proc. Sec. Conf. Sci. Res. Nat. Parks. 5:642-666. P
Hiebert, R.D., D.A. Uilcox, and N.B. Pavlovic. 1986. Vegetation patterns in and among pannes (calcareous
intradunal ponds) at the Indiana Dunes National Lakeshore. Amer. Nat. 116:276-281. PE
Landers, D.H. 1982. Effects of naturally senescing aquatic macrophytes on nutrient chemistry and chlorophyll
a of surrounding waters. Limnol. Oceanogr. 27:428-439. PM
Lindsey, A.A., R.O. Petty, O.K. Sterling, and W. Van Asdall. 1961. Vegetation and environment along the
Uabash and Tippecanoe rivers. Ecol. Monogr. 31:105-156. PW
258
-------
Inland Wetlands Having Biological
Community Measurements
Kansas
'&?-—' -."?.
^T* ^ » ' ; •
a ,
U- c
*
i
u-- V- ' ^ -j- -i- ,-----
-- -., , , •-' _ ; \ / + •
- +'''^'J _1_
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+ ' ^^ ' 1 vv r"^° ; ^*'i - ' * ' ; 1
' /~^ ( /' ''--T^''1 ! «^J j^~M
^ — -^Q-,' + -O^ : — - — 'r -) ^ ;t*)i Q
^ : +i ' i ^'O '-^3' *+" ^
• ' + • ' ^ __. , ': -J_— L±-J
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE * or - I 0m i
9 Research Study Site
I Migratory Shorebird Survey CBSB) site
Q Breeding Bird Census CBBO sit» that includes ueUand
O Annual Christmas 8
-------
KANSAS
Happed
KS1
Kansas Biological Survey and Kansas Geological Survey. 1987. Cheyenne Bottoms: An Environmental Assessment.
Univ. of Kansas, Rep. # 32. Submitted to the KS Fish and Game Comtn.
KS2-15
Shipley, F.S. 1980. Habitat distribution and fecundity in a marsh-upland bird species series. Ph.D. Oiss.,
Kansas State Univ. 140 pp. B
KSBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
Ithaca, NY. B
KSBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
KSBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, HA. B
KSBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
KSCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
Not Happed
Anderson, R.G. 1954. A palynological study of a fresh water marsh in Atchison County, Kansas. M.A. Thesis,
Dept. of Bot., Univ. Kansas, 21 pp.
Bellah, R.G. 1969. Forest succession on the Republican River floodplain in Clay County, Kansas. Ph.D. Diss.,
Kansas St. Univ., Manhatten. 43 pp.
Berger, T.J. 1985. Community ecology of pond-dwelling anuran larvae. Ph.D. Diss., Univ. Kansas, Lawrence.
150 pp.
Uorthen, G.L. 1976. The influence of weather on avian activity in an eastern Kansas riparian woodland. Ph.D.
Diss., Kansas St. Univ., Manhatten, KS.
261
-------
0>
o
00
o
X
"O
c
o
-------
KENTUCKY
Happed
KY1-3
Bosserman, R.W. and P.L. Hill, Jr. 1985. Community ecology of three wetland ecosystems impacted by acid mine
drainage, pp. 287-304 In: R.P. Brooks, D.E. Samuel, and J.B. Hill (eds.). Wetlands and Water Management on
Mined Lands. Penn. St. Univ., University Park, PA. P I
KY4
Baker, J.A., C.H. Pennington, C.R. Bingham, and I.E. Winfield. 1987. An Ecological Evaluation of Five
Secondary Channel Habitats in the Lower Mississippi River. U.S. Army Corps of Engr., Mississippi River Comm.,
Lower Mississippi River Environ. Prog., Rep. 7.
KY5
Sigrest, J.M. and S.P. Cobb. 1987. Evaluation of Bird and Mammal Utilization of Dike Systems along the Lower
Mississippi River. U.S. Army Corps of Engr., Mississippi River Commission, Lower Mississippi River Environ.
Prog. Rep. 10. 103 pp.
KYBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
KYBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, MA. B
KYBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
KYCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
Not Mapped
Hill, P.L. 1983. Wet I and-stream ecosystems of the western Kentucky coalfield: environmental disturbance and
the shaping of aquatic community structure. Ph.D. Diss., Univ. Louisville, Louisville, KY. 349 pp.
Steenis, J.H. 1947. Recent changes in the marsh and aquatic plant status at Reelfoot Lake. J. Tennessee
Acad. Sci. 22:22-27. RS P
Taylor, J.R., P.L. Hill, R.U. Bosserman, and W.J. Mitsch. 1982. Ecosystem analysis of selected wetlands in
the western Kentucky coalfield, pp. 75-85 In: B.R. Mcdonald (ed.). Proc. Symposium Wetland Unglaciated
Appalachian Reg., West Virginia Univ., Morgantown, WV. P R
Taylor, J.R. 1985. Community structure and primary productivity of forested wetlands in western Kentucky.
Ph.D. Diss., Univ. Louisville, Louisville, KY. 151 pp.
263
-------
Inland Wetlands Having Biological
Community Measurements
Louisiana
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE * of - I 0mi
0 Research Study Sitt
H Migratory Shorebird Survey (BSB> *,te
Q 9r**dtna 8>rci C»n»u» (BBC) «ii« that i nc I ud»« w»tland
Q AnnuaI CKri»tmo» Bird Count ar»a CIS-mil* diam*t *r >
Mo»t cover mainlv non-*4«tl and hob i tat
. „ , Thi« nap doe* NOT portray ALL wetland «a*plhng «it»*
T Br«%amg Bird Sur v«y Starting point* for 25mi tron»«et«
AND point* wh»r« tron««ct« «nl«r n.w county Mo.t cover Ef.pha«)« >• on •. U« uh«r« commun, t y- I «v« t data u*r«
mainly non-wetland habitat collected See chapter I for tnclu«ion criteria
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONS) NOT PLOTTED
• Slate/Federal waterfowl aurvey
Siie* are referenc*d by cod* number to the accompanying
•tale bibliography
USEPA En* i
Rt«*«r«h
Data Compilation Paul Adanue and Robin Renter to Cartography Jeff Iri*h
264
-------
LOUISIANA
Happed
LA1-14
Landin, M.C. 1985. Bird and mammal use of selected lower Mississippi River borrow pits. Ph.D. Diss.,
Mississippi State Univ., 405 pp. B MA
LA1-14
Buglewicz, E.G., W.A. Mitchell, J.E. Scott, M. Smith, and W.L. King. 1988. A Physical Description of Main
Stem Levee Borrow Pits Along the Lower Mississippi River; Lower Mississippi River Environmental Program, Rep.
2. U.S. Army Corps of Engr., Mississippi River Commission, Lower Mississippi River Environ. Prog. Vicksburg,
MS. 205 pp.
LA1-14
Cobb, S.P., C.H. Pennington, J.A. Baker, and J.E. Scott. 1984. Fishery and ecological investigations of main
stem levee borrow pits along the lower Mississippi River. Mississippi R. Comm., Vicksburg, MS. 120 pp. f
LA1 -14
U.S. Army Engineer Waterways Expt. St. Environ. Lab. 1986. Bird and Mammal Use of Main Stem Levee Borrow Pits
Along the Lower Mississippi River. U.S. Army Corps of Engr., Mississippi River Commission, Lower Mississippi
River Environ. Prog., Rep. 3. 137 pp.
LA15
Pollard, J.E., S.M. Melancon, and L.S. Blakey. 1983. Importance of bottomland hardwoods to crawfish and fish
in the Henderson Lake area, Atchafalaya Basin, Louisiana. Wetlands 3:1-25. F AI
LA16
Kemp, G.P. 1978. Agricultural runoff and nutrient dynamics of a swamp forest in Louisiana. M.S. Thesis,
Louisiana State Univ., Baton Route, LA. 58 pp.
LA16
Kemp, G.P., U.H. Conner, and J.W. Day, Jr. 1985. Effects of flooding on decomposition and nutrient cycling
in a Louisiana swamp forest. Wetlands 5:35-51. D I
LA18
Klimas, C.V. 1987. Baldcypress response to increased water levels, Caddo Lake, Louisiana-Texas. Wetlands
7:25-37. PW I
LA19-22
Lambou, V.W. 1959. Fish populations of backwater lakes in Louisiana. Trans. Amer. Fish. Soc. 88:7-15.
LA24
Conner, W.H., J.G. Gosselink, and R.T. Parrondo. 1981. Comparison of the vegetation of three Louisiana swamp
sites with different flooding regimes. Amer. J. Bot. 68(3):320-331. PW
LA25
Pollard, J.E., S.M. Melancon, and L.S. Blakey. 1983. Importance of bottomland hardwoods to crawfish and fish
in the Henderson Lake area, Atchafalaya Basin, Louisiana. Wetlands 3:1-25. F AI
LA26
Sklar, F.H. and W.H. Conner. 1979. Effects of altered hydrology on primary production and aquatic animal
populations in a Louisiana swamp forest, pp. 191-208 In: Proc. Coastal Marsh Estuary Manage. Symposium (3Rd).
Baton Rouge, LA. AI F P I
LA26
Sklar, F.H. and W.H. Conner. 1983. Swamp forest communities and their relation to hydrology: The impacts of
artificial canals, pp. 245-272 In: R.J. Varnell (ed.). Water Quality Wetland Manage. Conf. Proc., New Orleans,
LA. P I
LA27-28
Gunning, G.E. and R.D. Suttkus. 1984. Stream pollution monitoring using species composition of fish
populations and water quality data. Lousiana State Univ., Coast. Ecol. Lab., Pub. # LSU-CEL-83-13. F
265
-------
LOUISIANA (continued)
LA29-30
Faulkner, S.P. and W.H. Patrick,Jr.- n.d. Characterization of Bottomland Hardwood Wetland Transition Zones
in the Lower Mississippi River Valley. U.S. Army Engineers Waterways Exp. Stn., Vicksburg, MS. Appendix A,
14 pp. P
LA35,39
Baker, J.A., R.L. Kasul, L.E. Winfield, C.R. Bingham, C.H. Pennington, and R.E. Colentan. 1988. An Ecological
Investigation of Revetted and Natural Bank Habitats in the Lower Mississippi River. Rep.9, Lower Mississippi
River Environ. Prog., U.S. Army Engineers Waterways Exp. Stn., Vicksburg, MS. 81 pp.
LA38
Felton, M., J.J. Cooney, and W.C. Moore. 1966. A quantitative study of the bacteria of a temporary pond,
Florenville, Louisiana. J. Gen. Microbiol. 47:25-31. MI
LA42
Conner, W.H. and J.W. Day, Jr. 1984. The Impacts of Increased Flooding on Commercial Wetland Forests in the
Lake Verret watershed. Final Rep. to Louisiana Bd. Regents, Res. and Dev. Prog., Baton Rouge, LA. PW I
LA42
Conner, W.H. and J.W. Day, Jr. 1988. The impact of rising water levels on tree growth in Louisiana, pp. 219-
224. In: Hook, D.D., et. al (eds.). The Ecology and Management of Wetlands, Vol. 2: Management, Use, and
Value of wetlands. Croom Helm Ltd. Pub., England. PW I
LA42
Conner, W.H., W.R. Slater, K. McKee, K. Flynn, I.A. Mendelssohn, and J.W. Day, Jr. 1986. Factors controlling
the growth and vigor of commercial wetland forests subject to increased flooding in the Lake Verret, Louisiana
watershed. Final Rep. to Bd. of Regents Res. and Dev. Prog., Baton Route, LA. PW
LA42-43
Conner, W.H. and J.W. Day, Jr. 1988. Rising water levels in coastal Louisiana: Implications for two coastal
forested wetland areas in Louisiana. J. Coastal Res. 4(4).-589-596. PW I
LA44
Bowers, L.J. 1981. Tree ring characteristics of bald cypress in varying flooding regimes in the Barataria
Basin, Louisiana. Ph.D. Diss., Louisiana State Univ., Baton Rouge, LA, 159 pp. PW
LA45
Chambers, D.G. 1980. An analysis of nekton communities in the upper Barataria Basin, Louisiana. M.S. Thesis,
Louisiana State Univ., Baton Rouge, LA, 286 pp. AI
LA45
Butler, T.J. 1975. Aquatic metabolism and nutrient flux in a south Louisiana swamp and lake system. M.S.
Thesis, Louisiana State Univ., Baton Rouge, LA. 58 pp. PW
LA45
Day, J.W. Jr., T.J. Butler, and W.H. Conner. 1977. Productivity and nutrient export studies in a cypress
swamp and lake system in Louisiana. Estuarine Processes 11:255-269. P
LA45
Hopkinson, C.S. Jr. and J.W. Day, Jr. 1980. Modeling hydrology and eutrophication in a Louisiana swamp forest
ecosystem. Environ. Manage. 4(4):325-335. I
LA45
Kemp, G.P. and J.W. Day, Jr. 1981. Floodwater nutrient processing in a Louisiana swamp forest receiving
agricultural runoff. Louisiana Water Resour. Res. Inst., Louisiana State Univ., Baton Rouge, LA, Rep. No.
A-043-LA, 60 pp. PW
LA45
Kemp, G.P. and J.W. Day, Jr. 1984. Nutrient dynamics in a Louisiana swamp receiving agricultural runoff, pp.
286-293. In: K.C. Ewel and H.T. Odum (eds.). Cypress Swamps. Univ. Presses of Florida, Gainesville, FL. PW
266
-------
LOUISIANA (continued)
LA45
McNamara, S.J. 1978. Metabolism measurements of a flooded soil community in a Louisiana swamp forest. M.S.
Thesis, Louisiana State Univ., Baton Route, LA, 65 pp. PU
LA45
Sklar, F.H. 1983. Water budget, benthological characterization, and simulation of aquatic material flows in
a Louisiana freshwater swamp. Ph.D. Diss. Louisiana State Univ., Baton Rouge, LA. AI
LA45
Sklar, F.H. 1985. Seasonality and community structure of the backswamp invertebrates in a Louisiana
cypress-tupelo wetland, wetlands 5:69-86. AI
LA45
Sklar, F.H. and W.H. Conner. 1979. Effects of altered hydrology on primary production and aquatic animal
populations in a Louisiana swamp forest, pp. 191-210. In: J.W. Day, Jr., D.D. Culley, Jr., R.E. Turner, and
A.J. Humphry, Jr. (eds.). Proc. Third Coastal Marsh and Estuary Management Symposium, Lousiana State Univ.,
Div. Contain. Educ., Baton Route, LA. AI
LA45
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270
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MASSACHUSETTS
Happed
MA1
Peters, C.R. 1987. Peat stratigraphy evidence of the influence of hydrology on succession in a freshwater
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MA2
Pratt, J.M. and R.A. Coler. 1979. Ecological Effects of Urban Stormwater Runoff on Benthic Macroinvertebrates
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MA3
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MA4
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MA5-8,10
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MA11
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MA12
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MA13
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MA14
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MA15
Cole, G.A. and S.G. Fisher. 1978. Annual metabolism of a temporary pond ecosystem. Amer. Midi. Nat.
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MA15
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Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
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MABBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
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273
-------
MASSACHUSETTS (continued)
MABSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
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MABW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
MACBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
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274
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276
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MARYLAND
Happed
MD1-2
Whigham, D.F. and C.J. Richardson. 1988. Soil and plant chemistry of an Atlantic white cedar wetland on the
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MD3
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MD4
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MD4
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M05
Bartoldus, C. Unpub. Marley Creek, Maryland Wetland Monitoring, Annadale, VA. P
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Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
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MDBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
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MDBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
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MDBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
MDCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
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277
-------
Inland Wetlands Having Biologica
Community Measurements
Ma i ne
TK t « «\ap do»» NOT por trey ALL w«t I and *amp I i ng « > t«*
Empha*i• i* on *it«* wKnr• community-1«v*l dato w«r*
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Sit«» or* r«f«r«nc«d by cod* numb*r to th* accompanying
slat* bibIlography
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE + or - 10m,
• Research Study S< te
fl Mrgratory Shorebrrd Surv«y CBSB) site
("J Breedmg Bird Census (BBC) s,U that .ncludas wet I ond
O Annual Christmas Bird Count area (15-mile diameter)
Most cover mainly non-wet land habitat
~t- Breeding Bird Survey Starting pomts for 25m i transects
AND points where transects enter new county Most cover
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONCS) NOT PLOTTED
* State/Federal waterfowl survey
USEPA Environ«tni«l R»»*«rch Laboratory- Corvallls. Oregon
Data Compilotion PauI Adomus and Rob'n R*nl«ria
r Iogr aphy J*ff
278
-------
MAINE
Happed
ME1
Gibbs, K.E., T.M. Mingo, D.L. Courtemanch, and D.J. Stairs. 1981. The Effects on Pond Macroinvertebrates from
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ME2-3
Moring, J.R., G.C. Garman, and J. Mullen. 1985. The value of Riparian zones for protecting aquatic systems:
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ME4
Hunter, M.L., Jr., J.J. Jones, K.E. Gibbs, and J.R. Moring. 1986. Duckling responses to lake acidification:
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ME4
Hunter, M.L., Jr., J.J. Jones, and J.W. Witham. 1986. Biomass and species richness of aquatic macrophytes in
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ME5
Fefer, S.I. 1977. Uaterfowl Populations as Related to Habitat Changes in Bog Wetlands of the Moosehorn
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ME6
Ringelman, J.K. and J.R. Longcore. 1982. Movements and wetland selection by brood-rearing Black Ducks. J.
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ME7-13
Stockwell, S.S. 1985. Distribution and Abundance of Amphibians, Reptiles, and Small Mammals In Eight Types
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ME14
Jiffry, F. 1984. Loss of freshwater shellfish and some ecological impacts after water drawdown in Lake
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ME15.16
McAuley, D.G. and J.R. Longcore. 1988. Foods of juvenile Ring-Necked Ducks: Relationship to Wetland pH. J.
Wildl. Manage. 52(2):177-185. B R
ME15.16
McAuley, D.G. and J.R. Longcore. 1988. Survival of juvenile Ring-Necked Ducks on wetlands of different pH.
J. Wildl. Manage. 52(2):169-176. B R
ME19
Kenlan, K.H., G.L. Jacobson, Jr., and D.F. Brakke. 1983. Aquatic macrophytes and pH as controls of diversity
for littoral Cladocerans. pp. 63-83. In: Hendrey. (eds.). Early Biotic Responses to Advancing Lake
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MEBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
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MEBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
MEBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, MA. B
279
-------
MAINE (continued)
MEBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
MECBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
Not Mapped
Courtemanch, D.L. and K.E. Gibbs. 1979. The Effects of Sevin-4-Oil(R) on Lentic Communities: A Continuation
Study. Environ. Monitoring Coop. Spruce Budworm Control Proj., Maine 1978. Maine Dept. Conserv., Bur. For.
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Spencer, H.E. 1963. Man-made Marshes for Maine Waterfowl. Bull. No. 9, Maine Dept. Inland Fish. & Wildl.,
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Whitman, W.R. 1974. The response of macroinvertebrates to experimental marsh management. Ph.D. Dissertation,
Univ. Maine, Orono. 114 pp.
280
-------
Inland Wetlands Having Biological
Community Measurements
This nap do** NOT portray ALL w*tland soup Iing «it«
E»pha«i« is on «it*« wh«r« community-I«v*I data w«r
Co I I*ct»d S«« chapter 1 for incIu«ion crit«r t o
Sit** or* r*f*r»nc*d by cod* nu«b*r to th* accompany ing
•tat* bib'lography
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE * or - t0m,
9 Research Study Site
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* State/Federal waterfowl survey
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Do to Coup i I ot i ori PauV Adomu* ond Rob i n Renter i o Cartography Jeff Irish
282
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288
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Q Annual Christmas Bird Count area CIS-mile diameter)
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Data Coop i I at t on Pau I Adantu* and Rob i n R«n t«ria Cartography J»ff Irish
290
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MINNESOTA
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MN38,46,4
Siegel, D.I. 1981. Hydrogeologic setting of the evolution of patterned mires, Glacial Lake Agassiz peatland,
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292
-------
MINNESOTA (continued)
MN40
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HN41
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MN42
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MN43
Probst, J.R., D. Rakstad, and K. Brosdahl. 1983. Diversity of vertebrates in wildlife water-impoundments on
the Chippewa National Forest. North Center For. Serv., U.S. Dept. Agric., For. Serv., N. Central For.
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MN44
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MN45
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MN45
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MN46
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raised bog in the Lost River peatland, northern Minnesota. J. of Ecol. 74:393-401.
MN46
Boldt, D.R. 1985. Computer simulations of groundwater flow in a raised bog system. Glacial Lake Agassiz
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MN46
Chason, D.B. and Siegel, D.I. 1986. Hydraulic conductivity and related physical properties of peat, Lost
River peatland, northern Minnesota. Soil Sci. 142:91-99.
MN46
Glaser, P.H., J.A. Janssens, and D.I. Siegel. In Press. The response of vegetation to hydrological and
chemical gradients in the Lost River Peatland, northern Minnesota.
MN46
Janssens, J.A. and P.H. Glaser. 1986. The bryophyte flora and major peat-forming mosses at the Red Lake
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MN46
Seigal, D.I. and P.H. Glaser. 1987. Groundwater flow in a bog-fen complex. Lost River peatland, northern
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MN47
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Kirby, R.E. 1980. Waterfowl production estimates on forested wetlands from pair and brood counts. Wildl.
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MN48
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293
-------
MINNESOTA (continued)
MN49
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MN50
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MN50
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MN50
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MN50
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MN50
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MN51
Reiners, W.A. 1972. Structure and energetics of three Minnesota forests. Ecol. Monogr. 42:71-94. PW
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Reiners, W.A. and N.M. Reiners. 1970. Energy and nutrient dynamics of forest floors in three Minnesota
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MNBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
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MNBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
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MNBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
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MNBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
MNCBC1-
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MNLTR
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294
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MINNESOTA (continued)
Not Happed
Brown, J.M. 1972. Effect of overstory removal on production of shrubs and sedge in a northern Minnesota bog.
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Heuschele, A.L.S. 1968. The phenology of macrobenthos in a Mississippi River floodplain lake. Ph.D. Diss.,
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295
-------
MINNESOTA (continued)
Moyle, J.B. 1961. Aquatic invertebrates as related to larger water plants and waterfowl. Invest. Rep. 233,
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296
-------
Inland Wetlands Having Biological
Community Measurements
Missour i
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE * or -
9 Research Study Site _
| Migratorv Shorebird Survey (BSB) site
Q Breeding B.rd Census CBBC) site that incIudes wetland
O Annual Christmas Bird Count area (15-nnle diameter)
+ Breeding Bird Survey Starting points for 25mi transects
AND po i nt s *h»re transects ert-jr new county Most cover
SITE LOCATED IN COUNTY. SPECIFIC LXATIONCS) NOT PLOTTED
+ State/Federal waterfowl sur vey
01 U* ar» referenc*
slot* bib'logrophy
TKi* nap do** NOT portray ALL wetland *amp1ing *it«»
Emphas i * is on si t«s where coramuni lyl»v«l data were
coll «ct*d See chapter I for i nc I us < or» cr < i «r i a
USE PA EnvironBtntal R«»*i(*ch Libor«torjr> Cffi"* a I I l » , Or toon
Data Cortp ' I at ion Paul Adomus and Robin R*r>t«ri
Car t ography Jeff Ir ish
298
-------
MISSOURI
Mapped
M01-2
Landin, M.C. 1985. Bird and Mammal Use of Selected Lower Mississippi River Borrow Pits. Ph.D. Diss.,
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M05
Jones, O.W., M.J. McEUigott, and R.H. Mannz. 1985. Summary of Biological, Chemical, and Morphological
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M06
Robertson, P.A., M.O. Mackenzie, and L.F. Elliott. 1984. Gradient analysis and classification of the woody
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M06
Yanosky, T.M. 1982. Effects of Flooding Upon Woody Vegetation Along Parts of the Potomac River Flood Plain.
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M09
Kofron, C.P. and A.A. Schreiber. 1987. Observations on aquatic turtles in a northeastern Missouri marsh. SW
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M010-12
Berkman, H.E., C.F. Rabeni, and T.P. Boyle. 1986. Biomonitors of stream quality in agricultural areas: Fish
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M013
Stewart, E.M. and T.R. Finger. Diel Activity Patterns of Fishes in Lowland Hardwood Wetlands. Univ. Missouri,
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M014-15
U.S. Fish & Wildl. Serv. Ongoing studies.
M016
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M017-18
Neuswanger, D.J., W.W. Taylor, and J.B. Reynolds. 1982. Comparison of macroinvertebrate herpobenthos and
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M019
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M019
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Heitmeyer, M.E., L.H. Fredrickson, and G.F. Krause. 1989. Water and Habitat Dynamics of the Mingo Swamp in
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299
-------
MISSOURI (continued)
M019
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M019
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M019 «
Combs, D.L. 1987. Ecology of mallards during the winter in the upper Mississippi Alluvial Valley. Univ. of
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Ketley, J.R., Jr. 1986. Management and biomass production of moist-soil plants. Univ. of Missouri-Columbia,
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M019
McKenzie, D.F. 1987. Utilization of rootstocks and browse by waterfowl on moist-soil impoundments in Missouri.
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M022
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M023
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Reid, f. A. 1989. Differential habitat use by waterbirds in a managed wetlands complex. Univ. of
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U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
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300
-------
MISSOURI (continued)
MOBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
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MOBU1-
Missouri Department of Conservation. Unpub. Waterfowl census data. B
MOBW1-
U.S. Fish & Uildl. Service. Unpub. Waterfowl Survey Data. B
MOCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
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Not Mapped
Enns, W.R. 1967. Insects Associated with Midwestern Oxidation Lagoons. Terminal Prog. Rep., Univ. Missouri,
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Fredrickson, L.H. 1979. Floral and Faunal Changes in Lowland Hardwood Forests in Missouri Resulting from
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I B
Harvey, E.J. and J. Skelton. 1978. Relationship Between Hydrology and Bottom Land Vegetation in the Ozark
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301
-------
Inland Wetlands Having Biologica
Community Measurements
M t ssissipp
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE + or -
9 Research Study S.te
| Migraiory Shorebtrd Survey (BS3J site
Q Breed. ng 8>rd Census (BBC) s,'e that includes wetland
O Annual Christmas Bird Count area CIS-mil* diameter)
Most cover mainly non-wet land habitat
Thl* <>QP do** NOT Portray ALL wetland
, * r ic * .
T Breeding Bird Survey Starting points for 25m i transects
AND points where transects enter new county Most cover E«pha»i« <• on *it*« where conmun i t y- I eve 1
rnamlv non-wetland habitat collected Se» chapter I for inclusion c
SITE LOCATED IN COUNTY SPECIFIC LOCATIONCS) NOT PLOTTED
* State/Federal waterfowl survey
3< t«« are referenced by code number to the
•tale bibiiography
1ing »it*»
da to were
iter i a
accompanying
USE PA Environmental R*»e«rch Laboratory. Corv a tIi 9. Oregon
Data Conp > lot ion PauI Adonus and Rob in Renter i a Car tograpny Jeff Ir i*h
-------
MISSISSIPPI
Happed
MS1-5
Landin, M.C. 1985. Bird and Mammal Use of Selected Lower Mississippi River Borrow Pits. Ph.D. Diss.,
Mississippi State Univ. 405 pp. B MA
MS6-29
Pennington, C.H., H.L. Schramm,Jr., M.E. Potter, and M.P. Farrell. 1980. Aquatic Habitat Studies on the Lower
Mississippi River, River Mile 480 to 530. Rep. 5., Environ. Lab. U.S. Army Engr. Waterw. Expt.Stn. Vicksburg,
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MS6-29
Mathis, D.B., S.P. Cobb, L.G. Sanders, A.D. Magoun, and C.R. Bingham. 1981. Aquatic Habitat Studies on the
Lower Mississippi River, River Mile 480 to 530. Rep. 3. Environ. Lab. U.S. Army Engr. Waterw. Expt.Stn.,
Vicksburg, MS. Misc. Papers E-80-1. 83 pp. F AI
MS 6-29
Schramm, H.L., Jr. and C.H. Pennington. 1981. Aquatic habitat studies on the lower Mississippi River, River
Mile 480 to 530. Rep. 6., Environ. Lab. U.S. Army Engr., Uaterw. Expt.Stn., Vicksburg, MS. Misc. Paper
E-80-1. 74 pp. F
MS6
Pennington, C.H., H.L. Schramm,Jr., M.E. Potter, and M.P. Farrell. 1980. Aquatic Habitat Studies on the Lower
Mississippi River, River Mile 480 to 530. Rep. 5., Environ. Lab. U.S. Army Engr. Waterw. Expt.Stn. Vicksburg,
MS. Misc. Paper E-80-1. 101 pp. f
MS7
Newling, C.J. 1981. Ecological Investigation of a Greentree Reservoir in the Delta National Forest,
Mississippi. Environ. Lab., US Army Engr. Uaterw. Expt.Stat., Vicksburg, MS. 59 pp.. Misc. Paper El-81-5. P
MS10
Baker, J.A. and S.T. Ross. 1981. Spatial and temporal resource utilization by southeastern cyprinids. Copeia
1981:178-189. F
MS10
Ross, S.T. and J.A. Baker. 1983. The response of fishes to periodic spring floods in a southeastern stream.
Amer. Midi. Nat. 109(1):1-15. *
MS11
Cooper, C.M. 1987. Benthos in Bear Creek, Mississippi: Effects of habitat variation and agricultural
sediments. J. Freshw. Ecol. 4(1):101-113. AI I
MS11
Cooper, C.M. and J.W. Burns. 1984. Bryozoans--possible indicators of environmental quality in Bear Creek,
Mississippi. J. Environ. Qual. 13(1):127-130. AI
MS12
Faulkner, S.P. and W.H. Patrick,Jr. 1983. Characterization of Bottomland Hardwood Wetland Transition Zones
in the Lower Mississippi River Valley. U.S. Army Corps Engr., V-icksburg, MS. Appendix A, 14 pp. P
MS13,37
Cobb, S.P. and J.R. Clark. 1981. Aquatic Habitat Studies on the Lower Mississippi River, River Mile 480 to
530. Rep. 2, Environ. Lab. U.S. Army Engr. Waterways Expt.Stn. Vicksburg, MS, Misc. Paper E-80-1. 24 pp.
AI F
MS13.37
Cobb, S.P., C.H. Pennington, J.A. Baker, and J.E. Scott. 1984. Fishery and Ecological Investigations of Main
Stem Levee Borrow Pits Along the Lower Mississippi River. Mississippi R. Comm., Vicksburg, MS. 120 pp. F
MS18-20
Teels, B.M., G. Anding, D.H. Arner, E.D. Noorwood, and D.E. Wesley. 1976. Aquatic plant-invertebrate and
waterfowl associations in Mississippi. SE Assoc Game & Fish Comm. 13th Ann. Conf. AI
303
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MISSISSIPPI (continued)
MS22-27
Cooper, C.M. 1987. Benthos in Bear Creek, Mississippi: Effects of habitat variation and agricultural
sediments. J. Freshw. Ecol. 4<1):101-113.
MS23
Baker, J.A., C.H. Pennington, C.R. Bingham, and I.E. Winfield. 1987. An Ecological Evaluation of Five
Secondary Channel Habitats in the Lower Mississippi River. U.S. Army Corps of Engr., Mississippi River Conrc.,
Lower Mississippi River Environ. Prog., Rep. 7. Vicksburg, MS.
MS28
Ross, S.T. and J.A. Baker. 1983. The response of fishes to periodic spring floods in a southeastern stream.
Amer. Midi. Nat. 109(1):1-15. F
MS30
Kaminski, R.M. nd. Waterbird Use of "Moist-Soil" Impoundments in Noxubee National Wildlife Refuge,
Mississippi. Unpub. Report. Dept. Wildl. and Fisheries, Mississippi S. Univ., 32 pp. B
MS31
U.S. Environmental Protection Agency. 1983. Hydrographic, Water Quality and Biological Studies of Freshwater
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MS32-39
Webb, J.W. and C.V. Klimas. 1988. Vegetation Development on Revetments Along the Lower Mississippi River.
U.S. Army Corps of Engr., Mississippi River Commission, Lower Mississippi River Environ. Prog., Rep. 15.
Vicksburg, MS.
MS36
Beckett, D.C., C.R. Bingham, and L.G. Sanders. 1983. Benthic macroinvertebrates of selected habitats of the
lower Mississippi River. J. Freshw. Ecol. 2(3):247-261. AI
MS37
Conner, J.V., C.H. Pennington, and T.R. Bosley. 1983. Larval Fish Selected Aquatic Habitats on the Lower
Mississippi River. Tech. Rep. E-83-4, U.S. Army Engr. Waterw. Expt.Stn. CE, Vicksburg, MS. F
MS40-43
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304
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305
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MONTANA
Happed
MT1
Lee, L.C., T.M. Hinckley, and M.L. Scott. 1985. Plant water status relationships among major floodplain sites
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MT2
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MT3
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MT5, 6
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MT7
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MT8
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MT9
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MONTANA (continued)
Foote, G.G. 1965. Phytosociology of the bottomland hardwood forests in western Montana. M.S. Thesis, Univ.
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308
-------
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NORTH CAROLINA
Mapped
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NC6
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NC8
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NC13
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NC13
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NC14
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NC14
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NORTH CAROLINA (continued)
NC16.24
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NORTH CAROLINA (continued)
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Inland Wetlands Having Biological
Community Measurements
North Dakota
'$ l^^vT;*aS
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE « or - IBim
C Research Study S.te
| Migratorv Shorebird Survey (BSB) s,t»
Q Breeding Bird Census
-------
NORTH DAKOTA
Happed
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ND13
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NORTH DAKOTA (continued)
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Dakota. Ph.D. Diss., North Dakota St. Univ., Fargo, ND. 251 pp.
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Larson, G.E. 1979. The aquatic and wetland vascular plants of North Dakota. Ph.D. Diss., North Dakota State
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Malterer, T.J., A.J. Duxbury, and J.L. Richardson. 1987. Soil character of three calcareous fens in North
Dakota and Minnesota. N. Dakota Acad. Sci. 41:65. P
Reily, P.W. and U.C. Johnson. 1982. The effects of altered hydrologic regime on tree growth along the Missouri
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Rossiter, J.A. and Crawford R.D. 1981. Evaluation of Constructed Ponds as a Means of Replacing Natural
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Woodin, M.C. 1987. Wetland selection and foraging ecology of breeding diving ducks. Ph.D. Diss., Univ.
Minnesota, Minneapolis. 125 pp.
318
-------
0
o
o>
0
(0
~D
C
O
TJ
C
O
0)
-------
NEBRASKA
Happed
NE1-4
Erickson, N.E. and D.M. Leslie, Jr. 1987. SoiI-vegetation correlations in the Sandhills and Rainwater Basin
wetlands of Nebraska. U.S. Fish & Wildl. Serv. Biol. Rep. 87(11):73. P
NE5-8
Kallemeyn, L.S. and J.F. Novotny. 1977. Fish and fish food organisms in various habitats of the Missouri
River in South Dakota, Nebraska and Iowa. U.S. Fish & Uildl. Serv., Washington, D.C. FWS/OBS-77/25. AI F
NE9-13
Mahoney, D.L. 1977. Species richness and diversity of aquatic vascular plants in Nebraska with special
reference to water quality parameters. M.S. Thesis, Univ. of Nebraska-Lincoln. 38 pp.
NE14.15
Ducey, James E. 1987. Biological features of saline wetlands in Lancaster County, Nebraska. Trans. Nebraska
Acad. Sci. 15:5-14.
NE16
Golden, D.R. 1987. An ichthyological survey of Weeping Water Creek, Nebraska. Trans. Nebraska Acad. Sci.
15:15-22.
NE17
Maret, T.R. 1988. A water-quality assessment using aquatic macro-invertebrates from streams of the Long Pine
Creek watershed in Brown County, Nebraska. Trans. Nebraska Acad. Sci. 69-84 pp.
NEBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
Ithaca, NY. B
NEBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
NEBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, MA. B
NEBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
NECBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
Not Mapped
Currier, P.J. 1982. The floodplain vegetation of the Platte River: phytosociology, forest development, and
seedling establishment. Ph.D. Diss., Iowa St. Univ., Ames, IA. 341 pp.
321
-------
Inland Wetlands Having Biological
Community Measurements
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE + or - 10m,
6 Research Study Site
| Migratory Shorebird Survey CBSB) stte
Q Breeding Bird Census (BBC) site that includes wetland
O Annual Christmas Btrd Count area CIS-mile diameter)
~t~ Breading Bird Survey Star ting points for 25 mi transects
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONS) NOT PLOTTED
+ State/Federal waterfowl survey
New Hampshire
Th i « nap doe* NOT por tray ALL wet I and samp I i no, site* Site* are referenced by code number to the accompany > ng
Emphas is is on 3it«« where comttivjn i ly-1evel da to were state b i b I i ogr aphy
coI Iected See chapter I for incIu»ion crit«ria
USEPA En
nt«l R«l««rch Lsbor • t or r . C«rv«lllt. Oregon
Data Compilation Paul Adarnus and Robin R«nt*ria Cartooraphy J«ff Irish
322
-------
NEW HAMPSHIRE
Happed
NH1
Moeller, Robert E. 1975. Hydrophyte biomass and community structure in a small, oligotrophic New Hampshire
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NH2
Fahey, T.J. et al. In Process. Long Term Environmental Research Wetland Site: Mirror Lake.
NH3
Glime, J.M. and R.M. demons. 1972. Species diversity of stream insects on Fontinalis spp. compared to
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NHBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
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NHBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
NHBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
NHCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
Not Mapped
Nevers H.P. 1968. Waterfowl utilization of beaver impoundments in southeastern New Hampshire. Masters
Thesis, Univ. New Hampshire. 87 pp.
323
-------
Inland Wetlands Having Biologica
Community Measurements
New Jersey
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE - or - 10mi
i) Research Study Site
| Migratory Shorebird Survey CBSB) site
Q Breeding Bird Census (BBC) site that includes wetland
O Annual Christmas Bird Count area CIS-mile diameter)
_!_„_„ , Th ( * mop does NOT portray ALL w«tI and sampling site*
+ Breeding Bird Survey Starting points for 25mi transects
AND points where transects enter new county Most cover Emphasis • * on site* where cowmuni ty-1 eve 1 data wer*
mainly non-wet I and nab ttat
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONS) NOT PLOTTED
* Stale/Federal waterfowl survey
coiUcted S*« chapter I for mclu.ion cr.t*r.a
Site* are referenced by code number to t he accompany ing
*tole bibl'ogrophy
USEPA Environmental R«»*arch Laboratory, Cor vaI I f *•
Data Coopi lotion Pau1 Adomu* and Pobin Ren t er i o Cartography Jeff Jrish
324
-------
NEW JERSEY
Happed
NJ1
Garie, H.L. and A. Mclntosh. 1986. Distribution of benthic macroinvertebrates in a stream exposed to urban
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NJ3
Ehrenfeld, J.G. and M. Gulick. 1981. Structure and dynamics of hardwood swamps in the New Jersey Pine Barrens:
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NJ3-6
Morgan, M.D. and K.R. Philipp. 1986. The effect of agricultural and residential development on aquatic
macrophytes in the Mew Jersey Pine Barrens. Biol. Conserv. 35:143-158, P I
NJ7-13
Ehrenfeld, J.G. 1983. The effects of changes in land use on swamps of the New Jersey Pine Barrens. Biol.
Cons. 25:353-375. P R
NJ16-19
Ehrenfeld. J.G. 1986. Wetlands of the New Jersey Pine Barrens: The role of species composition in community
function. Amer. Midi. Nat. 115(2):301-313. P
NJ19
Cole, C.A. 1988.
NJ20
Buchholz, K. 1981. Effects of minor drainage on woody species distributions in a successions I floodplain
forest. Can. J. For. Res. 11:671-676. PW 1
NJ21
Jervis, R.A. 1969. Primary production in the freshwater marsh ecosystem of Troy Meadows, New Jersey. Bull.
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NJ22
Scott, D. and L. Bush. 1969. A study of a pond in the Great Swamp. Drew Univ., Madison, NJ.
NJ22
Gatter, R. 1986. A survey of the benthic macroinvertebrates of the Great Swamp National Wildlife Refuge and
its immediate environs Morris County, New Jersey. M.S. Thesis, Graduate Program in Ecology. Rutgers Univ.,
New Brunswick, NJ. AI
NJ24
Kaminsky, M., P. Scelsi, C. Kanakis, and D. Fanz. 1986. Route 130, Section 9F Rancocas Creek bridge: site
III Wetland replacement. N.J. Dept. of Trans., Bureau of Environ. Analysis, Trenton, NJ. P
NJ25
Scelsi, P. nd. Small mammal and bird utilization of New Jersey highway interchanges containing wetland
habitat. NJ Dept. of Transportation, Trenton. MA B
NJ25
Schneider, J.P. and J.G. Ehrenfeld. 1987. Suburban development and cedar swamps: Effects on water quality,
water quantity, and plant community composition, pp. 271-288. In: A.D. Laderman (ed.). Atlantic White Cedar
Wetlands, Westview Press, Inc. PM
NJ26-27
Karlin, E.F. 1985. The vegetation of the low-shrub bogs of northern New Jersey and adjacent New York:
ecosystems at their southern limit. Bull. Torrey Bot. Club 112(4):436-444.
NJ26-28
Andrus, R. E. 1986. Sphagnum vegetation of the low shrub bogs of northern New Jersey and adjacent New York.
Bull. Torrey Bot. Club 113(3):281-287.
325
-------
MEW JERSEY (continued)
NJ29-34
Hastings, R.U. 1984. The fishes of the Mullica River, a naturally acid water system of the New Jersey Pine
Barrens. Bull. New Jersey Acad. Sci. 29(1):9-23. F
NJ35
Brush, T. 1988. Foliage arthropods of the New Jersey Pine Barrens: Seasonal variation in abundance in
different plant taxa. Bull. New Jersey Acad. Sci. 33(1):1-6.
NJ36
O'Herron, J.C., 11 and R.G. Arndt. 1987. Fish Studies in the Manumuskin River Drainage Basin - and Portions
of the Maurice River and Manantico Creek, Maurice River Township, Cumberland County, New Jersey. Herptological
Associates, Inc., Environmental Consultants. HA File No. 87.01-B.
NJ36
Sutton.C.C., R. Barber, and J. Dowdell. 1987. An Inventory and Habitat Assessment of the Birds of the
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NJ36
Zappalorti, R.T. and R.D. Barber. 1987. Mammalogical and Herpetological Studies in the Manumuskin River
Drainage Basin in Cumberland and Atlantic Counties, New Jersey between 1986 and 1987. Herptological Associates,
Inc., Environmental Consultants. HA File No. 87.01-C.
NJBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
Ithaca, NY. B
NJBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
NJBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, HA. B
NJBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
NJCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
Not Mapped
Atkin, D. 1980. The age structures and origins of trees in a New Jersey floodplain forest. Ph.D. Diss.,
Princeton Univ., Princeton, NJ. 119 pp.
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of cedar swamps. Center for Coastal and Environ. Studies, Rutgers Univ., New Brunswick, NJ. P I
Frye, R.J., II and J.A. Quinn. 1979. Forest development in relation to topography and soils on a floodplain
of the Raritan River, New Jersey. Bull. Torrey Bot. Club 106:334-345. PW
Lomax J.L. 1982. Wildlife use of mineral industry sites in the coastal plains of New Jersey, pp. 115-121
In: W.D. Svedarsky and R.D. Crawford (eds.). Wildlife Values of Gravel Pits. Proc. Symp., Minnesota Agric.
Exp. Stn. Misc. Publ. 17.
Lynn, L.M. and E.F. Karlin. 1985. The vegetation of the low-shrub bogs of Northern NJ and adjacent NY:
Ecosystems at their southern limit. Bull, of the Torrey Bot. Club 112:436-444. P
Schneider, J.P. 1988. The effects of suburban development on the hydrology, water quality and community
structure of Chamaecyparis thyoides wetlands in the New Jersey Pinelands. Ph.D. Diss., Rutgers Univ. P
326
-------
NEW JERSEY (continued)
Smith, R.F. 1960. An ecological study of an acid pond in the New Jersey coastal plain. Ph.D. Diss., Rutgers
Univ.. New Brunswick, NJ. 197 pp.
Sutton, C.C., Jr. and R.T. Zappalorti. 1988. Wintering Raptors and Waterfowl along the Maurice River on the
Delaware Bayshore, Cumberland County, New Jersey. Herpetological Associates, Inc., HA File No. 87.44. H
327
-------
Inland Wetlands Having Biological
Community Measurements
+ j
-t-
.a
-f--"'--A
X^A. J
n
New Mexico
This mop do*« NOT periray ALL w«Uand aaffpling si
Enphasi* iv on • i !•• wK«r« commvjn i ly~ I sv*! data w
coll«ct«d S«« chapter t for inclusion erit«ria
S,t.« or. r.f.r.nc.d by cod*
•tat* bibliography
.b.r to th*
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE « or - I8»,
• R*>*arch Study S t.
I Migratory Shor.bird Surv*y «.t«
Q Br..d:ng Bird C.nsus CB3C) sit* that mclud*s u.I land
Q Annual Christmas Bird Count arsa (!5-mi!» diam*t*r)
Most cover mainly non-wetland habitat
4" Breeding Bird Surv*y Starting points for 25»i trans*cts
AND points whsr* transects enter new county Host cover
mainly non-uetland habitat
SITE LOCATED IN COUNTY. SPECIFIC LOCATIONS) NOT PLOTTED
* State/Federal waterfowl survey
USEPA Env I
Re»««rch L«bor»t«rx> Cor>elll», Oregon
Data Compilation Paul Adamus and Robin R*nt*ria
Cartography Jeff Irish
328
-------
NEW MEXICO
Mapped
NM1, 2
Dick-Peddie, W.A., J.V. Hardesty, E. Muldavin, and B. Sallach. 1987. SoiI-vegetalion correlations on the
riparian zones of the Gila and San Francisco Rivers in New Mexico. Biol. Rep. 87(9). U.S. Fish & Wildl. Serv.,
Washington, D.C. 30 pp. P
NM3
Hink, V.C. and R.D. Ohmart. 1984. Middle Rio Grande Biological Survey. U.S. Army Corps, of Engr., Contract
No. DAC W47-81-C-0015. B
NM4, LTR
Schlesinger, W., et at. In Process. Long Term Environmental Research Wetland Site: Jornada LTER Site. Dept.
of Bot., Duke Univ., Durham, NC. P
NMBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
NMBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, MA. B
NMBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
NMCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
Not Happed
Szaro, R.C. and J.N. Rinne. 1988. Ecosystem approach to managment of southwestern riparian communities. Tran.
N. Amer. Wildl. Nat. Resour. Conf. 53:502-511.
Szaro, R.C. 1989. Riparian forest and scubland community types of Arizona and New Mexico. Desert Plants 9:70-
138.
329
-------
Inland Wetlands Having BioIogica
Community Measurements
.81*
Nevada
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE » or -
0 Research Study Site
ff Migratory Shorebird Survey CBSB) site
Q Breeding Bird Census (BBC) site that includes wetland
O Annual Christmas Bird Count area (l5-i».le diameter)
Most cover mainly non-wetland habitat
,
T Breeding Bird Suryey Starting points for 25mi transects
Thi« map doe» NOT portray ALU uetland »anp I i no
AND
mainly non-wettand Habitat
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONCS) NOT PLOTTED
* Slate/Federal waterfowl survey
collected See chapter I for inclusion criteria
Sitee are referenced by code number to the accompanying
state bibllography
USEPA Env I r on««nt« I R«»««rch L»bor»tor». Corollls. Ortgon
Data Compilation Paul Adamue and Robin Renteria
Cartography Jeff Irish
330
-------
NEVADA
Happed
NV1
Platts, W.S. 1985. The effects of large storm events on basin-range riparian stream habitats.
NV2
Nachlinger, J.L. 1988. SoiI-Vegetation Correlations in riparian and emergent wetlands, Lyon County, Nevada.
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NV7
Bouffard, S.H. 1983. Canvasback and redhead productivity at Ruby Lake National Wildlife Refuge. Cal-Neva
Wildlife Trans. 84-90 pp.
NV7
Port, M.A. and L.K. Ports. 1988. Associations of small mammals occurring in a pluvial lake basin, Ruby Lake,
Nevada. U.S. Fish & Wildl. Serv., Ruby Valley Natl. Wildl. Refuge, 12 pp.
NV7
Young, J.A., R.A. Evans, B.A. Roundy, and J.A. Brown. 1986. Dynamic landforms and plant communities in a
pluvial desert rodent communities: Habitats and environmental complexity. Ecol. 50:558-572.
NV7
Young, J.A., R.A. Evans, B.A. Roundy, and J.A. Brown. 1986. Dynamic landforms and plant communities in a
pluvial lake basin. Great Basin Nat. 46(1):1-32.
NV8-11
Jensen, S., R. Ryel and W.S. Platts. 1989. Classification of riverine/riparian habitat and assessment of
nonpoint source impacts north fork Humboldt River Nevada. USDA Forest Service, Intermountain Research Station,
Fisheries Unit, Boise, ID.
NVBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
NVBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
NVCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
Not Happed
Finger, S.E., S.J. Olson, and A.C. Livingstone. 1989. On-site toxicity of irrigation drainwater from
Stillwater National Wildlife Refuge to aquatic organisms. 1988 Progress Report. Nat. Fish. Contam. Res.
Center. U.S. Fish & Wildl. Serv., Columbia, MO.
Ingersoll, C.G., F.J. Dwyey, M.K. Nelson, S.A. Burch, and D. R. Buckler. 1988. Whole effluent toxicity of
agricultural irrigation drain water entering Stillwater National Wildlife Refuge, NV: Acute toxicity studies
with fish and aquatic invertebrates.Fish and Invert. Toxicol. Sect., Nat. Fish. Contam. Res. Center. U.S. Fish
& Wildl. Serv., Columbia, MO.
Hedin, D.E. and W.P. Clary. 1989. Small Mammal Populations in a Grazed and Ungrazed Riparian Habitat in
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Tiehm, A. 1978. A taxonomic and floristic study of selected aquatic plants of Nevada. M.S. Thesis, Univ.
Nevada, Reno, NV. 169 pp.
331
-------
o
u
O>
0
~0
o
"D
C
o
c
n
.
332
-------
NEW YORK
Mapped
NY1
Menzie, C.A. 1980. The chironomid (Insecta: Diptera) and other fauna of a MyriophyUum spicatum L. plant bed
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NY2
Snow, P.O., R.P. Mason, C.J. George, and P.L. Tobiessen. 1978. Monitoring of hydraulic dredging for lake
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NY2-4
Malecki, R.A. and J.D. Sullivan. 1987. Assessment of an agricultural drainage improvement program in New York
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NY2-4
Malecki, R.A. and J. Eckler. 1980. An Environmental Assessment of the Wayne County Water Management Program
in the Melvin Brook Watershed. Agric. Water Manage. Program, Series No. 80-1. Cornell Univ., Ithaca, NY. 30
PP-
NY7
Gruendling, G.K. and D.J. Bogucki. 1978. Assessment of the physical and biological characteristics of the
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NY8
Mikol, G.F. 1982. Effects of mechanical control of aquatic vegetation on biomass, regrowth rates, and juvenile
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NY9
Cowardin, L.M. 1969. Use of flooded timber by waterfowl at the Montezuma National Wildlife Refuge. J. Wildl.
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NY9
Haramis, G.M. 1975. Wood Duck ecology and management within the green-tree impoundments of Montezuma National
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NY13
Paratley, R.D. and T.J. Fahey. 1986. Vegetation--environment relations in a conifer swamp in central New
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NY14
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NY15
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NY16
Pierce, G.J. nd. The influence of flood frequency on wetlands of the Allegheny River flood plain in Cattaraugus
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NY18-19
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333
-------
NEW YORK (continued)
NY29
Pierce, G.J. 1983. Annual Report for 1982 Southern Tier Expressway Allegheny River Valley Wetland Development
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B F
NY30-32
Sheldon R.B. and C.W. Boylen. 1975. Factors affecting the contribution by epiphytic algae to the primary
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NY33
Eckblad, J.W. 1973. Population studies of three aquatic gastropods in an intermittent backwater.Hydrobiol.
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NY33
Huenneke, L.F. 1982. Wetland forests of Tompkins County, New York. Bull. Torrey Bot. Club 109(1):51-63.
PW
NY33
Oglesby, R.T., A. Vogel, J.H. Perverly, and R. Johnson. 1976. Changes in submerged plants at the south end
of Cayuga Lake following tropical storm Agnes. Hydrobiol. 48(3):251-255. PM
NY34
Lynn, L.M. 1984. The vegetation of Little Cedar Bog, southeastern New York. Bull, of the Torrey Bot. Club
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NY36-42
Andrus, Richard E. 1986. Sphagnum vegetation of the low shrub bogs of northern New Jersey and adjacent New
York. Bull. Torrey Bot. Club 113(3):281-287.
NY39-42
Karlin, E.F. 1985. The vegetation of the low-shrub bogs of northern New Jersey and adjacent New York:
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NY46
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NY48
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NYBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
Ithaca, NY. B
NYBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
334
-------
NEW YORK (continued)
NYBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, MA. B
NYBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
NYCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
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335
-------
NEW YORK (continued)
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336
-------
NEW YORK (continued)
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337
-------
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-------
OHIO
Happed
OH1-3
Van Hassel, J.H., R.J. Reash, and H.W. Brown. 1988. Distribution of upper and middle Ohio River fishes,
1973-1975: I. Associations with water quality and ecological variables. J. Freshw. Ecol. 4(4):441-458. F
OH 1-3
Reash, R.J. and J.H. Van Hassel. 1988. Distribution of upper and middle Ohio River fishes, 1973-1985: II.
Influence of zoo-geographic and physiochemical tolerance factors. J. Freshw. Ecol. 4(4):459-476. F
OH4-16
Garono, R.J. and D.B. Maclean. 1988. Caddisflies (Trichoptera) of Ohio wetlands as indicated by
light-trapping. Ohio J. Sci. 88(4):143-151. AI
OH17
Anderson, J.H. 1950. Some aquatic vegetation changes following fish removal. J. Wildl. Manage. 14(2)-.206-209.
PM
OH18,37,38
Tramer, E.J. and P.M. Rogers. 1973. Diversity and longitudinal zonation in fish populations of two streams
entering a metropolitan area. Amer. Midi. Nat. 90(2):366-374. F I
OH19
Loveland, D.G. and I.A. Ungar. 1983. The effect of nitrogen fertilization on the production of halophytes in
an inland salt marsh. Amer. Midi. Nat. 109(2):346-354. I PE
OH20
Farney, R.A. 1982. Vegetation changes in a Lake Erie marsh (Winous Point, Ottawa County, Ohio) during high
water years. Ohio Acad. Sci. 82:103-107. PE
OH21
Meeks, R.L. 1969. The effect of drawdown date on Wetland plant succession. J. Wildl. Manage. 33(4):817-821.
P
OH22
Stuckey, R.L. 1971. Changes of vascular aquatic flowering plants during 70 years in Put-In-Bay Harbor, Lake
Erie, Ohio. Ohio J. Sci. 71(6):321-342. PM TS
OH23-36
Wentz, W.A. and R.L. Stuckey. 1971. The changing distribution of the genus Najas (Najadaceae) in Ohio. Ohio
J. Sci. 71(5):292-301. PM TS
OH37
Dames and Moore, Inc. 1985. Floodplain-Wetlands Quantitative Studies, Terrestrial ecological sampling program
results 1985, Zinner Conversion Project for American Electric Power Service Corp.
OH37.38
Tramer, E.J., and P.M. Rogers. 1973. Diversity and longitudinal zonation in fish populations of two streams
entering a metropolitan area. Amer. Midi. Nat. 90(2):366-374. F I
OH38
Reeder, B.C. and W.J. Mitsch. 1989. Seasonal patterns of planktonic and macrophyte productivity of a
freshwater coastal wetland, pp. 49-55 IN: Wetlands of Ohio's Coastal Lake Erie: A Hierarchy of Systems.
Environ Sci. Prog., School of Nat. Resources, Ohio State Univ., Columbus, OH.
OH39
Stuckey, R.L. and W.A. Wentz. 1969. Effect of industrial pollution on the aquatic and shore angiosperm flora
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OHBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
Ithaca, NY. B
339
-------
OHIO (continued)
OHBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data, Office of Migratory Bird
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OHBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
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OHBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
OHCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
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Not Mapped
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Hardin, E.D. 1982. Patterns in floodplain herbaceous vegetation and some aspects of the population biology
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72 PP.
340
-------
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342
-------
OKLAHOMA
Happed
OK1-5
Felley, J.D. and L.G. Hill. 1983. Multivariate assessment of environmental preferences of cyprinid fish of
the Illinois River, Oklahoma. Amer. Midi. Nat. 109(2):209-221. F
OK6-10
Heitmeyer, M.E. and P.A. Vohs, Jr. 1984. Distribution and habitat use of waterfowl wintering in Oklahoma. J.
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OK11
Erickson, N.E. and D.M. Leslie, Jr. 1988a. Impacts of the rule curve change on shoreline vegetation and
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OK11
Erickson, N.E. and D.M. Leslie, Jr. 1988b. Shoreline vegetation and general wildlife values around Grand
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OKBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
Ithaca, NY. B
OKBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
OKBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
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OKBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
OKCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
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Mot Mapped
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343
-------
Inland Wetlands Having Biological
Community Measurements
Or egon
.EPA
• 02.23
*.EPA
• 21
.25
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE » or - 10m,
• Research Study S . l«
| Migratory Shor.bird Survey CBS8) s.t.
Q Breeding Bird Census (BBC) site that includes wetland
O Annual Christmas Bird Count area (15-mile diameter)
Moel cover mainly non-uetland habitat
Breeding Bird Sur
Starting points for 25m i transects
SITE LOCATED IN COUNTf. SPECIFIC LOCATIONCS) NOT PLOTTED
* State/Federal waterfowl survey
Tft!« Map daes NOT portray ALL wetland sampling sites
EnphaAifl te I Rtstlrch Laboratory CorxllK. Qr«f«n
Data Compilation Paul Adamus and Robin Renter 10 Cartography Jeff Irish
344
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OREGON
Happed
OR1
Franklin, K.T. and R. Frenkel. 1987. Monitoring a Wetland Treatment System at Cannon Beach. Oregon. U.S. EPA
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OR2-4
Perdue, E.M., C.R. Lytle, M.S. Sweet, and J.W. Sweet. 1981. The Chemical and Biological Impact of Ktamath
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Inst., Oregon State Univ. WRRI-71, 199 pp. I A
OR5
Comely, J.E. 1982. Waterfowl production at Malheur National Wildlife Refuge, 1942-1980. 47th N.A. Wildl.
Conf. 47:559-571. B
OR6-7
Sanville, W.D., H.P. Eilers, T.R. Boss, and T.G. Pfleeger. 1986. Environmental gradients in northwest
freshwater wetlands. Environ. Manage. 10(1):125-134. P
OR8-13
Kreis, R.D. and W.C. Johnson. 1968. The response of macrobenthos to irrigation return water. J. Water Poll.
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OR10
Geiger, N.S. 1983. Winter drawdown for the control of Eurasian water milfoil in an Oregon oxbow lake, (Blue
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OR14
Bull, E.L. and J.M. Skovlin. 1982. Relationships between avifauna and streamside vegetation. 47th N. Amer.
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OR15
Cross, S.P. 1985. Responses of small mammals to forest riparian perturbations, pp. 269-275. In: R.R.
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OR16
Fishman Environmental Services, and Steffen, Robertson, and Kirsten (Colorado), Inc. 1988. Technical Report
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Nat. For. F B P
OR17
Fishman Environmental Services, Ogden Beeman and Associates, Inc., Shannon and Wilson, Inc., Scientific
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OR18
Fishman Environmental Services. 1989. Columbia Slough water quality management plan aquatic biology final
report: Benthic invertebrates. Fish and Bioaccumulation. City of Portland, Oregon, Bureau of Environ. Serv.
P B F
OR19
Fishman Environmental Services. 1989. Force Lake fisheries evaluation. Western Columbia Wetlands Conservancy.
OR20
Fishman Environmental Services. 1985. Wetland Assessment. Columbia Steel Casting Co., Inc. P B F
OR21
Sharp, L. 1987. Birds of the eastside development area and water features at office park sites. Port of
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345
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OREGON (continued)
OR22-25
Lippert, B. E. and D. L. Jameson. 1964. Plant succession in temporary ponds of the Willamette Valley, Oregon.
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OR31
U.S. Army Corps of Engineers. 1986. Malheur Lake Flood Damage Reduction Study, Harney County, Oregon. Draft
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ORBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
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ORBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
ORBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
ORCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
OREPA1-
U.S. Environmental Protection Agency, in press. Comparison of constructed and reference wetlands.
ORLTR
Swanson, F.J. et al. In Process. Long Term Environmental Research Wetland Site: H.J. Andrews Experimental
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Not Mapped
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346
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OREGON (continued)
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Seyer, S.C. 1979. Vegetation ecology of a montane mire. Crater Lake National Park, Oregon. M.S. Thesis,
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Taylor. A.H. and R.E. Frenkel. 1979. Ecological inventory of Joe Ney Slough Marsh mitigation site. Oregon
Dept. Land Conservation and Development, Salem, OR.
Young, R.P. 1986. Fire ecology and management in plant communities of Malheur National Wildlife Refuge,
southeastern Oregon. Ph.D. Diss., Oregon St. Univ., Corvallis. 183 pp.
347
-------
Inland Wetlands Having Biologica
Community Measurements
PennsyI van i a
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE * or - 10m,
0 Research Study Site
H Migratory Shorebird Survey CBSB) 9'te
Q Br«»dmg Bird Census CB8O ait* that \ oc I udes wetland
Q Annual CKri*tmaa Bird Count area (15-mile diameter)
Most cover mainly non-wetland habitat
+ Breeding Bird Survey Starting points for 25mi transects
AND points wher • transects *rtt*r o«w county Ho»t cover
mainly non-wetland habitat
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONS) NOT PLOTTED
* State/Fedaral waterfowl survey
Thi• map does NOT por tray ALL wet I and *ampI ing s't«s
col!*cl«d S»» chapter 1 for inc f usion crit er i a
Site* ar« referenced by code nunb*r to the accompanying
state b i b t i a Cartography J«ff Iri»h
348
-------
PENNSYLVANIA
Happed
PA1
Brenner, F.J., W. Kantour, B. Weston, G. Valeric, and K.R. Grayburn. 1986. Impact of flood control reservoirs
and pollution influx on the Sandy Creek Watershed, Mercer County, Pennsylvania, USA. Environ. Manage.
10(2):Z41-253. I
PA3-11
Brooks, R.P., D.E. Arnold, E.D. Bellis, C.S. Keener, and M.J. Croonquist. 1989. A methodology for biological
monitoring of cumulative impacts on wetland, stream, and riparian components of watersheds. In: Assoc. of
Wetland Managers, Inc., Berne, NY. T B MA
PA3-8
Brooks, R.P., J.B. Hill, F.J. Brenner, and S. Capets. 1985. Wildlife use of wetlands on coal surface mines
in Western Pennsylvania, pp. 337-352 In: R.P. Brooks, D.E. Samuel, and J.B. Hill (eds.). Wetlands and Water
Management on Mined Lands. Penn. St. Univ., University Park, PA. MA B H P
PA3-8
Hill, J.B. 1986. Wildlife use of wetlands on coal surface mines in Western Pennsylvania. M.S. Thesis,
Pennsylvania State Univ., School of Forest Resources, University Park. H MA P B
PA9-11
Brooks, R.P., D.E. Arnold, and E.D. Bellis. 1987. Wildlife and plant communities of selected wetlands in the
Pocono region of Pennsylvania. U.S. Fish & Wildl. Serv. NWRC Rep. 87-02. 41 pp. MA B R
PA12
Seelbach, P.W. and W.F. McDiffett. 1983. Distribution and abundance of zooplankton in an alkaline freshwater
marsh in Northumberland County, Pennsylvania. Int. Revue ges. Hydrobiol. 68(3):379-395. AI
PA13
Bott, T.L. 1975. Bacterial growth rates and temperature optima in a stream with a fluctuating thermal regime.
Limnol. & Oceanogr. 20(2):191-197. MI
PA14-16
Bradt, P.T. and M.B. Berg. 1987. Macrozoobenthos of three Pennsylvania lakes: Responses to acidification.
Hydrobiol. 150:63-74. AI 1
PA14-16
Bradt, P.T., J.L. Dudley, M.B. Berg, and D.S. Barrasso. 1986. Biology and chemistry of three Pennsylvania
lakes: Responses to acid precipitation. Water, Air and Soil Poll. 30:505-513. AI I
PA20
Hepp, J.P. 1987. An ecological survey of four newly created surface-mine wetlands in Central Pennsylvania.
M.S. Thesis, Pennsylvania State Univ., School of Forest Resource, University Park. H MA B AI P
PABBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
Ithaca, NY. B
PABBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
PABSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, MA. B
PABW1 -
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
PACBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
349
-------
PENNSYLVANIA (continued)
Not Mapped
Boulay, E.A. 1978. The effects of heavy metals on the abundance of aquatic insects and terrestrial plants.
Ph.D. Diss., Penn. St. Univ., University Park, PA. 138 pp.
Brenner, F.J. and J. Kelly. 1981. Characteristics of bird communities on surface mine lands in Pennsylvania.
Environ. Manage. 5:441-449.
Brenner, F.J., R.B. Kelly, and J. Kelly. 1982. Mammalian community characteristics on surface mine lands in
Pennsylvania. Environ. Manage. 6:241-149.
Cole, R.A. 1969. The effect of macrophytes on the abundance and diversity of macroinvertebrates in an enriched
stream. Ph.D. Diss., Penn. St. Univ., University Park, PA. 76 pp.
Denoncourt, R.F. and J.W. Stambaugh. 1974. An Ichthyofaunal Survey and Discussion of Fish Species Diversity
as an Indicator of Water Quality, Codorus Creek Drainage, York County, Pennsylvania. Proc. Penn. Acad. Sci.
48:71-78. F
Dinsmore, B.H. 1958. Ecological studies of twelve strip mine ponds in Clarion County, Pennsylvania. Ph.D.
Diss., Univ. Pittsburgh, Pittsburgh, PA. 118 pp.
Gehris, C.U. 1964. Pollen analysis of the Cranberry Bog Preserve, Tannersvilie, Monroe County, Pennsylvania.
Ph.D. Diss., Penn. St. Univ., University Park, PA. 82 pp.
Graffius, J.H. 1958. An ecological comparison of two bog areas with specific references to the algal flora.
M.S. Thesis, Univ. of Pittsburgh, Pittsburgh, PA. A
Greenwald, C.M. 1981. Prediction of songbird responses to habitat alteration resulting from wastewater
irrigation. School For. Resour., Penn. St. Univ., University Park, PA. 81 pp. B
Halma, J.R. 1974. An ecological investigation of the breeding avian populations in the TannersviUe Bog area,
Monroe County, Pennsylvania. Ph.D. Diss., Lehigh Univ., Bethlehem, PA. 164 pp.
Lewis, S.J. 1977. Avian communities and habitats on natural and wastewater irrigated vegetation. M.S. Thesis
Pennsylvania State Univ., University Park, PA. B P
Moore, J.R. 1965. Productivity and standing crop of vascular hydrophytes. Ph.D. Diss., Univ. Pittsburgh,
Pittsburgh, PA. 190 pp.
Mujamdar, S.K., R.P. Brooks, F.J. Brenner, and R.U. Tiner (eds.). Wetlands Ecology and Conservation: Emphasis
in Pennsylvania. Penn. Aca. Sci., Philadelphia.
Snider, J.R. and G.W. Wood. 1975. The effects of waste water irrigation on the activities and movements of
songbirds, pp. 20-49 In: Wood, G.W. et al (eds.). , Faunal Response to Spray Irrigation of Chlorinated Sewage
Effluents. Pub. 87 Inst. Res. Land and Water Resour. Res., Pennsylvania State Univ., Philadelphia, PA. B I
Sopper, W.E. and L.T. Kardos. 1973. Vegetation responses to irrigation with treated municipal wastewater.
pp. 271-294. In: Recycling Treated Municipal Uastewater and Sludge Through Forest and Cropland, 271-294 pp.,
Pennsylvania State Univ. Press, Univ. Park, PA. P
Van Dersal, W.R. 1933. An ecological study of Pymatuning Swamp. Ph.D. Diss., Univ. Pittsburgh, Pittsburgh,
PA.
Walker, P.C. 1958. The forest sequence of the Hartstown Bog area. Ph.D. Diss., Univ. of Pittsburgh,
Pittsburgh, PA. 96 pp.
350
-------
Inland Wetlands Having Biological
Community Measurements
Rhode Is I and
Thi• map do*« NOT por tray ALL w»tI and samp I ing *it*«
E»pha»i« i• on cite* where community-I*v*I data were
co1l*ct*d S*« chapter 1 for ioclu»*on criteria
Sit*« or* r«f«r«nc*d by cod* numb«r to th« accomponying
•tat* bibIiogrophy
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE + or - 10m,
9 R«3*arch Study Site
fl Migratory SKorebird Survey ($SB) site
Q Breeding Bird Census (BBC) site that includes wetland
O Annual Christmas B
-------
RHODE ISLAND
Happed
RI1
Sheath, R.G., J.M. Burkholder, J.A. Hambrook, A.M. Hogeland, E. Hoy, M.E. Kane, M.O. Morison, A.D. Steinman,
and K.L. Van Alstyne. 1986. Characteristics of softwater streams in Rhode Island. III. Distribution of
macrophytic vegetation in a smalt drain. Hydrobiologia 140:183-191. PM R
RI6-22
Theses and Dissertations, Dept. Forest & Wildlife Management, Univ. Rhode Island, Kingston, RI.
RIBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
Ithaca, NY. B
RIBBS1-
U.S. Fish & Uildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
RIBU1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
R1CBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
Not Happed
Doty, T.L. 1978. A study of larval amphibian population dynamics in a Rhode Island vernal pond. Ph.D. Diss.,
Univ. Rhode Island, Kingston, RI. 146 pp.
Lowry, D.J. 1984. Water regimes and vegetation of Rhode Island forested wetlands. M.S. Thesis, Univ. of
Rhode Island.
353
-------
Inland Wetlands Having Biological
Community Measurements
South Caro t ina
This map do*» NOT portray ALL w*tland sampling «it*»
Eiipha* i « i * on *i t*» uh*r• eommun iky~)*v*l data were
coll*ct*d See chapter 1 for inclusion criteria
Sit** or* r*f*r*nc*d by cod* numb*r to th* accompanying
•tat* bibllography
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE + or - 10m,
9 Research Study Sit*
| Migratory Shorebird Survey (BSB) Site
Q Bre.dmg Bird C«nsu* CB8C^ si'* that includ** «*tland
O Annual Christmas B.rd Count ar*a CIS-mil* dv i ron«*n id I R«»t»rch Laboratory. Cor»»lll». Ortgon
Data Compilation Paul Adanu* and Robin Renteria Cartography Jeff Irish
354
-------
SOUTH CAROLINA
Mapped
SC1
Scott, M.L., R.R. Sharitz, and L.C. Lee. 1985. Disturbance in a cypress-tupelo wetland: An interaction
between thermal loading and hydrology. Wetlands 5:53-68. I PW
SC2
McLeod, K.W., L.A. Donovan, N.J. Stumpff, and K.C. Sherrod. 1986. Biomass, photosynthesis, and water use
efficiency of woody swamp species subjected to flooding and elevated water temperature. Tree Physiol. 2:341-
346.
SC2
Sharitz, R.R. and L.C. Lee. 1985. Limits on regeneration processes in southeastern riverine wetlands, pp.
139-143 In: Riparian Ecosystems and Their Management: Reconciling Conflicting Uses. Gen. Tech. Rep. RM-120,
USDA Forest Serv., Fort Collins, CO. PW
SC2
Sharitz, R.R., J.E. Irwin, and E.J. Christy. 1974. Vegetation of swamps receiving reactor effluents. Oikos
25:7-13. P I
SC2
Congdon, J.D., J.L. Greene, and J.W. Gibbons. 1986. Biomass of freshwater turtles: A geographic comparison.
Amer. Midi. Nat. 115(1):165-173. H
SC2
Gibbons, J.W. and D.H. Bennett. 1974. Determination of anuran terrestrial activity patterns by a drift fence
method. Copeia 1:236-243. H
SC3
Sheldon R.B. and C.W. Boylen. 1975. Factors affecting the contribution by epiphytic algae to the primary
productivity of an oligotrophic freshwater lake. Appl. Microbiol. 30(4):657-667. A
SC4
Christensen, E.J., J.R. Jensen, E.W. Ramsey, and H.E. Mackey, Jr. 1988. Aircraft MSS data registration and
vegetation classification for wetland change detection. International J. Remote Sensing 9(1):23-38. RS
SC5
Dunn, C.P. and M.L. Scott. 1987. Response of wetland herbaceous communities to gradients of light and
substrate following disturbance by thermal pollution. Vegetatio 70:119-124. P I
SC5
Dunn, C.P. and R.R. Sharitz. 1987. Revegetation of a Taxodium-Nyssa forested wetland following complete
vegetation destruction. Vegetatio 72:151-157. PW
SC7
Smock, L.A., E. Gilinsky, and D. L. Stoneburner. 1985. Macroinvertebrate production in a southeastern United
States blackwater stream. Ecol. 66(5)-.491-503. AI
SC8
Patterson, G.G., G.K. Speiran, and B.H. Whetstone. 1985. Hydrology and its effects on distribution of
vegetation in Congaree Swamp National Momument, South Carolina. U.S. Geol. Surv. Rep. 85-4256. 31 pp. P I
SC11
Christy, E.J. and R.R. Sharitz. 1980. Characteristics of three populations of a swamp annual under different
temperature regimes. Ecol. 6:454-460. PE I
SC11
Oden, B.J. 1977. Comparative spatial and temporal variations among freshwater littoral meiofauna in a
reservoir receiving thermal effluents (Par Pond, Aiken, SC). Ph.D. Diss., Univ. South Carolina, Columbia, SC.
60 pp.
Gibbons, J.W., J.L. Greene, and J.D. Congdon. 1983. Drought-related responces of aquatic turtle populations.
J. Herpetology 17(3):242-246. H
355
-------
SOUTH CAROLINA (continued)
SC11
Fallen, M.H. 1987. Distribution of larval fish macrophyte beds and open channels in a southeastern floodplain
swamp. J. Freshw. Ecol. 4(2):191-200. F
SC12
Thorp, J.H., E.M. McEwan, M.F. Flynn, and F.R. Hauer. 1985. Invertebrate colonization of submerged wood in
a cypress-tupeIo swamp and blackwater stream. Amer. Midi. Nat. 113(1):56-68. AI
SC13
McArthur, J.V., L.G. Leff, O.A. Kovacic, and J. Jaroscak. 1986. Green leaf decomposition in coastal plain
streams. J. Freshw. Ecol. 3(4):553-559. D
SC15
Pechmann, J.H.K., D.E. Scott, J.W. Gibbons, and R.D. Semlitsch. 1988. Influence of wetland hydroperiod on
diversity and abundance of metamorphosing juvenile amphibians. Wetland Ecol. Manage. 1(1):3-11. H
SC15
Semlitsch, R.D. and J.H.K. Pechmann. 1985. Diel pattern of migratory activity for several species of
pond-breeding salamanders. Copeia 1985:86-91. H
SC16
James, W.F., R.H. Kennedy, W.E. Shain, and R.K. Myers. 1988. Leaf litter breakdown in a recently impounded
reservoir. Water Res. Bull. 24(4):831-837. D
SC17
U.S. Environmental Protection Agency. 1983. Hydrographic, Water Quality and Biological Studies of Freshwater
Canal Systems, South Carolina, Mississippi, and Florida. Environ. Protection Agency, Environ. Serv, Div.,
Athens, GA. AI
SC18-23
Woodwell, G. 1956 (unpub.). wetland vegetation data.
SC24
Schalles, J.F. andD.J. Shure. 1989. Hydrology, community structure, and productivity patterns of a dystrophic
Carolina bay wetland. Ecol. Monogr. 59(4)-.365-385.
SC24
Pechmann, J.H.K. and R.D. Semlitsch. 1986. Diel activity patterns in the breeding migrations of
winter-breeding anurans. Rainbow Bay, Barnwell County, South Carolina. Can. J. Zool. 64:1116-1120. H
SC25
Gibbons, J.W., J.L. Greene, and J.D. Congdon. 1983. Drought-related responces of aquatic turtle populations.
J. Herpetology 17(3):242-246. H
SC26
Congdon, J.D., J.L. Greene, and J.W. Gibbons. 1986. Biomass of freshwater turtles: A geographic Comparison.
Amer. Midi. Nat. 115(1):165-173. H
SC27
Bates, R.D. 1985. Biomass and primary productivity measurements of mature and early successionaI forest sites
on the Santee River floodplain. M.S. Thesis, Dept. of Environ. Health Sci., Univ. South Carolina, Columbia,
SC. PW
SC28
Harvey, R.M., J.R. Pickett, P.G. Mancusi-Ungaro, and G.G. Patterson. 1983. Aquatic Macrophyte Distribution
in Upper Lake Marion: 1983 Growing Season. Dept. of Health & Environ. Control, Columbia, SC. 61 pp. PM
SC29
Homer, M.L. and J.B. Williams. 1986. The Effects of Aquatic Macrophyte Control on Fish Populations Inhabiting
an Abandoned Rice Field in the Upper Cooper River, South Carolina. Dept. of Environ. Health Sci., Univ. of South
Carolina, Columbia, SC. 170 pp. PM F
356
-------
SOUTH CAROLINA (continued)
SC30
Mcllvaine, C.M. 1986. Seasonal abundance and diversity of zooplankton in upper Lake Marion, South Carolina,
M.S. Thesis, Dept. of Environ. Health Sci., Univ. of South Carolina, Columbia, SC. 73 pp. AI
SC31
Welch, R. and H.M. Remit lard. 1986. Aquatic Macrophyte Distributions in Lake Marion, South Carolina: June
and September, 1985. Lab. for Remote Sensing and Mapping Sci., Dept. of Geography, Univ. of Georgia, Athens,
GA. 15 pp. PM
SC31
Welch, R., S.S. Fung, and M.M. Remillard. 1985. Aquatic Macrophyte Distribution in Lake Marion, South
Carolina: 1983-1984. Lab. for Remote Sensing and Mapping Sci., Dept. of Geography, Univ. of Georgia, Athens,
GA. 18 pp. PM
SC31
Welch, R., S.S. Fung, and M.M. Remillard. 1986. Changes in the Distribution of Aquatic Macrophytes: Lake
Marion, South Carolina: 1972-1984.Lab. for Remote Sensing and Mapping Sci., Dept. of Geography, Univ. of
Georgia, Athens, GA. 16 pp. PM
SC32
CH2M Hill. Unpub. Central Slough Pilot Study. Grand Strand Water & Sewer Auth., Central Wastewater Treatment
Plant Wetlands Discharge, Charleston, SC. P
SCBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
Ithaca, NY. B
SCBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
SCBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, MA. B
SCBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
SCCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
Not Mapped
Bergen, J.F. and L.M. Smith. 1989. Differential habitat use by diving ducks wintering in South Carolina. J.
Wildl. Manage. 53:1117-1126.
Christy, E.J. and R.R. Sharitz. 1980. Characteristics of three populations of a swamp annual under different
temperature regimes. Ecol.61(3):454-460. P
Duncan, R.E. 1975. Wando River Aerial Imagery and Marsh Productivity Study. South Carolina Water Resour.
Comm. Spec. Study Rep. #120 28 pp. RS P
Grey, W.F. 1973. An analysis of forest invertebrate populations of Santee-Cooper Swamp, a floodplain habitat.
M.S. Thesis, Univ. of South Carolina, Columbia. AI
Hall, R.J. 1976. A preliminary report on the herpetological survey conducted in Four-Hole Swamp, 25 Match-18
October, 1976. Nat. Audubon Soc. (unpub). H
Name I, P.B. 1989. Breeding bird populations on the Congaree Swamp National Monument, South Carolina, pp. 617-
628 In: R.R. Sharitz and J.U. Gibbons (eds.). Freshwater Wetlands and Wildlife, Proceedings of a Symposium.
CONF-8603101 (NTIS No. DE90005384). U.S. Dept. Energy, Washington, D.C.
357
-------
SOUTH CAROLINA (continued)
Hauer, F.R., N.L. Poff, and P.L. Firth. 1986. Leaf litter decomposition across broad thermal gradients in
southeastern coastal plain streams and swamps. J. Freshw. Ecol. 3:545-552.
Homer, M.L. 1988. The impact of habitat loss on freshwater fish populations. Ph.D. Diss., Univ. South
Carolina, Columbia. 184 pp.
Jones, R.H. 1981. A classification of lowland forests in the northern coastal plain of South Carolina. M.S.
thesis, Clemson Univ., Clemson, SC.
Knight, R.L., J.S. Bays, and F.R. Richardson. 1989. Floral composition, soil relations, and hydrology of a
Carolina Bay in South Carolina, pp.219-234 In: R.R. Sharitz and J.W. Gibbons (eds.). Freshwater Wetlands and
Wildlife, Proceedings of a Symposium. CONF-8603101 (NTIS No. DE90005384). U.S. Dept. Energy, Washington, D.C.
Muzika, R.M., J.B. Gladden, and J.D. Haddock. 1987. Structural and functional aspects of succession in
southeastern floodplain forests following a major disturbance. Amer. Midi. Nat. 117:1-9.
Pendleton, W.O. 1974. A synecological study of the spiders of Santee Swamp South Carolina. M.S. Thesis,
Univ. of South Carolina, Columbia, SC. 23 pp. AI
Rikard, M.W. 1988. Hydrologic and vegetative relationships of the Congaree Swamp National Monument. Ph.D.
Diss., Clemson Univ., Clemson, SC. 113 pp.
Schalles, J.F. 1979. Comparative limnology and ecosystem analysis of Carolina Bay ponds on the upper coastal
plain of South Carolina. Ph.D. Diss., Emory Univ., Atlanta, GA. 290 pp.
Smith, G.C., J.B. Gentry, D. W. Kaufman, and M.L. Smith. 1980. Factors affecting distribution and removal
rates of small mammals in a lowland swamp forest. Acta Theriologica 25(5):51-59. MA
Smock, L.A., D.L. Stoneburner, and D.R. Lenat. 1981. Littoral and profundal macroinvertebrate communities of
a coastal brown-water lake. Arch. Hydrobiol. 92(3):306-320. AI
Taylor, B.E., D.L. Mahoney, and R.A. Estes. 1989. Zooplankton production in a Carolina Bay. pp. 425-436 In:
R.R. Sharitz and J.W. Gibbons (eds.). Freshwater Wetlands and Wildlife, Proceedings of a Symposium. CONF-
8603101 (NTIS No. DE90005384). U.S. Dept. Energy, Washington, D.C.
White T.R. and R.C. Fox. 1980. Recolonization of Streams by Aquatic Insects Following Channelization, Vol.
I. Clemson Univ., South Carolina Water Resour. Res. Inst., WRRI-87, W81-01157, OWRT-A-037-Sc(2) (NTIS
PB81-150252), 129 pp. AI
358
-------
Inland Wetlands Having Biologica
Community Measurements
Sou t h Dako t a
-K
.^Vt2
',+
_ +
-)-
+
+
»«»
"W.
+
+ •.
--K
• «
,H ,
+,
»2<
• 23
.*r
,25
+
•^ ^^ +
+ \ + "
.17 +-•
- ^ . +
v— -' !
+ , - ;^-.. +
1 — — — — , ' _ .. ^' '
1
-t- ,19 ~-
-t(
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE + or - 10m,
9 Research Study Site
| Migratory Shorebird Surv»y CBSB) s < i>
Q Braed^ng Btrd Cen»w» C8BC) sit* that mcludas wetland
O Annual Chri«tmoa Bird Count area (15-mile diameter)
Most cover ma i r. I y ^on-w*tlar>d hobitat
"I" Breed ing Bird Survey Storting po ir>ts for 25mi transects
AND point* where iran*e-ls enter new county Most cover
ma i n I y no-n-uel I and hat i tat
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONS) NOT PLOTTED
+ State/Federal waterfowl survey
Th i • mop doe» NOT portray ALL wetland »amp t i nai »ite*
Emphoais i* on sites where connunity-I«v«1 data -ere
collected See chapter 1 for inclusion criteria
31i»» are referenced by cod* number to th# accompany
-------
SOUTH DAKOTA
Mapped
SD1
Duebbert, H.F. and A.M. Frank. 1984. Value of prairie wetlands to duck broods. Wildl. Soc. Bull. 12:27-34.
SD2
Rumble, M.A., and L.D. Flake. 1983. Management considerations to enhance use of stock ponds by waterfowl
broods. J. Range. Manage. 36(6).-691-694. B
SD3
Hubbard, D.E. 1982. Breeding birds in two dry wetlands in eastern South Dakota. Prairie Hat. 14(1):6-8.
B
SD4
Mack, G.D. and L.D. Flake. 1980. Habitat relationships of waterfowl broods on South Dakota stock ponds,J.
Wildl. Manage. 44:695-700. B
SD5
Bue, I.G., L. Blankenship, and W.H. Marshall. 1952. The relationship of grazing practices to waterfowl
breeding populations and production on stock ponds in western South Dakota. Trans. N. Amer. Wildl. Conf.
17:396-414. B I
SD13-14
Hubbard, D.E., J.B. Millar, and D.D. Mayo. 1988. Soil Vegetation Correlations in Prairie Potholes of Beadle
and Deuel Counties, South Dakota. U.S. Fish & Wildl. Serv., Biol. Rep. 88(22):98. P
SD15-18
Klett, A.T., T.L. Shaffer, and D.H. Johnson. 1988. Duck nest success in the Prairie Pothole region. J.
Wildl. Manage. 52(3):431-440. B R
SD19
Benson, N.G. and P.L. Hudson. 1975. Effects of a reduced fall drawdown on benthos abundance in Lake Francis
Case. Trans. Amer. Fish Soc. 104:526-528. AI 1
SD20
Hawkes, C.L. 1979. Aquatic habitat of coal and bentonite clay strip mine ponds in the northern Great Plains.
Ecol. Coal Res. Dev.2:609-614. I P
SD21
Klett, A.T., T.L. Shaffer, and D.H. Johnson. 1988. Duck nest success in the Prairie Pothole region. J. Wildl.
Manage. 52(3):431-440. B R
SD22
Uresk, D.W. and K. Severson. 1988. Waterfowl and shorebird use of surface-mined and livestock water
impoundments on the Northern Great Plains. Great Basin Nat.48(3):353-357. B
SD23-25
USDA Soil Conservation Serv. 1985. Duck and Pheasant Use of Water Bank Program Agreement Areas in East-Central
South Dakota. SCS, Huron, South Dakota.
SDBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
SDBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, MA. B
SDBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
361
-------
SOUTH DAKOTA (continued)
SDCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
Not Mapped
Beck, D.A., D.E. Hubbard, and K.F. Higgins. 1987. Effects of Haying on Seasonal Wetland Hydrophyte and
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Brady, E.N. 1983. Birds on modified wetlands in eastern South Dakota. M.S. Thesis, South Dakota State Univ.,
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Brady, E.N. and B.A. Giron-Pendleton. 1983. Aquatic bird use of wetlands in Brookings County, South Dakota.
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Brewster, W.G. 1975. Breeding waterfowl population in South Dakota. M.S. Thesis, South Dakota State Univ.,
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Brewster, W.G., J.M. Gates, and L.D. Flake. 1976. Breeding waterfowl populations and their distribution in
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Broschart, M.R. and R.L. Under. 1986. Aquatic invertebrates in level ditches and adjacent emergent marsh in
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Donaldson, W.K. 1976. The aquatic ecology of two seasonal marshes in eastern South Dakota. M.S. Thesis,
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Dornbush, J.N. 1984. Suitability of Selected Organisms for Monitoring Leachate at a Refuse Disposal Site.
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Evans, K.E. and R.R. Kerbs. 1977. Avian Use of Livestock Watering Ponds in Western South Dakota. USDA For.
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Evans, C.D. and K.E. Black. 1956. Duck production studies on the prairie potholes of South Dakota. U.S. Fish
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Flake, L.D., G.L. Peterson, and W.L. Tucker. 1977. Habitat relationships of breeding waterfowl of stock ponds
in northwestern South Dakota. Proc. South Dakota Acad. Sci. 56:135-151. B
Flake, L.D. and P.A. Vohs. 1979. Importance of wetland types to duck production and to non-game bird
populations. Completion Rep., Proj. No. B-045-SDAK, Agreement No. 14-34-0001-6118. S.D. St. Univ. 50 pp.
Fritzell, E.K. 1975. Effects of agricultural burning on nesting waterfowl. Can. Field-Nat. 89:21-27. B I
Gates, J.M. and L.D. Flake. 1976. Pilot investigations of the importance of various wetland types to duck
production. NTIS PB-258 780/6ST. South Dakota St. Univ., Brookings. 38 pp.
Hubbard, D.E. 1984. Avian response to recent wetland modification of the Burke Game Production Area, Miner
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Kallemeyn, L.S. and J.F. Novotny. 1977. Fish and fish food organisms in various habitats of the Missouri
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Mack, G.D. 1977. Factors affecting waterfowl brood use of stock ponds in South Dakota. M.S. Thesis, South
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McEnroe, M. 1976. Factors influencing habitat use by breeding waterfowl in South Dakota. M.S. Thesis, South
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362
-------
SOUTH DAKOTA (continued)
Pendleton, G.U. 1984. Small mammals in prairie wetlands: Habitat use and the effects of wetland modification.
M.S. Thesis, South Dakota St. Univ., Brookings, SD. 54 pp. MA
Peterson, G.L. and L.D. Flake. 1977. Observations of wetland bird use of stock ponds in northwestern South
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Robertson, J.A. 1977. Variables associated with breeding waterfowl on South Dakota stock ponds. M.S. Thesis,
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Rumble, M.A. 1979. Habitat preferences and censusing of waterfowl broods on stock ponds in south central
South Dakota. M.S. Thesis, South Dakota St. Univ., Brookings. 42 pp. B
uwaldt, J.J., Jr., L.D. Flake, and J.M. Gates. 1979. Waterfowl pair use of natural and man-made wetlands in
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Schultz, B.D. 1987. Biotic responses of Typha-inonodominant semipermanent wetlands to cattle grazing. M.S.
Thesis, South Dakota St. Univ., Brookings. 92 pp. AI
Searls, D.A. 1974. Influence of vegetation of the distribution of small mammals on a waterfowl production
area. M.S. Thesis, South Dakota St. Univ., Brookings. 47 pp. MA PE
Shearer, L.A. and H.G. Uhlig. 1965. The use of stock-water dugouts by ducks
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Smith, R.I. and Flake L.D. 1985. Movements and habitats of brood-rearing wood ducks on a prairie river. J.
Wildl. Manage. 49:437-442. B
Swanson, J.D. 1959. Wildlife utilization of stock ponds in Minnehaha County, South Dakota. M.S. Thesis, South
Dakota St. Univ., Brookings. 43 pp.
Weber, M.J. 1978. Non-game birds in relation to habitat variation on South Dakota wetlands. M.S. Thesis,
South Dakota St. Univ., Brookings. 54 pp. B
Weber, M.J., P.A. Vohs Jr., and L.D. Flake. 1982. Use of prairie wetlands by selected bird species in South
Dakota. Wilson Bull. 94(4):550-554.
363
-------
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o
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1
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o
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- o
~o a
DO
364
-------
TENNESSEE
Happed
TN1
Fowler, O.K., D.H. Hill, and L.J. Fowler. 1985. Colonization of coal surface mine sediment ponds in Southern
Appalachia by aquatic organisms and breeding amphibians, pp. 261-285. Penn. St. Univ., University Park, PA.
In: R.P. Brooks, D.E. Samuel, and J.B. Hill (eds.). Wetlands and Water Management on Mined Lands. Penn. St.
Univ., University Park, PA. AI H I
TN2
Pierce, C.L., P.M. Crowley, and D.M. Johnson. 1985. Behavior and ecological interactions of larval odonata.
Ecol. 66(5):1504-1512. Al
TN3
Landin, M.C. 1985. Bird and Mammal Use of Selected Lower Mississippi River Borrow Pits. Ph.D. Diss.,
Mississippi State Univ., 405 pp. B MA
TN3
Cobb, S.P., C.H. Pennington, J.A. Baker, and J.E. Scott. 1984. Fishery and ecological investigations of main
stem levee borrow pits along the lower Mississippi River. Mississippi R. Comm., Vicksburg, MS. 120 pp. F
TN5-39
Durham, D., E. Bridges, P. NameI, S. Pearsall, L. Smith, and P. Sorners. 1985. Conserving Natural Communities:
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TN40
Robinson, J.C. and D. Orr. 1988. A quantitative evaluation of moist soil management areas on Cross Creeks
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TN41-43
Young, R.C. and W.M. Dennis. 1983. Productivity of the Aquatic Macrophyte Community of the Holston River:
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Water Res.,TVA/ONR/WR-83/12. Muscle Shoals, AL.
TN44
Turner, L.J. and D.K. Fowler. 1981. Utilization of Surface Mine Ponds in East Tennessee by Breeding
Amphibians. Off. Nat. Res., Div. of Land and Forest Res., TVA, Norris, TN.Contract # 14-16-0009-78-708.
FWS/OBS-81/08. H
TN45
James, W.K., D.R. Lowery, D.H. Webb, and W.B. Wrenn. 1989. Supplement to White Amur Project Report. Tennessee
Valley Authority. Resource development, River Basin Operations, Water Resources.TVA/WR/AB--89/1. Muscle
Shoals, AL.
TN46
Sigrest, J.M. and S.P. Cobb. 1987. Evaluation of Bird and Mammal Utilization of Dike Systems along the Lower
Mississippi River. U.S. Army Corps of Engr., Mississippi River Commission, Lower Mississippi River Environ.
Prog. Rep. 10. Vicksburg, MS. 103 pp.
TNBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
TNBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, MA. B
TNBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
*
TNCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christinas Bird Count Data. Cornell University,
Ithaca, NY. B
365
-------
TENNESSEE (continued)
Not Mapped
Barstow, C.J. 1971. Impact of channelization on wetland habitat in the Obion-Forked Deer Basin, Tennessee.
Trans. N. Amer. Wildl. Nat. Res. Conf. 36:362-376. I B
Bryan, B.A. and C.R. Hupp. 1984. Dendrogeeomorphic evidence of channel changes in an East Tennessee coal area
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Cox, R.J. 1988. A study of the microinvertebrate communities associated with real and artificial bryophytes
in lotic ecosystems. Ph.D. Diss., Univ. Tennessee, Knoxville. 180 pp.
Hall, T.F., W.T. Penfound, and A.D. Hess. 1946. Waterlevel relationships of plants in the Tennessee Valley
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Hupp, C.R., W.C. Carey, and D.E. Bazemore. 1988. Tree growth and species patterns in relation to wetland
sedimentation along a reach of the Middle Fork, Forked Deer River, West Tennessee. Assoc. Southeast. Biol.
Bull. 35:64.
Steenis, J.H. 1947. Recent changes in the marsh and aquatic plant status at Reelfoot Lake. J. Tennessee Acad.
Sci. 22:22-27. RS P
Summers, P.B. 1982. An ecological assessment of 21 sediment ponds at Ollis Creek Mine, Campbell County,
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Webb, D.H. and A.L. Bates. 1989. The aquatic vascular flora and plant communities along rivers and reservoirs
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366
-------
Inland Wetlands Having BioIogica
Community Measurements
Te
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE * or - !Bn,
9 Research Study Sit*
| Migratory Shorebird Surv*y (BSB) sit*
fj Br**ding Bird C*n«u«
-------
TEXAS
Happed
TX1
Wells, F.C., G.A. Jackson, and U.J. Rogers. 1988. Reconnaissance Investigation of Water-Quality, Bottom
Sediment, and Biota Associated with Irrigation Drainage in the Lower Rio Grange Valley and Laguna Atascosa
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TX2
Durocher, P.O., W.C. Provine, and J.E. Kraai. 1984. Relationship between abundance of Largemouth Bass and
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TX34
Klimas, C.V. 1987. Baldcypress response to increased water levels, Caddo Lake, Louisiana-Texas. Wetlands
7:25-37. PW I
TX35
Streng, D.R., J.S. Glitzenstein, and P.A. Harcombe. 1989. Woody seedling dynamics in an East Texas floodplain
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TX36
Institute of Applied Sciences, Univ. of North Texas. 1988. Pre-Impoundment Environmental Study of Ray Roberts
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TX37
Slack, R.D. and L.E. Marcy. 1983. Pre-Impoundment Environmental Study of Aquilla Lake.Final Report.
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TX37
Slack, R.D., B.R. Murphy, W.J. Spearman, and J. Hinson. 1989. Aquilla Lake Environmental Study (Year
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TX37
Slack, R.D., Maceina, M.J., and M.D. Hoy. 1986. Post-Impoundment Environmental Study (Year-Two) of Aquilla
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TX38
Briggs, R. 1982. Avian use of small aquatic habitats in south Texas. M.S.- Thesis, College Agriculture, Texas
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TX39
Til ton, D.A. 1986. Rock Iand Dam Initial Reevaluation Study - Potential Impacts to Inland Fish and Wildlife
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TX40
Baldassarre, G.A. and E.G. Bolen. 1984. Field-feeding ecology of waterfowl on the southern High Plains of
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TX41
Hobaugh, W.C. and J.G. Teer. 1981. Waterfowl use characteristics of flood-prevent ion lakes in north-central
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TXBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
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TXBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
369
-------
TEXAS (continued)
TXBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
Manomet, MA. B
TXBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
TXCBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
Ithaca, NY. B
Not Happed
AUard, D.W. 1982. Littoral microcrustacean population dynamics in Post Oak Lake. Ph.D. Diss., Texas A&M
Univ., College Station, TX. 134 pp.
Allen, C.E. 1975. Bioeconomics and feeding habits of ducks in flooded bottomlands of eastern Texas. M.S.
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Bettoli, P.U. 1987. The restructuring of a forage fish community following large-scale aquatic vegetation
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370
-------
TEXAS (continued)
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Texas. M.S. Thesis, Stephen F. Austin St. Univ., Nacogdoces, TX. 185 pp.
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Mohler, C.L. 1979. An analysis of floodplain vegetation of the Lower Neches Drainage, southeast Texas, with
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371
-------
TEXAS (continued)
Traweek, M.S., Jr. 1981. An introduction to the aquatic ecology of Texas panhandle playas. U.S. Fish Wildl.
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Welter, M.W. 1989. Plant and water-level dynamics in an East Texas shrub/hardwood bottomland wet I and. Wet lands
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Waver, R.H. 1977. Significance of Rio Grande riparian systems upon the avifauna, pp. 165-174 In: R.R. Johnson
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Wellborn, G.A. 1987. The effects of fish predation and thermal regime on an aquatic macroarthropod community.
M.S. Thesis, Univ. Texas, Arlington, TX. 139 pp.
White, D.H., and D. James. 1978. Differential use of fresh water environments by wintering waterfowl of
coastal Texas. Wilson Bull. 90(1):99-111.
Wiest, J.A. 1982. Anuran succession at temporary ponds in a post oak-savanna region of Texas, pp. 39-47 IN:
N.J. Scott (ed.). Wildl. Res. Rep. 13, U.S. Fish & Wildl. Serv., Washington, D.C.
372
-------
Inland Wetlands Having Biologica
Community Measurements
{ i
+ +',-.--*
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Utah • Research Study Site
• Migratory Shorebird Survey (BSB) site
Q Breed.ng Bird Census C8BO site that includes wetland
O Annual Christmas Bird Count area CIS-mile diameter)
Thi• «ap do** NOT portroy ALL wet I and sanpl ing *ite* ,
' Breading Bird Sur vey Storting points for 25mi transects
E»pha«.s is on sites where comrtun i ty-1 .v«l data were AND poir>u wher e transects enter new county Most cover
CO I Iected See chapter I for incI usion cr t teria mainlynon-wetlandKabitat
Sites are referenced by cod* number to the accompanying
state bibliogrophy
USEPA £rtv)ren«*nt«l R«s««
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONCS) NOT PLOTTED
* State/Federal waterfowl survey
Data Conpi I ation PauI Adonu* and Rob»n R*ni»r > o Cartography Jeff Ir i»h
37A
-------
UTAH
Mapped
UT1
Jensen, S., R. Ryel and U.S. Platts. 1989. Classification of riverine/riparian habitat and assessment of
nonpoint source impacts North Fork Humboldt River, Nevada. USDA Forest Service, Intermountian Research Station,
Boise Fisheries Unit.
UT2
RobeI, R.J. 1962. Changes in submersed vegetation following a change in water level. J. WildI. Manage.
26(2):221-224. PM I
UT3
Nelson, N.F. 1954. Factors in the development and restoration of waterfowl habitat at Ogden Bay Refuge, Weber
County, Utah. Utah State Dept. Fish & Game, Pub. # 6. 87 pp.
UT4
Platts, W.S., K.A. Gebhardt, and W.L. Jackson. 1985. The effects of large storm events on basin-range riparian
stream habitats, pp. 30-34 In: R.R. Johnson, C.D. Ziebell, D.R. Patton, P.F. Ffolliott, R.H. Hamre (tech.
coords.). Riparian Ecosystems and Their Management: Reconciling Conflicting Uses. Gen. Tech. Rep. RM-120, USOA
Forest Serv., Fort Collins, CO. PW F
UTS
Stephens, D.U., B. Waddell, and J.B. Miller. 1988. Reconnaissance Investigation of Water Quality, Bottom
Sediment, and Biota Associated with Irrigation Drainage in the Middle Green River Basin. U.S. Geol. Surv.
Water-Resour. Invest. Rep. 88-4011. 70 pp. AI BA I
UT6
Wolf, K. 1955. Some effects of fluctuating and falling water levels on waterfowl production. J. Wildl.
Manage. 19(1):13-23. B I
UT7
Moulton, D.W., W.I. Jenson, and J.B. Low. 1976. Avian botulism epizootiology on sewage oxidation ponds in
Utah. J. Wildl. Manage. 40(4):735-742. I B
UTS
Foote, A.L., J.A. Kadlec, and B.K. Campbell. 1988. Insect herbivory on an inland brackish wetland. Wetlands
8:67-74. AI P
UTS
'hornton, F.G. 1982. Concealment as a factor in nest site selection by seven species of Anatidae in Utah.
l.S. Thesis, Univ. Guelph, Guelph, Ontario. 80 pp.
UTBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
Ithaca, NY. B
UTBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
UTBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
UTCBC1-
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UTAH (continued)
Coombs, R.E. 1970. Aquatic and semi-aquatic plant communities of Utah Lake. Ph.D. Diss., Brigham Young Univ.,
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376
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o
o
CO
O>
c
o
CO
~o
c
o
378
-------
VIRGINIA
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VA3
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VA4-6
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VA7
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VA8-11
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VA12
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VA12
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VA12
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VA12
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VA12
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VA12
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-------
VIRGINIA (continued)
VA14
Gomez, M.M. and F.P. Day, Jr. 1982. Litter nutrient content and production in the Great Dismal Swamp. A.J.
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VA12
Carter, V., M.K. Garrett, and P.T. Gammon. 1988. Relation of hydrogeology, soils and vegetation on the
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VA13
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VIRGINIA (continued)
Ferguson, H.L., R.U. Ellis, and J.B. Uhelan. 1976. Effects of stream channelization on avian diversity and
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381
-------
Inland Wetlands Having Biological
Community Measurements
Vermon t
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE ' or - I 0m i
• Research Study Si(•
Migratory Shorebird Survey CBSB) S'te
Breeding Bird Census (BBC) site that -nc Iud«s wetland
Most cover mainly non-net land habitat
ireedtng Bird Survey Starting points for 25mi tronsects
main1/ non-wfttland habitat
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONCS) NOT PLOTTED
State/Federal waterfowl survey
Th t « map do** NOT por tray ALL weI I and samp 1 ing *i t«*
cotl«ct«d $•• chapter 1 for inclu»ion criteria
USEPA Environmental R««««rcK
S' l«« or* ref«r-«nc«d by cod* numb«r to th« occonpanying
state faibtiogrophy
retory* CorvaHl*. Ortgon
Doto Coupi I ation PauI Adamu* and Rob'n R«nt«r ia Car tograpHy J»ff Ir i*h
382
-------
VERMONT
Mapped
VT2
Gascon, D. and W.C. Leggett. 1977. Distribution, abundance, and resource utilization of littoral zone fishes
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VT3
Gruendling, G.K. and D.J. Bogucki. 1978. Assessment of the physical and biological characteristics of the
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VT3-4
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VT5
Duarte, C.M. and J. Kalff. 1986. Littoral slope as a predictor of the maximum biomass of submerged macrophyte
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VT5
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VT5
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VT8
Possardt, E.E. and W.E. Dodge. 1978. Stream channelization impacts on songbirds and small mammals in Vermont.
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VT8
Dodge, U.E., E.E. Possardt, R.J. Reed, and W.P. MacConnell. 1976. Channelization Assessment, White River,
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Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
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U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
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VTBSB1-
International Shorebird Survey. Unpub. digital data. Shorebird Survey Data. Manomet Bird Observatory,
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U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
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Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
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383
-------
VERMOKT (continued)
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384
-------
Inland Wetlands Having Biologica
Community Measurements
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE * or - 10m,
9 Research Study Site
I Migratory Shorebird Survey C8SB} site
Q Breeding Bird Census (BBC) stte that .ncIudes wetland
O Annual Christmas Bird Count area (15-mife diameter}
~i~ Breeding Bird Survey Starting points for 25m i transects
AND points where transects wnter new county Most cover
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONS) NOT PLOTTED
+ State/Federal waterfowl survey
Thi* viop do«« NOT por tray ALL we II and •amp I i ng * i t«*
Emphaa's is on sites where commoni Iy-I«v«I data were
coI I*c t*d See chap t er t for inclusion criteria
Si t*» ar• referenced by code nu»b«r to the accompany j ng
sVol« bib!lography
USEPA Env I ronfltntat R*s«ercK Laboratory. Corvallla.
Ooto Compilation Pou1 Adamu* and Robm Renter
Cartography Jeff Ir
386
-------
WASHINGTON
Mapped
WAI-6
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WA7
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WA10
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WA12
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WA14
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WA15
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WA17
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WA18
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WA19
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WA20
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387
-------
WASHINGTON (continued)
WA21
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Inland Wetlands Having Biological
Community Measurements
Wisconsin
Th i * nap do«« NOT per tr ay ALL w* 11 and «amp 1 i rig • j t ••
Enpha*i• is on *it*• wh*r« commun11 y-1*v«I dot a w*r•
coI I«c t«d S«* chapter I for ineIu*ion crit«ria
Sit«« or* r«f«r«nc«d by cod* number to th* occompony < ng
•let* bibliography
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE * or - 10m
• Research Study Sit«
I Migratory Sho'r*bird Survey CBSB) Sit»
Q Bretdmfi, B.rd Census
-------
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WISCONSIN (continued)
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396
-------
Inland Wetlands Having Biological
Community Measurements
Wes t Virginia
Thim map do»» NOT porI ray ALL u*tI and sanpI ing * t t«»
Empha*i* i» on site* wh*r e commonity-I eve 1 data were
CO 11ect*d Se* chapter I for t ncIu»ion crit*ria
Site* or* ref«i-«r»c*d by cod* nunb«r to th» accompanypng
•tat* bfbticgraphy
ACCURACY OF SITE LOCATIONS ESTIMATED TO BE + or - I 0m(
9 Research Study SiIm
| Migratory Shorebird Survey CBSB) site
Q Breeding Bird Census CBBC) site that includes wetland
O Annual Christmas Bird Count area CIS-mil* diameter)
Most cover mainly non-wet land habi lot
-f- Breeding 8»rd Survey Starting points for 25m i transects
mainly non-wet I and habitat
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONCS) NOT PLOTTED
* State/Federal waterfowl survey
USEPA Environaenta I R*s*«rch Laboratory, Corvallia. Qrtflon
Da to Compi I ation PauI Adanu* and Robin Renteria Car tography Jeff Irish
398
-------
WEST VIRGINIA
Happed
WV1
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WV12
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WV13
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WEST VIRGINIA (continued)
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Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
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WVBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
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WVBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
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400
-------
Inland Wetlands Having Biologica
Community Measurements
Wyoming
-^ i^ T ^
'"""'' jOi
0
^VlfT
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0 Research Study Site
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O Annual Christmas Bird Count area CIS-mil* diameter)
Mo«t cover ma inly non-wetland habitat
•f* Breeding Bird Survey Star ting points for 25mi transects
AND points where transects enter new county Most cover
SITE LOCATED IN COUNTY, SPECIFIC LOCATIONS) NOT PLOTTED
+ State/Federal waterfowl survey
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Emphas 13 19 on 3 < tes wher* connun i ty- I ev« t data were
collected Se* chap t *r 1 for inclusion criteria
USEPA Environ««ftt«l R«»«*rch
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state bibli ography
Data Compilation Paul Adamu* and Robin R*nt*ria Cartography J*ff Irish
402
-------
WYOMING
Happed
WY1
Henszey, R.J., S.U. Wolf, T.A. Uesche, and Q.D. Skinner. 1988. Assessment of a flow enhancement project as
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WYBBC1-
Cornell Laboratory of Ornithology. Unpub. digital data. Breeding Bird Census Data. Cornell University,
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WYBBS1-
U.S. Fish & Wildl. Service. Unpub. digital data. Breeding Bird Survey Data. Office of Migratory Bird
Management, Washington, D.C. B
WYBW1-
U.S. Fish & Wildl. Service. Unpub. Waterfowl Survey Data. B
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Cornell Laboratory of Ornithology. Unpub. digital data. Christmas Bird Count Data. Cornell University,
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Youngblood, A.P., W.G. Padgett, and A.H. Winward. 1985. Riparian Community Classification of Eastern Idaho-
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404
-------
APPENDIX C. Inland Wetland Community Profile Reports of the U.S. Fish and Wildlife Service.
Brinson, M.M., B.L. Swift, R.C. Plantico, and J.S. Barclay. 1981. Riparian Ecosystems: Their Ecology and
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Profile. Biol. Rep. No. 85(7.14). U.S. Fish and Wildl. Serv., Washington, D.C. 98 pp.
Herdendorf, C.E. 1987. The Ecology of Coastal Marshes of Western Lake Erie: A Community Profile.
Biol. Rep. No. 85(7.9). U.S. Fish and Wildl. Serv., Washington, D.C. 240 pp.
Herdendorf, C.E., C.N. Raphael, and E. Jaworski. 1986. The Ecology of Lake St. Clair wetlands: A
Community Profile. Biol. Rep. No. 85(7.7). U.S. Fish and Wildl. Serv., Washington, D.C. 187 pp.
Manny, B.A 1988. The Ecology of the Detroit River, Michigan: An Ecological Profile. Biol. Rep. 85(7.13).
U.S. Fish and Wildl. Serv., Washington, D.C. 86 pp.
Jahn, L.A., and R.V. Anderson. 1986. The Ecology of Pools 19 and 20, Upper Mississippi River: A
Community Profile. Biol. Rep.85(7.6). U.S. Fish Wildl. Serv., Washington, D.C. 142 pp.
Kantrud, H.A., G.L. Krapu, and G.A Swanson. 1989. Prairie Basin Wetlands of the Dakotas: A
Community Profile. Biol. Rep. 85(7.28). U.S. Fish and Wildl. Serv., Washington, D.C. Ill pp.
Laderman, AD. 1989. The Ecology of Atlantic White Cedar Wetlands: A Community Profile. Biol. Rep.
85(7.21). U.S. Fish and Wildl. Serv., Washington, D.C. 114 pp.
Sharitz, R.R. and J.W. Gibbons. 1982. The Ecology of Southeastern Shrub Bogs (Pocosins) and Carolina
Bays: A Community Profile. Rep. No. FWS/OBS-82/04. U.S. Fish and Wildl. Serv., Washington, D.C. 93
pp.
Vince, S.W., S.R. Humphrey, and R.W. Simons. 1989. The Ecology of Hydric Hammocks: A Community
Profile. Biol. Rep. 85(7.26). U.S. Fish and Wildl. Serv., Washington, D.C. 81 pp.
Wharton, C.H., W.M. Kitchens, and T.W. Sipe. 1982. The Ecology of Bottomland Hardwood Swamps of
the Southeast: A Community Profile. Rep. No. FWS/OBS-81/37. U.S. Fish and Wildl. Serv., Washington,
D.C. 133 pp.
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Windell, J.T., B.E. Willard, D.J. Cooper, S.Q. Foster, C.F. Knud-Hansen, L.P. Rink, and G.N. Kiladis. 1986.
An Ecological Characterization of Rocky Mountain Montane and Subalpine Wetlands. Biol. Rep. 86(11).
U.S. Fish and Wildl. Serv., Fort Collins, CO. 298 pp.
Zedler, P.H. 1987. The Ecology of Southern California Vernal Pools: A Community Profile. Biol. Rep.
85(7.11). U.S. Fish and Wildl. Serv., Washington, D.C. 136 pp.
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