United States Off ice of EPA-600/8-83-016BF
Environmental Protection Research and Development July 1984
Agency Washington, DC 20460
Research and Development
oEPA The Acidic Deposition
Phenomenon and
Its Effects
Critical Assessment
Review Papers
Volume II Effects Sciences
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS
CRITICAL ASSESSMENT REVIEW PAPERS
VOLUME II
Aubrey P. AltshuHer, Editor
Atmospheric Sciences
Rick A. Linthurst, Editor
Effects Sciences
Project Staff
Rick A. linthurst-Director
Betsy A. Hood-Coordinator
Gary B. Blank-Afanuscript Editor
Production
Clara B. Edwards
Wanda Frazier
Elizabeth McKoy
Benita Perry
Graphics
Mike Conley
David Urena
Steven F. Vozzo
C. Willis Williams
Advisory Committee
David A. Bennett-U.S. EPA
Project Officer
John Bachmann-U.S. EPA
Michael Berry-U.S. EPA
Ellis B. Cowling-NCSU
J. Michael Davis-U.S. EPA
Kenneth Demerjian-U.S. EPA
J. H. B. Garner-U.S. EPA
James L. Regens-U.S. EPA
Raymond Wilhour-U.S. EPA
This document has been prepared through the NCSU Acid Deposition Program,
a cooperative agreement between the United States Environmental Protection
Agency', Washington, D.C. and North Carolina State University, Raleigh, North
Carolina. This work was conducted as part of the National Acid Precipitation
Assessment Program and was funded by U.S. EPA.
U.S. Environ.-
Region V i •
230 Soun L
Chicago, lliir!U
Agency
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DISCLAIMER
This document has been reviewed in accordance with U.S. Environmental
Protection Agency policy and approved for publication. Mention of trade
names or commercial products is not intended to constitute endorsement or
recommendation for use.
Agency
n
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AUTHORS
Chapter A-l Introduction
Altshuller, Aubrey Paul, Environmental Sciences Research Laboratory, U.S.
Environmental Protection Agency, MD 59, Research Triangle Park, NC,
27711.
*Nader, John S., 2336 New Bern Ave., Raleigh, NC 27610.
*Niemeyer, Larry E., 4608 Huntington Ct., Raleigh, NC 27609.
Chapter A-2 Natural and Anthropogenic Emission Sources
Homolya, James B., Radian Corp., P. 0. Box 13000, Research Triangle Park, NC
27709.
Robinson, Elmer, Civil and Environmental Engineering Dept., Washington State
University, Pullman, WA, 99164.
Chapter A-3 Transport Processes
*Gillani, Noor V., Mechanical Engineering Dept., Washington University,
Box 1185, St. Louis, MO 63130.
Patterson, David E., Mechanical Engineering Dept., Washington University,
Box 1124, St. Louis, MO 63130.
Shannon, Jack D., Bldg. 181, Environmental Research Div., Bldg. 181, Argonne
National Laboraory, Argonne, IL 60439.
Chapter A-4 Transformation Processes
Gillani, Noor V., Mechanical Engineering Dept., Washington University,
Box 1185, St. Louis, MO 63130.
Hegg, Dean A., Atmospheric Sciences, AK-40, University of Washington,
Seattle, WA 98195.
Hobbs, Peter V., Dept. of Atmospheric Sciences, AK-40, University of
Washington, Seattle, WA 98195.
*Miller, David F., Desert Research Institute, University of Nevada, P. 0. Box
60220, Reno, NY 89506.
, * ' ~*
*Served as co-editor.
i i i
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WhHbeck, Michael, Desert Research Institute, University of Nevada, P. 0. Box
60220, Reno, NV 89506.
Chapter A-5 Atmospheric Concentrations and Distributions
of Chemical Substances
Altshuller, Aubrey Paul, Envlromental Sciences Research Laboratory, U.S.
Environmental Protection Agency, MD 59, Research Triangle Park,
NC 27711.
Chapter A-6 Precipitation Scavenging Processes
Hales, Jeremy M., Geosciences Research and Engineering, Battelle, Pacific
Northwest Laboratories, P. 0. Box 999, Rlchland, WA 99352.
Chapter A-7 Dry Deposition Processes
Hicks, Bruce B., NOAA/ERL, Atmospheric Turbulence and Diffusion D1v., ARL,
P. 0. Box E, Oak Ridge, TN 37830.
Chapter A-8 Deposition Monitoring
Hicks, Bruce B., U.S. Dept. of Commerce, National Oceanic and Atmospheric
Administration, Environmental Research Laboratories, P. 0. Box E,
Oak Ridge, TN 37830.
Lyons, William Berry, Dept. of Earth Sciences, James Hall, University of New
Hampshire, Durham, NH 03824.
Mayewski, Paul A., Dept. of Earth Sciences, James Hall, University of New
Hampshire, Durham, NH 03824.
Stensland, Gary J., Illinois State Water Survey, 605 E. Springfield Ave.,
P. 0. Box 5050, Station A, Champaign, IL 61820.
Chapter A-9 Deposition Models
Bhumralkar, Chandrakant M., Atmospheric Science Center, SRI International,
333 Ravenswood Ave., Menlo Park, CA 94025.
Ruff, Ronald E., Atmospheric Science Center, SRI International, 333
Ravenswood Ave., Menlo Park, CA 94025.
IV
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Chapter E-l Introduction
Linthurst, Rick A., Kllkelly Environmental Associates, Inc., P. 0. Box 31265,
Raleigh, NC 27622.
Chapter E-2 Effects on Soil Systems
Adams, Fred, Dept. of Agronomy and Soils, Auburn University, Auburn, AL
36849.
Cronan, Christopher S., Land and Water Resources Center, 11 Coburn Hall,
University of Maine, Orono, ME 04469.
Firestone, Mary K., Dept. Plant and Soil Biology, 108 Hilgard Hall,
University of California, Berkeley, CA 94720.
Foy, Charles D., U.S. Dept. of Agriculture, Agricultural Research Service,
Plant Stress Lab-BARC West, Beltsville, MD 20705.
Harter, Robert D., College of Life Sciences and Agriculture, James Hall,
University of New Hampshire, NH 03824.
Johnson, Dale W., Environmental Sciences Div., Oak Ridge National Laboratory,
Oak Ridge, TN 37830.
*McFee, William W., Natural Resources and Environmental Sciences Program,
Purdue University, West Lafayette, IN 47907.
Chapter E-3 Effects on Vegetation
Chevone, Boris I., Dept. of Plant Pathology, Virginia Polytechnic Institute
and State University, Blacksburg, VA 24060.
Irving, Patricia M., Environmental Research Div., Bldg. 203, Argonne
National Laboratory, Argonne, It 60439.
Johnson, Arthur H., Dept. of Geology D4, University of Pennsylvania,
Philadelphia, PA 19104.
*Johnson, Dale W., Environmental Sciences Div., Oak Ridge National
Laboratory, Oak Ridge, TN 37830.
Lindberg, Steven E., Environmental Sciences Div., Bldg. 1505, Oak Ridge
National Laboratory, Oak Ridge, TN 37830.
McLaughlin, Samuel B., Environmental Sciences Div., Bldg. 3107, Oak Ridge
National Laboratory, Oak Ridge, TN 37830.
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Raynal, Dudley J., Dept. of Environmental and Forest Biology, College of
Environmental Science and Forestry, State University of New York (SUNY),
Syracuse, NY 13210.
Shrlner, David S., Environmental Sciences D1v., Oak Ridge National
Laboratory, Oak Ridge, TN 37830.
S1gal, Lorene L., Environmental Sciences Div., Oak Ridge National Laboratory,
Oak Ridge, TN 37830.
Skelly, John M., Dept. of Plant Pathology, 211 Buckhout Laboratory,
Pennsylvania State University, University Park, PA 16802.
Smith, William H., School of Forestry and Environmental Studies, Yale
University, 370 Prospect Street, New Haven, CT 06511.
Weber, Jerome B., Dept. of Crop Science, North Carolina State University,
Raleigh, NC 27650.
Chapter E-4 Effects on Aquatic Chemistry
Anderson, Dennis S., Dept. of Botany and Plant Pathology, University of
Maine, Orono, ME 04469.
*Baker, Joan P., NCSU Acid Deposition Program, North Carolina State
University, 1509 Varsity Dr., Raleigh, NC 27606.
Blank, G. B., School of Forest Resources, Biltmore Hall, North Carolina State
University, NC 27650.
Church, M. Robbins, Corvallis Environmental Research Laboratory, U.S.
Environmental Protection Agency, 200 SW 35th Street, Corvallis, OR
97333.
Cronan, Christopher S., Land and Water Resources Center, 11 Coburn Hall,
University of Maine, Orono, ME 04469.
Davis, Ronald B., Dept. of Botany and Plant Pathology, Univeristy of Maine,
Orono, ME 04469.
Dillon, Peter J., Ontario Ministry of the Environment, Limnology Unit, P. 0.
Box 39, Dorset, Ontario, Canada, POA 1EO.
Driscoll, Charles T., Dept. of Civil Engineering, 150 Hinds Hall, Syracuse
University, NY 13210.
*Galloway, James N., Dept. of Environmental Sciences, University of Virginia,
Charlottesville, VA 22903.
Gregory, J. D., School of Forest Resources, Biltmore Hall, North Carolina
State University, NC 27650.
vi
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Norton, Stephen A., Dept. of Geological Sciences, 110 Boardman Hall,
University of Maine, Orono, ME 04469.
Schafran, Gary C., Dept. of C1v1l Engineering, 150 Hinds Hall, Syracuse
University, Syracuse, NY 13210.
Chapter E-5 Effects on Aquatic Biology
Baker, Joan P., NCSU Add Deposition Program, North Carolina State
University, 1509 Varsity Dr., Raleigh, NC 27606.
Drlscoll, Charles T., Dept. of Civil Engineering, 150 Hinds Hall, Syracuse
University, Syracuse, NY 13210.
Fischer, Kathleen L., Canadian Wildlife Service, National Wildlife Research
Centre, Environment Canada, 100 Gamelin Blvd., Hull, Quebec, Canada,
K1A OE7.
Guthrie, Charles A., New York State Department of Environmental Conservation,
Div. of Fish and Wildlife, Bldg. 40, SUNY-Stony Brook, Stony Brook, NY
11790.
*Magnuson, John J., Laboratory of Limnology, University of Wisconsin,
Madison, WI 53706.
Peverly, John H., Dept. of Agronomy, University of Illinois, Urbana, IL 61801
*Rahel, Frank J., Dept. of Zoology, Ohio State University, 1735 Neil Ave.,
Columbus, OH 43210.
Schafran, Gary C., Dept. of Civil Engineering, 150 Hinds Hall, Syracuse
University, Syracuse, NY 13210.
Singer, Robert, Dept. of Civil Engineering, 150 Hinds Hall, Syracuse
University, Syracuse, NY 13210.
Chapter E-6 Indirect Effects on Health
Baker, Joan P., NCSU Acid Precipitation Program, North Carolina State
University, 1509 Varsity Dr., Raleigh, NC 27606.
Clarkson, Thomas W., University of Rochester School of Medicine, P. 0. Box
RBB, Rochester, NY 14642.
Sharpe, William E., Land and Water Research Bldg., Pennsylvania State
University, University Park, PA 16802.
vi i
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Chapter E-7 Effects on Materials
Baer, Norbert S., Conservation Center of the Institute of Fine Arts,
New York University, 14 East 78th Street, New York, NY 10021.
Kirmeyer, Gregory, Economic and Engineering Services, Inc., 611 N. Columbia,
Olympia, WA 98507.
Yocom, John E., TRC Environmental Consultants, Inc., 800 Connecticut Blvd.,
East Hartford, CT 06108.
vi ii
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PREFACE
The Acidic Deposition Phenomenon and Its Effects: Critical Assessment Review
Papers was written at suggestion in the summer of 1980, by the Chairman of
the Clean Air Scientific Advisory Committee of EPA's Science Advisory Board.
The document was prepared for EPA through the Acid Deposition Program at
North Carolina State University. This document is the first of several
documents of increasing sophistication that assess the acidic deposition
phenomenon. It will be succeeded by assessment documents in 1985, 1987, and
1989, based largely on research of the National Acid Precipitation Assessment
Program.
The document's original charge was to prepare "a comprehensive document which
lays out the state of our knowledge with regard to precursor emissions, pol-
lutant transformation to acidic compounds, pollutant transport, pollutant
deposition and the effects (both measured and potential) of acidic deposi-
tion." The decision of the editors provided the following guidelines to the
authors writing the Critical Assessment Review Papers to meet this overall
objective of the document:
1. Contributions are to be written for scientists and informed lay
persons.
2. Statements are to be explained and supported by references;
i.e., a textbook type of approach, in an objective style.
3. Literature referenced is to be of high quality and not every
reference available is to be included.
4. Emphasis is to be placed on North American systems with
concentrated effort on U.S. data.
5. Overlap between this document and the SOX Criteria Document
is to be minimized.
6. Potential vs known processes/effects are to be clearly noted to
avoid misinterpretation.
7. The certainty of our knowledge should be quantified, when
possible.
8. Conclusions are to be drawn on fact only.
9. Extrapolation beyond the available data is to avoided.
10. Scientific knowledge is to be included without regard to
policy implications.
IX
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11. Policy-related options or recommendations are beyond the scope
of this document and are not to be included.
The reader, to avoid possible misinterpretation of the information pre-
sented, is advised to consider and understand these directives before
readi ng.
Again, the document has been designed to address our present status of know-
ledge of the acidic deposition phenomenon and its effects. It is not a
Criteria Document; it is not designed to set standards and no connections to
regulations should be inferred. The literature is reviewed and conclusions
are drawn based on the best evidence available. It is an authored document,
and as such, the conclusions are those of the authors after their review of
the literature.
The success of the Critical Assessment Review Papery has depended on the
coordinated efforts of many individuals. The document involved the partici-
pation of over 60 scientists contributing material on their special areas of
expertise under the broad headings of either atmospheric processes or ef-
fects. Coordination within these two areas has been the responsibility of A.
Paul Altshuller and Rick A. Linthurst, the atmospheric and effects section
editors, respectively. Overall coordination of the project for EPA is under
David A. Bennett's direction. Dr. Altshuller is an atmospheric chemist, past
recipient of the American Chemical Society's Award in Pollution Control,
and recently retired director of EPA's Environmental Sciences Research
Laboratory; Dr. Linthurst is an ecologist and served as Program Coordinator
for the Acid Precipitation Program at North Carolina State University. He is
currently at Kilkelly Environmental Associates, Inc. Dr. Bennett is the
Director of the Acid Deposition Assessment Staff in EPA's Office of Research
and Development.
The written materials that follow are contributions from one to eight authors
per chapter, integrated by the editors. Approximately 75 scientists, with
expertise in the fields being addressed, reviewed early drafts of the chap-
ters. In addition, 200 individuals participated in a public workshop held
for the technical review of these materials in November 1982. Numerous
changes resulted from these reviews, and this document reflects those com-
ments. A public review draft of this document was distributed in June 1983
for a 45-day comment period. During that period, 130 sets of comments from
53 reviewers were received. These comments were summarized and evaluated by
a technical and editorial panel, and then provided to the authors who ad-
dressed them by revision and rewriting to produce this final document, in
response to the comments received, revisions were made to all chapters in-
cluding a major revision of Chapter E-4, Effects on Aquatic Chemistry, and
the addition of a section on Corrosion in water piping systems in Chapter
E-7, Effects on Materials.
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ACKNOWLEDGMENTS FROM NORTH CAROLINA STATE UNIVERSITY
The editorial staff wishes to extend special thanks to all the authors of
this document. They have been patient and tolerant of our changes, re-
commendations, and deadlines, leading to this fourth and final version of
the document. These dedicated scientists are to be commended for their
efforts.
We also wish to acknowledge our Steering Committee, who has been patient with
our errors and deadline delays. These people have made major contributions
to this product, and actively assisted us with their recommendations on pro-
ducing this document. Their objectivity, concern for technical accuracy, and
support is appreciated. Dr. J. Michael Davis of EPA deserves special thanks,
as he directed the initial draft of the document in December of 1981. His
concern for clarity of thought and writing in the interest of communicating
our scientific knowledge was most helpful. Dr. David Bennett of EPA is
specifically recognized for his role as a scientific reviewer, and an EPA
staff member who buffered the editorial staff and the authors from the public
and policy concerns associated with this document. Dr. Bennett's tolerance,
patience, and understanding are also appreciated.
All the reviewers, too numerous to list, are gratefully acknowledged for
helping us improve the quality and accuracy of this document. These people
were from private, state, federal, and special-interest organizations in both
the United States and Europe. Their concern for the truth, based on the
available data, is a compliment to all the individuals and organizations who
were willing to deal objectively with this most important topic. It has been
a pleasure to see all groups, independent of their personal philosophies,
work together in the interest of producing a technically accurate document.
Dr. Arthur Stern is acknowledged for his contribution as a technical editor
of the atmospheric sciences early in the document's preparation. He has made
an important contribution to the final product.
Finally, EPA is acknowledged for its willingness to give the scientists an
opportunity to prepare this document. Its interest, as expressed through the
staff and authors, in having this document be an authored document to assist
in research planning, is most appreciated. Rarely does a group of scientists
have such a free hand in contributing independently to such an important
issue and in such a visible way. Although coordinating the efforts of so
many scientists can be a difficult and lengthy process, we feel the authored
scientific product makes a valuable contribution to the acidic deposition
issue.
The entire staff of the NCSU Acid Deposition Program and several part-time
workers have been involved in the production of this document since it began
in 1981. In addition to the people listed on the title page, these include:
William R. Alsop - Program Assistant
Ann Bartuska - Program Coordinator
XI
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Jody D. Castleberry - Receptionist/Secretary
Connie S. Harp - Secretary
Jeanie Hartman - Librarian
Helen Koop - Library Assistant
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS:
CRITICAL ASSESSMENT REVIEW PAPERS
Table of Contents
Volume I
Atmospheric Sciences
Page
AUTHORS HI
PREFACE 1 x
ABBREVIATION-ACRONYM LIST xxlx
GLOSSARY xl 111
A-l INTRODUCTION
1.1 Objectives 1-1
1.2 Approach—Movement from Sources to Receptor 1-1
1.2.1 Chemical Substances of Interest 1-1
1.2.2 Natural and Anthropogenic Emissions Sources 1-1
1.2.3 Transport Processes 1-1
1.2.4 Transformation Processes 1-2
1.2.5 Atmospheric Concentrations and Distributions of Chemical
Substances 1-2
1.2.6 Precipitation Scavenging Processes 1-2
1.2.7 Dry Deposition Processes 1-3
1.2.8 Deposition Monitoring 1-3
1.2.9 Deposition Models 1-4
1.3 Acidic Deposition 1-4
A-2 NATURAL AND ANTHROPOGENIC EMISSIONS SOURCES
2.1 Introduction 2-1
2.2 Natural Emission Sources 2-1
2.2.1 Sulfur Compounds 2-1
2.2.1.1 Introduction 2-1
2.2.1.2 Estimates of Natural Sources 2-2
2.2.1.3 Blogenlc Emissions of Sulfur Compounds 2-3
2.2.1.4 Geophysical Sources of Natural Sulfur Compounds 2-15
2.2.1.4.1 Volcanlsm 2-17
2.2.1.4.2 Marine sources of aerosol particles and
gases 2-19
2.2.1.5 Scavenging Processes and Sinks 2-21
2.2.1.6 Summary of Natural Sources of Sulfur Compounds 2-22
2.2.2 Nitrogen Compounds 2-23
2.2.2.1 Introduction 2-23
2.2.2.2 Estimates of Natural Global Sources and Sinks 2-24
2.2.2.3 Blogenlc Sources of NOX Compounds 2-28
2.2.2.4 Tropospherlc and Stratospheric Reactions 2-30
2.2.2.5 Formation of NOX by Lightning 2-30
2.2.2.6 Blogenlc NOX Emissions Estimate for the United States ... 2-32
2.2.2.7 Blogenlc Sources of Ammonia 2-33
2.2.2.8 Oceanic Source for Ammonia 2-36
2.2.2.9 Blogenlc Ammonia Emissions Estimates for the United
States 2-37
2.2.2.10 Meteorological and Area Variations for NOX and Ammonia
Emissions 2-38
2.2.2.11 Scavenging Processes for NOX and Ammonia 2-38
2.2.2.12 Organic Nitrogen Compounds 2-39
2.2.2.13 Summary of Natural NOX and Ammonia Emissions 2-39
xm
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Table of Contents (continued)
Page
2.2.3 Chlorine Compounds 2-39
2.2.3.1 Introduction 2-39
2.2.3.2 Oceanic Sources 2-40
2.2.3.3 Volcanl sm 2-44
2.2.3.4 Combustion 2-44
2.2.3.5 Total Natural Chlorine Sources 2-45
2.2.3.6 Seasonal Distributions 2-45
2.2.3.7 Environmental Impacts of Natural Chlorides 2-45
2.2.4 Natural Sources of Aerosol Particles 2-45
2.2.5 Precipitation pH 1n Background Conditions 2-48
2.2.6 Summary 2-52
2.3 Anthropogenic Emissions 2-53
2.3.1 Origins of Anthropogenlcally Emitted Compounds and
Related Issues 2-53
2.3.2 Historical Trends and Current Emissions of Sulfur Compounds 2-57
2.3.2.1 Sulfur Oxides 2-57
2.3.2.2 Primary Sulfate Emissions 2-62
2.3.3 Historical Trends and Current Emissions of Nitrogen Oxides 2-68
2.3.4 Historical Trends and Current Emissions of Hydrochloric Add (HC1) 2-72
2.3.5 Historical Trends and Current Emissions of Heavy Metals Emitted
from Fuel Combustion 2-76
2.3.6 Historical Emissions Trends In Canada 2-84
2t3.7 Future Trends In Emissions 2-93
2.3.7.1 United States 2-93
2.3.7.2 Canada 2-93
2.3.8 Emissions Inventories 2-96
2.3.9 The Potential for Neutralization of Atmospheric
Acidity by Suspended Fly Ash 2-97
2.4 Conclusions 2-102
2.5 References 2-106
A-3 TRANSPORT PROCESSES
3.1 Introduction 3-1
3.1.1 The Concept of Atmospheric Residence Times 3-2
3.2 Meteorological Scales and Atmospheric Motions 3-3
3.2.1 Meteorological Scales 3-3
3.2.2 Atmospheric Motions 3-4
3.3 Pollutant Transport Layer: Its Structure and Dynamics 3-10
3.3.1 The Planetary Boundary Layer (Mixing Layer) 3-10
3.3.2 Structure of the Transport Layer (TL) 3-12
3.3.3 Dynamics of the Transport Layer 3-16
3.3.4 Effects of Mesoscale Complex Systems on Transport Layer Structure
and Dynamics 3-27
3.3.4.1 Effect of Mesoscale Convectlve Precipitation Systems
(MCPS) 3-27
3.3.4.2 Complex Terrain Effects : 3-31
3.3.4.2.1 Shoreline environment effects 3-31
3.3.4.2.2 Urban effects 3-34
3.3.4.2.3 Hilly terrain effects 3-35
3.4 Mesoscale Plume Transport and Dilution 3-38
3.4.1 Elevated Point-Source Emissions (Power Plant Plumes) 3-38
3.4.2 Broad Areal Emissions Near Ground (Urban Plumes) 3-60
3.5 Continental and Hemispheric Transport 3-65
3.6 Conclusions 3-88
3.7 References 3-92
XIV
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Table of Contents (continued)
Page
A-4 TRANSFORMATION PROCESSES
4.1 Introduction 4-1
4.2 Homogeneous Gas-Phase Reactions 4-3
4.2.1 Fundamental Reactions 4-3
4.2.1.1 Reduced Sulfur Compounds 4-3
4.2.1.2 Sulfur Dioxide 4-4
4.2.1.3 Nitrogen Compounds 4-11
4.2.1.4 Halogens 4-17
4.2.1.5 Organic Acids 4-17
4.2.2 Laboratory Simulations of Sulfur Dioxide and Nitrogen Dioxide
0x1 datl on 4-17
4.2.3 Field Studies of Gas-Phase Reactions 4-21
4.2.3.1 Urban Plumes 4-21
4.2.3.2 Power PI ant PIumes 4-24
4.2.4 Summary 4-29
4.3 Solution Reactions - 4-31
4.3.1 Introduction 4-31
4.3.2 Absorption of Add 4-32
4.3.3 Production of HC1 1n Solution 4-38
4.3.4 Production of HN03 In Solution 4-38
4.3.5 Production of H2S04 In Solution 4-42
4.3.5.1 Evidence from Field Studies 4-42
4.3.5.2 Homogeneous Aerobic Oxidation of S02'H20 to H2S04 4-43
4.3.5.2.1 Uncatalyzed 4-43
4.3.5.2.2 Catalyzed 4-45
4.3.5.3 Homogeneous Non-aerobic Oxidation of S02'H20 to H£S04 ... 4-47
4.3.5.4 Heterogeneous Production of H2S04 1n Solution 4-52
4.3.5.5 The Relative Importance of the Various H2S04
Production Mechanisms 4-53
4.3.6 Neutralization Reactions 4-61
4.3.6.1 Neutralization by NH3 4-61
4.3.6.2 Neutralization by Particle-Add Reactions 4-62
4.3.7 Summary 4-63
4.4 Transformation Models 4-63
4.4.1 Introduction 4-63
4.4.2 Approaches to Transformation Modeling 4-66
4.4.2.1 The Fundamental Approach 4-66
4.4.2.2 The Empirical Approach 4-68
4.4.3 The Question of Linearity 4-71
4.4.4 Some Specific Models and Their Applications 4-74
4.4.4.1 Detailed Chemical Simulations 4-74
4.4.4.2 Parameterized Models 4-67
4.4.5 Summary 4-81
4.5 Conclusions 4-82
4.6 References ••• 4-86
A-5 ATMOSPHERIC CONCENTRATIONS AND DISTRIBUTIONS OF CHEMICAL SUBSTANCES
5.1 Introduction 5-1
5.2 Sulfur Compounds 5-2
5.2.1 Historical Distribution Patterns 5-2
5.2.2 Sulfur Dioxide 5-3
5.2.2.1 Urban Measurements 5-3
5.2.2.2 Nonurban Measurements 5-4
5.2.2.3 Concentration Measurements at Remote Locations 5-12
5.2.2.4 Comparison of Sulfur Dioxide Emissions and Ambient
Air Concentration 5-12
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Table of Contents (continued)
Page
5.2.3 Sul fate 5-13
5.2.3.1 Urban Concentration Measurements 5-13
5.2.3.2 Urban Composition Measurements 5-15
5.2.3.3 Nonurban Concentration Measurements 5-16
5.2.3.4 Nonurban Composition Measurements 5-19
5.2.3.5 Concentration and Composition Measurements at Remote
Locations 5-22
5.2.3.6 Comparison of Sulfur Oxide Emissions and Ambient Air
Concentrations of Sulfate 5-23
5.2.4 Particle Size Characteristics of Partlculate Sulfur Compounds 5-24
5.2.4.1 Urban Measurements 5-24
5.2.4.2 Nonurban Size Measurements 5-27
5.2.4.3 Measurements at Remote Locations 5-27
5.3 Nitrogen Compounds 5-28
5.3.1 Introduction 5-28
5.3.2 Nitrogen Oxides 5-28
5.3.2.1 Historical Distribution Patterns and Current
Concentrations of Nitrogen Oxides 5-28
5.3.2.2 Measurements Techniques-Nitrogen Oxides 5-29
5.3.2.3 Urban Concentration Measurements 5-29
5.3.2.4 Nonurban Concentration Measurements 5-30
5.3.2.5 Measurements of Concentrations at Remote Locations 5-34
5.3.3 Nitric Add 5-38
5.3.3.1 Urban Concentration Measurements 5-38
5.3.3.2 Nonurban Concentration Measurements 5-40
5.3.3.3 Concentration Measurements at Remote Locations 5-44
5.3.4 Peroxyacetyl Nitrates 5-45
5.3.4.1 Urban Concentration Measurements 5-45
5.3.4.2 Nonurban Concentration Measurements 5-48
5.3.5 Ammonia 5-50
5.3.5.1 Urban Concentration Measurements 5-50
5.3.5.2 Nonurban Concentration Measurements 5-51
5.3.6 Partlculate Nitrate 5-51
5.3.6.1 Urban Concentration Measurements 5-53
5.3.6.2 Nonurban Concentration Measurements 5-55
5.3.6.3 Concentration Measurements at Remote Locations 5-56
5.3.7 Particle Size Characteristics of Partlculate Nitrogen Compounds .. 5-56
5.4 Ozone 5-58
5.4.1 Concentration Measurements Within the Planetary Boundary Layer
(PBL) 5-60
5.4.2 Concentration Measurements at Higher Altitudes 5-63
5.5 Hydrogen Peroxide 5-63
5.5.1 Urban Concentration Measurements 5-64
5.5.2 Nonurban Concentration Measurements 5-64
5.5.3 Concentration Measurements In Rainwater 5-65
5.6 Chiorlne Compounds , 5-65
5.6.1 Introduction 5-65
5.6.2 Hydrogen Chloride 5-66
5.6.3 Partlculate Chloride 5-66,
5.6.4 Particle Size Characteristics of Partlculate Chlorine Compounds .. 5-67
5.7 Metallic Elements 5-68
5.7.1 Concentration Measurements and Particle Sizes In Urban Areas 5-68
5.7.2 Concentration Measurements and Particle Sizes In Nonurban Areas .. 5-71
5.8 Relationship of Light Extinction and Visual Range Measurements to Aerosol
Composition 5.73
5.8.1 Fine Particle Concentration and Light Scattering Coefficients .... 5-73
5.8.2 Light Extinction or Light Scattering Budgets at Urban Locations .. 5-74
5.8.3 Light Extinction or Light Scattering Budgets at Nonurban
Locations 5-76
5.8.4 Trends In Visibility as Related to Sulfate Concentrations 5-78
5.9 Conclusions 5-78
5.10 References 5-84
XVI
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Table of Contents (continued)
Page
A-6 PRECIPITATION SCAVENGING PROCESSES
6.1 Introduction 6-1
6.2 Steps In the Scavenging Sequence 6-2
6.2.1 Introduction 6-2
6.2.2 Intermixing of Pollutant and Condensed Water (Step 1-2) 6-5
6.2.3 Attachment of Pollutant to Condensed Water Elements (Step 2-3) ... 6-6
6.2.4 Aqueous-Phase Reactions (Step 3-4) , 6-13
6.2.5 Deposition of Pollutant with Precipitation (Steps 3-5 and 4-5) ... 6-13
6.2.6 Combined Processes and the Problem of Scavenging Calculations .... 6-16
6.3 Storm Systems and Storm Climatology 6-16
6.3.1 Introduction 6-16
6.3.2 Frontal Storm Systems 6-17
6.3.2.1 Warm-Front Storms 6-19
6.3.2.2 Cold-Front Storms 6-23
6.3.2.3 Occluded-Front Storms 6-23
6.3.3 Convectlve Storm Systems 6-23
6.3.4 Additional Storm Types: Nonldeal Frontal Storms, OrograpMc
Storms and Lake-Effect Storms 6-27
6.3.5 Storm and Precipitation Climatology 6-28
6.3.5.1 Precipitation Climatology 6-28
6.3.5.2 Storm Tracks 6-28
6.3.5.3 Storm Duration Statistics 6-31
6.4 Summary of Precipitation-Scavenging Field Investigations 6-31
6.5 Predictive and Interpretive Models of Scavenging 6-41
6.5.1 Introduction 6-41
6.5.2 Elements of a Scavenging Model 6-50
6.5.2.1 Material Balances 6-50
6.5.2.2 Energy Balances 6-52
6.5.2.3 Momentum Balances 6-52
6.5.3 Definitions of Scavenging Parameters 6-53
6.5.4 Formulation of Scavenging Models: Simple Examples
of Microscopic and Macroscopic Approaches 6-58
6.5.5 Systematic Selection of Scavenging Models:
A Flow Chart Approach 6-61
6.6 Practical Aspects of Scavenging Models: Uncertainty Levels and Sources
of Error 6-64
6.7 Conclusions 6-68
6.8 References 6-71
A-7 DRY DEPOSITION PROCESSES
7.1 Introduction 7-1
7.2 Factors Affecting Dry Deposition 7-1
7.2.1 Introduction - 7-1
7.2.2 Aerodynamic Factors -. 7-6
7.2.3 The Quasi-Laminar Layer 7-9
7.2.4 Phoretlc Effects and Stefan Flow 7-13
7.2.5 Surface Adhesion 7-14
7.2.6 Surface Biological Effects 7-15
7.2.7 Deposition to Liquid Water Surfaces 7-16
7.2.8 Deposition to Mineral and Metal Surfaces 7-17
7.2.9 Fog and Dewfall 7-19
7.2.10 Resuspenslon and Surface Emission 7-20
7.2.11 The Resistance Analog 7-21
7.3 Methods for Studying Dry Deposition 7-27
7.3.1 Direct Measurement 7-27
7.3.2 Wind-Tunnel and Chamber Studies 7-29
7.3.3 Mlcrometeorologlcal Measurement Methods 7-33
XVI1
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Table of Contents (continued)
Page
7.4 Field Investigations of Dry Deposition 7-37
7.4.1 Gaseous Pollutants 7-37
7.4.2 Partlculate Pollutants 7-44
7.4.3 Routine Handling In Networks 7-50
7.5 Mlcrometeorologlcal Models of the Dry Deposition Process 7-51
7.5.1 Gases 7-51
7.5.2 Particles 7-53
7.6 Summary 7-54
7.7 Conclusions 7-58
7.8 References 7-60
A-8 DEPOSITION MONITORING
8.1 Introduction 8-1
8.2 Wet Deposition Networks 8-2
8.2.1 Introduction and Historical Background 8-2
8.2.2 Definitions 8-3
8.2.3 Methods, Procedures and Equipment for Wet Deposition Networks .... 8-5
8.2.4 Wet Deposition Network Data Bases 8-7
8.3 Monitoring Capabilities for Dry Deposition 8-12
8.3.1 Introduction 8-12
8.3.2 Methods for Monitoring Dry Deposition 8-18
8.3.2.1 Direct Collection Procedures 8-19
8.3.2.2 Alternative Methods 8-20
8.3.3 Evaluations of Dry Deposition Rates 8-22
8.4 Wet Deposition Network Data With Applications to Selected Problems 8-31
8.4.1 Spatial Patterns 8-31
8.4.2 Remote Site pH Data 8-50
8.4.3 Precipitation Chemistry Variations Over Time 8-60
8.4.3.1 Nitrate Variation Since 1950's 8-60
8.4.3.2 pH Variation Since 1950's 8-63
8.4.3.3 Calcium Variation Since the 1950's 8-67
8.4.4 Seasonal Variations 8-67
8.4.5 Very Short Time Scale Variations 8-69
8.4.6 A1r Parcel Trajectory Analysis 8-69
8.5 Glaclochemlcal Investigations as a Tool In the Historical Delineation of
the Acid Precipitation Problems 8-71
8.5.1 Glaclochemlcal Data 8-72
8.5.1.1 Sulfate - Polar Glaciers 8-73
8.5.1.2 Nitrate - Polar Glaciers 8-73
8.5.1.3 pH and Acidity - Polar Glaciers 8-74
8.5.1.4 Sulfate - Alpine Glaciers 8-74
8.5.1.5 Nitrate - Alpine Glaciers 8-74
8.5.1.6 pH and Acidity - Alpine Glaciers 8-75
8.5.2 Trace Metals - General Statement 8-75
8.5.2.1 Trace Metals - Polar Glaciers 8-76
8.5.2.2 Trace Metals - Alpine Glaciers 8-77
8.5.3 Discussion and Future Work 8-78
8.6 Conclusions 8-80
8.7 References 8-85
A-9 LONG-RANGE TRANSPORT AND ACIDIC DEPOSITION MODELS
9.1 Introduction 9-1
9.1.1 General Principles for Formulating Pollution Transport and
D1ffusion Models 9-1
9.1.2 Model Characteristics 9-3
9.1.2.1 Spatial and Temporal Scales 9-3
9.1.2.2 Treatment of Turbulence 9-3
xvm
-------
Table of Contents (continued)
Page
9.1.2.3 Reaction Mechanisms 9-5
9.1.2.4 Removal Mechanisms 9-5
9.1.3 Selecting Models for Application 9-6
9.1.3.1 General 9-6
9.1.3.2 Spatial Range of Application 9-6
9.1.3.3 Temporal Range of Application 9-6
9.2 Types of LRT Models 9-9
9.2.1 Eulerlan Grid Models 9-9
9.2.2 Lagranglan Models 9-9
9.2.2.1 Lagranglan Trajectory Models 9-9
9.2.2.2 Statistical Trajectory Models 9-11
9.2.3 Hybrid Models 9-13
9.3 Modules Associated with Chemical (Transformation) Processes 9-13
9.3.1 Overview 9-13
9.3.2 Chemical Transformation Modeling 9-14
9.3.2.1 Simplified Modules 9-14
9.3.2.2 Multlreactlon Modules 9-15
9.3.3 Modules for NOX Transformation 9-16
9.4 Modules Associated with Wet and Dry Deposition 9-17
9.4.1 Overview 9-17
9.4.2 Modules for Wet Deposition 9-20
9.4.2.1 Formulation and Mechanism 9-20
9.4.2.2 Modules Used In Existing Models 9-21
9.4.2.3 Wet Deposition Modules for Snow 9-23
9.4.2.4 Wet Deposition Modules for NOX 9-23
9.4.3 Modules for Dry Deposition 9-24
9.4.3.1 General Considerations 9-24
9.4.3.2 Modules Used In Existing Models 9-25
9.4.3.3 Dry Deposition Modules for NOx 9-26
9.4.4 Dry Versus Wet Deposition 9-26
9.5 Status of LRT Models as Operational Tools 9-26
9.5.1 Overview 9-26
9.5.2 Model Application 9-27
9.5.2.1 Limitations In Applicability 9-27
9.5.2.2 Regional Concentration and Deposition Patterns 9-27
9.5.2.3 Use of Matrix Methods to Quantify Source-Receptor
Relationships 9-28
9.5.3 Data Requirements 9-33
9.5.3.1 General 9-33
9.5.3.2 Specific Characteristics of Data Used 1n Model
Simulations 9-36
9.5.3.2.1 Emissions 9-36
9.5.3.2.2 Meteorological Data 9-37
9.5.4 Model Performance and Uncertainties 9-37
9.5.4.1 General 9-37
9.5.4.2 Data Bases Available for Evaluating Models 9-39
9.5.4.3 Performance Measures 9-39
9.5.4.4 Representativeness of Measurements 9-40
9.5.4.5 Uncertainties 9-40
9.5.4.6 Selected Results 9-40
9.6 Conclusions 9-46
9.7 References 9-48
XIX
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS:
CRITICAL ASSESSMENT REVIEW PAPERS
Table of Contents
Volume II
Effects Sciences
Page
AUTHORS 111
PREFACE 1x
ABBREVIATION-ACRONYM LIST xxlx
GLOSSARY xllll
E-l INTRODUCTION
1.1 Objectives 1-1
1.2 Approach • 1-1
1.3 Chapter Organization and General Content 1-3
1.3.1 Effects on Soil Systems 1-3
1.3.2 Effects on Vegetation 1-4
1.3.3 Effects on Aquatic Chemistry 1-5
1.3.4 Effects on Aquatic Biology 1-5
1.3.5 Indirect Effects on Health 1-6
1.3.6 Effects on Materials 1-6
1.4 Acidic Deposition 1-6
1.5 Linkage to Atmospheric Sciences 1-7
1.6 Sensitivity 1-7
1.7 Presentation Limitations 1-7
E-2 EFFECTS ON SOIL SYSTEMS
2.1 Introduction 2-1
2.1.1 Importance of Soils to Aquatic Systems 2-1
2.1.1.1 Soils Buffer Precipitation Enroute to Aquatic Systems ... 2-2
2.1.1.2 Soil as a Source of Acidity for Aquatic Systems 2-2
2.1.2 Soil's Importance as a Medium for Plant Growth 2-2
2.1.3 Important Soil Properties 2-2
2.1.3.1 Soil Physical Properties 2-3
2.1.3.2 Soil Chemical Properties 2-3
2.1.3.3 Soil Microbiology 2-3
2.1.4 Flow of Deposited Materials Through Soil Systems 2-3
2.2 Chemistry of Add Soils 2-5
2.2.1 Development of Add Soils 2-5
2.2.1.1 Biological Sources of H+ Ions 2-6
2.2.1.2 Acidity from Dissolved Carbon Dioxide 2-6
2.2.1.3 Leaching of Basic Cations 2-7
2.2.2 Soil Cation Exchange Capacity ~. 2-8
2.2.2.1 Source of Cation Exchange Capacity 1n Soils 2-8
2.2.2.2 Exchangeable Bases and Base Saturation 2-8
2.2.3 Exchangeable and Solution Aluminum 1n Soils 2-9
2.2.4 Exchangeable and Solution Manganese In Soils 2-12
2.2.5 Practical Effects of Low pH 2-12
2.2.6 Neutralization of Soil Acidity 2-13
2.2.7 Measuring Soil pH 2-14
2.2.8 Sulfate Adsorption 2-15
2.2.9 Soil Chemistry Summary 2-18
2.3 Effects of Addle Deposition on Soil Chemistry and Plant Nutrition 2-18
2.3.1 Effects on Soil pH 2-19
2.3.2 Effects on Nutrient Supply of Cultivated Crops 2-24
2.3.3 Effects on Nutrient Supply to Forests 2-28
XX
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Table of Contents (continued)
Page
2.3.3.1 Effects on Cation Nutrient Status 2-28
2.3.3.2 Effects on S and N Status 2-31
2.3.3.3 Acidification Effects on Plant Nutrition 2-33
2.3.3.3.1 Nutrient deficiencies 2-33
2.3.3.3.2 Metal Ion toxlcltles 2-33
2.3.3.3.2.1 Aluminum toxlclty 2-34
2.3.3.3.2.2 Manganese toxlclty 2-35
2.3.4 Reversibility of Effects on Soil Chemistry 2-35
2.3.5 Predicting Which Soils will be Affected Most 2-36
2.3.5.1 Soils Under Cultivation 2-36
2.3.5.2 Uncultivated, Unamended Soils 2-36
2.3.5.2.1 Basic catlon-pH changes In forested soils .... 2-37
2.3.5.2.2 Changes 1n aluminum concentration 1n soil
solution In forested soils 2-40
2.4 Effects of Addle Deposition on Soil Biology 2-40
2.4.1 Soil Biology Components and Functional Significance 2-40
2.4.1.1 Soil Animals 2-40
2.4.1.2 Algae 2-40
2.4.1.3 Fungi 2-41
2.4.1.4 Bacteria 2-41
2.4.2 Direct Effects of Acidic Deposition on Soil Biology 2-42
2.4.2.1 Soil Animals 2-42
2.4.2.2 Terrestrial Algae 2-42
2.4.2.3 Fungi 2-43
2.4.2.4 Bacteria 2-43
2.4.2.5 General Biological Processes 2-44
2.4.3 Metals--Mob1l1zat1on Effects on Soil Biology 2-45
2.4.4 Effects of Changes In Mlcroblal Activity on Aquatic Systems 2-46
2.4.5 Soil Biology Summary 2-46
2.5 Effects of Acidic Deposition on Organic Matter Decomposition 2-47
2.6 Effects of Soils on the Chemistry of Aquatic Ecosystems 2-52
2.7 Conclusions 2-54
2.8 References 2-57
E-3 EFFECTS ON VEGETATION
3.1 Introduction 3-1
3.1.1 Overview 3-1
3.1.2 Background 3-1
3.2 Plant Response to Acidic Deposition 3-3
3.2.1 Leaf Response to Acidic Deposition 3-3
3.2.1.1 Leaf Structure and Functional Modifications 3-5
3.2.1.2 Foliar Leaching - Throughfall Chemistry 3-8
3.2.2 Effects of Acidic Deposition on Lichens and Mosses 3-13
3.2.3 Summary 3-16
3.3 Interactive Effects of Acidic Deposition with Other Environmental
Factors on Plants 3-17
3.3.1 Interactions with Other Pollutants 3-17
3.3.2 Interactions with Phytophagous Insects 3-20
3.3.3 Interactions with Pathogens 3-20
3.3.4 Influence on Vegetative Hosts That Would Alter Relationships
with Insect or Mlcroblal Associate 3-23
3.3.5 Effects .of Acidic Deposition on Pesticides 3-23
3.3.6 Summary 3-25
3.4 Blomass Production 3-26
3.4.1 Forests 3-26
3.4.1.1 Possible Mechanlslms of Response 3-27
3.4.1.2 Phonological Effects 3-29
3.4.1.2.1 Seed germination and seedling establishment .. 3-29
3.4.1.2.2 Mature and reproductive stages 3-32
3.4.1.3 Growth of Seedlings and Trees 1n Irrigation
Experiments 3-32
XXI
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Table of Contents (continued)
Page
3.4.1.4 Studies of Long-Term Growth of Forest Trees 3-33
3.4.1.5 Dleback and Decline 1n High Elevation Forests 3-36
3.4.1.6 Recent Observations on the German Forest Decline
Phenomenon 3-39
3.4.1.7 Summary 3-41
3.4.2 Crops 3-41
3.4.2.1 Review and Analysis of Experimental Design 3-42
3.4.2.1.1 Dose-response determination 3-42
3.4.2.1.2 Sensitivity classification 3-44
3.4.2.1.3 Mechanisms 3-44
3.4.2.1.4 Characteristics of precipitation simulant
exposures 3-45
3.4.2.1.5 Yield criteria 3-45
3.4.2.2 Experimental Results 3-46
3.4.2.2.1 Field studies 3-46
3.4.2.2.2 Controlled environment studies 3-50
3.4.2.3 Discussion 3-58
3.4.2.4 Summary 3-61
3.5 Conclusions 3-61
3.6 References 3-64
E-4 EFFECTS ON AQUATIC CHEMISTRY
4.1 Introduction 4-1
4.2 Basic Concepts Required to Understand the Effects of
Acidic Deposition on Aquatic Systems 4-2
4.2.1 Receiving Systems 4-2
4.2.2 pH, Conductivity, and Alkalinity 4-3
4.2.2.1 pH 4-3
4.2.2.2 Conductivity 4-4
4.2.2.3 Alkalinity 4-5
4.2.3 Acidification 4-6
4.3 Sensitivity of Aquatic Systems to Acidic Deposition 4-7
4.3.1 Atmospheric Inputs 4-7
4.3.1.1 Components of Deposition 4-7
4.3.1.2 Loading vs Concentration 4-8
4.3.1.3 Location of the Deposition 4-8
4.3.1.4 Temporal Distribution of Deposition 4-9
4.3.1.5 Importance of Atmospheric Inputs to Aquatic Systems 4-9
4.3.1.5.1 Nitrogen (N). phosphorus (P), and
carbon (C) 4-9
4.3.1.5.2 Sulfur 4-10
4.3.2 Characteristics of Receiving Systems Relative to Being Able to
Assimilate Acidic Deposition 4-13
4.3.2.1 Canopy 4-13
4.3.2.2 Soil 4-14
4.3.2.3 Bedrock 4-16
4.3.2.4 Hydrology 4-17
4.3.2.4.1 Flow paths 4-17
4.3.2.4.2 Residence times 4-22
4.3.2.5 Wetlands 4-23
4.3.2.6 Aquatic 4-24
4.3.2.6.1 Alkalinity as an Indicator of sensitivity .... 4-24
4.3.2.6.2 International production/consumption
of ANC 4-28
4.3.2.6.3 Aquatic sediments 4-31
4.3.3 Location of Sensitive Systems 4-32
4.3.4 Summary—Sensitivity 4-35
4.4 Magnitude of Chemical Effects of Acidic Deposition on
Aquatic Ecosystems 4-38
XX11
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Table of Contents (continued)
Page
4.4.1 Relative Importance of HN03 vs H2S04 4-39
4.4.2 Short-Term Acidification 4-45
4.4.3 Long-Term Acidification 4-48
4.4.3.1 Analysis of Trends based on Historic Measurements of
Surface Water Quality 4-53
4.4.3.1.1 Methodological problems with the evaluation
of historical trends 4-53
4.4.3.1.1.1 pH 4-54
4.4.3.1.1.1.1 pH-early method-
ology 4-54
4 .4.3.1.1.1.2 pH-current method-
ology 4-55
4.4.3.1.1.1.3 pH-comparab1l1ty
of early and cur-
rent measurement
methods 4-56
4.4.3.1.1.1.4 pH-general
problems 4-57
4.4.3.1.1.2 Conductivity 4-60
4.4.3.1.1.2.1 Conductivity
methodology 4-60
4.4.3.1.1.2.2 Conductivity-com-
parability of
early and current
measurement
methods 4-60
4.4.3.1.1.2.3 Conductivity-gen-
eral problems .... 4-61
4.4.3.1.1.3 Alkalinity 4-61
4.4.3.1.1.3.1 Alkalinity-early
methodology 4-61
4.4.3.1.1.3.2 Alkalinity-current
methodology 4-62
4.4.3.1.1.3.3 Alkalinity-compar-
ability of early
and current meas-
urement methods .. 4-63
4.4.3.1.1.4 Sample storage 4-63
4.4.3.1.1.5 Summary of measurement
techniques 4-63
4.4.3.1.2 Analysis of trends 4-64
4.4.3.1.2.1 Introduction 4-64
4.4.3.1.2.2 Canadian studies 4-66
4.4.3.1.2.3 United States studies 4-74
4.4.3.1.3 Summary—trends In historic data 4-98
4.4.3.2 Assessment of Trends Based on Paleollmnologlcal
Technique 4-99
4.4.3.2.1 Calibration and accuracy of paleollmnologlcal
reconstruction of pH history 4-100
4.4.3.2.2 Lake acidification determined by
paleollmnologlcal reconstruction 4-100
4.4.3.3 Alternate Explanations for Acidification-Land Use
Changes 4-105
4.4.3.3.1 Variations 1n the groundwater tabje 4-105
4.4.3.3.2 Accelerated mechanical weathering or
land scarification 4-105
4.4.3.3.3 Decomposition of organic matter 4-106
4.4.3.3.4 Changes 1n vegetation 4-106
4.4.3.3.5 Chemical amendments 4-107
4.4.3.3.6 Summary—alternate explanations for
acidification 4-107
XX 111
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Table of Contents (continued)
Page
4.4.4 Summary—Magnitude of Chemical Effects of Acidic Deposition 4-109
4.5 Predictive Modeling of the Effects of Acidic Deposition
on Surface Waters 4-113
4.5.1 Almer/D1ckson Relationship 4-114
4.5.2 Henriksen's Predictor Nomograph 4-119
4.5.3 Thompson's Cation Denudation Rate Model (CDR) 4-121
4.5.4 "Trickle-Down" Model 4-122
4.5.5 Summary of Predictive Modeling 4-125
4.6 Indirect Chemical Changes Associated with Acidification
of Surface Maters 4-128
4.6.1 Metal s 4-128
4.6.1.1 Increased Loading of Metals From Atmospheric
Deposition 4-129
4.6.1.2 Mobilization of Metals by Acidic Deposition 4-130
4.6.1.3 Secondary Effects of Metal Mobilization 4-131
4.6.1.4 Effects of Acidification on Aqueous Metal Spedatlon .... 4-132
4.6.1.5 Indirect Effects on Metals In Surface Haters 4-132
4.6.2 Aluminum Chemistry 1n Dilute Acidic Haters 4-132
4.6.2.1 Occurrence, Distribution, and Sources of Aluminum 4-132
4.6.2.2 Aluminum Spedat1 on 4-136
4.6.2.3 Aluminum as a pH Buffer 4-136
4.6.2.4 Temporal and Spatial Variations In Aqueous
Levels of Aluminum 4-137
4.6.2.5 The Role of Aluminum 1n Altering Element Cycling
H1th1n Acidic Haters 4-140
4.6.3 Organlcs 4-141
4.6.3.1 Atmospheric Loading of Strong Acids and Associated
Organic Mlcropollutants 4-141
4.6.3.2 Organic Buffering Systems 4-142
4.6.3.3 Organo-Metalllc Interactions 4-142
4.6.3.4 Photochemistry 4-143
4.6.3.5 Carbon-Phosphorus-Aluminum Interactions 4-143
4.6.3.6 Effects of Acidification on Organic Decomposition
In Aquatic Systems 4-144
4.7 M1t1gat1ve Strategies for Improvement of Surface Hater Quality 4-144
4.7.1 Base Additions 4-144
4.7.1.1 Types of Basic Materials 4-144
4.7.1.2 Direct Water Addition of Base 4-148
4.7.1.2.1 Computing base dose requirements 4-148
4.7.1.2.2 Methods of base application 4-152
4.7.1.3 Hatershed Addition of Base 4-154
4.7.1.3.1 The concept of watershed
application of base 4-154
4.7.1.3.2 Experience 1n watershed liming 4-156
4.7.1.4 Water Quality Response to Base Treatment 4-158
4.7.1.5 Cost Analysis, Conclusions and Assessment of Base
Addition 4-160
4.7.1.5.1 Cost analysis 4-160
4.7.1.5.2 Summary—base additions 4-162
4.7.2 Surface Water Fertilization 4-162
4.7.2.1 The Fertilization Concept 4-162
4.7.2.2 Phosphorous Cycling In Acidified Water 4-164
4.7.2.3 Fertilization Experience and Water
Quality Response to Fertilization 4-164
4.7.2.4 Summary—Surface Hater Fertilization 4-166
4.8 Conclusions 4-166
4.9 References 4-169
XXIV
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Table of Contents (continued)
Page
E-5 EFFECTS ON AQUATIC BIOLOGY
5.1 Introduction 5-1
5.2 Biota of Naturally Acidic Waters 5-3
5.2.1 Types of Naturally Acidic Waters 5-3
5.2.2 Biota of Inorganic Ad dotrophic Waters 5-4
5.2.3 Biota 1n Addle Brownwater Habitats 5-5
5.2.4 Biota In Ultra-OHgotrophlc Waters 5-7
5.2.5 Summary 5-9
5.3 Benthlc Organisms 5-14
5.3.1 Importance of the Benthlc Community 5-14
5.3.2 Effects of Acidification on Components of the Benthos 5-16
5.3.2.1 M1crob1al Community 5-16
5.3.2.2 PeMphyton 5-17
5.3.2.2.1 Field surveys 5-17
S.3.2.2.2 Temporal trends 5-18
5.3.2.2.3 Experimental studies 5-20
5.3.2.3 Mlcrolnvertebrates 5-21
5.3.2.4 Crustacea 5-22
5.3.2.5 Insecta 5-24
5.3.2.5.1 Sensitivity of different groups 5-24
5.3.2.5.2 Sensitivity of Insects from different
mlcrohabltats 5-29
5.3.2.5.3 Acid sensitivity of Insects based on food
sources 5-29
5.3.2.5.4 Mechanisms of effects and trophic
Interactions 5-29
5.3.2.6 Mollusca 5-30
5.3.2.7 Annelida 5-31
5.3.2.8 Summary of Effects of Acidification on Benthos 5-32
5.4 Macrophytes and Wetland PI ants 5-37
5.4.1 Introduction 5-37
5.4.2 Effects on Acidification on Aquatic Macrophytes 5-41
5.4.3 Summary 5-43
5.5 Plankton 5.44
5.5.1 Introduction 5-44
5.5.2 Effects of Acidification on Phytoplankton 5-45
5.5.2.1 Changes In Species Composition 5-45
5.5.2.2 Changes 1n Phytoplankton Blomass and Productivity 5-52
5.5.3 Effects of Acidification on Zooplankton 5-55
5.5.4 Explanations and Significance 5-67
5.5.4.1 Changes 1n Species Composition 5-67
5.5.4.2 Changes 1n Productivity 5-70
5.5.5 Summary 5-72
5.6 Fish 5-74
5.6.1 Introduction 5-74
5.6.2 Field Observations 5-75
5.6.2.1 Loss of Populations 5-75
5.6.2.1.1 United States 5-75
5.6.2.1.1.1 Adirondack Region of
New York State 5-75
5.6.2.1.1.2 Other regions of the eastern
United States 5-79
5.6.2.1.2 Canada 5-79
5.6.2.1.2.1 LaCloche Mountain Region of
Ontario 5-79
5.6.2.1.2.2 Other areas of Ontario 5-83
5.6.2.1.2.3 Nova Scotia 5-83
5.6.2.1.3 Scandinavia and Great Britain 5-88
5.6.2.1.3.1 Norway 5-88
5.6.2.1.3.2 Sweden 5-93
5.6.2.1.3.3 Scotland 5-93
XXV
-------
Table of Contents (continued)
Page
5.6.2.2 Population Structure 5-93
5.6.2.3 Growth 5-98
5.6.2.4 Episodic F1sh Kills 5-99
5.6.2.5 Accumulation of Metals In F1sh 5-101
5.6.3 Field Experiments 5-101
5.6.3.1 Experimental Acidification of Lake 223 Ontario 5-102
5.6.3.2 Experimental Acidification of Norrls
Brook, New Hampshire 5-104
5.6.3.3 Exposure of Fish to Acidic Surface Waters 5-104
5.6.4 Laboratory Experiments 5-108
5.6.4.1 Effects Of Low pH 5-109
5.6.4.1.1 Survival 5-109
5.6.4.1.2 Reproduction 5-112
5.6.4.1.3 Growth 5-119
5.6.4.1.4 Behavior 5-119
5.6.4.1.5 Physiological responses 5-120
5.6.4.2 Effects of Aluminum and Other Metals In Acidic Waters ... 5-122
5.6.5 Summary 5-125
5.6.5.1 Extent of Impact 5-125
5.6.5.2 Mechanism of Effect 5-127
5.6.5.3 Relationship Between Water Acidity and F1sh
Populat1on Response 5-128
5.7 Other Related Biota 5-129
5.7.1 Amphibians 5-129
5.7.2 Birds 5-134
5.7.2.1 Food Chain Alterations 5-134
5.7.2.2 Heavy Metal Accumulation 5-134
5.7.3 Mammals 5-135
5.7.4 Summary 5-136
5.8 Observed and Anticipated Ecosystem Effects 5-139
5.8.1 Ecosystem Structure 5-139
5.8.2 Ecosystem Function 5-141
5.8.2.1 Nutrient Cycling 5-141
5.8.2.2 Energy Cycling 5-141
5.8.3 Summary 5-142
5.9 Mitigative Options Relative to Biological Populations at Risk 5-143
5.9.1 Biological Response to Neutralization 5-143
5.9.2 Improving Fish Survival in Acidified Waters 5-145
5.9.2.1 Genetic Screening 5-145
5.9.2.2 Selective Breeding 5-145
5.9.2.3 Acclimation 5-146
5.9.2.4 Limitations of Techniques to Improve Fish Survival 5-148
5.9.3 Summary 5-149
5.10 Conclusions 5-149
5.10.1 Effects of Acidification on Aquatic Organisms 5-149
5.10.2 Processes and Mechanisms by Which Acidification
Alters Aquatic Ecosystems 5-155
5.10.2.1 Direct Effects of Hydrogen Ions 5-155
5.10.2.2 Elevated Metal Concentrations 5-156
5.10.2.3 Altered Trophic-Level Interactions 5-156
5.10.2.4 Altered Water Clarity 5-157
5.10.2.5 Altered Decomposition of Organic Matter 5-157
5.10.2.6 Presence of Algal Mats 5-157
5.10.2.7 Altered Nutrient Availability 5-157
5.10.2.8 Interaction of Stresses 5-157
5.10.3 Biological Mitigation 5-158
5.10.4 Summary 5-159
5.11 References 5-160
XXVI
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Table of Contents (continued)
Page
E-6 INDIRECT EFFECTS ON HEALTH
6.1 Introduction 6-1
6.2 Food Chain Dynamics 6-1
6.2.1 Introduction 6-1
6.2.2 Availability and B1oaccumulat1on of Toxic Metals 6-2
6.2.2.1 Speclatlon (Mercury) 6-2
6.2.2.2 Concentrations and Speclatlons In Water (Mercury) 6-4
6.2.2.3 Flow of Mercury 1n the Environment 6-4
6.2.2.3.1 Global cycles 6-4
6.2.2.3.2 Blogeochemlcal cycles of mercury 6-5
6.2.3 Accumulation In Fish 6-10
6.2.3.1 Factors Affecting Mercury Concentrations In Fish 6-10
6.2.3.2 Historical and Geographic Trends In Mercury Levels In
Freshwater Fish 6-20
6.2.4 Dynamics and Toxlclty In Humans (Mercury) 6-22
6.2.4.1 Dynamics In Man (Mercury) 6-22
6.2.4.2 Toxlclty 1n Man 6-23
6.2.4.3 Human Exposure from F1sh and Potential for Health
Risks 6-27
6.3 Ground, Surface and Cistern Waters as Affected by Acidic Deposition 6-31
6.3.1 Water Supplies 6-32
6.3.1.1 Direct Use of Precipitation (Cisterns) 6-32
6.3.1.2 Surface Water Supplies 6-34
6.3.1.3 Groundwater Supplies 6-37
6.3.2 Lead 6-39
6.3.2.1 Concentrations In Noncontamlnated Waters 6-39
6.3.2.2 Factors Affecting Lead Concentrations
In Water, Including Effects of pH 6-39
6.3.2.3 Speclatlon of Lead In Natural Water 6-41
6.3.2.4 Dynamics and Toxlclty of Lead In Humans 6-41
6.3.2.4.1 Dynamics of lead 1n humans 6-41
6.3.2.4.2 Toxic effects of lead on humans 6-42
6.3.2.4.3 Intake of lead In water and potential for
human health effects 6-49
6.3.3 Aluminum 6-51
6.3.3.1 Concentrations In Uncontamlnated Waters 6-53
6.3.3.2 Factors Affecting Aluminum Concentrations In Water 6-53
6.3.3.3 Speclatlon of Aluminum 1n Water 6-54
6.3.3.4 Dynamics and Toxlclty In Humans 6-54
6.3.3.4.1 Dynamics of aluminum In humans 6-54
6.3.3.4.2 Toxic effects of aluminum In humans 6-55
6.3.3.5 Human Health Risks from Aluminum In Water 6-55
6.4 Other Metals 6-55
6.5 Conclusions 6-56
6.6 References 6-58
E-7 EFFECTS ON MATERIALS
7.1 01 rect Effects on Material s 7-1
7.1.1 Introduction 7-1
7.1.1.1 Long Range and Local Effects 7-2
7.1.1.2 Inaccurate Claims of Acid Rain Damage to Materials 7-2
7.1.1.3 Complex Mechanisms of Exposure and Deposition 7-5
7.1.1.4 Deposition Velocities 7-6
7.1.1.5 Laboratory vs Field Studies 7-6
7.1.2 Damage to Materials by Acidic Deposition 7-8
7.1.2.1 Metals 7-9
7.1.2.1.1 Ferrous Metals 7-11
7.1.2.1.1.1 Laboratory Studies 7-13
7.1.2.1.1.2 Field Studies 7-14
XXV11
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Table of Contents (continued)
Page
7.1.2.1.2 Nonferrous Metal s 7-17
7.1.2.1.2.1 Aluminum 7-17
7.1.2.1.2.2 Copper 7-19
7.1.2.1.2.3 Zinc 7-19
7.1.2.2 Masonry 7-20
7.1.2.2.1 Stone 7-20
7.1.2.2.2 Concrete 7-26
7.1.2.2.3 Ceramics and Glass 7-27
7.1.2.3 Paint 7-27
7.1.2.4 Other Materials 7-31
7.1.2.4.1 Paper 7-32
7.1.2.4.2 Photographic Materials 7-32
7.1.2.4.3 Textiles and Textile Dyes 7-32
7.1.2.4.4 Leather 7-34
7.1.2.5 Cultural Property 7-34
7.1.2.5.1 Architectural Monuments 7-34
7.1.2.5.2 Museums, Libraries and Archives 7-34
7.1.2.5.3 Medieval Stained Glass 7-35
7.1.2.5.4 Conservation and Mitigation Costs 7-35
7.1.2.6 Economic Implications 7-37
7.1.2.6 MHIgatlve Measures 7-38
7.2 Potential Secondary Effects of Acidic Deposition on Potable Water
Piping Systems 7-39
7.2.1 Introduction 7-39
7.2.2 Problems Caused by Corrosion 7-40
7.2.2.1 Health 7-40
7.2.2.2 Aesthetics 7-40
7.2.2.3 Economics 7-40
7.2.3 Principles of Corrosion 7-40
7.2.4 Factors Affecting Internal Piping Corrosion 7-41
7.2.5 Corrosion of Materials Used In Plumbing and Water
D1strlbutlon Systems 7-47
7.2.5.1 Corrosion of Iron Pipe 7-47
7.2.5.2 Corrosion of Galvanized Pipe 7-49
7.2.5.3 Corrosion of Copper Pipe 7-49
7.2.5.4 Corrosion of Lead Pipe 7-50
7.2.5.5 Corrosion of Non-Metallic Pipe 7-50
7.2.6 Metal Leaching 7-50
7.2.6.1 Standing vs Running Samples 7-51
7.2.6.2 Metals Surveys 7-51
7.2.7 Corrosion Control Strategies 7-53
7.2.8 Economics 7-53
7.3 Conclusions 7-54
7.4 References 7-58
xxvm
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ABBREVIATION-ACRONYM LIST
6-ALA
ACHEX
ADI
Ag
AI
Al
A1203
Al2Si205(OH)4
AL
A1(OH)2H2P04
A1(OH)3
ANC
APN
ARL
ARS
As
ASTRAP
AWWA
B
BCF
BLM
BLMs
6-aminolevulinic acid
Aerosol Characterization Experiment
Acceptable daily intake
Silver
Aggresiveness index
Aluminum
Aluminum ion
Aluminum oxide
Aluminosilicate
Aeronomy Laboratory, NOAA
Varascite
Aluminum hydroxide
Acid neutralizing capacity
Air and Precipitation Monitoring Network
Air Resources Lab, NOAA
Agricultural Research Service, DOA
Arsenic
Advanced Statistical Trajectory Regional Air
Pollution Control Model
American Water Works Association
Boron
Bioconcentration factor
Bureau of Land Management, DOI
Boundary layer models
xxix
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BM
BNC
BNC aq
BOD
Br
BS
BSC
BUREC
BWCA
CB
Ca
CaCl
CaC03-MgC03
CaO
Ca(OH)2
CaS04
CAMP
CANSAP
CAPTEX
CCN
Cd
CDR
Bureau of Mines, DOI
Base neutralizing capacity
Aqueous base neutralizing capacity
Biologic oxygen demand
Bromine
Base saturation
Base saturation capacity
Bureau of Reclamation, DOI
Boundary Water Canoe Area
Base cation level
Calcium
Calcium ion
Calcium chloride
Calcium carbonate or crystalline calcite - limestone
Dolomite
Calcium bicarbonate
Calcium oxide - lime
Calcium hydroxide - lime
Calcium sulfate, sulfate salt
Syngenite
Continuous Air Monitoring Program
Canadian Network for Sampling Acid Precipitation
Cross-Appalachian Transport Experiment
Cloud condensation nuclei
Cadmium
Cation denudation rate
xxx
-------
CEC
CEQ
CH3Br
CH3C1
'CH3COOH
(CH3)2Hg
CH3Q
(CH3)2S
(CH3)2S2
CH3SH
CH4
cr
ci2
cm3 molecule"* s~
cm
cm s~l
cm yr~l
CO
C02
-COOH
COS
Cr
CS2
CSI
CSRS
Cu
Cation exchange capacity
Council on Environmental Quality
Methyl bromide
Methyl chloride
Acetic acid
Dimethyl mercury
Methoxy radical
Dimethyl sulfide (also CH3SCH3)
Dimethyl disulfide
Methyl sulfide (or methyl mercaptan)
Methane
Chloride ion
Elemental chlorine
Cubic centimeters per molecule per second
Centimeter
Centimeters per second
Centimeters per year
Carbon monoxide
Carbon dioxide
Carboxyl
Carbonyl sulfide
Chromium
Carbon disulfide
Calcite saturation index
Cooperative States Research Service, DOA
Copper
xxxi
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DEC
DPI
DO
DOA
DOC
DOD
DOE
DOI
DOS
ELA
emf
ENAMAP
EPA
EPRI
eq
eq ha"1 y1
ERDA
ESRL
F-
FA
FDA
FDA
Fe
FeS2
Fe2$i04
Department of Environmental Conservation, NY
Driving force index
Dissolved oxygen
Department of Agriculture
Dissolved organic carbon
Department of Defense
Department of Energy
Department of Interior
Department of State
Experimental Lakes Area
Electromotive force
Eastern North America Model of Air Pollutants
Environmental Protection Agency
Electric Power Research Institute
Equivalent
Equivalents per hectare per year
Energy Research and Development Agency (defunct)
Environmental Sciences Research Laboratory, EPA
Fluoride ion
Fulvic acid
Flourescein diacetate
Food and Drug Administration
Iron
Pyri te
01ivine (and
XXX11
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Ferrous sulfate
FEP Free erythrocyte protoporphyrin
FGD Flue gas desulfurization
FS Forest Service, DOA
FWS Fish and Wildlife Service, DOI
g Gram
g r1 Grams per liter
g dry wt m~2 Grams dry weight per square meter
g m~2 Grams per square meter
g m-2 s-l Grams per square meter per second
g m-2 yr-l Grams per square meter per year
g ha"1 hr~* Grams per hectare per hour
GAMETAG Global Atmospheric Measurement Experiment of
Tropospheric Aerosols and Gases
GTN Global Trends Network
H Hydrogen
H+ Hydrogen 1on
H2C03 Carbonic acid
H202 Hydrogen peroxide
H2o Water
H2S Hydrogen sulfide
H2S04 Sulfuric add
H3P04 Phosphoric acid
ha Hectare
HAOS Houston Area Oxidant Study
HC Hydrocarbon
xxxm
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HC1
HC03"
HCOH
HCOOH
HF
Hg
HIVOL
HgCl2
HgS
HHS
HN02
HN03
H02
H02N02
HO
HONO
HOS02
hr
ILWAS
IRMA
K
K+
KC1
K2S04
keq ha'1
Hydrochloric acid
Bicarbonate ion
Formaldehyde
Formic acid
Hydrogen fluoride
Mercury
High-volume
Mercuric ion
Mercuric chloride
Mercuric sulfide
Department of Health and Human Services
Nitrous acid
Nitric acid
Peroxy radical
Pernitric acid
Hydroxyl
Nitrous acid
Bisulfite
Hours
Integrated Lake Watershed Acidification Study
Immission rate measuring apparatus
Potassium
Potassium ion
Potassium chloride
Potassium sulfate, sulfate salt
Klloequivalents per hectare
XXXIV
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keq ha-1 yr-l Kiloequivalents per hectare per year
kg Kilogram
kg ha-1 Kilograms per hectare
kg ha-1 wk~l Kilograms per hectare per week
kg km-2 yr-1 Kilograms per square kilometer per year
kg ha-1 yr-1 Kilograms per hectare per year
KHM Kol-Halsa-Miljo Project
KJ mol-1 Kilojoule per mole
km Kilometer
km2 Square kilometer
km hr-1 Kilometers per hour
KMn04 Potassium permanganate
£ Liter
(£) Liquid phase
Si m-3 Liters per cubic meter
LAI Leaf area index
LI Langelier's index
LIMB Limestone Injection/Multistage Burner
LR Larson's ration
LRTAP Long-Range Transport of Air Pollutants
LSI Langelier Saturation Index
m2 Square meter
m3 yr-1 Cubic meter per year
peq Microequivalent
yeq £-1 Microequivalents per liter
xxxv
-------
yg
,-1
yg 100 ml-1
yg dl-1
yg nr3
urn
urn £-1
uM
ym yr-1
umho cm~l
m
M
m s-1
m yr~l
MAP3S
mb
MCC
MCL
MCPS
ME
meq jr1
meq 100 g-1
roeq m~2 yr~^
METROMEX
Mg
Micrograms
Micrograms per liter
Micrograms per 100 mill litters
Micrograms per decaliter
Micrograms per cubic meter
Micrometer
Micrometers per liter
Micromolar
Micrometers per year
micromhos per centimeter (conductivity)
Meter
Molar
Meters per second
Meters per year
Multi-State Atmospheric Power Production
Pollution Study
Millibars
t
Mesoscale convective complex
Maximum contaminant level
Mesoscale convective precipitation systems
Momentary excess
Milliequlvalents per liter
Mil11 equivalents per 100 grams
Mi 111 equivalents per square meter per year
Metropolitan Meteorological Experiment
Magnesium
xxxvi
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M92+ Magnesium 1on
mg Milligram
rag I'* Milligrams per liter
mg nr3 hr'1 Milligrams per cubic meter per hour
MgC(>3 Magnesium carbonate
Mg2$104 Oil vine and (F62S104)
M9$04 Magnesium sulfate, sulfate salt
mho cm'1 mhos per centimeter (conductivity)
MISTT Midwest Interstate Sulfur Transport and
Transformations
mm Millimeter
mm hr'1 Millimeters per hour
m S'1 Millimeters per second
mm yr'1 Millimeters per year
mM Millimolar
Mn Manganese
Mo Molybdenum
MOI Memorandum of Intent on Transboundary Air Pollution
mol Mole
mol £-1 Moles per liter
mol £-1 atm"1 Moles per liter per atmosphere
mT Metric ton
mT y"1 Metric tons per year
MW Megawatt
N204 N02 dimer
N20$ Nitrogen pentoxide
xxxvn
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N20
(-NH)
N
N(III)
Na
Na+
Nad
NaN02
Na2S04
NADP
NAS
NASA
NASN
NATO
NBS
NCAC
NCAR
NECRMP
NEDS
ng £-1
ng kg"1
ng m-3
NH3
NH4+
Nitrous oxide
Imide
Nitrogen
Liquid phase nitrogen
Sodium
Sodium ion
Sodium chloride
Sodium carbonate
Sodium nitrite
Sodium sulfate, sulfate salt
National Atmospheric Deposition Program
National Academy of Sciences
National Aeronautics and Space Administration
National Air Sampling Network
North Atlantic Treaty Organization
National Bureau of Standards, DOC
National Conservation Advisory Council
National Center for Atmospheric Research
Northeast Corridor Regional Modeling Program
National Emissions Data System
Nanograms per liter
Nanograms per kilogram
Nanograms per cubic meter
Ammonia
Ammonium ion
xxxvm
-------
NH4C1
NH40AC
(NH4)2HP04
(NH4)2S04
NH4OH
Ni
nm
NMAB
N02
N03'
NO
NOX
NOAA
NFS
NRCC
NSF
NSPS
NTN
NWS
0
°2
03
(-OH)
Ammonium chloride
Ammonium acetate
Letorlclte
Ammonium phosphate
Ammonium nitrate
Ammonium sulfate
Ammonium hydroxide
Nickel
Nanometer
National Materials Advisory Board
Nitrogen dioxide
Nitrate 1on
Nitric oxide
Nitric oxides
National Oceanic and Atmospheric Administration
National Park Service, DOI
National Research Council Canada
National Science Foundation
New Source Performance Standards
National Trends Network
National Weather Service, NOAA
Oxygen
Elemental oxygen
Ozone
Phenol
xxxix
-------
°ECD Organization for Economic Cooperation and
Devel opment
OH Hydroxyl
OMB Office of Management and Budget
ORNL Oak Ridge National Laboratory
OSM Office of Surface Mining, DOI
P Phosphorus
PAH Polycyclic aromatic hydrocarbons
PAN Peroxyacetyl nitrate
Pb Lead
Pb2+ Lead ion
PBCF Practical bioconcentration factor
PBL Planetary boundary layer
P6S04 Lead sulfate
PCB Polychlorinated biphenyl
PGF Pressure gradient force
PHS Public Health Service
P043" Phosphate ion
ppb Parts per billion
ppm Parts per million
RAM St. Louis Regional Air Modeling Study
RAPS St. Louis Regional Air Pollution Study
RI Ryznar index
RSN Research Support Network
s Second
S cm-1 Seconds per centimeter
xl
-------
5
Sulfur
$2- Sulfide
S(IV) Gas-ph?se sulfur, an oxidation state
SAC Sulfate adsorption capacity
SAES State Agricultural Experiment Station, DOA
Sb Antimony
SCS Soil Conservation Service, DOA
Se Selenium
Si Silicon
Si02 Silicon dioxide
SMA Swedish Ministry of Agriculture
S02 Sulfur dioxide
S032- Sulfite
SQ42- Sulfate ion
STP Standard temperature and pressure
SURE Sulfate Regional Experiment, EPRI
IDS Total dissolved solids
TFE Total fixed endpoint alkalinity
Tg Teragram (10*2 gram)
Tg yr-1 Teragrams per year
TIC Total inorganic carbon
TIP Total inflection point alkalinity
TPS Tennessee Plume Study
TSP Total suspended particulates
xli
-------
TVA Tennessee Valley Authority
USGS United States Geological Survey, DOI
V Vanadium
V20s Vanadium pentoxide
V cm-1 Volts per centimeter
VDI Verein Deutcher Ingenieure
VOC Volatile organic compounds
WHO World Health Organization
WHO World Meteorological Organization
yr Year
Zn Zinc
ZnS Zinc sulfide
xlii
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GLOSSARY
Acceptable dally Intake (ADI) - rate of safe consumption of a particular
substance or element 1n human food or water, as determined by the U.S. Food
and Drug Administration.
Acidic deposition - the deposition of acidic and acidifying substances from
the atmosphere.
Acid neutralizing capacity (ANC) - equivalent sum of all bases that can be
titrated with a strong acid; also known as alkalinity.
Adlabatlc - occurring without gain or loss of heat by the substance
concerned.
Adsorption - adhesion of a thin layer of molecules to a liquid or solid
surface.
Advectlon - horizontal flow of air to the surface or aloft; one of the means
by which heat Is transferred from one region of the Earth to another.
Aerosols - suspensions of liquid or solid particles In gases.
All quoting - dividing Into equal parts.
Alkalinity - measure of the ability of an aqueous solution to neutralize acid
(also known as acid neutralizing capacity or ANC).
Allochthonous inputs - substances introduced from outside a system.
Ambient - the surrounding outdoor atmosphere to which the general population
may be exposed.
Ammonium - cation (NH4+) or radical (Nfy) derived from ammonia by
combination with hydrogen. Present in rainwater, soils, and many commercial
fertilizers.
Anlon - a negatively charged ion.
Aqueous phase - that part of a chemical transformation process when
substances are mixed with water or water vapor in the atmosphere.
Antagonistic effects (less-than-additive) - results from joint actions of
agents so that their combined effect is less than the algebraic sum of their
individual effects.
Anthropogenic - manmade or related to to human activities.
Artifact - a spurious measurement produced by the sampling or analysis
process.
-------
Atmospheric residence time - the amount of time pollutant emissions are held
In the atmosphere.
Autochthonous Inputs - Indigenous, formed or originating within the system.
Autotrophic - able to synthesize nutritive substances from inorganic
compounds.
Background measurement - pollutants in ambient air due to natural sources;
usually taken in remote areas.
Base neutralizing capacity - equivalent sum of all acids that can be titrated
with a strong base.
Base saturation (BS) - the fraction of the cation exchange capacity satisfied
by basic cations.
Benthic organisms - life forms living on the bottoms of bodies of water.
Bioaccumulation - the phenomenon wherein toxic elements are progressively
amassed in greater quantities as individuals farther up the food chain ingest
matter containing those elements.
Bioconcentration factor (BCF) - the ratio of the concentration of a substance
in an organism to the concentration of the substance in the surrounding
habitat.
Bioindicators - species of plants or animals particularly sensitive to
specific pollutants or adverse conditions.
Biomass - that part of a given habitat consisting of living matter.
Biosphere - the part of the Earth's crust, waters, and atmosphere where
living organisms can subsist.
Brownian diffusion - spread by random movement of particles suspended in
liquid or gas, resulting from the impact of molecules of the fluid
surrounding the particles.
Brownwater lakes and streams - acidic waters associated with peatlands,
cypress swamps; acidity is caused by organic acids leached from decayed plant
material and from hydrogen ions released by plants such as Sphagnum mosses.
Budget - a summation of the inputs and outputs of chemical substances
relative to a given biological or physical system.
Buffer - a substance in solution capable of neutralizing both acids and bases
and thereby maintaining the original pH of the solution.
Buffering capacity - ability of a body of water and its watershed to
neutralize introduced acid.
xliv
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Bulk sampling - method for collecting deposition that does not separate dry
and wet deposition (see Chapter A-8).
Calcareous - resembling or consisting of calcium carbonate (lime), or growing
on limestone or lime-containing soils.
Caldte saturation Index (CSI) - measure of the degree of saturation of water
with respect to CaCOa, integrating alkalinity, pH, and Ca concentration.
Cation - a positively changed 1on
Cation exchange capacity (CEC) - the sum of the exchangeable cations,
expressed in chemical equivalents, in a given quantity of soil.
Chemoautotrophic - having the ability to synthesize nutritive substances
using an inorganic compound as a source of available energy.
Colorlmetric - a chemical analysis method relying on measurement of the
degree of color produced in a solution by reaction of the compound of
interest with an indicator.
Conductivity - the ability to conduct an electric current; this is a function
of the individual mobilities of the dissolved ions in a solution, the concen-
trations of the ions, and the solution temperature; measured in mho cirri.
Continental scale - measurement of atmospheric conditions over an area the
size of a continent.
Coriolis effect - an effect caused by the Earth's eastward rotation in which
the speed of the movement falls off as the circumference of the Earth gets
progressively smaller at higher latitudes; this results in the movement of
winds, and subsequently ocean currents, to the right in the northern
hemisphere and to the left in the southern hemisphere.
Cosmic ray - a stream of ionizing radiation of extraterrestrial origin,
chiefly of protons, alpha particles, and other atomic nuclei but including
some high energy electrons and protons, that enters the atmosphere and
produces secondary radiation.
Coulomb - a meter/kilogram/second unit of electric charge equal to the
quantity of charge transferred in one second by a steady current of one
ampere.
Coarse particles - airborne particles larger than 2 to 3 micrometers 1n
diameter.
Cultivar - cultivated species of crop plant produced from parents belonging
to different species or different strains of the same species, originating
and persisting under cultivation.
Cuticular resistance - the resistance to penetration of a leaf cuticle.
xlv
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Cyclone track - the path of a low pressure system.
Denitrification - a bacterial process occurring in soils, or water, in which
nitrate is used as the terminal electron acceptor and is reduced primarily to
N2. It is essentially an anaerobic process; it can occur in the presence
of low levels of oxygen only if the microorganisms are metabolizing in an
anoxic microzone (an oxygen-free microenvironment within an area of low
oxygen levels).
Deposition velocity - rate at which particles from the atmosphere contact
surfaces and adhere.
Detritus - loose material resulting directly from disintegration.
Diffusiophoresis - an effect created when particles approaching an
evaporating surface are impacted by more molecules on the side nearer the
surface.
Dissolved organic carbon (DOC) - the amount of organic carbon in an aqueous
solution.
Dissolved inorganic carbon (DIC) - the amount of inorganic carbon in an
aqueous solution.
Dose - the quantity of a substance to be taken all at one time or in
fractional amounts within a given period; also the total amount of a
pollutant delivered or concentration.
Dose-response curve - a curve on a graph based on responses occurring in a
system as a result of a series of stimuli intensities or doses.
Edaphic differences - soil differences.
Eddies - currents of water or air running contrary to the main current.
Eddy diffusities - dispersive movements of particles, caused by circular
motions in air currents.
Ekman layer - a layer of the atmosphere typically extending between 1 and 3
kilometers above the surface; see Section A-3.2.2 for detailed discussion.
Electromotive force (emf) - the amount of energy derived from an electrical
source per unit quantity of electricity passing through the source (as a cell
or generator).
Entrainment - the process of carrying along or over (as in distillation or
evaporation).
Epifaunal - organism living on an animal.
Epilimnion - the upper layer of a lake in which the water temperature is
essentially uniform.
xlvi
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Episodic precipitation event - a period during which rain, snow, etc., is
occurring.
Ericaceous - heathlike or shrubby; a member of the Ericaceae family.
Eucaryotlc algae - algae composed of one or more cells with visibly evident
nuclei.
Eulerian models - models with reference frames fixed on the source or at the
surface.
Eurytopic - having a wide range of tolerance to variation of one or more
environmental factors.
Eutrophic - relating to or being in a well nourished condition; a lake rich
in dissolved nutrients but frequently shallow and with seasonal oxygen
deficiency in the hypolimnion.
Eutrophication - the process of becoming more eutrophic either as a natural
phase in the maturation of a body of water or artificially, as by
fertilization.
Exposure level - concentration of a contaminant with which an individual or
population is in contact.
Extinction coefficient - a measure of the space rate of diminution, or
extinction of any transmitted light; thus, it is the attenuation coefficient
applied to visible radiation.
Fine particles - airborne particles smaller than 2 to 3 micrometers in
diameter.
Fly ash - fine, solid particles of noncombustible ash carried out of a bed of
solid fuel by a draft.
Foliar - referring to plant foliage (leaves).
Fumigate - to subject to smoke or fumes.
Gas-phase mechanism - a process occurring when pollutants are in a gaseous
state, as opposed to being combined with moisture.
Geostrophic - of or pertaining to the force caused by the Earth's rotation.
Global scale - measurement of atmospheric conditions on a world-wide basis.
Ground loss - the effect of deposition of pollutant from atmposhere to
Earth's surface.
Ground sink - the Earth's surface, where airborne substances may be
deposited.
xlvii
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Haze - an aerosol that Impedes vision and may consist of a combination of
water droplets, pollutants, and dust.
Hemispheric scale - measurements of activity covering half of the Earth.
Heterotrophic - obtaining nourishment from outside sources, requiring complex
organic compounds of nitrogen and carbon for metabolic synthesis.
Humic acid - any of various organic acids that are insoluble in alcohol and
organic solvents and that are obtained from humus.
Hydrocarbons - a vast family of compounds containing carbon and hydrogen in
various combinations; found especially in fossil fuels.
Hydrologic residence time - the amount of time water takes to pass from the
surface through soil to a lake or stream.
Hydrometeor - a product of the condensation of atmospheric water vapor (e.g.,
raindrop).
Hydrophilic - of, relating to, or having a strong affinity for water; readily
wet by water.
Hydrophobic particles - particles resistant to or avoiding wetting; of,
relating to, or having a lack of affinity for water.
Hydroxyl radical - chemical prefix indicating the [OH] group.
Hygroscopic particles - absorbing moisture readily from the atmosphere.
Hypolimnion - the lowermost region of a lake, below the thermocline, in which
the temperature from its upper limit to the bottom is nearly uniform.
Hysteresis - the failure of a property to return to its orginal condition
after the removal of the causal external agent (i.e., irreversibility).
Infauna - population of organisms living in sediments.
Inorganic acidotrophic lakes - waters associated with geothermal areas or
lignite burns; extremely acidic, often heated, and frequently containing
elevated metal concentrations.
Interstitial water - water in the space between cells.
Isopleth - 1. a line of equal or constant value of a given quantity with
respect to either space or time, also known as an isogram; 2. a line drawn
through points on a graph at which a given quantity has the same numerical
value as a function of the two coordinate variables.
Labile - readily or continually undergoing chemical or physical or biological
change or breakdown.
xlviii
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Lacustrine sediments - deposits formed in lakes.
Lagrangian models - models with reference frames fixed on the puff cf
pollutants.
Langmuir equations - empirical derivations from kinetic treatment of the
physical adsorption of gases or solids by soils; relating to the relative
adsorption capacity of a soil for a specific anion.
Leaf area index (LAI) - ratio of the total foliar surface area to the ground
surface area that supports it.
Lentic - of, relating to, or living in still waters.
Lidar - a laser-radar system operated from a mobile van.
Ligands - those molecules or anions attached to the central atom in a
complex.
Limnological - of or relating to the scientific study of physical, chemical,
meteorological, and biological conditions in freshwaters, especially ponds
and lakes.
Lipophilicity - the strong affinity for fats or other lipids.
Liquid-phase mechanism - a process occurring when pollutants are combined
with moisture, as opposed to being in a purely gaseous state.
Littoral - the shore zone between high and low watermarks.
Loading rate - the amount of a nutrient available to a unit area or body of
water over a given period.
Long-range transport - conveyance of pollutants over extensive distances,
commonly referring to transport over synoptic and hemispheric scales.
Macrophytes - higher plants.
Manometer - an instrument for measuring pressure of gases or work.
Mean (arithmetic) - the sum of observations divided by sample size.
Median - a value in a collection of data values which is exceeded in
magnitude by one-half the entries in the collection.
Mesoscale - of or relating to meteorological phenomena from 1 to 100
kilometers in horizontal extent.
Metalimnion - the thermocline.
Microbial pathogens - microscopic organisms capable of producing disease,
such as viruses, fungi, etc.
xlix
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Microflora - a small or strictly localized plant.
Micrometeorological - referring to conditions specific to a very small area,
such as a surface, a particular site, or locale.
Mist - suspension of liquid droplets formed by condensation of vapor or
atomization; the droplet diameters exceed 10 micrometers and in general the
concentration of particles is not high enough to obscure visibility.
Mixing layer - also called the planetary boundary layer (PBL); usually the
domain of microscale turbulance.
Mobile sources - automobiles, trucks, and other pollution sources that are
not fixed in one location.
Mole - The mass, in grams, numerically equal to the molecular weight of a
substance.
Morphology - structure and form of an organism at any stage of its life
history.
Mycorrhizal - relating to symbiotic association of a fungal mycelium with the
roots of a seed plant.
Nitrification - the principal natural source of nitrate, in which ammonium
(NH4+) ions are oxidized to nitrates by specialized microorganisms.
Other organisms oxidize nitrites to nitrates.
Nocturnal jet - phenomenon in the atmosphere of a high-velocity air stream
occuring at night above the nocturnal inversion layer.
Non-humic lakes - lakes without significant inputs of humic acid.
Ohm's law - a law in electricity: the strength or intensity of an unvarying
electrical current is directly proportional to the electromotive force and
inversely proportional to the resistance of the circuit.
Oligochaete worms - an annelid worm of the class Oligochaeta, i.e., having a
segmented body.
Oligotrophic - a body of water deficient in plant nutrients; also generally
having abundant dissolved oxygen and no marked stratification.
Ombrotrophic peat bog - a peat bog fed solely by rain water.
Oxic condition - the presence of oxygen.
Oxidant - a chemical compound that has the ability to remove electrons from
another chemical species, thereby oxidizing it; also a substance containing
oxygen which reacts in air to produce a new substance, or one formed by the
action of sunlight on oxides of nitrogen and hydrocarbons.
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Palearctic lake - a lake in the biogeographic region that includes Europe,
Asia north of the Himalayas, northern Arabia, and Africa north of the Sahara.
Particle morphology - the structure and form of substances
suspended in a medium.
Particulates - fine liquid or solid particles such as dust, smoke, mist,
fumes, or smog found in air or in emissions.
Ped surfaces - surfaces of natural soil aggregates.
Pelagic - of, relating to, or living in the open sea.
Periphyton - organisms that live attached to underwater surfaces.
Photoautotrophic organisms - autotrophic organisms able to use light as an
energy source.
Photochemical oxidants - primarily ozone, N02, PAN with lesser amounts of
other compounds formed as products of atmospheric reactions involving organic
pollutants, nitrogen oxides, oxygen, and sunlight.
Phytophagous insects - insects feeding on plants.
Phytoplankton - autotrophic, free-floating, mostly microscopic organisms.
Planetary boundary layer (PBL) - first layer of the atmosphere extending
hundreds of meters from the Earth's surface to the geostrophic wind level,
including, therefore, the surface boundary layer and the Ekman layer; above
this level lies the free atmosphere.
Plume - emission from a flue or chimney, normally distributed streamlike
downwind of the source, and which can be distinguished from surrounding air
by appearance or chemical characteristics.
Plume touchdown - point of a plume's contact with the Earth's surface.
Podzol - any of a group of zonal soils that develop in a moist climate,
especially under coniferous or mixed forests.
Point source - a single stationary location for pollutant discharge.
Precipitation scavenging - a complex process composed of four distinct but
interactive steps: intermixing of pollutant and condensed water within the
same airspace, attachment of pollutant to the condensed water, chemical
reaction of pollutant within the aqueous phase, and delivery of
pollutant-laden water to surfaces.
Precursor - a substance from which another substance is formed, specifically
one of the anthropogenic or natural emissions or atmospheric constituents
that reacts under sunlight to form secondary pollutants comprising
photochemical smog.
li
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Primary particles (or primary aerosols) - dispersion aerosols formed from
particles emitted directly into the air that do not change form in the
atmosphere.
Quasi-laminar layer - the internal viscous boundary layer above non-ideal or
natural surfaces; it is frequently neither laminar nor constant with time.
Rayleigh scattering - spread of electromagnetic radiation by bodies much
smaller than the wavelength of the radiation; for visible wavelengths, the
molecules constituting the atmosphere cause Rayleigh scattering.
Secondary particles (or secondary aerosols) - dispersion aerosols that form
in the atmosphere as a result of chemical reactions, often involving gases.
Sensitivity - the degree to which an ecosystem or organism may be affected by
inputs or stimuli.
Sequential sampling - repeated, periodic collection of data concerning a
phenomenon of interest.
Sinks - reactants with or absorbers of substances; collection surfaces or
areas where substances are gathered.
Steady state exposure - exposure to air pollutants whose concentration
remains constant for a period of time.
Stefan flow - results from injection into a gaseous medium of new gas
molecules at an evaporating or subliming surface; Stefan flow is capable of
modifying surface deposition rates by an amount that is larger than the
deposition velocity appropriate for many small particles to aerodynamically
smooth surfaces.
Stokes's law - a law in physics: the force required to move a sphere through
a given viscous fluid at a low uniform velocity is directly proportional to
the velocity and radius of the sphere.
Stoma - opening on a leaf surface through which water vapor and other gases
diffuse; often term applies to the entire stomatal apparatus including
surrounding specialized epidermal cells, guard cells.
Stream order - positions a stream in relation to tributaries, drainage area,
total length, and age of water. First-order streams are the terminal twigs
(headwaters or youngest segments of a stream system, having no tributaries).
Second-order streams are formed by the junction of two first order streams,
and so on. At least two streams of any given order are required to form the
next highest order.
Sub-optical range - particles too small to be seen with the naked eye.
Surfactant - a substance capable of altering the physiochemical nature of
surfaces, such as one used to reduce surface tension in a liquid.
lii
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Symbiotic - a close association between two organisms of different species,
1n which at least one of the two benefits.
Synergistic effects (more-than-additive) - result from joint actions of
agents so that their combined effect is greater than the algebraic sum of
their Individual effects.
Synoptic scale - relating to or displaying atmospheric and weather conditions
as they exist simultaneously over a broad area; the scale of weather maps.
Teragram (Tg) - one million metric tons, 1012 grams.
Thermocline - the stratum of a lake below the epillmnion in which there is a
large drop 1n temperature per unit depth.
Thermophoresis - a force near a hot surface that drives small particles away
from that surface.
Throughfall - precipitation falling through the canopy of a forest and
reaching the forest floor.
Titration - the process or method of determining the concentration of a
substance in solution by adding to it a standard reagent of a known
concentration in carefully measured amounts until a reaction of definite and
known proportion is completed, as shown by a color change or by electrical
measurement, and then calculating the unknown concentration.
Total fixed endpoint alkalinity (TFE) - a measure of acid neutralizing
capacity involving acidimetric titrations performed to an endpoint of pH 4.5
determined electrometrically or to an endpoint determined by either a
colorimetric indicator or mixed indicators.
Total inflection point (TIP) - a measure of acid neutralizing capacity,
Involving acidimetric titration to the HC03-H+ equivalence point of the
titration curve.
Total suspended particulates (TSP) - solid and liquid particles present in
the atmosphere.
Toxicity - the quality, state, or relative degree of being poisonous.
Trajectory - a path, progression, or line of development, as from a plume of
pollutant carried through the atmosphere from a source to a receptor area.
Transport layer - the layer between the earth's surface and the peak mixing
height of the day; for any given instant, it 1s made up of the current mixing
layer below and the relatively quiescent layer above.
Troposphere - that portion of the atmosphere in which temperature decreases
rapidly with altitude, clouds form, and mixing of air masses by convection
takes place; generally extending to about 11 to 17 km above the Earth's
surface.
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Ultra ollgotrophlc lakes - lakes 1n areas where glaciatlon has removed
calcareous deposits and exposed weather resistant granitic and siliceous
bedrock; such lakes have little carbonate-bicarbonate buffering capacity and
are very vulnerable to pH changes; they tend to be small and have low
concentrations of dissolved Ions.
Variance - a measure of dispersion or variation of a sample from Its expected
value.
Washout - the capture of gases and particles by falling raindrops.
Wet deposition - the combined processes by which atmospheric substances are
returned to Earth In the form of rain or other precipitation.
Wind shear - a sudden shift In wind direction.
X-ray diffraction - technique by which patterns of diffraction can be used to
Identify a substance by Its structure.
Zooplankton - minute animal life floating or swimming weakly 1n a body of
water.
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS
E-l. INTRODUCTION
(R. A. Linthurst)
1.1 OBJECTIVES
The basic and applied scientific knowledge that can be gained through the
study of the acidic deposition phenomenon will undoubtedly advance our under-
standing of emissions, transport, scavenging, and deposition interactions.
This knowledge is essential for a more complete understanding of the causes
of acidic deposition and for defining the loadings of acidic and acidifying
substances that ultimately interact with the ecosystem. However, it is the
perception that acidic deposition may be harming our natural and managed
environment that has stimulated world-wide interest. As a result, the
effects and/or the potential effects of acidic deposition are the primary
motivation for public concern and research activities now designed to learn
more about this phenomenon.
The objectives of the effects portion of this document are to define the
logic behind the concerns of potential effects, present the support, or lack
of support, for these concerns and draw conclusions relative to the effects
of acidic deposition based on the best available evidence. Special attention
is given to quantitative information on the magnitude and extent of effects.
However, it will become evident that placing statistical confidence limits on
the data presently available is difficult, and in most instances, impossible.
A lack of quantitative cause and effect data, in itself, defines the state of
knowledge in many of the research areas.
1.2 APPROACH
An ecosystem approach to the acidic deposition effects issues has been used.
Figure 1-1 diagramrnatically presents a conceptual flow of wet and dry depo-
sition through a forested system. As most of the terrestrial landscape is
covered by vegetation, most acidic inputs to a system pass through the canopy
or down the stems of plants, to the soil, and finally, over or through the
soil to aquatic systems, lakes and/or streams, or into the groundwater
system. At any point along this pathway, the chemistry of precipitation can
be significantly altered. As a result, attempts to quantify effects in
relation to a chemical dose become increasingly complex and difficult.
Direct deposition of acidic and acidifying substances to soils and aquatic
systems also occurs. The size of the receiving system of interest, in
relation to the size of any other ecosystem component which may alter the
deposition chemistry prior to contact, becomes important. A common example
of this concept is lake and watershed interactions. Small lakes surrounded
1-1
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INPUTS
GASEOUS
OUTPUT
DRYFALL WETFALL
LEACHING
(biological export)
/ STREAMFLOW
GEOCHEMICAL EXPORT
Figure 1-1. Conceptual diagram of wet and dry deposition pathways in
an ecosystem context (from Johnson et al. 1982. The effects
of acid rain in forest nutrient status. Water Res.
18(3):449-461)
1-2
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by large watersheds are more greatly Influenced by those waters which pass
through the terrestrial landscape prior to entering the lake; most of the
water received is from the terrestrial pathway. Thus, the effect of the
terrestrial system on precipitation/deposition chemistry becomes a variable
which ultimately defines the chemistry of the water entering the aquatic
systems via this path. If a lake is large in relation to the area it drains,
direct deposition to the lake surface becomes increasingly important and the
terrestrial component of the system plays a less important role.
Having defined a representative flow path through a system from a chemical
perspective, one must recognize that any part of the system which alters the
chemistry of precipitation can be affected. Thus, the vegetation, the soil,
and the waters may be altered by incoming wet and dry deposition. In addi-
tion to these direct alterations of the system components, indirect effects
can also occur. Soils, for example, if chemically altered, ultimately affect
vegetation responses. If water chemistry is affected, the biota in those
waters are then subject to change. Subsequently, these changes can be of
significance to human health since both vegetation and aquatic organisms are
part of the human food chain.
This ecosystem perspective, with all its complexities and linkages, should be
kept in mind throughout the reading of the chapters. The concept of acidic
deposition effects can be understood fully only with this perspective in
mind. However, for convenience of presentation, each major ecosystem com-
ponent has been somewhat artificially separated from the others and subse-
quently discussed in partial isolation from the holistic approach.
1.3 CHAPTER ORGANIZATION AND GENERAL CONTENT
Because soils affect both vegetation and water, the effects of acidic depo-
sition on soils are discussed first. Secondly, vegetation effects are
evaluated from a more direct influence perspective, capitalizing on the
knowledge of soils/nutrient cycling, i.e., the potential indirect effects.
Next, the water chemistry component of the system is reviewed from a water-
shed perspective, continuing to build the ecosystem perspective. Having
defined the effects of acidic deposition on water chemistry, a discussion of
aquatic organism responses to changing water chemistry follows.
Indirect effects on human health and a discussion of acidic deposition on
materials, man's structures of art and shelter, are also presented. Although
manmade structures are not part of the 'natural ecosystem1 concept, they are
certainly a part of our landscape and any effects of acidic deposition on
them are of concern.
The general content of the chapters is presented briefly below to establish a
general sense of what will be found in more detail in the chapters that
follow.
1.3.1 Effects on Soil Systems
Soils are natural integrators of ecosystem structure and function. They
provide a pathway for water delivered to aquatic systems or for uptake by
1-3
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vegetation. Therefore, in this chapter, emphasis is placed on the natural
processes that contribute to acidification, nutrient status, and metal move-
ment in soils. The effects of acidic and acidifying substances on these
natural processes is then superimposed as an additive factor, and their con-
tribution to these processes is examined. Natural and managed systems are
discussed separately. Reversibility concepts are presented and predictions
of changes over time are made after making several assumptions. These sec-
tions of the chapter are chemically oriented and some basic soil chemistry is
also included.
Nutrient cycling aspects of acidic deposition influences on soils are the
primary emphasis of the chapter. Both the chemical and biological compo-
nents of this process are discussed in detail. The importance of changing
nutrient/metal mobilization activity in soils is discussed as it relates to
both vegetation response and water chemistry. Soil organisms, their role in
nutrient cycling, and the potential and measured effects of acidic deposition
are also discussed.
Soils are chemically and biologically complex systems. The effect that
acidic deposition will have on such systems is dependent on numerous vari-
ables. Because of this complexity and the expectation that potential effects
may be long-term, the definitive conclusions one can draw are not as numerous
as some might expect.
1.3.2 Effects on Vegetation
Most of the terrestrial landscape is covered by vegetation. Because vegeta-
tion collectively includes the primary producers in the food web, its impor-
tance to man is without question. Thus, any change in plant productivity,
whether it be an increase or decrease, can have significant implications for
man's food and fiber system.
The material presented in the vegetation chapter discusses a diverse range of
acidic deposition-plant interactions. These include direct effects on the
smallest scale, i.e., physiological and cell/leaf response, to the gross
scale of forest and crop productivity. The potential effects of acidic depo-
sition, plant, and environmental condition interactions, leading to quantifi-
cation of plant response, are presented. Special attention is given to the
concept of cumulative effects on forests over time and the current lack of
data in this field of acidic deposition effects. The effects of vegetation
on deposition chemistry, as it passes through/over vegetation to soils, is
not discussed in detail.
Plants are subject to more environmental stress factors than most other com-
ponents of the system. Their fixed position in the system causes them to be
exposed regularly to changes in air quality, precipitation chemistry, soil
physiochemical characteristics, disease influence, and climate, to which
their limited avoidance/tolerance mechanisms may or may not be able to
respond. This immobility and dependence on air, soil, and water regimes of
high variability make it difficult to isolate single causes of response,
whether they be beneficial or detrimental. At the present level of under-
standing of plant response as influenced by general stress factors, the
1-4
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direct and indirect effects of acidic deposition that can be definitively
stated are extremely limited.
1.3.3 Effects on Aquatic Chemistry
Most of the present concern for the potential effects of acidic deposition,
and the significance of these effects, has been derived from the aquatics
literature. As already noted, lakes and streams in an ecosystem are not
isolated units. They are directly subject to acidic deposition inputs, but
they are also dependent on the terrestrial system buffering, or lack of
buffering, of these inputs. Unlike the longer term, chronic changes in soils
and vegetative productivity, evidence suggests that aquatic systems are
responsive to both episodic shocks of acidity (e.g., during snow melt) and
chronic inputs of acidic and acidifying substances over time.
The discussion of aquatic chemistry is designed to deal with the complexity
of processes that influence water quality and the relative importance of
these processes/events. Because considerable emphasis has been placed on
aquatic resources in the study of acidic deposition, rather lengthy dis-
cussions of methodology and historical trends are relevant to drawing con-
clusions regarding impacts of acidic deposition and are included. These
topics have been an important source of controversy and are therefore dealt
with in detail in this section. Predictive models, sensitive regions,
significance of metals, and mitigative strategies are also discussed
extensively.
The data base for defining historical changes in aquatic chemistry as a
result of acidic deposition is among the strongest for the ecosystem com-
ponents discussed in this document. Like any of the other system components,
however, predictions of water quality require an understanding of a large
number of other influencing variables, e.g., soils. Unfortunately, our pres-
ent ability to predict changes expected from acidic deposition is limited
since predictive models have yet to be adequately validated.
1.3.4 Effects on Aquatic Biology
The emphasis of the aquatic biology chapter is placed on the response of
aquatic organisms to acidification. For the most part, these discussions do
not attempt to link the acidic deposition phenomenon to observed biological
changes, but instead define the link between biological response and
acidification, whatever the cause.
The chapter discusses the biota found in naturally acidic systems, recog-
nizing that such systems have and will always exist. Such information proves
useful for comparing naturally vs artificially acidified systems and the
biota that are found in both. The components of the food chain in oligo-
trophic water systems most susceptible to change are discussed relative to
their importance and response to acidity. Benthos, macrophytes, plankton and
fish are included. Organisms which are dependent on aquatic systems, for at
least a portion of their life cycle, are also discussed. Mechanisms of re-
sponse, field and laboratory evidence for changes in aquatic biota resources,
and biological mitigation options are also presented and evaluated.
1-5
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Although predictions of species survival as a function of water quality are
feasible, the limited resource inventory and lack of predictive chemistry
models inhibits quantification of the magnitude and extent of acidic depo-
sition impacts on aquatic resources. Quantification of direct impacts of
acidification is most likely for the higher trophic levels, e.g., fish,
especially as better resource inventories become available. However, the
effects of acidification on the interactions between trophic levels remain
unclear at this time.
1.3.5 Indirect Effects on Health
Limited data are available on the potential and known effects of acidic
deposition on human health. Food chain dynamics are discussed in a bioac-
cumulation context. Particular emphasis is placed on aquatic organisms of
importance to man, and drinking water from ground, surface, or cistern
systems. Those metals suspected as being influenced by acidity are high-
lighted. These include mercury, lead, and aluminum.
Although the acidic deposition oriented 'toxicity data base1, is somewhat
limited, the authors have capitalized on the extensive toxicity literature
and research in other fields of science. Superimposed on these concepts is
the effect of acidification, and conclusions are drawn.
1.3-6 Effects on Materials
Like the natural ecosystem, materials, both natural and manmade, are subject
to many environmental influences. Among them are the effects of acidic and
acidifying substances. This chapter of the document reviews the rather
limited data available on the specific topic of acidic deposition effects, as
defined in this document, and discusses the major building and construction
materials that might be affected by acidic deposition. A separate section
discusses corrosion on water piping systems. Mechanisms of damage, economic
implications, and mitigative measures are presented and evaluated. The
importance of dry deposition over wet deposition is highlighted.
1.4 ACIDIC DEPOSITION
The previous sections refer to acidic deposition without definition. Volume
I, Chapter A-l defines this term for technical use in the atmospheric/
deposition sciences. However, from an effects point of view, the chemical
quality of precipitation is as, if not more, important than the pH. Depo-
sition, both wet and dry, contains both essential and nonessential substances
needed by ecosystems as part of their natural nutrient cycle. Therefore, the
materials presented in the effects chapters concentrate on the generic
concept of acidification and the importance of sulfate and nitrate loadings
to the ecosystem. Whether these substances are deposited in dry or wet form
is not differentiated. Because inputs of sulfur and nitrogen can be acidic
upon delivery, or can become acidifying as they are cycled through the
system, these substances are the critical elements for discussion. Because
the data bases were not sufficient to conclusively define input limits for
'protection1 of biological resources, there was no need to deal with a sepa-
ration of wet and dry forms of deposition. When simulated treatments are
1-6
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involved, differentiation of deposition forms is noted as necessary, e.g., in
the crop productivity discussion. Although an effort to separate the com-
ponents of deposition was not undertaken, this does not minimize the
potential for differential effects of wet vs dry deposition exposures.
Therefore, reference to acidic deposition will refer to total deposition of
acidic or acidifying substances. Differentiation is made only as deemed
appropriate by the authors on an issue-by-issue basis.
1.5 LINKAGE TO ATMOSPHERIC SCIENCES
Every effort to use information from the atmospheric chapters of the document
was made. Reference to deposition changes over time, emissions levels,
natural vs anthropogenic sources of sulfur and nitrogen, and/or sulfur and
nitrogen loadings are consistent with those presented in Volume I. Any con-
clusion which would have been drawn using data not consistent with the
atmospheric/deposition chapters was modified or removed. Therefore, Volume I
appropriately sets the stage for the levels of acidity/deposition, the
'cause1, that was considered in the development of the effects presentations.
References to chapters in Volume I are made, as necessary.
1.6 SENSITIVITY
In addition to problems of interpreting the meaning of the acidic deposition
concept, other terminology is equally subject to misinterpretation. In
particular, the term 'sensitivity1 lends itself to varied interpretations.
Sensitivity, as used in the effects chapters, refers to the relative poten-
tial for changes to occur within an ecosystem or one of its components. A
highly sensitive portion of an ecosystem will change more noticeably, or
rapidly, in response to acidic inputs than will one that is generally classi-
fied as having moderate, low, or no sensitivity. However, the reader must be
cautious in many of the effects areas to be certain the reference to sensi-
tivity is clear. For example, reference to a sensitive soil is not meaning-
ful. Acidic deposition effects must be considered with respect to a specific
physiochemical property of the soil. Soil-metal mobility or pH, for example,
can be classified as 'sensitive' to change. Likewise, particular tree
species, aquatic organisms, processes, and/or materials can be sensitive to
change due to acidic deposition. However, developing sensitivity classifi-
cations for larger units of the ecosystem can be misleading, and comparing
dissimilar ecosystem components, e.g., soils and fish, is inappropriate. In
addition, quantification of 'sensitivity1 is defined in the aquatic chemistry
chapter but only qualitative relative usage of the word appears in discus-
sions of other ecosystem components.
1.7 PRESENTATION LIMITATIONS
A phenomenon as complex as acidic deposition cannot be presented with respect
to every environmental factor that might influence ecosystem response. In
the discussions that follow, it is recognized that acidic deposition is
treated as if it were isolated from other pollutants with which it might
interact. Thus, not every possible link between the ecosystem and influenc-
ing has been considered. What is presented is the authors'/editors'
1-7
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unimportant. Rather, an absence of discussion suggests that the issue has
not, as yet, been recognized as essential to our understanding or that data
to support any relevant comments were lacking.
1-8
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS
E-2. EFFECTS ON SOIL SYSTEMS
(W. VI. McFee, F. Adams, C. S. Cronan, M. K. Firestone,
C. D. Foy, R. D. Harter, and D. W. Johnson)1
2.1 INTRODUCTION
Soil plays a key role in ecosystems. It is one of their most stable com-
ponents and, when combined with climate, defines a terrestrial ecosystem's
productivity limits. Moreover, because much of the water entering streams
and lakes directly contacts soil, soil properties also exert important
influences on aquatic systems.
Because of soil's importance to most ecosystems, the impact of acidic
deposition on soils assumes prominence in our discussion. Defining soil
sensitivity to acid inputs depends on understanding soil properties and
chemistry, which are discussed early in this chapter. Thereafter, we can
locate vulnerable soils and determine expected and potential effects on
various soil components. Types and rates of changes can be determined, and
the effects of soil changes on aquatic and terrestrial ecosystems can be
considered. Specifically, questions concern impacts on soil fertility;
nutrient, toxic substance, and organic acid availability; plant vitality; and
water quality. Both short and long-term implications must be considered in
relation to numerous soil components, to soil-plant relationships, and to
soil-water relationships.
2.1.1 Importance of Soils to Aquatic Systems
Aquatic systems receive diverse inputs from terrestrial ecosystems. In-
fluences of acidic deposition on transfers from terrestrial to aquatic
systems may be direct, when material deposited from the atmosphere passes
over or through the soil with little interaction, or they may be indirect,
when deposited materials cause changes in soil processes, such as weathering,
leaching, and organic matter decomposition. Thoroughly assessing effects of
atmospheric deposition on any element transferred from a terrestrial to an
aquatic system requires extensive measurements of inputs, internal processes,
and outflows (Gorham and McFee 1980). These authors note that our under-
standing of the processes is rather incomplete.
of these authors have contributed to this chapter. Because of sub-
sequent integration of the material,*these authors are not identified by
section.
2-1
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2.1.1.1 Soils Buffer Precipitation Enroute to Aquatic Systems—Soil systems
are generally strongly buffered against changes in pH. They are usually
thousands of times more resistant than water to pH shifts (Brady 1974).
Therefore, pH of deposited precipitation tends to shift toward that of the
soil if the water comes into intimate contact with the soil. The cation
exchange capacity (CEC) of the soil and the extent of its saturation with
basic cations (e.g., Ca2+, Mg2+, K+) determine the soil buffering
capacity in moderately acid soils (see Section 2.2). Strongly acid soils may
be buffered by the soil minerals. In general, soils with high clay content,
especially smectite clays, and with high organic matter content are strongly
buffered. These soils tend to deliver water that has come in intimate
contact with the soil matrix to aquatic systems at or near the soil pH. In
areas with alkaline, neutral, or slightly acid soils, the soil buffer system
removes much of the acidity in acidic deposition. Where the soils are near
the acidity of the incoming precipitation, they may not change the pH of
water as it passes through, especially if the soil solution remains rather
dilute.
2.1.1.2 Soil as a Source of Acidity for Aquatic Systems—Many of the soils
in the world's humid regions have been acid for very long periods. Bailey
(1933) pointed out that podzol soils (soil order Spodosol) were generally the
most acidic, followed by lateritic (Oxisols and Ultisols) and podzolic (Ulti-
sols and Alfisols) soils. He did not consider organic soils (Histosols),
many of which are quite acid. For example, all of those designated "Dysic"
at the family level of classification have a pH less than 4.5, and some have
a much lower pH (Soil Survey Staff 1975). Drainage waters from such acid
soils may be equally acidic as the soil and essentially control the pH of
receiving lakes or streams. In many cases, however, after percolating water
passes through acid soil, it interacts with more basic materials underneath
before reaching a stream. Thus, a lake may be surrounded with surface soil
considerably more acid than the water. Such is the case around many lakes in
the Adirondack mountains where most of the soils are strongly acid (Galloway
et al. 1980).
2.1.2 Soil's Importance as a Medium for Plant Growth
All of the other roles of soil fade into insignificance when compared to its
importance as a medium for plants. Soil provides the physical support, most
of the water, nutrients, and oxygen needed by plant roots for normal growth
and development. Well over 95 percent of our food and much of our fiber come
directly or indirectly from terrestrial plants. Soil properties set limits
on the productivity of terrestrial ecosystems. Even though soils tend to
resist rapid change, any significant reduction in their ability to support
plants, such as the increased Al toxicity cited by Ulrich et al. (1980) and
A. H. Johnson et al. (1981), is a serious matter.
2.1.3 Important Soil Properties
Any changes deleterious to the soil's role as a plant growth medium or that
alter its output to aquatic systems are causes for concern. These include
chemical changes, such as in acidity, nutrient supply, cation exchange
capacity, leaching rates of nutrients, or mobilization of toxic substances;
2-2
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physical changes, such as accelerated weathering rates or changes in aggre-
gation; or biological changes, such as reductions in nitrification or other
processes.
2.1.3.1 Soil Physical Properties—Soil physical properties are never inde-
pendent of chemical and biological properties; however, water movement, water
retention/storage capacity, and soil aeration are determined primarily by
physical properties. Controlling water flow is the most important influence
of soil physical properties on interaction of soil with acid rain. Soils
that have high surface runoff rates, such as those on steep slopes or with
low porosities, tend to transmit water rapidly without changing its compo-
sition. Likewise, if the soil has many coarse pores and is well drained, as
are many sands and loamy sands, water passing through may be changed only
slightly. Therefore, if the primary consideration is protection of a body of
water by the soil's buffering capacity, the two situations described are
"sensitive." On the other hand, if changes in the soil itself are the
concern, these soils are not particularly sensitive from the physical
standpoint.
2.1.3.2 Soil Chemical Properties--Resistance of soil chemical properties to
the effects of acidic deposition is measured in terms of the buffering
capacity, initial pH, sulfate adsorption capacity, and amount and type of
weatherable minerals. Soils with high buffering capacities due to high CEC
and high base status will be very slow to respond to acid inputs of the
magnitude acidic deposition introduces. Weatherable minerals containing
carbonates are common in lower horizons of the younger soils in many regions
and will effectively neutralize acids from all sources. Details of these
relations are discussed in later sections.
2.1.3.3 Soi 1 Microbiology--Biological processes in soils may be influenced
by acid deposition and, at the same time, provide some of the means of
resistance and/or recovery. If important soil biochemical processes, such as
N fixation, nitrification, organic matter decay, and nutrient release are
changed by acid deposition, the impact could be significant. Studies of
relationships of soil acidity to biochemical activity are plentiful. How-
ever, most have doubtful applications to the acid deposition problem because
they were studies of natural pH differences, not of shifts due to acid
inputs. A few recent studies indicate alteration in microbial activity near
the soil surface due to simulated acid precipitation (Strayer and Alexander
1981, Strayer et al . 1981). The capacity of most soils to buffer acid inputs
as well as the diversity and adaptability of microbes in the soil contribute
to resistance to acid deposition effects. A more complete discussion of soil
biology and acidic deposition follows in Section 2.4.
2.1.4 Flow of Deposited Materials Through Soil Systems
A generalized depiction of the flow of deposited materials through a ter-
restrial ecosystem is shown in Figure 2-1. In a forested ecosystem (and to a
lesser degree on cropland), a major portion of the precipitation is inter-
cepted by foliage. The chemical properties of the resultant throughfall and
stemflow can be substantially altered from the incipient precipitation (see
Section 3.2.1.2). While this alteration may be of no importance in
2-3
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ACID DEPOSITION
INTERCEPTION
DIRECT DEPOSITION
THROUGHFALL
SURFACE FLO
Minimum to
Moderate soil
interaction
CHANNELIZED FLOW
Minimum soil interaction
GROUNDWATER FLOW
DIFFUSION FLOW
Maximum soil interaction
T-r-— IMPERVIOUS ZONE —
>—— ( > - 1 1
-X- I ( —L_ ( L.
L.T 1. .1'.
Figure 2-1. Flow paths of precipitation through a terrestrial system.
2-4
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constructing the total system input-output balance, it has a big impact on
the nature of reactions expected at the soil surface.
Upon striking the surface, the water may infiltrate the soil or move later-
ally as surface flow. In a forested ecosystem, surface flow will usually not
be visible on the forest floor but will flow through the surface organic
layers. This provides opportunity for water to react chemically with sur-
ficial materials to a greater extent than does surface flow in cultivated
a'reas. The amount of interaction will be proportional to path length and
flow rate.
In cultivated or uncultivated areas, large channels can be established by
burrowing animals and decomposing roots. These are frequently open to the
surface and provide open conduits for flow of drainage water. These channels
may carry nearly all drainage water during saturated flow, and may be domi-
nant conduits during all rainfall events. Little opportunity for soil
interaction is provided, and the precipitation may be conducted through the
soil with little or no alteration.
Water movement by unsaturated flow will usually be through the capillary
pores where maximum opportunity exists for interaction with the soil. This
is the major source of water to plants. Flow through fine pores is necessary
in many deeper soil layers that have limited macropore space. The various
flow paths are depicted in Figure 2-1.
2.2 CHEMISTRY OF ACID SOILS
A brief discussion of important concepts in the chemistry of acid soils
presented here as background for understanding the sections that folli
Those already familiar with these concepts may wish to proceed to Sect
is
low.
r .. Section
2.3.
Although little is known about the impact of acidic deposition per se on
soils, much is known about acid soils in general. The factors which deter-
mine the natural acidification of soils are important to the development of
an adequate comprehension of recent and/or future acidic deposition impacts.
There are many acid soils in the United States, and it is appropriate to
capitalize on our understanding of these systems.
2.2.1 Development of Acid Soils
The eastern half of the United States has a climate in which rainfall exceeds
the combined losses of water by runoff, evaporation, and transpiration from
the soil. The excess water leaches through the soil, carrying with it basic
cations and other soluble materials. If leaching removes basic cations
faster than they are replenished by natural processes or human activities,
the soil profile becomes increasingly acid and impoverished of nutrients
(Pearson and Adams 1967). However, a prerequisite for leaching to cause soil
acidity is the addition of H+ ions to the system (Bache 1980, Ulrich 1980)
along with the presence of mobile anions. The H+ ions can be donated from
a variety of sources. (See Chapter A-8 for discussion of deposition of
acidic and acidifying substances.)
2-5
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2-2-1.1 Biological Sources of H+ Ions--Although H+ ions may be generated
by chemicalweathering of minerals through hydrolytic reactions, the sig-
nificant sources of H+ production in soils are all based on biological
reactions.
Oxidation of sulfur and sulfidas can be important natural sources of acidity.
Much of the sulfur in soils is present in a highly reduced state. This in-
cludes combined S in soil organic matter and such common minerals as pyrite,
FeS2. The release of sulfur from organic-matter in aerobic soils is
followed by the H+-producing oxidation reaction
S + 3/2 02 + H20 = S042- + 2H+.
Elemental S is sometimes used in agriculture for disease control and as a
fertilizer material. Its contribution to soil acidity is readily calculable
from the equation above, i.e., 16 kg of S per hectare is equivalent to one
hundred cm of pH 4.0 precipitation, 1 keq H+ ha-1.
When sulfide minerals, e.g., pyrite, are exposed to atmospheric oxygen, oxi-
dation of these minerals results in significant H+ production, according to
the reaction
2FeS2 + 7H20 + 7 1/2 02 = 2 Fe(OH)3+ 4S042- + 8H+.
Significant quantities of sulfide minerals are found only in recently exposed
soil materials or those that have been maintained in anaerobic conditions,
e.g., coastal marshes. Therefore, their influence is important only in very
limited areas.
Acidity from nitrification is an important contribution in most soils of the
humid regions. Nitrogen is one of the most abundant elements in plants and
in soil organic matter and is present mostly in a highly reduced state. It
is released from organic matter as NH3, which hydrolyzes to NH4+ in
soil solution. , Much of the NH4+ is oxidized to nitrate by bacteria,
according to the reaction
NH4+ + 202 = N03- + 2H+ + H20.
By this reaction, 9 kg NH4+ ha"1 could produce 1 keq H+ ha-1. The
theoretical maximum acidity from nitrification is never realized in soils
because concurrent or subsequent reactions involving N neutralize a portion
of the H+ produced. This process, when coupled with heavy additions of
ammoniacal fertilizers, can have significant effects (see Table 2-1).
Under poor aeration conditions, some oxidized forms of N and S can be
reduced, resulting in the addition of bases to the soil. This process
becomes dominant only in soils that are submerged or saturated for long
periods each year.
2.2.1.2 Acidity from Dissolved Carbon Dioxide—Atmospheric C02 contributes
some acid to soils; however, the respiratory activities of plant roots and
soil microbes result in soil air containing considerably more C02 than
2-6
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atmospheric air. Soil air commonly contains up to 1 percent C02 in com-
parison to the 0.033 percent of a normal atmosphere (Patrick 19/7). This
C02 lowers the pH of pure water according to the equation below, which can
be derived from the relationships among C02 content of air, dissolved
H2C03, and H+ activity.
(H+) = [1.50 x 10-10 x % CQ2]l/2.
If atmospheric C02 is 0.033 percent, then H+ activity of rainwater is 2.2
x 10-°M (moles per liter) or pH 5.65. If soil air contains 1.0 percent
C02, then H+ activity is 1.2 x 1Q-5M or pH 4.91. Thus, biologically
generated C02 is a source of H+ ions in soils but has very little
influence below a pH of about 5.0.
The dominant source of H+ in many soils used in nonleguminous agricultural
production in the United States is from the use of ammoniacal fertilizers,
e.g., NH3, NH4N03, (NH2)2CO, NH4H2P04, (NH4)2HP04, and (NH4)2S04. Because
nitrogen is often used at 100 to 200 kg ha'1, fertilizers alone may gener-
ate HT in soils at rates of 3.6 to 21.6 keq H+ ha'1. It should be
noted that the net acidification from ammoniacal N is frequently less than
the theoretical due to direct uptake of Nfy"1" by plants and H+-consuming
reactions in soils. Although these calculations are based on fertilizer
application to agricultural lands, these same relationships are applicable
for determining the acidification impact on soils from atmospheric N sources.
Nitrogen additions contribute to acidification by increasing basic cation
removal in plants harvested and by furnishing a mobile anion, N03~, for
leaching losses.
Acidity is also added from soil organic matter. The microbial process by
which plant residues are converted into soil humus generates many carboxyl
ligands, RCOOH, on the humus. The protons of such ligands partially dissoci-
ate, adding H+ to the soil solution. This source of H+ production
becomes increasingly important when large amounts of soil humus are present.
Roots can absorb unequal amounts of anions and cations because the uptake
mechanisms are relatively independent of each other. The electroneutrality
of the soil solution is maintained by plant release of H+ or HC03~
during the uptake process. Plants with N-fixing rhizobia absorb more cations
than anions from the soil when N is obtained almost entirely from N2. High
yielding legumes may produce H+ equivalent to several hundred kg CaC03
per hectare (several keq H+ ha'1).
2.2.1.3 Leaching of Basic Cations—Production of H+ resulting from the
various mechanisms does not produce acid soils unless it is accompanied by
leaching. In the absence of leaching (arid and semi-arid regions), HC03~
tends to accumulate in soil solution, leading to H+ neutralization and
precipitation reactions with Ca. In the presence of leaching, H+ in the
soil solution replaces some of the adsorbed basic cations (Ca, Mg, K) on the
exchange surfaces of soil particles. As the excess soil solution moves down-
ward through the soil profile, it carries basic cations equivalent to its
aniom'c content. Meanwhile, the adsorbed H+ remains in place with the soil
particles, causing the soil to become more acid. In this example, H+ was
2-7
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used for simplicity. Many of the basic ions may actually be replaced by Al
ions [A13+, A10H2+, and A1(OH)2+] as the acid is introduced, but the
net effect on soil acidity is the same (Section 2.2.3).
2.2.2 Soil Cation Exchange Capacity
Many differences in the sensitivity of soils to acidic inputs can be traced
to the extent of base saturation and to differences in cation exchange
capacity (CEC), the sum of the exchangeable cations, expressed in chemical
equivalents, in a given quantity of soil. It is the major characteristic of
soils that prevents them from becoming rapidly impoverished when leached.
This section is presented to explain the source of CEC and some of the
variables which affect it.
2.2.2.1 Source of Cation Exchange Capacity in Soils—To have a CEC, soil
particles must have a net negative charge.Soil clay particles may have a
negative charge due to isomorphous substitution of AP+ for Si4"1" in
tetrahedral layers and of Mg2+ or Fe2+ for A13+ in octahedral layers of
the clay structure. This charge is termed a "permanent charge" (Coleman and
Thomas 1967). A second mechanism is the result of the terminal metal atom's
reaction with water to complete its coordination with either OH~ or Hj?0.
At low pH, the coordinating ligand tends to be HgO, which results in a site
with a positive charge; at high pH, the coordinating ligand tends to be
OH", which results in a negatively charged site. Minerals with this kind
of negative charge as their primary source of CEC are referred to as having a
"pH-dependent charge." Therefore, these soil particles change CEC as the pH
changes.
In most soils, a significant component of the CEC comes from organic matter.
The major portion of soil humus is associated with the clay fraction, except
in extremely sandy soils (Schnitzer and Kodama 1977). Its pH-dependent CEC
is a major component of the CEC of surface soils and may be almost the sole
source of CEC in sandy soils. Soil humus has many ligands from which protons
dissociate, such as carboxyl (-COOH), phenol (-OH), and imide (-NH). In
acidic soils, however, only the carboxyl ligand ionizes enough to affect pH,
i.e., R-COOH -> R-COO" + H+, creating a negatively charged exchange site. The
fraction of H+ that ionizes from carboxyl ligands increases with increasing
pH, thereby increasing soil CEC.
The CEC of surface soils is determined by their clay and organic matter con-
tents. In the highly weathered Ultisol soils common to the Southeast,
surface-soil clays are usually kaolinite and hydroxy-Al intergrade vermicu-
lite. These soils contain a high percentage of sand and low contents of clay
and organic matter, and commonly have a CEC of about 5 meq 100 g"1. In
soils with a more temperate climate in the eastern half of the United States,
soil organic matter is usually greater and smectite clays are sometimes more
abundant; hence the CEC is normally higher, about 15 meq 100 g'1 (Coleman
and Thomas 1967).
2.2.2.2 Exchangeable Bases and Base Saturation—The exchangeable cations in
acid soils consist primarily of Ca, Mg, K, Al, H, and Mn. The basic cations
are Ca, Mg, and K, while Al and H are measures of soil acidity. The fraction
2-8
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of the CEC that is satisfied by basic cations is defined as "base satura-
tion." For a particular soil and CEC method, a well-defined, positive
correlation between pH and base saturation exists. Unfortunately, the CEC
reported in the literature is method dependent. The most common methods of
determining CEC are (1) sum of exchangeable cations by neutral salt ex-
traction, (2) NH4+ adsorption at pH 7.0, (3) Na+ adsorption at pH 8.2,
and (4) sum of exchangeable cations by neutral salt extraction plus
titratable acidity by triethanolamine at pH 8.0. The most commonly used
method is probably 1.0 N^ NH^OAc extraction at pH 7.0, method (2) above.
For soils with similar characteristics, pH can be used as a reasonable
estimate of base saturation. For example, the "soil pH" - "base saturation"
relationship of 111 tisols in Alabama is similar to the combined relationship
of Alfisols, Inceptisols, and Spodosols in New York (Figure 2-2).
Analogous to the base-saturation concept, quantities of individual exchange-
able cations can be expressed in terms of saturation of the CEC. This con-
cept is particularly useful in defining the relative availability of cations.
The cation-saturation concept is also useful in predicting probable toxic
levels of Al. Although Al phytotoxicity is a function of soil-solution Al
activity, it is more convenient to measure exchangeable Al.
2.2.3 Exchangeable and Solution Aluminum in Soilj
Aluminum mobility is a key area of concern for both aquatic impacts and
terrestrial vegetative response relative to acidic deposition. The soluble
Al in soils is a product of acid weathering of clay minerals and other solid
phases in acid soils. As H+ concentration increases in soil solution, the
stability of clay minerals decreases, resulting in the release of Al3+ ions
from their surface structure. Measurable amounts of soluble Al are found
only at a pH less than 5.5. Only a small portion of the dissolved Al resides
in the soil solution. Most becomes exchangeable, since cation-exchange sites
in soils have a strong affinity for A13+ ions.
Even though Al saturation of strongly acid soils (pH < 5.0) will normally
exceed 50 percent of the CEC, the concentration of Al in soil solution is
usually < 1 ppm. The significance of exchangeable Al is two-fold: (1) it is
the major component of exchangeable acidity in soils (i.e., acidity displaced
by a neutral-salt solution), and (2) it is the source for the immediate
increase of Al into soil-solution from an acid soil when replaced by other
cations on the exchange sites.
Soil-solution Al concentration is determined by the pH dependent solubility
of Al-containing clay minerals. For example, kaolinite dissolves according
to the reaction
Al2Si205(OH)4 + 6H+ = 2A13+ + 2Si(OH)4 + H20.
Thus, soil-solution Al concentration will be determined by the activities of
H+, Si(OH)4, or other products of weathering reactions.
Aluminum oxides are common in acid soils, and it is frequently assumed that
solution Al is controlled by A1(OH)3 solubility. In that case, Al3*
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8.0
7.0
6.0
2 5.0
4.0
=3
_l
O
O
UJ
Cr
Q.
2.0
1.0
0
0 10 20 30 40 50 60 70 80 90 100
PERCENT OF BASE SATURATION
Figure 2-2. Typical relationship of soil pH to the percent base
saturation. Adapted from Lathwell and Peech (1969).
2-10
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activity in soil solution is a function only of pH because of the reaction
A1(OH)3 + 3H+ = A13+ + 3H20.
The equilibrium log K for this reaction, log A13+ - 3 log H+, varies from
9.7 for the amorphous oxide to 8.0 for crystalline gibbsite. At pH 5.0, for
example, A13+ activity would vary from 5 yM for the more-soluble
amorphous oxide to 0.1 yM for gibbsite at equilibrium with the soil
solution.
In most acid soils of the United States, clays are primarily aluminosili-
cates, and solution Al is controlled by soil-solution Si as well as pH. When
both Al and Si are present in soil solution, their activities frequently
depend upon a solid-phase component with the general composition of
Al2Si205(OH)4. Its solubility in acid soils is expressed by the
equation
l/2Al2Si205(OH)4 + 3H+ = A13+ + Si(OH)4 + 1/2H20.
The equilibrium log K for this reaction, log A13+ + log Si(OH)4 - 3 log
K1", varies from 5.6 for amorphous halloysite to 3.25 for crystalline
kaolinite. If SI(OH)A in soil solution is 0.2 mM (a reasonable value for
acid soils), then AP+ activity at pH 5.0 would range from 2 yM for
amorphous halloysite to 0.01 yM for crystalline kaolinite at equilibrium
with the soil solution.
This generalized equation for aluminumosilicate weathering also illustrates
the H -consuming potential of the weathering process. Thus, in acid soils
(pH < 5.5) weathering of Al minerals may become the dominant buffering effect
in the soil (Section 2.2.9).
The relative solubilities of Al oxides and aluminosilicates in soils show
that soil-solution A13+ activity, at the same pH, varies according to the
solubility of the Al-control!ing mineral as follows: amorphous Al oxide >
amorphous halloysite > gibbsite > kaolinite > smectite. Consequently, the
level of soil-solution Al, and its phytotoxic effect on plants or its
transport to aquatic systems, varies among soils at the same pH, depending
upon which mineral is controlling solution Al.
Under nonagricultural ecosystems, soils generally contain too little solution
phosphorus (P) to affect soluble Al. However, fertilizer P is an effective
agent for lowering solution Al by forming insoluble precipitates such as
variscite, A1(OH)2H2P04. Dilute acid solutions of Al react with
sulfate to form insoluble compounds but these compounds will be the con-
trolling factor very infrequently. The influence of Al and Mn on plant
nutrition is discussed in Section 2.3.3.3.
In the presence of organic ligands, the solubility of aluminum can be greatly
enhanced (Lind and Hem 1975). Numerous reports emphasize the importance of
polyphenols and other components of soil organic matter in the transport of
Al within soils (Bloomfield 1955, Davies et al. 1964, Malcolm and McCracken
1968). In many cases, organic-aluminum complexes are the major form of
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mobile Al. Cronan (1980b,c) points out the importance of organic substances
in Al leaching and discusses the changes likely when strong acid anions such
as sulfate are present.
Inorganic aluminum is present in acid soil solutions primarily as monomeric
ions, the most common ones being A13+, A10H2+, A1(OH)2+, Al(OH)3°, A1S04
and A1H2P042+. in most acid soils, A1(OH)2+ is the most abundant solution
ion.
Since about 1920 soluble Al has been recognized as an important factor limit-
ing plant growth in acid soils (Adams and Pearson 1967). Because of the
pH-dependent solubility of Al, phytotoxic levels of solution Al can be ex-
pected in most mineral soils when soil pH is < 5.0 to 5.5. Only a fraction
of a ppm is needed for sensitive species to exhibit symptoms (see Section
2.3.3.3.2.1).
2.2.4 Exchangeable and Solution Manganese in Soils
Another result of acidification is associated with the mobility of manganese.
Manganese occurs in soils in three valency states. Since divalent Mn
(Mnz+) is the most soluble form, Mn availability depends upon the redox
potential of the system. The equilibrium between Mn oxides and solution
Mn2+ is subject to rapid shifts in the soil.
In most soils with significant levels of easily reducible Mn, toxic levels of
Mn2+ in soil solution can be expected when soil pH is < 5.5 (see Section
2.3.3.3.2.2). The lower the pH, the more likely phytotoxicity will occur.
Lower redox potentials favor Mn-oxide dissolution. In turn, lower redox
potentials are favored by waterlogged conditions, particularly when accom-
panied by the rapid decomposition of organic matter. Consequently, over the
short-term, toxic levels of Mn are more likely under poorly aerated condi-
tions. A long-term consequence of poor aeration, however, is the depletion
of easily reducible Mn and soluble Mn to quite low levels through leaching.
It is normal for Mn and Al phytotoxic symptoms to occur concurrently in many
acid soils because the pH-dependent solubility of Mn oxides and the Al-
containing soil minerals release toxic levels of Mn and Al at about the same
pH level, i.e., < pH 5.0 to 5.5. Whereas Al phytotoxic symptoms are not
generally evident on aerial plant parts, symptoms of Mn phytotoxicity are
quite severe before plant growth is affected significantly.
2.2.5 Practical Effects of Low pH
Low soil pH influences most chemical and biological reactions. It acceler-
ates mineral weathering and the release of phytotoxic ions to the soil
solution; it affects the downward migration of clay and organic-matter
particles in the soil-profile development process, and it affects the level
and availability of most plant nutrients in the soil-solution.
The solubility of soil minerals at low pH is important to plant growth and
transport of ions to aquatic systems. The common Al minerals or compounds in
acid soils are the aluminosilicates, hydrated oxides, phosphates, and
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hydroxy-sulfates. The relationship of low pH to Al and Mn solubility was
covered in Sections 2.2.3 and 2.2.4, and their influence on plant nutrition
is covered in Section 2.3.3.3.
Low soil pH affects the availability of all macronutrients (N, P, S, Ca, Mg,
K) to some extent (Adams and Pearson 1967, Adams 1978, Rorison 1980). These
effects, however, are seldom great enough to influence plant yields. Nitro-
gen availability is affected because low pH decreases the rate at which
organic matter decomposes and releases N to the soil solution. Phosphorus
availability is affected primarily via chemical solubilities. At low pH {<
pH 5.5), P is made increasingly less available because of its reaction with
Al and Fe. Sulfate availability is determined by both organic-matter decom-
position and by inorganic reactions with Al and Fe. The result of these
effects is that sulfate becomes progressively less available as pH decreases
below 6.0.
Cation (Ca, Mg, K) availability is not readily expressed as a function of
soil pH. The relative availability of these nutrients as a function of pH is
of no practical consequence in most cases, except that most soils become acid
only after depletion of these cations. In strongly acid soils, however,
toxic levels of solution Al render vegetation less able to utilize the Ca and
Mg.
Low soil pH affects the availability of all micronutrients (B, Cl, Cu, Fe,
Mn, Mo, Zn) except chloride (Adams and Pearson 1967, Rorison 1980). The
availability of Cu, Fe, Mn, and Zn is significantly increased by lower soil
pH in the range 6.5 to 5.0. Boron availability increases only slightly with
decreasing pH. Molybdenum availability decreases with decreasing pH because
of decreased solubility of molybdate forms. Additional information on soil
acidity and plant nutrition is given in Section 2.3.
2.2.6 Neutralization of Soil Acidity
In unamended soils, the natural forces that neutralize acidity are weathering
of neutral or basic minerals, the addition of basic materials from the atmos-
phere or floods, and the deposition of basic cations by vegetation recycling.
In humid temperate regions outside of floodplains, the uptake of basic
cations by plant roots and their deposition on the soil surface and weather-
ing are the important neutralizing forces. These forces do not normally
reverse the natural acidification trends, but modify the rate and distribu-
tion of acidification within the soil profile.
The effectiveness with which soil acidity can be neutralized by liming de-
pends upon the purity and particle size of the lime, the amount of lime
applied, the soil pH, the cation exchange capacity, the uniformity with which
the lime is spread, and the extent of soil-lime mixing (Barber 1967). Liming
materials are restricted to the Ca and Mg salts of carbonate, silicate, and
hydroxide. The bulk of agricultural lime comes from ground limestone.
The net reaction that causes lime to neutralize soil acidity is the result of
two separate reactions. One is the cation-exchange reaction that releases
Al^+ and H+ to the soil solution from exchange sites; the other is lime
2-13
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dissolution and the hydrolysis of CC^2-. when exchangeable A13+ is
displaced by Ca2+ from dissolving lime, it undergoes stepwise hydrolysis to
form a precipitate of A1(OH)3 and solution H+ ions. The overall
exchange-hydrolytic reaction is expressed by the equation
2 Al-soil + 3 CaCOs + 3 H20 = 3 Ca-soil + 2 Al(OH)3 + 3 C02«
With thorough mixing of small lime particles with an acid soil, the neutra-
lization reaction is quite efficient in raising soil pH to about 6.0. Lime
becomes increasingly less effective in dissolving and raising soil pH beyond
this value.
2.2.7 Measuring Soil pH
The term "soil pH" as it is commonly used refers to the pH of the solution in
contact with the soil. Soil pH is one of the most useful measurements made
on soils (Adams 1978). It is used to predict the likelihood of excessive
toxic ions, the need for liming a soil, a variety of soil microbial
activities, and the relative availability of several inorganic nutrients.
The usual method of measuring soil pH is to immerse a glass-electrode,
reference-electrode assembly into a soil-water suspension and measure the
electromotive force (emf) of the cell. Part of the measured emf is due to a
junction potential at the salt-bridge, test-solution interface. A basic
premise of soil pH measurements is that the junction potential between the
salt bridge and the test solution (or soil suspension) is the same as with
the standard solution. This equality is realized only where test solutions
and standard solutions are similar in ionic compositions. Soil suspensions
hardly meet this requirement, but they approximate it if the reference
electrode is placed in the supernatant while the glass electrode is immersed
in the settled suspension.
Because soil pH is an empirical value, the method of measurement must be
standardized. Samples should be either air-dried or oven-dried at low
temperature (< 50 C) ; oven drying at 105 C produces meaningless pH values.
When soil solution is separated from solid-phase soil, its pH seldom matches
that of the soil suspension. One reason for the discrepancy is the loss or
dilution of CO^ in the soil solution upon drying of the soil sample and the
subsequent addition of water.
Soil pH is influenced by the soil-water ratio and the salt concentration of
the water used. There is no universal agreement on what the ratio should be.
Soil to water ratios of 1:1 up to 5:1 are commonly used. Since most soils
are highly buffered, the differences obtained due to variations in soil:water
ratio are not of practical importance as long as the procedure is consistent
and stated with the results.
In acid soils, soil pH generally decreases temporarily with the addition of
fertilizer or other salts and increases with the dissipation of fertilizer,
either by crop removal or by leaching. In poorly buffered soils, this pH
change may be as much as 0.5 to 1.0 pH unit for normal fertilizer rates.
These changes in soil pH are not due to changes in total soil acidity but are
2-14
-------
due to shifts of Al and H ions from exchange sites to soil solution because
of cation-exchange reactions. Some of this variation can be overcome by use
of a 0.01M CaCl2 solution instead of water when measuring pH.
If soil acidity of an area is to be monitored over years, time of sampling
should be consistent with annual inputs of fertilizers, natural vegetative
cycles, and weather cycles. The most consistent values will be obtained if
samples are taken when salt content is at a minimum.
Spatial variation of soil pH within a field, both vertically and horizon-
tally requires careful sampling to obtain a sample that represents the area
of interest. The area to be represented should be reasonably uniform in
appearance within one soil series and uniform in history. Several identical
soil cores should be composited and thoroughly mixed before a subsample of
the composite for pH measurement is taken.
2.2.8 Sulfate Adsorption
As pointed out in Section 2.2.1.3, the presence of mobile anions is necessary
for the leaching of cations to occur. The dominant anion in the atmospheric
deposition in North America is sulfate (SCty2'). Therefore, the reaction
of sulfate, especially its adsorption or free movement, is an important soil
characteristic.
Soils containing large quantities of amorphous Fe and Al oxides or hydroxides
have a capacity to adsorb SO^-. Sulfate adsorption results from the
displacement of OH~ or the protonation of OH to form OH2+ on iron or
aluminum hydroxide surfaces (Rajan 1978). This results in an increased nega-
tive charge on the hydroxide surface which accounts for the simultaneous
retention of sulfate and associated cations in soil. Sulfate adsorption is
strongly affected by pH since deprotonization of amphoteric adsorption sites
can make them negatively-charged and cause repulsions of anions. Sulfate
adsorption is also affected by the cations present on exchange sites, with
the presence of polyvalent cations causing more adsorption than monovalent
ions. Soil pH is a more important factor than cation type, however (Chao et
al. 1963). Recently, it was shown that organic matter has a decidedly
negative influence on sulfate adsorption, even when free Fe and Al oxide
content is high (Johnson et al. 1979, 1980; Couto et al. 1979). This effect
is thought to be due to the blockage of adsorption sites by organic ligands.
The question of reversibility of sulfate adsorption is crucial to the long-
term effects of acidic deposition on soil leaching. If sulfate is irre-
versibly adsorbed, sulfate adsorption can be viewed as increasing the soil's
capacity to accept acidic deposition before significant leaching of cations
begins. If sulfate is reversibly adsorbed, however, its effects on reducing
leaching are only short-term, since desorption of sulfate will result in
equivalent losses of sulfate and cations from the soil.
The reversibility of sulfate adsorption varies with soil properties and the
desorbing solution used. In some cases, H20 recovers all adsorbed sulfate
whereas in other cases, full recovery is achieved only with phosphate or
2-15
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acetate extractions. Reasons for the better recovery with phosphate or
acetate include the greater affinity of these anions for adsorption sites
and, in the case of acetate, the increase in pH as well. Pre-treatment of
soils with phosphate (such as by fertilization in the field) is known to
reduce sulfate adsorption capacity since sulfate does not displace phosphate
from adsorption sites. However, phosphate does not always displace all
adsorbed sulfate, as shown by Bornemisza and Llanos (1967) for highly-
weathered tropical soils rich in Fe and Al oxides.
There is evidence that "aging" or prolonged contact between soil and solution
reduces the recovery of sulfate (Barrow and Shaw 1977). This effect is
attributed to slow reactions and occurs with other adsorbed anions as well.
Some soils are known to adsorb sulfate irreversibly (against ^0) under
field conditions but not in laboratory conditions (Johnson and Henderson
1979), a phenomenon likely related to slow reactions. Microbial immobiliza-
tion may be a factor in the "aging" phenomenon as well.
Sulfate adsorption is concentration-dependent i.e., sulfate adsorption
increases with solution sulfate concentration (Chao et al. 1963). Thus, for
any given input concentration, sulfate will adsorb on to soil sesquioxide
surfaces until the corresponding soil adsorbed sulfate value is reached on
the sulfate adsorption isotherm. When that point is reached, the soil should
be in steady-state with outputs equalling inputs. In the case where sulfuric
acid inputs increase, concentrations increase, thereby activating "new"
sulfate adsorption sites and causing a net sulfate retention in the soil.
With continued inputs, a new steady-state condition would eventually be
reached. This is schematically depicted in Figure 2-3 (Johnson and Cole
1980).
This concentration-dependent relationship will result in a "front" moving
downward through a sulfate adsorbing soil when a new, higher level of sulfate
concentration is introduced, and continually applied to the soil. Soil above
(or behind) the front will have a new higher level of sulfate on the soil in
response to the higher solution levels. Soil solution samples taken behind
the front might indicate signficant movement of cations and sulfate, while
samples at a lower depth indicate essentially no leaching of cations and
sulfate. Thus, the sulfate adsorbing soil delays cation leaching effects of
dilute sulfuric acid inputs until the adsorbing capacity (dependent on input
concentration) is satisfied down through the soil zones of interest. The
length of time associated with these processes likely ranges from a few weeks
for small changes in soils of low sulfate adsorbing capacity to decades for
large changes to occur in soils of high sulfate adsorbing capacity (Johnson
and Cole 1980, Lee and Weber 1982).
Where sulfuric acid inputs decrease, sulfate will desorb from the soil,
unless it is irreversibly adsorbed, to a point on the isotherm at which the
equilibriumsulfateconcentration equals input concentrations. At this
point, inputs and outputs are equal. Prior to this point, outputs exceed
inputs during sulfate desorption and the sulfate and cations previously
retained during adsorption are leached from the soil.
2-16
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CO
ce.
o
SULFATE ADSORPTION
UNTIL STEADY-STATE
IS ACHIEVED
SOLUTION SULFATE CONCENTRATION
Figure 2-3. Schematic representation of a soil sulfate adsorption
isotherm. U = undisturbed soil conditions, I = soil
conditions following increased ^504 input, SS and NON-SS
refer to steady-state and non-steady-state conditions,
respectively. Adapted from Johnson and Cole (1980).
2-17
-------
Sulfate adsorption capacity of soils is not routinely determined; therefore,
the extent of soils with significant capacity to adsorb sulfate has not been
established. Some adsorption is a common property of many ill ti sols, Oxisols,
some Alfisols, and is reported for other soils (Singh et al. 1980). The work
of Johnson and Todd (1983) shows sulfate adsorption is low in Spodosols. The
distribution of these soil orders within the United States is depicted in
Figure 2-4 in Section 2.3.5.
2.2.9 Soil Chemistry Summary
Acid soils are a natural consequence of long exposure to a climate of excess
rainfall because of the leaching action of natural inputs of acidic ions.
Unleached soils do not become acid. The rate at which leached soils become
acid depends upon soil characteristics, including buffer capacity, and the
rate of H+ input and the accompanying anion. Natural H+ inputs come from
002, Or9anic matter, nitrification, and sulfur oxidation. The buffer
capacity of soils partially neutralizes H+ input by reactions with carbo-
nates (> pH 7.0), with exchangeable bases (pH 5.5 to 7.0), and with clay
minerals (< pH 5.5). Soil-mediated injury to vegetation from H+ inputs
occurs only when pH is low enough to cause significant dissolution of Al- or
Mn-containing clay minerals (< pH 5.0 to 5.5).
The amount of H* required to lower pH of an acid soil depends upon the CEC
of that soil. For example, a loamy sand ill ti sol with the rather low CEC of
2.0 meq 100 g-1 requires about 1.1 meq H+ 100 g-1 to lower pH from 6.0
(65 percent base saturated) to 4.5 (10 percent base saturated). That would
be about 22 keq K* ha'1 to effect the change to a depth of 15 cm. A
finer textured Ultisol with a CEC of 10 meq 100 g-1 requires about five
times that amount. Soils high in smectites (expandible clays) or organic
matter require considerably more H+ for a comparable pH change.
The weathering of aluminosilicate clays will produce strong buffering in
soils that are already acid (5.5 or below) such that calculations of pH
changes, based on changes in basic cation removal by H+ additions, grossly
underestimate the amount of acid required to cause the changes in these
soils. The presence of sulfate adsorption capacity (see Section 2.2.8)
increases their capacity to absorb dilute H2S04 inputs before significant
change in pH or base status occurs.
2.3 EFFECTS OF ACIDIC DEPOSITION ON SOIL CHEMISTRY AND PLANT NUTRITION
It is not always clear what deposition is acidic or acidifying. From the
standpoint of the effects on neutral to acid soils, the following deposi-
tional materials could be expected to have acidifying effects: H2S04,
HN03, H2S03, S02, S, NH3, (NH4)2S04, whereas the following sulfate salts are
essentially neutral or slightly basic in effects on long-term soil pH: CaS04,
K2S04, Na2S04, MgS04. Carbonates of calcium and magnesium would raise the pH.
To alter the soil chemically, precipitation must bathe the soil particles.
Runoff water will minimally impact soil due to its brief contact with soil
particles. As Tamm (1977) has noted, water percolating through soil is not
necessarily at equilibrium with the soil solution but may move directly
2-18
-------
through old root channels, animal burrows, and large pores at ped surfaces.
Soils percolating similar quantities of water may differ in the extent of
their reaction with the water. Under unsaturated conditions, water tends to
move through the small pores of soil aggregates and has the best opportunity
to attain chemical equilibrium with the soil. During a rainfall, the flow
velocity in the small pores within aggregates becomes negligible relative to
that in the large pores between aggregates. Drainage water, therefore, only
reacts with the soil to the extent that dissolved constituents diffuse
between the small and large pores (Bolt 1979). This effect can be demon-
strated by comparing soil solution chemistry, obtained by porous ceramic
cups, with that of free leachate water. Using this system, Shaffer et al.
(1979) demonstrated that solutions applied to a saturated soil can pass
through the soil rapidly and nearly unchanged.
2.3.1 Effects on Soil pH
In considering the effects of acidic deposition, it is essential to realize
that acids are produced naturally within soils (Reuss 1977, Rosenqvist 1977,
Rosenqvist et al. 1980; also see Section 2.2.1). Atmospheric acidic inputs
must be viewed as an addition to natural, continual acidification and leach-
ing processes due to carbonic acid formation, organic acid formation, vege-
tative cation uptake, and a variety of management practices (Reuss 1977,
Johnson et al. 1977, Andersson et al. 1980, Sollins et al. 1980). In Table
2-1 several values are given for potential acidifying or neutralizing effects
of lime, N fertilizer, acidic precipitation, and internal acid production in
soils. Even though most of the values are only approximate, it is clear that
a year of rather heavy acidic deposition has potential acidifying effects
that are small compared to common agricultural amendments. For that reason,
it is generally concluded (McFee et al. 1977, Reuss 1977) that acidic depo-
sition will not have a measurable effect on the pH of soils that are under
normal cultivation practices.
The values for internal acidity production (see Table 2-5 in Section 2.3.3.1)
span a wide range. If the lower values occur, then acidic deposition is
potentially as influential as natural processes, but in other cases it would
be quite small and of little consequence in natural ecosystems. Unfortu-
nately, the data base for including natural acid formation in assessments of
impact on soils is extremely limited. Thus, current schemes, by default,
often assume that atmospheric inputs add significantly to internal acid
production, an assumption that is not universally accepted (e.g., Rosenqvist
1977, Rosenqvist et al. 1980). Carbonic acid is a major leaching agent in
some forest soils (McColl and Cole 1968, Nye and Greenland 1980), yet it does
not produce low pH (i.e., < 5.0) solutions under normal conditions (McColl
and Cole 1968; Johnson et al. 1975, 1977). Organic acids may contribute
substantially to elemental leaching in forest soils undergoing podzolization
(Johnson et al. 1977) and can produce low pH (i.e., < 5.0) in unpolluted
natural waters as well (Johnson et al. 1977, Rosenqvist 1977, Johnson 1981).
Experiments that directly indicate a change in pH due to acidic deposition
inputs (Tamm 1977, Abrahamsen 1980b, Farrell et al. 1980, Wainwright 1980,
Stuanes 1980, Bjor and Teigen 1980) either used accelerated application rates
far exceeding natural precipitation or applied concentrated acid. Both
2-19
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TABLE 2-1. RELATIVE ACIDIFYING AND NEUTRALIZING POWER OF
MATERIALS ADDED TO SOILS
Source
Potential acid or base effect
Agricultural liming operation Neutralizing or basic effect
5000 kg CaC03 ha"1 100 keq ha'1 10 eq nr2
Nitrogen fertilization with
reduced form of N, such as
urea or Nlty
70 kg N ha'1
Acidifying effect3
10 keq ha'1
1 eq m"2
Atmospheric deposition
1 year (100 cm) pH 4.0 rain
Acidifying effect
1 keq ha'1
0.1 eq m"2
16
kg S ha"1 dry deposition Acidifying effect
1 keq ha"1
0.1 eq n
Internal acid production in
soils due to carbonic and
organic acids in one year
from Table 2.5
Acidifying effect
0.23-22.7 keq ha'1 .023-2.27 eq nT2
aN fertilization usually has somewhat less actual acidifying effect.
This is the maximum assuming complete nitrification of the N fertilizer.
2-20
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create situations unlikely to exist in nature because they do not allow for
normal influences of weathering, and nutrient recycling. It is also clear
that soils exposed to concentrated acids over short periods will undergo
reactions and changes that would never occur with more dilute acid over
longer periods. Therefore, the effects of acidic deposition on soil pH are
often predicted from known soil chemical relationships, using input values
similar to those measured in recent years and without the benefit of
long-term experiments under simulated natural conditions.
McFee et al. (1976) calculated theoretical reductions in both soil pH and
base saturation from atmospheric H+ inputs, assuming no concurrent inputs
of basic cations. They concluded that most soils resist pH change and that
there is only a "small likelihood of rapid soil degradation due to acid
precipitation." However, they also suggest that long-term (e.g., 100 years)
soil acidification trends could have an impact on non-agricultural soils and
that these trends are very difficult to evaluate in short-term experiments.
Models of soil acidification processes range from complex ecosystem budget
approaches (Andersson et al. 1980, Sollins et al. 1980) to process-oriented
soil leaching models (Reuss 1978). Their quantitative applicability on a
wide range of sites has not been tested, but they can add to our
understanding of the concepts involved and may be applied to many terrestrial
ecosystems.
Despite uncertainties in estimating potential acidification rates, the
authors of this chapter provide some illustrations in Table 2-2. The data
illustrate that large differences in potential acidification rates can be
expected due to CEC alone, even without considering such other soil
properties as anion adsorption capacity or hydro!ogic characteristics. It
also illustrates how the assumptions concerning accompanying cations, H*
replacement efficiency, and weathering rates change estimates of
acidification rates.
Several considerations embodied in Table 2-2 must be understood if the data
are to be used correctly.
1) The input rates of acidic deposition are considerably higher than those
now reported for the United States.
2) Most natural ecosystems within humid regions have acid soils. Soils
with neutral to slightly-acid pH and with very low CEC, 3 to 6 meq 100
g~l, are uncommon in the humid regions.
3) A 50 percent decrease in base saturation in many mineral soils could
lower pH from the slightly acid (6.6 to 6.8) range to strongly acid
(5.0 to 5.5) range.
4) These estimates ignore anion adsorption capabilities and natural
acidifying processes.
5) Assumptions under scenario 1 are not realized in nature. Those under 2
and 3 are realistic for many soils and many deposition situations, but
cannot be considered universally applicable.
2-21
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TABLE 2-2. ESTIMATES OF TIME REQUIRED TO EFFECT A 50% CHANGE IN BASE
SATURATION IN THE TOP 15 CM OF SOIL. TIME REQUIRED FOR SIGNIFICANT
ACIDIFICATION OF UNCULTIVATED SOILS THAT ARE SLIGHTLY ACID OR NEARLY
NEUTRAL UNDER HIGH RATES OF ACIDIC DEPOSITION—100 CM OF PH 4.0
PRECIPITATION PLUS 16 KG S HA'1 YR'1 IN DRY DEPOSITION
(TOTAL ACID INPUT OF 2 KEQ H+ HA'l YR'l)
Soil
with low organic matter
CEC meq
100 g'1
Assumption
1 2 3
years
Midwestern Alfisol
Southeastern Ultisol
Quartzipsamnent
15
9
3
75
45
15
110
67
22
220
125
45-90
oo
oo
Assumption 1.
Assumption 2.
Assumption 3.
All incoming H+ exchanges for (replaces) basic cations on
the soil exchange complex. There are no accompanying basic
and no weathering or other input of basic cations.
situation and cannot exist in
cations
This is
nature.
the "worst case1
The incoming H+ is accompanied by 0.3-0.5 keq ha~l yr~^
of basic cations Ca, Mg, K (Cole and Rapp 1981), and the
replacing efficiency of H+ for basic cations drops below
1.0 as the base saturation of the soil drops (Wiklander
1975).
Same as under 2 except that acidification is further slowed
by release of basic cations from weathering 1 keq ha~l
yr'1 (for example, 20 kg Ca ha'1 of 15 kg Ca plus 3 kg Mg
ha-1 yr-1) within range calculated for Hubbard Brook
(Likens et al. 1977) and the cycling of basic cations back to
soil surface by plants.
2-22
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If we consider a soil with a low CEC of only 3 meq 100 g-1 and assume a
soil bulk density of 1.3 g cm"3, this soil would have a total of 60 keq
cation exchange capacity per hectare in the top 15 cm (third soil in Table
2-2). A significant pH change could be accomplished by reducing the percent
base saturation by 50 percent. This would seem to be theoretically possible
in 15 years: 15 yr x 2 keq ha'1 yr'1 = 30 keq ha"1. However, all of
the acid input would have to replace and leach an equivalence of bases
(Assumption 1 in Table 2-2). This is highly unlikely. Wiklander (1974)
indicates a replacement efficiency considerably less than 1.0 in acid soils,
pH 5.5 to 6.5. Further, accompanying salts of Ca, Mg, and K also reduce the
acid efficiency in lowering pH (Assumption 2). Such rapid change also
assumes no H+ consumption by weathering and no recycling of bases to the
surface soil whereas Abrahamsen (1980b) indicated weathering rates were
keeping pace with acid inputs in treatments with pH above 4.0. Moreover,
vegetation may deposit significant quantities of basic nutrient ions on the
surface. A more reasonable estimate of the years required to lower the soil
pH significantly, even in this very poorly buffered example, is 22 to 90
years. If a value of 9 meq CEC or higher is assumed (a more common value for
most surface soils in the United States) then the minimum time is 67 years
without weathering and much longer, or infinity, with normal weathering.
The magnitude of soil resistance to pH changes is illustrated by the small pH
changes that have resulted from natural acid inputs of 0.23 to 2.27 ke
ha"1 yr"1 generated by N-fixation metabolism, organic matter decay
COg from respiration (Table 2-1). These inputs have not caused rapid pH
changes and it is unlikely that an additional 2 keq ha"1 yr"1 or less
from acidic deposition will cause a significant change in many soils.
The evidence for acidification of soils by the present rate of acidic depo-
sition is not strong. If significant acidification is to occur within a few
decades, it will be in the limited soil areas that combine the following
characteristics: the soil is not renewed by fresh soil deposits; it is low
in cation exchange capacity, i.e., low in clay and organic matter; it is low
in sulfate adsorption capacity; it receives high inputs of acidic deposition
without significant basic cation deposition; it is relatively high in present
pH (neutral to slightly acid) and free of easily weatherable materials to one
meter depth (see Section 2.3.5.2.1).
As Section 2.3.3.1 discusses, acid precipitation cannot leach nutrient cat-
ions unless the associated sulfate or nitrate anions in the soil are mobile.
Evidence indicates that sulfate is not always mobile (Section 2.2.8) particu-
larly as soils become more acid (Johnson and Cole 1977, Johnson et al. 1979,
Abrahamsen 1980b, Singh et al. 1980).
It is also possible for a soil to be leached of cations without concurrent
acidification, if acidic inputs stimulate the weathering of cations from
primary minerals. Therefore, it is important to make a distinction between
cation leaching and the process of soil acidification. It is unrealistic to
assume either a steady-state condition for soil exchangeable cations or a
condition where weathering is zero and cations are depleted from exchange
sites in proportion to H* inputs. These common assumptions made in pre-
dictive models seriously limit the models' applicability to natural systems.
2-23
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Another important factor which models do not consider is the acidification
caused by natural processes. As noted in Section 2.2.1, atmospheric acid
inputs must be viewed as an addition to the natural acidification processes
of cation uptake by plants, nitrification, and soil leaching by organic and
carbonic acids (Johnson et al. 1977, Reuss 1977, Cronan et al. 1978,
Rosenqvist et al. 1980).
Section 2.1.3, on leaching is closely related because long-term pH changes
require leaching of basic cations as well as acidic inputs.
2.3.2 Effects on Nutrient Supply of Cultivated Crops
This section deals with the significance of atmospheric additions of S and N
to crop requirements. Few detrimental effects of acidic deposition are
expected on nutrient supply to cultivated crops (see Section 2.3.1) because
by comparison agricultural practices have a massive effect.
Input of nutrients to plant systems from rainfall has been documented since
the mid-19th century (Way 1855). Calculations made in a number of U.S.
regions have estimated the seasonal atmospheric deposition of nutrient
species, particularly sulfate and nitrate, to agricultural and natural
systems and the implications of this deposition on plant nutrient status.
Estimates by Hoeft et al. (1972) of 30 kg S ha'1 yr'1 and 20 kg N ha"1
yr~l deposited in precipitation in Wisconsin indicates the importance of
atmospheric sources of these elements. These values, however, are higher
than those usually reported in the United States (see Chapter A-8). Jones et
al. (1979) reported that atmospheric S is a major contribution to the agrono-
mic and horticultural crop needs for S as a plant nutrient in South Carolina.
The amount of annual S deposition at selected sites is presented in Table
2-3. Amounts of S recorded for 1953-55 in rural areas along the Gulf and
southern Atlantic coasts were usually less than 6 kg S ha"1 yr-1. In
northern Alabama, Kentucky, Tennessee, and Virginia the levels were much
higher (10 to 30 kg ha'1 yr-1) (Jordan et al. 1959). These can be
compared with the recent NADP data for wet deposition of S (see Figure 8-19,
Chapter A-8).
These amounts of S represent a significant portion of that required by crops.
The amounts of S absorbed by crops are summarized in Table 2-4. Terman
(1978) estimates an average crop removal of 18.5 kg S ha~l yr"1 and
concludes that if current rates of atmospheric S deposition are greatly
reduced, the need for applying fertilizer S for satisfactory crop yield will
increase.
The atmospheric deposition of N is usually lower than deposition of S, but
crop requirements ate much higher. Therefore, it is generally accepted that
atmospheric hj/deposition plays a small or insignificant role in nutrition of
cultivated-crops (see Chapter E-3, Section 3.4.2).
It is well known that foliar applications of plant nutrients can stimulate
plant growth (Garcia and Hanway 1976). It is possible, but unproven, that
repeated exposure of plants to small amounts of atmospheric deposition may be
2-24
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TABLE 2-3. AMOUNTS OF SULFUR DEPOSITED BY PRECIPITATION IN VARIOUS STATES
State
Southern States
Alabama
Arkansas
Florida
Kentucky
Louisiana
Mississippi
North Carolina
Oklahoma
Tennessee
Texas
Virginia
Location
in state
Prattville
Muscle Shoals
Muscle Shoals
Muscle Shoals
NW and SE
Gainesville
Others
Various
Various
Various
Statesville
Others
Still water
Various
Various
Beaumont
Others
Norfolk
Various
Sites
1
19
20
23
2
1
5
6
5
7
1
15
1
7
5
1
4
1
16
Years
1954-55
1954
1955
1956
1954-56
1953-55
1953-55
1954-55
1954-55
1953-55
1953-55
1953-55
1927-42
1955
1971-72
1954-55
1954-55
1954-55
1953-56
Major
source kg
General
General
Steam Plant
Steam Plant
General
Urban
General
General
General
General
Industry
General
General
General
General
Industry
General
Industry
General
Average
S ha-* yr"1
3.7
5.4
11.9
11.0
3.7
8.8
3.2
13.1
9.0
5.0
15.5
6.0
9.7
14.2
17.1
12.1
5.7
35.2
21.4
2-25
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TABLE 2-3. CONTINUED
State
Northern States
Indiana
Michigan
Nebraska
New York
Wisconsin
Location
in state
Gary
Others
Various
Various
Ithaca
Industrial Site
Urban
Rural
Si tes
1
10
5
7
1
1
9
13
Years
1946-47
1946-47
1959-60
1953-54
1931-49
1969-71
1969-71
1969-71
Major
source
Industry
General
Industry
General
Urban &
Industry
Industry
Urban
General
Average
kg S ha-i yr1
142.2
30.0
11.3
7.2
54.9
168.0
42.0
16.0
Adapted from Terman (1978). See original for data sources.
2-26
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TABLE 2-4. SULFUR CONTENT OF CROPS
Yield Total S Content
Crops tons ha'1 kg ha"1
Grain and oil crops
Barley (Hordeum vulgare L.)
Corn (Zea Mays L.)
Grain sorghum (Sorghum bicolor L. Moench)
Oats (Avena sativa L.)
Rice (Oryza sativa L.)
Wheat (Triticum aestivum L.)
Peanuts (Arachis hypogaea L.)
Soybeans (Glycine max Merr.)
Hay Crops
Alfalfa (Medicago sativa L.)
Clover-grass
Bermuda-grass (Cynodon dactyl on L.)
Common
Coastal
Orchardgrass (Dactyl is glomerata L.)
Timothy (Phleum pratense L.)
5.4
11.2
9.0
3.6
7.8
5.4
4.5
4.0
17.9
13.4
9.0
22.4
13.4
9.0
22
34
43
22
13
22
24
28
45
34
17
50
39
18
Cotton and tobacco
Cotton (lint + seed) (Gossypium hirsutum L.) 4.3 34
Tobacco {Nicotiana tabacum L.)
Burl ey
Flue-cured
Fruit, sugar, and vegetable crops
Beets
Sugar (Beta saccharifera)
Table (Beta vulgaris L.)
Cabbage (Brassica oleracea)
Irish potatoes (Solanum tuberosum L.)
Oranges (Citrus sp.)
Pineapple (Ananas comosus)
4.5
3.4
67
56
78
56
52
40
21
50
50
46
72
27
31
16
Estimates by Potash/Phosphate Institute of North America. Adapted from
Terman (1978).
2-27
-------
more effective in stimulating plant growth than a comparable amount applied
to soils (see Chapter E-3, Section 3.4).
2.3.3 Effects on Nutrient Supply to Forests
Nutrient supply may be influenced by acidic deposition effects on leaching of
cations or by pH-induced changes in mineral solubility, microbial processes,
or weathering rates in addition to the direct influence of additions of N and
S in deposition. Microbial processes are discussed in Section 2.4. Solu-
bility (availability) and weathering reactions are discussed in Section 2.2.
Acid precipitation has created a major concern because of the potential for
accelerated cation leaching from forest soils and eventual losses of pro-
ductivity (Engstrom et al. 1971). This concern was the driving force for
numerous empirical studies of acid precipitation effects on forest nutrient
status in general and cation leaching in particular (reviewed by Johnson et
al. 1982).
Perhaps because of the negative implications of the term "acid rain," initial
speculations about acid deposition effects on forest productivity devoted
little or no attention to concurrent sulfate and nitrate deposition on
forests deficient in S or N. Only recently has it been recognized that acid
deposition can cause increases as well as decreases in forest productivity
(Abrahamsen 1980b, Cowling and Dochinger 1980). The net effect of acid
deposition on forest growth depends upon a number of site-specific factors
such as nutrient status and amount of atmospheric acid input. (See also
Chapter E-3, Section 3.4.1.)
It is also very important to consider that ions such as S042~ and NOs" are
already in the ecosystem and that H+ is generated naturally by the plant
community (Ulrich 1980). Thus, the question is one of relating inputs to
natural levels; e.g., does atmospheric H+ input significantly add to or
exceed natural H+ production within the soil? Do the detrimental effects
of H+ deposition offset the benefits of N03~ deposition in an N-
deficient ecosystem or the benefits of SO/^- deposition in an S-deficient
ecosystem? In short, the problem of assessing the effects of acid deposi-
tion on forest nutrient status is largely a matter of quantification and
requires a nutrient cycling approach.
2.3.3.1 Effects on Cation Nutrient Status—Cation leaching is important to
soil properties because it may lead to a loss of plant nutrients and de-
pressed soil pH. It is important in hydrology because cations leached from
soils may be transferred to aquatic systems.
The basic cation status of a soil depends on the net effect of leaching and
other losses versus weathering and other inputs (Abrahamsen 1980a, Ulrich et
al. 1980). Weathering is stimulated by additional H+ input, offsetting
leaching to some extent. However, most acid irrigation studies (Abrahamsen
1980b) and one study under ambient conditions (Ulrich et al. 1980) indicate a
net decline in exchangeable basic cations with time. There is little doubt
that acid deposition can accelerate cation leaching rates, but the magnitudes
of these increases must be evaluated within the context of natural, internal
2-28
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leaching processes. The magnitude is quite variable, depending upon the
amount of acid input, the rate of soil leaching by natural processes (Cole
and Johnson 1977, Cronan et al. 1978), and the degree to which soils are
buffered against leaching (e.g., by anion adsorption; Johnson and Cole 1977).
Furthermore, the ultimate effects of accelerated cation leaching on cation
nutrient status depend upon a number of variables, most notably exchangeable
cation capital, primary mineral weathering rate (Stuanes 1980), forest cation
nutrient requirement, and management practices such as harvesting.
A comparison of the effects of some of these factors on cation nutrient
status is given in Table 2-5. Various schemes for evaluating internal acid
production have been proposed (Reuss 1977, Sollins et alI. 1980, Ulrich 1980),
but in this case, only the values reported by various investigators for soil
leaching (usually by carbonic acid) are considered. It is obvious that
atmospheric acid inputs vary not only in absolute magnitude, but also in
their importance relative to internal leaching processes and effects of
harvesting.
At the unpolluted site in Findley Lake, it is not surprising that internal
leaching processes and harvesting effects exceed atmospheric H+ inputs.
However, even in the beech stand at Soiling, West Germany, values for
H?C03 production reported by Andersson et al . (1980) exceed atmospheric
H* inputs as measured by open-bucket collectors. In this case, the com-
parison is misleading, however, since dry deposition to the forest canopy at
Soiling is known to be exceedingly high (Ulrich et al. 1980), and, conse-
quently, H+ inputs to the forest floor substantially exceed those deposited
above the canopy. It is also noteworthy that Ulrich et al. believe that
while internal H+-producing processes are important at Soiling, acid rain
is having serious, deleterious effects on forests there.
Studies of basic cation leaching due to acidic inputs sometimes give in-
consistent results. Under ambient conditions, Mayer and Ulrich (1977) noted
a net loss of Ca, Mg, K, and Na from the soils under a beech forest. Except
for Na, however, the loss was equal to or less than nutrient accumulation in
the trees. Roberts et al. (1980) reported that acidic precipitation on
Delamere forest (pine) of central England may produce small changes in litter
decomposition, but they found no effect on Ca, Mg, K, or Na leaching rate.
Cole and Johnson (1977) found no detectable effect of acid precipitation on
the soil solution of a Douglas-fir ecosystem. On the other hand, Andersson
et al. (1980) noted a net output of Ca from both a pine forest soil in Sweden
and a beech forest soil in West Germany; both soils accumulated N but not
sulfate. Cronan (1980a) reported net losses of Ca, Mg, K, and Na from
subalpine soil in New Hampshire, attributing losses to acidic precipitation.
Studies by Mollitor and Raynal (1982) suggest that leaching of K may be the
most serious problem of cation leaching in Adirondack forest soils.
Nitrate is sometimes associated with acidic deposition and differs con-
siderably from sulfate in that it is very poorly adsorbed to most soils
(Johnson and Cole 1977). However, biological processes in N-limited eco-
systems quickly immobilize nitrate, and since N limitations are common in
forested regions of the world, nitrate is rarely mobile (Abrahamsen 1980b).
2-29
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TABLE 2-5. ATMOSPHERIC H+ INPUTS VS CATION REMOVAL BY INTERNAL H+
PRODUCTION (CARBONIC AND ORGANIC ACIDS) AND POTENTIAL NET ANNUAL CATION
REMOVAL IN BOLE ONLY AND WHOLE-TREE HARVESTING (WTH) IN SELECTED FOREST
ECOSYSTEMS (ADAPTED FROM EVANS ET AL. 1981)
Site
Species
Precipitation Cation
Age H+ leaching by
(yr) input9 internal acid
production0
Cation
removal by
harvesting0
Bole WTH
(eq ha-1 yr~l)
Thompson,
Washington
Soiling,
W. Germany
Jadrass,
Sweden
Findley,
Washington
H.J. Andrews,
Oregon
Pseudotsuga 42
menziesn
Fagus sylvatica 59
Pinus sylvestris
Abies amabilis, 175
Tsuga mertensiana
Pseudotsuga 450
menziesn
240d
(4.8)
9009
1909
90"
(5.6)
289
420d
(5.9)
19509
2269
1410h
(4.5)
227009
380e 6606
2209 37Qe
272e 460e
60e 106e
aWeighted average [H+] times precipitation amount; weighted average [H+] as
pH appears in parenthesis where available.
^Calculated from net increase in weighted average HCO^~ or organic anion
concentration (the latter estimated by anion deficit) times water amount.
Weighted average [H+] as pH for solutions appears in parentheses where
available.
°Nutrient content divided by age; WTH = whole tree harvest, removal of all
aboveground biomass.
dFrom Cole and Johnson (1977).
eFrom Cole and Rapp (1981).
fFrom Lindberg et al. (1979).
9From Andersson et al. (1980). For comparison in this table, only H2C03
production values are included.
2-30
-------
On the other hand, nitrogen-rich ecosystems (where biological immobilization
of N03" is minimal) are susceptible to leaching by HN03-
With regard to North American forests, cation deficiencies are very rare
although they are known to occur in red pine (Pinus resinosa) on some sandy
soils in New York State (Stone and Kszystyniak 1977, Heiberg and White 1950,
Hart et al. 1969). Acid rain accelerated leaching could, in theory,
exacerbate this situation, but this possibility has not been investigated.
It should be noted, however, that these ecosystems are exceedingly conser-
vative with regard to potassium (Stone and Kszystyniak 1977), and biological
cycling and conservation may play major roles in resisting effects of acid
rain on K+ leaching (e.g., other cations may be leached while K+ is
conserved).
2.3.3.2 Effects on S and N Status—Deficiencies of S have been indicated in
forests remote from pollutant inputs in eastern Australia (Humphreys et al.
1975) and the northwestern United States (Youngberg and Dyrness 1965, Will
and Youngberg 1978). Humphreys et al. (1975) suggest that pollutant inputs
from power plants benefit S-deficient Australian forests, particularly when
the soils have little S042" adsorption capacity. In these situations
continual input of moderate amounts of H2S04 as acid rain may be a source
of fertilizer.
At the other extreme, continual atmospheric S inputs may help alleviate sub-
optimal sulfate availability in sulfate "fixing" soils that are rich in
hydrated Fe and Al oxides. Although adsorbed insoluble sulfate is thought to
be available to plants in the long run, the intensity or rate of supply to
the soil solution can be less than that required by plants, effecting an S
limitation (Hasan et al. 1970).
Research has shown that N fertilization, a practice in some forested regions
of the world, results in rapid use of ecosystem S supplies, possibly leading
to S limitations (Humphreys et al. 1975, Turner et al. 1980). It has been
suggested that forest N and S status must be evaluated because of the closely
related roles of these elements in protein synthesis (Kelly and Lambert 1972,
Turner and Lambert 1980, Turner et al. 1980). In relatively unpolluted
regions of the northwestern United States, evidence indicates that lack of
growth response to N by Douglas fir is due to marginal S status (Turner et
al. 1977, 1979). Thus, it seems evident that moderate amounts of S in depo-
sition could benefit forests undergoing N fertilization. In the United
States this currently involves a total of about 1,000,000 ha of forest lands,
primarily in the Northwest and Southeast (Bengston 1979).
Amounts of atmospheric S input sufficient to satisfy forest S requirements
are much smaller than many crop requirements. In general, S inputs of 5 kg
ha-1 yr'1 are sufficient to satisfy S requirements in most forest ecosys-
tems (Humphreys et al. 1975, Evans et al. 1981, Johnson et al. 1982). Inputs
of S04^~ in acid rain affected regions frequently exceed this value
(often by a factor of 2 to 4), implying that S is currently being deposited
in excess of forest requirements (Table 2-3 and Chapter A-8).
2-31
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Several studies have shown that excess S cycles within vegetation and accumu-
lates in soils as S042~ without any apparent harm {Kelly and Lambert
1972, Turner et al . 1980, Turner 1980). The plateau between S sufficiency
and toxicity in forest ecosystems appears to be quite broad. Inputs of S
usually constitute a more significant increment to the natural sulfur flux
within forest ecosystems than do equivalent inputs of H+ to the natural
flux of H+. Therefore, it would appear that further emphasis ought to be
given to effects from the S042~ component of acidic deposition. Simi-
larly, further emphasis ought to be given to the effects of N inputs, because
they appear to be increasing (Abrahamsen 1980b) and N is commonly the
limiting nutrient in forest ecosystems.
Nitrogen deficiencies are common in forests throughout the world (Abrahamsen
1980b). Inputs of N03~ (as well as NH4+ and other forms of N) are
likely to improve forest nutrient status and productivity in many cases.
Nearly all forest ecosystems for which nutrient budgets are available appear
to accumulate N03~ as well as other forms of N (i.e., inputs > outputs;
Abrahamsen 1980b). Since N03~ is very poorly adsorbed to most soils
(Vitousek et al. 1979), this accumulation is undoubtedly due to biological
uptake. The inhibiting effect of NOs" immobilization on the leaching
potential of HNO^ is the same as that of $042- immobilization on the
leaching potential of ^$04 even though the mechanisms of immobilization
for those two anions are different.
Because forest N requirements are relatively high compared to S requirements,
instances of atmospheric N inputs in excess of forest N requirements seldom
occur. An apparent exception is the Soiling site in West Germany, where
atmospheric inputs of N, S, and H+ are high (Ulrich et al. 1980).
If atmospheric N inputs increase to the point where N deficiencies are al-
leviated and excess N is available in soils, nitrification may be stimulated.
Nitrification pulses are thought to be responsible for a large percentage of
leaching at the heavily-impacted Soiling site in West Germany, for example
(Ulrich et al. 1980). Thus, nitrogen "saturation" of forest ecosystems could
result in significant increases in cation leaching and, under extreme circum-
stances, soil acidification. Such "saturation" would occur most readily in
forests with low N demand (i.e., boreal coniferous forests; Cole and Rapp
1981) or in forests with adequate or excessive N supplied (such as by N-
fixing species). Indeed, the naturally acidifying effects of red alder, an
N-fixing species indigenous to the northwestern United States, have been
noted by several investigators. However, there is not evidence of wide-
spread, imminent nitrogen saturation of forests since N deficiencies are
still quite common and most ecosystems are still accumulating N (Abrahamsen
1980b, Johnson et al. 1982).
Acidic deposition may indirectly affect N availability in forest soils. Tamm
(1976) predicted short-term increases in N availability (due to increased
decomposition and microbiological N immobilization) and tree growth due to
acidic precipitation. However, long-term declines in both N status and tree
growth could occur due to net N losses from the ecosystem. With regard to
decomposition, empirical results have been variable (see Section 2.5).
Whether this increase in N availability is due to changes in microbial
2-32
-------
activity or to the acid-catalyzed hydrolysis of labile soil N is unknown. In
either event, the results of the Norwegian studies, in which both N avail-
ability and nitrate leaching were stimulated by H^SCty inputs, strongly
suggest that, contrary to earlier predictions that nitrification would be
inhibited by acidic inputs (Tamm 1976), nitrification can be stimulated by
acidic inputs.
2.3.3.3 Acidification Effects on Plant Nutrition--It is unlikely that many
soils will be significantly acidified by acid rain at current input levels in
the United States (see Sectim 2.3.1). Should soil acidification occur,
however (e.g., in restricted areas with high acid inputs and very poorly
buffered soils), a great deal of information is available about plant
responses. Also, recent results from the heavily-impacted Soiling site in
West Germany suggest that slight changes in soil pH due to the combined
effects of acid precipitation and internal processes are causing serious
negative effects on forests there (Ulrich et al. 1980).
2.3.3.3.1 Nutrient deficiencies. In general, only those acidic soils that
are highly leached (sandy and/or low CEC) are likely to be sufficiently low
in Ca to affect growth of higher plants. That is, if Al and other toxic ions
are not present in excess, most acidic soils will have adequate Ca for good
growth of most plants (Foy 1964, 1974a). The evidence suggests that many, if
not all, of the Ca deficiencies reported on acidic soils in the field are due
to Al-Ca antagonisms rather than low Ca per se. For a fuller treatment of
the Ca-deficiency Al-toxicity argument, see earlier reviews (Kamprath and Foy
1972; Foy 1974a,b, 1981). Similarly, magnesium deficiencies observed in
plants grown on acid soils are often due to Al-Mg antagonisms rather than low
total soil Mg levels.
Phosphorus deficiency is a common problem in crops and forests grown on
acidic soils because such soils are often low in total P and because native
P, as well as fertilizer P, is combined with Al and Fe in forms that are only
sparingly soluble (Adams and Pearson 1967, Kamprath and Foy 1972, Pritchett
and Smith 1972, Graham 1978).
Unlike other micronutrients, Mo is less available in strongly acid soils
(Kamprath and Foy 1972). Molybdenum deficiencies such as those reported on
the Eastern Seaboard, in the Great Lakes states, and on the Pacific coast of
the United States generally occur on such soils (Kubota 1978).
2.3.3.3.2 Metal ion toxicities. Any metal can be toxic if soluble in suf-
ficient quantities. In near-neutral soils, heavy metals occur as inorganic
compounds or in bound forms with organic matter, clays, or hydrous oxides of
Fe, Mn, and Al. However, a decrease in soil pH can create metal toxicity
problems for vegetation. Zinc, Cu, and Ni toxicities have occurred fre-
quently in a variety of acid soils. Iron toxicity occurs only under flooded
conditions where Fe occurs as the reduced, soluble Fe2+ form (Foy et al.
1978). Toxicities of Pb, Co, Be, As, and Cd occur only under very unusual
conditions. Lead and Cd are of particular interest because they move into
the food chain and affect human and animal health. For further details, see
a recent review (Foy et al. 1978).
2-33
-------
Aluminum and Mn toxicities are the most prominent growth-1imiting factors in
many, if not most, acidic soils (Foy 1973, 1974b, 1981; Tanaka and Hayakawa
1975). Hence, this review will emphasize the harmful effects of these two
elements on plants. The chemistry of Al and Mn in soils was discussed in
Sections 2.2.3 and 2.2.4.
2.3.3.3.2.1 Aluminum toxicity. Because Al is a structural constituent
of soil clay mineral particles, Al toxicity is theoretically possible in
most, if not all, soils. The primary condition required to produce
solubility of excess Al is a low pH. As Section 2.2.3 pointed out, aluminum
may become soluble enough to be of concern when the soil pH is in the range
5.0 to 5.5 or below.
Aluminum toxicity is believed to be a primary factor in limiting plant root
development (depth and branching) in many acidic subsoils of the southeastern
United States (Foy 1981). For example, Kokorina (1977) noted that acid soil
toxicity was more harmful in dry years. This dry season phenomenon in
concert with acidic deposition may also be a factor in Ulrich's (1980) recent
reports on forest growth reduction in West Germany.
On the basis of some complex theories of ecosystem acidification processes on
and after a decade of monitoring at the Soiling site, scientists at the
University of Gottingen in West Germany state that the forests of Soiling (as
well as others like it in Germany) are being seriously impacted by acid rain
(Ulrich 1980). Most significantly, at the Soiling site Al concentrations in
soil solutions have increased twofold (from 1-2 mg £-1 to 2-5 mg jr1)
beneath the beech stand and ~ tenfold (from 1-2 mg £-1 to 15-18 mg
£-!) beneath the spruce stand over the last decade (Matzner and Ulrich
1981). It is hypothesized that Al concentrations are reaching toxic levels,
thereby damaging or killing tree roots and causing serious consequences to
the maintenance of these forest ecosystems. An important question relative
to toxicity of Al levels concerns the form of Al in soil solution. It would
be important to know the extent of chelation by organic materials.
Atmospheric H+ inputs must be viewed as additions to natural, internal acid
generation (Ulrich 1980). One very important internal H+ generating pro-
cess at Soiling is nitrification in mineral soil layers during warm, dry
years. Nitrification during these periods (thought to be caused by decompo-
sition of previously accumulated N-rich root residues) causes a pulse of acid
production. According to Ulrich et al. (1980), systems that have been acidi-
fied by acid precipitation are unable to withstand such pulses because their
buffering capacities are much reduced. Thus, Al is mobilized at such times,
creating toxic conditions for roots.
Undoubtedly, acid inputs to the Soiling site are very high. Inputs of H+
measured with open-bucket collectors are not themselves excessively high,
being approximately 700 eq ha~l yr'1 (0.7 kg ha"1 yr'1); compara-
tively, H* input values of this magnitude are not uncommon in forests of
the United States (Chapter A-8). However, at Soiling H+ flux in through-
fall is two to five times greater than in open precipitation due to dry
deposition in the forest canopy.
2-34
-------
In contrast to results and hypotheses at Gottingen, scientists with the
Norwegian SNSF Project demonstrated the ability of forest ecosystems to
tolerate acid inputs and Al levels exceeding those reported at Soiling. This
ability is shown by results of an intensive series of irrigation studies
involving inputs of t^SCty ranging from current background levels (ap-
proximately 0.8 keq ha-1 yr"1) up to approximately 30 times that amount
(26 keq ha~l yr~l). Although Al concentrations in soil solutions and in
tree foliage increased substantially, no indications of Al toxicity were
noted and growth effects were small (slight growth increases occurred in some
species, slight decreases in other species, and no effects in some species;
Abrahamsen 1980a,b; Tveite 1980a,b). It is also noteworthy that large
nitrification pulses occurred in most acid treatments (Abrahamsm 1980a) .
Finally, greenhouse studies involving acid irrigation and liming of Norway
spruce showed that this species (which occurs also at the Soiling site) is
extremely tolerant of high acid inputs and foliar Al concentrations.
Plant species and cultivars differ widely in their tolerances to excess Al in
the growth medium. Published references to such differences are too numerous
to cite individually, but access to the older literature is provided in re-
view papers (Foy 1974b, 1981). Aluminum tolerance has been associated with
pH changes in root zones, Al trapping in non-metabolic sites within plants, P
uptake efficiency, Ca and Mg uptake and transport, root cation exchange ca-
pacity, root phosphatase activity, internal concentrations of Si, NH4+ -
N0a~ tolerance or preference, organic acid contents, Fe uptake efficiency
and resistance to drought. For citations from the earlier literature, see
review papers (Foy 1974b, 1981, Foy and Fleming 1978, Foy et al. 1978).
2.3.3.3.2.2 Manganese toxicity. Manganese toxicity frequently occurs
in soils with pH values of 5.5 or below, if the soil parent materials are
sufficiently high in easily reducible Mn content. However, some soils do not
contain sufficient total Mn to produce toxicity, even at pH 5.0 or below.
Soils of the Atlantic Coastal Plain of the United States are lower in total
Mn than those of the Gulf Coastal Plain (Adams and Pearson 1967). However,
within any area, soils vary widely in Mn contents (Sedberry et al. 1978). In
that study, the DTPA extractable Mn varied more with parent material and clay
than with pH and organic matter. Reducing environments induced by poorly
aerated conditions in soils increase Mn availability and potential for
toxicity.
2.3.4 Reversibility of Effects on Soil Chemistry
Changes in soil chemistry caused by acidic deposition in unmanaged
terrestrial ecosystems must, in general, be considered irreversible, but
there are exceptions. Nutrients lost are not readily regained. However,
exchangeable basic cations in surface soils may be replaced gradually by
weathering, by recycling by deep rooted species, and by dust inputs if the
acidic inputs are reduced. Because basic cation depletion is the normal,
long-term trend in humid regions, the trend toward increased acidity would
probably not be reversed in such environments even if inputs stopped.
Because microbial activity in soils responds quickly to changing
environments, important soil processes it moderates can be expected to return
2-35
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to former levels when the environment changes as a result of reductions in
deposition.
Leaching of Al to aquatic systems in response to acidic inputs would likely
lessen with reduced acidic deposition.
2.3.5 Predicting Which Soils will be Affected Most
2.3.5.1 Soils Under Cultivation—It is unlikely that acidic precipitation
will adversely affect cultivated soils. Not only do many management prac-
tices result in acid production greater than that expected to be derived from
acidic deposition, but good management also requires controlling pH at a
level most conducive to plant growth (see Section 2.2.6). For example,
NH4+ is an important source of fertilizer N to soils. This form rapidly
oxidizes to N03~ in soil, resulting in significant acid production (see
Sectbn 2.2.1). Routine additions of N fertilizers may result in the release
of between one and two orders of magnitude more H+ than will be annually
derived from acidic deposition {McFee et al. 1976).
2.3.5.2 Uncultivated, Unamended Soils—As indicated in the soil chemistry
section, 2.2.1.3,arid orsemi-arid region soils that are not normally
leached do not naturally acidify, and adding acidic deposition will not
change that nor cause any foreseeable ill effects.
The soils that might be affected are those of the humid regions, which are
not normally amended with lime and/or fertilizers. This area includes most
of the forested land of the eastern United States, the Pacific Northwest and
some high altitude areas of the west. It is important to identify which
soils in these regions are likely to be adversely affected by acidic depo-
sition.
Various schemes for assessing site sensitivity to acidic deposition effects
have been proposed. Those directed toward aquatic effects have emphasized
bedrock geology (Hendrey et al . 1980, Norton 1980), while those concerned
with terrestrial effects have emphasized cation exchange capacity and base
saturation (McFee 1980, Klppatek et al. 1980). For the reasons previously
discussed, sulfate adsorption capacity should be included in the sensitivity
criteria for both aquatic and terrestrial impacts (Johnson 1980) , but unfor-
tunately, the data base for the latter is limited. In considering soil
sensitivity to adverse effects of acidic deposition, it is helpful to sepa-
rate the effects into two categories: (1) changes related to soil pH-basic
cation changes, which would include any direct losses of nutrients and
changes in processes or availability related to pH; (2) changes in soil
solution and/or leachate chemistry that might affect aquatic systems or be
toxic to plant roots, for which the primary concern is change in aluminum
concentration in solution.
McFee (1980) has suggested that cation exchange capacity (CEC) be used as the
primary criterion for determining soil sensitivity to acidic deposition. The
suggested classification considers soils with CEC greater than 15.4 meq 100
g~i, those subject to frequent flooding, or those with free carbonates in
the upper 25 cm of the solum to be insensitive. Non-calcareous, non-alluvial
2-36
-------
soils with CEC between 6.2 and 15.4 meq 100 g-1 are classed as slightly
sensitive, and those with CEC less than 6.2 meq 100 g'1 are classified as
sensitive.
Wiklander (1974, 19805) proposed a more complex classification system, which
considers soil buffering capacity as well as the ability of H+ to compete
for exchange sites in low pH, low base saturated soils. Buffering capacity
will, of course, be directly affected by CEC as well as by pH, base satura-
tion, and the presence of carbonates and ferromagnesium minerals. Consider-
ing base saturation separately recognizes that H+ competes best with base
ions on pH-dependent charge sites (Snyder et al. 1969, McLean and Bittencourt
1973). As base saturation decreases and a larger proportion of the pH-
dependent charge sites are filled with acidic ions, H+ inputs become less
effective in removing basic cations.
Wiklander1s classification scheme still does not include all known factors
that moderate effects of acidic deposition. For example, Wiklander (1975,
1980a,b) demonstrated that the presence of neutral salts, either in the
precipitation or in the soil, significantly moderates the effect of acidic
precipitation on soil. Sulfate adsorption capacity of the soil should also
be considered because mobile sulfate serves as a counter ion for cation
leaching (Cronan et al. 1978, Johnson 1980). Many acid soils have an anion
retentive capacity which can be related to both the presence of hydrated Fe
and Al oxides and to charge of the soil with decreased pH (Wiklander 1980a) .
High sulfate adsorption capacity will decrease soil sensitivity to cation
removal.
Comparisons of above systems indicate weakness in all, but a tendency to
agree when viewed on a national scale. The regions dominated by Ultisols,
Spodosols and some of the Inceptisols (Figure 2-4) encompass most of the
areas predicted to be sensitive to acidic deposition. All mapping efforts at
any level above the most detailed (county soil maps for example) will of
necessity include a wide range of conditions within any map unit. For that
reason, all of the efforts published thus far should be used with some
caution.
2.3.5.2.1 Basic cation-pH changes in forested soils. Based on the sensi-
tivity criteria proposed by MeFee(1980),Wiklander (1980b), and Johnson
(1980), it is clear that soils likely to undergo significant changes in basic
cation content or change in pH have these characteristics:
(1) they are not renewed by flooding or other processes;
(2) they are free of carbonates to considerable depth (1.0 meter or
more);
(3) they have low CEC but pH of at least 5.5 to 6.0; and
(4) they have a low sulfate adsorption capacity.
Because soils with low CEC (< 6.0 meq 100 g-1, McFee 1980, Klopatek et al.
1980) in humid climates tend to become acid naturally over time, few soils
2-37
-------
Figure 2-4. Generalized soil map of the United States (Soil Survey Staff 1975) show-
ing regions dominated by suborders or groups of suborders. The most
common suborder is named. Many other suborders exist within the bound-
aries of each area.
ro
i
CO
00
Alfisols
Al Aqua!fs
A2 Boralfs
A3 Udalfs
A4 Ustalfs
A5 Xeralfs
Aridisols
Dl Argids
D2 Orthids
En ti sols
El Aquents
E2 Orthents
E3 Psamments
Histosols
HI Hemists
H2 Hemists and Saprists
H3 Fibrists, Hemists, and Saprists
Inceptisols
II Andepts
12 Aquepts
13 Ochrepts
14 Umbrepts
Mollisols
Ml Aquolls
M2 Borol1s
M3 Udol1s
M4 Ustolls
M5 Xerolls
Spodosols
SI Aquods
S2 Orthods
Ul ti sol s
Ul Aquults
U2 Humul ts
U3 Udults
Vertisols
VI Uderts
V2 Usterts
-------
33IMI1S NO!W»3INOO HOC
S31V1S Q31INO 3H1 JO dVW 1IOS 1VH3N33
-------
meet criterion 3 above. So few have, in fact, that by the time we apply the
other criteria, it is clear that accelerated loss of basic cations and
lowered soil pH as a result of acidic deposition are unlikely to be extensive
problems. Maps prepared by Olsen et al. (1982) show areas of low CEC and
moderately high pH that are extensive enough to appear on a national map only
in the central portion of the United States. In that area, however, most
soils do not meet criterion 2 and do not currently receive significant acidic
deposition.
2.3.5.2.2 Changes in aluminum concentration in soil solution in forested
soils. Based on the discussion of soil chemistry in Section 2.2.3, it is
clear that soils most likely to have increased Al in solution or in leachate
due to acidic deposition are already acid, (pH < 5.5), and meet criteria 1,
2, and 4 above. Cation exchange capacity is not as important in this case,
but effects will be most pronounced where CEC is low. In such soils, the
buffer capacity is largely controlled by Al-mineral chemistry. Increased
acidic inputs may increase the rate of Al release and increase its concen-
tration in soil solution or leachate from the soil. This is most likely to
occur where total quantity of the controlling Al compounds exposed to chemi-
cal action is small, e.g., in a coarse-textured acid soil.
2.4 EFFECTS OF ACIDIC DEPOSITION ON SOIL BIOLOGY
2.4.1 Soil Biology Components and Functional Significance
The biological component of soil is of primary importance in the functioning
of the complete ecosystem. In this section, the soil biota will be briefly
described in terms of functional significance. For general reference, see
Alexander (1980a), Richards (1974), or Gray and Williams (1971).
2.4.1.1 Soil Animals--The most significant roles played by the invertebrate
soil fauna pertain to turnover of organic material and soil physical charac-
teristics. Many members of this group, such as earthworms, mites, ants, and
termites are involved in mixing the organic and inorganic soil constituents.
The quantity of organic material actually assimilated by these organisms is
small, generally less than 10 percent, but the relatively large quantity of
material consumed is frequently altered chemically by enzymes or micro-
organisms present in the animal's gut. Thus, by maceration and mixing, these
organisms play an important role in the conversion of plant material to soil
humus.
2.4.1.2 A1gae--Ch1orophyta (green algae), Cyanobacteria (blue-green algae)
and Chrysophyta (diatoms) are common inhabitants of the soil surface. Since
algae are dominantly photoautotrophic organisms (using light as an energy
source and C02 as a carbon source) they can colonize environments lacking
the organic carbon required by many life forms. In areas where higher life
forms are largely absent, such as fresh volcanic deposits, beach sands,
eroded areas, and freshly burned areas, algae commonly appear as the pioneer-
ing species, frequently supplying the organic material required for subse-
quent colonization by other life forms. Some blue-green algae (bacteria) can
convert atmospheric Ng to organic compounds. In many environments, such as
flooded paddy fields, this ability to fix nitrogen provides a critical input
2-40
-------
of nitrogen to the system. Lichens, an intimate association between certain
algae and fungi, are also important pioneering species, and some have the
ability to fix nitrogen. Ubiquitous on rock surfaces and other extremely
harsh environments, lichens are instrumental in the long-term breakdown and
dissolution of rocks ultimately to form soil.
2.4.1.3 Fungi--Soil fungi are involved in degrading a wide range of organic
compounds, from simple sugars to complex organic polymers. Many members of
this group possess the enzymatic capacity to attack the major plant consti-
tuents, such as cellulose, hemicellulose, and lignin. Fungi are normally the
dominant initial colonizers of plant debris and are ultimately responsible
for many of the steps occurring during the conversion of plant material to
soil organic matter. The complex network of fungal hyphae which totally
permeates the fabric of soil constitutes a major portion of the soil biomass
as well as binding together soil particles to form aggregates. Products of
fungal metabolism in soil, such as carbohydrates, can act as glues for
primary soil particles.
Certain types of soil fungi can play direct roles in nutrient availability to
plants by forming mycorrhizal associations with plant roots. The fungal
hyphae greatly expand the volume of soil from which plant roots can effec-
tively draw nutrients. In deficient soils, the fungal partner can substan-
tially improve phosphorus, copper, zinc, and possibly nitrogen (ammonium)
availability to plants. In addition, the mycorrhizal association may enhance
water availability, increase salt tolerance, enhance heavy metal resistance,
and affect plant growth via hormone production. Although relationships are
not yet well understood, each of these effects is currently under investiga-
tion.
2.4.1.4 Bacteria--The procaryotic microflora of soils are also extremely
important in the decomposition of plant litter and the synthesis and break-
down of soil organic matter. Bacteria are primarily responsible for making
organic forms of N, S, and P available to plants by mineralizing organic
matter. For substantial plant uptake to occur, S must be as S042~ and N as
either N03~ or NH4+. Oxidation of Nfy* to NOs" (nitrification) is dominantly
catalyzed by autotrophic soil bacteria. Nitrogen is lost from the soil
through anaerobic bacterial reduction of N03~ to the gaseous species N2
and N20 (denitrification). Most nitrogen enters ecosystems through bac-
terial reduction of atmospheric N2 to NH4+ (^-fixation). Fixation
by bacteria living symbiotically with plants can contribute significant
amounts of nitrogen to both agricultural and forest systems. Nitrogen nutri-
tion of many leguminous plants is enhanced through N2~fixation by bacteria
of the genus Rhizobium. Fixation by actinomycetes, such as Frankia, in asso-
ciation with woody species may contribute critical amounts of nitrogen to
some forest systems. The oxidation and reduction of S roughly parallel that
of N. In addition to bearing primary responsibility for the availability of
N and S to plants, soil microbes also strongly influence the avail- ability
of phosphorus, iron, and manganese through organic mineralizations and redox
reactions.
The distribution of microbial activity in soil generally reflects the fact
that many of these microbes are heterotrophs, that is, they require preformed
2-41
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organic compounds. Soil microbial activity is generally greatest in regions
of high organic carbon availability. While most types of microbial activity
do occur to some extent throughout the soil profile, recognizing that maximal
activity commonly occurs in somewhat discrete areas of the soil is important
to understanding potential effects of acidic deposition. Microbial attack on
plant debris takes place largely in the surface litter layer. Production and
breakdown of soil humus occur dominantly in the upper portion of the soil
profile, reflecting the site of initial leaf, stem, and root material depo-
sition. Heterotrophic microbial activity is also high in soil near plant
roots, where root-derived material provides carbon for soil bacteria and
fungi.
2.4.2 Direct Effects of Acidic Deposition on Soil Biology
The effects of acidic deposition should be expected to vary tremendously,
depending on the type of organism and the characteristics of the soil which
it inhabits. While soil acidification does affect many biological processes,
it is often impossible to distinguish direct effects of acidification from
secondary effects resulting from acid-induced changes in the soil solution.
The following section documents some effects which have been attributed to
soil acidification resulting from acid inputs.
2A.2.1 Soil Animals--Many classes of soil animals, such as earthworms
(Lumbricidae), millipedes (Myriapoda), and nematodes (Nematoda), are known to
be less abundant in acid soils than in neutral soils. However, large popu-
lations of other soil animals, such as springtails (Collembola) and potworms
(Enchytraeidae), are common in acid soils high in organic matter (Richards
1974).
Effects of simulated acid precipitation on soil fauna vary markedly according
to the species observed. Studies by Baath et al. (1980), in which soils were
treated with 50 or 150 kg ha"1 ^$04 for 6 years, showed that the num-
bers of Collembola increased, Enchytraeidae decreased, but mites (Acarina)
were generally unaffected by both application rates. In a 2-year exposure to
simulated rain of pH 2.5 to 6.0 (25 or 50 mm per month), Collembola, Acarina,
and Enchytraeidae were generally unaffected or increased in number with
the acid treatments. However, a few species of Acarina and the dominant
Enchytraeid were significantly reduced by the more extreme acidification
(Hagvar 1978, Abrahamsen et al. 1980). It should be noted that the soils
studied by these two groups were naturally very acidic; hence the indigenous
soil fauna may have been relatively acid tolerant. In less acid deciduous
woodland soils (Kilham and Wainwright 1981), the native population of soil
animals appeared to be much more sensitive to acid rain (pH 3.0) localized
near a coking works, but these results also reflect the presence of
substantial dry deposition on the litter.
2.4.2.2 Terrestrial Algae--While green algae (Chlorophyta) readily colonize
relatively acid soils, blue-green algae (Cyanobacteria) have been reported to
be particularly sensitive to soil acidity (Dooley and Houghton 1973, Wilson
and Alexander 1979). While there is little experimental verification in soil
systems, the general sensitivity of free-living Cyanobacteria to acidity
suggests they may be susceptible to acidic deposition. The sensitivity of
2-42
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blue-green algae to acid precipitation has been demonstrated in a lichen
symbiosis. Simulated acidic deposition of pH 4.0 or less substantially
reduced ^-fixation by the dominant ^-fixing lichen in a deciduous
forest (Denison et al. 1977).
2.4.2.3 Fungi—Fungi become increasingly important in acid soils as compared
to neutral-alkaline soils (Gray and Williams 1971). The commonly observed
dominance of fungi over bacteria in acid soils may, in part, result from a
greater sensitivity of heterotrophic bacteria to H+ concentration and the
consequent reduction in competition (Alexander 1980a).
The relative tolerance of fungi to acid precipitation was demonstrated by
Wainwright (1979), who isolated fewer heterotrophic bacteria but more fungi
from soils exposed to acid rain and heavy atmospheric pollution than from
similar but unaffected soils. The presence of nitrifying fungi in acid soils
lacking autotrophic nitrifiers (Remacle 1977, Johnsrud 1978) also appears to
indicate the relative resistance of fungi to soil acidity.
Most investigations of the effects of acidic deposition on soil fungi, how-
ever, have used traditional plate count methods, which do not necessarily
reflect viable fungal biomass. Baath et al. (1980) found that FDA (fluores-
cein diacetate) active fungal biomass decreased significantly under the two
acid regimes described earlier (Section 2.4.2.1) while total fungal mycelia
(the sum of viable and non-viable hyphae) increased.
To date, little information available concerns the response of mycorrhizal
associations to acidic deposition. Sobotka (1974) reported a reduction in
the fungal mantle of spruce mycorrhizae receiving heavy atmospheric pollu-
tion, including acid rain. In a short-term experiment, Haines and Best
(1975) found no visible damage to endomycorrhizae of sweetgum exposed to pH
3.0 treatments. To explain deviations in nutrient flux data, these research-
ers suggested that cation carriers of mycorrhizal roots may be more suscep-
tible to inhibition by H+ than are non-mycorrhizal roots.
2.4.2.4 Bacteria—The discussion in this section pertains largely to soil
bacteria. In many soil microbial processes, however, it is impossible or
meaningless to isolate bacterial functions from soil fungal and faunal
processes with which they are inherently integrated. For example, leaf
litter decomposition requires fungal, bacterial, and faunal attack.
Bacteria are generally considered to be less acid tolerant than fungi. Some
bacteria, however, are extremely acid tolerant. For example, species of the
chemoautotrophic thiobacilli can survive at pH 0.6 and thrive at pH 2.0
(But!in and Postgate 1954).
Acidic deposition may affect heterotrophic bacteria in soil by causing
changes in total numbers and/or species composition. Francis et al . (1980)
reported that the total number of bacteria and actinomycetes generally
declined in soil acidified from pH 4.6 to 3.0 with an addition of ^$04,
although the magnitude of these effects was not reported. In soils trans-
ferred to a site receiving pH 3.0 rain and dry deposition, Wainwright (1980)
2-43
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found that over a 1-year period bacterial numbers did not change signifi-
cantly, even though the soil pH fell from 4.2 to 3.7. Baath et al. (1980)
noted a shift towards spore-forming bacteria in soils receiving HgSOd
inputs for 6 years as compared to control soils, suggesting a response to
adverse conditions. In the same experimental series, total bacterial numbers
(by plate counts) did not change, but bacterial biomass and FDA-active
bacteria did decrease with increasing severity of treatment (Baath et al.
1979, 1980).
2.4.2.5 General Biological Processes—Net heterotrophic activity (bacterial,
fungal, and faunal) and the rate of organic matter decomposition are commonly
determined by measuring CO? evolution. The rate of glucose mineralization
was reduced in surface soils receiving 100 cm of simulated rain (pH 3.2 and
4.1), continually or intermittently, over a 7-week period (Strayer and
Alexander 1981). However, the 7-week treatments caused less significant
effects than did the continuous exposure, and the reductions were less severe
in soils of greater natural acidity. The authors therefore suggested that
some microbial adaptation was occurring over time.
Respiration in soils transferred to a site receiving pH 3.0 rain was reduced
by 50 percent after a one-year exposure (Wainwright 1980). Similar effects
of simulated acid precipitation have also been reported by Tamm et al.
(1977). Observed effects of simulated acid precipitation on litter decompo-
sition are summarized in Section 2.5.
Several reports now indicate that acid inputs can slightly accelerate miner-
alization of organic nitrogen (Wainwright 1980, Strayer et al. 1981) Tamm et
al. (1977) similarly found increased accumulation of NH4+ in acid-treated
humus samples, but they interpreted this to mean that immobilization was re-
tarded more than mineralization (a hypothesis for which no substantiating
data existed). Conversely, Francis et al. (1980) found lower Nfy* pro-
duction in a soil that had received an addition of (^$04. For all of
this work, the treatment periods were relatively short (from 1 hour to 1
year); longer exposures may yield more consistent results. The data, how-
ever, are compatible with the fact that "natural" soil acidity does not have
a uniform effect on N-mineralization (Alexander 1980b).
Because nitrification is generally believed to be catalyzed by relatively few
types of autotrophic nitrifiers (known to be acid-sensitive on laboratory
media), researchers have predicted that this process should be one of the
microbial processes most sensitive to acid precipitation (Tamm 1976,
Alexander 1980b). While evidence indicates that acid inputs to soil inhibit
autotrophic nitrification, the overall effects on NH4+ oxidation to
N03~ are neither uniform nor easily interpreted. Francis et al . (1980)
could detect little nitrifying activity in the naturally acid forest soil
studies (pH 4.6) or in the soil sample that had received an addition of
H2S04, but they concluded that further acidification of an acid forest
soil would lead to a significant reduction in nitrification. Wainwright
(1980) found essentially no effect on nitrifying activity in a soil exposed
to acid rain (pH 3.0) from a coking works. Strayer et al. (1981) examined
the effects of acute acidification on nitrification in surface soil from soil
2-44
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columns and found interesting but somewhat complex results. When high
NH4+ amendments (100 ppm N) were added to the nitrification assay, all
acid treatments tested (pH 3.3 to 4.1) caused substantial reductions in
nitrification rates. However, when NH4+ was not added to the soil, the
acid treatments caused no detectable effect, or in some cases, caused a
slight stimulation in N03~ production. Because forest soils would be
expected to have relatively low natural concentrations of NH4+, the
authors conclude that short-term exposures to acid rain should not
substantially affect nitrification in forest soils. The results reported by
Strayer et al . (1981) are consistent with the occurrence of heterotrophic
nitrifying organisms in naturally acidic forest soils; these heterotrophic
nitrifiers are considered much less sensitive to acidity than are autotrophic
nitrifiers (Remacle 1977, Johnsrud 1978).
Few published data concern effects of acidic deposition on soil denitrifi-
cation. While slight soil acidification may not alter the overall rate of
this process, it should be expected to increase N20 production relative to
N2 (Firestone et al. 1980).
A substantial amount of work on the sensitivity of ^-fixation by
legume-Rhizobium associations to soil acidity has been published. In some
cases, the bacterial symbiont appears to be sensitive to acidity (Bromfield
and Jones 1980, Lowendorf et al. 1981); in other cases, the nodule formation
or activity are affected (Evans et al. 1980, Munns et al. 1981). However,
work on the effects of acidic deposition on ^-fixation by legumes is
scant. Shriner and Johnston (1981) reported that simulated rain of pH 3.2
applied for 1 to 9 weeks caused decreased nodulation in kidney beans. The
authors suggest that similar effects would be unlikely to occur under normal
agricultural management practices but might be expected to occur in natural,
unmanaged ecosystems (Shriner and Johnston 1981). No data are available
concerning effects of acid rain on the associations of actinomycetes with
woody plants.
2.4.3 Metals—Mobilization Effects on Soil Biology
Two questions concerning mobilization of metals and effects on soil biology
must be addressed. First, the input of acidity to soil can cause mobiliza-
tion of Al and Mn from mineral forms indigenous to the soil. Can mobili-
zation of Al and Mn by acid inputs be expected to have toxic effects on the
soil biota? Second, acidic deposition is sometimes accompanied by atmospheric
deposition of various heavy metals. Does the acidity of the rain increase
the potential toxicity of these metals? While few data available directly or
realistically address these potential effects of acidic deposition, a small
body of pertinent background literature exists.
The toxicity of available Al to soil microbial activity has been reported by
Mutatkar and Pritchett (1966), who found that additions of Al to soils with
pH maintained below 4.0 created exchangeable Al levels of 1 ug g-1 or
higher and significantly reduced the rate of soil respiration. Ko and Hora
(1972) have identified Al3+ ions as being fungitoxic in acid soil extracts.
These workers found germination of ascospores to be totally inhibited by
aqueous solutions (pH 4.8) containing as little as 0.65 ppm Al . They did not
2-45
-------
identify Mn as toxic to the fungi tested, but the concentrations of this
metal in the soil extracts examined were low compared to Al concentrations.
In studies dealing with the growth of the Rhizobiurn-bean symbiosis in acid
tropical soils, Dobereiner (1966) found that additions of 40 ppm Mn to acid
soils reduced either ^-fixation efficiency or nodule numbers. Since
preliminary evidence suggests that the threshold concentrations for toxicity
of mobilized aluminum are relatively low, such an indirect consequence of
acid input to soil may be a possibility. However, acid rain, within current
pH limits, has not been shown to mobilize these metals in quantities toxic to
soil biota.
Soils in the vicinity of metal-smelting and coal-burning are likely to be
subject to atmospheric deposition of heavy metals (Little and Martin 1972,
Freedman and Hutchinson 1980) in addition to acidic deposition. The input of
heavy metals to these soils is significant because metal solubilization and
biological toxicity are pH dependent. Numerous pure culture studies demon-
strate increasing metal toxicity with decreasing pH of solution (e.g., Babich
and Stotzky 1979). However, many of these studies should not be extrapolated
to soils because of the complexity of the metal cation interactions with soil
constituents. Babich and Stotzky (1977) found that Cd toxicity to microbes
in soil was a function of soil pH; however, this may have been an anomaly,
since toxicity increased with increasing soil pH.
Metals vary in potential toxicity; work by Somers (1961) indicated that the
microbial toxicity of heavy metals is highly correlated with the electro-
negativity of the metal. When attempting to assess the potential effects of
acidic deposition in association with metal deposition, one must consider
several factors: 1) the toxicity potential of the metal, 2) the quantities
and speciation of metals deposited and degree of association with acid
inputs, and 3) the pH dependence of metal toxicity in the recipient soil
environment. Mobilization of metal ions in soils receiving acid inputs, and
subsequent toxicity of these metals, may be a mechanism by which acidic
deposition affects soil biological activity; but experimental evidence is
lacking.
Apparently certain plant-microbial associations are able to protect plants
from metal toxicity. Bradley et al . (1981) found that mycorrhizal infection
of an ericaceous, Gal lima species reduced heavy metal uptake by the plant.
The authors suggested that protection by the fungal symbiont allowed this
species to colonize heathland soils in which the low pH increases avail-
ability of metal cations to levels which are toxic to many non-ericaceous
species.
2.4.4 Effects of Changes in Microbial Activity on Aquatic Systems
Because our current understanding of the effects of acidic deposr
microbial activity in terrestrial ecosystems is limited, extrapolat
ition en
ations to
possible secondary effects on aquatic systems are tenuous at best. It is
important to recognize, however, that even a small change in microbial
activity in soil may cause profound changes in aquatic systems, into which
much of the soil water will ultimately drain.
2-46
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2.4.5 Soil Biology Summary
The following statements represent simplifications of complex and sometimes
contradictory trends in the existing data. They reflect both the complexity
of microbial processes and the variability in experimental protocols. The
extreme variability in pH and ionic composition of simulated rain, as well as
differences in important soil characteristics, makes comparing data diffi-
cult. Treatment durations in the experiments reported ranged from 1 hour to
6 years. Short-term "accelerated" treatments may not only overlook potential
long-term effects, but also may yield misleading predictions. The short-
comings of long-duration experiments involving infrequent sampling should
also be recognized. Acid precipitation rarely occurs in isolation; rather,
it occurs in association with other pollutants such as heavy metals and the
gaseous precursors of acid species. The potential synergisms among these
pollutants should not be overlooked. The following statements summarize or
interpret the limited data available and should be read with the above-
mentioned limitations in mind.
Acidic deposition will not substantially affect soil biological activity in
cultivated soils because of the much greater influence of soil amendments.
The following statements pertain to uncultivated soil systems:
° The effects of acidic deposition on animals in strongly acid soils
are not significant. In less acid soils, pH 3.0 simulated rain has
produced significant changes in litter animals.
* Certain types of soil microbial activity are more sensitive to soil
acidity than are others. Soil fungi are probably the components of
the soil biota least sensitive to acid inputs; but little is known
about effects on mycorrhizal symbionts.
0 Preliminary evidence indicates that ^-fixation by lichens is
inhibited by rain of pH less than 4.0. The evidence for acidic
deposition influences on Rhizobium or actinomycete symbiotic
N-fixation is insufficient for a conclusion.
° Autotrophic nitrification in surface soils is reduced by artificial
acid inputs; however, no evidence exists to prove that acidic
deposition at the rates currently common in the United States will
cause such a decrease. Net nitrification may not be similarly
decreased because of the acid tolerance of heterotrophic
nitrifiers.
0 Slight increases and decreases in N-mineralization rates result from
treatments of short duration, but little direct evidence concerning
long-term responses to realistic inputs exists.
2.5 EFFECTS OF ACIDIC DEPOSITION ON ORGANIC MATTER DECOMPOSITION
One of the long-standing hypotheses regarding the environmental effects of
acidic deposition has been that increased acid loading to forest soils will
2-47
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TABLE 2-6. REVIEW OF STUDIES CONCERNED WITH THE IMPACT OF ACIDIC DEPOSITION ON ORGANIC DECOMPOSITION
Author
Duration
Soil Type of Treatments
Experiment
Results
1. Abrahamsen et al. 1980 Lodgepole pine needles 75-90 days
Norway spruce needles
3-9 mos.
ro
Co
Raw coniferous humus
unspecified
Needles from field experiments
at pH 5.6 and 3.0 were incu-
bated in moist condition and
weighed.
Spruce needles in lysimeters
were watered 2x weekly with
pH 5.6, 3, or 2 water at a
rate of 100 mm mo."1 or 200
mm mo. •
Raw humus in litterbags ex-
posed to pH 5.3, 4.3, and
3.5 treatments.
Acid treatment increased decom-
position-29% greater at pH 3
than 5.6
Relatively small effects from
acid treatments. No signifi-
cance at 100 mm mo."*. At
200 mm mo.-l, the pH 3 and 2
treatments decreased decompo-
sition by < 5%
Increased leaching of K, Mq, Mn,
Ca.
pH 4.3 treatment caused 8% de-
crease in decomposition rate,
while pH 3.5 caused 10%
decrease.
2. Abrahamsen and Dollard 1978 General Review
3. Abrahamsen et al. 1976
Lodgepole pine needles 90 days
Needles moistened with di-
lute H2$04 solutions.
Cellulose/Wood
Unspecified
Decomposition of organic matter
in acidic coniferous forest
soils is apparently only
slightly sensitive to acidifi-
cation. Decomposition of
fresh litter and cellulose is
influenced only at pH _< 3.
Decomposition was depressed at
pH 1.8 as compared to 3.5. No
difference between pH 3.5 and
4.0
Unspecified acid treatments No consistent trends.
-------
TABLE 2-6. CONTINUED
Author
Soil Type
Duration
of
Experiment
Treatments
Results
4. Alexander 1980a
Strayer and Alexander
1981
Honeoye silt loam (pH 7.1) 2+ wk. Soils were exposed to pH 4.1 and pH 4.1 treatment had no
3.2 acid rain treatments and
were incubated with
glucose.
effect on glucose
mineral ization
pH 3.2 treatment decreased
glucose mineralization rate
by 30-66%.
5. Alexander 1980b
Spodosols from the 14-61
central Adirondacks days
ro
-------
TABLE 2-6. CONTINUED
Author
Soil Type
Duration
of
Experiment
Treatments
Results
8. Cronan 1980b
Coniferous and hardwood
forest floors
3 mo.
9. Hovland 1981
ro
i
en
O
Norway spruce needle
litter
5 yr.
Forest floor microcosms were
subjected to weekly 3.5 cm
simulated rains at pH 5.7
and 4.0
Field plots were exposed to
pH 6.1, 4.0, 3.0, and 2.5
rains over 5 yr. Litter
collected from these plots
was assayed.
Hardwood forest floors showed
60% more Ca leaching and 65%
more Mg leaching at pH 4.0.
Coniferous forest floors
showed 40% more Ca and 25%
more Mo. leaching at pH 4
compared to pH 5.7. In
general, cation fluxes from
the hardwood litter were much
greater than from coniferous
litter.
Acid rain treatments produced
very little effect on biolo-
gical activity in litter as
measured by respiration and
eellulose activity.
10. Hovland et al. 1980
Norway spruce needles 16-38 wk.
Lysimeters containing spruce
needles were exposed to pH
5.6, 3.0 and 2.0 solutions
at 100 and 200 mm mo"*.
Small effects on decomposition.
Treatments at pH 3 and 2 ini-
tially increased the decompo-
sition rate at 100 mm mo"1.
After 38 wk., decomposition
had decreased relative to
controls in pH 3 and 2 treat-
ments at 200 mm mo"*.
Effect of acid treatments on
monosaccharide content was not
consistent. However, there
was an indication of reduced
lignin decomposition at 200
mm mo~l for pH 3 and 2.
Acid treatments caused increased
leaching of Mg, Mn, and Ca.
Initially, acid rains decreased
P leaching; later, this
reversed.
-------
TABLE 2-6. CONTINUED
ro
en
Author
duration
Soil Type of Treatments
Experiment
Results
11. Francis et al. 1980
12. Lohm 1980
13. Roberts et al. 1980
14. Tamm et al. 1976
Oak-pine sandy loam (pH 4.6) 5 mo.
Coniferous iron Podzol
Coniferous Podzol
Coniferous Podzol
6 yr.
5 mo.
5-6 yr.
Soils were adjusted with
acid or base to give a
soil pH of 3.0 or 7.0,
and were then incubated
with controls.
Plots were exposed to 0,
50, and 150 kg ha'1
H2$04 per yr.
Litter bags were exposed
for 2 yr.
Field plots were subjected
to biweekly 5 mm appli-
cations of pH 3.1 and
2.7 acid rain.
Field plots received 0,
50, and 100 kg ha"1
yr~l applications
of H2S04.
The acidified soil showed 6-52%
less C02 production, depend-
ing upon amendments.
Acid treatments lowered the
decomposition rate by 5-7%.
No significant effect of acid
treatments on respiration.
Litterbags showed significant
increase in weight loss (15%)
with increased acidity.
Found decreased C02
respiration with increased
H2S04.
-------
result in decreased decomposition rates for organic matter. This hypothebi.
has been addressed by a number of investigators (Tamm et al. 1977; Abrahamsen
et al. 1976, 1980; Abrahamsen and Dollard 1978; Alexander 1980a,b; Baath et
al. 1980; Croran 1980a,b; Hovland et al. 1980; Francis et al. 1980; Lohm
1980; Roberts et al. 1980; Hovland 1981; Kilham and Wainwright 1981; Strayer
and Alexander 1981; Strayer et al. 1981). Unfortunately the results from
these studies have appeared mixed and inconsistent (Table 2-6). However, if
one screens the published studies and selectively excludes the results from
those investigations that represent extremely acute treatments, then the fol-
lowing summary statements emerge.
(1) Most decomposition studies related to acidic deposition have been
conducted with coniferous litter materials.
(2) Results suggest that it is important to interpret data from
decomposition studies in relation to H+ loading and not simply
with respect to the pH of the artificial rain treatments.
(3) It is important to distinguish between the physical-chemical and
the biological components of organic decomposition. Based upon
shorter-term studies (2 to 4 months or less), it has been shown
that increased H+ loading generally will increase leaching of
cations and organic constituents from forest litter. This re-
sponse may help to explain why acidic precipitation treatments
increase the initial rate of weight loss in some experiments. Over
the longer term (> 4 months), it appears that the biologi-
cally-mediated mineralization of organic matter in forest soils
will be only slightly inhibited by acidic deposition (< 1 to 2
percent decrease in decomposition rate).
(4) Overall, unless average precipitation inputs were to drop to pH
3.0 or below, one would not expect significant impacts of acidic
deposition on litter decomposition.
2.6 EFFECTS OF SOILS ON THE CHEMISTRY OF AQUATIC ECOSYSTEMS
Much of the evidence for atmospheric depositions' contribution to surface
water acidification, while convincing in many cases (e.g., Johnson 1979), is
circumstantial. Only recently have efforts been made to establish the
mechanisms by which atmospheric acid inputs are transferred to aquatic
ecosystems (Abrahamsen et al. 1979, Seip 1980, N. M. Johnson et al. 1981).
If acidic precipitation passes through soil prior to entering an aquatic
ecosystem, it will usually be strongly influenced by the chemical nature of
the soil. Even barren rock has some influence on the chemistry of runoff
water (Abrahamsen et al. 1979). The pH of water leaving the soil is not
necessarily the same as the soil solution pH in intimate contact with the
soil.
Rosenqvist (1977, 1978, Rosenqvist et al. 1980) has argued that the influence
of soil and bedrock on the chemistry of waters is overwhelming and that the
pH of runoff water would be the same whether snowmelt was acid or neutralized
by a suitable base. Seip et al. (1980) carried out an experiment to test
2-52
-------
Rosenqvist's hypothesis by applying NaOH to one of the mini-catchment water-
sheds in Norway; results showed that, indeed, the neutralization of snow with
NaOH had little effect on runoff pH. The investigators attributed the lack
of effect to differences in weather conditions and Na content of the depo-
sition.
Seip (1980) presented a hypothesis for surface water acidification which has
met with agreement among soil scientists as to its mechanism but not
necessarily to its magnitude. This has been termed the "mobile anion
mechanism." In essence, it states that the introduction of a mobile anion
into an acid soil will cause the pH of a soil solution to drop. This is
because of the requirement for cation-anion balance in solution and because
most exchangeable cations in acid soils are H+ and Al3"1". Thus, due to
cation exchange processes and the requirement for cation-anion balance,
increased anion concentration in an acid soil solution causes increased H+
and A13+ concentrations, regardless of whether the anion is introduced as a
salt or an acid. This mechanism has been known to soil scientists for de-
cades as the "salt effect," wherein soil pH is usually more acid in CaCl2
solutions than in ^0 (Yuan 1963). Field studies have confirmed that this
mechanism is valid (Abrahamsen et al. 1979; Seip et al. 1979a,b, 1980;
Abrahamsen and Stuanes 1980). However, doubt remains as to whether the
magnitude of pH change this mechanism can produce could cause the pH changes
reported for acidified surface waters (Abrahamsen and Stuanes 1980; Johnson
1981; Rosenqvist 1981, pers. comm.). It is clear, however, that neutral
salts can, when added to an acid soil, cause a flux of Al in a low-pH
solution to streams.
Natural acid production, changes in land use patterns, and management prac-
tices such as harvesting, burning, and fertilizing are suggested alternative
sources for surface water acidification (Rosenqvist 1977, 1978; Patrick et
al. 1981). These possibilities have been explored to some extent in southern
Norway, but we have no concrete evidence that changes due to harvesting and
land use have caused surface water acidification (Drablj6s et al. 1980)
although the debate continues. Evidence suggests, however, that fish kills
associated with acidic pulses have been occurring in at least one place in
southern Norway (Roynelandsvann) since the 1890's (Torgenson 1934). In this
instance, liming was successful as a mitigative measure for short-term
effects on fish populations (Abrahamsen, pers. comm.). The causes of these
acid pulses are unknown, but presumably acid rain effects were much smaller
nearly a century ago.
Some attention has been given to neutralization processes affecting acid rain
as it passes through terrestrial to aquatic ecosystems. N. M. Johnson et al.
(1981) found a two-stage process operative in the Hubbard Brook, NH ecosystem
in which H+ in acid rain is initially neutralized by dissolution of reac-
tive alumina in the soil before both H+ and A13+ are neutralized by
chemical weathering of alkali and alkaline earth minerals in bedrock.
Because stage 2 proceeds more slowly than stage 1, first- and second-order
streams may contain H+ and A13+, but neutralization is usually complete
before surface waters reach third-order streams.
2-53
-------
Kilham (1982) reports a case in which deposition appears to have caused an
increase in lake alkalinity. Alkalinity in Weber Lake, Michigan, has in-
creased two-fold over the last thirty years, and theoretical considerations
of acid-base budgets lead to the hypothesis that this alkalization has
resulted from plant nitrate uptake, bacterial sulfate reduction, and
carbonate mineral weathering, all enhanced by acid precipitation. This
effect, while no more desirable than acidification, contradicts the
assumption that acid rain always causes surface water acidification and is
ample testimony to the complexity of terrestrial-aquatic interactions.
Kilham (1982) indicates that alkalization is likely only in lakes of high
alkalinity with abundant carbonates in the watershed.
In view of the lack of understanding of terrestrial-aquatic transport
processes, assigning "sensitivity" ratings to acid deposition on a regional
scale is premature. Nonetheless, agencies alarmed by reports of ecological
effects of acid precipitation insist upon knowing something about the
geographical magnitude of the acid rain "problem," and scientists must make
their best guesses as to appropriate criteria, even though the mechanisms are
not completely understood. This situation reflects a gap in understanding
and a critical research need that encompasses not only soil and bedrock
chemical reactions but also hydrological processes. Recent studies have
shown the important contribution of variable source areas (i.e., portions of
watershed landscapes that contribute to streamflow during storm events) to
surface waters and their chemical composition during stormflow (Henderson et
al. 1977, Huff et al. 1977, Johnson and Henderson 1979).
Similarly, water flow through soil macropores (see Figure 2-1) can be a very
important component of soil water flux during periods of saturated flow
(Luxmoore 1981). Both variable source areas and macropore flow reduce the
amount of contact between soils or bedrock and waters passing through
terrestrial ecosystems. Integrated studies of terrestrial-aquatic transport
processes involving both hydrological and chemical components are essential
to an understanding of the effects of acid rain on aquatic ecosystems.
2.7 CONCLUSIONS
Effects of acidic deposition related to soils are in these general
categories: soil acidification, nutrient supply, Al and Mn mobility, and
microbial activity. The following conclusions, relative to these general
categories, can be drawn from Chapter E-2:
o Soils amended in agricultural practice will not be harmed by acidic
deposition (Section 2.3.5).
0 Soil acidification is a natural process in humid regions. It is
obvious that acidic deposition contributes to this process; how-
ever, at current levels, it is a minor contribution (Section
2.3.5).
° Most soils of low buffering capacity in areas of high rainfall are
already acid; therefore, few soils are likely to become perceptibly
more acid due to deposition. They are the soils that have low
2-54
-------
buffering capacity, a relatively high pH (slightly acid, pH 5.5 to
6.5), low sulfate adsorption capacity, no carbonates, and no basic
inputs (Section 2.3.5).
The availability of sulfur and nitrogen to plants will be enhanced
by their presence in the deposition. Because nitrogen limitations
are so common and cation limitations are so rare in forests of the
United States, it seems likely that HNOa inputs generally will be
beneficial. Exceptions may occur on sites with adequate or exces-
sive N supplies. Benefits of H2S04 deposition are probably
minimal, because S deficiencies are rare and probably easily
satisfied with moderate atmospheric S inputs (Section 2.3.2).
The long-term effect (i.e., over decades or centuries) of acidic
deposition can be expected to remove cations from forest soils,
but it is not clear whether this will reduce available cations and
enhance acidification of soils. For example, cation leaching
rates, although increased by acid precipitation, may remain
insignificant relative to total soil supplies and forest growth
requirements; furthermore, exchangeable cations may be replaced by
weathering from primary minerals at rates sufficient to maintain
their current status partially as a result of acid precipitation
inputs (Section 2.3.3).
Assessing acidic deposition effects on forest nutrient status
involves quantifying amounts of inputs involved and the S, N, and
cation nutrient status of specific sites. It cannot be stated
that forest ecosystems, in general, respond to acidic deposition
in a single predictable way. Indeed, the contrasting behavior of
Norway spruce in Germany and in Norway exemplifies the variable
response that can be expected from different sites (Section
2.3.3).
Aluminum toxicity may affect forests on already acid soils where
acidic deposition plus natural acidifying processes increase
acidity enough to cause a significant rise in Al availability. If
soil pH is low enough (< pH 5.0 to 5.5) in mineral soils to cause
the dissolution of Al- and Mn-containing minerals, H+ input will
increase release of Al and Mn to the soil solution (Section
2.3.3).
The increased mobility of Al in uncultivated, acid soils is prob-
ably the most significant effect of acidic deposition on soils as
they influence terrestrial plant growth and aquatic systems
(Section 2.3.3).
Short-term studies indicate that increased H+ loading will cause
increased loss of cations and organic components from forest
litter. Over the longer term, the biologically-mediated minerali-
zation of organic matter in forest soils will be only slightly
inhibited by acidic deposition (< 1 to 2 percent decrease in
2-55
-------
decomposition rate). Unless average precipitation inputs were to
drop to pH 3.0 or below, significant impacts of acidic deposition
on litter decomposition in natural systems are not expected
(Section 2.3.3).
Soil microbial activity may be significantly influenced near the
surface if inputs are great enough to affect pH or nutrient avail-
ability. Evidence for effects of acidic deposition on Rhizobium
or actinomycete symbiotic N-fixation remains inconclusive"Slight
decreases and increases in N mineralization rates result from
short-term acid inputs, but long-term responses are not docu-
mented. Important effects under field conditions have not been
clearly demonstrated (Section 2.4).
2-56
-------
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS
E-3. EFFECTS ON VEGETATION
3.1 INTRODUCTION
3.1.1 Overview (Eds.)
This chapter examines diverse plant-pollutant relationships to assess poten-
tial and recognized effects of acidic deposition as described in the extant
literature. Vegetation responses discussed include morphological and
physiological responses, species/varieties and life-stage susceptibilities,
disease and insect stresses, indirect effects of nutrient cycle alterations,
and crop and forest productivity.
Because of the close relationship between soils and plants, we must consider
how soil acidification affects productivity. It is important to recall the
following points from the previous chapter:
0 soils amended in agricultural practice will not likely be negative-
ly impacted by acidic deposition;
o soil acidification is a natural process in humid regions, so most
soils that are easily acidified are already acid; and
0 soils with low buffering capacity, relatively high pH, low sulfate
adsorption capacity, no carbonates, and no basic inputs are sus-
ceptible to increased acidification rates from atmospheric inputs
of acidic and acidifying substances.
With these points understood, Chapter E-3 will deal with the direct effects
of acidic deposition on plant response, and the interactive effects of acidic
deposition with other factors, such as other pollutants, insects, pathogens,
and pesticides.
Given the uncertainty still surrounding effects on plant productivity, how-
ever, this document does not attempt to make economic assessments of recog-
nized or potential damage to vegetation; nor does it consider mitigative
measures to counter acidic deposition inputs to plant systems. Discussions
of nutrient cycling and forest productivity are included in both this chapter
and the soils chapter, from slightly different perspectives. Both chapters
should be read carefully to gain a more complete understanding of the issues.
3.1.2 Background (P. M. Irving and S. B. Mclaughlin)
The observation that both gaseous and rain-borne pollutants affect vegetative
growth is not limited to recent years. Robert Angus Smith (1872) in his
3-1
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manuscript, "Air and Rain: The Beginnings of a Chemical Climatology," in-
cluded a section on "Effect of Acid Gases on Vegetation and Capability of
Plants to Resist Acid Fumes." As early as 1866 the Norwegian playwrite Ibsen
(1866) referred to the phenomenon in the drama "Brand":
"... A sickening fog of smoke from British coal
Drops in a grimy pool upon the land,
Befouls the vernal green and chokes to death
Each lovely shoot, .. ."
Of course the fog of smoke referred to by Ibsen was from imported British
coal and not from the long-range transport of pollutant gases. An intensive
effort to study the effect of acidic deposition was not initiated until the
Norwegian SNSF (Sur Nedbtfrs Virkning Pa Skog Og Fisk--"Acid Rain Effects on
Forests and Fish") Project was established in 1972. The phenomenon was first
widely recognized in North America at the First International Symposium on
Acid Precipitation and the Forest Ecosystem in Ohio (USDA 1976), and at the
NATO Conference on Effects of Acid Precipitation on Vegetation and Soils
(Toronto 1978). At the Ohio conference, Tamm and Cowling (1976) speculated
upon the potential effects of acidic deposition, but few existing studies
directly supported their hypotheses of damaging effects.
As the acid rain phenomenon gained increasing attention and its occurrence
was reported over large areas of North America, economic damage to vegetation
was predicted (i.e., Glass et al. 1979, U.S. EPA 1979) and a number of re-
search programs to investigate the effects were initiated in the mid-19701s.
Anthropogenic and natural air contaminants are usually inventoried on a sepa-
rate basis (e.g., chemical speciation) when information is sought as to
sources, dispersion, or induced effects (see Chapters A-2 and A-5). Cate-
gorically, the National Ambient Air Quality Standards (NAAQS) for criteria
pollutants (ozone and other photochemical oxidants, sulfur oxides, nitrogen
oxides, carbon monoxide, lead, and particulate matter) have been established
to protect human health and welfare. Comprehensive documents that describe
vegetation effects of the major phototoxic air pollutants are available (U.S.
EPA 1978; 1982a,b). As distances from pollutant sources increase, chances
for combinations to occur also increase, or, as in the case of large metro-
politan/industrial areas, pollutant combinations are the rule rather than the
exception. However, as distances from sources increase, concentrations of
pollutants generally decrease.
The wet deposition of acidic pollutants may consist of a number of variables
affecting vegetation (i.e., hydrogen, sulfur, and nitrogen doses). The
influence of predominant gaseous pollutants that may be present within the
defined isopleths of acidic precipitation must also be taken into account. If
results of such interaction studies are not available or understood, effects
may be attributed to acidic depositions but instead be due to gaseous
pollutants alone, or as combined with the influence of acidic depositions.
Because of the potential for interactions with biotic and abiotic entities,
factorial research designs and multivariate analyses may be necessary to gain
a more complete understanding of vegetative response to acidic deposition.
3-2
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In the United States, the eastern half of the country is the geographical
area of major concern for impacts of air pollution (both gaseous pollutants
and acid rainfall) on crop and forest productivity. Certain areas of the
western United States, such as the Los Angeles Basin, are also of concern,
however. The combination of a high density of fossil-fuel combustion plants,
a high frequency of air stagnation episodes, and elevated levels of both
photochemical oxidants and rainfall acidity over widespread areas of the
eastern United States have resulted in exposure of large acreages of forests
to increased deposition of atmospheric pollutants (Mclaughlin 1981). An
overlay of isopleths of air stagnation frequency (a measure of the potential
of pollutants to accumulate during periods of limited atmospheric disper-
sion), isopleths of rainfall acidity, and forest zones of the United States
is shown in Figure 3-1.
This overlay highlights this juxtaposition of stress potential and forest
types. While air stagnation episodes are not in themselves a measure of air
pollution stress, they do provide an indication of the potential for pollu-
tants from multiple sources to be concentrated within regional air masses.
The eastern half of the United States, with approximately 80 percent of the
total fossil-fueled electric power plants, thus has both the emissions and
the atmospheric conditions to create regional scale elevation of air pollu-
tants (see Chapter A-2). Comparable conditions also appear to exist in
coastal California, where severe air stagnation has led to very high levels
of photochemical oxidants.
The acidity of rainfall in much of the northeast quadrant of the United
States (Figure 3-1) averages about pH 4.1 to 4.3 annually--about 30 to 40
times as acid as the hypothetical carbonate-equilibrated natural rainfall
with a pH of 5.6 (Likens and Butler 1981). Vegetation in the high-altitude
boreal forests of New England experiences even greater inputs, being exposed
for hundreds of hours during the growing season to clouds with pH values in
the range of 3.5 to 3.7 (Johnson and Siccama 1983). Photochemical oxidants,
principally ozone, which are formed both naturally in reactions involving
ultraviolet radiation and from biogenic and anthropogenic hydrocarbon and
nitrogen oxide precursors, occur at potentially phytotoxic levels over the
entire eastern region (Westburg et al. 1976). Forest productivity losses
from this pollutant have not been quantified except in southern California,
where extreme urban pollution from the Los Angeles Basin and poor air disper-
sion have combined to produce the highest oxidant concentrations in the
United States and widespread forest mortality and decline in the nearby San
Bernadino Mountains (Miller et al. 1977).
3.2 PLANT RESPONSE TO ACIDIC DEPOSITION
3.2.1 Leaf Response to Acidic Deposition (D. S. Shriner)
Any discussion of foliar effects of acidic deposition must be prefaced by a
recognition that our knowledge of the potential effects is drawn from
experimental observations with simulated rain solutions rarely typical of
ambient events. As a result, in the absence of field observation of effects
due to ambient precipitation events, it is important to recognize that these
3-3
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't).70
hi.00
0.70
0.50
5.5C
1-BOREAL FOREST ECOSYSTEM
J-LAKE STATES FOREST ECOSYSTEM
3-EASTERN DECIDUOUS FOREST ECOSYSTEM
4-SOUTH EASTERN PINE FOREST ECOSYSTEM
S-TROPICAL FOREST ECOSYSTEM
6-WESTERN MONTANE FOREST ECOSYSTEM
7-SUBALPINE FOREST ECOSYSTEM
8-PACIFIC COAST FOREST ECOSYSTEM
9-CAUFORNIA WOODLAND
10-SOUTHWESTERN WOODLAND
Figure 3-1. Distribution of frequency isopleths for total number of
forecast days with high meteorological potential for air
pollution over a 5-year period (solid lines). Isopleths
are shown in relation to major forest types of the United
States (adapted from Miller and McBride 1975) and in rela-
tion to mean annual hydrogen ion deposition (kg ha~l yr'1;
dashed lines) in precipitation (adapted from Henderson et
al. 1981).
3-4
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experimental observations are most useful for understanding mechanisms of
effect, and less so for extrapolation to field-scale impacts.
Most of the terrestrial landscape being impacted by acidic deposition is
covered by a minimum of one layer of vegetation. As a result, a large pro-
portion of the incident precipitation ultimately affecting soils and surface
water chemistry has previously contacted vegetation surfaces. The fact that
vegetation surfaces are perhaps the most probable primary receptors of de-
posited pollutants raises two important issues regarding the interactions
between water droplet and receptor surface:
1) effects of incident precipitation chemistry on the receptor surface
structure and function; and
2) effects of the receptor surface on incident precipitation chemistry.
3.2.1.1 Leaf Structure and Functional Modifications--Based on experimental
evidence with simulated rain, a wide range of plant species is believed to be
sensitive to direct injury from some elevated level of wet acidic deposition
(Evans et al. 1981b, Shriner 1981; see also Section 3.4). Other species have
been noted to be tolerant of equally elevated levels (to pH 2.5 for up to 10
hours total exposure) without visible injury (Raines et al. 1980). These
results suggest that generalizations about sensitivity to injury may be dif-
ficult, and some understanding of the mechanisms by which injury may occur is
necessary. The sensitivity of an individual species of vegetation appears to
be influenced by structural features of the vegetation, which 1) influence
the foliage wettability; 2) make the foliage more vulnerable to injury (e.g.,
through differential permeability of the cuticle); or 3) retain rainwater due
to leaf size, shape, or attachment angle. In those instances where one or
more of the above conditions renders a plant potentially sensitive to acidic
deposition, effects may be manifested in alterations of leaf structure or
function.
Injury to foliage by simulated acidic precipitation largely depends on the
effective dose to which sensitive tissues are exposed. The effective dose,
that concentration and amount of hydrogen ion, and time period responsible
for necrosis of an epidermal cell, for example, are influenced by the contact
time of an individual water droplet or film on the foliage surface (Evans et
al. 1981b, Shriner 1981). Contact time, in turn, can be regulated by the
wettability of the leaf, or by leaf morphological features that prevent rapid
runoff of water from the surface. Physical characteristics of the leaf sur-
face (e.g., roughness, pubescence, waxiness) or the chemical compositions of
the cutin and epicuticular waxes determine the wettability of most leaves
(Martin and Juniper 1970).
For injury to occur at the cellular level, the ions responsible must pene-
trate these protective physical and chemical barriers or enter through
stomata (Evans et al. 1981b). Crafts (1961a) has postulated that cuticle
penetration occurs through micropores. Evidence indicates that these micro-
pores are most frequent in areas such as at the bases of trichomes and other
specialized epidermal cells (Schnepf 1965). However, the occurrence of such
micropores is not well documented for all plant cuticles (Martin and Juniper
3-5
-------
1970). Hull (1974) demonstrated that basal portions of trichomes are more
permeable than adjacent areas; cuticles of guard cells and subsidiary cells
are preferred absorption sites (Dybing and Currier 1961, Sargent and
Blackman 1962). In addition, Linskens (1950) and Leonard (1958) found that
the cuticle near veins is apparently a preferential site for absorption of
water-soluble materials.
Perhaps as important as the greater density of micropores associated with
these specialized cells is Rentschler's (1973) evidence that, at least in
certain species, epicuticular wax is less frequently present on certain of
these specialized epidermal cells. Such an absence of wax, in combination
with increased cuticular penetration at those sites, would tend to maximize
the sensitivity of those sites. Evans et al. (1977a,b; 1978) have determined
that approximately 95 percent of the foliar lesions occurring on those plant
species observed by them occurred near the bases of such specialized epi-
dermal cells as trichomes, stomatal guard and subsidiary cells, and along
veins. Stomatal penetration by precipitation, on the other hand, is thought
to be infrequent (Adam 1948; Gustafson 1956, 1957; Sargent and Blackman 1962)
and is considered a relatively insignificant route of entry of leaf surface
solutions (Evans et al. 1981b).
Solution pH has also been shown to influence the rate of cuticular penetra-
tion in studies with isolated cuticles (Orgell and Weintraub 1957, McFarlane
and Berry 1974). The rate of penetration of acidic substances increased with
a decrease in pH, while the rate of penetration of basic substances increased
with an increase in pH (Evans et al. 1981b).
Preliminary work by Shriner (1974) suggested that, in addition to the physi-
cal abrasion of superficial wax structure by raindrops, leaves exposed to
rainfall of pH 3.2 appeared to weather more rapidly than did leaves of pH 5.6
control treatment plants. However, it was impossible to determine from those
experiments whether chemical processes at the wax surface were responsible
for the differences or whether the acidic rain induced physiological changes
that retarded regeneration of the waxes and recovery from mechanical damage.
The latter explanation may be the most tenable because the waxes would be
expected to resist chemical reaction with dilute strong acids (Evans et al.
1981b), and because numerous reports of physiological imbalance resulting
from acidic precipitation exposure exist (Shriner 1981). Hoffman et al.
(1980) proposed a mechanism by which precipitation acidity can act as a
chemical factor in weathering epicuticular waxes. They pointed out that the
wax composition, as polymeric structures of condensed long-chain hydroxy
carboxylic acids, may result in an "imperfect" wax matrix in which the
uncondensed sites containing hydroxy functional groups are more readily
weathered. Strong acid inputs to such a system would oxidize and release a
wide range of carbon chain acids from the basic waxy matrix, conceivably
yielding the type of change in weathering rate Shriner observed.
Rentschler (1973) and, more recently, Fowler et al. (1980) have shown
relationships between the superficial wax layer of plants and plant response
to gaseous air pollution. The work of Fowler et al. compared the rate of
epicuticular wax degradation of Scots pine needles from "polluted" and
unpolluted sites in the field. These "polluted" sites included exposure to
3-6
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both dry deposition of gaseous pollutants and wet deposition as acid rain,
making it impossible to distinguish between relative effects of the two forms
of deposition. Needles at the polluted site showed greater epicuticular wax
structure degradation during the first eight months of needle expansion.
Determing the quantity of wax per unit leaf area showed very small
differences between polluted and clean air sites. Fowler et al. concluded
that observed differences (by scanning electron microscopy) were "due more to
changes in form than gross loss of wax." Since the fine structure of the wax
layer is controlled largely by the chemical composition of the wax (Jeffree
et al. 1975), the observed changes may also reflect stress-induced changes in
wax synthesis. Fowler et al. estimated that increased water loss due to
accelerated breakdown of cuticular resistance would only influence trees if
water were a limiting factor. They concluded that "the extra water loss may
reduce the period (or degree) of stomatal opening" and that the magnitude of
the effect on dry matter productivity would not be greater than 5 percent at
their polluted site. Because study sites used by Fowler et al. were exposed
to gaseous sulfur dioxide as well as to acidic precipitation, their work does
not allow identification of a single causative factor.
Histological studies of foliar injury caused by acidic precipitation have
revealed evidence of modification of leaf structure associated with plant
exposure to acidic precipitation (Evans and Curry 1979). Quercus palustris,
Tradescantia sp., and Pppulus sp. exposed to simulated acidic precipitation
experienced abnormal cellproliferation and cell enlargement. In Quercus
(oak) and Populus (poplar) leaves, prolonged exposure to treatment at pH 2.5
produced hypertrophic and hyperplastic responses in mesophyll cells. Lesions
developed, followed by enlargement and proliferation of adjacent cells,
resulting in formation of a gall on adaxial leaf surfaces. In poplar test
plants, this response involved both palisade and spongy mesophyll parenchyma
cells, while in oak test plants, only spongy mesophyll cells were affected
(Evans and Curry 1979). Because other similar histological studies have not
been reported, it is impossible to evaluate how frequent or widespread such
structural modification may be. Because species that have been reported to
show hyperplastic and hypertrophic response of leaf tissues were consistently
injured less than species that did not show these responses, gall formation
may be linked to characteristics common to species tolerant of acidic pre-
cipitation exposure.
Several studies have reported modification of various physiological functions
of the leaf as a result of exposure to simulated acidic precipitation.
Sheridan and Rosenstreter (1973), Ferenbaugh (1976), Hindawi et al. (1980),
and Jaakhola et al. (1980) reported reduced chlorophyll content as a result
of tissue exposure to acidic solutions. Ferenbaugh, however, observed that
significant reduction in chlorophyll content did not occur at pH 2.0, and
that chlorophyll content slightly increased at pH 3.0. Irving (1979) also
reported higher chlorophyll content of leaves exposed to simulated precipi-
tation at pH 3.1. Hindawi et al. observed a steady reduction in chlorophyll
content in the range between pH 3.0 to 2.0, and found no change in the ratio
of chlorophyll a:b.
Ferenbaugh (1976) determined photosynthesis and respiration rates of test
bean plants exposed to simulated acidic precipitation. Respiration and
3-7
-------
photosynthesis were significantly increased at pH 2.0. Ferenbaugh concluded
that because growth of the plants was significantly reduced, photophosphory-
lation was uncoupled by the treatments. Irving (1979) reported increased
photosynthetic rates in some soybean treatments, attributing them to in-
creased nutrition from sulfur and nitrogen components of the rain simulant,
which overcame any negative effect of the pH 3.1 treatment. Jacobson et al.
(1980) reported a shift in photosynthate allocation from vegetative to
reproductive organs as a result of acidic rain treatments of pH 2.8 and 3.4,
also suggesting that the primary effect was not on the photosynthetic process
itself.
3.2.1.2 Foliar Leaching - Throughfall Chemistry—Rain, fog, dew, and other
forms of wet deposition play important roles as sources of nutrients for
vegetation and as mechanisms of removal from vegetation of inorganic nutri-
ents and a variety of organic substances: carbohydrates, amino acids, and
growth regulators (Kozel and Tukey 1968, Lee and Tukey 1972, Hemphill and
Tukey 1973, Tukey 1975). Tukey (1970, 1975, 1980) and Tukey and Morgan
(1963) have extensively reviewed the leaching of substances from plants as
the result of water films on plant surfaces.
During periods between precipitation events, the vegetation canopy serves as
a sink, or collection surface, upon which dry particulate matter, aerosols,
and gaseous pollutants accumulate by gravitational sedimentation, impaction,
and adsorption. Throughfall can be defined as that portion of the gross, or
incident, precipitation that reaches the forest floor through openings in the
forest canopy and by dripping off leaves, branches, and stems (Patterson
1975). Throughfall generally amounts to between 70 and 90 percent of gross
rainfall, with the balance divided between stemflow and interception loss to
the canopy.
Chemical enrichment of throughfall has been well documented for a broad vari-
ety of forest species (Tamm 1951, Madgwick and Ovington 1959, Nihlgard
1970, Patterson 1975, Lindberg and Harriss 1981). TMs enrichment has three
potential sources: 1) reactions on the leaf surface in which catij^s on
exchange sites of the cuticle are exchanged with hydrogen from rainfali; 2)
movement of cations directly from the translocation stream within the ieaf
into the surface film of rainwater, dew, or fog by diffusion and mass flow
through areas devoid of cuticle (Tukey 1980); and/or 3) washoff of atmos-
pheric particulate matter that has been deposited on the plant surfaces
(Patterson 1975, Parker et al. 1980, Lindberg and Harriss 1981).
The exchange of hydrogen ions in precipitation for cations on the cuticle
exchange matrix can result in significant scavenging of hydrogen ions by a
plant canopy. Eaton et al. (1973), for example, found the forest canopy to
retain 90 percent of the incident hydrogen ions from pH 4.0 rain (growing
season average), resulting in less-acidic ( ~ pH 5.0) solutions reaching the
forest floor. The removal of H+ by exchange processes in the forest canopy
does not eliminate the effects of H+ deposition on the forest ecosystem,
however. Cations leached from the foliage may eventually be leached from the
ecosystem if the anion associated with H+ inputs (S042~ or N03~)
is mobile (see Figure 2-1, Chapter E-2). Plant response to this may be 1)
accelerated uptake to compensate for foliar cation losses, or 2) reduced
3-8
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foliar cation concentrations, if H+ inputs and foliar exchange are of
significant magnitude and duration. In either event, the introduction of
H+ with a mobile anion will cause the net loss of cations from the
ecosystem, whether the H+ cation exchange occurs in the forest canopy or in
the soil. Further aspects of cation leaching are discussed in Chapter E-2
(soils).
An example of the second case has recently been hypothesized by Rehfuess et
al. (1982) for Norway spruce in high elevation forests of eastern Bavaria.
Trees experiencing symptoms of decline and dieback (see Sections 3.4.1.5 and
3.4.1.6) were paired with non-symptomatic trees in the same stands and site
conditions. Large differences were noted in foliar content, particularly of
older leaves, of Ca and Mg, with declining trees consistently showing lower
levels of Ca and Mg content than healthy trees. The Mg contents were
characterized by the authors as in "extreme deficiency," with calcium in
"poor supply." The authors further speculated that since these nutrient
deficiences occurred on soils varying considerably in content of both
elements, that soil depletion was probably not the dominant contributing
factor, but rather that the deficiency is mainly a consequence of enhanced
leaching of Ca and Mg from the foliage as a result of acidic deposition of
strong acids. The authors further speculated that Ca and Mg uptake from soil
pools may be inadequate to replace this foliar leaching. Such nutritional
disorders have been reported to subsequently make foliage more susceptible to
additional leaching (Tukey 1970).
Separating relative contributions of internal (leached) and external (wash-
off) fractions of throughfall enrichment is difficult and has been attempted
infrequently. Parker et al. (1980) have reviewed those attempts to estimate
the importance of dry sulfur deposition to throughfall enrichment by sulfate-
sulfur (Table 3-1). For those studies that have attempted such an analysis,
the estimated percentage contribution of dry deposition to throughfall en-
richment ranged from 13 to 100 percent, or from 0.3 to 14.4 kg ha'1 yr'1.
Parker et al. concluded that for temperate hardwood forests in industrialized
regions, 40 to 60 percent of annual net throughfall (throughfall enrichment)
of sulfate is due to washoff of dry deposition, with 30 to 50 percent being
typical for conifers of the same regions. For hardwoods and conifers in
regions typified by low background levels of dry sulfur deposition, washoff
may range from 0 to 20 percent of throughfall enrichment. Similar data have
been developed for several trace elements (Lindberg and Harriss 1981).
Through the application of simulated rainfall in controlled experiments,
precipitation acidity has been studied as a variable influencing the leaching
rate of various cations and organic carbon from foliage (Wood and Bormann
1974, Fairfax and Lepp 1975, Abrahamsen et al. 1977). Foliar losses of
potassium, magnesium, and calcium from bean and maple seedlings were found to
increase as the acidity of simulated rain increased. Tissue injury occurred
below pH 3.0, but significant increases in leaching rates occurred as high as
pH 4.0 (Wood and Bormann 1974). Phaseolus vulgaris L. foliage exposed by
Evans et al. (1981a) to citrate-phosphate buffer solutions with a range in
acidity from pH 5.7 to pH 2.7 also demonstrated that greater acidity of these
solutions preferentially leached greater amounts of calcium, nitrate, and
sulfate, while less acidic solutions leached greater amounts of potassium and
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TABLE 3-1. REPORTED VALUES FOR SULFATE-SULFUR DEPOSITION RATES FOR THROUGHFALL AND INCIDENT
PRECIPITATION IN WORLD FORESTS
co
i
Forest system
Subalpine balsam fir,
New Hampshire
Hemlock,
British Columbia
Conifers,
southern Norway
Conifers,
southern Norway
Conifers,
southern Norway
Beech,
central Germany
Spruce,
central Germany
Hemlock-spruce,
Reference
Cronan 1978
Feller 1977
Haughbotn 1973
Haughbotn 1973
Haughbotn 1973
Heinrichs and Mayer
1977
Heinrichs and Mayer
1977
Johnson 1975
S deposition
Incident
24.4
11. Oa
32. 3b
17.7
10.0
24. ld
24.1
0
kg ha'1 yr"1
Throughfall
46.4
40.0
111.2
69.1
21.1
47.6
80.0
16.4
Precipitation
amount
(cm)
203d
245C
77
77
77
106
106
270
southeastern
Alaska
Tropical rain forest,
Costa Rica
Johnson 1975
12.5
23.3
390
-------
TABLE 3-1. CONTINUED
Forest system
Douglas fir,
Washington
Subalpine silver fir,
Washington
Hardwoods,
Amazonian Venezuela
Hardwoods,
Amazonian Venezuela
Hard beech,
New Zealand
Beech,
Southern Sweden
Spruce,
Southern Sweden
Oak,
Southern France
Douglas fir,
Oregon
Loblolly pine,
Reference
Johnson 1975
Johnson 1975
Jordan et al . 1980
Jordan et. al . 1980
Miller 1963
Nihlgard 1970
Nihlgard 1970
Rapp 1973
Soil ins et al . 1979
Wells et al. 1975
S deposition
Incident
4.0
16.8f
44.5
46.6
8.4
7.9d
7.9d
16.4
4.7
7.9a
kg ha"1 yr"1
Throughfall
5.2
5.3
16.7
19.6
10.4
18.5
54.2
22.6
2.4
9.9
Precipitation
amount
(cm)
165
300
391
412
135
95
95
NA
237
NA
North Carolina
-------
TABLE 3-1. CONTINUED
u>
I—"
ro
S deposition kg ha"* yr"*-
Forest system
Chestnut oak,
Tennessee
Mixed oak, Tennessee
Mixed oak, Tennessee
Reference
Lindberg et al . 1979
Kelly 1979
Kelly 1979
Incident
13.2b»e
8.7a
11.3a>b
Throughfall
32.0
15.0
14.0
Precipitation
amount
(cm)
143
154
75
Scaled up from a subannual estimate.
In vicinity of factory or power plant.
cMean of extreme estimates.
Includes stem flow.
eSeveral years data.
fLittle throughfall.
-------
chloride. Abrahamsen and Dollard (1979) observed that Norway spruce (Picea
abies (L.) Karst) lost greater quantities of nutrients under their most
acidic treatments, but no related change in foliar cation content occurred,
in contrast to the observations of Rehfeuss et al. (1982) discussed above.
Wood and Bormann (1977) noted results similar to those of Abrahamsen and
Dollard (1979) for eastern white pine (Pinus strobus L.).
3.2.2 Effects of Acidic Deposition on Lichens and Mosses (L. L. Sigal)
The objective of this section is to review the literature on the effects of
acidic deposition on lichens and mosses and also te review the literature
that describes the effects of realistic, low levels of gaseous sulfur dioxide
($03) on lower plants. Several researchers (Skye 1968, Turk and Wirth
1975) have concluded that SO? toxicity and pH effects are not independent
factors (Grennfelt et al. 1980).
Lichens and mosses are considered by some researchers (Nieboer et al. 1976)
to be among the most pollution-sensitive plants, and by others to be more
sensitive and better indicators of chronic pollution than vascular plants
(Hawksworth 1971, Nash 1976, Guderian 1977, Winner et al. 1978). In addition
to their roles in the ecosystem, they are also valuable as biomonitors of air
quality. However, it must be noted that lichens and mosses integrate the
effects of all ambient pollutants, and in most cases, their use as bioindi-
cators is only an index of general air pollution.
Lichens are sensitive to air pollutants such as sulfur dioxide, (Ferry et al .
1973), ozone and peroxyacetyl nitrate (PAN) (Nash and Sigal 1979, Sigal and
Taylor 1979), fluorine (Nash 1971, Roberts and Thompson 1980), and metals
(Rao et al. 1977; lead, Lawrey and Hale 1981; nickel, Nieboer et al. 1972;
mercury, Steinnes and Krog 1977; zinc, Nash 1975; and chromium, Schutte
1977). Scientists in many countries have demonstrated that it is possible to
correlate the distribution of lichens around air pollution sources with mean
levels of air pollutants. Laboratory and transplant studies have corro-
borated the data from field investigations. However, the importance of peak
concentrations of pollutants relative to long-term average levels has not
been established. Excellent summaries on the theory and application of
lichens in pollution studies have been published by Ferry et al. (1973),
Gilbert (1974), Hawksworth and Rose (1976), Le Blanc and Rao (1975),
Richardson and Nieboer (1981), Skye (1968, 1979), and Saunders (1970). In
addition, the air pollution literature is regularly indexed in the British
journal "The Lichenologist" (1974-81).
Moss species are also sensitive to air pollution (Gilbert 1968, 1970; Nash
1970; Nash and Nash 1974; Stringer and Stringer 1974; Turk and Wirth 1975;
Winner and Bewley 1978a,b). However, less attention has been given to mosses
in air pollution research. Laboratory studies with mosses have shown that 1)
photosynthesis decreases in relation to a decrease in pH of sulfuric acid
solutions (Sheridan and Rosenstreter 1973), 2) sulfite and bisulfite solu-
tions reduce photosynthesis (Inglis and Hill 1974, Ferguson and Lee 1979),
and 3) growth of four species of Sphagnum moss was reduced when they were
fumigated for several months with mean SOg concentration of 130 vig m~3
(Ferguson et al. 1978). It has been suggested that sul fate at "feasible"
3-13
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atmospheric concentrations has no effects upon photosynthesis in mosses;
however, the fall in pH that accompanies the oxidation of atmospheric S02
to $04 is capable of reducing photosynthesis (Ferguson and Lee 1979). The
phytotoxic effect of S02 for both mosses and lichens is known to be greater
at low pH (Gilbert 1968, Puckett et al. 1973, Inglis and Hill 1974, Hallgren
and Huss 1975).
The generally accepted mechanisms of injury are disruption of cell and
chloroplast membranes (Wellburn et al. 1972, Puckett et al. 1974, Malhotra
1976, Ferguson and Lee 1979), and destruction of chlorophyll (Rao and Le
Blanc 1966, Nash 1973, Puckett et al. 1973). Susceptibility to SO? injury
is greatest when lichens are in a moistened or saturated condition TRao and
Le Blanc 1966; Nash 1973, 1976; Turk et al . 1974). In an air-dried state,
lichens have been shown to be relatively insensitive to S02 (Showman 1972,
Nash 1973, Turk et al. 1974, Marsh and Nash 1979).
The sensitivity of lichens to air pollutants is due to a number of factors:
(1) they rapidly absorb moisture in different forms (e.g., rain, fog, dew)
and most toxic substances dissolved in the water (Richardson and Nieboer
1981); (2) they are long-lived, and accumulated sulfur metabolites, metals,
etc. are not eliminated seasonally (Nash 1976); (3) they lack a vascular
system with which to eliminate pollutants through translocation (Nieboer et
al. 1976); (4) they lack structures such as epidermis and stomata to exclude
pollutants (Sundstrom and Hallgren 1973); (5) they probably have less buf-
fering capacity than vascular plants (Nieboer et al. 1976); and (6) the
relationship of the alga and the fungus is delicately balanced; air pollution
probably disrupts that balance, resulting in disassociation and destruction
of the plant (Neiboer et al. 1976).
The ecology of lichens can be drastically changed by air pollutants. As a
result, ecosystems are affected because lichens are integral parts of many
relationships and processes. As pioneer species in disturbed areas (Treub
1888), lichens initiate soil formation (Ascaso and Gal van 1976) and stabilize
soil (Rychert and Skujins 1974, Drouet 1937). They fix an estimated 10 to 50
percent of the newly-fixed nitrogen in old growth forests in the United
States (Denison 1973, Becker 1980, Rhoades 1981). They act as sinks for air
pollutants and contribute to the cleansing of the atmosphere (A. C. Hill
1971).
Many invertebrates (mites, caterpillars, earwigs, snails, slugs, etc.) as
well as vertebrates (caribou, reindeer, squirrels, woodrats, voles) feed
partly or wholly on lichens (Llano 1948, Richardson 1975, Gerson and Seaward
1977, Richardson and Young 1977). Other animals have adaptive camouflage
that resembles lichen-covered trees or rocks (Richardson and Young 1977).
The interrelations among birds and lichens and insects are multifaceted.
Birds use lichens for nest-building, camouflage, and feeding behavior
(Kettlewell 1973, Ewald 1982), while many insects have co-evolved with
lichens to escape predation from birds (Cott 1940).
Reports of injury to lichens at low levels of S02 are found in several
recent studies. Showman (1975) found that Parmelia caperata and P_. rudecta
were absent in regions around a coal-fired power pi ant when the annual
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average exceeded 50 yg nr3. Will-Wolf (1980) found that Parmelia
caperata and P. bolllana showed morphological alterations in areaswhere
maximum $03 levels were 389 yg rrr3, and annual averages were 5 to 9
yg m~3. Eversman (1978) found decreased respiration rates in Usnea hirta
after field fumigations with S02 at about 47 yg rrr3 for 96 days, and
plasmolysis of algal cells in both JJ. hirta and Parmelia chlorochroa after 31
days of S02 at the same concentration. Le Blanc and Rao (1975) concluded
that long-range average concentrations for S02 between 16 to 79 yg m
(0.006 to 0.03 ppm) cause chronic injury to epiphytes.
In the Ohio River Valley, maximum annual averages of S02 ranged from about
50 to 80 yg m"3 in 1977 and 1978. Maximum 1-hr averages ranged from 300
to 500 yg nr3 (Mueller et al. 1980). At the same sites (Rockport and
Duncan Falls), mean rainfall pH's for August 1978 to September 1981 were 4.12
and 4.36, with ranges of 3.60 to 5.48 and 3.59 to 5.73, respectively [digital
(9 track tape) or hard copy (printout) versions of these data are available
upon request directly from Peter K. Mueller at EPRI]. Recent experimental
evidence shows that photosynthesis was reduced by 40 percent in the lichen
Cladina stellaris by field fumigations with fluctuating S02 concentrations
of less than 655 yg m~3 (0.25 ppm; Moser et al. 1980). Laboratory ex-
posures of the same lichen species wetted by artifical precipitation having a
pH = 4.0 and a sulfate concentration = 10.00 mg £-1 reduced photosyn-
thesis by 27 percent (Lechowicz 1982). From these and succeeding data, it
appears that at least some of the mechanisms of injury for S02 and acid
precipitation are similar and that existing, long-term low levels of the
pollutants are influencing lichen distribution on a regional scale.
The effect of direct acidic deposition on lichens is a new area of research
and therefore has produced few published results other than those of
Lechowicz (1982). Evidence from previous laboratory studies of the effects
of pH on lichens is indirect and based generally on aqueous solutions of
sulfur compounds. Puckett et al. (1973, 1974) found that low pH enhanced
aqueous sulfur dioxide toxicity in buffered solutions even when the exposure
times were brief. D. J. Hill (1971) found that sulfite in buffered solutions
was toxic at pH 4.0 and below but not toxic at pH 5.0 and above. Turk and
Wirth (1975) found that damage to lichens exposed to sulfur dioxide and
subsequently submersed in buffer solutions from pH 8.0 to pH 2.0 increased
with increasing acidity. Baddeley et al. (1971) studied the effect of pH in
buffered solutions on the respiration of several lichen species found in
eastern North America. Exposure times were short, about 15 minutes, but
respiration was clearly pH-dependent, and there were definite pH optima for
each species, mostly acidic (pH 4.0). Repeated exposures might show
different patterns of respiration.
Little is known about the effects of acidic deposition on nitrogen fixation
by lichens. Denison et al. (1977) reported a trend toward decreased nitrogen
fixation in the lichens Lobaria pulmonaria and J.. oregana as a function of
decreasing pH of the water in which the lichens were soaked. These results
must be considered preliminary, and additional work in this area is needed
because lichens can be important contributors of fixed nitrogen in forest
ecosystems (Forman 1975; Pike 1978; Becker 1977, 1980; Rhoades 1981), in
3-15
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tundra and grasslands (Alexander 1974), and in deserts (Shields et al. 1957,
Rychert and Skujins 1974).
Evidence from the few existing field studies of acid precipitation effects on
lichens (Robitaille et al. 1977, PIummer 1980) is inconclusive because
separating pH effects from potential ambient S02 (or other gaseous
pollutant) toxicity is impossible under natural conditions. Few of the
studies that suggest a pH response in lichens (Brodo 1974) actually include
the measurement of pH of the aqueous solutions in which the lichens are
bathed. Several field studies suggest that acidification of lichen
substrates may prevent establishment and development of lichen propagules
(Barkman 1958, Skye 1968, Gilbert 1970, Grodzinska 1979). Other studies
(Abrahamsen et al. 1979, Dahl et al. 1979) show that lichrens alter the
chemistry of "rainwater" flowing over granite surfaces partly covered with
lichens. Pyatt (1970) notes that lichens are capable, to some extent, of
exerting a modifying influence upon the environment. According to Gilbert,
the pH and buffer capacity of the lichen thai 1 us and substrate are important
for the survival and regeneration of lichens in polluted areas because pH and
buffer capacity control the distribution and proportions of toxic compounds
in solution and the rates of breakdown pf these compounds. Under conditions
of acid precipitation and reduced buffer capacity, heavy metal absorption by
lichens is increased (Rao et al. 1977).
3.2.3 Summary (D. S. Shriner and L. L. Sigal)
Leaf structure may play two roles in the sensitivity of foliar tissues to
acidic precipitation: 1) leaf morphology may selectively enhance
(broad-leaved species) or minimize (needle or laminar-leaved species) the
surface retention of incident precipitation; and 2) specific cells of the
epidermal surface, by virtue of a more permeable cuticle or the absence of
waxes, may be initial sites of foliar injury. Once such a lesion occurs,
further development of local lesions appears to be enhanced by water
collected in the depression formed by the necrotic tissue.
Information on the effects of acidic deposition on the accelerated weathering
of epicuticular wax of plants is very preliminary and at present must be
considered no more than a "testable hypothesis." Should further research
support the hypothesis, virtually all of the important functions of the wax
layer could be subject to alteration due to acidic deposition.
Chlorophyll degradation may occur following prolonged exposure to acidic pre-
cipitation. Conclusive linkage to decreased photosynthetic rates is current-
ly missing, but premature senescence resulting from chlorophyll degradation
may reduce overall photosynthetic capacity of plants affected in this manner.
Further study is needed before photosynthetic rate, chlorophyll content, and
premature senescence can be causally linked to acidic deposition exposure.
Because simulated acid precipitation experiments have been conducted at ex-
treme ranges, more attention must be paid to pH values commonly observed in
nature.
Acid deposition is frequently partially neutralized by cation exchange and
other reactions on leaf surfaces. These reactions reduce the direct inputs
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of H+ to soils, but they do not prevent cation losses from the ecosystem.
If the am'on associated with acidic deposition is mobile, cation losses will
occur whether H+ is exchanged in the canopy or soils.
Information on which to assess the effects of acidic deposition on lichens is
inadequate. Studies should investigate the direct effects of H+ concen-
tration and the other acidic deposition components (S, N) on lichens. A
comparison of process-level physiological mechanisms of response to acidic
deposition is necessary, followed by an analysis of the resulting effects, if
any, on the overall growth, yield, or ecosystem function of lichens. In
addition, the relevance of laboratory studies to field observations must be
established. Given the sensitivity of lichens to related stress agents, they
are probably sensitive to acidic deposition. In certain ecosystems (e.g.,
boreal forests) lichens are a major system component, and potential effects
should be regarded as a serious concern for long-term ecosystem stability.
3.3 INTERACTIVE EFFECTS OF ACIDIC DEPOSITION WITH OTHER ENVIRONMENTAL
FACTORS ON PLANTS
Several important, but often overlooked, indirect effects of acidic deposi-
tion are potential interactions with other pollutants, alterations of host-
insect interactions, host-parasite interactions, and symbiotic associations
(Figure 3-2). These relationships could involve a direct influence of acidic
deposition on a host plant; a direct influence of acidic deposition on an
insect, microbial pathogen, or microbial symbiont; or a direct influence of
acidic deposition on the interactive process of plant and agent, i.e.,
infestation, disease, or symbiosis (Figure 3-2).
3.3.1 Interactions with Other Pollutants (J. M. Skelly and B. I. Chevone)
The available literature concerning interactive effects of acidic precipita-
tion and gaseous air pollutants on terrestrial vegetatation consists of only
three separate studies as of late 1981. Shriner (1978b) examined the inter-
action of acidic precipitation and sulfur dioxide or ozone on red kidney bean
(Phaseolus vulgaris) under greenhouse conditions. Treatments with simulated
rain at pH 4.0 and multiple 03 exposures resulted in a significant reduc-
tion in foliage dry weight. Simulated precipitation and sulfur dioxide in
combination did not affect photosynthesis or biomass production. Troiano et
al. (1981) exposed two cultivars of soybean to ambient photochemical oxidant
and simulated rain at pH 4.0, 3.4, and 2.8 in a field chamber system. The
interactive effects of oxidant and acidic precipitation were inconclusive,
with seed germination greater in plants grown in the absence of oxidant at
each acidity level. Irving and Miller (1981) also examined the response of
field-grown soybeans to simulated acidic rain at pH 5.3 and 3.1 in combina-
tion with sulfur dioxide and ambient ozone concentrations. No interactive
effects on soybean yield occurred from acid treatments with sulfur dioxide.
Sulfur dioxide alone, however, resulted in substantial yield reductions.
With information from only three studies, current assessment of the potential
detrimental interactive effects of gaseous air pollutants and acidic rain on
terrestrial plants can be considered only preliminary. No studies have been
3-17
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ACID DEPOSITION
INSECT
FOLIAGE
FEEDER
MICROBE
FOLIAR
PATHOGEN
INSECT
BARK BEETLE
MICROBE
STEM PATHOGEN
|
MICROBE
ROOT PATHOGEN
SYMBIONT
INSECT
SOIL ARTHROPOD
Figure 3-2. Acid deposition may influence insects, pathogens, and
symbionts associated with forest trees by direct influence
(solid arrows) or indirect influence via host alteration
(dashed arrows). Direct influence on soil inhabiting
insects and microbes is judged less likely than direct
influence on aboveground organisms. Alterations of soil pH
or chemistry by acid deposition may indirectly impact soil
organisms.
3-18
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conducted with non-agricultural vegetation which, because of potential soil
impacts, is considered more sensitive to the indirect effects of acidic
precipitation.
Research efforts at present have addressed the indirect interaction of acidic
precipitation and gaseous pollutant stress to plants. Plants have been ex-
posed to pollutants individually so that any interactive effects are mediated
through the plant response, whether directly or indirectly to each pollutant.
With this exposure regime, each pollutant may predispose the plant to addi-
tional injury and elicit a more sensitive response to the second pollutant.
It is advantageous, under these conditions, to use experimental systems that
are most sensitive to both acidic inputs and gaseous pollutant stress. Due
to crop management practices, agronomic systems are probably least sensitive
to increased acidic input and alterations in soil physiochemical properties.
Additional research in which both acidic precipitation and gaseous pollutants
can exert their individual effects on the various components of an ecosystem
is required.
Effects of acidic deposition on soil chemistry and nutrient recycling are
unlikely to occur rapidly (Chapter E-2, Section 2.3) and unlikely to occur in
agricultural systems where soils are regularly amended (Section 2.3.5).
After more than a decade of research in Scandinavia, the observed changes in
forest soil chemical properties that can be attributed to acidic precipita-
tion still remain undetermined (Overrein et al. 1980). It is, therefore,
unlikely that interactive effects of acidic deposition and gaseous pollutants
on plants, which may be expressed through changes in soil properties, will
become evident within a single growing season. Because only annual plants
have been used in interactive studies, the effect of acidic rain in combi-
nation with other air pollutants stressing perennial plant species on a
yearly basis for several years is unknown. Also, research efforts have not
addressed the temporal relationship between precipitation events and the
occurrence of other gaseous air pollutants in the ambient atmosphere.
No information exists on the interaction of a gaseous air pollutant with a
wet leaf surface. Such direct interactions can occur only with the same
frequency as precipitation events (including fog, dew, and condensation), but
liquid-phase reactions, especially with S02, can alter the chemical form of
the pollutant species. Sulfur dioxide in water can exist as the hydrated
sulfur dioxide molecule, the bisulfite ion, or the sulfite ion, depending
upon the pH of the solution (Gravenhorst et al. 1978). At pH greater than
3.5, hydrated sulfur dioxide dissociates almost completely into hydrogen ions
and bisulfate ions. Increased solubility of sulfur dioxide can occur if the
bisulfite ion is oxidized irreversibly to the sulfate ion. This oxidation
process can be catalyzed by metal cations, specifically iron (Fuzzi 1978) and
manganese (Penkett et al. 1979). Particulate deposits on the leaf surface,
containing either iron or manganese, may act as sources of these catalysts.
Depending upon the rate of this oxidation and the mechanism(s) involved,
increased dissolution of gaseous sulfur dioxide will occur in leaf surface
water, generating additional hydrogen ions. Whether such reactions do occur
at the leaf surface, the extent to which they occur, and their importance in
pollutant stress to plants are unknown.
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3.3.2 Interactions with Phytophagous Insects (W. H. Smith)
The damaging influence of high population densities of certain insects can be
very visible and cause widespread forest destruction; however, substantial
evidence supports the hypothesis that forest insects, even those that cause
massive destruction in the short run, may play essential and beneficial roles
in forest ecosystems in a long-term context. These roles may involve regu-
lating tree species competition, species composition and succession, primary
production, and nutrient cycling (Huffaker 1974, Mattson and Addy 1975). As
a result, assessing interrelationships between acidic deposition and
phytophagous insects is important.
Air pollutants may directly affect insects by influencing growth rates, muta-
tion rates, dispersal, fecundity, mate finding, host finding, and mortality.
Indirect effects may occur through changes in host age structure, distribu-
tion, vigor, and acceptance. Few researchers have investigated the effects
of acidic deposition on insects. Some studies relative to acidity effects on
aquatic insects are available (e.g., Borstrum and Hendrey 1976). Terrestrial
arthropods, on the other hand, have been the subject of very few studies.
Hagvar et al. (1976) have concluded that acidic precipitation from western
and central Europe increases the susceptibility of Scots pine forests to the
pine bud moth (Exoteleia dodecella).
Various studies have presented data indicating that species composition or
population densities of insect groups are altered in areas of high air
pollution stress, for example, roadside (Przybylski 1979) or industrial
(Sierpinski 1967, Novakova 1969, Lebrun 1976) environments. Further specific
information is available on the general influence of polluted atmospheres on
population characteristics of forest insects (Templin 1962; Schnaider and
Sierpinski 1967; Sierpinski 1970, 1971, 1972a,b; Boullard 1973; Wiackowski
and Dochinger 1973; Hay 1975; Charles and Villemant 1977; Sierpinski and
Chlodny 1977; Dahlsten and Rowney 1980). Johnson (1950, 1969) has reviewed
much of the literature dealing with air pollutants and insect pests of
conifers. One of the most comprehensive literature reviews available
concerning forest insects and air contaminants has been presented by
Villemant (1979). Recently, Alstad et al. (1982) provided an excellent
overview of the effects of air pollutants on insect populations.
3.3.3 Interactions with Pathogens (W. H. Smith)
Abnormal physiology, or disease, in woody plants follows infection and
subsequent development of an extremely large number and diverse group of
microorganisms within or on the surface of tree parts. All stages of tree
life cycles and all tree tissues and organs are subject, under appropriate
environmental conditions, to impact by a heterogeneous group of microbial
pathogens including viroids, viruses, mycoplasmas, bacteria, fungi, and
nematodes. As with insect interactions, microbes and the diseases they cause
play important roles in succession, species competition, density, composi-
tion, and productivity. In the short term, the effects of microbial patho-
gens may conflict with forest management objectives and assume a considerable
economic or managerial as well as ecologic significance (Smith 1970).
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The interaction between air pollutants and microorganisms in general is
highly variable and complex. Babich and Stotzky (1974) have provided a
comprehensive overview of the relationships between air contaminants and
microorganisms. A specific air pollutant, at a given dose, may be
stimulatory, neutral, or inimical to the growth and development of a
particular virus, bacterium, or fungus. In fungi, fruiting body formation,
spore production, and spore germination may be stimulated or inhibited.
Microorganisms that normally develop in plant surface habitats may be
especially subject to air pollutant influence. These microbes have received
considerable research attention and have been the subject of review (Saunders
1971, 1973, 1975; Smith 1976). Numerous comprehensive reviews have sum-
marized the interactions between air contaminants and plant diseases
(Laurence 1981). Heagle (1973) summarized nearly 100 references and found
that sulfur dioxide, ozone, or fluoride had been reported to increase the
incidence of 21 diseases and decrease the occurrence of nine diseases in a
variety of nonwoody and woody hosts. Treshow (1975) has provided a detailed
review concerning the influence of sulfur dioxide, ozone, fluoride, and
particulates on a variety of plant pathogens and the diseases they cause.
Treshow lamented the fact that most of the data available deal with in vitro
or laboratory accounts of microbe-air pollutant interactions, while only a
few investigations have examined the influence of air pollutants on disease
development under field conditions.
A review provided by Manning (1975) pointed out that most research attention
has been directed to fungal pathogen-air pollutant interactions. Greater-
research perspective is needed concerning air pollution influence on viruses,
bacteria, nematodes, and the diseases they cause. Macroscopic agents of
disease, most importantly true- and dwarf-mistletoes, must also be examined
relative to air pollution impact, especially in the western part of North
America, where the latter are extremely important agents of coniferous
disease.
Forest trees, because of their large size, extended lifetimes, and widespread
geographic distribution are subject to multiple microbially-induced diseases
frequently acting concurrently or sequentially. The reviews of Heagle
(1973), Treshow (1975), and Manning (1975) considered a variety of pollutant-
woody plant pathogen interactions but were not specifically concerned with
forest tree disease. In their review of the impact of air pollutants on
fungal pathogens of forest trees of Poland, Grzywacz and Wazny (1973) cited
literature indicating that air pollution stimulated the activities of at
least 12 fungal tree pathogens while restricting the activities of at least
10 others.
Our understanding of the influence of acidic deposition on pathogens and the
diseases they cause is meager. Shriner (1974, 1975, 1977) has provided us
with some valuable perspectives in this important but understudied area.
Falling precipitation and the precipitation wetting of vegetative surfaces
(see Section 3.2.1), play an enormously important role in the life cycles of
many plant pathogens. Recognizing this, Shriner (1974, 1975, 1977) has
examined the effects of simulated rain acidified with sulfuric acid on
several host-parasite systems under greenhouse and field conditions. The
3-21
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simulated precipitation he employed had a pH of 3.2 and 6.0, approximating
the common range of ambient precipitation pH.
Applying simulated precipitation of pH 3.2 resulted in (1) an 86 percent
restriction of telia production by Cronartium fusiforme (fungus) on willow
oak, (2) a 66 percent inhibition of Meloidogyne hapla (root-knot nematode) on
kidney bean, (3) a 29 percent decrease in percentage of leaf area of kidney
bean affected by Uromyces phaseoli (fungus), and (4) both stimulated and
inhibited development of halo blight of kidney bean caused by Pseudomonas
phaseolicol a (bacterium). In the latter case, the influence of acidic
precipitation varied and depended on the particular stage of the disease
cycle when the exposure to acidic precipitation occurred. Simulated sulfuric
acid rain applied to plants prior to inoculation stimulated the halo blight
disease by 42 percent. Suspension of inoculum in acidic precipitation
decreased inoculum potential by 100 percent, while acidic precipitation
applied to plants after infection occurred inhibited disease development by
22 percent.
Examining willow oak and bean leaves with a scanning electron microscope
revealed distinct erosion of the leaf surface by rain of pH 3.2 (see Section
3.2). This may suggest that altered disease incidence may be due to some
change in the structure or function of the cuticle (see Section 3.2.1.1).
Shriner has also proposed that the low pH rain may have increased the
physiological age of exposed leaves. Shriner (1978a) concluded his initial
experiments by suggesting that he had not established threshold pH levels at
which significant biological ramifications to pathogens occur from acidic
precipitation. He did suggest, however, that artificial precipitation of
extremely low pH probably alters infection and disease development of a
variety of microbial pathogens.
In recent years, a very serious disease of hard pines caused by a twig and
leaf pathogen called Gremmeniella abietina has increased in importance in the
northeastern United States. The disease, termed Scleroderris canker, was
first reported on red pine in New York in 1959. Currently, ^. abietina is
causing significant large tree mortality in Vermont and New York. Because it
may be more than coincidence that this region is included within the highest
acidic precipitation zone of North America, Paul D. Manion, SUNY, Syracuse,
initiated an acidic rain Scleroderris research project. The laboratory and
field studies reported to date indicate the disease may be affected by
precipitation pH, but there was no indication that abnormally high acidified
rain increased disease incidence. In fact, the opposite may be true. That
is, acidic rain may reduce the importance of the canker disease (Raynal et
al. 1980, Bragg 1982, Manion and Bragg 1982).
Armillaria mellea is an extremely important forest tree root pathogen
throughout the temperate zone. The fungus is geographically very wide-
spread, has an extremely broad host range, and is especially significant in
causing disease in trees under stress. Shields and Hobbs (1979) have indi-
cated that soil pH is related to disease development caused by /\. mellea. If
acidic deposition influences soil pH (see Chapter E-2) or tree vigor, it may
indirectly impact tree susceptibility to /\. mellea infection. In the north-
east, spruce decline in high elevation forests has been a recent concern.
3-22
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/\. mellea is associated with spruce trees exhibiting dieback and decline
symptoms in northern New England and may play an important role in the
morbidity and mortality of this species. The habitats of soil pathogens such
as _A. mellea are buffered relative to plant-surface habitats, so for acidic
deposition to influence these pathogens an-alteration of soil pH or chemistry
or host susceptibility would have to occur.
Fusiform rust caused by Cronartium fusiforme is the most important disease of
managed pines in the southeast. Bruck et al. (1981) applied simulated rain
of various pH levels to loblolly pine at the time of inoculation with rust
basidiospores. Significantly fewer galls formed on trees treated with
simulated rain at pH 4.0 or less than formed on trees treated with rain at pH
5.6.
Various bacterial species are important components of leaf microfloras. Lacy
et al. (1981) observed that populations of Erwinia herbicola and Pseudomonas
syringae were reduced on soybean leaves when host plants were treated with
water acidified to pH 3.4 relative to leaves exposed to distilled water (pH
5.7).
3.3.4 Influence on Vegetative Hosts That Would Alter Relationships with
Insect or Microbial Associate (W. H. Smith)
As Section 3.2 discussed, exposure to acidic deposition may lead to acidifi-
cation of plant surfaces, leaf cuticle erosion, and foliar lesions. Foliar
lesions could release plant volatiles attractive or repulsive to insect pests
or may serve as infection courts for microbial disease agents.
The influence of acidic deposition leached chemicals on insects infesting
tree leaves or bark could prove attractive, repulsive, or provide chemical
orientation. In the case of surface microbes, leached compounds may inhibit
vegetative growth or spore germination (alkaloids, phenolic substances) or
stimulate vegetative growth (as nutrients) or spore germination (as inducers
or nutrients--sugars, ami no acids, vitamins). Leaching of toxic radio-
elements from plant surfaces could have a restrictive impact on plant surface
biota (Myttenaere et al. 1980).
Plant growth and yield may be stimulated or inhibited by acidic deposition.
If growth is either stimulated or suppressed, it is probable that differen-
tial influence on insects and pathogens would follow. In the case of some
host-pathogen and host-insect relationships, a tree under stress is more
vulnerable to infestation or infection. Bark beetles and root-infecting or
canker-forming fungi are generally more successful in less vigorous individ-
uals. Trees exhibiting vigorous growth, on the other hand, may be predis-
posed to more serious impact from certain rust fungi and other disease
agents.
3.3.5 Effects of Acidic Deposition on Pesticides (J. B. Weber)
Pesticides are used annually to manage pests in terrestrial and aquatic
They are applied directly to animals, vegetation, soils, and/or inland
waters, but ultimately they end up in soils and/or waters. The behavior and
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fate of pesticides in the environment depend upon the following:
(1) method of application of the pesticide;
systems. The majority of these materials are organic chemicals that
selectively control unwanted and injurious insects, pathogens, or weeds.
(2) chemical properties of the pesticide;
(3) edaphic properties of the system;
(4) dissipation routes of the pesticide; and
(5) climatic conditions.
Butterfield and Troiano (1982) reported that increased acidity of simulated
rainfall (pH 5.6 to 3.0) increased the removal of a fungicide [triphenyltin
hydroxide (TPTH)] from the leaves of snap bean for both field-grown and
greenhouse-grown plants. Additional studies (Troiano and Butterfield 1982)
showed that elevated concentrations of H+, S042-, and N03- in simu-
lated rain also increased removal of fungicide from the bean leaves. It is
likely that acidic rain would increase the removal of other ionizable pesti-
cides like TPTH.
No studies on effects of acidic deposition on pesticides were found in the
literature; however, pH changes have been reported to affect factors 2
through 4 listed above.
Foliar absorption and injury from herbicides applied directly to vegetation
have been reported to be greatly enhanced by lowering the pH for both phen-
oxyacetic acid (Crafts 1961b) and dinitrophenol (Crafts and Reiber 1945) type
compounds. Acidic conditions promote formation of the un-ionized species
that more readily penetrate and injure vegetative membranes than do ionized
species. Thus, acidic deposition could conceivably result in enhanced injury
to weeds and/or crops in certain instances. The most likely possibility of
this occurring would be in herbicide applications to forests, pastures,
minimum-tillage crop production systems, or aquatic systems where the foliage
has had ample time to accumulate acidic deposition.
Significantly lowering pH of inland waters would have a substantial effect on
the direct biological activity and longevity of herbicides used in aquatic
weed and algae control. One would expect a significant increase in the herb-
icidal activity of the phenoxyacetic acid compounds. Aquatic herbicides such
as simazine would perform less satisfactorily under acidic conditions. Many
investigators (Armstrong et al. 1967, Jordan et al. 1972) have reported that
chloro-s_-triazines decompose at a much faster rate under acidic conditions.
This would make it necessary to increase the rates of chloro-£-triazine
herbicides and to make more frequent applications for satisfactory aquatic
weed control in waters where the pH levels were below normal levels.
Organic pesticides are categorized into five major types depending on ioniz-
ing characteristics (Weber 1972, Weed and Weber 1974). Examples of the five
types are:
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(1) cationic (diquat, paraquat);
(2) basic (atrazine, simazine, prometryn);
(3) acidic (2,4-D, fenac, picloram)
(4) phosphates and arsenates (glyphosate, DSMA); and
(5) nonionic (alachlor, carbaryl, methomyl).
The behavior of these materials in soils is analagous to that described for
the organic ions in Chapter E-4, Section 4.6.3. Cationic pesticides behave
similarly to inorganic cations like calcium and magnesium, basic pesticides
behave like ammonia, acidic pesticides behave like nitrates, and phosphates
and arsenates behave like phosphate and sulfate anions. Soil behavior of
non-ionic pesticides is dependent upon the water solubility, lipophilicity,
molecular size, and other properties. Changes in pH levels of waters or soil
solutions affect the ionizing properties of basic and acidic pesticides to
the greatest extent. At lowered pH levels acidic and basic pesticides tend
to be more readily adsorbed by soil particulate matter, and hence less bio-
logically active and less mobile (Weber 1972, Weber and Weed 1974). Under
such circumstances, higher rates of these pesticides would be required to
provide satisfactory performance, and the longevity of the chemicals may be
affected, depending on their modes of decomposition.
Pesticides degraded biologically would be affected by changes in microbial
populations. Captan, dicamba, amitrole, vernolate, chloramben, crotoxyphos
(Hamaker 1972), metribuzin (Ladlie et al 1976), 2,4-D and MCPA (Torstensson
1975), and prometryn (Best and Weber 1974) were reported to persist longer
under acidic conditions than under neutral conditions. Conversely, diazinon
and diazoxon (Hamaker 1972) were degraded more readily at lower pH levels.
Pesticides degraded chemically are directly affected by soil pH levels.
Malathion and parathion (Edwards 1972) persisted much longer in acidic soils
than in neutral soils, while atrazine (Best and Weber 1974) and simazine were
degraded much more rapidly under acidic conditions than under neutral
conditions.
3.3.6 Summary (W. H. Smith and J. B. Weber)
A review of the evidence on the interaction of acidic deposition with other
pollutants, and insect and microbial pests does not allow generalized state-
ments concerning stimulation or restriction of biotic stress agents, or their
activities, by acidic deposition. Certain studies report stimulation of pest
activities associated with acidic deposition treatment, while other studies
report restriction of pest activities following treatment. No studies report
significant interactive effects between acidic deposition and other pollu-
tants although potential for such effects exists.
Future research must combine both field and control 1ed-environment studies.
Mechanisms for acidic deposition impact on predisposition/protection of
forest trees to/from disease caused by microbial pathogens, and infestation
3-25
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caused by phytophagous insects must be examined. Evidence available comes
from laboratory and controlled environment studies, but no evidence on this
topic from studies employing large trees under field conditions exists.
We cannot, however, rule out the possibility of indirect, subtle interaction
of acidic deposition with other pollutants, phytophagous insects, and
microbial pathogens.
Two studies have shown that increased acidity of simulated rain increases the
removal of an iom'zable fungicide (TPTH) from plant surfaces, suggesting that
pest control may be diminished by acid precipitation. Thus, it may be neces-
sary to apply higher rates or make more frequent applications of certain
pesticides under acidic precipitation conditions. No known studies demon-
strate that acidic deposition on plant surfaces directly affects the biologi-
cal activity of pesticides. However, ample evidence shows that pH of aqueous
solutions of certain herbicides greatly affects herbicidal activity, and
observed effects were greatest between pH levels of 6.0 and 3.0. These
occurrences have been reported for herbicides applied to terrestial and
aquatic weeds.
No studies show indirect effects of acidic deposition on pesticide inacti-
vation, mobility, and decomposition in soils; however, ample evidence shows
that soil pH greatly affects all of these processes. It is likely that if
acidic deposition is found to affect soil and water pH, then pesticide
behavior and fate will likewise be affected.
3.4 BIOMASS PRODUCTION
3.4.1 Forests (S. B. Mclaughlin, D. J. Raynal, A. H. Johnson and S. E.
Lindberg)
Changing levels and patterns of emissions of atmospheric pollutants in recent
decades have resulted in increased exposure of extensive forests in Europe
and North America to both gaseous pollutants and acid precipitation. Reports
of decreased growth and increased mortality of forest trees in areas receiv-
ing high rates of atmospheric pollutant deposition have stressed the need to
quantify the rates of changes in forest productivity and identify the causes
of such changes. The complex chemical nature of combined pollutant exposures
and the fact that these pollutants may have both direct effects to vegetation
and indirect (possibly beneficial) effects makes quantification of such
effects particularly challenging. The complexity of forest growth and suc-
cession and the sensitivity of forest trees to natural environmental stresses
add further to the challenge of quantifying effects of anthropogenic pollu-
tants on forest productivity.
Such quantification requires that several critical tasks be addressed: (1)
definition of the chemical nature of the present and past air quality within
the regions of principal concern, (2) documentation of the basis for assuming
that detectable effects may be occurring within those regions, and (3)
identification of the types of effects that might be produced under present
and likely future exposure regimes.
3-26
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A critical need in evaluating stress effects on perennial forest systems is
documenting the magnitude, rate, and point of inception of historical changes
in air quality. Unfortunately, the maximum period of record for the present
National Atmospheric Deposition Program (NADP) network is four years, while
ozone monitoring data have not been collected by standardized methods in
network fashion before 1975. The most recently published estimates of his-
torical changes in isopleths of precipitation acidity (Likens and Butler
1981) suggest that initial intensification of acidity of northeastern precip-
itation may have begun in the 1950's. However, because of the limited data
points and the uncertain chemical techniques used, the validity of these
earliest data has been questioned (see Chapter A-8). Other sources of
information currently being developed include emissions inventories coupled
with regional air dispersion modeling, evaluation of historical stream and
lake chemistry data, historical reconstruction of weathering rates of marble
monuments, and analysis of changes in elemental composition of annually-
formed lake sediments and tree rings. Collectively, these techniques offer
possibilities for documenting the period of intensification of atmospheric
deposition of anthropogenic pollutants. (Further discussion of such documen-
tation can be found in Chapter A-8).
3.4.1.1 Possible Mechanisms of Response—A wide variety of potential direct
and indirect responses of forest trees to acid deposition have been hypothe-
sized based on fundamental responses of biological systems to acidity and
other stresses (Tamm and Cowling 1976). Included among these are increased
leaching of nutrients from foliage, accelerated weathering of leaf cuticular
surfaces, increased permeability of leaf surfaces to toxic materials, water,
and disease agents, altered reproductive processes, and altered root-
rhizosphere relations. In addition to the direct effects of acidity from
contact with foliage, roots, and rhizosphere organisms, a major area of
interest is the indirect effects of increased acidity on soil nutrient
availability to vegetation and the consequences of soil leaching losses to
aquatic systems (SMA 1982). Many of the key processes to be considered in
evaluating the effects of acidic deposition on forest systems are identified
schematically in Figure 3-3. The diversity of these processes illustrates
the complexity of potential interactions of acidic deposition with forest
systems and the need for better understanding of system level integration of
potential effects on multiple processes.
Forest responses must be examined both from the perspective of today's mature
forests which have been produced over the last 50 to 100 years (a period of
significant changes in atmospheric emissions) as well as with respect to the
forests of the future, which by contrast are growing under atmospheric
stresses that will likely span their entire life cycle. Thus, productivity
of these forests may be more influenced by alteration of the potentially more
sensitive life stages including reproduction, seedling establishment, and
growth.
Seedling emergence, establishment, and early growth phases are considered to
be potentially among the most susceptible stages affected (Abrahamsen et al.
1976, Likens 1976, Lee and Weber 1979, Raynal et al. 1980). Additionally,
reproductive phases of growth may be the most sensitive to acidic deposition
(Likens 1976, Cowling 1978, Jacobson 1980). Various controlled field and
3-27
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ACID DEPOSITION
-ACID RAIN (WET)
-POLLUTANT GASES (DRY)
f
DEPOSITION
[DIRECT EFFECTS!
GROWTH
VIGOR
REPRODUCTION
THROUGHFALL
I FOREST
PRODUCTIVITY
PNDIRECT EFFECTSI
NUTRIENT AVAILABILITY
TOXIC EFFECTS
MICRQBIAL PROCESSES
SOIL-PLANT
NUTRIFICATION
DENITRIFICATION
IMMOBILIZATION
RELEASE
MYCORRHIZAE
MINERALIZATION
, AQUATIC SYSTEMS N
Figure 3-3. Key components and processes to be considered in evaluating
effects of acidic deposition on forested ecosystems.
3-28
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laboratory studies in Scandinavia and in the United States have been
conducted to quantify possible effects of simulated acid rain on seed
germination, seedling establishment, and growth of trees in field plots.
3.4.1.2 Phenological Effects—Plants may respond to the deposition of acidic
substancesTnamannerwhich depends on genetic characteristics of the
species; sensitivity of individual developmental stages; timing, duration,
frequency, and severity of deposition events; and nature of meteorological
and microenvironmental conditions (Cowling 1978). Thus, a complete assess-
ment of the influences of acidic deposition on plants must include considera-
tion of phenology--changes in life cycle stages as affected by environment
and season. Seed germination and seedling emergence and establishment are
early growth phases potentially susceptible to acidic deposition (Abrahamsen
et al. 1976; Lee and Weber 1979; Raynal et al. 1982a,b). As well, mature and
reproductive phases of growth may be sensitive to acidic deposition (Likens
1976, Cowling 1978, Jacobson 1980, Evans 1982). However, differences in the
sensitivity of vegetation to acidic deposition are not documented from
natural field studies.
Plant growth, development, and reproduction may be affected by acidic depo-
sition both positively and negatively. Response depends upon species
sensitivity, plant life cycle phase, and the nature of exposure acidity.
Considerable variation in plant species susceptibilty exists, and at the
individual level the effect of acidification on different plant organs or
tissues may vary widely. Controlled environment studies indicate that the
deposition of acidic and acidifying substances from the atmosphere may have
stimulatory, detrimental, or no apparent effects on plant growth, devel-
opment, and reproduction. Both stimulatory and detrimental effects may
simultaneously occur, making determination of both acute and chronic effects
quite difficult. For example, tree seedling growth may be enhanced by
deposition of nitrate and possibly sulfate when soils are deficient in these
while, concomitantly, foliar injury may occur due to hydrogen ion deposition.
Because many biotic and abiotic factors interact to influence plant per-
formance, plant dieback or reduction in growth or yield must be evaluated in
terms of physiological stress, soil toxicity and nutrient deficiency prob-
lems, plant disease, and direct and indirect effects of acidic precipitation,
if chronic effects of deposition of acidic substances are to be fully
characterized.
3.4.1.2 1 Seed germination and seedling establishment. Laboratory studies
indicate that .a wide range of sensitivity of seed germination to acidic
substrate conditions exists (Abrahamsen et al. 1976, Lee and Weber 1979,
Raynal et al. 1982a). Studies focused on woody plants reveal that seed
germination of some species, including yellow birch and red maple, is
inhibited, but other species, such as sugar maple, are not affected when
exposed to substrate acidity of pH 3.0 or less (Raynal et al. 1982a). In
some coniferous species such as white pine and white spruce, substrate
acidity of pH 3.0 may promote seed germination, but it produces no effect in
other species such as eastern hemlock. Figure 3-4 illustrates the con-
trasting response of seed germination of three tree species to different
substrate acidity levels.
3-29
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100
90
SO
70
60
50
40 -
30 -
20 -
10 -
0
(a)
SUGAR MAPLE
& pH 5.6
a pH 4.0
• pH 3.0
_L
10 20 30 40
DAYS SINCE START OF GERMINATION
50
.
3
70
60
50
40
30
20
10
WHITE PINE
\
2 4 6 8 10 12 14 16
DAYS SINCE START OF GERMINATION
05 10 15 20 25
DAYS SINCE START OF GERMINATION
Figure 3-4. Mean cumulative percent germination of sugar maple, yellow
birch, and white pine seeds subjected to different substrate
acidity levels. Arrows indicate point at which differences
in response become significant (p < 0.05) determined by
Tukey's test for mean separation following analysis of
variance. Data show contrasting responses of species to
increasing acidity: (a) no significant difference at pH 3.0,
4.0, and 5.6 for sugar maple, (b) decreased germination in
yellow birch at pH 3.0 compared with that at pH 4.0 and 5.6
(no significant difference between pH 4.0 and 5.6), and
(c) increased germination in white pine at pH 2.4 and 3.0
compared with that at pH 4.0 and 5.6 (no significant differ-
ence between 2.4 and 3.0 or 4.0 and 5.6). Adapted from
Raynal et al. (1982a).
3-30
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Interaction of substrate solution reaction (pH) and osmotic potential may be
significant, and the effect of acidity may vary due to differences in ionic
characteristics of the germination medium (Chou and Young 1974, Abouguendia
and Redmann 1979). Leaching of various substances from the seed or fruit
coat by acidic solutions may also occur, subsequently causing neutralization.
The necessity of continually adjusting the pH of in vitro solutions to
maintain constant acidity levels in germination studies suggests that seed
tissues may effectively buffer the germination medium, thus reducing
potential detrimental effects of acidic deposition (Raynal et al. 1982a).
Under natural field conditions, vegetation canopy, litter, organic matter,
and mineral soils may further buffer emerging seedlings from highly acidic
deposition (Raynal et al. 1982b, Mollitor and Raynal 1982). Thus, seeds are
often protected from direct influence by acidic deposition and seed
germination typically may be minimally affected, as indicated by much of the
research to date.
Emergence and establishment of the seedling have been shown to be more sensi-
tive to low substrate pH than is seed germination itself (Abrahamsen et al.
1976, Lee and Weber 1979, Raynal et al. 1982b). Certain species, such as
sugar maple, show no detrimental effect of acidity on seed germination at pH
3.0 but may be inhibited at the establishment phase, as shown in studies of
effects of simulated acidic precipitation on juvenile growth (Raynal et al.
1982a,b). Injury to the emerging seedling radicle and hypocotyl may be
direct, due to hydrogen ion concentration, and/or indirect, resulting from
increased susceptibility to microbial pathogens that tolerate acidic con-
ditions and changing nutrient levels (Raynal et al. 1982b). Seedling growth
studies in which young plants are exposed to simulated acidic precipitation
have shown that juvenile plants may exhibit reduced or stimulated growth,
depending on the species (Wood and Bormann 1974, Raynal et al. 1982b).
Possible changes in soil nutrient status associated with acidic deposition
must be considered in evaluating plant growth response to acidification (see
Section 2.3). Some workers (Benzian 1965, Abrahamsen et al. 1976, Abrahamsen
1980a) have demonstrated that optimal height growth of coniferous seedlings
(including species of pine, spruce, and fir) occurs in soils having a pH
between 4.0 and 5.0. Whether hydrogen ion deposition directly influences
seedling growth or whether it, in association with the deposition of other
cations and anions, causes variation in soil nutrient characteristics af-
fecting growth is not fully known (Abrahamsen 1980a). However, at low
fertility levels, simulated acidified canopy throughfall of pH 3.0 or less
has been found to promote seedling growth in some species (Raynal et al.
1982b). Such a benefical response could result from deposition of nitrate or
other nutrients. (See Chapter E-2 for detailed discussions of forest
nutrient effects.)
Even where growth is stimulated by simulated acidic deposition (Raynal et al .
1980, 1982b), however, foliar injury may simultaneously occur in some
species. Thus, competitive promotive and inhibitory effects of acidic depo-
sition may concomitant!y affect seedling growth and development. It is,
therefore, not surprising that studies of the effects of simulated acidic
precipitation or forest canopy throughfall on plant growth have produced
3-31
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variable results, ranging from no apparent effects, to stimulation of growth,
to inhibition of growth (Wood and Bormann 1974, Matziris and Nakos 1977,
Raynal et al . 1980).
3.4.1.2.2 Mature and reproductive stages. Studies of interference of acidic
deposition on flower or cone development in flowering plants and conifers
have not been made. Should highly acidic precipitation events coincide with
floral or gamete development, pollination, or fruit or seed set, plant
populations and regeneration processes could possibly be altered. Numerous
studies reveal that various air pollutants, including sulfur dioxide and
ozone, may cause reductions in cone size and weight (Smith 1981). Studies of
air pollutant effects on pollen germination and pollen tube elongation
suggest that pollen function may be altered because of acidification of
floral tissues, including stigmas (Karnosky and Stairs 1974). Findings that
red and white pine pollen germination and tube elongation were greater in a
relatively unpolluted site compared with one of high pollution incidence
provide circumstantial evidence that pollen gametogenesis and development
potentially may be altered by acidic deposition (Houston and Dochinger 1977).
Evaluating acidic precipitation effects on plant reproduction demands that
the coupling of effects of air pollution and acidification be understood.
3.4.1.3 Growth of Seedlings and Trees in Irrigation Experiments--Abrahamsen
(1980b) has reviewed field experiments in Sweden and Norway designed to de-
termine the effects of artificial acidification on growth of forest trees and
tree seedlings. In Swedish experiments (Tamm and Wiklander 1980), young
(18-yr-old) Scots pines were irrigated below the canopy with dilute sulfuric
acid (0.16N; annual application, 50 to 150 kg ha~i ^$04 in one appli-
cation per year) both with and without prior addition of fertilizer. After 6
years of application a negative correlation between treatment acidity and
basal area growth was found on the fertilized plots (£ 10 percent decrease at
highest acidity) whereas growth responded positively (approximately 30 per-
cent increase at highest acidity level) on the unfertilized plots. Increased
nitrogen uptake was considered a probable cause of positive responses. Re-
sults of these studies were complicated by changes in nutrient availability
in the soil and associated with the effects of high acidity on soil fungi,
bacteria, and competing understory vegetation (Tamm and Wiklander 1980).
In Norwegian experiments (Abrahamsen et al. 1976, Tveite and Abrahamsen
1980), a variety of combinations of acidified groundwater treatment (pH
values between 6.0 and 2.0 by H2SOd addition), treatment volume (25 to 50
mm per month) application technique (below or above canopy), lime application
(500 to 4500 kg CaO ha"1), and tree species (lodgepole pine, Norway spruce,
silver birch, and Scots pine) were used. The overall effects of these exper-
iments were small where treatment effects were found after 4 to 7 years of
treatment application (Tveite 1980a). In studies with Scots pine, positive
growth effects were found at pH levels of 3.0, 2.5, and 2.0 after 4 years of
treatment, followed by significant growth reduction by pH 2.0 in the 5th
year. Norway spruce showed reduced diameter growth at all acid treatment
levels in the year after 6 years of prior treatment. Height growth of silver
birch was stimulated by rainfall acidity. Lime application had little or no
3-32
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effect on observed responses. Effects of acid irrigation on foliar nutrient
levels were also generally small (Tveite 1980b).
In evaluating the results of the Scandinavian irrigation experiments
Abrahamsen (1980b) concluded that the data give "no substantial evidence of
effects on tree growth at acidity levels presently found in precipitation."
However, he cautions that acid effects produced, particularly at highest
acidity levels, may be partly attributable to soil effects that were arti-
facts of the highly acid treatment levels and hence not representative of
longer-term responses to be expected under actual field conditions.
Such results corroborate findings of researchers in the United States who
have demonstrated differential effects of simulated acidic precipitation on
plant growth (Wood and Bormann 1977; Raynal et al. 1980, 1982b). Conclusions
regarding plant growth response from experiments where vegetation and soils
have been subjected to accelerated acidic deposition rates or concentrated
acidic inputs must be viewed with caution, however, for reasons discussed in
Chapter E-2, Section 2.3.1.
3.4.1.4 Studies of Long-Term Growth of Forest Trees—The evidence for
effects of regional-scale anthropogenic pollutants on productivity of forests
comes from a limited number of studies in the United States and Europe in
which long-term growth trends determined from tree rings have been analyzed.
In Scandinavia, where acid precipitation was first recognized and studied as
an environmental problem, research on changing patterns of tree growth based
on tree-ring chronologies have provided circumstantial evidence of growth
declines that occurred at about the time acidity of rainfall is thought to
have intensified. In Norway, research by Abrahamsen et al. (1976) and Strand
(1980) showed a decrease in growth (generally less than 2.3 percent per year)
of Norway spruce and Scots pine that became evident around 1950, primarily in
the eastern third of the country. These responses could not be clearly
associated with the geographical patterns of most acid rainfall, which
occurred in the southern (pH average = 4.3) rather than the eastern (pH
average = 4.5) part of the country. Some drawbacks of these studies,
however, were that individual sites were not characterized with respect to
soil chemical characteristics, and neither the influences of climate nor
aging trends were removed from the data.
Preliminary analysis of differences in responses between sites of differing
productivity class (high vs low) in southern Norway showed no differences in
response to acidic precipitation (Abrahamsen et al. 1976). On the other
hand, studies in Sweden by Jonsson (1975) and Jonsson and Sundberg (1972)
involving Scots pine and Norway spruce showed similar temporal trends in
growth reduction beginning around 1950, and these effects were most pro-
nounced in areas of greatest expected susceptibility to acidic deposition.
Site susceptibility was estimated based on the average pH of precipitation
and pH and ion content of lakes and rivers in 1965 and 1970 and the distri-
bution of soil types. Jonsson (1975) concluded from these studies that
"acidification cannot be excluded as a possible cause of poorer growth
development, but may be suspected to have had an unfavorable effect on growth
within the more susceptible regions." Differences in growth reductions
3-33
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between susceptible and non-susceptible regions were estimated to be in the
range of 0.3 to 0.6 percent per year.
A second study has been initiated covering an additional 9 years, 1965-74,
since the first survey was completed (Jonsson and Svensson 1983). These data
confirmed the earlier downward trend beginning in 1950 but showed a period of
increased productivity beginning in the mid- to late 1960's. At sites of
relatively poor quality, growth of both pine and spruce in the 1970's had
increased substantially since its minimum in the mid-60's but was still sub-
stantially less than that attained up to 1940. The overall trend was still
downward over the interval 1910-74. By contrast, growth of these species on
good sites showed an upswing in the 1965-74 interval which resulted in a
growth rate equal to or above that attained during the preceding 50 years.
In explaining these trends and summarizing the results of their own and the
Norwegian SNSF project the Swedes make the following statements (SMA 1982).
"A conceivable explanation of these changes is that the mathemati-
cal model that was used has not compensated for or caught those
effects in the ground that are the results of more long-term
cyclical changes in climate. These changes may, for example,
affect the supply of nitrogen in the ground that is available to
plants. It must also be noted that the Swedish forests have to
take increased quantities of nitrogen that are deposited along
with precipitation. This gives a fertilizing effect. There are at
the present time no clear signs or evidence of either increased or
reduced forest production resulting from the effects of acid
precipitation on Scandinavian forestland and its fertility."
The final report on the Norwegian SNSF project makes the point that:
"decreases in fores't growth due to acid deposits have not been
demonstrated. The increased nitrogen supply often associated with
acid precipitation may have a positive growth effect. This does
not exclude, however, the possibility that adverse influences may
be developing over time in the more susceptible forest ecosystems.
The most serious consequence for terrestrial ecosystems of re-
gional acidification at levelscurrently observed in Norway may be
the increased rate of leaching of major elements and trace metals
from forest soils and vegetation. This also has a bearing on the
aquatic systems receiving these effluents. From an ecological
point of view it is difficult to forecast the ultimate results of
the atmospheric acidification and related air pollutants on ter-
restrial systems and to judge the rate and even the direction of
changes. In the more susceptible areas it seems, however, to be a
question of proportion and time required rather than whether any
ecological effects appear or not."
In examining the Scandinavian work it is important to note that the character
of their atmospheric emissions and the chemistry of their rainfall have
changed dramatically in recent years, resulting in substantial increases in
nitrogen inputs from the atmosphere. Emission of SOg in Sweden increased
3-34
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85 percent (from 240 to 445 thousands of tons of S yr-1) during the
interval 1950 to 1970, but had decreased back to 240 tons yr'1 by 1978.
Sulfate in precipitation showed a substantial (65 percent) increase (from 55
to 90 microequivalents per liter) during the interval 1955 to 1964, but then
remained constant through 1974. By contrast, nitrate levels increased by 33
percent (15 to 20 meq £-1) from 1955 to 1964 and by 1974 had reached 35
meq r1, a level 133 percent above that in 1955 (SMA 1982). Thus, while
it will be difficult to interpret the Scandinavian tree-ring studies until
both climatic and age-related trends are removed from the data, the most
recent analysis suggests the possibility that relatively recent significant
increases in atmospheric inputs of nitrogen (coupled^with the trends in
atmospheric chemistry) may be an important factor in mostr^e$ent changes in
growth patterns.
In the United States, Cogbill (1976) examined growth of beech, birctv. and
maple in the White Mountains of New Hampshire and red spruce in the Smajcey
Mountains of Tennessee. From analysis of tree-ring chronologies, he con-
cluded that no synchronized regional decrease in radial growth had occurred.
The ring chronologies presented for all of the species he studied, however,
showed evidence of a decreasing growth trend from around 1960 until 1970.
More recent studies in New York by Raynal (1980) with red spruce and white
pine, and by Johnson et al. (1981) in the New Jersey pine barrens with pitch,
shortleaf, and loblolly pine, have shown patterns of decline among most of
these species during the past 26 years.
In New Jersey, a strong statistical relationship between annual variation in
stream pH and growth rates suggested that acidic precipitation may have been
a growth-limiting factor for the past two decades (Johnson et al. 1981).
Stream pH, in this poorly buffered soil system, was closely correlated with
precipitation pH during a 36-month period of concurrent records. Of the
trees examined, approximately one-third showed normal growth, one-third
showed noticeable abnormal compression of annual increments during the past
20 to 25 years, and the remainder showed dramatic reduction in annual growth
over this time interval. This effect was evident in trees of different
species and at different sites and occurred regardless of age or whether
trees were planted or native. An interesting response of both these trees
and the four species examined by Puckett (1982) in southeastern New York was
a change in the influence of climate on tree growth over the past 25 to 30
years. Increased sensitivity of trees in these studies to climatic variables
suggests the possibility that changes in the physiological relationship of
these trees to their growing environment may have occurred during recent
decades.
Of the above studies, only that of the pine barrens by Johnson et al. (1981)
examined the possible influences of gaseous pollutants on observed growth
trends. In those studies, growth reponses were demonstrably unrelated to
03 levels. Although uncertain, we might anticipate that gaseous air
pollutants would also have played only a minor influence on growth trends
observed in Scandinavia where the density of gaseous pollutant sources is
rather low and concentrated in coastal areas (SMA 1982). In central Europe
where dieback and decline of silver fir, Norway spruce, and beech has
occurred (German Federal Ministry of Food, Agriculture, and Forestry 1982)
3-35
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and in inland areas of the eastern United States, contributions of gaseous
pollutants, primarily 03 and S02, nave changed over the same time spans
as has acid precipitation and thus should be considered in any study of
long-term growth effects.
3.4.1.5 Dieback and Decline in High Elevation Forests—Within the United
States, the forests presently receiving the highest levels of acidic depo-
sition are those at high elevations in the northeast. Forests characterized
by varying proportions of spruce, fir, and white birch occur at the high ele-
vations of the Appalachian Mountains from eastern Canada to North Carolina.
The northern boreal forests of New York, Vermont, and New Hampshire have
received considerable attention with respect to the potential for acidic
deposition impacts. Although the mountain summits are remote from large
point sources of sulfur, they receive extraordinarily high rates of H+,
sulfur, and heavy metal deposition (Lovett et al. 1982, Friedland et al.
1983). In addition, the vegetation is subjected to very acid cloud moisture
for a considerable portion of the year (Johnson et al. 1984). Typically,
cloud moisture pH is in the range 3.5 to 3.7, whereas ambient precipitation
is about pH 4.1 to 4.3. Another cause for attention stems from the quanti-
tative documentation of a red spruce decline in the Green Mountains of
Vermont, the causes of which are obscure at present (Siccama et al. 1982).
The northern boreal forests are characterized by red spruce (Picea rubens),
balsam fir (Abies balsamea) and white birch tBetula papyrifera var.
cordifolia) in the canopy, mountain ash (Pyrus americana) and mountain maple
(Acer spfcatum) as important understory trees, and an herb layer dominated by
ferns TtTryopteris sp.) and Oxalis montana (Siccama 1974). The lowermost ele-
vation "to~whTch~~the boreal forests extend varies from 250 m above sea level
in Maine and Nova Scotia to 750 m in New Hampshire and Vermont, 900 to 1000 m
in the Adirondack and Catskill Mountains of New York, and about 1500 m in
North Carolina (Costing 1956, Siccama 1974). The presence of boreal vegeta-
tion is believed to be related to the incidence of cloud moisture, with the
boreal vegetation occupying the often cloud-capped upper slopes, and hard-
woods holding the lower elevation sites (Nichols 1918, Davis 1966, Vogelmann
et al. 1968, Siccama 1974). In the Green Mountains of Vermont, the boreal
forests are above cloud base for 800 to 2000 hours per year, depending on
elevation (Johnson et al. 1984).
Although there is considerable interest in cloud moisture pH and there are
several ongoing studies in the mountains of the Northeast (H. Vogelmann,
University of Vermont; F. H. Bormann, T. G. Siccama, Yale School of Forestry;
G. E. Likens, J. Eaton, Cornell University; V. Mohnen, J. Kadlecek, State
University of New York, Albany; C. V. Cogbill, Center for Northern Studies),
there are few published data. Data from especially designed cloud moisture
collectors at Mt. Moosilauke, NH, indicate that growing season cloud moisture
pH is generally in the mid-3 range (Lovett et al. 1982). The few reported
cloud pH measurements obtained from airplane flights suggest that growing
season cloud moisture pH is distinctly lower than moisture precipitated from
the cloud, and that clouds are most acid near cloud base (Scott and Laulainen
1979). The current indication is that cloud moisture pH is approximately 0.5
pH units lower than ambient rain or snow pH, but considerably more data are
3-36
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needed to characterize the nature of cloud acidity. The implication is that
boreal forest vegetation is exposed to moisture with pH of 3.0 to 4.0 fre-
quently and for a total of 30 to 80 days per year.
In the mountainous areas of New England, precipitation increases with alti-
tude. Lovett (1981, in Cronan 1984) estimates precipitation rates of 240 cm
yr-1 in the balsam fir forests of New Hampshire. Low-elevation precipita-
tion in New England ranges from about 100 to 150 cm yr"1. Siccama (1974)
determined that growing season throughfall increased by 2.9 cm 100 m-2 in
the Green Mountains of Vermont due to increased rainfall and an increase in
the cloud moisture intercepted by vegetation. Vogelmann et al. (1968) report
that at 1087 m in the Green Mountains, open collectors fitted with screens to
intercept cloud moisture collected 66.8 percent more water than control
collectors without screens. Throughfall collectors placed under balsam fir
at 1250 and 1300 m in the White Mountains collected 8 percent more water than
precipitation collectors placed in the open at the same elevation, and 36
percent more water than precipitation collectors located at 520 and 640 m.
Thus, high precipitation rates coupled with intercepted cloud moisture
probably produce H+ deposition rates far in excess of the regional rates
reported by precipitation collection networks based on samples collected at
lower elevation.
Cronan (1984) estimated H+ input to the canopy at 77 to 100 meq nr2 for
the 6 month period May through October, 1978 in the high elevation fir
stands. The hardwood canopy at 520 and 640 m received 50 to 62 meq H+
nr2 during this period. Based on Cronan1 s data, it appears that the boreal
forest canopy is not effective at neutralizing atmospherically deposited H+
as throughfall collectors indicated that the H+ input to the forest floor
under fir was 98 mg nr2 for the growing season. Probably the best esti-
mate of H+ deposition has been made by Lovett et al. (1982), who used field
collection of cloud moisture samples and modeling of cloud droplet inter-
ception to estimate H+ deposition in the subalpine zone of the White
Mountains to be ~ 340 meq nr2 yr-1.
As a result of the substantial input and the inferred low neutralization
capacity of the canopy (Cronan 1984), the potential for accelerated leaching
of bases is high, but to date, no quantitative data from high elevation
forests indicate that the rate has actually increased over the past few
decades. Changes in soil pH are not expected to be rapid, as the forest
floor of the boreal zone soils is naturally extremely acid. Siccama (1974)
reported soil pH in HgO of 3.4 to 3.7 in the forest floor (0 horizons) at
Camels Hump, Vermont in the mid-1960's. Johnson et al. (1984) found that at
the same sites, pH was slightly but not significantly higher in 1980.
Estimates of dry deposition have not been made for high-elevation forests,
but as wind velocities increase with altitude (Siccama 1974) and as conifers
have a high surface area and have foliage all year, dry deposition may add
substantially to the quantity of atmospherically deposited H+ processed.
A decline of red spruce (but not fir or white birch) has been quantitatively
documented in the Green Mountains of Vermont (Siccama et al. 1982) and
3-37
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observed in New York and New Hampshire (Johnson et al. 1984). An overall
reduction of approximately 50 percent in basal area and density was observed
in the Green Mountains between 1965 and 1979. Trees in all size classes were
affected. The primary cause is presently unknown, but it is not likely to be
successional dynamics, climatic changes, insect damage, or primary pathogens
(Hadfield 1968, Roman and Raynal 1980, Siccama et al. 1982). Studies of
pathogens in declining spruce indicate the presence of secondary fungal
pathogens, with Armlllarla mellea, Forces pini, and Cytospora kunzii most
prominant (Hadfield iyb«). RadTield 119681speculated that the infected
trees had been weakened by the drought of the early 1960's prior to invasion
by the fungi. Using the framework of Manion (1981), the spruce decline has
the characteristics of a complex biotic-abiotic disease related to environ-
mental stress. Currently, there are no data which implicate acidic depo-
sition as a contributing stress, nor are there data which rule out all of the
possible pathways by which acidic deposition could affect forest trees.
At present, serious dieback of spruce (Picea abies) and fir (Abies alba) is
under study in Germany (see also Section 3.4.1.6). From long-term, inten-
sive, ecosystem-level studies, Ulrich (Ulrich et al. 1980; Ulrich 1981a,b^
1982) suggested that acidic deposition has contributed to changes in H
generation and consumption which have caused soil acidification, mobilization
of Al, mortality of fine roots, and ultimately, dieback and decline in
spruce, fir, and beech (Fagus sylvatica). That contention is based on care-
ful documentation of changes in soil solution chemistry, a nearly parallel
decrease in fine root biomass and increase in soil solution Al concentrations
during the growing season, and nutrient solution studies which indicated that
the ratio of uncomplexed Al (i.e., A13+) to Ca found in the soil solution
was sufficient to cause abnormal root growth and development. While those
findings suggest the possibility of Al toxicity, they are not definitive.
Bauch (1983) determined that the roots of declining spruce and fir were Ca
deficient, but had the same levels of Al as healthy spruce and fir. Rehfuess
(1981) has observed declining fir on calcareous soils which would seem to
preclude Al toxicity or Ca deficiency in those cases. More recently, how-
ever, Rehfuess et al. (1982) noted Mg and possible Ca deficiencies by foliar
analysis even in base-rich soils. They speculate that accelerated foliar
leaching may be responsible (see Section 3.2.1.2). Rehfuess points out that
the parallel change in soil solution Al and fine root biomass noted by Ulrich
was not synchronized in that marked decreases in fine root biomass preceded
the increase in soil solution Al. Rehfuess cites several studies (Goettsche
1972, Deans 1979, Persson 1980) in support of his contention that late summer
declines in fine root biomass are naturally controlled, and need not be
related to Al levels. Ulrich's extrapolation of nutrient solution Al:Ca
levels to the field situation are also questionable because the soil matrix
may alter the availability of those and other plant-essential or phytotoxic
elements.
The hypothesis of Ulrich appears to have limited applicability to the North
American spruce decline, where dieback and decline is most prominent in the
high elevations where soils are Borofolists or Cryofolists which have ~ 80
percent organic matter by weight (Friedland et al. 1983), and Al toxicity
would likely be masked by complexation with organic matter (Ulrich 1982).
3-38
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Data on spruce root chemistry from Camels Hump, VT, indicate that Ca:Al
ratios increase with increasing elevation. As mortality increases with
elevation, it is not likely that imbalances of Al and Ca in root tissue are
the major cause of spruce decline (Lord 1982, Johnson et al. 1984).
Whether the red spruce decline is related to acidic deposition has been the
focus of considerable speculation. The decline is widespread, easily dis-
cerned, dramatic, and of unknown origin. It has occurred in an environment
that receives very high annual input of H+ from the atmosphere and where
trees are frequently subject to extremely acid cloud moisture; hence, it is
logical that research on acidic deposition effects in high-elevation forests
has been initiated.
At present, there are few testable hypotheses regarding how acidic deposition
could have contributed to spruce mortality. The Al toxicity proposed by
Ulrich (1981a,b; 1982) is not supported by the data collected to date. The
foliar leaching hypothesis of Rehfuess et al. (1982) remains untested as yet,
however.
The spruce decline appears to be a stress-related disease. The trees are
probably predisposed to decline by the site conditions whereby some short-
term stress, possibly the drought of the early 1960's, triggered a loss of
vigor, and where biotic stress imposed by fungal attack is sufficient to
cause widespread mortality. Acidic deposition could act to intensify the
predisposing stresses, exacerbate the effects of the triggering stress, or
increase the susceptibility to fungal attack, and these possibilities warrant
research in the future.
3.4.1.6 Recent Observations on the German Forest Decline Phenomenon--
Summaries of technical presentations at an international conference on acidic
deposition (VDI 1983) and observations made during a guided field trip
through forests of West Germany have recently become available (Lindberg
1983). These observations indicate the serious nature of the forest decline
in Europe and suggest several hypotheses for the observed effects. Recent
surveys of West German spruce forests indicate extensive areas of dead and
dying trees (Knabe 1983). The problem is thought to be air pollution plus
drought stress. Effects were seen as early as 1972 but became much more
extensive from 1979, when fully vital needles were 20 percent of the total
tree in affected areas, to 1981 when they were only 3 percent of the total
needles. Symptoms include yellow and red-brown needles, crown death, and
branch loss in the middle of the trees. Discolored needles are low in Ca and
Mg compared to green, while dead branches are enriched in Cu, Mn, and Si.
Surveys of plots in the Black Forest in southwestern West Germany indicate
considerable damage (Schroeter 1983). Approximately 30 plots of 2.5 x 10s
m2 each were checked (750 spruce and fir trees) every six months, with the
result that 65 percent of silver fir trees and 100 percent of spruce checked
in 1980 were without damage, while only 1 percent of fir and 5 percent of
spruce fell into this category in 1982. The author felt that no single
factor could account for such drastic losses.
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Sulfur dioxide levels in the Black Forest ranged up to 150 yg nr3 with
means in the 20 to 40 range; westerly winds result in highest levels (Arndt
1983). Ozone episodes of greater than 240 yg nr3 (hourly average) also
occur with SW winds; 03 levels at 900 m in the Black Forest exceed those in
large cities in the valley.
Damage to trees in the Black Forest began at the higher elevations, but has
moved downslope rapidly (Krause et al. 1983). The effects are not age
specific, with affected trees ranging in age from 3 to 100 years. There
seems to be a shading influence on needle chlorosis with undersides of
needles and shaded branches not affected. There are differences in Ca, Mg,
and $04^- levels between green and yellow needles for silver fir, Douglas
fir, and spruce (factors of 3 to 8 lower in yellow). Laboratory experiments
were used to test the hypothesis that 03 was involved in conjunction with
acid (or any) rain. Results showed that Ca and Mg leaching increased,
needles yellowed, and photosynthesis decreased with increasing 03 exposure.
Rehfuess (1983) believes the problem for Norway spruce to be Ca and Mg
deficiency in foliage due to enhanced leaching from dry deposition of $03
and HN03 plus rain and fog deposition of acids, and that soil-mediated
effects are not the only explanation. Ulrich (1983) continues to discuss his
Al toxicity/fine root death theories (summarized earlier) but adds that his
recent data indicate that increased acid deposition can also lead to deple-
tion of Ca and Mg with replacement by Al, can mobilize toxic heavy metals,
can exceed the normal buffering capacity of the canopy, and can act in
conjunction with $02, 03, and climatic effects to cause such acute
problems as are occurring in the German forests.
Considerable discussion continues at this time concerning ideas that such
rapid demise of large forest areas could not be solely pollution related, but
must involve a plant disease as well (e.g., lichens normally sensitive to
some air pollutants are unaffected in these forests, supporting this theory).
On the other hand, this could be a rapid manifestation of a chronic problem
of exposure to gaseous pollutants and wet/dry deposited acids and metals over
several years. In the Black Forest, one or more factors are adversely af-
fecting the vitality of numerous forest stands. Plant pathologists and
physiologists are beginning to study the vegetation, along with atmospheric
and soil chemists, to unravel the complex mixture of factors influencing the
trees. The higher 03 levels, considerable rain, and numerous fog days
combined often with poor soils, previously disturbed sites, and non-native
vegetation in many areas are, not surprisingly, all factors which can have
adverse effects.
Nearly all scientists present at the recent German conference on acid depo-
sition agreed that further research was needed, but some insisted that the
problem is serious enough to warrant immediate federal action. The German
forest dieback phenomenon is widespread and increasing in area affected, and
it is apparent that the role of heavy metals and gaseous pollutants in
conjunction with acid deposition is being increasingly considered in the
analysis of forest death, is related to a complex mixture of site charac-
teristics, climatic conditions, and air pollution, and is being studied
3-40
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vigorously. The results of ongoing, but only recently initiated, research in
West Germany should begin to address the many possible hypotheses regarding
the forest dieback during the next few years.
3.4.1.7 Summary--At present there is no proof that acidic deposition is
currently limiting growth of forests in either Europe or the United States.
From field studies of mature forests trees it is apparent that altered growth
patterns of principally coniferous species examined to date have occurred in
recent decades in many areas of the northeastern United States and in some
areas of Europe with high atmospheric deposition levels. Recent increases in
mortality of red spruce in the northeastern United States and Norway spruce
and beech in Europe add further to the concern that forests are undergoing
significant adverse change; however, no clear link has been established
between these changes and anthropogenic pollutants, particularly acidic
rainfall. This must be presently viewed from the perspective of two possible
hypotheses: (1) recent changes are purely circumstantial and not in an way
linked to acid precipitation, or (2) we have not yet adequately studied a
very complex association in which multiple and interactive factors may be
involved and responses may be subtle and chronic.
It is too early to conclude that acidic deposition has not nor will not
affect forest productivity. Irrigation studies with seedlings and young
trees provide no indication for immediate alarm but they are difficult to
interpret because of potential artifacts of experimental protocols. De-
tecting responses of mature forest trees is made difficult by the complexi-
ties of competition, climate, and site factors, the potential interactions
between acid precipitation, gaseous pollutants, and trace metals, and the
lack of control or unattended sites with which acid precipitation impacted
sites can be compared. Although the task of assessing potential impacts of
forest productivity will assuredly be difficult, the potential economic and
ecological consequences of even subtle changes in forest growth over large
regions dictates that it should be attempted.
To address these problems it will be necessary to evaluate the long-term
dynamics of forest systems over a broad enough range of environmental
conditions to document both whether systematic changes have occurred and the
extent to which such changes are linked to variables such as levels of
deposition of anthopogenic pollutants, soil fertility, moisture status,
species composition, and stand stocking. A combination of approaches will be
needed: dendroecological studies to document past growth patterns of trees
in a broad range of conditions, permanent long-term growth plots to study
changes in stand dynamics, and forest growth models to examine the potential
long-term significance of changing growth rates to forest growth and
compensation. The above approaches will be correlative in nature and should
be used to focus on the range of conditions in which responses have occurred.
However, they must also be coupled with mechanistic studies aimed at specific
mechanisms of effect before acid precipitation effects on forest productivity
can ever be conclusively established or refuted.
3.4.2 Crops (P. M. Irving)
A considerable number of studies on the vegetative effects of acidic pre-
cipitation have been published in the last 5 years. However, because of
3-41
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limitations in research design, few of these studies can be used to estimate
crop loss realistically. Among the large scale field studies which are most
potentially useful for estimating yield effects, differences in methodologies
make intercomparisons difficult and results appear to be inconsistent. The
following is a discussion of the approaches used in acid precipitation
effects studies, an analysis of the design limitations of those studies, and
a comparison of their methodologies and results.
3.4.2.1 Review and Analysis of Experimental Design—The most widely used
method for making crop loss assessments in the past has been field surveys in
which observers estimate vegetation injury from visible symptoms under
ambient conditions and subjectively relate leaf damage to yield loss.
Because visible injury to crops has never been reported as the result of
ambient acid precipitation, experiments using simulated rain in field or
controlled environment (i.e., greenhouse, growth chamber, laboratory) studies
have been used to determine the threshold acidity levels that produce visible
injury.
Three general approaches have been used to determine impacts on plants from
acidic deposition: (1) Determination of a dose-response function for a
specific species in a defined environment; (2) classification of relative
sensitivity based on morphological, physiological, or genetic characteris-
tics; and (3) determination of mechanisms of action. Both field and
control!ed-environment methodologies with simulated rain have been used in
these approaches. Only dose-response studies provide quantitative data to
estimate growth and yield effects.
3.4.2.1.1 Dose-response determination. Current methods for determining
whether crop yield losses are occurring due to acid rain exposure include
dose-response studies to mathematically relate yield to pollutant dose. The
term 'dose-response' suggests a univariate relationship; however, a number of
potentially important variables comprise 'acid rain dose1 (see next section).
Complex factorial designs and multivariate analyses may be necessary to
describe the relationships adequately. Dose-response studies of pollutant
effects on crops fall into two basic categories: (1) field studies and (2)
controlled-environment studies. Each type of study has its advantages and
limitations.
Field studies are often a more realistic means of estimating actual effects
because the experimental plants can be grown under normal environmental
conditions, especially if common agricultural practices are used. Because
different environmental conditions related to geography (i.e., temperature,
soil type, and water availability) may lead to different responses, field
studies are useful in estimating regional impacts of pollutants when similar
experiments are performed in various regions and then compared. Field
research, however, demands considerable time and labor and is thus expensive.
Adding to the expense is the need for either a high degree of replication so
that the sometimes subtle treatment effects can be observed above the dif-
ferences caused by environmental variability or for a large number of treat-
ment plots for response surface analyses. Reliable dose-response predictions
cannot usually be made without at least 2 to 3 years of replicate studies
conducted using normal agronomic practices.
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A lack of comparable unpolluted (control) plots is also a problem for field
studies in most regions. This has led to the use of such devices as open-top
chambers for the elimination of gaseous pollutants from field plots and to
the use of rain exclusion shelters. Experiments using these devices must be
designed properly for valid comparisons to be made. For example, in a study
by Kratky et al. (1974), plots of tomato plants were placed inside and
outside plastic rain shelters in the Kona district of Hawaii during a
volcanic eruption. The plants growing outside the rainshelter received rain
with a pH of 4.0 and produced no salable yield, while plants under the
shelter averaged 5 kg per plant of salable fruit. However, an explanation
other than acid rain should be considered for the Kratky study because of a
possible shelter effect. Dry deposited materials from the volcanic eruption,
possibly acidic, may have been dissolved by rainfall on leaf surfaces outside
the shelter but remained in the nonreactive dry form inside the shelter.
Thus rainfall, acidic or not, would have had an effect by acting as a wetting
agent. The problem of separating the effects of dry deposition when it
occurs in conjunction with wet deposition is one facing all field
researchers.
Controlled-environment studies are useful indicators of potential effects and
may suggest subtle changes not measureable in an uncontrolled situation.
Controlled studies also allow the investigator to reduce the dimensionality
or number of variables in the experiment. These types of studies, for
example, may be necessary to determine which characteristics of rain (i.e.,
intensity, droplet size, ionic composition) must be simulated in field
studies. Their use is limited, however, because plants may be more sensitive
to stress when grown under short photoperiod, low light intensity, medium
temperature, and adequate soil moisture (Leung et al. 1978), conditions which
frequently occur in a growth chamber or greenhouse as compared to the field.
Since controlled-environment studies may overestimate acid rain stress
because of greater plant sensitivity, they should be used with caution when
assessing potential damage. For example, Lee and Neely (1980) found
chamber-grown radish and mustard greens to be more sensitive to simulated
acidic rain than were field-grown plants. Troiano et al. (1982) observed
that greenhouse-grown plants developed foliar injury more readily from acid
rain simulants than did field-grown plants. Since light intensity and wind
speed affect cuticular development (Juniper and Bradley 1958), which in turn
affects leaf wettability, greenhouse-grown plants may be affected more by
acidic deposition than field plants because of decreased wax development (see
Section 3.2.1.1). On the other hand, under some conditions plants may be
more stressed in controlled environments (due to restricted root growth or
lower photosynthetic rate) and thus less susceptible to treatment stress
because of lower metabolic rates and thus lower pollutant uptake.
Soil factors, nutrition and cultural practices (i.e., application of
fertilizer, pesticides and other chemicals, irrigation, planting schedules)
may all affect the sensitivity of a plant to pollution and therefore should
be recorded in experimental methods and, for greater accuracy, should reflect
common agricultural conditions as closely as possible. To determine the
interaction of these factors with pollutant effects, controlled-environment
studies are necessary.
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Pollutants rarely occur alone, and because pollutant combinations have been
found to cause more-than-additive or less-than-additive effects (Ashenden and
Mansfield 1978, Jacobson et al. 1980), the concentrations of other pollutants
should be monitored and reported in conjunction with acid precipitation
studies. Exposures of various pollutant combinations in controlled studies
are necessary to determine interactive effects.
3.4.2.1.2 Sensitivity classification. There may be considerable variability
in sensitivity to pollutant stress between plant communities, species within
communities, cultivars within species, and growth stages of cultivars
(Heggestad and Heck 1971; see also Section 3.4.1.2). Gaseous pollutants
(i.e., ozone, st>lfur dioxide) have been found to affect certain crop culti-
vars more than others, and limited information indicates that this is also
true for cultivar response to acidic precipitation (see following section).
Because it would be prohibitively Expensive and time-consuming to perform
dose-response studies on all crop cultivars, some experimental studies are
aimed at identifying giant characteristics that can be used to indicate a
plant's relative sensitivity or resistance to acidic deposition. For
example, leaf wettability, which is related to surface morphology, has been
suggested as a parameter that may indicate sensitivity to acidic precipita-
tion (Evans et al. 1977a).
It has been suggested that crop classes can be grouped according to their
sensitivity to acid precipitation. Based on a study of 28 different crops,
Lee et al. (1981) reported that inhibition of marketable yield was observed
only in the dicotyledons that were studied, and within this group root crops,
leaf crops, cole crops, tuber crops, legumes and fruit crops were ranked in
decreasing order of sensitivity. But the data are contradicted by other
studies. For example, Evans et al. (1982) in a study of two root crops found
radishes to be resistant and garden beets to be sensitive to simulated acidic
precipitation.
Plant response may also be related to stage of development when exposure
occurs. The possibility that a particular life stage may be more susceptible
to an acid precipitation event than other stages must be considered when
researchers investigate and report acid precipitation effects.
3.4.2.1.3 Mechanisms. Studies that attempt to determine mechanisms of
action of an air pollutant (mechanistic) can provide information to explain
the basis of an observed plant growth response. In studies of this type,
measurements are made to determine effects on basic processes such as photo-
synthesis, respiration, transpiration, and metabolism. Examples of such
measurements include C02 uptake and emission, leaf diffusive resistance,
metabolite pools, and enzyme activities. This information may then be
interpreted and applied through the .use of plant growth models to predict
total plant response. Physiological measurements may also be used to support
and explain plant yield response. For example, Irving and Miller (1980),
using a 14C02 assimilation technique in the field, reported that SOj?
exposures reduced both photosynthesis and yield of soybeans but that acid
rain treatments had apparently stimulated the photosynthetic rates with no
effect on soybean yield. Usually physiological determinations alone are
inadequate to estimate the economic damage of pollutants to crops.
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3.4.2.1.4 Characteristics of precipitation simulant exposures. The effects
of a pollutant on crop yield may be defined by correlating yield variations
with variations in pollutant dose. Acidic precipitation, however, consists
of a number of variables that may have an effect on crop yield. For example,
the sulfate and nitrate concentrations, which are frequently correlated with
the hydrogen ion concentration of the rain, may be more important in affect-
ing plant response than the pH of the rain (Irving and Sowinski 1980). Lee
and Neely (1980) found that simulated rain acidified with sulfuric acid
resulted in a different effect on the growth of mustard green, onion, fescue,
radish, lettuce, and orchard grass than simulated rain at the same pH,
acidified with sulfuric and nitric acids (2:1 equivalent weight ratio; refer
to Tables 3-2 and 3-3 in Section 3.4.2.2). Acid rain dose should therefore
be described by concentrations of sulfate, nitrate, and other important ions
(e.g., NH4+, Ca2+, Mg2+, etc.), as well as hydrogen ion (pH). For a
complete analysis, it may be necessary to determine the effect of each indi-
vidual ion as well as their combination so that all important ions are
simulated at levels found in polluted and unpolluted rain.
Plant injury responses are a function of pollutant concentration and exposure
time or quantity (i.e., acid rain dose = [H+ x cm rain] + [S042" x cm]
+ [N03~ x cm]). Response to a given dose of gaseous pollutant is fre-
quently greater if deposited in a shorter exposure time. Response to acid
rain, however, may be positively correlated with the amount of time the leaf
is wet. When comparing experimental results, one must compare concentration
and duration of exposure to understand the response in terms of dose and
rate. In the case of acid rain, reporting the pH of applied precipitation is
inadequate without total dose or deposition of important ions (i.e., kg ha~l
of $042-, N03~, and H+), rate or intensity (i.e., cm hr-1), duration, and
and frequency. Physiological systems can be quite resilient due to activa-
tion of defense and repair systems during periods of stress. Therefore, time
between stress events may be important for repair functions. It has been
reported that the "recovery" period between gaseous pollutant exposures may
affect the total plant response. Similarly, the number of "dry" days between
precipitation events may influence the net response of a plant to acidic
deposition. Because of differences in leaf wettability, plants may respond
differently to a rain or mist; thus droplet size is yet another important
characteristic (see Section 3.2.1.1).
3.4.2.1.5 Yield criteria. Because crop production is measured in terms of
the yield of a marketable product, it is useful to express pollutant injury
in terms of the economically valuable portion of the crop. However, this is
not easily applied uniformly in experimental studies. Leaf injury estimates
have been commonly used to assess pollution damage, but economic loss is not
always closely related to leaf damage (Brandt and Heck 1968). Assessing loss
based on visible injury may overestimate or underestimate the economic loss.
For example, in a study of defoliation effects on yield, Jones et al. (1955)
found no reduction in root yield or sugar content of sugar beets after
removal of 50 percent of the leaves. Irving and Sowinski (1980) reported
increased yield of greenhouse-grown soybeans that had also exhibited necrosis
as a result of acid rain exposures. Increased yield was also reported by Lee
et al. (1980) for alfalfa that exhibited foliar injury from acidic rain.
Conversely, chlorosis or necrosis of leaves could result in considerable
3-45
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economic loss of a crop such as lettuce or mustard greens without causing
measurable changes in leaf weight.
3.4.2.2 Experimental Results—To allow comparisons of acid precipitation
effects research by investigators using various techniques, it is necessary
(although perhaps not sufficient) to describe the experimental conditions,
the dose, and the responses for each investigation in comparable units.
Accordingly, calculations were made, based on information in the literature
or by personal communication, to describe each investigation in comparative
terms. These changes in units were made only for comparison purposes. None
of the experimental results described below have been changed from those of
the original author. Given the experimental design limitations discussed in
the previous section, conclusions based on the following research results
must be made cautiously.
3.4.2.2.1 Field studies. The studies described in Table 3-2 were performed
in the field, using accepted agricultural practices to the extent experi-
mental design would permit. Because hydrogen, sulfate, and nitrate ions are
the components of precipitation that are believed to most likely affect the
growth and yield of crops, they were used to describe the precipitation dose.
In all experiments, simulated rain was applied at regular intervals during
the life cycle of the crop and, except for 'Beeson' and 'Williams' soybeans,
was applied in addition to ambient precipitation. Thus, total deposition
received by the crop is the sum of simulant plus ambient loadings.
Among the 14 crop cultivars (9 species) studied, only one exhibited a con-
sistently negative yield effect at all acidity levels used (garden beet),
three were negatively affected by at least one of the acidity levels used in
the study ('So. Giant Curled1 mustard green, 'Pioneer 3992' field corn, and
'Amsoy' soybean), and six had higher yields from at least one acidity level
('Champion1 and 'Cherry Belle' radish, 'Vernal' alfalfa, 'Alta' fescue,
'Beeson1 soybean, and 'Williams' soybean). The most frequent response
reported to result from simulated acidic rain was "no effect" ('Red Kidney'
kidney bean, 'Davis1 and 'Wells' soybean, 'Cherry Belle' radish, 'So. Giant
Curled1 mustard green, 'Improved Thick Leaf spinach, and 'Vernal' alfalfa).
Some experiments demonstrated both positive and negative response to acid
rain, depending on the H+ concentration. There is little evidence for a
linear response function, however, because no effect frequently occurred at
doses greater than those producing positive or negative response. Except for
garden beet, this was true for each study that reported a negative response
to at least one level of acidic deposition. For example, a 9 percent de-
crease in the yield of corn resulted from treatments with 42 times the
ambient H+ deposition (six times ambient H+ concentration), but no effect
occurred at 132 and 187 times (pH 4.0, 3.5, 3.0, respectively). In the
garden beet study, the yield decrease from acid rain was not the result of
lower beet root weights but because of fewer number of marketable roots per
plot. Perhaps the acid rain treatments affected germination or seedling
establishment. The ratio of sulfate to nitrate ions in the precipitation
simulant also affected the response of some plants (i.e., alfalfa, fescue,
mustard green; Table 3-2), independent of pH.
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TABLE 3-2. FIELD RESEARCH ON CROP GROWTH AND YIELD AS AFFECTED BY ACID PRECIPITATION
co
i
Total deposition
kg ha-1
(simulant + ambient)
H+ S042- N03-
Simulant concentration
mg fl
H* S042- N03- S042~:N03~
Rate
cm hr-1
No.
Events
hr/
events
Droplet
size
urn
PH
Effect*
Alfalfa, 'Vernal', Hedlcago saliva L. (Lee and Neely 1980)
0.017
0.171
0.833
3.611
0.011
0.017
0.271
0.833
2.611
0.011
2.13
13.31
38.89
120.84
0.75
2.13
9.07
30.44
89.15
0.75
(Garden) Beet, '
0.077
0.078
0.082
0.090
0.077
Corn, '
0.028
0.594
1.847
5.814
0.014
Fescue
0.017
0.271
0.833
2.611
0.011
0.017
0.271
0.833
2.611
0.011
Kidney
1.14
2.26
2.26
2.26
2.26
0.30
2.26
7.89
19.64
60.85
0.30
Perfected
0.0025 0.53 0.753 0.7
0.10 4.83 0.753 6.4
0.316 14.67 0.753 19.5
1.00 46.19 0.753 61.3
0.016 1.07 0.434 2.5
0.0025 0.53 0.753 0.7
0.10 3.20 2.92 1.1
0.316 11.42 7.44 1.5
1.00 34.00 23.29 1.5
0.016 1.07 0.434 2.5
Detroit V-904', Beta vulgarls L. (Evans
0.002 1.26 ' 3.04 0.4
0.010 5.47 3.04 1.8
0.079 37.07 3.04 12.2
1.995 106.6 3.04 35.1
0.087
0.67
0.67
0.67
0.67
0.67
0.67
0.67
0.67
26
26
26
26
26
26
26
26
1
1
1
1
1
1
1
1
.5
.5
.5
.5
.5
.5
.5
.5
1200
1200
1200
1200
1200
1200
1200
1200
5
4
3
3
4
5
4
3
3
.6
.0
.5
.0
.8
.6
.0
.5
.0
4.8
Control
9% greater yield than pH 5.6
No effect on yield compared to pH 5.6
No effect on yield compared to pH 5.6
Ambient
Control
No effect on yield compared to pH 5.6
No effect on yield compared to pH 5.6
No effect on yield compared to pH 5.6
Ambient
et al. 1982)
35.0
35.0
35.0
35.0
19
19
19
19
0
0
0
0
.001
.001
.001
.001
353
353
353
353
5
4
3
2
4
.7
.0
.1
.7
.1
10% greater shoot growth, 16% greater root
yield than ambient
Lower number of marketable roots per plot
than ambient or pH 5.7
Lower number of marketable roots per plot
than ambient or pH 5.7
Lower number of marketable roots per plot
than ambient or pH 5.7
Ambient
Pioneer 3992' 7ea mays L. (Lee and Neely 1980)
4.03
19.51
67.20
198.16
0.96
(Tall), '
2.13
13.31
38.89
120.84
0.75
2.13
9.07
30.44
89.15
0.75
477?-
17.33
43.54
135.47
0.39
(T.0025 0.53 0.753 0.7
0.10 3.20 2.92 1.1
0.316 11.42 7.44 1.5
1.00 34.00 23.29 1.5
0.016 1.07 0.434 2.5
Alta', Festuca elatlor L. var. arundlnacea Schreb
2.26
2.26
2.26
2.26
0.30
2.26
2.26
2.26
2.26
0.30
Bean, 'Red Kidney' ,
13.02
98.05
5.57
5.57
0.0025 0.53 0.753 0.7
0.10 4.83 0.753 6.4
0.316 14.67 0.753 19.5
1.00 46.19 0.753 61.3
0.016 1.07 0.434 2.5
0.0025 0.30 0.753 0.7
0.10 3.20 2.92 1.1
0.316 11.42 7.44 1.5
1.00 34.00 23.29 1.5
0.016 1.07 0.434
0.67
0.67
0.67
0.67
. (Lee and
0.67
0.67
0.67
0.67
0.67
0.67
0.67
0.67
58
58
58
58
Neely
26
26
26
26
26
26
26
26
1
1
1
1
1980)
1
1
I
1
1
1
I
1
.5
.5
.5
.5
.5
.5
.5
.5
.5
.5
.5
.5
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
5
4
3
3
4
5
4
3
3
4
5
4
3
3
4
.6
.0
.5
.0
.8
.6
.0
.5
.0
.8
.6
.0
.5
.0
.8
Phaseolus vulgarls L. (Shrlner and Johnston 1981)
0.001 0.02 0.12 0.2
0.631 50.0 0.12 417
3.0
3.0
27
27
0
0
.17
.17
900
900
6
3
.0
.2
Control
9% lower yield; no effect on growth compared
to pH 5.6
No effect on growth or yield compared to
pH 5.6
No effect on growth or yield compared to
pH 5.6
Ambient
Control
24% greater yield than pH 5.6
19% greater yield than pH 5.6
No effect on yield compared to pH S.6
Ambient
Control
No effect on yield compared to pH 5.6
No effect on yield compared to pH 5.6
No effect on yield compared to pH 5.6
Ambient
Control
No effect on growth or yield compared to
12.99
5.37
2.30 0.95
2.4
pH 6.0
Ambient
aEffects are reported when statistical significance Is _< 0.05 level.
-------
TABLE 3-2. CONTINUED
co
i
-P»
CO
Total deposition Simulant concentration
kg ha-1 mg r1
(simulant + ambient)
H+ S0»2- H03- H* SOa2-
Mustard
0.033
0.189
0.535
1.629
0.039
0.033
0.189
0.535
1.6Z9
0.029
Radish,
0.106
0.130
0.231
0.733
0.105
0.139
0.169
0.243
0.915
0.138
Radish,
Radish,
0.018
0.081
0.090
0.129
0.081
Green, 'So. Giant Curled', Brasslca
2.78 1.95 0.0025 0.53
9.66 1.95 0.10 4.83
25.40 1.95 0.316 14.67
75.83 1.95 1.00 46.19
1.93 0.78 0.016 1.07
2.78 1.95 0.0025 0.53
7.05 5.45 0.10 3.20
20.20 12.68 0.316 11.42
56.33 38.04 1.00 34.00
1.93 0.78 0.016 1.07
N03-
S042':N03~
Rate
cm hr-1
Events
No.
hr/
events
Droplet
size
urn
pH
Effect*
japonlca Hort. (Lee and Neely 1980)
0.753
0.753
0.753
0.753
0.434
0.753
2.92
7.44
23.29
0.434
'Champion', Raphanus satlvls L. (Trolano et
U.UUZb 0.72
0.06 2.9
0.32 11.7
1.585 55.6
0.17
0.0025 0.72
0.06 2.90
0.32 11.70
1.58 55.60
0.16
'Cherry Belle', Raphanus satlvus L.
0.0025 0.53
0.10 4.83
0.316 14.67
1.00 46.17
0.026 0.96
0.0025 0.53
0.10 3.20
0.316 11.42
1.00 34.00
0.026 0.96
'Cherry Belle', Raphanus satlvus L.
0.002 1.26
0.010 5.47
0.079 37.07
1.995 106.6
0.087
0.31
1.4
5.8
27.6
0.31
1.40
5.80
27.6
0.7
6.4
19.5
61.3
2.5
0.7
1.1
1.5
1.5
2.5
al. 1982)
2.3
2.1
2.0
2.0
2.3
2.1
2.0
2.0
(Lee and Neely 1980]
0.753
0.753
0.753
0.753
0.471
0.753
2.92
7.44
23.29
0.471
(Evans
3.04
3.04
3.04
3.04
0.7
6.4
19.5
61.3
2.5
0.7
1.1
1.5
1.5
2.5
et al. 1982)
0.4
1.8
12.2
35.1
0.67
0.67
0.67
0.67
0.67
0.67
0.67
0.67
1.0
1.0
1.0
1.0
1.0
1.0
1.0
1.0
0.67
0.67
0.67
0.67
0.67
0.67
0.67
0.67
35.0
35.0
35.0
35.0
16
16
16
16
16
16
16
16
5
5
5
5
9
6
6
6
6
11
12
12
12
12
12
12
12
12
9
9
9
9
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1
1
1
1
1
1
1
1
1
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
0.001
0.001
0.001
0.001
1200
1200
1200
1200
1200
1200
1200
1200
1900
1900
1900
1900
1900
1900
1900
1900
1200
1200
1200
1200
1200
1200
1200
1200
353
353
353
353
5.6
4.0
3.5
3.0
4.8
5.6
4.0
3.5
3.0
4.8
5.6
4.2
3.5
2.8
3.8
5.6
4.2
3.5
2.8
3.8
5.6
4.0
3.5
3.0
5.6
5.6
4.0
3.5
3.0
5.6
5.7
4.0
3.1
2.7
4.1
Control
No effect on growth or yield compared to pH 5.6
No effect on growth or yield compared to pH 5.6
No effect on growth or yield compared to pH 5.6
Ambient
Control
31% lower yield; 29% lower root wt than pH 5.6
No effect on yield or growth compared to pH 5.6
33% lower yield; 24* lower root wt than pH 5.6
Ambient
No effect on yield but 51 higher shoot wt than
ambient
7% higher root wt (yield) than pH 5.6
7% higher root wt (yield) than pH 5.6
13% higher root wt (yield) than pH 5.6
Ambfent
12% lower root wt (yield), 7% higher shoot wt
than ambient
3% higher root wt (yield) than pH 5.6
11% higher root wt (yield) than pH 5.6
17% higher root wt (yield) than pH 5.6
Ambient
Control
No effect on growth or yield compared to pH 5.6
25% greater yield than pH 5.6
No effect on growth or yield compared to pH 5.6
Ambient
Control
No effect on growth or yield compared to pH 5.6
No effect on growth or yield compared to pH 5.6
No effect on growth or yield compared to pH 5.6
Ambient
No effect on growth or yield compared to pH
4.06 (ambient)
No effect on growth or yield compared to pH 5.7
No effect on growth or yield compared to pH 5.7
No effect on growth or yield compared to pH 5.7
Ambient
"Effects are reported when statistical significance Is £ 0.05 level.
-------
TABLE 3-2. CONTINUED
OJ
vo
Total deposition
kg ha-1
(simulant + ambient)
H+ S042- N03-
Soybean, 'Amsoy1,
Soybean
0.229
0.916
3.262
0.218
Soybean
0.198
0.496
1.976
4.965
0.216
0.431
1.717
10.834
Soybean
0.077
0.464
0.076
Soybean
0.229
0.916
3.262
0.218
Spinach
0.033
0.134
0.503
1.529
0.029
0.033
0.134
0.503
1.529
0.029
,° 'Beeson
2.88
10.21
39.97
10.51
, 'Davis1,
6.19
7.25
93.19
256.31
11.13
25.25
127.33
683.91
, 'Wells'.
9.02
18.72
8.90
Glyclne
Simulant concentration
mgrl
H* 504^- N03-
max (L.) Merr. (Evans et al
0.10 1.4 3.90
0.794 28.3 3.90
1.995 83.0 3.90
10.0 265.0 3.90
0.79 2.64 1.62
Events
S042':N03~
. 1982)
0.4
7.3
21.3
67.9
1.6
', Glyclne max (L.) Merr. (ToHano et al. 1983)
07J2 OTlO 4.321 2.18 2.1
1.16
4.49
1.40
Glyclne
4io6
13.19
32.75
6.68
7.60
14.5
63.31
Glyclne
4.26
2.82
0.40 15.28 7.84
1.58 59.18 30.45
0.10
1.9
1.9
max (L.) Merr. (Heagle et al . 1983)
0.005 0.27 0.15 1.8
0.10 1.55 0.46
0.63 28.10 3.39
1.58 80.30 9.65
0.06 1.71 0.83
0.004 0.20 0.15
0.08 0.54 0.48
0.63 13.00 3.00
3.98 248.00 21.00
0.08 3.90 2.31
max (L.) Merr. (Irving and
~~ff.0025 4.80 2.48
0.871 39.18 4.96
0.081 6.07 3.00
,b 'Williams', Glyclne max (L.) Merr. (Trolano
2.88 0.32 O.TTT 4.32 2.18
10.21
39.97
10.51
1.16
4.49
1.40
0.40 15.28 7.84
1.58 59.18 30.45
0.10
, 'Improved Thick Leaf, Splnada oleracea L.
2.72 1.91 0.0055 (JT53 O53
9.17
23.93
71.18
1.93
2.72
6.76
19.06
52.93
1.93
1.91
1.91
1.91
0.78
1.91
5.16
11.94
35.71
0.78
0.10 4.83 0.753
0.316 14.67 0.753
1.00 46.17 0.753
0.016 1.07 0.434
0.0025 0.53 0.753
0.10 3.20 2.92
0.316 11.42 7.44
1.00 34.00 23.29
0.016 1.07 0.434
3.4
7.9
8.3
2.1
1.3
1.1
14.3
11.8
1.7
Miller 1981)
1.9
7.9
Z.O
et al. 1983)
2.0
1.9
1.9
(Lee and Neely
0.7
6.4
19.5
61.3
2.5
0.7
1.1
1.5
1.5
1.0
Rate
cm hr'l
35.0
35.0
35.0
35.0
2.7
1.27
1.27
1.27
1.5
1.5
1.5
1.5
1.7
1.7
1.7
1.7
2.0
2.0
1.27
1.27
1.27
1980)
0.67
0.67
0.67
0.67
0.67
0.67
0.67
0.67
No.
41
41
41
41
29
18
18
18
13
30
30
30
30
41
25
25
25
25
29
11
11
15
18
18
18
13
15
15
15
15
15
15
15
15
hr/
events
0.001
0.001
0.001
0.001
4.7
1
1
1
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.33
0.33
1
1
1
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
Droplet
size
urn
353
353
353
353
730
730
730
900
900
900
900
900
900
900
900
1800
1800
730
730
730
1200
1200
1200
1200
1200
1200
1200
1200
pH
4.0
3.1
2.7
2.3
4.1
4.0
3.4
2.8
4.0
5.3
4.0
3.2
2.8
4.2
5.4
4.1
3.2
2.4
4.1
5.6
3.06
4.1
4.0
3.4
2.8
4.0
5.6
4.0
3.5
3.0
4.8
5.6
4.0
3.5
3.0
4.8
Effect*
No effect on growth or yield compared to pH 4.
Mo effect on growth or yield compared to pH 4.
1
1
11.5% lower seed wt; lower seeds and pods/plant
than 4.1
No effect on yield; lower number pods/plant
than 4.1
Ambient-control
Control
No effect on yield; 8% lower seed size 18%
greater seed/pods than pH 4.0
32% greater yields; 17% greater seed size than
pH 4.0
Ambient
Control
No effect on growth of yield
No effect on growth of yield
No effect on growth of yield
Ambient
Control
No effect on growth or yield
No effect on growth or yield
No effect on growth or yield
Ambient
No effect on yield; 4% greater wt/seed than
ambient
No effect on yield; 4% greater wt/seed than
5.6
Ambient
Control
No effect on yield; 171 lower seed size than
4.0
24% greater yield; 22% greater seed size than
pH 4.0
Ambient
Control
No effect on growth or yield
No effect on growth or yield
No effect on growth or yield
Ambient
Control
No effect on growth or yield
No effect on growth or yield
No effect on growth or yield
Ambient
•Effects are reported when statistical significance Is < 0.05 level.
"Field plots sheltered from ambient deposition. ~
-------
A comparison of studies on five different cultivars of soybeans by four dif-
ferent investigators appears to indicate that the 'Amsoy1 cultivar may be
more susceptible to acidic deposition than 'Beeson', 'Davis1, 'Williams' or
'Wells'; however, the experimental conditions such as soil type and charac-
teristics of the rain simulant were different for each study. Figure 3-5
indicates the location and results of the four soybean field studies in
relation to the principal production regions and soil types. The one cul-
tivar that responded negatively to acid rain treatments ('Amsoy') was grown
in an area with a sandy soil, while the other studies were in a loam soil.
The simulated rain used in the 'Amsoy1 study was applied more frequently and
also had high concentrations of heavy metals (i.e., 20 ppb Cd, 50 ppb Pb, 100
ppb F; Evans et al. 1977a) that were not present in the rain simulants used
by other investigators. The 'Beeson1 and 'Williams' cultivars, which were
studied in a location near the 'Amsoy' study, responded positively to the
acid rain treatments when ambient ozone was removed. The 'Davis' and 'Wells'
cultivars were studied in major soybean-growing areas with highly buffered
soils and had no response to acid rain treatments as much as ten times more
acidic than ambient. This comparison suggests that the region may be an
important component of response to acid precipitation because of differences
in major soil types, cultivars grown, climatic conditions, and ozone con-
centrations.
In the five separate studies of radish (two cultivars), a positive linear
correlation between yield and acidity was observed in two studies (Troiano et
al. 1982), a non-linear positive correlation was observed in another study
(Lee and Neely 1980), and no effect was reported in two studies (Lee and
Neely 1980, Evans et al. 1982). The differences in results could be due to
factors such as cultivar differences, environmental variability, or differ-
ences in total deposition of H+, S042', N0a~, or S042~ to N0s~ ratios. Ex-
perimental results from some of these studies also demonstrate that the re-
sponse of unharvested biomass is not a reliable predictor of yield response.
Effects on marketable yield will not necessarily be reflected in changes in
shoot or root growth. For example, field corn (Table 3-2) exhibited lower
grain yield at pH 4.0 but no effect on shoot growth. The results from these
studies are inadequate to indicate whether the average concentration or total
deposition of H+, S042~, and NOs" is important in determining yield response.
3.4.2.2.2 Controlled environment studies. As with the field studies, exper-
imental conditions, dose, and response in all controlled environment studies
are expressed in comparable units, based on calculations from published and
private communications (Table 3-3). To compare total deposition in Tables
3-2 and 3-3 multiply g nr2 (Table 3-3) by 10 to obtain kg ha'1 (Table
3-2). A comparison of effects on the same species grown in a controlled
environment as opposed to in the field indicates a similar response in most
species (alfalfa, spinach, mustard green, soybean) although radishes ex-
hibited a negative effect in a controlled environment and a positive effect
in the field. In general, total deposition of H+, S042-, and N03~ applied
was greater in the controlled environment studies than in the field studies
because of a higher deposition rate or greater number of exposures.
There were 34 crop varieties (28 species) studied in controlled-environment
experiments; six exhibited a negative response from acid rain exposure (pinto
3-50
-------
= 0
SOYBEANS
Crop yield - kg fur* (harvested)
1978 Census of Agriculture
0 - 1500
1500 - 2000
2000
oo
en
ANL-Argonne National Laboratory
Soil: silt loam 'Martinton
Cultivar: 'Wells'
Acidity Effect: None
Irving and Miller 1981
BNL-Brookhaven National Laboratory
Soil: loamy sand 'Plymouth
Cultivar: 'Amsoy'
Acidity Effect: Negative
Evans et al. 1981c
BTI-Boyce Thompson Institute
Soil: sandy loam ,...„..
Cultivar: 'Beeson', 'Williams
Acidity Effect: Positive
Troiano et al. 1983
NCS-North Carolina State University
Soil: sandy clay loam 'Appling
Cultivar: 'Davis1
Acidity Effect: None
Heagle et al. 1983
F,gure 3-5. Location of four 0
production regions and soil
-------
TABLE 3-3. CONTROLLED ENVIRONMENT STUDIES ON CROP GROWTH AND YIELD AS AFFECTED BY ACID PRECIPITATION
en
ro
Total deposition
gm-2
H+
Alfalfa.
0.001
0.056
0.177
0.56
Barley, '
0.001
0.045
0.142
0.45
( l events)
5042- M03-
Simulant concentration
H+
' Vernal', Hodlcago saliva
0.300 0.417
2.744 0.417
17.35 0.417
54.92 0.417
0.0025
0.10
0.316
1.00
Steptoe' , Hordeum vulgare
0.238 0.335
2.205 0.335
13.945 0.335
44.13 0.335
Beet, 'Detroit Dark Red1,
0.001
0.026
0.082
0.26
0.140 0.193
1.274 0.193
8.057 0.193
25.50 0.193
B1bb lettuce, 'Limestone'
0.0002
0.009
0.028
0.09
Bluegrass
0.002
0.072
0.227
0.720
Broccol 1 ,
0.0006
0.022
0.070
0.22
0.048 0.067
0.441 0.067
2.789 0.067
8.826 0.067
, 'Newport' , Poa
0.382 0.472
3.528 0.472
22.31 0.472
70.61 0.472
'Italian Green
0.117 0.164
1.078 0.164
6.819 0.164
21.58 0.164
0.0025
0.10
0.316
1.00
mg f1
5042- NOa"
L. (Cohen et al .
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
L. (Cohen et al .
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
Beta vulgarls L. (Cohen et
0.0025
0.10
0.316
1.00
, Lactuca
0.0025
0.10
0.316
1.00
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
satlva L. (Cohen
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
pratensts L. (Cohen et al.
0.0025
0.10
0.316
1.00
Sprouting'
0.0025
0.10
0.316
1.00
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
S042':N03
1981, Lee
0.7
6.6
41.8
132.5
1981, Lee
0.7
6.6
41.8
132.5
al. 1981,
0.7
6.6
41.8
132.5
Rate
cm hr"1
et al. 1981)
0.67
0.67
0.67
0.67
et al. 1981)
0.67
0.67
0.67
0.67
Events
No.
56
56
56
56
45
45
45
45
hr/
events
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
Droplet
size
urn
1200
1200
1200
1200
1200
1200
1200
1200
Fertilizer
N-P-K
67-252-252b
67-252-252
67-252-252
67-252-252
112-224-224D
112-224-224
112-224-224
112-224-224
pH
5.6
4.0
3.5
3.0
5.6
4.0
3.5
3.0
Effects*
Control
No effect market yield, Increased
shoot wt
31% greater market yield, Increased
shoot/ root wt
No effect growth or market yield
Control
No effect growth or market yield
No effect growth or market yield
No effect growth or market yield
Lee et al . 1981)
0.67
0.67
0.67
0.67
et al. 1981, Lee et al.
0.7
6.6
41.8
132.5
1981, Lee
0.7
6.6
41.8
132.5
, Brasslca oleracea L. var.
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
0.7
6.6
41.8
132.5
0.67
0.67
O.b7
0.67
et al. 1981)
0.67
0.67
0.67
0.67
Botry tl s L.
0.67
0.67
0.67
0.67
26
26
26
26
1981)
9
9
9
9
72
72
72
72
(Cohen
22
22
22
22
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
' 1.5
1.5
1.5
et al. 1981
1.5
1.5
1.5
1.5
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
112-224-224&
112-224-224
112-224-224
112-224-224
112-224-224b
112-224-224
112-224-224
112-224-224
224-448-448b
224-448-448
224-448-448
224-448-448
5.6
4.0
3.5
3.0
5.6
4.0
3.5
3.0
5.6
4.0
3.5
3.0
Control
No effect growth or market yield
No effect growth or market yield
43% decrease market yield; decrease
root/ shoot growth
Control
No effect growth or market yield
No effect growth or market yield
No effect growth or yield; decrease
root growth
Control
No effect market yield or growth
No effect market yield or growth
No effect market yield or growth
, Lee et al . 1981 )
1200
1200
1200
1200
168-224-224°
168-224-224
168-224-224
168-224-224
5.6
4.0
3.5
3.0
Control
No effect market yield or growth
No effect market yield or growth
25% lower market yield
aEffects are reported when statistical significance Is < 0.05 level.
^Fertilizer as kg ha-1 of N-P205-K20.
fertilizer as percentage of N-P205-K20.
-------
TABLE 3-3. CONTINUED
en
co
Total deposition
gm-2
H+
( i events)
S042- 1103-
Bush bean, 'Blue Lake 274'
0.000004
0.00017
0.000004
0.00017
0.00083
Cabbage,
0.001
0.051
0.067
0.51
Carrot, '
0.001
0.044
0.139
0.44
0.001 0.001
0.008 0.001
0.001 0.001
0 .007 0 .001
0.041 0.001
Simulant concentration
K*
mg r1
S042- N03"
, Phased us vulgarls L
0.0025
0.10
0.0025
0.10
0.631
'Golden Acre', Brassica
0.270 0.379
2.499 0.379
15.80 0.379
50.02 0.379
Danvers Half Long
0.230 0.327
2.156 0.327
13.636 0.327
43.15 0.327
0.0025
0.10
0.316
1.00
0.60 0.83
5.33 0.70
0.60 0.83
5.33 0.70
30.70 0.75
oleracea L. var
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
S042~:N03~
. (Johnston et al .
0.7
7.6
.7
7.6
40.9
Rate
cm hr-1
1982)
1.64
1.64
1.64
1.64
1.64
. Capita ta L. (Cohen et al.
0.7
6.6
41.8
132.5
0.67
0.67
0.67
0.67
', Caucus carota L. var. Satlva DC (Cohen et al .
0.0025
0.10
0.316
1.00
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
Cauliflower, 'Early Snowball', Brassica oleracea
0.0006
0.023
0.073
0.23
0.122 0.171
1.127 0.171
7.128 0.171
22.56 0.171
Corn, 'Golden Midget1, Zea
0.0005
0.020
0.063
0.20
Fescue, '
0.001
0.059
0.186
0.59
Ar ««_«*_
0.11 0.149
0.980 0.149
6.198 0.149
19.61 0.149
0.0025
0.10
0.316
1.00
mays L
0.0025
0.10
0.316
1.00
Alta', Festuca elatlor L
0.31 0.439
2.891 0.439
18.20 0.439
57.86 0.439
0.0025
0.10
0.316
1.00
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
. (Cohen et al.
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
0.7
6.6
41.8
132.5
L. var. Botry tl s
0.7
6.6
41.8
132.5
1981, Lee et al .
0.7
6.6
41.8
132.5
. var. arundlnacea Schreb. (Cohen
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
0.7
6.6
41.8
132.5
0.67
0.67
0.67
0.67
L. (Cohen
0.67
0.67
0.67
0.67
1981)
0.67
0.67
0.67
0.67
Events
No.
18
18
16
16
16
1981,
51
51
51
51
1981,
44
44
44
44
et al
23
23
23
23
20
20
20
20
hr/
events
0.67
0.67
0.67
0.67
0.67
Lee et al
1.5
1.5
1.5
1.5
Lee et al .
1.5
1.5
1.5
1.5
Droplet
size
vim
900
900
900
900
900
. 1981)
1200
1200
1200
1200
1981)
1200
1200
1200
1200
Fertilizer
H-P-K
0-20-0C
0-20-0
0-20-0C
0-20-0
0-20-0
224-224-224b
224-224-224
224-224-224
224-224-224
224-224-224>>
224-224-224
224-224-224
224-224-224
pH
5.6
4.0
5.6
4.0
3.2
5.6
4.0
3.5
3.0
5.6
4.0
3.5
3.0
Effects*
Control
No effect yield; older leaves aged
more rapidly
Control
No effect
Higher trifoliate chlorophyll;
lower shoot wt/pod number; no
effect pod wt
Control
No effect growth or yield
No effect growth or yield
No effect growth or yield
Control
27% lower market yield
45% lower market yield; decrease
shoot wt
44% lower market yield; decrease
shoot wt
. 1981, Lee et al. 1981)
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
et al . 1981 , Lee et al .
0.67
0.67
0.67
0.67
59
59
59
59
1.5
1.5
1.5
1.5
1200
1200
1200
1200
1200
1200
1200
1200
1981)
1200
1200
1200
1200
224-224-224b
224-224-224
224-224-224
224-224-224
l68-336-336b
168-336-336
168-336-336
168-336-336
l68-336-336»
168-336-336
168-336-336
168-336-336
5.6
4.0
3.5
3.0
5.6
4.0
3.5
3.0
5.6
4.0
3.5
3.0
Control
No effect growth or yield
No effect growth or yield
No effect growth or yield
Control
No effect market yield or growth
No effect market yield or growth
13% greater market yield
Control
No effect market yield; decreased
root growth
No effect market yield or growth
No effect market yield; decreased
root growth
^— ••»^.»* »>•»» i bprvi bbVI TTMVII 9
"Fertilizer as kg ha-1 of N-
'Fertilizer as percentage of N-
-------
TABLE 3-3. CONTINUED
Total deposition
g m-2 ( i events)
H+ S042- N03-
Simulant concentration
H+
•gi-1
SO*2' N03"
Green pea, 'Marvel', PI sum satlvum L. (Cohen et al
0.001
0.028
0.088
0.28
0.150
1.372
8.677
27.46
0.208
0.208
0.208
0.208
Green pepper, 'California
0.001
0.038
0.128
0.380
Kidney
0.0004
1 0.029
cn
•**• 0 .095
0.105
0.107
0.134
0.200
0.229
Lettuce
0 .00096
0.03024
0.03024
0.20
1.86
11.78
37.27
0.283
0.283
0.283
0.283
bean, 'Red Kidney' ,
0.007
2.274
7.564
8.32
9.831
10.59
15.88
18.14
, 'Oakland
0.02304
0.13824
0.78384
0.043
0.043
0.043
0.043
0.043
0.043
0.043
0.043
0.0025
0.10
0.316
1.00
Wonder1 ,
0.0025
0.10
0.316
1.00
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
Capsicum annum
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
S042':N03~
. 1981, Lee et al
0.7
6.6
41.8
132.5
L. (Cohen et al.
0.7
6.6
41.8
132.5
Events
Rate No. hr/
cm hr-1 events
Droplet Fertilizer
size N-P-K pH
um
Effects*
. 1981)
0.67
0.67
0.67
0.67
1981,
0.67
0.67
0.67
0.67
28
28
28
28
Lee et al.
38
38
38
38
1.5
1.5
1.5
1.5
1981)
1.5
1.5
1.5
1.5
1200
1200
1200
1200
1200
1200
1200
1200
67-224-2240 5.6
67-224-224 4.0
67-224-224 3.5
67-224-224 3.0
224-448-4480 5.6
224-448-448 4.0
224-448-448 3.5
224-448-448 3.0
Phaseolus vulgarls L. (Shrlner 1978a)
0.001
*
*
*
*
*
*
0.02 0.12
0.02/50 0.12
0.02/50 0.12
0.02/50 0.12
0.02/50 0.12
0.02/50 0.12
0.02/50 0.12
0.631 50 0.12
' , Lactuca satfva
0.02976
1.7856
0.95232
0.002
0.63
0.63
L. (Jacobson et
0.48 0.62
2.88 37.20
16.33 19.84
0.2/416
0.2/416
0.2/416
0.2/416
0.2/416
0.2/416
0.2/416
416.67
al. 1980)
0.77
0.08
0.82
3.0
3.0
3.0
3.0
3.0
3.0
3.0
3.0
0.80
0.80
0.80
24
24
24
24
24
24
24
24
3
3
3
0.17
0.17
0.17
0.17
0.17
0.17
0.17
0.17
2.0
2.0
2.0
900
900
900
900
900
900
900
900
900
900
900
6.0
6. 0/3. 2/6. Od
3.2/6.0/6.0
6.0/6.0/3.2
3.2/3.2/6.0
6.0/3.2/3.2
3.2/6.0/3.2
3.2
Half- strength 5.7
Hoaglands 3.2
3.2
Control
No effect growth or market yield
No effect growth or market yield
No effect growth or market yield
Control
No effect market yield or growth
20% greater market yield; Increased
shoot growth
No effect market yield; decreased
shoot growth
Control
75% Increased pod number; greater
shoot and root wt
50% lower pod number; greater
shoot wt
50% lower pod number; greater
shoot wt
No effect pod number; lower shoot/
root wt
75% greater pod number; greater
root wt
No effect pod number; lower shoot/
root wt
50% greater pod number; lower
shoot wt; greater root wt
Control
No effect growth or yield
7% Increase root wt; 24% Increase
0.03024 1.38336 0.11904 0.63 28.82
2.48
11.6
0.80
2.0 900 3.2
Mustard green, 'So. Giant Curled,1 Brasslca Japonlca Hort. (Cohen et al. 1981V Lee et al. 1981)
0.0004 0.074 0.104 0.0025 0.53 0.74 0.7 0.67 14 1.5 1200 112-224-224b 5.6
0.014 0.687 0.104 0.10 4.90 0.74 6.6 0.67 14 1.5 1200 112-224-224 4.0
0.044 4.339 0.104 0.316 30.99 0.74 41.8 0.67 14 1.5 1200 112-224-224 3.5
0.14 13.73 0.104 1.00 98.07 0.74 132.5 0.67 14 1.5 1200 112-224-224 3.0
*0.001/0.631.
•Effects are reported when statistical significance Is < 0.05 level.
fertilizer as kg ha"1 of N-P^Os-KzO-
"pH sequence Is: 10 events prior to Halo blight Infection/3 events during Infection period/11 events post Infection.
apical leaf wt
iO% Increase root wt; 29% Increase
apical leaf wt
Control
14% lower market yield
No significant effect
31% lower market yield
-------
TABLE 3-3. CONTINUED
Ol
tn
Total deposition
g m-2 (z events)
H* 504^- M03'
Oats,
0.001
0.048
0.152
0.48
Onion,
0.002
0.06S
0.205
0.65
Simulant concentration
rag i-l
H+ $042-
N03-
S042':N03'
'Cayuse', Avena satlva L. (Cohen et al. 1981, Lee et al .
0.254 0.357
2.354 0.357
14.87 0.357
47.07 0.357
'Sweet Spanish1,
0.34 0.484
3.185 0.484
20.14 0.484
63.75 0.484
Orchardgrass, 'Potomac'
0.001
0.035
0.111
0.35
Pinto
0.003
1.355
2.149
3.405
5.396
Potato
0.001
0.052
0.164
0.52
Radish
0.0003
0.012
0.033
0.12
0.19 0.260
1.715 0.260
10.85 0.260
34.32 0.260
0.0025 0.53
0.10 4.90
0.316 30.99
1.00 98.07
0.74
0.74
0.74
0.74
Allllun cepa L. (Cohen et al .
0.0025 0.53
0.10 4.90
0.316 30.99
1.00 98.07
, Dactyl Is glomerate
0.0025 0.53
0.10 4.90
0.316 30.99
1.00 98.07
0.74
0.74
0.74
0.74
L. (Cohen
0.74
0.74
0.74
0.74
bean, 'Univ. Idaho 111', Phaseolus vulgar 1s L.
8.533 1.365
64.033 1.365
102.16 1.365
162.49 1.365
258.16 1.365
0.002 5.0
0.794 37.52
1.259 59.86
1.995 95.21
3.162 151.27
, 'White Rose', Solarium tuberosun L.
0.276 0.387
2.548 0.387
16.11 0.387
51.00 0.387
, 'Cherry Belle1,
0.064 0.089
0.588 0.089
3.719 0.089
11.77 0.089
0.0025 0.53
0.10 4.90
0.316 30.99
1.00 98.07
Raphanus satlvus L.
0.0025 0.53
0.10 4.90
0.316 30.99
1.00 98.07
Red clover, 'Kenland', Trlfollim pratense L
0.001
0.056
0.177
0.56
0.300 0.417
2.744 0.417
17.35 0.417
54.92 0.417
0.0025 0.53
0.10 4.90
0.316 30.99
1.00 98.07
0.80
0.80
0.80
0.80
0.80
(Cohen et
0.74
0.74
0.74
0.74
(Cohen et
0.74
0.74
0.74
0.74
0.7
6.6
41.8
132.5
1981, Lee
0.7
6.6
41.8
132.5
Rate
cm hr-1
1981)
0.67
0.67
0.67
0.67
Events '
Mo. hr/
events
48
48
48
48
1.5
1.5
1.5
1.5
Droplet Fertilizer
size N-P-K
urn
1200
1200
1200
1200
112-224-224b
112-224-224
112-224-224
112-224-224
pH
5.6
4.0
3.5
3.0
Effects*
Control
No effect market yield or growth
Ho effect market yield; Increased
root growth
No effect market yield or growth
et al. 1981)
0.67
0.67
0.67
0.67
et al. 1981, Lee et
0.7
6.6
41.8
132.5
(Evans and
6.2
46.9
74.8
119.0
189.1
0.67
0.67
0.67
0.67
65
65
65
65
al. 1981)
35
35
35
35
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1200
1200
1200
1200
1200
1200
1200
1200
336-336-336b
336-336-336
336-336-336
336-336-336
112-224-224b
112-224-224
112-224-224
112-224-224
5.6
4.0
3.5
3.0
5.6
4.0
3.5
3.0
Lewln 1981)
0.72
0.72
0.72
0.72
0.72
al. 1981, Lee et al .
0.7
6.6
41.8
132.5
al. 1981,
0.7
6.6
41.8
132.5
. (Cohen et al . 1981
0.74
0.74
0.74
0.74
0.7
6.6
41.8
132.5
0.67
0.67
0.67
0.67
Lee et al .
0.67
0.67
0.67
0.67
. Lee et al
0.67
0.67
0.67
0.67
45
45
45
45
45
1981)
52
52
52
52
1981)
12
12
12
12
. 1981)
56
56
56
56
0.33
0.33
0.33
0.33
0.33
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
353
353
353
353
353
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
manure and
limestone
added
247-224-224b
224-224-224
224-224-224
224-224-224
112-224-224&
112-224-224
112-224-224
112-224-224
67-336-336b
67-336-336
67-336-336
67-336-336
5.7
3.1
2.9
2.7
2.5
5.6
4.0
3.5
3.0
5.6
4.0
3.5
3.0
5.6
4.0
3.5
3.0
Control
No effect market yield or growth
No effect market yield or growth
No effect market yield; Increased
shoot growth
Control
No effect market yield; decreased
root growth
No effect market yield or growth
231 greater market yield; Increased
root growth
Control
No effect yield
28% lower seed yield
291 lower seed yield
39% lower seed yield
Control
No effect yield; Increased shoot
growth
11% greater market yield; Increased
shoot growth
8% lower market yield
Control
No effect growth or market yield
Lower market yield
Lower market yield; decreased shoot
growth
Control
No effect growth or market yield
No effect growth or market yield
No effect growth or market yield
•Effects are reported when statistical significance 1s <
""Fertilizer as kg ha"1 of N-P205-K20.
0.05 level.
-------
TABLE 3-3. CONTINUED
Total deposition Simulant concentration
gm-2
H+
Ryegrass
0.001
0.055
0.183
0.58
Spinach,
0 .0004
0.014
0.044
0.14
Soybean,
0.004
w 1 .549
I
£! 6-169
CT>
Soybean,
0.002
0.002
0.002
0.105
0.105
0.105
0.700
0.700
0.700
(l events)
S042- N03- H+
, "L1nn', Loll urn perenne
0.31 0.432 0.0025
2.842 0.432 0.10
17.97 0.432 0.316
56.88 0.432 1.00
rag i-l
S042-
L. (Cohen
0.53
4.90
30.99
98.07
N03'
S042':N03~
et al. 1981, Lee et al
0.74
0.74
0.74
0.74
'Improved Thick Leaf, Splnada oleracea L.
0.074 0.104 0.0025
0.687 0.104 0.10
4.339 0.104 0.316
13.73 0.104 1.00
'Arasoy 71' , Glydne max
9.755 1.561 0.002
73.20 1.561 0.794
295.12 1.561 3.162
'Hells', Glyclne max (L.
0.669 0.721 0.0025
0.980 0.490 0.0025
1.113 0.371 0.0025
3.738 3.731 0.15
2.485 5.530 0.15
5.600 3.619 0.15
20.55 18.55 1.0
25.52 12.82 1.0
29 .40 9 .800 1 .0
0.53
4.90
30.99
98.07
(L.) Herr.
5.0
37.52
151.27
0.74
0.74
0.74
0.74
(Evans et
0.80
0.80
0.80
) Merr. (Irving and
0.96
1.40
1.6
5.34
7.09
8.00
29.36
36.45
42.00
Strawberry, 'Qulnalt1, Fragarla chlloensls
0.002
0.080
0.253
0.800
0.42 0.595 0.0025
3.920 0.595 0.10
24.79 0.595 0.316
78.46 0.595 1.00
0.53
4.90
30.99
98.07
1.03
0.70
0.53
5.33
3.55
5.17
26.50
18.32
14.00
0.7
6.6
41.8
132.5
(Cohen et al
0.7
6.6
41.8
132.5
al. 1981c)
6.2
46.9
189.1
Rate
cm hr-
. 1981)
0.67
0.67
0.67
0.67
. 1981,
0.67
0.67
0.67
0.67
0.72
0.72
0.72
Events
No.
1
58
58
58
58
Lee et al .
14
14
14
14
78
78
78
hr/
events
1.5
1.5
1.5
1.5
1981)
1.5
1.5
1.5
1.5
0.17
0.17
0.17
Droplet
size
utn
1200
1200
1200
1200
1200
1200
1200
1200
353
353
353
Fertilizer
N-P-K
112-224-224b
112-224-224
112-224-224
112-224-224
112-224-224b
112-224-224
112-224-224
112-224-224
manure and
limestone
added
pH
5.6
4.0
3.5
3.0
5.6
4.0
3.5
3.0
5.7
3.1
2.5
Effects*
Control
No effect market yield; decreased
root growth
No effect market yield; decreased
root growth
No effect market yield; decreased
root growth
Control
No effect growth or yield
No effect growth or yield
No effect growth or yield
Control
11? greater seed yield; decreased
shoot growth
lit lower seed yield; decreased
shoot growth
SowlnsM 1980)
1.0
2.0
3.0
1.0
2.0
1.5
1.0
2.0
3.0
Duchesne var. ananassa
0.74
0.74
0.74
0.74
.7
6.6
41.8
132.5
21.2
21.2
21.2
21.2
21.2
21.2
21.2
21.2
21.2
(Cohen
0.67
0.67
0.67
0.67
10
10
10
10
10
10
10
10
10
et al. 1981
80
80
80
80
0.33
0.33
0.33
0.33
0.33
0.33
0.33
0.33
0.33
, Lee et
1.5
1.5
1.5
1.5
2300
2300
2300
2300
2300
2300
2300
2300
2300
15-30-15C
15-30-15
15-30-15
15-30-15
15-30-15
15-30-15
15-30-15C
15-30-15
15-30-15
5.6
5.6
5.6
3.8
3.8
3.8
3.0
3.0
3.0
1:1 504:1103; control
2:1 S04: N03; control
3:1 S04:N03; control
No effect growth or yield compared
to 1:1 control
No effect growth or yield compared
to 2:1 control
Lower root nodule wt compared to 3:1
control; no effect yield
No effect growth or yield compared
to control
No effect growth or yield compared
to control
25% greater yield than 1:1 control,
19% greater than 2:1 control
al. 1981)
1200
1200
1200
1200
224-336-336b
224-336-336
224-336-336
224-336-336
5.6
4.0
3.5
3.0
Control
51% greater market yield; Increased
shoot growth
72% greater market yield; Increased
shoot/ root growth
72% greater market yield; Increased
shoot/root growth
•Effects are reported when statistical significance Is £0.05 level.
bFertHUer as kg ha"1 of N-PzOs-KzO.
cFert111zer as percentage of N-PzOs-K^O.
-------
TABLE 3-3. CONTINUED
CO
I
en
Total deposition
g m-2 ( z events)
H+ $04?- N03-
Simulant concentration
mg t ••
H* S042-
Strlss chard, 'Lucullus', Beta vulgarls var
0.001
0.032
0.101
0.32
Timothy
0.001
0.033
0.104
0.33
Tobacco
0.001
0.024
0.076
0.24
Tomato,
0.001
0.051
0.161
0.510
Wheat,
0.001
0.046
0.145
0.46
0.17
1.568
9.92
31.38
, 'Climax
0.17
1.617
10.23
32.36
, 'Burley
0.127
1.176
7.438
23.537
'Patio',
0.27
2.50
15.80
50.02
'Fleldwln'
0.244
2.254
14.255
45.'11
0.238
0.238
0.238
0.238
', Phleum
0.256
0.256
0.256
0.256
0.0025 0.53
0.10 4.90
0.316 30.99
1.00 98.07
L
N03-
. dcla L.
0.74
0.74
0.74
0.74
S04Z~:N03~
(Cohen et al.
0.7
6.6
41.8
132.5
pratense L. (Cohen et al. 1981, Lee et al
0.0025 0.53
0.10 4.90
0.316 30.99
1.00 98.07
21', Nlcotlana tabacum L.
0.179
0.179
0.179
0.179
0.0025 0.53
0.10 4.90
0.316 30.99
1.00 98.07
0.74
0.74
0.74
0.74
(Cohen et
0.74
0.74
0.74
0.74
Lycoperslcon esculentum Mill. (Cohen
0.379
0.379
0.379
0.379
0.0025 0.53
0.10 4.90
0.316 30.99
1.00 98.07
0.74
0.74
0.74
0.74
, Trltlcum aestlvum L. (Cohen et al.
0.342
0.342
0.342
0.342
0.0025 0.53
0.10 4.90
0.316 30.99
1.00 98.07
0.74
0.74
0.74
0.74
0.7
6.6
41.8
132.5
al. 1981, Lee
0.7
6.6
41.8
132.5
et al. 1981,
0.7
6.6
41.8
132.5
1981, Lee et
0.7
6.6
41.8
132.5
Rate
cm hr-1
Events
Mo. hr/
events
1981 , Lee et al .
0.67
0.67
0.67
0.67 .
. 1981)
0.67
0.67
0.67
0.67
32
32
32
32
33
33
33
33
1981)
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
Droplet Fertilizer
size N-P-K
jjm
1200
1200
1200
1200
1200
1200
1200
1200
168-224-224b
168-224-224
168-224-224
168-224-224
112-Z24-224b
112-224-224
112-224-224
112-224-224
pH Effects*
5.6 Control
4.0 Mo effect market yield or growth
3.5 Ho effect market yield or growth
3.0 No effect market yield or growth
5.6 Control
4.0 No effect market yield or growth
3.5 No effect market yield or growth
3.0 241 greater market yield
et al. 1981)
0.67
0.67
0.67
0.67
Lee et al
0.67
0.67
0.67
0.67
il. 1981)
0.67
0.67
0.67
0.67
24
24
24
24
. 1981)
51
51
51
51
46
46
46
46
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
,
224-448-448°
224-448-448
224-448-448
224-448-448
112-224-224&
112-224-224
112-224-224
112-224-224
5.6 Control
4.0 No effect growth or yield
3.5 No effect growth or yield
3.0 No effect growth or yield
5.6 ' Control
4.0 No effect market yield, Increased
shoot growth
3.5 No effect market yield, Increased
shoot growth
3.0 .311 greater market yield, decreased
root growth
5.6 Control
4.0 No effect market yield; decreased
root growth
3.5 No effect market yield; decreased
root growth
3.0 No effect market yield; decreased
'Effects are reported when statistical significance Is < 0.05 level.
"Fertilizer as kg ha'1 of N-P
-------
bean, mustard green, broccoli, radish, beet and carrot), eight exhibited a
positive response (alfalfa, tomato, green pepper, strawberry, corn, orchard
grass, timothy, and 'Oakland lettuce1), 17 showed no effect (bush bean,
Wells' soybean, spinach, 'Limestone1 lettuce, cabbage, cauliflower, onion,
fescue, bluegrass, ryegrass, swiss chard, oats, wheat, barley, tobacco, green
pea, and red clover), and three species showed both positive and negative
yield response depending on the H ion concentration (potato, 'Amsoy 71'
soybean), or conditions of exposure (kidney bean).
3.4.2.3 Discussion—Interpreting and comparing results of experiments on the
effects of acidic deposition on crop plants must include considering the
exposure conditions, simulant characteristics, dose rate, and total dose of
important ions (H+, S042~, and N03~). Unexplained inconsistencies
among experimental results could be due to differences in experimental design
or exposure conditions. For example, in all field studies except those of
'Champion' radish and "Beeson1 and 'Williams' soybeans, the ratio of sulfate
to nitrate in the rain simulant differed among treatments and was usually
much higher than the sulfate:nitrate ratio in ambient rain. Rain chemistry
data from the National Atmospheric Deposition Program (NADP) indicate that
weekly precipitation pH values can vary widely for a particular area (i.e., a
range of pH 3.7 to 6.8 for New York) while the S042~ to N03" ratio
appears to be independent of pH (Figure 3-6). Because preliminary evidence
indicates that plants are affected by the sulfate:nitrate ratio in rain
(Irving and Sowinski 1980, Lee et al. 1980), the differences reported among
treatments in these investigations may be the result of this ratio rather
than the hydrogen ion deposition. All published experiments used treatments
having the same chemistry from event to event although the chemistry of
ambient rain can fluctuate greatly from one event to another (Figure 3-6).
Some crops may be affected by peak concentrations of acidity while others may
respond to the total deposition of ions. No experiments separating the peak
versus total loading response have yet been reported, although Irving et al.
(1982) found that rain with a chemistry that varied from event to event had a
different effect on plant growth than did a constant rain chemistry with the
same mean pH. Johnston et al. (1982) reported that bushbeans tended to weigh
less when treated with acid rain (average pH = 3.2) in which the acidity
varied during the event as compared to a constant rain chemistry having the
same average acidity.
The majority of the 14 crop cultivars studied in the field and the 34 studied
in controlled environments exhibited no effect on growth or yield as a result
of exposure to simulated rain more acidic (usually up to 10 times more
acidic) than ambient. The growth and yield of some crops, however, were
negatively affected by acidic rain while others exhibited a positive re-
sponse. The 9 percent reduction in the yield of field corn exposed to pH 4.0
rain (0.594 kg ha'1 depositon of H+) is an alarming result; however,
treatments with greater acidity levels produced no effect on the corn yield.
The experiment was repeated a second and a third year with no statistically
significant effects observed (J. J. Lee, pers. comm.). The reduction in the
yield of one ("Amsoy") of the five cultivars of soybeans that have been stud-
ied suggests that genetic factors may control plant response to acidic rain.
If the results of these two studies are substantiated by further research,
ramifications of the negative effects of acid rain could be considerable
3-58
-------
i
o
13
12
11
10
9
8
co 7
N I*
N
N N 0 0 0 0 N
N 0 N N N
OP 0 0 ONOPNN N
NPPOPNN POO P N °
- N PONONOOOO NNPN OOON
P PON N PP P
N ON P N N
lh N P
N SYMBOL IS LETTER OF STATE
3;5 4.0 4.5 5.0 5.5 6.0 6.5 7.0
Figure 3-6. Ratio of S042~ to NO^" versus pH of precipitation in New
York (N), Pennsylvania (P), and Ohio (0) during the
growing season. Data are from the NADP network, 1979.
3-59
-------
because soybeans and field corn are two of the most economically important
agricultural crops In the United States. For reasons discussed in this
review, however, these studies do not offer definitive proof that ambient
acidic precipitation is damaging corn and soybean productivity of all cul-
tivars in all agricultural regions.
The positive response of some crops to acidic rain suggests a fertilizer
•response to the sulfur and nitrogen components of the rain. The net response
of a plant to acid rain appears to result from the interaction between the
positive effects of sulfur and nitrogen nutrition and the negative effects of
acidity. Input of nutrients to plant systems from rainfall has been docu-
mented since the mid-19th century (Way 1855). Calculations made in a number
of regions in the United States estimate the seasonal atmospheric deposition
of nutrient species, particularly sulfate and nitrate, to agricultural and
natural systems and the implications of this deposition on plant nutrient
status.
Estimates by Hoeft et al. (1972) of 30 kg S ha'1 per year and 20 kg N
ha"1 per year deposited in precipitation in Wisconsin indicated the im-
portance of atmospheric sources of these elements, although N requirements
certainly could not be completely satisfied in this way. Jones et al. (1979)
reported that atmospheric S is a major contribution to the agronomic and
horticultural crop needs for S as a plant nutrient in South Carolina.
Although the amount of S and N in a single rain event is small compared to a
fertilizer application, it is known that foliar applications of plant nutri-
ents may stimulate plant growth and yield (Garcia and Hanway 1976). The
repeated exposure of plants to rain, especially during the critical repro-
ductive stage, suggests that nutritional benefits from rain may be signifi-
cant, even in comparison to a one-time fertilizer application.
Reports of most acid rain field studies contain little or no characterization
of the soil conditions. Soil fertility may determine whether a plant re-
sponds positively or negatively to acidic precipitation. Long-term effects
of acidic deposition on poorly managed, unamended agricultural soils may have
negative effects on crop productivity through the leaching of soil nutrients
or mobilization of toxic metals. This effect has more potential for becoming
significant in those soils with low cation exchange capacity (low in clay and
organic matter), low sulfate retention capacity, and high permeability
(sands). Although such an effect may not become measurable for decades or
more, it will be most important in forage crops that are not usually highly
managed. Some speculation exists that agricultural management practices may
be modified as a result of acidic deposition but agricultural soil scientists
generally accept that the influence of acidic deposition on the need for
additional fertilizer and lime application is probably miniscule.
Another consideration that may be important in controlling the impact of
ambient acid precipitation, is that crop cultivar recommendations are based
on productivities obtained under ambient conditions of acidic deposition.
Therefore, crops currently being grown may have been selected, indirectly,
for their adaptations to rainfall acidity and the presence of other pollu-
tants.
3-60
-------
3.4.2.4 Summary--
1) Because of limitations in research design, differences in meth-
odologies and inconsistent results, it is difficult to compare
research results directly or arrive at an overall conclusion re-
garding crop response to acidic deposition without a thorough
description and comparison of experimental methods.
2) Complex factorial research designs and multivariate analyses may be
necessary to describe adequately the relationship between acid rain
dose and plant response rather than the simple univariate approach
(treatment pH vs yield) used in the past.
3) Given the above limitations to making generalizations about past
research, analysis of experimental results from field and con-
troll ed-envi ronment experiments indicates that the majority of crop
species exhibited no effect on growth or yield as a result of expo-
sure to simulated acidic rain (acidity treatments had pH values of
4.2 or less). Growth and yield of a few crops in some studies,
however, were negatively affected by acidic rain, while other crops
exhibited a positive response.
4) Interpretation of available research results suggests that the net
response of a crop to acidic deposition is the result of the inter-
action between the positive effects of sulfur and nitrogen ferti-
lization, the negative effects of acidity, and the interaction
between these factors and other environmental conditions such as
soil type and presence of other pollutants. Available experimental
results appear to indicate that the effects of acidic precipitation
on crops are minimal and that when a response occurs it may be
positive or negative. However, many crops and agricultural systems
have not been adequately studied.
3.5 CONCLUSIONS
Chapter E-3 has examined vegetative response to acidic deposition, reviewing
literature from studies that shed light on diverse plant-pollutant relation-
ships. Documented experiments concern widely varying situations, from
control 1ed-environment studies to field studies, and from intensively managed
agricultural systems to natural plant communities. Controlled-environment
studies are useful indicators of potential effects and may suggest subtle
changes not easily measurable in an uncontrolled situation. Field studies,
however, are a more realistic means of estimating actual effects because in
these studies experimental plants are grown under normal agricultural
conditions.
The following statements summarize Chapter E-3:
o Leaf structure may play two roles in the sensitivity of foliar
tissues to acidic deposition: 1) leaf morphology may selectively
enhance or minimize surface retention of incident precipitation,
3-61
-------
and 2) specific cells of the epidermal surface may be initial sites
of foliar injury. Information on the effects of acidic deposition
on the accelerated weathering of epicuticular wax of plant leaves
is very preliminary. Chlorophyll degradation may occur following
prolonged exposure to acidic precipitation (Section 3.2.1).
Leaching mechanisms are major factors in nutrient cycling in ter-
restrial ecosystems and are critical to the redistribution of
nutrients within these cycles. If the rate of leaching exceeds th
rate of mineral nutrient uptake, plant growth and yield reductions
are likely (Section 3.2.1.).
Information on which to assess the effects of acidic deposition on
nonvascular plants is inadequate. Field and laboratory studies
show that lichens and mosses are sensitive to S02; however, it
appears that the uptake of S02 is limited by the S02-induced pH
of the surface water on the plants (see Chapter A-7). Because
nonvascular plants are dependent on surface water for metabolism,
the modification of that surface water chemistry by wet and dry
deposition may be a factor in the expression of phytotoxic re-
sponse. Laboratory studies are needed to determine the rates of
uptake and physiological responses to direct acidic deposition.
These studies must be related to field observations and to deter-
mination of effects on the growth, yield, and ecosystem function of
the plants (Section 3.2.2).
Under laboratory conditions, gaseous pollutant combinations and
integration have well defined effects. However, ozone is the
single most important gas pollutant to plant life located at great
distances from the industrial and urban origin of nitrogen oxides
and hydrocarbon precursors. Direct effects due to ozone include
foliar injury and growth and yield reductions in numerous agronomic
and forest species (Section 3.3.1).
A review of the evidence on the interaction of forest trees, insect
and microbial pests, and acidic deposition does not allow gener-
alized statements concerning stimulation or restriction of biotic
stress agents, or their activities, by acidic deposition. Certain
studies report stimulation of pest activities associated with
acidic deposition treatment, while other studies report restriction
of pest activities following treatment. Further research must com-
bine field and control!ed-environment studies. Available evidence
suggests that the threshold of ambient pH capable of influencing
certain insect and microbial pests lies within the range of pH 3.0
to 4.0 (Sections 3.3.2, 3.3.3, and 3.3.4).
Performance and longevity (persistence) of certain pesticides de-
pend on the pH of the systems to which these pesticides are applied
or in which they ultimately reside; thus, it is likely that acidic
deposition will have significant but limited effects (Section
3.3.5).
3-62
-------
At present we have no direct evidence that acidic deposition cur-
rently limits forest growth in either North America or Europe, but
we do have indications that tree growth reductions are occurring,
principally in coniferous species that have been examined to date,
that these reductions are rather widespread, and that they occur in
regions where rainfall acidity is generally quite high, or pH is
low (~ pH 4.3) for an annual average (Section 3.4.1).
Controlled-enyironment studies indicate that the deposition of
acidic and acidifying substances may have stimulatory, detrimental,
or no apparent effects on plant growth and development. Response
depends upon species sensitivity, plant life cycle stage, and the
nature of exposure to acidity. Some simulation studies have indi-
cated that acidic deposition may result in simultaneous stimulation
of growth and the occurrence of visible foliar injury (Section
3.4.1).
The majority of crop species studies in field and controlled-
environment experiments exhibited no effect on growth or yield as a
result of exposure to simulated acidic precipitation (pH 3.0). In a
few studies, though, growth and yield of certain crops were nega-
tively affected by acidic deposition, while others exhibited
positive responses (Section 3.4.2).
A crop's net response to acidic deposition results from a combi-
nation of the positive effects of sulfur and nitrogen fertilization,
the negative effects of acidity, and the interaction between these
factors and other environmental conditions such as soil type and
presence of other pollutants (Section 3.4.2).
Available experimental results do not appear to indicate that the
negative effects of acidic precipitation outweigh the positive
effects, however, many crops and agricultural systems have not been
properly or adequately studied (Section 3.4.2).
3-63
-------
3.6 REFERENCES
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growth of four conifers on acidic and alkaline substrates. Forest Science
25:358-360.
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and A. Tollan, eds. Proc. of an International Conference, Sandefjord,
Norway. SNSF Project, Oslo.
Abrahamsen, G. 1980b. Acid precipitation effects on forest and fish. _^n_
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Abrahamsen, G., K. Bjor, R. Horntvedt, and B. Tveite. 1976. Effects of acid
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Adam, N. K. 1948. Principles of penetration of liquids into solids.
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Alexander, V. 1974. A synthesis of the IBP tundra biotne circumpolar study
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Alstad, D. N., G. F. Edmunds, Jr., and L. H. Weinstein. 1982. Effects of
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-------
Ascaso, C. and J. Galvan. 1976. Studies on the pedogenetic action of lichen
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Ashenden, T. W. and T. A. Mansfield. 1978. Extreme pollution sensitivity of
grasses when S02 and NO? are present in the atmosphere together. Nature
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Babich, H. and G. Stotzky. 1974. Air pollution and microbial ecology.
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Baddeley, M. S., B. W. Ferry, and E. J. Finnegan. 1971. A new method of
measuring lichen respiration: Response of selected species to temperature,
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Barkman, J. J. 1958. Phytosociology and Ecology of Cryptogamic Epiphytes.
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Bauch, J. 1983. Biological alterations in the stem and root of fir and
spruce due to pollution influences. j[n Workshop on the Effects of Accumula-
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Becker, V. E. 1980. Nitrogen fixing lichens in forests of the southern
Appalachian mountains of North Carolina. The Bryologist 83:29-39.
Becker, V. E., J. Reeder, and R. Stetler. 1977. Biomass and habitat of
nitrogen fixing lichens in an oak forest in the North Carolina Piedmont.
Bryologist 80:93-99.
Benzian, B. 1965. Forestry Commission Bulletin. London 1(37):1-251.
Best, J. A. and J. B. Weber. 1974. Disappearance of s-triazines as affected
by soil pH using a balance-sheet approach. Weed Sci. 72:364-373.
Borstrom, F. and G. R. Hendrey. 1976. pH tolerance of the first larva
stages of Lepidurus arcticus (Pallas) and adult Gammarus lacustris. G. 0.
Sars. Report, Zoological Museum, Oslo Univ., Oslo, Norway, 37 pp.
Boullard, B. 1973. Interactions entre les pollutants atmospheriques et
certains parasites des essences forestiees. (Champignons et insects). For.
Privee 94:31,33,35,36.
Bragg, R. J. 1982. Interactions of Schleroderris canker disease of red pine
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS
E-4. EFFECTS ON AQUATIC CHEMISTRY
4.1 INTRODUCTION (J. N. Galloway)
In the last decade, the relationship between acidic deposition and
acidification of streams and lakes, with subsequent biological damage, has
been thoroughly reviewed and debated (NAS 1981, U.S./Canada 1983, NRCC 1981).
Despite this attention, confusion and uncertainties still exist, particularly
with regard to past, current, and future trends in the acidification of
aquatic systems, key processes that control acidification, the role of acidic
deposition relative to natural acid-generating processes, and the degree of
permanency of chemical and biological effects. The intent of this chapter is
to provide a critical review of available data and, to the extent possible,
an assessment of chemical responses of aquatic ecosystems to acidic
deposition.
The aquatic response to acidic deposition is largely controlled by processes
within the terrestrial ecosystem. Thus this chapter draws heavily on
discussions and conclusions from Chapter E-2 (Effects on Soil Systems). In
turn, it forms a basis for the assessment of impacts on aquatic biota in
Chapter E-5.
Chapter E-4 is arranged according to eight major topics:
0 basic concepts and definitions
0 characteristics of terrestrial and aquatic systems that determine the
sensitivity of surface waters to acidic deposition
0 locations of sensitive surface waters
0 the roles of sulfur (S) and nitrogen (N) in the acidification process
0 documentation of acidification and locations of lakes and streams already
acidified
0 evidence linking acidification to acidic deposition; alternative expla-
nations for acidification
0 predictive modeling of the chemical response to acidic deposition, and
0 interactions between acidification and metal and organic biogeochemical
cycles
4-1
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Much of the evidence for effects of acidic deposition on aquatic chemistry is
empirical. Discussions of cause-and-effect must rely largely on theoretical
considerations. The bulk of data results from studies in the northeastern
United States and eastern Canada. Thus, extrapolation of results to the
United States as a whole is difficult and introduces additional uncertain-
ties.
4.2 BASIC CONCEPTS REQUIRED TO UNDERSTAND THE EFFECTS OF ACIDIC DEPOSITION
ON AQUATIC SYSTEMS
The following concepts concerning effects of acidic deposition on aquatic
systems will serve as a foundation for critically assessing our current
knowledge.
4.2.1 Receiving Systems (J. N. Galloway)
Receiving systems are terrestrial, wetland, and aquatic. Their component
parts include:
a. Terrestrial Components
(1) forest, crop, or grass canopy
(2) litter layer
(3) organic soil layer
(4) inorganic soil layer
(5) bedrock
b. Wetland Components
(1) vegetation - mosses and other semi-submerged plants
(2) water - stream, pond, swamp
c. Aquatic Components
(1) stream
(2) lake
(3) sediment
These systems and their components are linked, so the effects of atmospheric
deposition on one component can cause secondary effects in another component.
The hydro!ogic pathway controls which components are affected by (or linked
to) other components. Water (precipitation) first hits the tree canopy, then
travels through successive layers of the terrestrial system before it enters
wetlands adjacent to the terrestrial system and then finally the lake.
Therefore, the effects of atmospheric deposition on any one component of the
terrestrial-wetland-aquatic system depend not only on the composition of the
atmospheric deposition but also on the effect of the atmospheric deposition
on every system 'upstream' from the component of interest. For example, the
effect of acidic deposition on aquatic systems depends on the quantity and
quality of atmospheric deposition and its effect on all components of the
terrestrial and wetland systems that it contacts prior to discharge into the
4-2
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aquatic system. A decrease in acidic deposition may not result in a decrease
in lake acidification until the terrestrial system above the lake recovers.
Instances where the terrestrial-wetland system is less important are (1) lake
systems with a large lake/watershed area ratio and (2) lake and stream
systems that receive runoff or snowmelt that has had little contact with the
terrestrial-wetland system.
The .composition of aquatic systems is controlled not only by physical and
chemical processes but also by biological processes. In discussing the
concept of system sensitivity and determining the degree of acidification, we
cannot ignore the biological component because, depending on location, type,
and productivity, the biological component can make waters more sensitive,
less sensitive, more acidified, or less acidified. Specific details on the
importance of the biological systems in mediating the chemical response of an
aquatic system to acidic deposition can be found in Section 4.3.2.6.
Additional details on terrestrial systems are found in the following sections
and in Chapters E-2 and E-3 on soils and vegetation, respectively. The next
chapter, Chapter E-5, discusses the effects of acidification of aquatic
systems on biota.
4.2.2. pH. Conductivity, and Alkalinity (M. R. Church)
Three analytical measures of importance in evaluating acidification of ground
or surface waters are pH, conductivity, and alkalinity. Definitions of these
three quantities are briefly given here. A later section (4.4.3.1.1) exam-
ines problems concerning the comparability of historical and more recent pH,
conductivity, and alkalinity data.
4.2.2.1 £H--In 1909 the Danish chemist, S. P. L. Srfrensen, introduced the
term pH when he used exponential arithmetic to express the concentration of
hydrogen ions in aqueous solution. He formulated his definitive equation as
CH = 10-p [4-1]
where CH was the hydrogen ion concentration and P was the hydrogen ion
exponent, which Stfrensen then wrote as PH and which we now write as pH
(Bates 1973). For a number of reasons, too detailed to explore here, pH as
originally defined by Srfrensen is not a measure of either hydrogen ion
activity or concentration (Feldman 1956, Bates 1973). By 1924, Stfrensen
and K. Linderstr(5m-Lang had realized that activity and not concentration
was the driving force for electromotive force (emf) changes in galvanic cells
(Feldman 1956, Bates 1973), so they defined a second term (pan)
paH = -log an = - log mHYH, [4-2]
where aH is the hydrogen ion activity, mu is the hydrogen ion molality,
and YH is the hydrogen ion activity coefficient. A theoretical problem
with this definition is that the activity of one ionic species by itself is
conceptually undefined, a problem not alleviated by the subsequent definition
= log mHY±, [4-3]
4-3
-------
where ^+ was chosen to represent the mean activity coefficient of a
dissolved (or of an average dissolved) uni-univalent electrolyte (Bates
1973). In an applied sense there obviously exists a need for a pH scale and
measurement system that can be used in day-to-day work by those not concerned
with strictly thermodynamic considerations (e.g., biologists, industrial
chemists, clinicians, water quality personnel). Such a practical pH (Feldman
1956) or operational pH (Bates 1973) is defined in standard fashion as
PH(x)=pH(s)
RT In 10
where pH(s) is the assigned pH of a standard solution, Es the electro-
motive force (emf) produced in a pH cell by the solution, F the Faraday
constant, R the universal gas constant, T the temperature in °K, and Ex the
potential produced in the pH cell by an unknown solution X, which then by
definition has a pH of pH(x).
Devising both a conceptually strict definition of pH and a pH scale
consistent with physical methods of measurement has proved exceedingly
difficult (Feldman 1956, Bates 1973). As Bates so succinctly put it:
The choice of a pH scale must take into account both the theoret-
ical and the experimental aspects. Unfortunately, no convenient
experimental method exists for the routine measurement of pH values
on the scales that are the most satisfactory in theory. Further-
more, the pH obtained by the convenient experimental techniques has
no simple exact meaning.
Fortunately, neither of these facts, in and of itself, adversely affects
estimations of acidification of surface waters by comparison of pH deter-
minations made over time. For purposes of such acidification estimations,
using the practical or operational pH scale defined above is sufficient.
4.2.2.2 Conduct!' v i ty—Conduc ti v i ty (or specific conductance) is the measure
of the ability of a solution to conduct an electric current. This capacity
is a function of the individual mobilities of the dissolved ions, the
concentrations of the ions, and the temperature of the solution. As the
"ohm" is the standard unit of resistance, the "mho" (ohm spelled backwards)
is the standard unit of conductance. Conductivity is conductance per unit
length of a substance of unit cross section and is usually reported as ymho
cm"* or the equivalent \i Siemens cm. Distilled water may have a con-
ductivity as low as 0.5 ymho cm"1, and some naturally-occurring surface
waters in the United States may have conductivities as high as 1500 ymhos
cm"r (Golterman 1969, American Public Health Association 1976, Skougstad et
al. 1979).
The rationale for measuring conductivity in relation to acidification of
surface waters is threefold. First, low conductivity values in surface
waters generally indicate a lack of buffering and thus susceptibility to
4-4
-------
acidification (Ontario Ministry of the Environment 1979). (In some cases,
however, organic compounds may contribute to buffering, but only very little
to conductivity. Such may also be the case with any nonelectrolyte "solid"
in suspension. Second, low conductivity has been correlated with sparsity of
fish populations in low pH lakes (Leivestad et al. 1976, Wright and Snekvik
1978). Third, increases in conductivity over time in surface waters are
sometimes associated with acidification (Nilssen 1980). Hydrogen ions have
extremely high mobilities in solution and contribute greatly to conductivity.
As a a body of water becomes acidified over time, increases in hydrogen ion
concentrations could lead to an appreciable increase in conductivity (e.g.,
from pH 5.0 to pH 4.5, an increase of approximately 7 ymho cm"1, using a
value of 0.313 ymho cm"1 per ueq £"1 free H+; see Wright and Snekvik 1978).
4.2.2.3 Alkalinity—Alkalinity or acid-neutralizing capacity (ANC) is
operationally defined as the equivalent sum of all of the bases that can be
titrated with strong acid to a preselected equivalence point or reference
proton level (Stumm and Morgan 1981). At least one author (NRCC 1981) has
sought to distingish alkalinity as that portion of ANC contributed by
dissolved carbonate species and hydroxide only. In their authoritative text
on aquatic chemistry, Stumm and Morgan (1981) use the terms alkalinity and
ANC interchangably, however, and that is the convention followed here.
In the very dilute surface waters often studied in relation to acidification,
total inorganic carbon concentrations are low; therefore, ANC due to the
carbonate system is also low. It is not unusual to find in these systems
that other species, such as naturally-formed weak organic acids (when
dissociated) and aluminum-hydroxy compounds leached from soils and sediments,
contribute measurably to ANC. For such waters an approximate expression for
ANC is
ANC = (HC03-) + (A10H2+) + 2 (A1(OH)2+)
+ 4 (A1(OH)4~) + (RCOO-) - (H+) [4-5]
where (RCOO") represents dissociated organic acids (Bisogni and Driscoll
1979). This expression neglects those dissolved and suspended protolytes
that commonly contribute very little to ANC. Also, this expression pertains
only to solutions isolated from important natural solid phases, such as lake
sediments.
In some select waters organic acids may dominate both the pH and buffering of
natural waters. Areas of North America that contain some waters of this type
include parts of the south and southeast, the upper midwest, locations in the
northeast, and the Atlantic maritime provinces of Nova Scotia and
Newfoundland. The relative abundance of such waters in the above areas is,
of course, quite variable. Naturally acidic brownwater lakes and streams are
discussed further in Section 5.2.1, Chapter E-5. For discussion of buffering
due to organic systems see Bisogni and Driscoll (1979), Wilson (1979), and
Section 4.6.3.2.
4-5
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The operational procedure for determining ANC is acidimetric titration with
strong acid to an appropriate end point. Methods for performing such
titrations and theoretical treatment of the pertinent equilibria have been
detailed in many publications (e.g., Golterman 1969, American Public Health
Association 1976, Loewenthal and Marais 1978, Skougstad et al. 1979, Stumrn
and Morgan 1981).
4.2.3 Acidification (J. N. Galloway)
Acidification is defined as the loss of alkalinity. Those aquatic systems
for which acidic deposition may cause acidification or loss of alkalinity to
levels that result in biological change are termed sensitive. Non-sensitive
systems may, therefore, experience a loss of alkalinity (i.e., acidification)
but are unlikely to experience any major biological effects. Note also that
use of the term acidification does not automatically imply acidification as a
result of acidic deposition.
Loss of alkalinity can be either chronic or acute, identified as long-term
acidification and short-term acidification, respectively. Short-term acidi-
fication refers to the development of strong acidity (i.e., alkalinity
< 0 yeq r1) during acid episodes (e.g., spring snowmelt) lasting for
periods of days or weeks. Because of the relatively short exposure periods,
biological effects occur only at those very low alkalinity levels {< 0 ueq
r1). Long-term acidification refers to the gradual loss of alkalinity
over periods of years or decades. As a result of chronic exposures, bio-
logical effects may occur at alkalinity < 50 yeq a~l (Chapter E-5,
Section 5.10.4), and waters with alkalinity < 200 yeq JT1 are generally
considered sensitive as defined in Section 4.3.2.6.1.
For the purposes of this chapter, acidic deposition refers to precipitation
with a pH below that attributable to natural processes. Galloway et al.
(1982a), based on measurements of precipitation chemistry in remote areas of
the world, noted:
The reference level commonly used is pH 5.6—the pH that results
from the equilibration of atmospheric C02 with precipitation....
Although pH 5.6 has been a useful reference level, it should not be
considered the pH of precipitation in all natural areas but only in
those areas that have no other acidic or basic precursors. In
reality, such areas are probably rare since small amounts of acids
or bases would either lower or raise the pH. What then is the
natural pH of precipitation? We believe that there is no single
natural pH of precipitation applicable to the whole globe.... The
lower limits of the natural mean pHs of precipitation in marine and
continental areas were pH >_ 5.
Carlson and Rodhe (1982) also concluded that the pH of natural precipitation
is highly variable, perhaps in the range of pH 4.5 to 5.6 (Chapter A-8,
Section 8.4.2). Thus, the definition of acid precipitation must also be
site-specific.
4-6
-------
In general, however, acidic deposition will refer to precipitation with a pH
< 5.0. Use of this arbitrary definition does not, however, preclude the
possibility that acidic deposition associated with precipitation at a pH
somewhat greater than 5.0 could result in acidification of surface waters
especially sensitive to atmospheric inputs of H+ and associated ions.
4.3 SENSITIVITY OF AQUATIC SYSTEMS TO ACIDIC DEPOSITION
The previous sections pertaining to aquatic systems have presented concepts
and definitions required to assess our knowledge of how aquatic systems are
affected by acidic deposition. This section and the ones following begin our
assessment by identifying important components in deposition processes and
receiving systems that will control the response of aquatic systems to acidic
deposition. Later sections will examine what is known about this response.
4.3.1 Atmospheric Inputs (J. N. Galloway)
Five factors must be considered when we assess the role of atmospheric
deposition in the acidification of aquatic and terrestrial ecosystems. These
are the components (total vs wet vs dry) of the deposition that are measured,
the chemical species in the deposition, the concentration of the substances
in the deposition relative to their loading (input rate), the location of the
deposition [considering a geographic scale as well as considering the
different components (e.g., leaf vs soil) of any system], and the temporal
distribution of the loadings.
4.3.1.1 Components of Deposition—To assess the impact of acidic deposition
we must know the total input (wet and dry). A major part of the current
North American effort regarding deposition monitoring is devoted to 'wet-
only' measurements. These data are inadequate for assessing impacts on
aquatic and terrestrial ecosystems; total deposition is underestimated not
only near major point sources of SOX, NOX (Dillon et al. 1982) but also
in remote areas (Galloway et al. 1982a). Relatively few attempts have been
made to measure dry deposition separately (Lindberg et al. 1982). In a few
cases (e.g., Dillon et al . 1982) 'calibrated1 lakes and watersheds have been
used to infer dry or total deposition of acidic substances. In other cases,
'bulk1 deposition measurements (made with a continuously open collector) have
been used. Although these collect an undefined portion of the dry depo-
sition, this information is more useful for chemical budget calculations than
'wet only1 measurements unaccompanied by dry deposition measurements. See
Chapter A-8 for further discussion of deposition monitoring.
In addition to H+ deposition, it is also important to measure the
atmospheric deposition of sulfur (as S042~ and 503), nitrogen (as
NOi- and Nfy"1"), and basic cations (see Section 4.4.1 and Chapter
A-8). Chemical and biological transformations of NOa" within the ter-
restrial or aquatic system (Section 4.3.2.6.2) may result in significant
internal production of ANC (NRCC 1981, Dillon et al. 1982). In some cases,
SO/^- is stored in terrestrial watersheds by the process of sulfate
adsorption (Chapter E-2, Section 2.2.8; Johnson and Cole 1980 Galloway et
al. 1983a), a process that may also generate ANC if the SO^- is reduced
or if strong acid is simultaneously stored. S042~ may also be reduced
4-7
-------
in lakes, resulting in production of ANC (Section 4.3.2.6.2; Cook 1981).
This production of ANC is only important on a long-term basis if it is net
production, i.e., a net reduction of N03~ and S042~ on an annual
basis. In other systems $042- apparently acts as a conservative sub-
stance. Within the limits of error in the measurement of the dry deposition
fluxes, the amount of S042~ leaving the watershed and entering in
deposition are approximately equal (Likens et al . 1977, Galloway et al .
1983c).
Once wet or dry deposited, S02 and SCty" have similar pathways through
the terrestrial and aquatic systems; therefore, the effect of S on aquatic
systems is not dependent on chemical speciation or type of deposition (MAS
1983).
Virtually all of the ammonium ion (Nty"1"), deposited on terrestrial and
aquatic systems is used chemically or biologically in those systems (Likens
et al . 1977, MAS 1981). Many of these reactions result in a decrease in ANC
(Chapter E-2, Section 2.2). Nfy"1" deposition is 'significant1 (25 percent
to 50 percent) relative to H+ deposition. For example, at Harp Lake,
Ontario, about 25 percent of the net input of acid was from NH4+ depo-
sition (Dillon et al . 1979). Therefore, measuring only free acid (H+) is
inadequate for assessing the impact of acidic deposition on systems.
The input rate of basic cations (e.g., Ca2+, Mg2+) is required for cal-
culation of the net loss of base cations from the watershed. In addition,
the effects of acid and acidifying ions (H+, S042", N0s~, and HN4+) are de-
pendent in part on the accompanying rates of deposition of neutralizing
cations (e.g., Ca2+) (NAS 1983).
4.3.1.2 Loading vs Concentration—Because the ANC of some components of
systems receiving acidic deposition is not renewed (other than over geologic
time), the total loading (or input rate) is the factor that determines how
long those components will be able to assimilate acidic deposition. The
ability of some other components to assimilate acidic deposition may depend
on concentration as well as total load of acids. The assimilation capacities
of components that have a continually renewed ANC (e.g., a lake epilimnion
that has ANC produced through primary production), or those where reaction
rates are controlled by hydrologic factors (e.g., reaction between acidic
deposition and silicate bedrock), are sensitive to the amount of water pas-
sing through components as well as to the concentration of acid.
In general, current measurements of acidic deposition include both concen-
trations of important substances and total loading rates of those substances,
with the exception of dry deposition as discussed in Section 4.3.1.1.
4.3.1.3 Location of the Deposition- -Wet deposition of acidic substances is
well measured in most areas of North America where the geological terrain has
a low capability to neutralize acids and where wet deposition is known to be
relatively high (> 20 meq strong acid nr2 yr'1; see Chapter A-8).
On a smaller scale, the relative magnitude of deposition on different
components (leaf, soil, water surfaces, etc.) of specific ecosystems is less
4-8
-------
understood. For example, the ability of the vegetation in a terrestrial
system, particularly the forest canopy, to modify deposition of acidic
substances has been demonstrated (Parker et al. 1980) but needs to be
quantified in further studies.
Other factors, such as the relative deposition to the terrestrial component
of a watershed vs directly onto the surface water, are also important. These
factors determine the relative importance of the pathways that the deposited
substances follow, which in turn controls the overall assimilation capacity
of the system.
4.3.1.4 Temporal Distribution of Deposition—To assess their impact on
receiving systems, the input rates of acids or acidifying substances must be
considered on a seasonal and a short-term (i.e., episodic) basis as well as
on a long-term (annual) basis.
Seasonal inputs are particularly important in areas where snowpack formation
occurs, with the subsequent release of a major portion of the annual depo-
sition during snowmelt (Jeffries et al. 1979, Galloway et al. 1980b). In
some cases (e.g., central Ontario; Jeffries et al. 1979), during snowmmelt
the ground may be partially frozen. As a result, the release of ions occurs
at a time when the terrestrial system cannot assimilate the ions as effi-
ciently as it can at other times.
Short-term variations in deposition, on even an episodic basis, may be
important in some instances. Flow paths may be altered on a short-term
basis, resulting in shortened reaction times and less assimilation of the
acidic deposition.
The seasonal variation in deposition has been frequently investigated; short-
term variations are less poorly studied and need further quantification.
4.3.1.5 Importance of Atmospheric Inputs to Aquatic Systerns--
4.3.1.5.1 Nitrogen (N), phosphorus (P) and carbon (C). Only recently have
researchers appreciated the importance of precipitation inputs of various
cations and anions, especially N and P, to the nutrient balance of inland
freshwaters (e.g., Gorham 1958, 1961; Vollenweider 1968; Schindler and
Nighswander 1970; Likens 1974; Likens and Borman 1974). Concentrations of
inorganic and organic N and P in rain and snow may be small, but the total
input by storm, by season, or by year may be a significant source of these
nutrients for aquatic organisms, particularly in nutrient-poor lakes (Likens
et al . 1974). Direct inputs of nutrients in precipitation to lakes are
particularly important in areas with granitic geologic substrates, especially
if the ratio of lake surface area to terrestrial drainage area is large
(Likens and Bormann 1974). In addition, the gaseous exchanges of nitrogenous
compounds in many lakes may be important but are poorly understood (Likens
1974).
Based on relatively few data, some 50 percent of the P and 56 percent of the
dissolved N for oligotrophic lakes may come from direct precipitation (Likens
at al. 1974). With human influences in the watershed (urbanization,
4-9
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agriculture, etc.) runoff inputs to aquatic ecosystems increase and direct
precipitation inputs become much less important to the total budget, even
though the absolute amount provided by precipitation remains the same. Where
terrestrial inputs of N and P dominate, lakes are usually much more
biologically productive, if not eutrophic (Likens et al. 1974).
Preliminary data suggest that organic carbon inputs in precipitation may be
ecologically significant for some aquatic ecosystems, particularly oligo-
trophic lakes. Mean concentrations averaged about 6 mg C &"1 in pre-
cipitation and accounted for 28 percent of the total allochthonous inputs of
organic carbon for a small oligotrophic lake in New Hampshire (Jordan and
Likens 1975). Data are insufficient, however, to extrapolate concerning the
importance of atmospheric inputs of organic carbon to oligotrophic lakes in
general .
4.3.1.5.2 Sulfur. Two sources provide sulfur for surface waters: rock
weathering and atmospheric deposition. In the absence of reactive sulfur
sources in bedrock, atmospheric deposition is the primary source (Cleaves et
al. 1970, Wright 1983). This is especially true in areas without significant
sources of reactive sulfur in the watershed and receiving acidic deposition,
where atmospheric sulfate becomes the dominant anion in low alkalinity waters
(Gjessing et al . 1976, Oden 1976a, Henriksen 1979, Wright et al . 1980,
Galloway et al . 1983c, Wright 1983). This dependence is illustrated by
plotting the mean and range of excess S0^~ (over and above that supplied
by sea salt cycling) export from watersheds across North America on a line
that transects the region of large atmospheric deposition of $04^'
(Figure 4-1). The wet deposition of excess SO*2" at each location is
shown in the same figure, with estimated total SO^' deposition shown at
four locations. There is a clear positive relationship between excess
$0^2- deposition and S042~ in the runoff although S04^~ export
exceeds deposition in the areas of highest deposition. This deficiency of
sulfate measured in precipitation as compared to sulfate export from water-
sheds may, at least in part, be due to dry deposition of S04Z~ and $03 •
The dry deposition would be greater in regions nearer to or downwind from
industrial sources (U.S./Canada 1983).
The dependence of surface water values of S042~ on atmospheric deposition of
S042~ is also denoted by the significant (p < 0.001) correlation between
S042" concentrations in surface waters and $04*' concentrations in precip-
itation over a wide range of concentrations, illustrated in Figure 4-2.
Areas of North America receiving precipitation with high concentrations of
S042" (southeastern Canada, northeastern United States) have higher
S042' concentrations in lakes, while areas receiving precipitation with
low values of S04Z~, have surface waters with low concentrations of
S042' (Rocky Mountains, Colorado, Labrador, northern Quebec). Using the
latter areas as baseline for North America, the estimated background
S042' concentration in North American lakes is 20 to 40 ueq fc'1. In con-
trast, lakes in eastern North America receiving acidic deposition have $04^-
values of 100 to 167 peq JT1, suggesting that about 80 to 120 yeq S042~ a~l
(average of 100 yeq £-I) is derived from anthropogenic atmospheric deposition.
This applies for a relatively large region of eastern North America,
4-10
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s
o
o
CM
O
O
LABRADOR
ISLAND OF
NEWFOUNDLAND
HALIFAX
NEW BRUNSWICK
LAFLAMME
MAURICIE
ADIRONDACK
N. OF OTTAWA
ALGONQUIN
HALIBURTON
SUDBURY
ALGOMA
THUNDER BAY
QUETICO
ELA
,
baiu
Figure 4-1. Mean and range of basin specific yield of excess sulfate
) compared with atmospheric excess sulfate deposition
) in precipitation for 1980 (Thompson and Mutton
1981, 1982) and the range of estimated wet deposition for
1977-80 from the CANSAP precipitation network (Barrie and
Sirois 1982). Also shown are the ranges of wet plus dry
deposition of sulfate (| — |) calculated from the 1980
measurements of SOX in the atmosphere at 4 Canadian Acid
Precipitation Network Stations (Barrie 1982). Adapted
from U.S./Canada (1982).
4-11
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180
160
140
120
~ 100
I
CM
80
o" 60
40
20
E. Ontario
Connecticut
Adirondack*
Maine
- Florida •
Laurent1an
Mts.
Nova Scotia
New Hampshire
- W. Ontario
.Labrador
Kereke
• Newfoundland
* Quebec
• Rocky Mts.
> Labrador
20
40
60
80
100
SO.2" PRECIPITATION (ueq i1)
Y « 1.92X + 14.08 R « 0.86 P <. 0.001
Figure 4-2. Mean concentration of 864* (excess S042-, over and above
that supplied by sea salt cycling) for 15 lake groups in
North America and mean $04* in wet deposition at nearby
deposition monitoring stations. Adapted from Wright (1983)
4-12
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with some areas relatively far away from S sources. Those waterbodies in
areas closer to S emission sources will have larger increases in Sfty2'
concentrations. For example, lakes near Sudbury, Ontario have 400 peq
jr1 of S042~ from atmospheric deposition while lakes east of the Rhine-Ruhr
industrial region of Germany can have > 1000 yeq r1 of S042' from atmos-
pheric deposition (Schoen et al. 1984).
o
The influence of the increased $04 deposition on aquatic, chemistry is
large, for on an equivalent basis, the increase in $04 " in surface
waters has to be matched by an increase in a cation, either protolytic
(proton-donating; e.g., H+, Aln+) or non-protolytic (e.g., Ca2+,
Mg2+, etc.) (Galloway et al. 1983a). An increase in the former will result
in loss of alkalinity (acidification) of the waterbody. An increase in the
latter will result in a loss of basic cations from the terrestrial system.
Both effects can potentially alter biological communities in the respective
ecosystems and are discussed in greater detail in Section 4.4.3.
4.3.2 Characteristics of Receiving Systems Relative to Being Able To
As s i mi Tate~ Acidic Deposition (P. J. Dillon and J. N. Galloway)
The anthropogenic acids transported via the atmosphere may be deposited
directly onto aquatic systems (lakes, streams, wetlands) or onto terrestrial
systems that drain into the aquatic systems. Each of the components or
subsystems of these systems may be capable of assimilating some or all of the
acidic deposition received. This section discusses the factors that
determine the quantitative capability of the subsystems to assimilate acidic
deposition.
4.3.2.1 Canopy--Throughfall and stemflow have elevated levels of most
elements relative to incident rainfall (Miller and Miller 1980) and even, in
at least one report, relative to snowfall (Fahey 1979). The changes in
chemical content result from washdown of particles filtered from the
atmosphere by the vegetation, and from leaching of the vegetation (the crown
in the case of throughfall, the bark as well in the case of stemflow). The
process of particle washdown is, of course, completely independent of any
ability of the canopy to assimilate acidic deposition. On the other hand,
leaching of cations from the canopy may represent a significant assimilation
capacity. However, the relative importance of each process is generally
unknown (see Chapter E-3, Section 3.2.1.2). Although there are conflicting
reports, some generalizations may be made.
Stemflow often has a lower pH than does incident precipitation, either
because of leaching of organic acids or washdown of acidic aerosols (Miller
and Miller 1980).
Throughfall in deciduous forests has usually been found to have elevated pH
and increased cation (Ca2+, Mg2+) concentration (Likens et al. 1977, Cole
and Johnson 1977). The relative importance of washdown of filtered particles
and of cation exchange with the leaf is unknown. Direct uptake of S02
(Fowler 1980) and ammonium (Miller and Miller 1980) also may contribute to
the acidity of the throughfall. The pH of throughfall in coniferous forests
4-13
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has been reported to be decreased relative to pH of precipitation (Horntvedt
and Joranger 1976), although the basic cation content is increased.
The amount of throughfall or stemflow is, however, less than incident
precipitation (Ford and Deans 1978, Miller and Miller 1980). Therefore, an
increase in concentration of substances in throughfall relative to
precipitation does not necessarily indicate that the canopy has supplied
materials as a result of either washdown or leaching. The loading of each
substance beneath the canopy must be compared to that above the canopy before
the occurrence of either process can be ascertained.
4.3.2.2 Soil—The surficial material accumulated on the bedrock of North
America is extremely complex in both physical and chemical properties. This
surficial material assimilates acidic deposition through dissolution, cation
exchange, sulfate adsorption, and biological processes. Further detail on
these processes in soils is provided in Chapter E-2, and the effects of soils
on the chemistry of aquatic ecosystems is discussed in Chapter E-2, Section
2.6. Major concepts are summarized below.
In general, surficial materials containing carbonate minerals have abundant
exchangeable bases and can assimilate acidic deposition to an almost unlim-
ited extent. Regions of North America with soils formed in situ on lime-
stone, dolomite, or marble provide adequate neutralizing capacity under all
loading conditions. Soils formed in situ on carbonate-cemented, carbonate-
interbedded, or carbonate clastic sedimentary rocks may have reduced assim-
ilation capacity under very high acidic deposition conditions, but effects of
acidic deposition on streams and lakes are probably minimal. As a result of
the transport of surficial material in the glaciated areas, it is possible to
find carbonate-containing deposits on non-carbonate bedrock.
The ability of surficial materials that contain no carbonate minerals to
assimilate acidic deposition results from cation exchange reactions, sili-
cate-mineral dissolution reactions and, in some cases, Fe and Al oxide
dissolution. The result of these reactions is an increase in the concen-
trations of major cations (particularly Ca2+, Mg2+, and possibly Na+,
and K+), and Al and Fe in the runoff water leaving the watersheds. This
ability is affected by:
1) the chemical nature of the surficial material, in particular the cation
exchange capacity (CEC) and the base saturation (BS),
2) the permeability of each layer of the soil,
3) the surface area (or grain size) of the soil particles, and
4) the amount (depth and/or mass) of soil in the watershed.
The most important of these factors are the CEC (the total amount of cations
that can be exchanged for H+; Table 4-1) and the BS (the proportion of the
total exchangeable cations that consists of Ca2+, Mg2+, Na+, and K+)
(see Chapter E-2, Section 2.2.2). The organic layer of the soil has a high
4-14
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TABLE 4-1. TYPICAL CATION EXCHANGE CAPACITIES OF SOIL COMPONENTS
(FROM MCFEE ET AL. 1976)
SOIL COMPONENTS CECa
(meq per TOO g)
Organic matter (humus) 200
Silicate clays
vermiculite 150
montmorillorite 100
kaolinite 10
illite 30
Hydrous oxide clays 4
Silts and sands negligible
Variation is commonly 40% of these mean values.
4-15
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CEC (McFee et al. 1976). Fresh organic litter has a substantial BS
component. Soils with high BS have a greater potential to assimilate acidic
deposition, all other factors being equal, than soils with low BS.
The permeability of the soil layers is also important because it determines
the contact time of the percolating water with the soil pajjticles (Chapter
E-2, Section 2.1.3.1). Loosely-packed organic material in/the upper layer is
usually highly permeable and so may provide little assimilation capacity,
especially in cases of high input of water. As the surface area of the soil
particles in the organic layer increases, the permeability of the layer
decreases; both factors increase the H+ assimilation capacity of the soil,
whether it is a result of surface cation exchange reactions or silicate or
metal oxide dissolution reactions. However, the proportion of the soil
consisting of very small particles (i.e., clays) may increase to the point
where permeability of a specific layer is decreased very significantly. In
some cases, impermeable layers may effectively eliminate the potential for
assimilation of acidic deposition by deeper soil layers.
The depth of the surficial material in a watershed is, of course, also very
important. Areas with extremely shallow (1 m) till often have only an
organic layer and a well-weathered layer (horizon) that may have little
assimilation capacity left (i.e, have low BS). Areas with deep tills (e.g.,
till plains, kames, moraines, eskers, spillways, outwash, and alluvial
formations) will almost always have high capacity for assimilating acidic
deposition because of their moderate to high BS at greater depth, combined
with their large amounts of unweathered material.
Another soil process important in controlling the response of aquatic systems
to acidic deposition is sulfate adsorption. Soils with large sulfate
adsorption capacities will essentially act as sinks for the atmospheric
sulfur, preventing it from reaching the aquatic system. As noted in Chapter
E-2, Section 2.2.8, sulfate adsorption capacity of soils is not routinely
determined; therefore, the extent of soils with significant capacity to
adsorb sulfate has not been established. Some adsorption capacity is a
common property of many Ultisols, Oxisols, some Alfisols, and is reported for
other soils (Singh et al. 1980). The work of Johnson and Todd (1983) shows
sulfate adsorption is low in Spodosols. The distribution of these soil
orders within the U.S. is depicted in Figure 2-4 (Chapter E-2). Spodosols
are common in the glaciated regions of the northeastern United States and
upper Midwest, and in much of Florida. Ultisols are prevalent in much of the
southeastern United States.
4.3.2.3 Bedrock—The ability of bedrock to neutralize acidic deposition is
control 1ed by:
1) chemical composition of the bedrock,
2) effective reaction surface area, and
3) retention time or contact time of water with the bedrock.
4-16
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Carbonate minerals in the bedrock result in rapid assimilation of the strong
acids by dissolution and in production of bicarbonate ion. Bedrock types
containing no carbonate minerals may neutralize acidic deposition by the
dissolution of silicate minerals, which is an extremely slow process relative
to carbonate dissolution.
Massive, impermeable bedrock's effective surface area for chemical reaction
is minimal. Acidic deposition contacts only the upper surface layer, so the
slow dissolution process will modify water chemistry only marginally,
regardless of which silicate material is involved. Bedrock exhibiting only
jointing or fracturing will provide relatively greater surface area for
reaction, but complete assimilation will only occur at considerable depth,
probably affecting the chemistry of the groundwater pool but having little
effect on stream and lake chemistry. The maximum extent of surface reactions
will be attained by silicate bedrock having a porous nature, e.g., weakly
cemented sandstone.
Slower movement of acidic waters through silicate bedrock will result in
greater assimilation. Massive igneous beds will shed water with only a short
contact time, while more permeable sandstone beds will increase contact time.
Table 4-2 summarizes the assimilation capacity of various bedrock types. The
ratings are qualitative only and are meant to reflect 'characteristic1 values
for each bedrock type. Surficial geology, including glacial deposits, soils,
and unconsolidated material, has a greater influence on a system's ability to
assimilate acidic deposition. Bedrock influence on surface water chemistry
is mainly indirect through derived unconsolidated material.
4.3.2.4 Hydro!ogy (G. B. Blank, P. J. Dillon, J. D. Gregory) —
4.3.2.4.1 Flow paths. The extent to which strong acid components of
deposition react with each component of the substrate (i.e., bedrock, soil)
depends in most cases on the time of contact with that substrate; thus the
flow path of water is important in determining the total assimilating
capacity of the terrestrial system. Time of contact is important because
only surface reactions (adsorption, ion exchange) occur rapidly for
aluminosilicate minerals; slow diffusion processes control subsequent
reaction rates. Reaction rates with carbonate (bedrock, or in soil) are
rapid; therefore, these areas are not sensitive to acidic deposition.
Because the groundwater pool often has a slow turnover rate (i.e., contact
time is long), assimilation of H+ is expected.
A generalized depiction of the flow of water and associated materials through
a terrestrial ecosystem (eventually discharging into a lake or stream) is
shown on Figure 2-1 and discussed in Section 2.1.4, Chapter E-2. Additional
details are presented here because of the importance of these hydrologic
processes in determining chemical changes (both short-term and long-term) in
aquatic systems in response to acidic deposition (cf. Section 2.6, Chapter
E-2).
Upon striking the land surface, water may either infiltrate the soil or move
laterally as surface (overland) flow. In temperate climates, about 75
4-17
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TABLE 4-2. APPROXIMATE BUFFERING CAPACITY OF VARIOUS BEDROCK TYPES
(ADAPTED FROM HENDREY ET AL. 1980b)
Buffering capacity
Bedrock type
Low to none
Granite/Syenite or metamorphic
equivalent
Granitic gneisses
Quartz sandstones or metamorphic
equivalent
Medium to Low
Sandstones, shales, conglomerates or
their metamorphic equivalents (no
free carbonate phases)
High-grade metamorphic felsic to
intermediate volcanic rocks
Intermediate igneous rocks
Calc-silicate gneisses with no free
carbonate phases
Medium to high
Slightly calcareous rocks
Low-grade intermediate to mafic
volcanic rocks
Ultramafic rocks
Glassy volcanic rocks
'Infinite1
Highly fossiliferous sediments or
metamorphic equivalents
Limestones or dolostones
4-18
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percent of all precipitation enters the soil to become soil moisture or
groundwater (Hewlett 1982). At any location this amount varies, of course,
depending on precipitation intensity and the type of surface the
precipitation contacts. Bare rock outcrops, for instance, shed water to
nearby soils and aquatic systems almost immediately.
The following factors have been shown to influence infiltration rates:
0 organic matter and biologic activity,
0 soil texture and structure,
0 slope gradient,
0 type of colloids in the soil,
0 whether the soil is frozen,
0 presence of hygroscopic or hydrophobic layers,
0 season of the year, and
0 vegetative cover.
The type of forest floor can also alter the rate at which water may move into
the mineral soil. The infiltration rate under hardwoods is generally higher
than under conifers on the same soils because of the greater activity of soil
fauna in hardwood litter (Armson 1977). In addition, to the degree that
forest floors are disturbed by cultivation, grazing, repeated burning,
logging, and road building, infiltration may be hindered so that overland
flow occurs.
Factors controlling infiltration also govern percolation rates, or soil water
movement and distribution during and after the infiltration process. Soil
texture and structure affect the distribution of pore space, which in turn
affects infiltration, detention storage (gravitational water moving through
the soil profile) vs retention storage (water held in capillary pores and
surface films against the force of gravity), and water movement (Hewlett
1982).
In uncultivated areas, large channels are often established in the soil
system as a result of burrowing animals and decomposition of tree roots.
These channels are frequently open to the surface and provide open conduits
for flow of drainage water (Section 2.1.4, Chapter E-2). Such direct inflow
to deeper soil layers and bedrock or directly to aquatic systems, lessens
soil-water contact time.
Hursch and Hoover (1941) noted that "the annual decay of some roots each
year, and their subsequent channeling by microorganisms and small insects
create relatively large continuous openings that serve as hydraulic pathways
for the rapid movement of water." Weaver and Kramer (1932) traced one 1.3 cm
4-19
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diameter channel at the 3 m level for a distance of 1.2 m, and noted that
many smaller channels were found. They asserted that large root channels
seem to remain in place for a long time. Gaiser (1952) found > 4000 cavities
per acre on one site in southeastern Ohio. His study showed channels
penetrating as deep as 0.8 m, with channel diameters at deepest penetration
ranging from 2 to 30 cm. Depths of channels are limited by the nature of the
soil and parent material through which roots have grown.
At any one time, for a given soil system and terrain, movement of water from
an individual storm through the watershed is largely controlled by the degree
of soil saturation. Saturation levels are determined by numerous factors
such as length of time since the last storm event, drainage character of the
soils and underlying material, cover vegetation type and evapotranspiration
potential, slope of the terrain, and land use. Variations in soil saturation
through time, and the associated variations in water flow path, result in
temporal variations in the quality of water discharged to the receiving
system (e.g., stream). Factors that determine shifts in saturation levels
thus influence the susceptibility of the aquatic system to short-term
acidification.
Hydro!ogists identify two primary components of streamflow: baseflow and
stormflow (or quickflow). Baseflow is continuous flow between storm events
and includes slow drainage of soil water directly from the vadose zone (the
unsaturated zone above the water table—also called the zone of aeration; see
Figure 2-1, Chapter E-2) and slow drainage of groundwater from the saturated
zone below the water table (the result of deep percolation from the vadose
zone). Because of its extended period of interaction with soil before
discharge, baseflow has relatively high alkalinity and pH levels. Stormflow
is the high flow associated with a storm event and comprises channel
precipitation, overland flow, and interflow (Ward 1975). Of these three,
interflow (rapid subsurface lateral flow to a channel) is the most important
in raising stormflow discharge above baseflow rates. Subsurface stormflow
includes the following:
0 flow through large connected macropores in unsaturated conditions,
0 rapid saturated flow through the forest floor or coarse-textured
soil layers,
0 lateral flow above slowly permeable zones,
0 piping through channels made by decayed roots or by burrowing
animals, and
0 in flat terrain with a high water table, lateral flow resulting from
a rise in the water table.
Subsurface stormflow is mainly contributed to surface water by the saturated
zone, termed the source area, adjacent to the channel or lake. Hewlett and
Hibbert's (1967) variable source area model has been widely accepted to
define the relationship between precipitation and stormflow (Ward 1975).
4-20
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According to this concept, as rainfall proceeds during a storm, the saturated
zone (source area) expands due to infiltration and lateral flow through the
soil from upslope; it then contracts as rainfall ends. Expansion of the
source area creates ephemeral stream channels feeding the perennial channel
from higher upslope. In upland areas with good infiltration, runoff does not
come from all areas of a watershed equally, and peak contribution areas may
change with time. Contact time of interflow with the soil is much less than
for baseflow, so there is less opportunity for neutralization of acidity.
. Flow rates through the soil are much higher and discharge to ephemeral
channels decreases average distance of flow through the soil. Flow paths for
interflow are larger than for baseflow, so the area of water-soil interface
per unit of volume also decreases. Runoff to ephemeral channels may be
particularly rapid from thin, rocky mineral soils (with or without deep humus
layers) high up in a watershed. Thus, precipitation may be released to
streams and lakes without having passed through the deeper mineral soils
downslope where neutralization reactions can occur.
The following factors affect the size of the source area and rate of drainage
to channels:
0 hydrologic depth (soil volume for storage),
0 antecedent water content (soil moisture conditions),
0 soil hydraulic conductivity (infiltration and percolation rates),
and
0 rainfall intensity and duration (total quantity of falling water).
According to Harr (1977), steep slopes and highly permeable surface soils are
conducive to rapid, shallow subsurface flow, which would account for quick
response of streams to storms.
Overland flow derives from water failing to infiltrate the surface and
instead running to the nearest stream channel. Hewlett and Hibbert's (1967)
calculations indicate about 2.7 percent of stormflow in forested watersheds
comes from overland flow, with about 1.0 percent contributed by channel
precipitation (water falling directly in the stream channel or lake). Harr
(1977) notes that overland flow rarely occurs in forested watersheds in humid
regions. What is commonly believed to be overland flow from a storm is often
rapid interflow (also called translatory flow) displaced from soil storage by
new rainfall and infiltration farther upslope. The overland flow component
during snowmelt runoff, on the other hand, has been measured to be as high as
100 percent and as low as 0 percent (Colbeck 1981). During snowmelt the
overland flow component often travels through the bottom layer of the snow.
Minerals dissolved in precipitation remain in the watershed and, if not taken
up by vegetation or otherwise absorbed, follow the natural flow paths
downslope toward a stream channel or lake. Krug and Frink (1983) refer to
this downward migration but also note the varied disposition of acidity in
soil layers. They maintain that thinner soils farther upslope produce thick
4-21
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humus layers likely to be much more acid than thicker soils downslope. In
the event of heavy rain or rapid snowmelt, a greater proportion of the
streamflow will have been in contact with the most acid soil layers higher up
in the watershed. Krug and Frink (1983) also note the disproportionately
large effect peats bordering lakes and streams can exert on water chemistry,
a view supported by the variable source area concept (Hewlett and Hibbert
1967, Hewlett 1982).
In areas with snowpacks, contact time is reduced during snowmelt because of
the quick saturation of the soils by the first stages of melting. In areas
where the soil freezes, contact time is even further reduced. In both cases,
the impact of snowmelt on runoff (and therefore on stream and lake) chemistry
is great (Jeffries et al. 1979, Johannessen et al. 1980, Overrein et al.
1980). In some areas of central Ontario, the upper 1.0 to 1.5 m of the soil
is generally frozen each winter (Jeffries, D., Ontario Ministry of the
Environment, Rexdale, Ontario, personal communication 1981), so spring runoff
may flow principally over the soil layer or through only the top few cm. In
other areas (e.g., Adirondacks, White Mountains in New Hampshire), surface
soil layers freeze only when little snowpack develops during winter.
Obviously, factors that control the movement of water through the terrestrial
system to the aquatic system are extremely complex. While it is possible to
generalize concerning watershed characteristics that influence aquatic
sensitivity to long-term and short-term acidification, it is difficult to
impossible to utilize these criteria in assessing the geographic extent of
sensitive waters. In addition, in some cases, our understanding of critical
concepts of hydrology, for example the importance of macropore and
channelized flow, is insufficient. Each watershed exhibits unique
characteristics, and these characteristics translate to unique impacts on
water quality and quantity available to aquatic systems.
4.3.2.4.2 Residence times. It is often assumed that headwater lakes are
more sensitive to acidic deposition than are other lakes (Gjessing et al.
1976, Minns 1981). This assumption may arise, in part, because headwater
lakes
a) often have longer hydrologic residence times than lakes downstream,
simply because their total catchment area-lake area ratio is smaller
(hydrologic residence time is a function of lake volume rather than
lake area so lake morphometry must also be considered);
b) often are at higher elevations (on a regional basis) and therefore
have few or no soil deposits in their watersheds; and
c) often have poorly developed soils in their watersheds.
Lakes with smaller catchment area-lake area ratios will usually receive a
greater porportion of their total input of water via deposition directly on
the lake surface. The acids in the deposition on the lake surface have not
been assimilated by any other system. On the other hand, even in systems
with small watersheds, assimilations of hydrogen ion in the terrestrial
systems can be > 50 percent of the total deposition on an annual basis
4-22
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(Galloway et al. 1980a, Wright and Johnannessen 1980, Jeffries et al. 1981).
As the catchment area-lake area ratio increases, the ability of the overall
watershed (terrestrial catchment + lake) to assimilate the acidic deposition
falling on it increases.
A long hydrologic residence time is favorable (i.e., makes a lake less sen-
sitive) if a major portion of the ANC that enters the lake results from
internal processes. If water renewal rate is slow, the ANC provided by
processes such as primary production will build up from year-to-year rather
than be lost from the lake via outflow.
In summary, the relative importance of the ANC supplied by internal processes
in a lake vs the acid assimilation capability of the terrestrial watershed
will determine, for a particular lake, whether a long hydrologic residence
time is beneficial or detrimental.
4.3.2.5 Wetlands--Very little is known about the role of wetlands in
assimilating acidic deposition. In addition to neutralization by alkalinity
present in the aqueous component of the wetland, other processes may
contribute to assimilation, including 1) reduction reactions and 2) ion
exchange reactions.
Reduction reactions (e.g., N03~ reduction, S042' reduction, Fe3+ reduction)
occur in the aqueous portion of the wetland under anaerobic conditions, e.g.,
under ice-cover during the winter. They may also occur in the sediments,
which are typically high in organic content and are anaerobic at all times.
The ANC produced by these reduction reactions may, however, be temporary
(Section 4.3.2.6.2) if the reactions are reversed when the water is oxic, or
if the water is removed (e.g., by evaporation) exposing the sediments to the
atmosphere. Some of the ANC produced is permanent if, for example, sulfide
produced from SO^- reduction is stored as FeS. In other cases, oxygen
demand in the wetland may be high enough at all times to keep the aqueous
component anoxic. The reduction processes may, in these cases, produce
permanent ANC.
Cation exchange reactions with the sediments or detrital material in the
wetland may result in significant assimilation of strong acid if the BS is
appreciable. However, this is probably seldom the case. In fact, some
wetlands, particularly Sphagnum bogs, nave been shown to produce mineral
acidity (Clymo 1963) by means of cation exchange reactions.
Hemond (1980) examined sources of acidity and alkalinity in a small bog
system in central New England. The process of ion exchange increased the
mineral acidity of water in the bog, but only to a modest degree when
compared with other influences. Inputs of H+ from atmospheric deposition
were by far the largest contributor to mineral acidity. The influence of
acidic deposition, however, was largely (> 90 percent) counteracted by
biological processes within the bog, specifically reduction of S04^~ and
the biological uptake of N03~. The resulting mineral acidity of the
bog water was quite low ( -0.05 meq &~1). By far the dominant in-
fluence on the acidity of the bog was the presence of weak organic acids at
concentrations of ~ 1 meq £-1.
4-23
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It cannot be inferred that all wetland systems respond in similar fashion.
Gorham et al. (1984) emphasize the large number of unknowns concerning the
biogeochemistry of wetlands, and note that some wetlands may be among the
ecosystems most vulnerable to acidic deposition.
Waters in wetlands often naturally have low pH. Thus, in addition to their
potential role as assimilators of acid deposition, wetlands are of interest
as potential contributors to acidity and acidification of surface waters.
This topic is considered within Section 4.4.3.3, Alternative Explanations For
Acidification.
4.3.2.6 Aquatic—The ability of aquatic systems to assimilate atmospheric
deposition is dependent on several factors, among them, the amount, timing,
and rate of acidic deposition, and the hydrologic flowpath and the rate of
alkalinity generation in the watershed and aquatic system. To understand the
effects of these processes on any given aquatic system's response to acidic
deposition would require a process oriented model. However, to determine, on
a regional scale, the ability of aquatic systems to assimilate acidic depo-
sition, a simpler indicator is required. The following section discusses the
past use of alkalinity as such an indicator, presents an analysis of the use
of 200 yeq r1 alkalinity as the boundary between sensitive and non-
sensitive systems for long-term and short-term acidification, and presents an
assessment of the validity of this threshold.
4.3.2.6.1 ATJcalinity as an indicator of sensitivity. Threshold alkalin-
ities, below which an aquatic system receiving acidic deposition would have
the potential for becoming acidic to a point where biological effects might
occur, have been estimated. Thresholds should be such that both long-term
and short-term acidification effects are considered. The following material
provides past estimates and support for a qualitative estimate of such a
threshold.
In the past, subjective criteria have been established to 'classify' lakes;
e.g., lakes in Ontario were classified as having extreme sensitivity if 0 to
40 yeq xr1 alkalinity was measured, moderate sensitivity if 40 to 200
yeq &-1 alkalinity was measured, etc. (Anon 1981). Altshuller and
MacBean (1980) classified lakes as 'susceptible1 if alkalinity was measured
as < 200 yeq £-1. Calcite saturation index (CSI)--a measure of the
degree of saturation of water with respect to CaCOs (calcite) that inte-
grates alkalinity, pH, and Ca concentration—has also been used (NRCC 1981).
In another case (Minns 1981), simple assessment of lake sensitivity has been
based on ionic strength (conductivity), with the unstated assumption that
ionic strength must be a good correlate of alkalinity.
The boundary between 'sensitive1 and 'insensitive' that is often used is 200
yeq JT1 of alkalinity before the onset of acidification (Hendrey et al.
1980b). The justification for this value is as follows. Acidified aquatic
ecosystems have been defined as those that have lost alkalinity. Sensitive
acidified aquatic ecosystems have been defined as systems where alkalinity
reductions have resulted in biological changes. Biological effects due to
acidification become apparent as pH declines to near 6.0 (Chapter E-5,
Section 5.10.4). However, to relate alkalinity changes to biological
4-24
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effects, it is necessary to first relate pH to alkalinity. To do this, data
from 928 streams and lakes in New York State have been compiled (Figure 4-3;
Hendrey 1982). These data are for New York State only; a similar compilation
for 1936 sites in New York, Pennsylvania, North Carolina and New England
shows a nearly identical relationship between pH and alkalinity (Hendrey
1982). Data for 201 lakes in New England (Haines and Akielaszek 1983) are
plotted in Figure 4-4. These data show that pH 6.0 corresponds to an
alkalinity of approximately 40 yeq r1 (range 10 to 90 yeq £-!).
Therefore aquatic systems that are acidified to an alkalinity of 40 yeq
£-1 or below have a good chance of experiencing biological effects. It
should be noted, however, that for this relationship to be applied to other
areas of North America, additional data compilations may need to be per-
formed.
To determine the threshold between sensitive aquatic systems and nonsensitive
ones it is necessary to add to 40 yeq r1 of alkalinity, the amount of
alkalinity loss that an aquatic system would experience from acidic depo-
sition. On a regional basis in the northeastern United States, the maximum
increase of SO^- due to acidic deposition in aquatic systems is ~ 100
yeq £-! (Harvey et al. 1981, Wright 1983; Section 4.3.1.5.2 and Section
4.4.3). Therefore, the maximum alkalinity decrease that could have occurred
over time is 100 yeq £~1 (although in areas of eastern North America
e.g., West Virginia, Pennsylvania, that are closer to S emissions sources
than the northeastern United States, the maximum increase of S04~2 due to
acidic deposition can be much greater than 100 yeq £-1, see explana-
tion, Section 4.4.3; also see Section 4.3.1.5.2). Given the above two
points, if systems with original alkalinities < 140 yeq £-1 are acidi-
fied to the maximum amount (alkalinity loss of 100 yeq jT1), then
resulting alkalinities will be < 40 yeq £"1, which is the threshold for
biological effects on a long-term basis (< 40 yeq £"!).
This value of 140 yeq £~1 alkalinity considers only long-term acidifi-
cation. When the phenomena of short-term acidification is considered, 200
yeq £~! appears to be a reasonable value because, during spring snow-
melt, alkalinity reductions of > 100 yeq £-1 lasting several weeks have
been reported (Galloway et al. 1980b, Galloway and Dillon 1983). In fact,
200 yeq £~* may underestimate sensitive water bodies sensitive to
short-term acidification.
The use of 200 yeq £~1 as the boundary between sensitive and nonsensi-
tive aquatic systems has some deficiencies. Although it may underestimate
aquatic systems sensitive to short-term acidification, it is probably an
overestimate for long-term acidification. One reason has already been men-
tioned, namely 140 yeq a~l is more reasonable than 200 yeq £~1.
There are two other reasons. First, the computed threshold ignores any
assimilation of acidic deposition by the watershed. Henriksen (1982a) states
that for some low alkalinity systems, up to 40 percent of the increase in
$04^" may be matched by an increase in base cation rather than an
increase in H+ (loss of alkalinity). Secondly, the computed threshold is a
static measure. It represents the instantaneous ability of lake or stream
water to assimilate acidic deposition and is quantitatively measured as the
4-25
-------
5.5
5.0-
4.5 -
4.0
-100
100 200 300
ALKALINITY (ueq A"1)
400
Figure 4-3. The change in pH for a given change alkalinity at two
alkalinity levels and an example of pH-alka'iinity relation-
ship for aquatic systems. The alkalinity data were obtained
by a single or multiple endpoint titrations using a pH meter.
The solid S-shaped line represents the median values. The
dashed lines form a 68% band (analogous to one standard
deviation). Each line is a smoothed (cubic spline) moving
average of five points of the appropriate percentiles (2, 16,
50j 84, 98) computed from the data at each 0.1 pH point.
Data are for 928 lakes and streams in New York State.
Adapted from Hendrey (1982).
4-26
-------
PH
8.0
7.5
7.0
6.5
6.0
5.5
5.0
4.5
4.0
-200
o o
O 00
o o o o
o ooooooo •
o o o
ooo • o • o
o oo «o o o oo o*
o 0*000 o o
• o* «o o ooo ooo
oo «ooo *o o«o o
• o ooo
ooo oo • o • o
000
•• o
00 • O
o »o
•• o
•o o
O 00
o
00*
o*
••
•
o o
oo
o o
o* o
o
-100
100
200
300
400
500
ALKALINITY (peq t"1)
Figure 4-4. Plot of pH as a function of alkalinity (inflection point
alkalinity) for 201 lakes in New England. The residuals
from the regression of pH on alkalinity were not signifi-
cantly correlated with water color. Adapted from Haines
and Akielaszek (1983).
4-27
-------
ANC or alkalinity of the water (Stumm and Morgan 1970). Since it is a
static measure, it ignores processes that control the rate of alkalinity
generation in the watershed and aquatic systems. Thus an aquatic system
could have a low alkalinity but it may be quite resistant to alkalinity loss
due to acidic deposition because the rate of alkalinity generation in the
watershed or water body may be quite large. Therefore, for both cases, an
increase of 100 yeq i"1 of S042" would not result in a decrease of
100 y eq &"1 alkalinity. As a result of these two deficiencies in the
use of alkalinity as a sensitivity indicator, aquatic systems may not be as
sensitive to acidic deposition as the alkalinity value estimates. However,
it is true that lower alkalinity systems are generally less able to assimi-
late acidic deposition than higher alkalinity systems. And since there is
wide spatial variability in the processes that control rates of alkalinity
generation, the static measurement of alkalinity has been used as a general
indicator of aquatic system sensitivity.
It should be noted that alkalinity, as a measure of sensitivity to acidic
deposition, is unaffected by the presence of organic acid/base systems.
Alkalinity reflects the total acid neutralizing capacity of the water, both
inorganic and organic. However, as noted above, assuming that two waters
with equal alkalinity but different organic content are equally sensitive
implies that both systems generate alkalinity at equal rates. Data available
are inadequate to test this assumption, but the primary processes involved in
alkalinity generation in wetlands dominated by organic acids may be markedly
different from comparable processes in other systems (cf. Sections 4.3.2.2,
4.3.2.4, and 4.3.2.5).
In summary, the boundary between sensitive and nonsensitive aquatic systems
commonly used is 200 yeq £~! (value before the onset of acidifica-
tion, which in areas receiving acidic deposition is greater than current
alkalinity). This value has been selected after consideration of (1) current
levels of acidic deposition, (2) the increase in SO^- levels of surface
waters due to acidic deposition, (3) the relationship between pH and alka-
linity in oligotrophic systems, and (4) the pH and alkalinity values at which
acidification will result in biological effects. The choice of 200 yeq
£~1 of alkalinity identifies all aquatic systems possibly sensitive to
long-term acidification as a result of current levels of acidic deposition
but may underestimate those systems sensitive to short-term acidification.
Watershed/aquatic systems having low alkalinities (< 200 yeq i-1) but
rapid rates of alkalinity regeneration, may not be acidified as much by an
increase of 100 yeq &-1 of S042~ from acidic deposition. To estab-
lish true sensitivity, alkalinity generation processes in the watershed/
aquatic system may have to be considered. However, until it is possible to
generalize these processes to regional scales, the static measure of surface
alkalinity remains the best indicator of sensitive aquatic systems.
4.3.2.6.2 Internal production/consumption of ANC. The internal production of
alkalinity is usually overlooked in considerations of lake sensitivity, but
it may be very important, especially in lakes with low alkalinity. In the
epilimnion, the major pathway for the production of alkalinity is primary
production (photosynthesis) (Brewer and Goldman 1976, Goldman and Brewer
4-28
-------
1980). The generation of alkalinity depends upon the use of NOa" as a
nitrogen source by algae:
106 C02 + 16 NOs" + HP042' + 122 H20 + 18 H+ •* [4-6]
Cl06H263°110N16pl + 138 $2..
Any NH4+ use results in a decrease in alkalinity in the lakewater
(Schindler, D. W., Department of Fisheries and Oceans, Winnipeg, Manitoba,
unpub. studies). Although it is well known that NH4+ is preferred over
N03~ (Lui and Rolls 1972, McCarthy et al . 1977), mass balances of the two
species in many north-temperate lakes are such that N03~ use often
surpasses NH4+ use (NRCC 1981; Dillon, P., Ontario Ministry of the
Environment, Rexdale, Ontario, unpub. studies). For example, Zimmerman and
Harvey (1979, NRCC 1981), observed an increase in pH of the epilimnion of
Croisson Lake (Ontario) from 5.1 in May to 6.6 by August 1978. Neither pre-
cipitation pH nor the pH of water supplied by/inflowing streams could account
for this decrease in H+ concentration. Over the same period, NOs" concen-
tration decreased from 15 to 1.0 yeq jr1, while NH4+ concentration varied be-
tween 0.3 and 0.5 yeq jr1. N03~ uptake during photosynthesis, therefore may
have generated about 14 yeq a~*- of alkalinity, sufficient to raise the
pH of the epilimnion (cf. Figure 4-3).
On the other hand, the reverse of the photosynthetic reaction (i.e., aerobic
respiration) is a source of H+ (consumes alkalinity). Thus, N03~ up-
take during primary production results in a net gain in ANC only to the
extent that photosynthesis exceeds aerobic respiration (decomposition), i.e.,
to the extent that the inorganic N03~ converted to organic nitrogen is
stored permanently in the lake's sediments (or transported downstream). The
uptake of N03~ (corrected for uptake of NH4+) is often in the range
of 10 to 20 yeq a~ over the summer in oligotrophic north-temperate
lakes (Dillon 1981). The net uptake calculated on a whole-year basis, on the
other hand, may be closer to 5 y eq £-1. Even this lesser amount may be
significant; e.g., in a lake with mean depth of 10 m, this represents a
production of 50 meq alkalinity m~2 yr~*, an amount comparable to the
deposition of strong acids in many parts of eastern North America.
Therefore, an increase in nutrient levels may increase the alkalinity gener-
ation if N03~ is used as the N-source, on a net basis, and the organic N
is lost permanently to the sediments. Fertilization with NH4+, on the
other hand, may result in lake acidification. Nutrient status is therefore
very important in determining the sensitivity of a lake to acidic deposition.
Some lakes classified as potentially sensitive based on their geologic and
hydrologic setting may, in fact, be insensitive as a result of cultural
eutrophication.
Internal processes within the hypolimnion may also deplete or produce alka-
linity (Schindler et al. 1980b, Cook 1981, NRCC 1981).
Acidification of lakes by acidic deposition results in increased transparency
(Dillon et al. 1978, Schindler et al . 1980b, NRCC 1981, Schindler and Turner
4-29
-------
1982, Van 1983; Section 4.6.3.4). Therefore, hypolimnetic primary production
(by phytoplankton or periphyton), and associated production of ANC, may be
elevated relative to non-acidic lakes of equivalent nutrient and morphometric
status.
Under oxic conditions, mineralization of organic matter (produced principally
in the epilimnion and metalimnion) results in a decrease in alkalinity or de-
pletion of ANC:
CioeHzesOiioNiePi + 138 02 + 106 £02 + 122 H20 [4-7]
+ 16 HN03 + H3P04-
This reaction may occur in the hypolimnetic water or at the sediment water
interface. As mentioned earlier, some of the organic matter produced in the
lake is permanently stored in the sediments (i.e., respiration < production).
Under anoxic conditions, several microbial processes that occur in the hypo-
1imnion (or in the surficial sediments) and that require organic material
produce alkalinity:
S042' reduction
C106H263°110N16pl + 53 S042' + 106 H+ + / [4-8]
106 C02 + 16 NH3 + H3P04 + 106 H20 + 53 H2$ \
~ reduction
+ 24 N03~ + 24 H++ 30 C02 + 12 N2 + 42 H20 [4-9]
Mn4+ reduction
Cl06H2630l!ONl6Pl + 236 Mn02 + 472 H+ -> [4-10]
236 Mn2+ + 105 C02 + 8N2 + H3P04 + 366 H20
4-30
-------
reduction
C106H263°110N116P1 + 424 FeOOH + 848 H+ -* [4-11]
424 Fe2+ + 106 C02 + 16NH3 + H3P04 + 742 H20.
However, the alkalinity produced by some of these processes may be temporary.
Fe2+ and Mn2+ (and Nfy"1") production is probably largely temporary,
with the reverse reaction occurring as soon as oxic conditions again prevail
at overturn. N03~ reduction occurs in hypolimnia or in lake sediments,
but the N2 evolution makes the reaction irreversible; therefore, this rep-
resents a source of permanent alkalinity. S042~ reduction results in
permanent alkalinity if the S2" formed is irreversibly lost to the sedi-
ments. Any S2~ (HS~, H2S) left in the water column at fall circulation
is re-oxidized to S042", with concurrent loss of alkalinity.
The critical factor with respect to the ability of a lake's hypolimnion to
assimilate acidic deposition is its oxygen regime. At the Experimental Lakes
Area (ELA), Schindler et al. (1980b), Kelly et al. (1982), and Cook (1981)
studied fertilized and unfertilized lakes that had anoxic hypolimnia and
consequent summer alkalinity production. Increased S042~ input result-
ed in increased alkalinity generation. During the experimental acidification
of Lake 223 (anaerobic hypolimnion) at ELA, acid additions (H2S04) were
only 31 to 38 percent effective at depleting alkalinity, in large part due to
$04^- reduction in the hypolimnion (Schindler et al. 1980b). In Muskoka
and Haliburton counties (Dillon et al., Ontario Ministry of the Environment,
Rexdale, Ontario, unpub. results) and in the Sudbury area (Van and Miller
1982), most study lakes did not have large anoxic zones in their hypolimnia
and appreciable S042~ reduction was not observed. Fertilized lakes (Van
and Lafrance 1982) were an exception, however. Kilham (1982) calculated an
acid-base budget for Weber Lake, a small seepage lake (with an anaerobic
hypolimnion during summer stagnation) in northern Michigan. According to
Kilham (1982), H+ deposition to the lake has increased approximately
20-fold over the last 25 years, yet lake alkalinity has increased. Alkalin-
ity production resulting from NOs" uptake and S042' reduction has
been sufficient to completely neutralize the H+ entering the system as
atmospheric deposition. A similar response was described for a bog envi-
ronment in Section 4.3.2.5. The occurrence of a reducing environment within
the aquatic system may, in part, therefore, determine the aquatic response to
acid inputs.
4.3.2.6.3 Aquatic sediments. The potential for lake sediments to assimilate
acidic deposition is not quantitatively understood. The same microbial
processes that occur in hypolimnia occur in lake sediments, but the contri-
bution of alkalinity to the overlying waters is controlled by slow diffusion
processes.
That sediments also supply ANC by chemical pathways can be inferred from
neutralization experiments near Sudbury, Ontario (Dillon and Smith 1981).
The acidified lakes studied had reduced pH (of - 4.0 to 4.5) in the upper 5
cm of the sediments, with pH of 6.0 to 7.0 at greater depth. Following
4-31
-------
neutralization of three study lakes with CaC03 plus Ca(OH)o, the pH of
the upper sediments increased to the same levels as the deep sediments. Sed-
iment consumption of the added ANC varied from 33 to 60 percent of the total
added to the lake. The sediments were therefore able to supply 0.9 to 3.0 eq
m"^ of BNC. Over the subsequent five years, one of the three neutralized
lakes reacidified. The pH of the upper 5 cm of sediment decreased to levels
comparable to those measured prior to neutralization of the lake.
The same processes that occur in soils may occur in lake sediments. Hongve
(1978) has suggested that cation exchange in lake sediments may contribute to
acidification of lakewater as a result of Caz+ exchange for H+. He sug-
gested, however, that the reverse process will occur with increased lake
acidity. These results were demonstrated in laboratory experiments only.
4.3.3 Location of Sensitive Systems (J. N. Galloway)
Identification of aquatic systems sensitive to acidic deposition ideally
should take into account all factors outlined in Section 4.3.2. Unfortu-
nately, for most of these parameters, regional data are not available nor do
we have a clear understanding of how parameters interact. The alkalinity of
a surface water reflects a combination of many relevant factors. Aquatic
systems with an initial alkalinity < 200 yeq &"1 (before the onset of
acidification) have been defined in Section 4.3.2.6.1 as potentially sensi-
tive to acidification by acidic deposition. For regions not yet receiving
acidic deposition, these systems can be located by direct analyses of alka-
linity over large areas. For regions currently receiving acidic deposition,
present day measurements of alkalinity must be corrected for the estimated
acidification (decrease in alkalinity) to date. Alternatively, geological,
soil, and land use maps can be used to identify aquatic systems with natu-
rally low alkalinity and high sensitivity to acidic deposition. The advan-
tage of the first method is that the alkalinity is determined by an actual
measurement. The disadvantage is that thousands of measurements have to be
made of lower order streams and headwater lakes to determine sensitivity on a
regional basis. In the absence of measurements, no mechanism exists to esti-
mate the alkalinity. In addition, estimates of acidification to date are
approximate, at best. The advantage of the second method is that broad re-
gional determinations can be made. The major disadvantage, however, is that
fine detail is unavailable. Therefore, the proper way to address this issue
is to use regional data on bedrock, soil, and land use characteristics to de-
termine general areas of sensitivity, and follow up with alkalinity surveys
in regions designated as sensitive.
The state of our knowledge is illustrated with four figures. Using bedrock
geology as a criterion, Galloway and Cowling (1978) made a rough approxi-
mation of sensitive areas in North America (Figure 4-5). Their identifi-
cation was improved by the addition of information on soils and surficial
geology for eastern Canada (NRCC 1981; Figure 4-6). Unfortunately, a similar
map for the United States is not yet available.
As a check on the use of soil characteristics and bedrock geology as
predictors of low alkalinity waters, Hendrey et al. (1980b), using methods
developed by Norton (1980), compared surface water alkalinities with
4-32
-------
Figure 4-5. Regions in North America containing lakes that are
potentially sensitive, based on bedrock geology, to
acidification by acidic deposition. Adapted from
Galloway and Cowling (1978).
4-33
-------
HIGH SENSITIVITY
Granite, granite gneiss,
orthoquartzite, syenite
INTERMEDIATE-HIGH SENSITIVITY
Volcanic rocks, shales, greywacke
sandstones, ultramafic rocks, gabbro,
mudstone, and metamorphic equivalents
INTERMEDIATE-LOW SENSITIVITY
Calcareous clastnc rocks, carbonate rocks
interbedded or interspersed with non-calcareous
sedimentary, igneous and metaporphic rocks
Limestone, dolomite and metamorphic
equivalents
Figure 4-6. Map of areas containing aquatic systems in eastern Canada that are potentially sensitive ,
based on bedrock geology and surficial soils, to acidic deposition. Adapted from NRCC (1981).
-------
sensitivity predicted on the basis of geology, county by county (U.S.); they
found clear correlations. Haines and Akielazsek (1983) surveyed New England
lakes and compared alkalinities with predictions (by Norton) of sensitivity
on a drainage basin basis; high correlations existed.
As mentioned earlier, instead of using maps of soil characteristics and bed-
rock geology to predict areas of low alkalinity, actual values of alkalinity
may be measured and displayed on a map. Omernik and Powers (1982) used such
an approach, as is shown in Figures 4-7 and 4-8.
The maps are a useful presentation of regions where waters of low alkalinity
might be found. In essence, they were created using a predictive technique.
Specifically, existing data on surface water alkalinity were compiled and
then correlated with geologic, soil, climatic, physiographic, and human fac-
tors. These correlations were then used to predict mean annual alkalinity
for areas without alkalinity data. There are, however, problems with this
predictive technique. First, if the compiled data are not themselves repre-
sentative of a region (e.g., if they are weighted towards small or large
watersheds instead of a representative mixture), the resulting correlations
and predictions will also be biased. Second, it is difficult to estimate the
errors involved in the prediction. Third, as the authors note, a certain
degree of averaging was required to create a map on the scale of the United
States. Therefore, the ranges cited are for the mean annual alkalinity of
most surface waters in a given region. In areas where substantial
heterogeneities in soil, geology, elevation, land use, etc. occur there may
be large variations from the mean. Unfortunately, sensitive areas generally
occur in regions with large variations in elevation and soil thickness.
Regional maps are currently being developed and scaling problems associated
with these maps may be less.
Several regions in North America contain aquatic systems with low alkalinity
that are sensitive to acidic deposition: much of eastern Canada and New
England, parts of the Allegheny, Smoky, and Rocky Mountains, the northwestern
and north central United States (Galloway and Cowling 1978, NAS 1981, NRCC
1981, McCarley 1983), and the south and east coasts of the United States
(Omernik and Powers 1982). A large amount of more detailed survey work is,
however, required to determine the levels of alkalinity and the degree of
sensitivity of individual aquatic systems.
4.3.4 Summary—Sensi ti vi ty
The sensitivity of aquatic systems to acidic deposition depends on the compo-
sition of the deposition, the total rate of the loading (wet plus dry depo-
sition), the temporal distribution, and the characteristics of the receiving
system.
Atmospheric deposition is a major source of ions to aquatic systems. The
elements supplied by atmospheric deposition in important quantities include
P, S, N, and H. The effects of S on aquatic systems are independent of type
of deposition (wet or dry) or chemical speciation (e.g., S02, S04 )•
4-35
-------
LEGEND
•c200 yeq £-1
200 - 399 peq a~
400 - 599 jjeq x,'
Figure 4-7. Total alkalinity of surface waters. Adapted from Omernik
and Powers (1982).
4-36
-------
I
GO
DRAFT
TOTAL ALKALINITY OF
SURFACE WATERS
SOURCE JAMES M OMERNIK
CORVALLIS ENVIRONMENTAL RESEARCH LABORATORY
U S ENVIRONMENTAL PROTECTION AGENCY
TOTAL ALKALINITY*
Iw eg III
<50
50 TO 98
100T0189
| | OVER 200
REPRESENTATIVE OF MEAN ANNUAL VALUES
Figure 4-8.
Total alkalinity of surface waters in the eastern United States. Compiled by U.S. DOE
from regional alkalinity maps developed by Omernik and Powers (in press).
-------
The ability of receiving systems to assimilate acidic deposition depends upon
many factors, three of which are size, composition, and hydrologic residence
time. In general, the greater the watershed to surface water ratio, the
greater the ability to assimilate acids. The composition and characteristics
of the soil are also important. Soil systems derived from calcareous rock,
for example, are better able to assimilate acidic deposition than soils
derived from granite bedrock, with low CEC, percent BS, and sulfate adsorp-
tion capacity. The hydrologic residence time and flow path are also impor-
tant. Generally, the longer acidic deposition stays in contact with the
terrestrial system, the less the effect of acidic deposition on the aquatic
system. Aquatic systems that tend to be the most sensitive to acidic
deposition are located 'downstream' of terrestrial systems that are small,
have slowly weathering soil and bedrock, have short hydrologic residence
times, and as a result assimilate only a part of the acidic deposition that
falls on them.
The above are broad generalizations concerning complex systems; additional
details are provided in the preceding sections and in Chapter E-2. Unfor-
tunately, for most of these parameters, regional data are not available nor
do we have a perfect understanding of how parameters interact. Thus, sur-
face water alkalinities are often used as a simple (and approximate) indi-
cator of sensitivity. After consideration of the maximum loss of alkalinity
that could be caused by acidic deposition and the alkalinity range where
biological effects begin, sensitive aquatic systems are defined as those with
alkalinity < 200 yeq a ~l (prior to the onset of acidification) (see
Section 4.3.2.6.1). Such systems are located throughout much of eastern
Canada and New England, parts of the Allegheny, Smoky and Rocky Mountains,
the northwestern and north central United States, and the south and east
coasts of the United States.
4.4 MAGNITUDE OF CHEMICAL EFFECTS OF ACIDIC DEPOSITION ON AQUATIC ECOSYSTEMS
The previous sections have laid a foundation of important definitions, con-
cepts and characteristics of deposition and receiving systems. The following
sections discuss what is known about the degree of acidification of sensitive
systems, and the methods used to determine the degree and rate of acidifi-
cation.
Mechanisms by which atmospheric acid inputs are transferred to aquatic sys-
tems are not completely understood; the available literature is summarized in
Section 2.6, Chapter E-2. Seip (1980) outlined three possible conceptual
models for acidification:
0 A model based on direct effects, i.e., assuming that a substantial
fraction of the precipitation reaches streams and lakes essentially
unchanged,
0 A model emphasizing the increased deposition of mobile anions, par-
ticularly $042-, and
0 A model based on effects on aquatic systems as a result of increased
soil acidity.
4-38
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Precipitation may reach lakes and streams with minimal contact with soils and
bedrock (i.e., essentially unchanged) via deposition directly onto surface
waters, or via overland flow, or rapid interflow, especially through soil
macropores and channels. Our understanding of terrestrial-aquatic transport
processes and the prevalence of macropores and channelized flow (Section
4.3.2.4; Section 2.1.3.1, Chapter E-2) is insufficient for a final analysis
of the importance of the above processes. It is unlikely, however, that
direct effects, by themselves, can explain the magnitude of acidification
observed to date. Likewise, soil acidification, with consequent effects on
aquatic systems, although theoretically important has yet to be demonstrated
in the field. Therefore most soil scientists favor the proposed 'mobile
anion mechanism' (Section 2.6, Chapter E-2). In essence it states that the
introduction of a mobile anion into an acid soil will cause the pH of the
soil solution to drop, and a decrease in the pH of surface waters 'down-
stream', regardless of whether the anion is introduced as a salt or an acid.
Increased concentration and movement of an anion, e.g., S04 ,
through a catchment results in increased concentrations of H+ and Al3+
simply as a result of the requirement for cation-anion balance and because
most exchangeable cations in acid soils are H+ and A13+ (see Section 2.6,
Chapter E-2 for further details). Consideration of possible mechanisms for
acidification facilitates interpretation of the following sections.
4.4.1 Relative Importance of HN03 vs H2S04 (J. N. Galloway)
H2S04 is generally more important than HN03 in acidification of aquatic
systems for two reasons. First, in most areas impacted by acidic deposition,
atmospheric H2S04 loading exceeds HN03 loading (Table 8-7, Chapter
A-8J. Second, in systems impacted to date N0a~, more so than
S04 , tends to be retained within the terrestrial ecosystem (Table
4-3). Thus, S042~ often acts as the 'mobile anion' described above.
Retention of anions within the watershed may be associated with biological
and chemical transformations similar to those described in Section 4.3.2.6.2.
resulting in the production of alkalinity and thus neutralization of H
input as HN03 or H2S04-
Sulfate retention in the terrestrial ecosystem is controlled largely by
sulfate adsorption in soils (Sections 4.3.2; Section 2.2.8, Chapter E-2). In
general, in granitic watersheds common in the northeastern United States and
eastern Canada, the sulfate adsorption capacity (SAC) of soils is low.
Sulfate in deposition may move through the terrestrial ecosystem and thus
play an important role in the movement of cations, including H+, from the
terrestrial to the aquatic system. In certain kinds of soils, however, such
as those common in the southeastern United States, SAC is high, retarding the
movement of cations. Systems with high SAC are less sensitive to
acidification at this time. Depending on the extent and magnitude of future
sulfur deposition, such systems will, however, become more sensitive.
Nitrate retention in soils, on the other hand, results primarily from
biological activity, conversion of N03" to organic nitrogen by plants
and bacteria. Both S and N are essential plant nutrients. However, S, as
opposed to N, is usually present in soils at levels adequate for plant
4-39
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TABLE 4-3. THE RETENTION OF NITRATE, AMMONIUM, AND SULFATE IONS BY
FORESTED WATERSHEDS IN THE NORTHEASTERN U.S. AND EASTERN CANADA
% retention in the watershed on an annual basisa
Location N03~ NH4+ S042~
Adirondacks, NY
Hubbard Brook, NHC
Muskoka-Haliburtond
Ontario
Kejimkujik National Park
Nova Scotia6
White Oak Run, VA^
35
15
81
99
99
92
90
83
98
99
-8
-39
21
69
Negative values indicate net loss from the watershed.
bGalloway et al. 1983c; inputs include wet and dry deposition.
°Likens et al. 1977 inputs include weet deposition only.
dScheider et al. 1979b; inputs based on measurements of bulk deposition.
eKerekes 1980; inputs include wet deposition only.
^Shaffer and Galloway 1982; inputs include wet deposition only.
4-40
-------
growth. In areas of very high N03~ deposition or after long periods of
atmospheric additions of N03~, N03~ might also be present in
excess of plant requirements. In such cases, N03~ mobility would be
increased, and HN03 could play a greater role in the acidification of
surface waters.
Acidification from HN03 varies seasonally, reflecting in part seasonal
variations in biological activity and in part seasonal variations in
hydrologic residence time. During most of the year, the residence time of
water in the soil is sufficient to allow for rapid uptake of N03
(Likens et al. 1977). Of the N03~ released from the terrestrial sys-
tem, most comes during periods of high flow (spring snowmelt, large intense
rainstorms). During these types of events the rate of nitrogen transport
through the system is faster than the rate of biological uptake. In addition
to the effect of hydrologic residence time on N03~ transport through
soil systems, a temperature dependency also exists. During warm periods
(e.g., summer), when biological activity is highest, N03~ is effi-
ciently retained within the terrestrial systems. During colder periods
(e.g., winter), maximum N03~ concentrations often occur (Likens et al.
1977, Galloway and Dillon 1983).
Therefore, larger fluxes of N from the soil system to surface waters with
potential impacts on the acidity of lakes and streams dccur principally
during two periods: winter base flow and spring snowmelt. Seasonal var-
iations in N03~ concentration are illustrated for the outlet of Woods
Lake, a small oligotrophic lake in the Adirondack Mountains, NY, (Figure 4-9)
and for two of the inflows to Harp Lake in Southern Ontario (Figure 4-10).
N03~ values are highest in the winter and during spring snowmelt
(usually in March and April). Galloway et al. (1980b) studied the role of
N03~ in the acidification of Woods and Panther Lakes during the 1979
snowmelt. Decreased alkalinity in the two lakes during snowmelt (Figure
4-11) was related to dilution of base cations (CB) and an increase in
HN03 in the lake epilimnion (Section 4.4.2). Although S042~ concen-
trations changed only slightly in Woods and Panther Lakes during snowmelt,
S04^~ still probably contributed to the acidification in an indirect
manner, namely, by causing long-term alkalinity reductions (as opposed to
episodic) (Galloway et al. 1983c). Thus, the episodic reduction of alkalin-
ity due to N03" is apparently adde3~to the long-term reduction of alka-
linity due to S042" (See Sections 4.4.2 and 4.4.3).
Galloway et al. (1980b) concluded that the primary cause of the increased
N03" concentration was release from the snowpack. An analysis of two
additional snowmelt periods (1978, 1980) supports this conclusion (Galloway
et al. 1983b). N03~ from nitrification in the soil could also contrib-
ute to the increase in N03" observed in surface waters (Likens et al.
1977). To better determine the source of increased N03" levels during
snowmelt, information is needed on snowmelt flow path and on mechanisms that
add or subtract NOs" to (or from) the snowmelt as it travels to the
stream or lake.
4-41
-------
.£>
ro
JAN APR JUL OCT JAN
78 79
APR JUL OCT
JAN APR JUL OCT JAN
80 81
Figure 4-9. The concentration of N0s~ in the outlet of Woods Lake, Adirondack Mts., NY. Adapted
from Galloway and Dillon (1983).
-------
CT)
co
o
CD
3.
2000
1600
1200
• 800
n
o
400
0
HARP INFLOW 5
...A AA.IA
1976 1977
1978 1979 1980
Figure 4-10. Nitrate concentration in inflow 3A and inflow 5 to Harp
Lake, Ontario, for a 4-year period (June 1976 - May 1980)
Adapted from Galloway and Dillon (1983).
4-43
-------
240
200
160
CT
O)
-80
SPRING
, THAW _
LEGEND
PANTHER LAKE
WOODS LAKE
MIDWINTER
JHAW.
SPRING
THAW
J I I I 1 1 1 1 1 1 L
J L
J L
JFMAMJJASONDIJFMAM
1978 ' 1979
Figure 4-11. Temporal trends in alkalinity at outlets of Woods and Panther Lakes. Adapted from
Galloway et al. (1980a).
-------
The chemical changes that accompany the seasonal decreases in pH and
alkalinity, however, are not consistent from study area to study area. For
example, Jeffries and Snyder (1981) found that S042" levels increase in
several streams in the Muskoka-Hali burton area of Ontario at peak flow during
snowmelt. On the other hand, Johannessen et al . (1980) reported decreasing
S042~ during snowmelt in streams in Norway. Three of the six streams
studied by Jeffries and Snyder (1981) exhibited declining NOr-
concentrations associated with peak H+ concentrations, a finding opposite
to that of Galloway et al . (1980b) in the Adirondacks. To better understand
the processes involved in short-term acidification (discussed in more detail
in Section 4.4.2) data on N03- and S042- behavior during snowmelt
(and other times of the year) are needed for areas of North America other
than the Adirondacks and southern Ontario.
In summary, during most of the year S042- is the most important
associated with acidification related to acidic deposition. However, in
winter and in the spring, in areas studied in the Adirondack Mountains, NY
and in southern Ontario, N03~ may become more important both in an
absolute sense and relative to $042-. in general, the effects of
H£S04 and HNOs, on acidification of aquatic ecosystems are:
0 H2S04 causes long-term (decades) alkalinity reductions on a regional
basis.
HN03 can cause episodic short-term (weeks) alkalinity reductions that
are in addition to the long-term reductions caused by H2S04-
4.4.2 Short-Term Acidification (J. N. Galloway and J. P. Baker)
Acidification of lakes and streams during major hydrologic events, apparently
as a result of acidic deposition, has been demonstrated in Norway (Gjessing
et al. 1976, Henriksen and Wright 1977, Johannessen et al . 1980), Sweden
(Oden and Ahl 1970, Hultberg 1977), Finland (Haapala et al . 1975), Ontario
(Scheider et al . 1979a, Jefferies et al . 1979, Jeffries and Snyder 1981) and
the northeastern United States (Johannessen et al . 1980; Galloway et al .
1980b, 1983c). The hydrologic event leading to acidification has usually
been snowmelt; however, periods of heavy rain also can result in decreases in
alkalinity and pH (e.g., Scheider et al . 1979a).
Episodic events have resulted in decreases in pH of greater than or equal to
one pH unit in several reported cases (Table 4-4). For example, the change
in pH of Harp Lake Inflow #4 during the snowmelt of 1978 was 1.2 pH units
(Jeffries et al . 1979) while the alkalinity decrease was 100 ueq i'1.
During the 1979 spring snowmelt, the pH and alkalinity decreases in the epi-
limnion of Panther Lake were 1.8 and 180 yeq £~S respectively
(Galloway et al . 1980a,b). Streams in Ontario and New York with lower
pre-melt pH's and alkalinities had correspondingly smaller decreases (Table
4-4).
It is important to note, however, that not all aquatic systems within areas
receiving acidic deposition experience significant episodic pH depressions.
Likewise, in areas not currently impacted by acidic deposition, pH
4-45
-------
depressions during snowmelt or heavy storms occur naturally in some streams
(Table 4-4). Both 'natural' and 'anthropogenic' (i.e., acidic deposition)
factors contribute to short-term acidification.
Two processes play primary roles in natural acidic episodes during snowmelt
or storms: dilution and hydrologic flowpath. Simple mixing of dilute pre-
cipitation, even 'non-acidic1 (pH > 5.0) precipitation, into stream waters
can result in declines in pH and alkalinity. .For example, given a stream at
pH 7.0 with an alkalinity of 100 yeq &~l that receives during snow-
me}t an equal flow of meltwater at pH 5.6 with an alkalinity of 0 yeq
£~ , the endpoint (assuming in this simple example no interaction with
soils or stream sediments) will have an alkalinity of 50 yeq a~ and
a pH of approximately 6.0 (based on Figure 4-3). Dilution of stream water
with large quantities of precipitation or meltwater can result in distinct pH
declines, particularly in low alkalinity waters with pre-episodic pH _^ 6.0.
Shifts in hydrologic flowpath during storm events and snowmelt may also play
a role in stream acidification. As discussed in Section 4.3.2.4.1, the ma-
jority of water reaching a 'typical' stream during a storm event results from
rapid interflow, and may pass principally through upper soil horizons. If
these upper soil horizons are acidic, this shift in hydrologic flowpath can
result in pH depressions in the receiving water. Low pH soil solutions are
often dominated by organic acids. In addition, however, nitrification, espe-
cially during long drought periods in summer or under the snowpack, may
supply additional H+ ions (Section 4.4.1). In this instance, the pH
depression would be accompanied by an increased flux of ~
The mechanism resulting in short-term acidification must be evaluated for
each system. To what degree does acidic deposition add to the effects of
natural acidifying processes? This question can be approached in several
ways.
Catchments with similar physical, chemical, and biological features should
exhibit different magnitudes of response given different atmospheric acid
loadings. Data in Table 4-4 generally support this hypothesis. Systems with
an initial pH between 6 and 7 and receiving < 25 kg S042~ ha-1 yr"1
maintained a pH > 5.5 even during episodes. In contrast, in similar systems
receiving > 30 kg S042' ha'1 yr'1, p^l levels dropped below 5.5, and
at times below 5.0.
Mass balance calculations have also been used in an attempt to evaluate the
relative contribution of acidic deposition to acid episodes. Galloway and
Dillon (1983) estimated that dilution accounted for 74 percent (125 yeq
r1) of the alkalinity decline (170 yeq jr1; Table 4-4) observed at
Panther Lake durinq snowmelt; HNOa from the snowpack accounted for 18 per-
cent (13 yeq J2,'1); and H2S04 from the snowpack accounted for 5
percent (8 yeq &'1) . In contrast in Woods Lake, with a lower pre-melt
pH and alkalinity (Table 4-4), the total alkalinity decline (41 yeq r1)
was smaller, primarily as a result of a smaller dilution effect (10 yeq
a~ ). The HNO-j contribution from the snowpack was equal to that for
Panther Lake (31 yeq a'1). Jones et al . (1983) also relied princi-
pally on mass balance assumptions in their evaluation of the cause of acid
4-46
-------
TABLE 4-4. MAGNITUDE OF pH AND ALKALINITY (yeq r1) DECREASES IN LAKES
AND STREAMS DURING SPRING SNOWMELT OR HEAVY RAINFALL. SURFACE ALKALINITIES
IN THESE AREAS ARE GENERALLY < 200 yeq JT1.
Location
Adirondack;, NY
Panther Lake, 1979a
Sagamore Lake, 1979a
Woods Lake, 1979»
Little Moose Lake, outlet,
New Hampshire
the Bowl-upstream, 1973C
The Bowl-downstream, 1973°
South-Central Ontario*1
Harp Lake *4, 19/8
Paint Lake il, 1978
Dickie Lake «0, 1978
Southern Blue Ridge Province
White Oak Run, VA, 1980e
Raven Fork, NC, 198H
Enloe Creek, NC, 1981 f
West Prong of the Little
Pigeon River, 19789
Southwestern Ontario"
Speckled Trout Creek, 1981
Barrett River, 1981
Quebec1
Ste. -Marguerite River, 1981
Minnesota.}
FIT son Creek, 1977
Washington
Ben Canyon Creek1
Idaho
"STTver Creek*
Approximate annual
sulfate loading
(kg ha'1 yr"1)
38
1977b
38
30
27
25
22
17
<20
<20
Crlor
PH
6.6
6.1
4.8
7.0
5.6
6.2
6.6
5.5
4.8
6.0
5.7
5.9
6.3
6.7
6.6
6.7
6.6
7.0
6.1
to episode
Alkalinity
162
29
-39
108
61
-16
20
60
40
76
Water Chemistry
During
PH
4.8
4.9
4.5
4.9
5.0
5.8
5.4
5.0
4.5
5.7
4.4
5.5
5.8
5.1
5.0
5.9
5.5
5.8
5.7
episoae
Alkalinity
-18
-17
-42
8
8
-32
<20
<20
10
70
A pH
1.8
1.2
0.3
2.1
0.6
0.4
1.2
0.5
0.3
0.3
1.3
0.4
0.5
1.JS
1.6
0.7
1.1
1.2
0.4
Change
A Alkalinity
180
46
4
100
53
16
30
6
»Gal1oway et al. •1980b
bSchofield 1977
'Martin 1979
dJeffries et al. 1979
«Shaffer and Galloway 1982
Mones et al. 1983
9S1lsbee and Larson 1982
hKeller 1983
^Brouard et al. 1982
iSlegel 1981
KLefohn and Klock 1983
4-47
-------
episodes in the Raven Fork watershed, NC (Table 4-4). They concluded that
weak acids (i.e., organic acids and aluminum) were the dominant sources of
H+.
Although mass balance calculations can suggest relationships between inputs
and outputs, they cannot establish cause-and-effect. Even in the above
studies the sources of short-term acidification have not been conclusively
determined. For example, data in Galloway and Dillon (1983) cannot disprove
the hypothesis that elevated N0a~ levels resulted from nitrification
(Section 4.4.1), and thus that the observed pH depressions were almost
entirely due to natural processes. Similarly in the Raven Fork watershed, it
is possible that the driving force behind the generation of weak acids
(particularly Al3+) during storm events is acidic deposition.
Perhaps the only way to establish clearly the role of acidic deposition in
acid episodes is through field experiments. Unfortunately the costs and
logistical problems associated with large-scale watershed acidification (or
neutralization) experiments have precluded such studies to date. Seip et al.
(1979, 1980) have conducted several small-scale, short-term watershed experi-
ments in Norway. In 4 to 5 day experiments with simulated rain (pH 3.8 to to
5.2) on mini-catchments (- 80 n£), changes in runoff pH of 0.2 to 0.4
units occurred in response to changes in precipitation pH (Figure 4-12) (Seip
et al. 1979). Seip et al. (1980) also attempted an experiment with neu-
tralized snow, adding sufficient NaOH to neutralize the snowpack (pH 4.3) on
78 m2. No significant increase in runoff pH occurred. However, approx-
imately 65 percent of the Ma* was retained within the watershed, probably
through ion exchange for H+. Thus, the neutralization process was only
moderately ( ~ 35 percent) effective. Additional field experiments are
needed that avoid additions of extraneous cations and that encompass larger,
more diverse watersheds over periods of months or years.
Based on the studies cited above and on other available data sets (Leivestad
and Muniz 1976, Schofield 1980; Table 4-4) it is reasonable to expect that pH
levels during spring snowmelt or heavy rain events may approach as low as pH
4.5 to 5.0. This is the same pH range observed in some cases for chronic,
long-term acidification (Pfeiffer and Festa 1980, Haines and Akielaszek 1983;
Section 4.4.3.1.2). The difference is that in episodic acidification,
aquatic systems with pH's as high as 7.0 can be acidified to pH _< 5.0; in
long-term acidification, aquatic systems with pH's of > 6.5 are, on the
average, too well buffered to be acidified to pH < 5.0.
4.4.3 Long-Term Acidification (J. N. Galloway)
Aquatic systems at risk due to acidic deposition must (1) have low alkalinity
(< 200 yeq jr1) and (2) receive acidic deposition (Figure 4-13). 69111-
bining the information from Figures 4-5, 4-7 and Figure 4-13 identifies
systems with both characteristics. To document actual acidification requires
additional data. These data can be obtained from three types of studies:
(1) analysis of temporal trends in alkalinity and pH, (2) paleolimnological
analysis, and (3) investigation of the importance and source of SO^- in
aquatic systems.
4-48
-------
ARTIFICIAL PRECIPITATION
*.
i
PH
5.0
4.0
3.0
PH
4.6
4.4
4.2
4.0
15 mm acid prec.
5 mm neutr. "
40 mm acid prec. 70 mm neutr.
« 10 mm neutr. "
5 mm neutr. prec. 20 mm acid prec.
• 35 mm acid "
1
17
Runoff
o°o
O o
18
0°=
19
20
0<>
.00
00
23
-»-
24 25
Date
Figure 4-12. pH in artificial rain and in runoff during "minicatchment" experiment, o = oulet 1,
+ = outlet 2. Adapted from Seip et al. (1979).
-------
CANADA
• CANSAP
• APN
AOME
UNITED STATES
• NADP
• MAP3S
Figure 4-13. pH from weighted-average hydrogen concentration for 1980 for
wet deposition samples. Adapted from Barrie et al. (1982);
also Figure 8-17 in Chapter A-8.
4-50
-------
Studies that have used the first technique, historical pH/alkalinity data, to
identify waters acidified by acidic deposition are reviewed in Section
4.4.3.1. Although problems exist with the comparison of historic to recent
data in some studies, significant evidence has been presented suggesting that
the chemistry of deposition can exert a strong influence on the chemistry
(specifically the acidity) of some surface waters.
Supporting this circumstantial evidence is the analysis of diatoms in lake
sediment cores. Although such analysis has been used successfully in
Scandinavia, the technique is still being developed for use in the United
States (Section 4.4.3.2).
A technique for implicitly avoiding the problems of incomplete or imprecise
trend data has been proposed by Galloway et al. (1983c). The approach is
based on considerations of solution electrical neutrality (EC-J = £a*
where c-j is the normality of the ith cation and a^ is the normality of
the ith anion). It is most applicable to clear water lakes and streams (no
organic ions) with no source of sulfur in the bedrock within the drainage
basin. Marine aerosol content corrections should also be performed.
The basis for the technique is that the concentration of $042- in clear-
water lakes and streams has increased due to atmospheric deposition (cf.
Figure 4-1). With the increase in S042- there must be an increase in a
positive ion, H+, Ca2+, Mg2+, etc. If H+ increases, the aquatic sys-
tem is acidified (i.e., alkalinity decreases). If the concentration of
Caz+ or another non-protolytic cation increases, only, then no loss of
alkalinity occurs. For example, Figure 4-14 shows the two extremes of chem-
ical changes that can occur in an aquatic system associated with a five-fold
increase in $042- concentration. At 9ne extreme, the increase in the
S04Z" anion is balanced by an increase in the non-protolytic base cations
(Alternative 1, Figure 4-14). At the other extreme, the increase in
S042~ is balanced totally by an increase in H+, which causes a reduc-
tion of alkalinity (Alternative 2, Figure 4-14). These are extremes; the
real world lies somewhere in between and depends on the characteristics of
the soil and the hydrologic pathway. In sensitive systems (bedrock and soil
with low ANC, low SAC, and short hydrologic path lengths), Alternative 2
appears to be a closer approximation to the process that has occurred. As
support of this Henriksen (1982a), in an analysis of long-term time series
for concentrations of Ca2+ and Mg2+ over gradients of acidic deposition,
concludes that increases in S042~ in lakes are balanced by increases in
H+ (^60 percent) and increases in base cations (£40 percent). Therefore,
in aquatic systems with a predominantly atmospheric source of S042" and
with alkalinity less than 200 y eq a -1, increases in S042- will
cause decreases in alkalinity, i.e., acidification, although the magnitude/
significance of this decrease is dependent on watershed characteristics.
The acidification of freshwaters is the result of a series of complex inter-
related processes. The series begins with increased emissions of S to the
atmosphere, followed by a relatively 'instantaneous1 increase in S deposi-
tion. Eventually the watershed-lake system will attain a new steady-state
condition in balance with these higher S inputs, but attainment of steady-
state may be delayed by several factors. For example, the terrestrial system
4-51
-------
PRE-ACIDIC
DEPOSITION
PERIOD
BASE
CATIONS
HC0
Alternative
1
Alternative
2
ACIDIC
DEPOSITION
PERIOD
BASE
CATIONS
HCO,
SO
2-
BASE
CATIONS
Figure 4-14.
Two extremes for the response of aquatic systems to a
5-fold increase in SO/p-. The height of the boxes relates
to yeq £-1.
4-52
-------
can act, through SCty2" adsorption, as a sink for anthropogenic S
(Section 2.2.8), thus precluding acidification. All of the deposited S does
not enter the aquatic system until the SAC of the soil is saturated. Given a
saturated SAC, an amount of S04 equivalent to the S deposition from
the atmosphere will be discharged to the aquatic system. As described above,
the increased S04^~ in the aquatic system must result in decreased
alkalinity, increased base cation (BC) concentrations, or a combination of
the two. Both, however, may not change at the same rate. An initial
increase in SO^- may result in proportionately large increases in BC
concentrations relative to decreases in alkalinity until the easily weathered
or exchangeable reservoirs of BC's in the soil are depleted. Then, the
concentration of H+ (and possibly aluminum, see Section 4.6.2) increases
more rapidly with a concurrent decrease in alkalinity. Time to reach
steady-state depends in part on the SAC and quantity of readily available
BC's. Values at steady-state depend in part on the rate at which BC's are
resupplied through primary weathering and other processes. Further details
on this conceptual model of freshwater acidification are available in
Galloway et al. (1983a).
It is not known whether systems respond at the same rate to decreases in S
deposition as they respond to increases in S deposition. They may respond
faster, or slower, or not at all. In addition, it is not known whether sys-
tems in the northeastern United States are now at steady-state with current
levels of acidic deposition.
The maximum degree of acidification by acidic deposition depends on the total
increase in acid anions (primarily S042", Section 4.4.1). For each
peq £~ , the maximum loss of alkalinity is 1 yeq £-1. Studies of S04 in
aquatic systems across depositional gradients (Figure 4-1, NRCC 1981, Bobee
et al. 1982), and sulfur budget studies for watersheds and lakes (Dillon
1981, Dillon et al. 1982, Galloway et al. 1983c), indicate that S042-
levels are elevated in aquatic systems receiving acidic deposition and that
the maximum increase in $042- to date on a regional basis (and therefore
maximum loss of alkalinity as a result of acidic deposition) is 100 yeq
A"1 (Section 4.3.1.5.2). The actual loss will certainly be less. The
maximum alkalinity decrease is merely a boundary condition that can be
compared to measured or estimated levels of acidification.
4.4.3.1 Analysis of Trends based on Historic Measurements of Surface Water
Quality (M. R. Church)— ——
4.4.3.1.1 Methodological problems with the evaluation of historical trends.
In assessing the effects of acidic deposition on the chemistry of surface
waters, investigators have searched laboratory records and the literature for
historical data with which to compare present day measurements. The three
water chemistry variables most widely cited in this regard are pH, conduc-
tivity, and alkalinity (Section 4.2.2). A discussion of how methodology for
their determination has changed with time and the comparability of historical
and current data is presented here. For other discussions of this topic see
Kramer and Tessier (1982, 1983).
4-53
-------
4.4.3.1.1.1
4.4.3.1.1.1.1 pH-early methodology--Many of the early measurements of
surface water pH in areas of North America and Scandinavia were made color-
imetrically with acid-base indicators. Materials for visual colorimetry are
inexpensive and readily portable, and, thus, highly amenable to use in rug-
ged, remote field locations, often the site of 'acidification1 problems and
studies. An excellent discussion of acid-base colorimetric indicators is
presented by Bates (1973), who recommends the works of Kolthoff (1937) and
Clark (1922, 1928) for even more exhaustive accounts, descriptions, and dis-
cussions of the proper use of colorimetric indicators.
Acid-base indicators are weak acids or bases that change color with the loss
or gain of a proton (or protons). Such behavior may be represented by the
simplified equilibrium formulation
HIn (Color A) t In- (Color B) + H+. [4-11]
Indicators are used to measure the pH of an unknown aqueous solution as fol-
lows. When the optical characteristics or 'color tone' of an unknown (with
indicator added) match the color tone of a standard reference solution (to
which indicator has also been added), then the two solutions are assumed to
have the same pH. Sometimes the color tone of the unknown solution plus
indicator is matched with calibrated colored discs, each indicating a dif-
ferent pH. The band of pH over which the color change of an indicator is
detectable (by a colorimeter or by the human eye) is called the transfor-
mation range. For visual color comparisons using two-color indicators,
transformation range is generally on the order of two pH units (Golterman
1969, Bates 1973). As Haines et al. (1983) noted, the best results are
achieved near the mid-point of the transformation range of each indicator.
The key assumption in indicator use is that identical color tone of an un-
known and a standard solution of the same temperature to which indicator has
been added implies identical pH. Under some circumstances, however, as Bates
(1973) explains, this is not true.
One reason this assumption may be false can be explained with the aid of the
equation
PaH = pKHIn + log —«— + log Y In [4-12]
l-« Y HIn
where pan is defined by Equation 4-2, pKHIn 1s tne thermodynamic disso-
ciation constant of the acid form of the indicator, a is the fraction of
the indicator in the form In, and YIn and THIP are *ne activity coeffi-
cients in the dissociated and undissociated forms of the indicator, respec-
tively. Color matching (by eye or instrument) indicates only that the term
log a /(I- a) is the same for the unknown and the standard solutions. If
the activity coefficient ratio (the last term in Equation 4-12) of the
indicator is not the same in both the standard and the unknown solutions,
however, the pH of the solutions will not be the same when the colors are
4-54
-------
identical. This is called the 'salt error1 and can be estimated by comparing
the 'true' or electrometric (hydrogen electrode) pH of a series of solutions
having different ionic strengths with their respective values of pH as
implied by use of the indicator (Bates 1973). Salt effects can be minimized
by adjusting the ionic strengths of the buffer solution or the unknown
solution so they are nearly equal. Such adjustments, however, may cause
changes in the reference or unknown pH, introducing further uncertainties.
Another potential source of error is that the addition of an indicator to a
solution may change the pH of that solution. This is most likely in poorly
buffered waters, such as those readily susceptible to acidification. To
overcome this problem the pH of the indicator solution can be adjusted to be
sufficiently close to the pH of the unknown solution so that little pH change
occurs when the two are mixed. This can be accomplished by an iterative
technique using portions of the sample to be determined plus a variety of
indicators (Haines et al. 1983). Alternatively, a quantitative correction
may be applied (Kramer and Tessier 1982).
Bates (1973) indicates that when the above cautions are observed and correc-
tion or adjustment is made for salt effects, an accuracy and a precision of
0.05 and 0.1 pH unit, respectively, can be expected "in properly standardized
routine measurements of buffered solutions." It is likely that colorimetric
determinations of pH made in the field, often under adverse conditions and
often on poorly buffered solutions, may not approach such accuracy or preci-
sion. Although prescribed standarized procedures for performing pH analysis
may be located and critically examined (e.g., see Kramer and Tessier 1982),
the exact procedures actually used to perform specific historical pH measure-
ments are often impossible to reconstruct with certainty. Fortunately, many
of the investigators of acidification trends in surface water pH values
appreciate such considerations (e.g., Wright 1977, Overrein et al. 1980).
4.4.3.1.1.1.2 pH-current methodology—Today, most pH measurements are
made electrometrical ly (potentiometrically) both in the laboratory and, with
the advent of more reliable portable pH meters, in remote field locations as
well.
The 'practical1 or 'operational1 pH was defined in Equation 4-4 (see Section
4.2.2.1). To define standard potentials and set the pH scale, cells of the
following type are used:
Pt; H2(g), Soln. X | KCKsatd.) I reference electrode. [4-13]
The reference electrode is usually either a calomel or silver-silver chloride
electrode (Bates 1973, Durst 1975), which is a primary cell. For most day-
to-day laboratory measurements and all field measurements researchers use
secondary cells in which the hydrogen gas electrode is replaced by a glass
electrode. The proper use of commonly available commercial pH assemblies
(cell plus meter circuitry) has been discussed in many books, journal arti-
cles, and laboratory manuals (e.g., Feldman 1956, Golterman 1969, Bates 1973,
Durst 1975, American Public Health Association 1976, Westcott 1978, Skougstad
et al. 1979).
4-55
-------
An important potential source of error in electrometric pH measurements of
surface waters is the residual liquid-junction potential. Liquid-junction
potentials arise at the point of contact of the reference electrode and the
solution being tested. Such potentials are a function of, among other
things, the ionic strength of the solution being tested. Therefore, the
liquid-junction potential formed in a high ionic strength medium (e.g.,
buffer) is different from that formed in a low ionic strength medium (e.g.,
dilute acidification-prone surface water). The difference between these
liquid-junction potentials is the "residual liquid-junction potential" (Bates
1973). Such a potential can introduce errors on the order of 0.04 pH unit
when ignored in measurements of dilute precipitation samples (Galloway et al.
1979). This type of error can be minimized by equalizing the ionic strength
of the test and reference solutions. Three ways to do this are to 1) add
inert salts (e.g., KC1) to the dilute test solution (this may introduce impu-
rities, thus altering the pH), 2) dilute the standard solution (which alters
its pH~a correction must be applied), or 3) use dilute strong acid standards
(these are not normally reliable pH standards--they must be frequently cali-
brated by titration) (Bates 1973, Galloway et al. 1979).
Another potential source of error in electrometric pH measurements of dilute
solutions is the streaming potential. Errors arise when measurements are
made on dilute solutions while they are flowing or being agitated. Errors of
this sort as large as 0.5 pH unit have been reported for precipitation sam-
ples (Galloway et al. 1979). To eliminate such error, measurements should be
made only on quiescent solutions.
Under rigorous conditions in a properly equipped laboratory, routine elec-
trometric pH measurements can probably approach, at best, an accuracy and a
precision of +_ 0.02 pH unit. Most field measurements of the pH of dilute
surface waters probably have an accuracy and precision of no better than +_
0.05 unit.
4.4.3.1.1.1.3 pH-comparability of early and current measurement methods
—Colorimetric and eTectrometric measurements(usingsecondary cells)are
both based on operational or practical pH (designated by the primary pH cell
and scale) and thus the methods, when applied in an unbiased fashion, are
directly comparable. Attention has been, and should continue to be, placed
on the limits of reliability of the measurement methods as discussed above.
In most of the studies of pH changes cited in the following section (Section
4.4.3.1.2) historical measurements of pH were performed using either Hellige
or Pennwalt color comparator kits. A number of researchers have made direct
comparisons of electrometric pH measurements with measurements obtained using
either or both of these colorimetric kits.
Pfeiffer and Festa (1980) reported considerable bias in such comparisons,
with the measurements by the Hellige kit consistently overestimating pH.
Schofield (1982), however, performed a similar analysis in his Cornell labo-
ratory and found only a slight positive bias (increasing with decreasing pH)
associated with using the Hellige kit (Figure 4-15). He ascribed the
differences in the results of the two studies to "errors in pH meter
measurements obtained by Pfeiffer and Festa (1980)" (Schofield 1982).
4-56
-------
Burns et al. (1981) reported comparisons between measurements by a Hellige
kit and pH meter "agreed to within + 0.15 of a pH unit" and Davis et al.
(1978) reported that measurements using a Pennwalt kit and pH meter "were
found to agree within 0.1 pH units."
Haines et al. (1983) and Norton et al. (1981a) compared values obtained using
both the Pennwalt and Hellige kits to electrometric measurements. They found
that good agreement (within 0.2 pH unit) could be obtained in even low
alkalinity (20 yeq &"1) waters if careful attention was paid to the
use of overlapping indicators and the eventual selection of an indicator such
that the sample pH was near the midpoint of the indicator operating range.
They obtained good agreement between methods in a field survey of New England
lakes (Figure 4-16) when pH was first measured electrometrically and then
colorimetrically with the appropriate indicator (i.e., that indicator with
the midpoint of its operating range closest to the sample pH). Haines et al.
(1983) noted correctly that comparisons based on such a priori knowledge do
not directly mimic historical sampling and analysis conditions in which such
detailed a priori knowledge was not available to guide investigators in the
selection of appropriate indicators. They also noted, however, that "in the
absence of a pH meter, equivalent accuracy could be obtained by repeating pH
measurements with a series of indicators until a result near the midpoint of
an indicator is obtained, or until two indicators with overlapping ranges
agree on the result" (Haines et al. 1983).
Further comments by Haines (1982) bear repeating here. "The early textbooks
on pH measurement (Clark 1922, 1928; Doyle 1941) discussed the problem that
colorimetric indicators might change the pH of poorly buffered samples and
described ways of dealing with this problem. As these texts were the stan-
dard reference works of the period, I assume that reputable scientists were
aware of the problem and took the appropriate corrective measures. Juday et
al. (1935) report just such an occurrence, and cite four references that
showed similar results."
In conclusion, historic measurements of pH certainly deserve scrutiny to
insure their quality but out-of-hand dismissal of such measurements because
of their age and the technique used is not justified. All measurements of
pH, whether historic or current, should be carefully evaluated for lack of
bias when used in a comparative manner to evaluate changes in water quality.
4.4.3.1.1.1.4 pH-general problems—Independent of the methodology em-
ployed, several factors can influence pH measurements of surface waters and
the use of such measurements to estimate the degree of acidification over
time. Principal among these factors is the variation in the pH of surface
waters over relatively short time intervals. The most dramatic and important
"short-term" changes in surface water pH values are those seasonal changes
associated with spring snowmelt and ice-out periods, during which pH may drop
sharply due to release of acid held in ice and snow (Wright 1977, Overrein et
al. 1980, Galloway et al. 1980b, Hendrey et al. 1980a) (Table 4-4). Surface
water pH values during the rest of the year may be considerably higher than
those during snowmelt. Obviously, time of year must be taken into account
when comparisons are made of past and present pH measurements.
4-57
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8
CL
CORNELL
DEC
8
DEC HELLIGE pH
Figure 4-15.
Comparison of colorimetric and meter pH values for
Adirondack lakes waters. Meter measurements by N.Y,
DEC (Department of Environmental Conservation) and
Cornell. Adapted from Schofield (1982).
4-58
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o
I—I
on
O
O
7.5
7.0 -
6.5
6.0
5.5
5.0 —
4.5
I I
I I
4.5 5.0 5.5 6.0 6.5 7.0 7.5
NEW ELECTRODE (pH)
Figure 4-16. Recent lake surface water electrode pH vs recent colorimetric
pH. Adapted from Norton et al. (1981a) and Haines et al.
(1983).
4-59
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Less important, but potentially meaningful, effects are pH changes associated
with the uptake and release of C02 and/or HC03" by aquatic plants.
Most lakes studied in conjunction with acidification problems are usually
oligotrophic, and these changes are probably small. Yet another factor to
consider (especially in streams) is the occurrence of local sources of
groundwater high in C02- One method sometimes used to account for variable
C02 concentrations is to report the pH value after a sample has been
thoroughly agitated to equilibrate its C02 partial pressure with that in
the laboratory. It must be noted, however, that the C02 concentration in a
laboratory can vary considerably from day to day and may be well above that
commonly considered to be the global mean (Church 1980). A number of methods
may be employed to overcome this problem and to insure comparability both
between laboratories and within a laboratory on a day to day basis. These
methods include equilibrating solutions with outside air or determining the
partial pressure of C02 in solutions or in the laboratory atmosphere.
Better yet would be to equilibrate all samples by bubbling with bottled air
of standardized COg content.
4.4.3.1.1.2 Conductivity.
4.4.3.1.1.2.1 Conductiy1ty methodology--The apparatus for measuring
conductivity consists of a cell of two electrodes (often platinum) and a
Wheatstone bridge. The latter is used to balance the resistance of standard
or unknown solutions in which the cell is immersed. Solutions of KC1 are
used to standardize the instrument by calculation of the cell constant.
Important corrections due to temperature variation are also required. Con-
ductivity is routinely reported as umho cm-1 at 25.0 C. Detailed in-
structions for the measurement of the conductivity of surface water samples
can be found in standard laboratory manuals (e.g., Golterman 1969, American
Public Health Association 1976, Skougstad et al. 1979). The precision of
conductivity measurements of surface water samples seems inversely related to
the sample conductivity, with relative standard deviations being as great as
10 percent at levels of conductivity as low as those often reported in stud-
ies of acidification of surface waters (American Public Health Association
1976, Skougstad et al. 1979). Inasmuch as this figure pertains to meas-
urements made under laboratory conditions it is to be expected that meas-
urements made with portable battery-powered conductivity meters in the field
would be less precise.
4.4.3.1.1.2.2 Conductivity-comparability of early and current measure-
ment methods—Routine measurements of conductivity are always made with the
type of apparatus described above, so historical and recent data should be
roughly comparable, if the instrumentation has been properly calibrated and
used. Data published in the literature concerning otherwise comparable lakes
lying in acidic and unaffected areas show that acidified lakes tend to have
higher conductivities (Wright and Gjessing 1976, Dillon et al. 1979), most
likely reflecting the higher hydrogen (and to a much lesser extent sulfate
and nitrate) ion concentrations found in those lakes. Continuous monitoring
of some surface waters in southern Norway has shown increases in conductivity
over a period of decades coinciding with decreases in pH and increases in
transparency of lakes (Nilssen 1980), all changes associated with effects of
acidic deposition.
4-60
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It must be noted here that many factors, not just inputs of acids, may cause
increases in the concentrations of dissolved salts, and thus conductivity, in
surface waters. In fact, increases in conductivity certainly may be associ-
ated with either increases or decreases in pH and alkalinity. For this rea-
son observed increases in conductivity should not be used by themselves to
infer that acidification has occurred.
4.4.3.1.1.2.3 Conductivity-general problems—Conductivity can be expect-
ed to vary seasonally (e.g., it may be much higher during snowmelt than at
other times). Therefore, comparison of historical and recent measurements to
assess acidification should take into account time of year when the measure-
ments were made. Temporary changes in conductivity of surface waters may
also occur during rainfall events. In short, any factor that alters ionic
concentrations will alter conductivity.
4.4.3.1.1.3 Alkalinity. Procedures routinely used to determine ANC of
surface waters have changed significantly over the years, so estimating acid-
ification as the decrease in ANC with time may be extremely difficult (Dillon
et al. 1978, Ontario Ministry of the Environment 1979, Zimmerman and Harvey
1979, Jeffries and Zimmerman 1980, NRCC 1981).
4.4.3.1.1.3.1 Alkalinity-early methodology—Historically, acidimetric
titrations have usually been performed to an endpoint of pH 4.5 determined
electrometrically or to an endpoint determined by a colorimetric indicator
(usually methyl orange) or mixed indicators (e.g., bromcresol green-methyl
orange). ANC measured in this way has been termed total fixed endpoint
alkalinity or TFE (Dillon et al. 1978, Ontario Ministry of the Environment
1979, Jeffries and Zimmerman 1980). These procedures can lead to two types
of problems.
The first problem is associated with the fact that equivalence point pH is a
function of the concentration of the species being titrated. For example,
for inorganic carbon species the exact relationship is
[H+]4 + [H+]S K! + [H+]2 (K! K2 - ct KI- KW) [4-14]
- [H+] iq (2 Ct K2 + Kw) - iq K2 Kw = 0
(Stumm and Morgan 1981),
where
[H+] = hydrogen ion concentration
KI = first dissociation constant of carbonic acid
K2 = second dissociation constant of carbonic acid
Kw = dissociation constant of water
C^ = total inorganic carbon concentration (moles a~ )
4-61
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Appropriate approximate relationships are
[H+] = (CtKi + Kw)°'5 for Ct > 10-6 M [4-15]
[H+] = (CtKi)°-5 for Ct > 10-5 M [4-16]
(Stutnm and Morgan 1981).
Thus, it can been seen that routine titration of all samples to a preselected
pH will not yield accurate values for ANC for all samples, only for those
with specified values of ANC. For example, titration to pH 4.5 is appro-
priate for samples with total inorganic carbon (TIC) concentrations on the
order of 2.5 mM. Samples with lesser TIC will be overestimated with respect
to ANC, and samples with greater TIC will be underestimated with respect to
ANC if titrated to this pH.
A second problem to note with regard to many historical alkalinity titrations
is that unless detailed notes have been kept of titrations to some endpoint
determined with a colorimetric indicator, it may be impossible to determine
exactly what the pH was at the finish of the titration. For example, the
indicator methyl orange has a pKa of 3.5. The transition range for this
indicator is usually given as pH 4.5 to 3.1 (e.g., Bell 1967, Golterman
1969), over which range the color changes from yellow to orange to pink to
red. Careful analysts prepare standard solutions of known pH to which indi-
cator is added so they can tell by comparison to the sample being titrated
precisely when the titration has reached the pH that they have a priori
selected as the endpoint. Unfortunately, many early titration data are
accompanied by notations only to the effect that "such and such" an indicator
was used. In such cases it may be impossible to determine the endpoint pH of
the titration.
Kramer and Tessier (1982, 1983) reported an experiment in which three
analysts were instructed to perform independent replicate colorimetric
titrations of a low alkalinity sample with methyl orange as the indicator per
instructions given in the 1933 edition of "Standard Methods for the
Examination of Water and Sewage." The endpoint specified for the titrations
was the subjective judgment "until faintest pink coloration." For a total of
24 titrations this judgment corresponded to a pH of 4.04 +_ 0.10 (Kramer and
Tessier 1982, 1983). In such methyl-orange alkalinity titrations, where it
is unambiguously clear that analysts were careful to use the exact method-
ological instructions (i.e., "until faintest pink coloration"), this value of
pH 4.04 may, with appropriate reservations, be used as a reference endpoint
pH. For other less well documented methyl orange titrations, pH 4 may serve
for crude comparative purposes as an estimate of the possible lower limit for
titration endpoint.
4.4.3.1.1.3.2 Alkali ni ty-current methodology—Determi ni ng ANC of
surface water samples is now commonly done by acidimetric titration to the
(HC03~-H+) equivalence point (inflection point) of the titration
curve. This point can be readily determined by using differential elec-
trometric titration methods or the procedure of Gran (1952) (see also Stumm
and Morgan 1981, Butler 1982). ANC determined in this fashion is sometimes
4-62
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termed total inflection point alkalinity or TIP (Dillon et al. 1978, Ontario
Ministry of Environment 1979, Jeffries and Zimmerman 1980).
As discussed previously in Section 4.2.2.3 (and later in Section 4.6.3.2)
organic compounds may contribute significantly to ANC in waters low in total
inorganic carbon and of low pH. This contribution becomes important in the
pH range below that of the (HC03--H+) equivalence point (see Bisogni
and Driscoll 1979, Wilson 1979). Because of this fact it is likely that most
TFE alkalinity titrations fail to measure any possible contribution of
organics to the buffering of natural waters. Gran's procedure in which the
solution is titrated to quite low pH values and total ANC determined by
linear back-extrapolation is able to account for such buffering, should it
exist.
4.4.3.1.1.3.3 Alkalim'ty-comparability of early and current measurement
methods—If the endpoint pH of a titration for ANC is known, then approximate
corrections may be applied to determine the value of ANC that would have
resulted if the titration had been carried to some preselected equivalence
point (e.g., the (HC03--H+) equivalence point for systems with ANC
dominated by inorganic carbon species). Derivations and instructions for
carrying out such corrections have been provided by NRCC (1981) and Henriksen
(1982b). The procedure of Henriksen apparently does not take into account a
correction for bases remaining untitrated at some endpoint pH greater than
the equivalence point pH but this is a very minor concern because, almost
always, the correction to be applied in acidification studies is that due to
overtitration rather than undertitration.
Application of the procedures given by NRCC (1981) and Henriksen (1982b) thus
allow direct comparison of historical ANC values (titration endpoint pH
known) with more recent Gran's titrations or fixed endpoint titrations (again
if endpoint pH is known).
As always, when one compares samples taken years apart care must be taken so
that short-term variability in ANC (e.g., due to snowmelt, rainstorms, uptake
of HCOs" by aquatic plants) will not distort evaluations of long-term
trends.
4.4.3.1.1.4 Sample storage. Kramer and Tessier (1982, 1983) have re-
cently examined the possible importance of container type on the chemistry of
stored water samples. As pointed out by those authors and also by Bacon and
Burch (1940, 1941) soft-glass sample bottles may contribute very significant
amounts of alkalinity (relative to acidification studies regarding dilute
unbuffered waters) to samples stored in such containers. This is especially
true the younger or 'fresher' the bottle and the longer the storage time.
Pyrex brand glass and laboratory-quality plastics, which were introduced and
became popular about 1960, do not contaminate samples in this manner (Kramer
and Tessier 1982, 1983; Schock and Schock 1982). Thus the possibility that
historical samples may have been contaminated during storage (but not during
field measurements; Kramer and Tessier 1982, 1983) contributes uncertainty as
to the overall accuracy of such measurements.
4-63
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4.4.3.1.1.5 Summary of measurement techniques. Each of the three types
of measurements (i. e,pPT^ conductivity, alkalinity) discussed here has
something to recommend its continued use in the study of surface water
acidification. Conductivity seems to be the least informative of the
measurements, but it is likely that historical measures of this variable are
the most accurate and consistent (with current data) of the measures
discussed. Although historical measures of pH are somewhat unreliable, in
comparison to current pH data, a relative wealth of pH measurements exists in
comparison to early data for conductivity and alkalinity. As discussed
above, early measurements of alkalinity are often of little use due to
procedural problems. In addition, they are relatively scarce. Knowledge of
the alkalinity of surface waters and changes in alkalinity with time,
however, are important considerations in the study of acidification.
4.4.3.1.2 Analysis of trends
4.4.3.1.2.1 Introduction. Numerous studies of temporal trends in the
pH, alkalinity, or conductivity of selected North American surface waters
have appeared in the peer reviewed scientific literature or in readily
available technical reports. The following is a brief review of the material
presented in these reports and articles.
In considering each of these studies the critical reader should bear in mind
all of the potential problems of bias (in both sampling and chemical
analysis) that may or may not have been taken into account, reported, and
discussed by the principal investigators. As an example of the kinds of
problems that may exist with regard to unbiased sampling, Figures 4-11 and
4-17 serve to illustrate the kinds of seasonal variations that may occur in
alkalinity and pH at the outlets of Adirondack lakes. Not shown in these
figures are the kinds of shorter term variations that may occur over a day
due to biological activity or the longer term variations that may result from
extended periods of either drought or greater than usual precipitation.
Given the kinds and ranges of variation that occur, it is clear that
significant potential often exists for sampling bias and resulting
misinterpretation of observed temporal "difference" in pH or alkalinity.
This potential is, of course, greatest when data from two discrete points in
time are compared, rather than a more complete time series of data.
Each of the following reviews presents the pertinent information given by the
authors in their original manuscripts. The authors may possess considerably
more information concerning their research than they were able to present in
their original publications. The location and evaluation of such unreported
information is clearly outside of the scope of this review. Only that infor-
mation presented in the original technical report or journal article is re-
viewed here. In some cases the information presented by the original authors
does not demonstrate "beyond a shadow of a doubt" that their sampling or
analyses were completely unbiased, but this does not mean then that their
sampling or analyses were necessarily biased. It is neither the duty nor the
intent of this reviewer to focus unduly on such omissions or to speculate
irresponsibly on presumptions of their importance. Major critical discus-
sions are presented here only on important points of reasonable debate for
which sufficient information was presented by the authors.
4-64
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-pa
I
01
4.0
3.0
_L
j L
LEGEND
PANTHER LAKE
SAGAMORE LAKE
WOODS LAKE
J 1 1 1 1 1 1-
1 1 U
ND|JFMAMJJASOND|JFMAM
1977 1978 1979
Figure 4-17. Temporal trends in pH at outlets of Woods, Panther, and Sagamore Lakes. Adapted from
Galloway et al. (1980b).
-------
For other reviews and commentary of studies of changes in pH and alkalinity
in surface waters and the potential relationship of such changes to acidic
deposition see Haines (1981, 1982), Howells (1982), and Turk (1983).
4.4.3.1.2.2 Canadian studies.
South-Central Ontario (Beamish and Harvey 1972)
Beamish and Harvey (1972) were the first investigators to present evidence of
decreases in lake pH in North America attributable to acidic deposition.
They studied chemistry changes and loss of fish populations in lakes of the
La Cloche Mountains, an area that has quartzite geology and that receives
acidic precipitation. The acidity of the precipitation is directly attri-
butable to smelters at Sudbury, Ontario, 65 km to the northeast. During the
period of their study (1969-71) Beamish and Harvey found the pH of rainwater
ranged from 3.6 to 5.5 and the pH of melted snow ranged from 2.9 to 3.8.
The authors began their study with Lumsden Lake, a small oligotrophic lake in
a watershed devoid of either human habitation or industry. The study was
then expanded to include a total of 150 lakes in the region. For some of
these other lakes earlier data (pre-1968) were available from studies per-
formed by the Ontario Department of Lands and Forests.
In all of the studies, samples were taken between April and November (most
often in August and September). Beamish and Harvey (1972) measured pH in the
field with a Sargent-Welch Model PBL portable pH meter standardized at pH 7.0
and 4.0 before and after each series of readings. Prior to 1970 they
repeated their pH measurements on shore with a Fisher Model 310 expanded-
scale pH meter. All measurements were made promptly in the field to avoid
the kind of pH changes they observed with time (probably due to C02 degas-
sing). In studies prior to 1968 the Ontario Department of Lands and Forests
measured pH with a Hellige comparator (Beamish and Harvey 1972). At the
pre-1968 pH values found by Beamish and Harvey (1972) the Hellige comparator
values apparently agree well with electrometric pH values (e.g., see
Schofield 1982, Norton et al. 1981a, Burns et al. 1981). No other details of
sampling or analytical procedures were given.
Beamish and Harvey (1972) found "little vertical stratification" in pH in
Lumsden Lake and nearby George Lake and only "some seasonal variation."
Their principal finding with regard to lake chemistry was that for lakes in
and to the east of the La Cloche Mountains pH had decreased with time (Table
4-5). For 11 lakes sampled prior to 1961 H+ concentration had increased
10- to 100-fold by 1971. The average annual change in mean pH for all 22
lakes was minus 0.16 unit. The authors found that 26 lakes in a region just
north of the La Cloche Mountains were less acidic and had apparently experi-
enced lesser decreases in pH (Table 4-6). They attributed these facts, at
least partially, to the presence of outcrops of carbonate-bearing rocks in
that area. The authors concluded that "the increases in acidity appear to
result from acid fallout in rain and snow. The largest single source of this
acid was considered to be the sulfur dioxide emitted by the metal smelters of
Sudbury, Ont." (Beamish and Harvey 1972).
4-66
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South-Central Ontario (Beamish et al. 1975).
Beamish et al (1975) reported on the relationship between various fish popu-
lations and water chemistry in George Lake, Ontario, for the period 1967-73.
In that report they cited evidence for a trend of pH decrease in the lake.
Over the period 1968 to 1973 they measured pH electrometrically in the field
or in the laboratory within 12 hours of sampling. From a regression of 28
such measurements plus one measurement "using a dye indicator method" in 1961
they arrived at a linear decline in lake pH of 0.13 unit per year, on the
average. The correlation coefficient for this regression was 0.85. Dis-
carding the 1961 data point [apparently done by Hellige Kit, see Beamish and
Harvey (1972) and discussion above], they arrived at a linear mean annual
decline of 0.13 with a correlation coefficient of 0.65 (Beamish et al. 1975).
In their report they provided no other details of their sampling methods or
analytical procedures.
South-Central Ontario (Dillon et al. 1978).
As part of a study on the effects of acidic deposition on lakes in south-
central Ontario, Dillon et al. (1978) collected alkalinity data for four
lakes for which some historical data existed. These lakes were Walker Lake,
Clear Lake, Harp Lake, and Jerry Lake. Precipitation in the region has a
mean pH between 3.95 and 4.38.
The authors sampled Clear Lake three times in the period June-August 1977 and
found TIP alkalinities ranging from 2 to 25 (yeq n-l). This was a
decrease from a TIP alkalinity of 33 (peg £-!) reported for the year
1967 by Schindler and Nighswander (1970).
Dillon et al. (1978) reported TFE alkalinities (measured potentiometrically
to pH 4.5) of 153 (weq &'1) for the epilimnion and 130 (yeq i "M
during a non-stratified period for Walker Lake in 1976. These were decreases
from TFE values of approximately 180 (yeq 5,"1) during 1974 (from unpub-
lished data of the Ontario Ministry of the Environment) and approximately 400
(ueq jr1) during 1971 (several samples on a single date; Michalski
The authors did not find any noticeable differences between the TFE alkalin-
ities of Harp Lake (137 to 152 peq J^1) or Jerry Lake (137 to 168 yeq
A"1) in 1978 and earlier values reported by Nicholls (1976).
Dillon et al. (1978) discussed in detail their analytical methodology but did
not give any details of their sampling procedures or any information on pos-
sible short-term variations in alkalinity.
Halifax, Nova Scotia (Watt et al. 1979).
Gorham (1975) reported on the chemistry of 23 lakes near Halifax, Nova
Scotia, sampled in December 1955. Twenty-one years and two weeks later, Watt
et al. (1979) attempted to sample these same lakes to look for water
chemistry changes that may be associated with sulfur emissions from
4-67
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TABLE 4-5. EARLIEST AND 1971 pH MEASUREMENTS ON LAKES IN AND TO THE
EAST OF THE LA CLOCHE MOUNTAINS (BEAMISH AND HARVEY 1972)
Lake
Brokers
Carlyle
David
Free! and
George
Greya
Johnnie
Kakakise
KUlarney
L & F 24
Lumsden
Lumsden II
Lumsden III
Mahzenazing9
Nellie
Norway
O.S.A
Spoon9
Sunfisha
Threenarrows
Township
Attl ee
Carlyle
Stalin and Goschen
Killarney
Killarney
Sale
Goschen and Carlyle
Killarney
Killarney
Carlyle
Killarney
Killarney
Killarney
Carlyle and Humboldt
Roosevelt
Killarney
Killarney
Kil patrick and
Humboldt
Humbol dt
Killarney, Roosevelt,
and Stalin
Date
Sept/6lb
Aug/71
May/68b
Aug/71
Aug/61b
Aug/71
June/69
Sept/71
Sept/61b
Sept/71
Sept/59b
Sept/71
Aug/6lb
Aug/71
June/68b
Aug/71
Aug/69
Sept/71
Sept/67b
Aug/71
Sept/61b
Aug/71
June/69
Oct/71
June 69
Oct/71
Sept/6 lb
Aug/71
Sept/69
Aug/71
Sept/69b
Aug/71
/61b
Sept/71
Sept/61b
Aug/71
Sept/61b
Apr/71
Nov/69b
Aug/71
PH
6.8
4.7
5.5
5.1
5.2
4.3
5.2
4.8
6.5
4.7
5.6
4.1
6.8
4.8
6.0
5.7
4.5
4.4
6.0
5.0
6.8
4.4
4.6
4.0
4.6
4.0
6.8
5.3
4.5
4.4
4.5
4.5
5.6
4.3
6.8
5.5
6.8
4.4
5.2
5.2
Avg
annual
change in
pH units
-0.21
-0.13
-0.09
-0.20
-0.18
-0.13
-0.20
-0.10
-0.05
-0.25
-0.24
-0.30
-0.30
-0.15
-0.05
0.00
-0.12
-0.12
-0.24
0.00
4-68
-------
TABLE 4-5. CONTINUED
Avg
annual
change in
Lake Township Date pH pH units
Tysona Unnamed Sale and Humboldt Aug/55b 7.4 -0.16
Lake3
(46001'30"NSr24'W) Killarney June/69 5.7 -0.25
Oct/71 5.2
Mean of 22 lakes -0.16
aLocated east of the La Cloche Mountains.
determined by the Ontario Department of Lands and Forests.
4-69
-------
TABLE 4-6. EARLIEST AND 1971 pH MEASUREMENTS ON LAKES NORTH OF THE
LA CLOCHE MOUNTAINS (BEAMISH AND HARVEY 1972)
Lake
Anderson
Annie
Bear
Brazil
Deerhound
Elizabeth
Frank
Fox
Griffin
Hannah
Hanwood
Lang
Leech
Little Bear
Little Hannah
Little Panache
Long
Loon
Plunge
St. Leonard
Township
Merritt
Bevin and Sale
Roosevelt and Dieppe
Foster
Curtin
Foster
Goschen
Goschen
Merrit
Foster, Truman,
Curtin, and Roosevelt
Roosevelt
Curtin
Roosevelt
Roosevelt
Truman
Louise and Dieppe
Eden, Waters, and
Broder
Merritt and Foster
Roosevelt
Foster
Date
Aug/60a
Oct/71
/61a
Aug/71
Aug/68a
Aug/71
Aug/67a
Aug/71
Sept/68a
Aug/71
Sept/68a
Sept/71
/60a
Oct/71
July/60a
Sept/71
Aug/60a
Oct/71
Aug/68a
Aug/71
Aug/67a
Oct/71
Aug/68a
Oct/71
Aug/67*
Oct/71
Aug/68a
Oct/71
Aug/68a
Aug/71
July/68
May/70
Nov/69a
Sept/71
Sept/68a
Oct/71
Aug/68a
Oct/71
Sept/68a
Aug/71
pH
7.4
6.4
5.6
4.7
6.5
6.3
7.5
6.7
7.0
6.7
6.5
7.5
6.9
5.6
6.1
5.3
7.8
6.7
7.0
6.7
7.0
6.0
6.5
6.8
6.5
6.0
6.5
5.7
7.5
6.5
8.5
7.8
6.5
6.8
6.5
6.5
6.6
6.0
6.8
6.7
Avg
annual
change in
pH units
-0.09
-0.09
-0.07
-0.20
-0.10
+0.33
-0.03
-0.07
-0.10
-0.10
-0.25
+0.10
-0.13
-0.27
-0.33
-0.35
+0.15
0.00
-0.20
-0.03
4-70
-------
TABLE 4-6. CONTINUED
Lake
Simon
Spring
Stratton
Walker
Whitefish
White Oak
Mean of 26 lakes
Township
Graham
Merritt
Foster
Truman and Roosevelt
Whitefish Indian
Reserve
Til ton and Halifax
Date
Aug/60a
Sept/71
Aug/66a
Oct/71
Sept/68a
Aug/71
Aug/68a
Aug/71
Aug/60a
Oct/71
Nov/693
Oct/71
PH
6.1
6.4
7.0
6.2
7.0
6.7
6.5
6.3
6.3
6.4
4.2
4.1
Avg.
annual
change in
pH units
+0.03
-0.16
-0.10
-0.07
+0.01
-0.05
-0.08
apH determined by the Ontario Department of Lands and Forests.
4-71
-------
industrial sources near Halifax. They found one lake to be filled, one to be
inaccessible, and five to have significant local disturbances—leaving 16
lakes to be compared to the 23 studied by Gorham.
Watt et al. (1979) took considerable care to sample in the manner Gorham
(1957) used. They measured pH with a Fisher Accumet Model 230 pH meter
before and after sample C02 equilibration with the laboratory atmosphere
and stated that "since both studies used glass-electrode pH meters, the
combined error for the pH differences should be less than +_0.07" (Watt et
al. 1979). They also measured specific conductivity, alkalinity and acidity,
even though the last two variables were not determined by Gorham (1957).
Watt et al. (1979) performed variance analysis on the samples from the 16
lakes and found that pH differences associated with geology had not changed
since the study by Gorham (1957) but that pH values of the lakes did differ
significantly from those found in 1955. They found current pH values from
3.89 to 6.17 (before air equilibration). In 1955, pH values in these lakes
ranged from 3.95 to 6.70 (before air equilibration) (Gorham 1957). Watt et
al. (1979) plotted 1977 pH values vs 1955 pH values (Figure 4-18) and found
that all points were below the 1:1 line, that the pH drop was significant to
the p < 0.001 level, and that the slope was significantly less than one (p <
0.001). They also found that conductivity in the lakes increased signifi-
cantly (p < 0.001) over the 21-year period. The authors reported that recent
pH data from other Nova Scotia lakes and from lakes in New Brunswick and on
Prince Edward Island, when compared with data reported by Hayes and Anthony
(1958), tend to confirm a trend towards lake acidification in these areas.
Watt et al. (1979) did not measure precipitation pH but did note that mean
sulfur emissions from the Halifax metropolitan area were approximately double
in 1977 the amount they were in 1955. The authors concluded that it was
"clearly unnecessary to look beyond local sources (i.e., to long-range atmos-
pheric transport) for an explanation of the acidic condition of lakes in the
Halifax areafl (Watt et al. 1979).
Nova Scotia and Newfoundland (Thompson et al. 1980).
V
Thompson et al. (1980) reported temporal trends in the pH of Nova Scotia and
Newfoundland rivers. In their report they discussed data given by Thomas
(1960) for the years 1954-56 and more recent data reported by the Water
Quality Branch of Environment Canada. The more recent data are stored in the
data archive NAQUADAT.
Three Nova Scotia rivers were studied—the Tusket River, the Medway River,
and the St. Mary's River. Samples were taken approximately monthly in 1955
(Thomas 1960) and in the years 1965-74. Samples were kept tightly stoppered
in the dark, and "the pH's used for comparison were measured in the labora-
tory, at room temperature" (Thompson et al. 1980). Thompson et al. (1980)
compared the discharges on days of sampling to mean annual discharges and
concluded that "although sampling in various years was commonly biased toward
either high or low flow, there was no consistent relationship between mean pH
and such bias ... the calculated pH's are reasonable, representative and
comparable." No other information was provided on sampling or analysis. The
4-72
-------
7.0
6.0
? 5.0
Q.
4.0
3.0
I
3.0
4.0
5.0
pH 1955
6.0
7.0
Figure 4-18.
Relationship between pH values for 16 lakes (near Halifax,
Nova Scotia) in 1977 and 1955. Dashed line is line of no
change; all values are below this line and drop in pH is
significant to p < 0.001 level. Slope of least-squares
equation (solid line) is significantly less than that of
dashed line (p < 0.001) indicating greater pH declines in
in higher pH lakes. Adapted from Watt et al. (1979).
4-73
-------
value of discharge-weighted mean pH of the rivers decreased from roughly 5.2
to 4.4 (Tusket River), 5.7 to 4.9 (Medway River), and 6.2 to 5.5 (St. Mary's
River).
The three Newfoundland rivers studied were the Isle Aux Morts River, the
Garnish River, and the Rocky River. Sampling and analysis were as for the
Nova Scotia rivers. Although plots of discharge weighted mean annual pH of
these rivers over the period 1971-78 appear quite variable, the authors
believe that these data together with the data for the Nova Scotia rivers
indicate a general steady decrease in pH until 1973 and a steady increase
afterwards. The increase is apparently attributed to decreased acid loading
to the Atlantic Provinces since 1973 presumably because of changed weather
patterns" (Thompson et al. 1980). The authors presented no appropriate
statistical evidence in support of any of the "apparent" trends.
4.4.3.1.2.3 United States studies.
New England (Maine) (Davis et al. 1978).
Davis et al. (1978) studied 1936 pH readings taken from 1368 Maine lakes
during the period 1937-74 in an effort to see if they could find pH decreases
associated with the acidic precipitation of that area (4.4 < pH < 5.0 since
at least 1956; Cogbill 1976; Likens 1976). Samples and data were from a
variety of sources (Davis et al. 1978) but apparently most samples were taken
over the deepest portion of each lake,-jje'ar mid-day, during the summer.
"Until the 1960's" pH was measured using a Pennwalt colorimetric kit. After
that time pH was measured with portable meters. "The two methods were found
to agree within 0.1 pH units" (Davis et al. 1978).
The authors noted initially that the mean pH of 296 samples from 1937-42 was
6.81 and that the mean for 289 samples from 1969-74 was 6.09—a 5.2-fold
acidity increase. They also noted that most of the change seemed to occur in
the early 1950's and that overall the change might have been greater if it
had not been for some cultural euthrophication beginning in the 1950's. The
authors realized that these preliminary results might have been affected by
regional edaphic differences in lake types and also by differences in
precipitation acidity across the state. Amounts and seasonal patterns of
precipitation also may have played a part (Davis et al. 1978). In an attempt
to minimize such potential regional distortions, they analyzed the data by
using three procedures based on H+ concentration changes in individual
lakes.
They found 258 lakes had pH readings separated by at least a year. There was
a mean of 2.9 readings per lake and a mean of 12.7 years between successive
readings (pairs) for a total of 376 pairs during the period 1937-74.
Procedure I of Davis et al. (1978) was as follows. They used data pairs to
calculate slopes (H+ concentration vs time) for individual lakes and then
mean slopes from 1937-74. The mean slopes were added to obtain a total H+
concentration change for the entire period. Given a starting pH of 6.89
(mean of 123 values 1937-42), the final (1974) pH would be 5.79, an increase
in acidity of 12.6 times. Using a t-test, the authors also found that the
4-74
-------
mean annual increase in H+ concentration based on the mean slopes for each
year was significantly different from zero change with p < 0.0001. The
authors noted, however, that this procedure more strongly weights data pairs
with long time separations, thus possibly invalidating the use of a t-test.
The second procedure Davis et al. (1978) used was to average the 376 single
slope values. This gave a mean of 0.115 peq £-1 yr"1 H+ concen-
tration change. By t-test, this mean is significantly different from zero at
p < 0.1, but not at p < 0.05. If a disproportionately greater decrease in pH
occurred in the 1950's (as the authors hypothesized), this procedure would
give greater weighting to the more frequent data pairs beginning about that
time and would thus overestimate total change (Davis et al. 1978).
Procedure III the authors used was to weight each data pair (H+ concen-
tration) slope linearly in inverse proportion to the time interval between
each reading. These weighted slopes were then averaged for each year that
they applied. Using an initial pH of 6.89 in 1937, the authors noted that pH
decreased by 1950 to only 6.83. By 1961, however, the pH had decreased to
5.91, so 73 percent of the increase in acidity occurred in this latter time
period. The authors believed that this 73 percent increase in acidity was
actually an underestimate for this time period.
Davis et al. (1978) also discussed some alkalinity data they had for 44 of
the 258 lakes cited above. These data were from the period 1939-71, a total
of 96 values and 52 pairs. No information was given on the analytical
method(s) used to determine alkalinity. Applying their Procedure I to those
data, they obtained a decrease of about 6.34 ppm (as CaC03; from 11.82 to
5.48 ppm, typically; corresponding to a decrease of 127 veq £~1 from
236 to 109 yeq jr*) over the period. This was much less than expected
from pH changes from the same period and from observed relationships between
pH and alkalinity. The authors noted that "the discrepancy may be due in
large part to the inadequate sampling and great variance of the alkalinity
data, including the fact that 67 percent of the pairs had their initial
member in 1960 or later" (Davis et al. 1978).
The authors concluded from their study that between the years 1937-74 H+
concentration in Maine lakes increased about 1 yeq &"1 and pH decreased
from about 6.85 to 5.95. Further, nearly three-quarters of this change
occurred in the 1950's. "This is the first demonstration of a pH decrease
due to acidic precipitation on a large region of lowland lakes in the United
States" (Davis et al. 1978).
New England (Maine, New Hampshire, Vermont) (Norton et al. 1981a).
Norton et al. (1981a) measured pH in 94 New England lakes (82 in Maine, 8 in
New Hampshire, 4 in Vermont) for which historical pH existed from the period
1939-46. The lakes sampled were small, oligotrophic-mesotrophic, and located
in forested areas on non-calcareous bedrock. The recent sampling (1978-80)
was done during July-October but not on the same monthly dates as the
historic sampling. These samples were collected at 1 m depths, and the lakes
were stratified at the time of sampling.
4-75
-------
The pH values of the recent samples were measured in the field with (1) a
portable pH meter with combination electrode, and (2) a Hellige color
comparitor. Historical pH values were obtained using a Hellige color
comparator. Except for three spurious cases of low pH lakes, the authors
found that "reasonable agreement exists for these two methods, especially at
higher pH's" (Norton et al. 1981a).
The authors presented their results in plots of (1) old colorimetric pH vs
recent colorimetric pH, and (2) recent colorimetric pH vs recent electro-
metric pH (Figures 4-16 and 4-19). They concluded that, qualitatively, their
study "confirms the results of Davis et al. (1978) regarding an overall
decrease in the pH of Maine lakes" (Norton et al. 1981a).
New England (Maine, New Hampshire, Vermont, Connecticut, Massachusetts, Rhode
Island) (Haines and Akielaszek 1983)~
Haines and Akielaszek (1983) recently surveyed the chemistry of 226 headwater
lakes and low order streams in the six New England states. The waters sam-
pled were low in color and were, for the most part, free from human disturb-
ance. Most of the sampling took place from mid summer to early winter of
1980.
For 95 of the lakes sampled historical (1938-78) data exist for pH. Most of
these data (66 of 95 values) predate 1960. For 56 of the lakes historical
data exist for alkalinity (38 values predating 1960). Colorimetric proce-
dures were used to determine pH for all but one (electrometric) of the
historical values. A portable pH meter with gel combination electrode was
used for the recent survey. Historical alkalinity was determined by agencies
of the six states by acidimetric titration to some methyl orange endpoint (pH
unspecified). Haines and Akielaszek (1983) used both a fixed endpoint proce-
dure (pH 4.5 determined electrometrically) and the procedure of Gran (1952)
to determine alkalinity for their survey samples.
The mean pH of the historical samples was 6.07 (mean H+ 0.8 yea JT1) and the
mean pH of the recent samples was 5.37 (mean H+ 4.3 peq I ). By paired
t-test (t = 4.17, p < 0.0001), the recent pH values were significantly lower
than the historical values (Figure 4-20) (Haines and Akielaszek 1983).
The mean alkalinity of the historical samples was 198 yeq i"1 and the
mean alkalinity of the recent survey samples was 68 y eq £~i. Employ-
ing the assumption that the methyl orange titration endpoint roughly coinj
cided with a pH of 4.5 Haines and Akielaszek subtracted 32 yeq £
from the historical alkalinity data and then compared the .mean of the
adjusted data (166 yeq JT1) to the mean (68 yeq SL~L) of their
survey data (Figure 4-21). By paired t-test (t = 4.03, p = 0.002) the
decrease was significant. If a worst case estimate of methyl orange endpoint
of pH 4 is assumed (see Section 4.4.3.1.1.3.1) for all historical, data then
the decrease would be less (from approximately 98 u eq £"i histor-
ically to 68 yeq £"1) than calculated by the authors. Still, how-
ever, there would be evidence of a decrease in alkalinity and this would be
qualitatively consistent with the observed decrease in pH.
4-76
-------
8.5
8.0
3 7.5
o;
7.0
ce:
o
o
0 6.5
6.0
5.5
I
4.5 5.0 5.5 6.0 6.5 7.0 7.5
NEW COLORIMETRIC (pH)
Figure 4-19. Old lake water pH (colorimetnc) vs recent lake water pH
(colon'metric). Adapted from Norton et al. (1981a).
4-77
-------
8
-p.
CO
GL
at
J I I I I I I
Figure 4-20.
HISTORICAL pH
Current vs historical pH for 95 New England lakes
Adapted from Haines and Akielaszek (1983).
Solid line indicates equivalent pH,
-------
CT
Ol
UJ
a:
o:
100 -
0
100
200
300
400
500
600
HISTORICAL ALKALINITY (yeq jf1)
Figure 4-21.
Current vs historical alkalinity for 56 New England lakes. Solid line indicates
equivalent alkalinity. Adapted from Haines and Akielaszek (1983).
-------
New England (New Hampshire) (Hendrey et al. 1980b, Burns et al. 1981)
During 1936-39 the Mew Hampshire Department of Fish and Game conducted a
biological survey of waters in the White Mountains of that state. Their
survey included measurement of pH of headwater streams and measurement of
alkalinity and pH for small lakes. In 1979 Burns et al. (1981) resampled 38
of these waters and made determinations of alkalinity and pH (note: the data
for this study were also presented and discussed by Hendrey et al. 1980b).
Since at least 1955-56 this area has been receiving precipitation with a
weighted annual pH less than 4.5 (Cogbill and Likens 1974).
The sampling rationale and analytical methodology used by Burns et al. (1981)
were exactly the same as used in their study of North Carolina streams. A
detailed discussion of these methods is presented in that section of this
review.
Burns et al. (1981) found that 90 percent of the 38 samples showed a decrease
in pH between the late 1930's and 1979 (mean pH 6.66 in 1936-39 and mean pH
6.06 in 1979). Mean H+ concentration was 0.22 Ueq n~l) in 1936-39
and 0.87 (yeq r1) in the 1979 samples. A t-test showed this increase
in H+ to be significant at the p < 0.02. "However, when the errors asso-
ciated with comparing the colorimetric data to the electrometric data are
considered, the difference in pH between the 1960's (sic—the authors meant
1930's, Burns pers. comm.) and 1979 may not be significant" (Burns et al.
1981). The authors had historical alkalinity values for only five lakes in
New Hampshire. Alkalinity decreased at all five sites (mean decrease 103
percent of original), but the authors noted that there were not enough
samples to make a valid statistical comparison. (See also the review of the
North Carolina study by the same authors for a critical discussion of
comparison of their alkalinity values with historical measurements.)
New York (Schofield 1976a,b).
Schofield (1976a,b) reported on a 1975 survey of water chemistry and fish
status of 217 Adirondack lakes located at elevations greater than 610 m. For
40 of these lakes, pH data exist from the period 1929-37. Frequency
distribution plots (Figure 4-22) of lake pH for the two data sets illustrate
the apparent pH decrease with time (Schofield 1976b). During the period
September 5, 1974-April 9, 1975 the weighted mean pH of precipitation on this
area on a storm-by-storm basis was 4.23 (range 3.94 to 4.83) (Scnoneld
1976c). Schofield (1976a) presented a complete discussion of sampling and
analytical methods for the 1975 survey. Schofield (1976b) did not present
any information on sampling or analytical methodology for pH for the 1929-37
data, but did state that the two sets were "comparable".
New York (Pfeiffer and Festa 1980).
In the summer of 1979 the New York Bureau of Fisheries Lake Acidification
Studies Unit sampled 396 ponded Adirondack waters. For 138 of these waters
historical pH data from the period 1930-34 existed. As part of their report
on the acidity status of Adirondack lakes, Pfeiffer and Festa (1980) compared
the pH values of these lakes in 1979 to the values of the period 1930-34.
4-80
-------
20
10 -
1930's
8
20
10
1975
8
NO FISH PRESENT
FISH PRESENT
Figure 4-22.
Frequency distribution of pH fish population status in 40
Adirondack lakes greater than 610 m elevation, surveyed
during the period 1929-37 and again in 1975. Adapted from
Schofield (1976b).
4-81
-------
The 1979 sampling was done via helicopter and samples were taken at a depth
of 1 m. No information was given on the sampling during the period 1930-34.
For the samples taken in 1979, pH was determined in the laboratory, using
both a pH meter and a Hellige colorimetric comparitor. These determinations
were made on the samples after each sample had been equilibrated with the
laboratory atmosphere. The only information given on the pH determinations
of the 1930-34 samples was that the measurements were made using a Hellige
comparitor.
Pfeiffer and Festa (1980) reported that their colorimetric and electrometric
measurements on the samples taken in 1979 disagreed markedly and that the
Hellige comparitor consistently overestimated pH throughout the range of
sample values and especially drastically at the lower values. Schofield
(1982) compared Hellige comparitor measurements to pH meter measurements for
similar samples, concluding that agreement between the two methods was much
better than found by Pfeiffer and Festa (1980) and that the discrepancies
found by these authors were due to "errors in pH meter measurements."
To minimize any potential bias in the comparison of pH measurements over
time, Pfeiffer and Festa (1980) used only colorimetric measurements in their
data analysis. They presented their results graphically (Figure 4-23). They
concluded that "historic readings obtained in the 1930's were generally
higher than comparable current determinations for the same group of waters.
This reflects a general deterioration of water quality during the 40-year
time frame between samplings" (Pfeiffer and Festa 1980). The authors
attributed the observed deterioration of water quality to the acidic
precipitation in the region.
Two additional comments may be made on the data presented by Pfeiffer and
Festa (1980). First, the parallel trend shown in Figure 4-23 is curious when
one considers that buffer intensity varies non-1inearly as a function of pH
(Stumm and Morgan 1981). Perhaps a more careful analysis or plotting of the
data would show the expected effects.
Second, the distribution of 1979 pH values reported by Pfeiffer and Festa
(1980) differs markedly from that shown by Schofield (1976b) for 1975. This
is because Schofield (1976b) was interested in examining changes in lakes
higher in elevation (i.e., relatively more sensitive to acidic deposition)
and, as he clearly noted, he chose his data set accordingly (Schofield
1976b).
New Jersey (A. H. Johnson 1979).
Searching for evidence of temporal trends, A. H. Johnson (1979) examined 17
years of pH data for two small headwater streams (McDonalds Branch and Oyster
Creek) in the New Jersey Pine Barrens. Precipitation in the area had a mean
pH of 4.4 in 1970, 4.25 for seven months in 1971, and 3.9 from May 1978 to
April 1979. Nearly all of the data for the study came from two sources:
U.S. Geological Survey sampling and analyses from 1963-78 and a University of
Pennsylvania trace metal study in 1978-79. The USGS samples were collected
randomly with a frequency of 2 to 12 per year. This sampling was not biased
seasonally for McDonalds Branch but was slightly biased consistently
4-82
-------
TOO
Q.
CD
o
CO
UJ
_J
D_
<
u_
o
a*
40 -
20 -
5.5
6.0 6.5 7.0 7.5
DETERMINED COLORIMETRICALLY (pH)
8.5
Figure 4-23. Cumulative comparison of historic and recent pH values for a
set of 138 Adirondack lakes. Adapted from Pfeiffer and
Festa (1980).
4-83
-------
throughout the study towards a greater representation of spring samples for
Oyster Creek. The University of Pennsylvania samples were collected weekly
in McDonalds Branch only from 1978 through 1979. Johnson presented little
information on sample pH analyses except that "all pH values were measured
with a glass electrode."
Johnson (1979) had varying levels of confidence in the pH data. Those data
he considered most reliable were from samples on which cations balanced
anions within 15 percent and calculated conductance balanced measured
conductance within 15 percent. He performed regressions of stream pH vs time
for different groups of data (Table 4-7 and Figure 4-24) and found for most
groups that a significant decrease existed. Johnson noted no evidence that
oxidation of geological sulfides, changes in land use, or changes in the
amount of precipitation were responsible for the long-term trends. He
concluded "it appears that the decrease in stream pH is a real phenomenon and
not attributable to differences or bias in sampling or measurement. The data
collected to date are consistent with the postulation of an atmospheric
source for the increased H+."
Pennsylvania (Arnold et al. 1980).
In an effort to assess temporal changes in pH and alkalinity of Pennsylvania
surface waters, Arnold et al. (1980) examined five existing water quality
data bases. Nearly all of the data examined were from streams. Arnold et
al. found 314 instances where data were taken at least one year apart at the
same location or "sufficiently close (generally within one mile with no major
tributaries or influences between)." Of these 314 cases, 107 (34 percent)
showed decreases in pH, alkalinity, or both. The mean pH of the "earliest"
of these 107 cases of decrease was 7.31 (range 5.8 to 8.8), whereas the mean
pH of the "most recent" was 6.94 (range 4.9 to 8.3). The mean change in pH
was a decrease of 0.37 unit, and the range of change was -1.3 to +0.2 units.
For alkalinity, the mean of the "earliest" samples was 834 (yep JT1)
(range 100 to 4000 yeq £-!), and the mean of the "most recent" was 532
(yeq r1) (range 40 to 3720 yeq jr1). The mean net change was a
decrease of 302 (yeq £-1) and the range was (-2100 to +360 yeg
£~1). The average time span between the "earliest" and "most recent1
samples was 8 1/2 years; the range was 1 to 27 years. Arnold et al. (1980)
concluded that "although the data upon which this report is based are not
sufficiently strong to define statistically valid relationships, it seems
clear that there is a definite overall trend toward increasing acidity in
many Pennsylvania streams ...."
Although the authors presented and discussed the means and ranges of pH and
alkalinity decreases for those cases where decreases were found (34 percent
of the total), they did not present or discuss the overall changes for the
314 total cases examined. If 34 percent of the total cases decreased, then
66 percent must have remained the same or increased. This, plus the fact
that five separate data bases were used, that very little information was
presented concerning sampling, and that no information was presented about
analytical procedures gives rise to some serious questions concerning this
study. Also of concern is the fact that decreases over a period as short as
one year are considered part of a "definite overall trend" (Arnold et al.
4-84
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TABLE 4-7. REGRESSIONS OF STREAM pH ON TIME: N IS THE NUMBER OF SAMPLES,
r IS THE CORRELATION COEFFICIENT, AND P IS THE LEVEL OF SIGNIFICANCE;
ao AND ai ARE COEFFICIENTS IN THE REGRESSION pH = ao + aix,
WHERE x IS THE NUMBER OF MONTHS AFTER JUNE 1963 (A. H. JOHNSON 1979)
Data source N ao
ai r P
A yeq H+
per liter
(1963-
1978)
USGS data, 1963-78
USGS data + UP dataa
USGS data, am'on equiva-
lents balance cation
equivalents; measured
and calculated specific
conductances are equal
All USGS data
USGS data, anion equiva-
lents balance cation
equivalents; measured
and calculated specific
conductances are equal
McDonalds Branch,
New Jersey Pine
Barrens
90 4.42 -0.0022
100 4.49 -0.0030
36 4.35 -0.0012
Oyster Creek,
New Jersey Pine
Barrens
78 5.10 -0.0047
26 4.89 -0.0027
-0.22 0.05
-0.32 0.01
-0.29 nsb
-0.56 0.01
-0.53 0.01
+57
+80
+29
+48
+26
aincludes all data collected by the U.S. Geological Survey (USGS) from
1958 to 1978 and the monthly average pH of University of Pennsylvania
(UP) samples.
bNot significant.
4-85
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I
OYSTER CREEK
A
Q
O
.:•,
MCDONALDS BRANCH
1960
1970
1980
Figure 4-24.
Stream pH 1979. Closed circles represent samples in which
anion and cation equivalents balanced and calculated and
measured specific conductances were equal. Open circles
are samples for which the chemical analyses were incomplete
or for which discrepancies in anion and cation and con-
ductivity balances could not be attributed to errors in pH.
The closed triangle represents the average pH determined in
a branch of Oyster Creek in a 1963 study. Open triangles
are monthly means of pH data collected weekly from May 1978
to January 1979 during a University of Pennsylvania trace
metal study. Adapted from A. H. Johnson (1979).
4-86
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1980). Yet another consideration in studies such as this has been noted by
Schofield (1982); "it is obvious that detection of significant, long-term pH
changes in acidifying systems, still in a bicarbonate buffered state, cannot
be made reliably because normal metabolism induced changes in C02 levels
would likely obscure any pH change resulting from decreased alkalinity. Thus
interpretations of long-term pH changes in the range of 6-7 must be viewed
with caution." Given the possible variability (not to mention potential
bias) in data taken from a variety of sources (perhaps arrived at by a vari-
ety of procedures), the mean decreases in pH and alkalinity of only those
cases that did decrease in the study seem not so profound. They may, indeed,
only represent inherent scatter in such a data set. To cite these as evi-
dence of a "definite overall trend" (Arnold et al. 1980) seems premature.
North Carolina (Hendrey et al. 1980b, Burns et al. 1981).
In the period 1961-64 the North Carolina Division of Inland Fisheries meas-
ured the pH and alkalinity of a number of North Carolina headwater mountain
streams. Burns et al. (1981) resampled 38 of these streams in 1979, attempt-
ing to discern any changes in stream chemistry that might have occurred in
association with the acidic precipitation that falls in the area (weighted
annual pH 4.7 to 5.2 in 1955-56 and < 4.5 in 1979). The data discussed by
Burns et al. (1981) were also presented and discussed by Hendrey et al.
(1980b).
Burns et al. (1981) used detailed maps to resample at exactly the locations
of the original samples. The authors considered the possible sampling bias
inherent in representing by a single sample the chemistry of a stream "where
pH could fluctuate daily as well as seasonally. It was assumed that daily
and seasonal fluctuations were random and normally distributed if the new
samples were taken during the day and at the same time of year as the pre-
vious ones."
For the historical samples (1961-1964) pH was measured with a Hellige color-
imetric kit and for the recent samples (1979) pH was determined electro-
metrical ly. The authors compared pH measurements by Hellige kit to those
with their pH meter and found agreement within +_ 0.15 pH unit (Burns et al.
1981). The authors did not find a significant temporal trend in pH (mean
6.77 in 1961-64 and mean 6.51 in 1979).
Alkalinity was determined in the historical studies by acidimetric titration
with methyl orange as the indicator. No endpoint pH was given by the
authors. In the recent study the procedure of Gran (1952) was used to
determine the alkalinity. From the historical titrations to methyl orange
endpoint the mean alkalinity was determined to be 146 y eq £ . Using
this value and the equations given by Stumm and Morgan (1981) (see Section
4.4.3.1.1.3.1, Equations 4-14 to 4-16) it can be determined that the true
equivalence point pH is at least as-great as 5.1. Thus, some overtitration
of the historical samples occurred. To correct for overtitration the authors
subtracted 32 yeq a~l from each of the historical values. This
correction assumes that the actual titration endpoint was pH 4.5. If, in
fact, the methyl orange titration endpoint was as low as pH 4 (see Section
4.4.3.1.1.3.1 and Kramer and Tessier 1982, 1983) then the correction should
4-87
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be on the order of 92 yeq r*1. This would Indicate that historical
alkalinity values had a mean of approximately 54 yeq i~l compared to
the mean (by Gran's method) in 1979 of 80 yeq £~1. In conclusion,
the uncertainty in the endpoint pH of the historical alkalinity determi-
nations casts doubt upon the findings of the authors that "the decrease in
alkalinity between the 1960's and 1979 was statistically significant at the
0.02 probability level using a t-test" (Burns et al. 1981).
Florida (Crisman et al. 1980).
Crisman et al. (1980) reported pH changes in 13 poorly buffered oligotrophic
lakes (known as the Trail Ridge lakes) in northern Florida. They monitored
the lakes quarterly (1978-79) and found a mean annual pH of 4.98. The mean
annual precipitation pH at the time of the study was 4.58. "Comparison of
the present data with that collected over the past 20 years indicates that
the mean pH of the Trail Ridge lakes has declined an average of 0.5 pH units
(sic) since 1960" (Crisman et al. 1980). The authors neither presented
further information on their sampling or analytical methods for pH, nor did
they present any historical data or their sources for such data.
Colorado (Lewis 1982).
In an effort to examine the possible effects of deposition (bulk precipi-
tation of less than pH 5.0 for 89 percent of the weeks during the interval
June 1979-80) Lewis (1982) compared data taken in 1974 with data from 1938-42
and 1949-52 for 64 lakes in the Colorado Rockies. Historical data were takei
by Pennak and consisted of 152 samples analyzed for pH, alkalinity and res-
idue. Historical pH was determined with colorimetric indicators and alkalin-
ity by methyl orange titration.
Lewis (1982) sampled each of the lakes in 1979 on the same day of the year as
did Pennak originally. For the 1979 data pH was determined electrometrically
and Lewis (1982) noted "the meter was checked against Pennak1s indicator
method and the two found to be in good agreement . . .". Alkalinity was
determined by titration with 1/44 N HCl (as with the historical data) "but
the endpoint (4.4) was determined electrometrically with the pH meter rather
than with the methyl orange indicator used by Pennak" (Lewis 1982). This
last statement seems to indicate that the original alkalinity titrations were
performed to an endpoint pH of 4.4 (same as the 1979 titrations) but this is
never explicitly stated by the author (Lewis 1982).
Not all of the data were used in the comparison presented. Lakes of eleva-
tion below 2000 m were omitted as well as any lakes with changes in alka-
linity or total residue > 60 percent or changes in pH > 1.5 units. Changes
of such magnitude were judged to be "evidence of the operation of factors
other than precipitation chemistry" (Lewis 1982).
Results of the analysis of the remaining data are shown in Table 4-8. The
mean alkalinity change is roughly -97 yeq £ . Lewis (1982) analyzed
runoff patterns for the year 1979 and concluded that roughly 5 percent of the
22 percent decrease in alkalinity could be attributed to above long-term
average discharge in that year. He concluded "it seems doubtful that the 17
4-88
-------
percent decline can be explained by any reasonable mechanism other than acid-
ification of the water reaching the lakes."
California (McColl 1981).
The San Francisco Bay area of California receives part of its water supply
from two Sierra Nevada reservoirs—Pardee and Hetch Hetchy. These reservoirs
are located in an area underlain principally by Mesozoic granite and they
receive deposition affected by NOX and S02 pollution generated in the San
Francisco Bay area (McColl 1980, 1981). Precipitation chemistry apparently
is not measured at the reservoirs but if the data taken from other California
sites (see McColl et al. 1982, McColl 1982) can be used as a crude guide,
then pH of precipitation at the reservoirs may be in the range of 5.0 to 5.2.
Measurements of pH have been made weekly in untreated reservoir outlet waters
for the two reservoirs since 1954. Alkalinity has been measured weekly (by
titration to a pH 4.5 endpoint) in Pardee outlet water since 1944. McColl
(1981) reported on results of analyses of these data up to the year 1979.
McColl (1981) performed linear regressions of both the pH data (as annual
average H+ concentration) and the alkalinity data vs. time. The results of
the regression analyses are shown in Figures 4-25 and 4-26. The increases in
(H+) and decreases in alkalinity are clear. Further analyses by McColl
showed that (1) mean annual [H+] of the two reservoirs was correlated (r =
0.51, p < 0.02), (2) that rates of increase of [H+] did not vary signifi-
cantly on a seasonal basis, and (3) yearly precipitation did explain a small
percentage of the variance in mean annual [H+J of the release water but
that time was by far the most important factor.
McColl (1981) considered the possible influence of logging and mining within
the reservoir watersheds on the observed trends in [H+] and alkalinity,
concluding that these activities could not account for the trends. He sim-
ilarly considered and dismissed as unimportant the possible effects of
increases in the concentration of atmospheric C02.
McColl (1981) concluded from his analyses "It is clear that the [H+] of
waters in both reservoirs has increased since at least 1954, if not 1944. On
the basis of indirect evidence and correlative data discussed ... I conclude
that the most likely cause is the increased acidity of atmospheric depo-
sitions, especially those resulting from emissions of nitrous oxides by
automobiles."
National - U.S. Geological Survey Hydrologic Bench-Mark Network (Smith and
Alexander 1983).
The U.S. Geological Survey Hydrologic Bench-Mark Network consists of 47 water
quality and discharge monitoring stations located on streams in small, mostly
undeveloped watersheds in 37 states (Cobb and Biesecker 1971). At these
sites, sampling and water quality analyses (Skougstad et al. 1979) have been
applied beginning as early as 1964 (as late as 1974). Noting that the
watersheds apparently have experienced "little or no changes" in land use
since then and that these records "are particularly appropriate for
investigating atmospheric influences on water quality", Smith and Alexander
4-89
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TABLE 4-8. COMPARATIVE CHEMICAL DATA FOR COLORADO LAKES ABOVE 2,000 m (ADAPTED FROM LEWIS 1982)
•4S,
I
VO
o
Variable
PH
Alkalinity as C02
(mg-liter-1)
z residue
(mg-liter-1)
1938-1960 1979 Change* Percent Change
N Mean SE Mean SE Mean SE Mean SE
104 7.09 0.04 6.87 0.04 -0.22 0.04
104 22 2.3 18 2.3 -4.2b 0.61 -22 2.2
64 35 2.9 29 2.8 -6.1 1.44 -16 2.8
aChange in averages may not equal average change because of missing data.
^Alkalinity is given here in terms of C02. As a bicarbonate ion concentration, decline is
-5.9 mg-liter'1.
-------
200
160
7* 120
cr
+x 80
40
LEGEND
O HETCH HETCHY
• PARDEE
1955 1960
1965 1970
YEAR
6.7
6.8
6.9
7.0
7.1
7.2
7.3
7.4
7.6
7.8
1975 1980
Figure 4-25. Increasing acidity at Pardee and Hetch_Hetchy, shown by
hydrogen ion activity vs year, for the period 1954-79.
Adapted from McColl (1981).
4-91
-------
co
o
o
re
o
20
18
16
14
12
10
1945 1950 1955 1960 1965 1970 1975 1980
YEAR
Figure 4-26. Decreasing alkalinity at Pardee, shown by alkalinity as
CaC03 vs year, for the period 1944-79. Adapted from
McColl (1981).
4-92
-------
(1983) applied the Seasonal Kendall test for trend (Hirsch et al. 1982) to
monthly determinations (through 1981) of pH, alkalinity and sulfate
concentrations as well as to the ratio alkalinity: sum base cations. Smith
and Alexander (1983) gave a concise description of the statistics they used
in their analyses,
The Seasonal Kendall test is nonparametric and is intended for
analysis of time trends in seasonally varying water-quality data
from fixed, regularly sampled monitoring sites such as those which
the Bench-Mark Network comprises (Hirsch and others, 1982; see also
Smith and others, 1982). In addition to a test for trend, the
statistical procedure includes an estimate of the median rate of
change of quality over the sampling period (trend slope) and a
method for adjusting the data to correct for effects of changing
stream flow on trend in the water-quality record. Trend is defined
here simply as monotonic change with time, occurring either as an
abrupt or gradual change in water quality.
Because of the nationwide extent of the Hydrologic Bench-Mark data it is con-
venient to examine the results of the analyses of Smith and Alexander (1983)
by "region". In doing so, this review will follow the approach taken by
those authors as well as a previous reviewer (Turk 1983).
In most cases the pH trend information is somewhat ambiguous in relation to
the other trend results and therefore it will not be emphasized here. This
ambiguity is possibly because (1) pH data have been gathered for fewer years
than other records, and (2) in many cases the pH values of the waters are in
the range where buffer intensities are high. Also, in such pH ranges, vari-
ations in dissolved C02 can add significant noise to measured values.
Northeast - General trends in the northeast are for decreases in sulfate
concentrations and increases in alkalinity and the ratio alkalinity: sum
base cations since the mid 1960's (see Figures 4-27 to 4-29 and Table 4-9).
As Smith and Alexander (1983) noted, "In the northeastern quarter of the
country, SOg emissions have decreased over the past 15 years and the trends
in the cited chemical characteristics of Bench-Mark streams are consistent
with a hypothesis of decreased deposition in that region."
South and West - In these regions there seem to be increases in sulfate
concentration at the Bench-Mark stations and a tendency for decreases in
alkalinity. Emissions of SO^ have increased in the regions over the same
period (Smith and Alexander 1983, Gschwandtner et al. 1983; Chapter A-2,
Section 2.3.2). In the southeast there is evidence that precipitation has
recently become more acidic (Turk 1983) on a regional scale. In the west,
data on precipitation chemistry are scarce and local sources may predominate
in its control (Wisniewski and Keitz 1983).
To summarize the results of Smith and Alexander (1983), the trends in both
regional emissions and surface water chemistry at the Bench-Mark stations are
consistent with the hypothesis that the chemistry of precipitation can, and
has, significantly influenced the chemistry of streams in small, relatively
undisturbed watersheds.
4-93
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SIGNIFICANCE LEVEL
NO TRENDS p > 0.2 | 0.1
0.01 < p < 0.1 A p < 0.01
< 0.2
Figure 4-27.
Comparison of trends in stream sulfate concentrations at
Bench-Mark stations for the period of record through 1981
with trends in S02 emissions to the atmosphere by state,
1965-80. Triangles indicate-dorection and significance
levels of trends in stream sulfatex. Numbers give percent-
age change in S02 emissions from 1965 to 1980 for each
state. States showing increasing levels of S02 emissions
are shaded. States showing decreasing levels of SOg emis-
sions are unshaded. Adapted from Smith and Alexander
(1983).
4-94
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(a) ALKALINITY
SI6MF1CAMCE LEVEL
• NO TRENDS p > 0.2 ^ 0.1 < p < 0.2
0.01 < p < 0.1 A p < 0.01
Figure 4-28.
Trends in (a) alkalinity and (b) the ratio of alkalinity
to total major cation concentration at Bench-Mark stations
for the period of record through 1981. Symbols indicate
direction and significance level of trends. Dark symbols
indicate stations with mean alkalinity less than 1 meq rl.
Adapted from Smith and Alexander (1983).
4-95
-------
SIGNIFICANCE LEVEL
• NO TRENDS p > 0.2 ^ 0.1 < p < 0.2
0.01 < p < 0.1 A p < 0.01
A
Figure 4-29.
Trends in pH at Bench-Mark stations for the period of
record through 1981. Symbols indicate direction and
significance level of trends. Dark symbols indicate
stations with mean alkalinity less than 1 meq rl.
Adapted from Smith and Alexander (1983).
4-96
-------
TABLE 4-9.
SUMMARY OF TRENDS FOUND BY SMITH AND ALEXANDER (1983)
Alkalinity:
"Region" Sulfate Alkalinity sum base cations pH
Northeast 4 + -t- + M
South and West + 4- 4-M + M
M - mixed
4-97
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4.4.3.1.3 Summary—trends in historic data. Numerous studies have examined
temporal changes in the chemistry of streams and lakes in relation to the
chemical composition of incident precipitation. A frequent (and sometimes
major) drawback of these studies is a lack of clear documentation of the
historic data used. Often it is unproven that these crucial data are unbi-
ased, either by sampling or by the analytical procedures used. Many authors
recognize this problem; for example, Davis et al. (1978) stated of their
work, "the unconventional and imperfect means which we used to reconstruct
the pH history of Maine lakes were made necesary by the deficiencies of the
only data set available."
Listed below (and on Figure 4-33,
studies conducted on this topic.
reviewer and is based primarily on:
Section 4.4.4) are the "most reliable"
Inclusion is by best judgment of this
0 results of data analyses performed by the original author (including
information on ranges of values, total changes observed, and rates of
change)
0 lack of potential for bias in
sampli ng
analytical procedures
0 depth of documentation of all facets of the study as presented in
published form.
LOCALE
LaCloche Mt. region, Ontario
Halifax, Nova Scotia
Maine
New England
Adirondack region, New York
New Jersey Pine Barrens
Sierra Nevada Mts.,
California
USGS Hydrologic
Bench-Mark Stations
REFERENCE
Beamish and Harvey 1972
Beamish et al. 1975
Watt et al. 1979
Davis et al. 1978
Norton et al. 1981a
Haines and Akielaszek 1983
Schofield 1976a,b
Pfeiffer and Festa 1980
A. H. Johnson 1979
McColl 1981
Smith and Alexander 1983
CHEMICAL
VARIABLES
CONSIDERED
PH
PH
pH, alkalinity
PH
pH, alkalinity
PH
pH
PH
pH, alkalinity
pH, alkalinity,
S04
4-98
-------
In every case reviewed, the scientists who performed these studies concluded
that changes in surface water chemistry reflected, at least partly, either
(1) trends in regional emissions of S02, or (2) changes in chemical
composition of incident precipitation. This reviewer finds the body of
evidence presented here convincing. Particularly noteworthy by its absence
is any body of data indicating consistent decreases in alkalinity or pH of
surface waters at otherwise undisturbed sites not receiving acidic depo-
sition. Furthermore, this reviewer is unaware of any natural process that
would cause decreases in pH and/or alkalinity at the rates indicated by these
studies. Until appropriate evidence is presented in support of some such
natural process or until some better explanation of the data presented above
is put forth, the only logical conclusion is that acidic deposition (of
either remote or local origin) at these sites has caused, or is now causing,
acidification of some surface waters. It is only reasonable to assume that
other surface waters of similar sensitivity that receive similar levels of
acidic deposition have become or are now being acidified.
4.4.3.2 Assessment of Trends Based on Paleolimnological Technique (R. B.
Davis and D. S. Anderson)--
To assess the impact of acidic deposition and associated pollutants on lake
ecosystems, scientists have been analyzing the record contained in lake
sediments (Miller 1973; Berge 1979; Norton and Hess 1980; Davis et al. 1980,
1983). The sediment contains a diversity of physical, chemical, and
biological evidence which starts in deep sediments deposited thousands of
years ago and proceeds upward toward the sediment surface to cover the period
of the industrial revolution and recent technological activities. By
applying paleolimnological techniques including the dating of the sediment
(Birks and Birks 1980, Davis et al. 1984), researchers can reconstruct
chronological sequences of pollution inputs to lakes (e.g., lead) and
responses of the lake biota (e.g., plankton). Among the specific studies
being carried out is the identification and enumeration of the many kinds of
diatom remains (their siliceous shells) preserved in the sediments. Diatoms
are sensitive indicators of water pH; the various species differ in that each
is more or less restricted to a different pH range. By careful study of
these pH relationships for present-day diatom assemblages, it is possible to
calibrate the sedimentary diatom record so that the past pH of lake waters
can be inferred (Battarbee 1984, Davis and Anderson 1984). Similarly, the
deposition rate of some elements (e.g., Zn, Mn) (inferring increased leaching
in response to acidification, see Section 4.6.1.2) (Kahl and Norton 1983,
Kahl et al. 1984) has been used to estimate the range and direction of
historic pH change. Thus, a dated record of lake acidification can be
constructed by studying sediment cores.
The paleolimnological approach is useful for assessing the impact of acidic
deposition, because for the vast majority of lakes susceptible to
acidification no record of past, direct pH measurement exists. Where such
direct data exist they are generally of limited value because (1) pH readings
of lakes did not begin until after 1920, (2) the readings are usually only
for one year or a short series of years (Wright 1977), (3) they are
ordinarily only for mid-summer when pH's are usually highest (Davis et al.
1978), and (4) prior to about 1965 the readings were usually made by means of
4-99
-------
colorimetric pH indicators that may have altered the pH of poorly buffered
waters (Bates 1973, Blakar and Digerness 1984, Haines et al. 1983, Section
4.4.3.1.1). Paleolimnological reconstructions, on the other hand, use a
single technique that can provide a nearly continuous record of past pH
extending back thousands of years. While such reconstructions lack the
accuracy of properly taken, direct readings, they can circumvent the problem
of direct sampling of short-term variation in pH by integrating daily,
seasonal, and annual variation in single sediment samples encompassing an
entire year or small number of years. (The reconstructions therefore are
unsuitable for resolving short-term [< 3 yr] variation in pH, except possibly
for detailed studies of varved sediment.)
4.4.3.2.1 Calibration and accuracy of paleolimnological reconstruction of
pH history. Various publications (reviewed by Battarbee 1984) have presented
calibrations of the sedimentary diatom record of pH by deriving "transfer
functions" (Webb and Clark 1977) from the study of subfossil diatoms in
surface-sediments (uppermost 0.5 to 1.0 cm) of lakes. Davis et al . (1983)
and Davis and Anderson (1984) obtained surface-sediments from the deepest
parts of 31 lakes in northern New England and 36 lakes in Norway. These
authors developed regression equations relating such subfossil diatom
assemblages to pH of the surface waters in the lakes. The regressions have
r2 values of 0.27 to 0.91 and standard errors (Se) of +0.24 to +0.51 pH
unit. The regression coefficients have been used as transfer functions to
infer down-core pH. The errors for the New England data are greater than
those for Norway, partly because of the greater diversity of the New England
lakes. Regressions based on Hustedt (1937-39) diatom pH groups provide the
least accurate pH inferences, especially for lakes pH < 6.0. This probably
results from the semi-quantitative nature of HustecTt's groups and the
uncertainty in assigning individual taxa to groups. Charles (1982, 1984)
carried out oH calibrations based on diatoms in 38 Adirondack Mt., NY, lakes
obtaining r2 values of 0.61 to 0.94 and Se of 0.28 to 0.60 pH unit for
the regressions. Several factors responsible for variance in the surface-
sediment data sets would have remained more or less constant at any given
lake during the past two or three centuries. For example, elevation and lake
morphometry would have been constant, and concentrations of certain elements
in the water (e.g., K and CD are likely to have changed little. Thus, any
relative changes in pH inferred down-core at individual lakes are probably
more accurate and precise than the regression statistics for the
surface-sediment data would suggest.
4.4.3.2.2 Lake acidification determined by paleolimnological reconstruction.
Quantifying this paleolimnological approach and applying it to lakes affected
by acidic deposition are quite recent techniques (Battarbee 1984). The
methods are time-consuming. By early 1984 pH reconstructions have been
completed for about 40 lakes of which about half are in North America. This
approach is now being applied to more than 50 additional lakes in North
America (e.g., EPRI 1983). In southern Norway, reconstructions for seven
acidic (defined in this section as pH < 5.5) lakes indicate that
acidification started between 1850 and 19301 different dates at different
lakes) and that the total decrease in pH by 1980 was 0.1 to 0.8 unit
(depending on lake; average decrease about 0.40 unit) (Davis et al. 1983).
Before this acidification, these lakes were "naturally" all quite acidic (pH
4-100
-------
5.0 to 6.0) and were highly susceptible to further acidification. Flower and
Battarbee (1983) applied the Index B regression equation of Renberg and
Hell berg (1982) to diatom counts in 210pb-dated cores from two Scottish
lakes where they inferred pH decreases of 0.7 to 1.0 unit since about 1850 in
one lake and since about 1925 in the other. In southern Sweden, Aimer et al.
(1974) estimated a pH decrease from "about 6.0 to 4.5" for Stora Skarsjon
occurring between 1943 and 1973. Also in southern Sweden, Renberg and
Hellberg (1982) report for Gardsjon a pH decrease from 6.1 to 4.5 starting in
the 1950's; in Harsvatten, a decrease from 5.9 to 4.1 (no dates); and in
Lysevatten, from 6.2 to 5.3 (no starting date) until liming occurred in 1974.
The results for the northeastern United States are, so far, less clear.
Reconstructions for 6 acidic lakes in northern New England (Davis et al.
1983) indicate that acidification started between 1900 and 1970 (different
dates at different lakes) and that the total decrease in pH by 1980 was 0.2
to 0.35 unit (depending on lake; average decrease 0.26 unit). In an
additional two acidic lakes in the same region, no pH decrease was found in
one (Unnamed Pond, ME, now pH 4.7) and a decrease of about 0.2 unit in about
1965 was found in the other (Branch Pond, VT, now pH 4.7) (Davis and Anderson
pers. comm.). Analyses of metal content in sediment cores in the same lakes
support these conclusions. Sediments in acidic lakes (pH £ about 5.5)
consistently have decreasing concentrations of Zn toward the top of the
sediment (beginning about 20-50 years ago) (Figure 4-30; Davis et al. 1983,
Kahl et al. 1984). Lakes with pH > 5.5 and in regions not receiving acidic
deposition (e.g., Iskander and Keeney 1974) exhibit no decrease in Zn in
modern sediments. Davis et al. (1983), however, caution than in at least
three of the six acidic lakes examined, the pH decline may have resulted in
part from a recovery from an earlier, mild eutrophication (and elevated pH)
associated with lumbering or other disturbances (Section 4.4.3.3.2).
In the Adirondack region of New York, paleolimnological studies based on
diatom analyses are available for 10 lakes. Del Prete and Schofield (1981)
examined sediment cores for three lakes: Honnedaga Lake, pH 4.7; Woodhull
Lake, pH 5.2; Seventh Lake, pH 6.5. Honnedaga Lake had a marked increase in
acidophilous taxa (prefer pH <_ 7) in the surface-sediment compared to deeper
in the core. The sediments were not dated. The two other lakes showed no
significant change in estimated pH. Del Prete and Galloway (1983) presented
a preliminary analysis of pH changes in Woods Lake, pH 4.7; Sagamore Lake, pH
5.5; and Panther Lake, pH 6.0. None of these lakes exhibited a dramatic
shift in inferred pH in recent years (Figure 4-31). Whitehead et al. (1981)
reported on deepwater cores from three lakes in the High Peaks region of the
Adirondacks: Heart Lake, pH 6.5; Upper Wai Iface Pond, pH 5.0; and Lake
Arnold, pH 4.9. The emphasis in this instance was on long-term changes in pH
(late-glacial and Holocene), resulting from natural processes. All three
lakes were basic (pH > 8.0) in the late-glacial period, gradually becoming
more acid during the early-Holocene (pH of about 6.8 for Heart Lake, 6.0 or
below for the higher elevation lakes—Upper Wall face and Arnold). Whitehead
et al. (1981) did remark, however, that both Upper Wallface Pond and Lake
Arnold have acidified markedly in recent years. Charles (1984) examined the
recent pH history of Big Moose Lake (current pH 4.6 to 5.0). From about 1800
until about 1950, the inferred pH of the lake remained fairly constant at
about pH 5.7 (Figure 4-32). After 1950, however, the inferred pH dropped
4-101
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Pb
Zn
0
10-
o
no
CC
r>
oo
o
Lul
00
UJ
CO
40-
0
10
20
30
40 -
ii iii i
i r
400 0
PPM OF IGNITED WEIGHT
5(
Figure 4-30.
Profiles of Pb and Zn concentrations (as ppm of ignited weight) at Speck Pond, ME.
Vertical scale in cm below the sediment surface. Adapted from Davis et al. (1980).
According to 210Pb and pollen chronostratigraphic dating, 20 cm would be about 1810,
15 cm about 1870, 10 cm about 1930, and 5 cm about 1958 (R. B. Davis and S. A. Norton
pers. comm.).
-------
O
CO
4.0
10
o
o.
UJ
o
50 -
INFERRED pH
5.0 6.0
5.0
WOODS LAKE
SAGAMORE LAKE
PANTHER LAKE
Figure 4-31.
Historical, inferred pH values vs sediment core depth (cm) for Woods, Sagamore, and
Panther Lakes. Adapted from Del Prete and Galloway (1983).
-------
pH
n.
UJ
a
0
2
4
6
8
10
12
14
16
18
20
22
24
26
28
30
32
34
36
38
40
I I I I I I
* Ambrosia rise
** 1903 fire
+ pH measurement
1982
1970
1960
1950
1940
1920
**
1910
1880
*
1860
1850
1840
1820
1800
Figure 4-32.
Profile of inferred pH for Big Moose Lake, NY, based on
analysis of diatom taxa in sediment cores. Adapted from
Charles (1984).
4-104
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steadily and relatively quickly to about 4.7. This decrease in inferred pH
is corrobated by historical water chemistry data and by the associated
decline and loss of fish populations in the lake. Charles (1984) concluded
that, given the magnitude and timing of the pH shift the most reasonable
explanation for the decline in inferred pH is acidic deposition.
In early 1984, detailed paleolimnological analyses of pH change in the past
300 years have been completed for 15 acidic lakes in the northeastern United
States. Based on the sediment diatom record, 9 of these lakes have experi-
enced pH decreases of < 0.3 unit in recent years and two have experienced
decreases of about 0.6 and about 1.0 unit (beginning about 9 to 80 years ago,
depending on the lake). For at least 3 of these 9 lakes, long-term trends in
pH suggest that the pH decline may have resulted in part from a recovery from
an earlier, mild eutrophication (and elevated pH) associated with lumbering
or other disturbances (Davis et al. 1983). For 4 of the 9 acidified lakes,
however, no such period of pH increase followed by pH decrease has been noted
(Del Prete and Schofield 1981, Charles 1984).
Additional sediment cores have been collected both in northern New England
and the Adirondacks and are currently being processed. The paleolimnological
data published to date are too limited and variable to provide firm estimates
of the extent and magnitude of acidification, natural or anthropogenic.
4.4.3.3 Alternate Explanations for Acidification-Land Use Changes (S. A.
Norton)--
Land use changes and natural processes may directly affect the pH (and rela-
ted chemistry) of surface waters via several mechanisms, including variations
in the groundwater table; accelerated mechanical weathering or land scarifi-
cation; decomposition of organic matter; long-term changes in vegetation; and
chemical amendments. Details of each are presented below.
4.4.3.3.1 Variations in the groundwater table. The water table in mineral
or organic soils generally marks a transition from aerobic to anaerobic con-
ditions. This transition is particularly sharp in saturated, organic-rich
soils. With a lowering of the groundwater table due to drought, lowered lake
levels, or drained terrestrial systems (e.g., bogs), previously anaerobic and
reduced material is exposed to oxygen. The following types of reactions may
occur:
FeS2 + 02 + H20 + Fe(OH)3 or FeO(OH) or Fe203 + 2H+ + S042'
Mn2+X + 02 + H20 -> Mn02 + H2X
Organic matter + Decay + N03~ + H+ + C02
The associated H+ production is commonly accompanied by accelerated loss of
cations from the ecosystem (Likens et al . 1966, Damman 1978).
4.4.3.3.2 Accelerated mechanical weathering or land scarification. These
processes may result from logging, fires, slope failure, and other distur-
bances of the land surface. The exposure of relatively unweathered material
4-105
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to chemical weathering results in accelerated leaching of cations from
watersheds. If uptake of nitrogen from decaying organic material occurs, the
pH of surface waters may rise along with cation concentrations. This results
in eutrophication trends in downstream waters (Pierce et al. 1972). Read-
justment of the system may take decades, with concurrent long-term changes in
surface water chemistry, including pH. For example, Dickman et al. (1984)
found that in individual lakes with burned drainage basins or where logging
had occurred, fossil diatom assemblages indicate that a period of 30-40 years
of elevated pH occurs in downstream lakes with pH gradually returning to
pre-disturbance values (see also Section 4.4.3.2). Recent acidification
unrelated to deforestation or recovery from fires, however, has resulted in
diatom-inferred lake pHs lower than at any time since about 1890 (also see
Section 4.4.3.2.2).
4.4.3.3.3 Decomposition of organic matter. As discussed in Section
4.3.2.6.2 (Equation 4-7), a net loss of organic matter generally results in
accelerated production of nitric acid, C02, and increases in cations—all
other conditions being the same. Following experimental deforestations at
Hubbard Brook, NH, Likens et al. (1970) observed decreases in stream pH,
apparently as a result of increased nitrification associated with accelerated
decomposition rates. Regrowth of vegetation after deforestation may,
however, induce a rapid (2-4 years at Hubbard Brook) return to ambient pH
values. A change in stored biomass is generally accompanied by other
alterations, such as a shift in canopy interception of aerosols, changes in
evapotranspiration, or changes in surface water temperatures, so the individ-
ual effects are difficult to sort out.
4.4.3.3.4 Changes in vegetation. Long-term changes in vegetation bring
about various physical and chemical changes in the soils and watershed, which
in turn result in long-term changes in surface water chemistry. Soil acidi-
fication can clearly be caused by either reforestation (after grasses) or
changes in forest type on otherwise equivalent sites (Raynal et al. 1983,
Overrein et al. 1980). Malmer (1974) reviewed the Swedish literature and
concluded that chemical changes in soils associated with reversion of farm-
land to forest (increased organic content, lower pH, lower exchangeable
metals) are much the same as those that have been attributed to acidic dep-
osition.
Shifts in the dominant vegetation affect surface water chemistry in a variety
of ways. Many researchers have suggested that in aggrading forest ecosys-
tems, when plants take up an excess of cations over anions, protons will be
released in order to maintain electroneutral ity (Reuss 1977, Rosenqvist
1977). However, Nilsson et al. (1982) and Gorham et al. (1979) both maintain
that root uptake of cations leads to soil acidification, not streamwater
acidification.
Certain vegetation types (e.g., conifers) produce abundant humic material
that can produce acidity. Thus the appearance of these vegetation types in
succession could yield long-term declines in pH as well as increased organic
matter concentrations. The appearance of Sphagnum sp., perhaps as a result
of changes in moisture regime, could also result in acidification of surface
waters due to the highly effective cation exchange capacity of Sphagnum with
4-106
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associated release of H+ (Clymo 1963). Hemond (1980) evaluated four
potential sources of acidity within Thoreau's Bog, MA (pH = 3.8): (1) acid
precipitation (pH about 4.1), (2) cation exchange capacity of the peat, (3)
biochemical transformations of S and N, and (4) production of organic acids.
Acid precipitation contributed about 0.43 meq JT1, but was counteracted
by the subsequent reduction of the associated $642- and NOs" (gener-
ating 0.4 meq jr1 of alkalinity). Ion exchange contributed only modestly
(0.05 meq £-1) to bog acidity. The acidity of Thoreau's Bog was
maintained principally by organic acids (1 meq £~1). The role of organic
acids in effecting low pH is important in many bogs, but not universal
(Hemond 1980).
Variations in interception and evaporation associated with different vege-
tation types may affect the quantity of precipitation reaching the soil, as
well as quality, via changes in aerosol capture of acidic components. Con-
version from deciduous hardwood stands to pine resulted in a 20 percent
reduction in stream flow within the Coweeta watershed, North Carolina (Swank
and Douglas 1974). This change, by itself, would result in increased concen-
trations of all biologically conservative elements. Variations in H+
loading reaching the soil may range over a factor of 10 due to vegetation
type (Skeffington 1983). However, it is not known how these net deposition
fluxes are transmitted to surface water chemistry. The linkage between
natural soil acidification (due to the C02-H20 system and organic acid
production) and surface water acidification has not been demonstrated.
Harriman and Morrison (1980) observed that spruce reforestation in Scotland
resulted in acidification of streams. It is not clear, however, whether this
change in acidity is related to indigenous processes associated with the
spruce vegetation as compared to the previous peaty soil vegetation, or to
increased dry deposition as a result of greater aerosol capture of acidic
components, or to changes in hydrology.
4.4.3.3.5 Chemical amendments. Adding some fertilizers (such as ammonium
phosphate) to agricultural soils has an acidifying effect on soils, and this
could be transmitted to surface waters (along with elevated levels of phos-
phate). This potential acidification is generally recognized and the
affected soils are amended with a base, CaCOs, with subsequent elevation of
pH. In regions where agriculture is on the wane and reforestation is under-
way, the implicit cessation of CaCOs application might result in a decline
in the pH of surface waters, erroneously suggesting natural acidification.
4.4.3.3.6 Summary—alternate explanations for acidification. Certainly
natural processes and land use changes can result in slightly acidic waters
(see Section 5.2, Chapter E-5), and must be considered when assessing current
and potential damages related to acidic deposition. There is no evidence,
however, that land use changes in areas not receiving acidic deposition
produce clear surface waters with pH's much less than 5.5. Land use changes
may bring about dystrophication (production of organic-rich colored water)
and acidification due to organic acids, but this is a rare phenomenon and
unrelated to clearwater acidification. Thus natural acidification, or the
return of a system to its natural state will not produce clearwater
oligotrophic lakes with pH much less than 5.5.
4-107
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Any evaluation of the relative importance of natural processes and land use
changes vs the importance of acidic deposition with regard to the acidi-
fication of surface waters discussed in Section 4.4 must first consider the
fol1owi ng:
1. Acidification involves the loss (partial or complete) of alkalinity,
i.e., reduction in HC03-;
2. In clearwater, oligotrophic lakes and streams (i.e., sensitive to acidic
deposition as discussed in Section 4.3.2), complete loss of alkalinity is
only associated with the presence of a strong acid, normally H2S04,
but locally and temporally HN03 (Sections 4.4.2 and 4.4.3.1.2).
3. In areas presently not receiving acidic deposition (e.g., Rocky Mts.,
Colorado; Experimental Lakes Area, Ontario; Labrador) sensitive lakes
rarely have pH < 5.5; few have pHs between 5.5 and 6.0; most have pHs
between 6.0 and 7.0 (Wright 1983).
4. In areas not receiving acidic deposition, studies of the effects of land
use changes, particularly those related to changing vegetation, have
focused on effects on soil fertility and soil pH. Rarely has the effect
on surface waters been evaluated, and when it has the results have often
been ambiguous (e.g., Schindler et al. 1980a). Variations in lake
acidification due to land use changes typically involve shifts in pH
above 6.0, with HCOa' as the dominant anion.
5. In areas receiving acidic deposition, the roles of natural processes and
land use changes have not been carefully evaluated in the field because
of problems in maintaining control systems. However, several studies
have qualitatively (or semi-quantitatively) evaluated the contribution of
acidic deposition vs alternate H+ sources:
a) Harriman and Morrison (1982) noted that reforestation within some
watersheds resulted in streams more acidic than similar streams
draining moorland vegetation; however both types of streams were more
acidic than those in regions not receiving acidic deposition.
Additionally, the dominant anion, after sea-salt correction, was
S04 •
b) Drablrfs et al. (1980) examined historical land use changes in
southern Norway and their relationship to regional lake acidification
and decreasing fish populations. They found no relationship.
c) Charles (1984) examined 4 alternative hypotheses for the rapid
decrease of pH (based in diatom pH reconstruction methods) in Big
Moose Lake (Adirondacks, NY) since 1950 (Figure 4-32, Section
4.4.3.2.2): (1) long-term natural acidification caused by increased
leaching of cations from the soil; (2) increased development of
bog-vegetation (e.g., Sphagnum) in the watershed; (3) disturbance in
watershed vegetation (e.g., fires, logging) followed by rapid
regrowth; and (4) increased atmospheric loading of strong acids.
Alternative 1 is unlikely, and can not explain the break in the pH
4-108
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pattern around 1950, and the large and rapid decline in pH from 1950
to the present (Figure 4-32). There is no evidence of significantly
increased bog-type vegetation in the watershed (alternative 2).
Logging of the watershed in the late 1800s and early 1900s caused no
apparent sizeable shift in lake pH, thus ruling out alternative 3.
Charles (1984) concluded that the most reasonable explanation for the
decline in pH is acidic deposition.
d) Sharpe et al. (1982) investigated the potential causes of acidity in
four stream systems in the Laurel Mountains of Pennsylvania. Two
streams had bogs at their headwaters; two did not. All four streams
experienced decreases in pH during high flow. The estimated
contribution of bog discharge to stream acidity was least during
periods of peak flow, when H+ concentrations were highest.
Two generalizations result from this analysis. Acidification of clearwater,
oligotrophic surface waters to pH values below 5.0 occurs only in regions
receiving acidic deposition, and regional acidification only occurs where
acidic deposition is present. Secondly, regional surface water acidification
occurs without land use changes in areas receiving acidic deposition.
4.4.4 Summary—Magnitude of Chemical Effects of Acidic Deposition
At the beginning of 4.4.3, it was noted that systems impacted by acidic
deposition had three characteristics—they were sensitive, received acidic
deposition, and had been shown to be acidified.
Aquatic systems most likely to be influenced by atmospheric deposition (i.e.,
sensitive) are those with alkalinity of less than 200 yeq a~ . Large
areas of Canada and the United States contain such systems. For example,
approximately 80 percent of New England, by virtue of its geology, has
surface waters with less than 200 yeq JT1. Areas in provinces of
eastern Canada identified as sensitive cover from 90 percent (Quebec) to 20
percent (New Brunswick) of the total land area.
Of the aquatic systems that are potentially susceptible to acidification
(Figures 4-5 to 4-8), only those located in eastern North America and in
small regions of western North America are receiving acidic deposition (pH _<
5.0; Chapter A-8, Section 8.4; W-isniewski and Keitz 1983).
Acidification of aquatic systems receiving acidic deposition has been noted
in several instances. Using the information on temporal trends in Sections
4.4.3.1.2 and 4.4.3.2 and studies of the role of atmospheric sulfur in
aquatic systems (Section 4.4.3), Table 4-10 and Figure 4-33 indicate (with
numbers) areas that have been shown to be acidified by acidic deposition.
All numbers fall in sensitive areas receiving acidic deposition. In addi-
tion, the letters on Figure 4-33 represent studies in which acidic deposition
and atmospheric S, because of their low concentrations, have been shown not
to have acidified aquatic systems.
4-109
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TABLE 4-10. THE RESPONSE OF AQUATIC SYSTEMS TO ATMOSPHERIC DEPOSITION OF
ACIDIC AND ACIDIFYING SUBSTANCES FROM LOCAL OR REGIONAL (LONG-RANGE) SOURCES.
-p.
I
Locale
1.
2.
3.
4.
5.
6.
7.
8.
A.
B.
C.
D.
LaCloche Mts., Ont.
Halifax, N.S.
Northern New England
Adirondacks, NY
New Jersey Pine Barrens
Muskoka-Haliburton, Ont.
Laurentlde Park, Que.
Sierra Nevadas, CA 5
Experimental Lakes Area,
Ont.
Rocky Mts., CO
Cascades, WA
Labrador
Approximate
(kg ha'1
pH SO;2-
4.2
4.3 26
4.3 25
4.2 35
4.3 25
4.2 35
4.3 35
.0-5.2 -
5.0
4.9
5.0
4.9
Wet Deposition
yr-1)
N03-l
12
15
22
15
25
22
-
<10
<10
<10
10
Temporal Trends
Beamish and Harvey
1972
Beamish and Harvey
1978
Watt et al . 1979
Davis et al. 1978
Norton et al . 1981a
Halnes and Aklelaszek
1983
Schofleld 1976a
Pfeiffer and Festa
1980
A.H. Johnson 1979
McColl 1981
Schlndler and
Ruszcznski 1983
Type of Evidence
Paleolimnological Excess Sulfate
NRCC 1981
Davis et al. 1983 Wright 1983
Norton 1983
Johnson et al . 1981
Charles 1984 Galloway et al
Del Prete and 1983c
Schofleld 1981 Wright 1983
Wnitehead et al .
1981
Dillon et al .
1980
NRCC 1981
Wright 1983
Bobee et al .
1982
Wright 1983
Dillon et al.
1980
NRCC 1981
McCarley 1983
Wright 1983
Logan et al .
1982
Wright 1983
-------
Figure 4-33.
The response of aquatic systems to atmospheric deposition
of acidic and acidifying substances from local or regional
(long-range) sources. The numbers refer to references noted
in Table 4-10, which conclude that acidic deposition has
caused acidification of aquatic systems. The letters refer
to references in Table 4-10, where possible acidification of
aquatic systems has been studied but not found. Precipita-
tion pH isopleths are based on Chapter A-8, Section 8.4 and
Wisniewski and Keitz (1983).
4-111
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Acidification of aquatic systems by acidic deposition is supported by the
following lines of evidence:
0 Due to acidic deposition, S042" concentrations have increased in
aquatic systems in much of eastern North America. The increase in
S04 problem has to have been matched by an increase in CB or
H+. Since aquaticsystemswTth original low alkalinities are
characterized by watersheds with low CB/H+ ratios in the soil, a
large portion of the increase in $042- w111 have to be matched by an
increase in H+, i.e., decreased alkalinity.
o Although there can be significant problems with comparing old and new
data, overall, the analysis of temporal records shows recent decreases in
alkalinity and pH in some otherwise undisturbed streams and lakes in
areas receiving acidic deposition. As yet, no body of evidence exists
suggesting that changes of such magnitude, and at such rates, occur in
otherwise undisturbed areas not receiving acidic deposition.
0 The limited application of paleolimnological indicators (diatoms and
metals; in northeastern United States) shows decreases in pH over the
last 10 to 80 years for most (9 of 15) acidic lakes studied. For at
least three of these acidified lakes, the recent decline in pH may
reflect in part a recovery from an earlier increased pH due to temporary
eutrophication. For 4 of the 9 acidified lakes, however, no such pattern
of pH increase followed by pH decrease has been noted.
0 No other possibilities exist to explain the regional scale of acidifi-
cation that has occurred. For example, changing land use is at times
advanced as one explanation. However, in areas with comparable changes
in land use, it is only those areas receiving acidic deposition that are
acidified.
In some of the studies, the link between acidic deposition and surface water
acidification can be critized because of weakness in historical data or lack
of attention to specific processes occurring in the soil or water body that
could also result in acidification. However, given the fact that S04
values in clearwater oligotrophic lakes and streams of eastern North America
are substantially higher in areas receiving acidic deposition (Figures 4-1
and 4-2, Section 4.3.1.5.2) and that such an increase has to have been
associated with at least some increase in H+ (decrease in alkalinity), due
to the acidic nature of the soils surrounding lakes with low alkalinity, it
is reasonable to conclude that surface water acidification has resulted from
acidic deposition.
A further piece of evidence linking acidic deposition with increased
S04 and decreased alkalinity in aquatic systems is an analysis of
10-15 years of water quality records from a network of benchmark sampling
stations in the United States (Smith and Alexander 1983). The authors, based
on the seasonal Kendall test for trends in monthly records of stream
S04 and alkalinity, conclude that the regional pattern for stream
sulfate trends was similar to that reported for trends in S02 emissions to
the atmosphere. In addition, trends in stream alkalinity were the
4-112
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approximate inverse of stream S042- trends (Smith and Alexander 1983).
These results support the conclusion that not only do S02 emissions affect
the S042~ concentrations of surface waters but that increases in
surface water S04 result in decreases in surface water alkalinity.
H2S04 is the primary cause of the long-term acidification of aquatic
systems on a regional basis. The maximum decrease in alkalinity that can
occur due to acidic deposition depends on the maximum long term increase in
S042" in surface waters. In the northeastern United States and
southeastern Canada this is about 100 yeq £-1. In areas closer to
emission sources of sulfur, the maximum increase in S042~ may be 100 's
of yeq £~1. The actual decrease in alkalinity depends on how much of
the increased S042~ is balanced by increases in base cations. One
estimate (Henriksen 1982a) is that for a 100 yeq £ -1 increase in
S042' and N03" there will be, on the average, an approximately 60
ueq £-1 decrease in alkalinity. The pH change associated with an
alkalinity decrease of 60 yeq £-1 can range from a few tenths of a pH
unit to 2 pH units. Those systems with the lowest initial alkalinities will
show the greatest loss of alkalinity due to acidic deposition because of the
scarcity of exchangeable cations in the terrestrial system.
In addition to long-term acidification (years and decades) by
short-term acidification (days to weeks) occurs as a result of the combined
action of H2S04 and HNOa in areas that develop acidic snowpacks or
receive a large amount of rain over a short period of time. Losses of
alkalinity of 200 yeq r1 and reduction of pH from 7.0 to 4.9 have been
reported.
4.5 PREDICTIVE MODELING OF THE EFFECTS OF ACIDIC DEPOSITION ON SURFACE
WATERS (M. R. Church)
The predictive modeling of the effects of acidic deposition on the chemistry
of natural waters is an extremely complicated task requiring a great amount
of data, knowledge, insight, and skill. Two avenues exist for approaching
the problem—empirical modeling and mechanistic modeling. Each approach has
its advantages and disadvantages.
Empirical models, in general, have two principal advantages. First, they
integrate the processes between inputs and outputs, thus eliminating the need
for precise knowledge of the behavior of controlling mechanisms. Second, they
are usually very simple computationally. Empirical models do have certain
drawbacks, however. One drawback is that long periods of data may be
required to verify that an observed relationship between inputs and outputs
represents a steady state. Other drawbacks include the problems of verifying
the validity of applying a relationship observed in one geographic area to
another area and extrapolating from one observed loading rate (or regime) to
another. Finally, because they are almost always based on assumptions of
steady state, empirical models usually possess no time component.
Mechanistic models, of course, have a different set of pros and cons. The
principal attraction of mechanistic modeling is that if accurate mathematical
representations of all (or the most important) of the physical /chemical/
4-113
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biological processes involved can be devised and properly related to one
another, then a variety of extrapolations may be made with confidence. Such
extrapolations include the application of the model (with appropriate
calibration) to a variety of geographic areas; the use of the model to
estimate rates of change (e.g., of the alkalinity or pH of a lake); and the
prediction of responses to almost any loading scenario.
Along with this potential for widespread application, however, go certain
problems. The first, and perhaps the most obvious, is that the knowledge may
not exist to allow formulation of accurate representations of all (or even
the most important) physical/chemical/biological processes of interest.
Second, mechanistic models (especially of lake-watershed ecosystems) require
extensive calibration for the region to which they will be applied. Such
calibration can be very time consuming and expensive. Third, to be used
predictively, mechanistic models that operate with a relatively short
time-step (say, less than one week) require a correspondingly fine-scale
source of predicted input. This requires a separate method (or model) to
generate inputs of precipitation form, amount, and quality as stochastic
variations around annual (or even seasonal) means. This task, by itself, is
somewhat involved and time consuming. The last drawback to the mechanistic
approach is that as the representations of controlling processes become more
detailed and intertwined, the time and effort required to perform the
calculations increases substantially, even to the point where significant
amounts of computer time may be needed to perform long-term simulations.
A variety of models exist or are currently being developed to deal with the
problem of predicting the effects of various levels of acidic deposition on
the chemistry of surface waters (e.g., Aimer et al. 1978; Henriksen 1980,
1982a; Christophersen and Wright 1981; Thompson 1982; Chen et al. 1982;
Christophersen et al. 1982; Schnoor et al. 1982). The models range from
simple empirical approaches to very computationally complex formulations. A
comprehensive review of all of these efforts is beyond the scope of this
chapter. Instead, a brief review is presented of those four empirical models
that are, so far, the best known and most referenced of existing approaches.
4.5.1 Almer/Dickson Relationship
Aimer et al. (1978) plotted lake pH vs lake sulfur loading [g of S nr2
yr-1 "concentration of 'excess' (above sea salt contributions) sulfur
multiplied by yearly runoff"] for Swedish lakes. They found "titration
curve"-type patterns for data from sets of lakes occurring in areas of
similar bedrock. They plotted two curves (Figure 4-34): one for waters
"with extremely sensitive surroundings" and one for waters with "slightly
less sensitive surroundings" (Aimer et al. 1978). The authors did not
specify their procedures for lake selection nor did they define any objective
method for classifying lakes with regard to their surroundings and responses
to sulfur loadings (e.g., 'extremely sensitive1 or 'slightly less
sensitive1). This limits their approach as a general tool for predicting the
pH of lakes as a function of sulfur loadings.
At first glance, using such a treatment of data might seem to be a way to
help determine the levels of sulfate deposition (to watersheds) that may have
4-114
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adversely affected lake water quality (pH). Closer examination of the
approach, however, demonstrates that care must be taken in making such an
application. For example, the quantity "excess S in lake water" must be
carefully distinguished from the quantity "total excess S deposited".
Unfortunately, confusion about this question and the original designation of
the abscissa of Figure 4-34 has led to several mislabelings of reproductions
of the original figure (e.g., Glass 1980, Glass and Loucks 1980, Loucks et
al. 1981, U.S./Canada 1982, Loucks 1982). If Figures 4-35 and 4-36 {adapted
from Aimer et al. 1978) can be comparedTnote that they represent data
roughly four years apart), they show that the relationship is quite variable
for the regions of Sweden for which the "Almer/Dickson Relationship" was
derived. Not only is more excess sulfur deposited than shows up in lake
water (indicating some sulfate retention), but also the isopleths of the two
plots are not parallel, indicating that this retention is different in
different regions.
As this example illustrates, the crux of the problem in applying the
"Almer/Dickson Relationship" is the translation of the abscissa of Figure
4-34 from a representation of "excess S in lake water" to some more primary
or causative factor (e.g., areal rate of total excess sulfur deposition,
area! rate of wet excess sulfur deposition, concentration of sulfate in
precipitation, pH of precipitation, etc.). Such a translation requires
quantitative knowledge of the relationships among such things as concen-
trations in lake waters, concentrations in precipitation, ratios of wet to
dry deposition, amounts of precipitation, amounts of runoff, etc. In turn,
the statistical estimation of these types of relationships for any region
requires large amounts of data for that specific region.
Beyond the problems described above, other pertinent factors involved in the
use of the "Almer/Dickson Relationship" must be considered. It is important
to note that several assumptions are inherent in the approach.
First, Aimer et al. (1978) assumed that within each of the two sets of lakes
represented by the curves of Figure 4-34, initial (e.g., prior to deposition
of strong acids) steady-state values of alkalinity were all the same.
Second, they assumed that the current pH values and the current excess sulfur
concentrations they observed in lake water were both at steady state. No
evidence was offered in support of either of these assumptions. Finally,
there is the problem of hysteresis. No data exist to indicate that as a
result of decreases in S042' loading rates, previously acidified lakes
would "return" along the curves of Figure 4-34 to higher steady-state pH
values. Conditions extant have not permitted such observations to be made,
and there is perhaps no clear scientific consensus on this problem.
As a minimum condition, before the Almer/Dickson Relationship can be applied
to the problem of predicting the effects of changes in acidic deposition on
the chemistry of surface waters in any geographic region, reliable
quantitative relationships between primary factors (e.g., wet sulfate
deposition) and sulfate concentrations in surface waters must be developed.
Further, all assumptions inherent in the approach require testing and
validation.
4-115
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7.0
6.0
5.0
4.0
0 1 2
EXCESS S IN LAKE WATER (g m'2 yr"1)
Figure 4-34. The pH values and sulfur loads in lake waters with
extremely sensitive surroundings (curve 1) and with
slightly less sensitive surroundings (curve 2). Load =
concentration of 'excess1 sulfur multiplied by the yearly
runoff. Adapted from Aimer et al. (1978).
4-116
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Figure 4-35.
Atmospheric load of 'excess' sulfur from precipitation and
dry deposition, 1971-72 (g S m~2 yr~l). Dry deposition
calculated from a deposition velocity of 0.8 cm s~l.
Adapted from Aimer et al. (1978).
4-117
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Figure 4-36.
'Excess' sulfur in lake water per year (g S m~2 yr~l).
(Concentration of "excess sulfur multiplied by the yearly
runoff.) Adapted from Aimer et al. (1978).
4-118
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4.5.2 Henriksen's Predictor Nomograph
The contributions of Henriksen (1979, 1980, 1982a) to the empirical study of
the effects of atmospheric and edaphic factors on the chemistry of
oligotrophic lakes in Scandinavia are well known. Among his contributions is
the "predictor nomograph"—an empirical relationship intended to be used as a
tool in predicting effects of varying levels of acidic deposition on the pH
of lakes.
Using data from 719 lakes in southern Norway (Wright and Snekvik 1978),
Henriksen (1980) compared the concentration of excess (above sea salt
contributions) calcium plus excess magnesium with excess sulfate
concentrations in the pH ranges 4.6 to 4.8 and 5.2 to 5.4 (see Figure 4-37)
and found "highly significant" linear correlations. Axes of excess calcium
concentration (parallel to the axis of excess calcium plus magnesium) and
excess sulfate in precipitation and pH of precipitation (both parallel to the
axis of excess sulfate in lake water) complete the predictor nomograph.
These final axes were developed from local empirical relationships.
Henriksen (1980) used an independent data set from a survey of 155 Norwegian
lakes to test his nomograph and found that it correctly predicted pH
groupings approximately 85 percent of the time. Henriksen (1982a) concluded
that the relationships depicted by the predictor nomograph corroborated his
hypothesis that for the lakes he studied (clear headwater oligotrophic lakes
on granitic or siliceous bedrock) "acidified waters are the result of a large
scale acid-base titration." He further concluded that the nomograph was
capable of predicting the effects that a change in precipitation pH might
have on the pH status of lakes of the type he studied in the region he
studied.
As with all predictive constructs, or models, a number of key assumptions
(all clearly recognized and noted by Henriksen 1980, 1982a) are involved in
the use of the predictor nomograph.
One assumption or condition for using the model is that it not be used for
lake waters with significant concentrations of organic acids. This is
because (1) these acids may affect lake pH independent of precipitation
acidity and (2) analyses for calcium and magnesium include these ions bouml
to organics; thus, ionic concentrations of excess Ca2+ plus excess Mg
may be overestimated.
A second factor in the use of the nomograph involves the possible increased
leaching of base cations from soils by acidic precipitation. In his original
work, Henriksen (1980) assumed no increased leaching of base cations but
noted the possible importance such an event would hold for use of the nomo-
graph. He has subsequently studied this question in more detail, using data
from lakes in North American and Scandinavia (Henriksen 1982a).
He examined data from lakes in areas of similar geology over a gradient of
deposition acidity and he also compared time trend data of calcium and magne-
sium in certain waters. Unfortunately, he found no clear cut answer to the
question. In some cases, there was evidence of increases in base cation
concentrations (up to 0.63 for Lake Rishagerodvatten, Sweden). In other
4-119
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cr
O)
3.
*
re
O
0
0 100 200
S04* IN LAKEWATER (yeq £-1)
50
100
S04* IN PRECIPITATION (ueq
5.04.7 4.54.4 4.3 4.2 4.
pH OF PRECIPITATION
4.0
Figure 4-37.
A nomograph to predict the pH of lakes given the sum of
non-marine calcium and magnesium concentrations or non-
marine calcium concentration only and the non-marine
sulfate concentrations in lake water or either the
weighted-average non-marine sulfate concentration or
the weighted-average hydrogen ion concentration in
precipitation. * denotes sea salt corrected values.
Adapted from Henriksen (1980).
4-120
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cases, there was none. In an effort to overcome these difficulties and
conflicting data, Henriksen (1982a) used his best judgment to designate a
maximum value of "base cation increase factor" of 0.4 yeq (Ca* + Mg*)/yeq
$04*. That is, for every yeq £-1 increase in excess sulfate ($04*)
concentration in a lake, a maximal increase in excess calcium plus magnesium
(Ca* + Mg*) concentration would be 0.4 yeq £-1. It must be noted that
in at least one case, Henriksen (1982a) found a greater increase factor than
this—0.63 for Lake Rishagerodvatten, Sweden. Care should be exercised in
the application of this "base cation increase factor" for predictive pur-
poses. It may vary significantly from region to region (or watershed to
watershed within a region) as a function of soil chemical properties (e.g.
sulfate adsorption capacity, cation exchange capacity, base saturation), soil
depth, and the path of precipitation through the soil. In fact, it seems
reasonable to assume that for some regions initially experiencing acidic
deposition, the "increase factor" may be as high as 1.0. Certainly more
quantitative research is needed on this question.
Another condition noteworthy in the use of the predictor nomograph is the
premise that all data used in its construction and verification represent
steady state conditions. Due to the large number of lakes and deposition
events and periods sampled, the data requirements to verify this condition
for the nomograph are astronomical and virtually impossible to satisfy. As
an article of faith it must be assumed that the data employed do represent
steady state conditions. For many of the lake data (especially at the
"edges" or extremes of conditions) this probably is not a bad assumption.
Lake data representing transitory conditions are, perhaps, more suspect.
A final question to consider in regard to the predictor nomograph is its
application to geographic regions other than (but similar to) the one
for/from which it was developed. This is always a key question with such
empirical models. Even if the general approach is accepted as sound, common
sense dictates that the empirical relationships found in southern Norway and
Sweden may not pertain to even seemingly analogous conditions elsewhere.
(Certainly this is true of the axes relating precipitation chemistry to
excess sulfate concentrations in lakes. Most acidic precipitation in North
America contains relatively more nitric acid than does acidic precipitation
in Scandinavia.) The inconsistencies encountered by Bobee et al. (1982),
Haines and Akielaszek (1983), and Church and Galloway (1984), in attempting
to apply the nomograph to lakes in Quebec, New England, and the Adirondacks,
respectively, should be noted in this regard. It may very well be that the
predictor nomograph will have to be modified to accommodate local relation-
ships for whatever region application is attempted.
4.5.3 Thompson's Cation Denudation Rate Model (CDR)
As seen in the previous discussions of the Almer/Dickson relationship and
Henriksen1s predictor nomograph, the quantification of the interrelationships
of sulfate loading, base cation concentrations, and surface water pH seem to
hold promise for understanding and predicting surface water chemistry in some
situations. These interrelationships have been explored also by Thompson
(1982), who has related surface water pH to excess sulfate loading and the
rate of cation loss from watersheds (the Cation Denudation Rate or CDR). As
4-121
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with the prior models, her approach is restricted to relatively unbuffered
surface waters with low concentrations of organic acids in areas with acid-
resistant bedrock, till, and soils.
Thompson's model derives from charge balance and holds that a plot of excess
sulfate concentration vs the sum of base cation concentrations yields a
series of lines representing either constant bicarbonate concentration or
constant strong acidity. If C02 partial pressure is constant, then each
line also represents constant pH. If CDR (concentration x discharge *
watershed area) is plotted against atmospheric excess sulfate loading rate
(equivalent to acid loading) and if runoff is specified, then an equivalent
representation applicable to lakes or streams is generated (Thompson and
Mutton 1981, Thompson 1982) (see Figure 4-38).
A number of important assumptions apply to this approach. First, all non-sea
salt sulfate must come from atmospheric loading alone. Second, all sulfate
deposited in a watershed must flow through the watershed without being
retained (on a net basis). Third, all sulfate must be accompanied by protons
as it enters and leaves the watershed. The difficulties with each of these
assumptions and the everyday application of such a model have been thoroughly
described in the preceding discussions of the Almer/Dickson Relationship and
the Henriksen predictor nomograph. Another difficulty or necessary assump-
tion relates to both the constancy and quantification of PQQZ in any set of
waters to which the model may be applied. Significant variations in C02
partial pressures in surface waters are well known.
With regard to possible variations in cation leaching or weathering, Thompson
(1982) noted that CDR varies over short time scales (following discharge) but
that, "It is not known whether the CDR varies significantly from year to
year." The possible importance of such longer term variations to the
predictive use of such a model has been discussed above in relation to
Henriksen's predictor nomograph.
Another point worth considering is the fact that Thompson (1982) tested this
approach in some highly colored lakes and rivers of Nova Scotia (Figure
4-39). Although she noted that the pH values of these rivers "have been
thought to be dominated by naturally-occurring organic acids", Thompson
(1982) feels that "their low pHs can be explained quite well on the basis of
simple inorganic chemistry", as evidenced by the apparent agreement indicated
in Figure 4-39. A more direct way to resolve this question is through Gran
titrations for weak and strong acids. Apparently, such a study has not been
conducted. To the knowledge of this reviewer, the CDR model has not been
verified with any other data sets.
4.5.4 "Trickle-Down" Model
"Trickle-down" is the descriptive name given by Schnoor et al. (1982, 1983a,
b) to their mathematical model of the effects of acidic precipitation on the
alkalinity of surface and sub-surface waters. The name refers to the
"trickling-down" of acidic pollutants from the atmosphere first to the
terrestrial canopy, then to the soil surface, then to soil water etc... until
the acids are neutralized or leave the system in surface or sub-surface flow.
4-122
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Pco2 - 10~2'5
RUNOFF = 1 m yr
-1
sr 200
CO
o
1—4
5
co
s.
o
I
CNJ
I
cr
cu
cc
o
o
Intercepts
2.5
are
and
ACID LOAD (meq m-2 yr-1) or EXCESS S042' (yeq r1)
Figure 4-38.
A plot of the model that relates pH and sum of cations to
excess SO^- in concentration units, or pH and CDR (C02)
to rate of excess SO/^' loading in rate units. Note that
the author assumes a 10-fold supersaturation of C02, i.e.,
PCO? = 10"^ rather than PCO? = 10"3-^. Adapted from
Thompson (1982).
4-123
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WALLACE!
200
2
PpCO = 10~2'5
6.3
6.0
4.3
50 100
EXCESS S04Z" (meq
-2
150
yr-l)
Figure 4-39.
CDR plot for rivers with mean runoff near 1 m yr'l, 1973
excess $04^' loads, and mean or median river pH. Note that
the author assumes of 10-fold supersaturation of C02, i.e.,
PCO? = 10~2'5 rather than PCOo = 10'3-5. Adapted from
Thompson (1982).
4-124
C02
-------
The approach is based on continuity and the conservation of mass of a
single-state variable—alkalinity. In essence, the model consists of two
solutions (one time-variable, the other steady-state) to a mass-balance
equation for alkalinity in a lake. The mass-balance equation contains terms
for outflow of alkalinity from the lake, neutralization of acidity by lake
sediments, and inflow of alkalinity to the lake. This last term is inter-
esting in that it is written as a function of acidity loading and the
fraction of acid neutralized in the watershed. In general use, Schnoor et
al. (1983b) and Stumm et al. (1983) have concentrated on the steady-state
solution to the mass-balance equation.
To calculate a predicted steady-state alkalinity for a lake (neglecting
neutralization of acids by sediments) the following quantities need to be
known: the outflow rate of the lake, the precipitation rate (volume/time),
the precipitation acidity, and the "fraction" of acids neutralized in the
watershed. This last quantity is calculated from rainfall amount, precip-
itation acidity, and weathering rate; the weathering rate is calculated based
on pH and carbonate alkalinity and assumed to be at steady state. An
important question that must be answered before predicting a new steady-
state alkalinity value resulting from a change in loading is how loading
affects weathering. As noted in discussions of other models in this docu-
ment, this important question has yet to be precisely answered for field
situations.
As described above, predicted steady-state alkalinity values may be calcu-
lated algebraically from the solution to the mass-balance equation. Alterna-
tively, new values may be determined via nomograms presented by Schnoor et
al. (1983b) and Stumm et al. (1983). These nomograms are reproduced here as
Figures 4-40 and 4-41 [explicit use of the graphical method as well as data
for specific lakes, indicated as "a" through "h will not be discussed here;
see Schnoor et al. (1983b) and Stumm et al. (1983) for details]. To move
from one point (e.g., a current condition) to some other point (e.g., a
predicted condition) on the nomograms, the new loading rate as well as the
effect of changes in loading on weathering must be known.
As with all steady-state models, this "lumped-parameter" model is designed to
estimate a "final" alkalinity value resulting from a change in loadings; it
does not predict how long it will take to reach that value.
4.5.5 Summary of Predictive Modeling
As is evident in the preceding discussions, there is still much to learn
about a number of key factors that influence the ways in which lakes/water-
sheds respond to acidic deposition, and thus the ways in which these re-
sponses may be modeled and predicted, even on the most basic levels. Factors
that appear to be of primary importance but about which our knowledge is
still inadequate include: 1) the ability of soils to retain sulfur inputs
from atmospheric deposition; 2) the effects of acidic inputs on cation
exchange and leaching from soils; 3) the mobilization of aluminum compounds
from soils due to acidic deposition; 4) the effects of acidic inputs on
mineral weathering; and 5) the presence or absence of hysteresis in those
4-125
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100
A
O)
ID
C
£
yeq
-1
- 7.26
PH
Figure 4-40.
Steady-state model for alkalinity vs total acid deposition
concentration (ILaCy/Q) with iso-f, neutralization extent
lines. Values of pH are plotted for samples equilibrated
with atmospheric C02 (PC02 = 10~3-5 atm). Adapted from
Schnoor et al. (1983b).
a. 3 Minnesota BWCAW lakes (Warpaint, Agawato, Omaday)
b. 4 Wisconsin lakes 1979-80 (Sand, Greater Bass, Sugar
Camp, Ike Walton)
c. 3 Norwegian waters 1972-80 (Birkenes, Storgama, Langtjern)
d. 5 Swedish lakes 1979 (Gardsjon, St. Holmevatten, L.
Holmevatten, Bravatten, L. Otter)
e. 2 Northeastern U.S. waters 1979 (Woods Lake, Falls Brook)
g. 8 LaCloche Mountain lakes 1972 (Lumsden, Killarney,
Freeland, George, Kakakise, Norway, Threenarrow, OSA)
h. 7 Swiss Ticini lakes 1972 (Starlarescio, Orgnana, Piatto,
Zotta, Tomeo, Pianca, Cristallina)
4-126
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PH
6.0
5.5
5.0
4.5
4.0
600
500
400
300
n>
.0
Cv
50
100
150
200
ACID ADDED (yeq f1)
Figure 4-41.
Titration curve (dashed line) for the reaction pathway of
lake water beginning at point p. W (solid line) is the
chemical weathering (neutralization) rate in the water-
shed (including the sediments). Adapted from Stumm et al.
(1983).
4-127
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processes and their effects as a function of increasing or decreasing Inputs
of acids (Galloway et al. 1983a).
In short, predictive modeling of the acidification of surface waters is still
in an infant stage. Some interesting ideas have been put forth and some
progress is being made but there is still a very long way to go before any
model will be able to be used with quantitative confidence. Certainly none
of the four models discussed briefly here has been verified adequately for
"off-the-shelf" application in North American waters. Such an application
without a clear recognition and statement of all the assumptions and
limitations contained in these approaches would violate virtually every rule
concerning the prudent use of predictive models (Reckhow and Chapra 1981,
Bloch 1982).
4.6 INDIRECT CHEMICAL CHANGES ASSOCIATED WITH ACIDIFICATION OF SURFACE
WATERS
Acidic deposition is composed of NH4+, $042-, N03", H+, and basic cations.
The previous sections have discussed the chemical effects acidic deposition
has in aquatic systems by directly altering the concentrations of these same
chemicals. There are additional indirect effects on other chemicals.
Specifically, the addition of acidic deposition to terrestrial and aquatic
systems can disrupt the natural biogeochemical cycles of some metal and
organic compounds to such a degree that biological effects occur. The
following three sections discuss these chemical effects and assess the state
of our knowledge. The first section (4.6.1) focuses on metals in general;
the second (Section 4.6.2), specifically on aluminum. Elevated levels of
aluminum in acidified surface waters have been demonstrated to be toxic to
aquatic biota (Chapter E-5, Section 5.6.4.2) and thus are of particular
concern. Potential interactions between acidic deposition and organic carbon
cycles are discussed in Section 4.6.3.
4.6.1 Metals (S. A. Norton)
The impact of acidic deposition or, more broadly, atmospheric deposition on
metal mobility in aquatic ecosystems may be divided into four areas:
1) Increased loading of metals from atmospheric deposition to
terrestrial and aquatic ecosystems.
2) Direct effects of atmospheric deposition on metal release
rates from or to aquatic ecosystems.
3) Secondary effects of atmospheric deposition on metal release
rates from or to aquatic ecosystems—both positive and negative.
4) Changes in aqueous^speciation of metals and consequent
biological effects^
4-128
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4.6.1.1 Increased Loading of Metals From Atmospheric Deposition—In many
instances enhanced loadings of metals are associated with elevated levels of
NH4+, $042-, N03-, and H+ in acidic deposition. Although this
excess of metals is apparently related to industrial activities, nistone
measurements of metals in atmospheric deposition are not sufficient for
establishment of temporal trends. Indirect evidence for increasing
atmospheric deposition of metals is as follows:
a) Contemporary variations in atmospheric deposition of metals (e.g., Pb and
Zn) are closely related to the geographic distribution of fossil fuel
consumption, smelting, and transportation (by means of the internal the
internal combustion engine) (Lazrus et al. 1970). Where these sources are
absent, metal deposition rates are lower (Galloway et al. 1982b). Thus, as
fossil fuel consumption and other processes expand, injection of metals
into the atmosphere increases and atmospheric deposition increases.
b) Ombrotrophic peat bogs, those having no source of nutrients other than
precipitation, receive all their nutrients and non-essential metal from
atmospheric deposition. Some elements are relatively immobile (e.g., Pb)
and, after deposition, do not chemically migrate as the peat accumulates.
Increased concentrations of lead in recent peat in eastern Massachusetts
(up to 1.2 x over background) suggest increases in atmospheric deposition
of at least 3.5 x over the past few decades (Hemond 1980). Absolute
chronology in accumulating peat generally can only be estimated; thus
absolute increases cannot be rigorously established. Other elements
(e.g., Zn and Cu) are increased in concentration in modern peat as
compared to "old" peat, but chemical mobility at the low pH of peat
interstitial waters, variable redox conditions, and biological recycling
do not permit precise calculation of absolute increases of atmospheric
deposition of these metals.
c) 'Continuously1 accumulating snow is believed to record or reflect changes
in the chemistry of atmospheric deposition of metals. However, fractional
melting, ablation, erosion and deposition of snow, and other factors
obscure absolute deposition rates. Nonetheless, it is clear that the
deposition of Pb and Zn (fossil fuel-related elements) has greatly
accelerated over the last 100 to 150 years in areas as remote as Greenland
(Herron et al. 1976, 1977). The relative increases depend on background
(pre-pollution) values and the emission (and subsequent deposition) rates
for specific metals.
d) Galloway and Likens (1979) showed higher concentrations of Pb, Au, Ag, Zn,
Cd, Cr, Cu, Sb, and V in 'modern1 sediments relative to older sediments of
relatively undisturbed lakes. Norton et al. (1981a) and Johnston et al.
(1981) demonstrated that concentrations of Pb, Zn, Cu, Cd, and V are
higher in modern sediments (post-1850) than in older sediments and
established that the ubiquitous (in northern New England) and essentially
synchronous (ca. 1860-80) increases correlate with the initial rapid
increase in the consumption of fossil fuel in this country. Because these
lakes are relatively undisturbed, these changes are interpreted to be
caused by increases in the rate of atmospheric deposition of these metals,
starting prior to 1860.
4-129
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e) Hanson et al. (1982) have shown that Pb concentrations in the organic soil
horizons of high elevation spruce/fir forests of New England, New
Brunswick, and Quebec are related to the pH of precipitation. Low pH is
associated with high Pb. Lead in the northeastern United States is
probably derived from 2 major sources, fossil fuel burning and automobile
emissions (Lazrus et al. 1970). Consequently, Pb deposition rates may
vary independently of the pH of precipitation. Groet (1976) demonstrated
spatial variation in the northeastern United States of concentrations of
heavy metals in bryophytes, mosses, and liverworts (known concentrators of
atmospherically-deposited metals). Highest concentrations are related to
regional industrialization.
f) The litter, fermentation, and humic layers of organic soils of fir forests
represent successively longer time periods and progressively more decayed
material. The concentration of lead, which is chemically immobile
(probably because of adsorption), is highest in the fermentation layer
(but nearly the same as in the litter layer), suggesting increased
deposition of Pb (Reiners et al. 1975, Hanson et al. 1982). Although Pb
can be removed mechanically by erosion and vertical displacement, rates of
deposition can be derived if the age of litter is known and mechanical
erosion is nil. Siccama et al. (1980) studied white pine forest soils in
central Massachusetts collected at two times (separated by 16 years) and
found a higher rate of Pb accumulation in recent litter. Many workers
have demonstrated spatial and temporal trends for other elements (e.g.,
Zn) which parallel those for Pb, but increased deposition rates cannot be
assessed quantitatively because of the nonconservative nature of these
elements.
4.6.1.2 Mobilization of Metals by Acidic Deposition—The stoichiometry of
chemical weathering reactions and cation exchange and experimental evidence
(e.g., Cronan 1980), suggests that increasingly acidic deposition should
increase the release of cations (any positively charged aqueous species) from
soils and aquatic sediments. Empirical evidence from the United States for
accelerated release of cations due to acidic deposition over a long time
period, however, is rare. Oden (1976) cited evidence for long-term
increasing Ca concentrations in Swedish rivers, but long-term land use
changes on the scale of 10 to 100 years (including vegetational succession)
(Nilsson et al. 1982) may cause similar results (Section 4.4.3.3).
Paleolimnological evidence from sediment cores (Hanson et al. 1982) indicates
that detritus deposited in lakes has been, in undisturbed watersheds,
progressively more depleted in recent time with respect to easily mobilized
elements, e.g., Zn, Mn, Ca, and Mg. These decreases in concentration start
as early as about 1880 and are interpreted to result from increased leaching
of these elements from the terrestrial ecosystem. Similar changes are not
seen in areas that have only recently received acidic deposition (e.g.,
Swedish Lappland, Norton, unpub. data). Deposition rate and concentration
data for sediments from undisturbed lakes in New England and the Adirondack
Mountains of New York indicate continuously increasing values for Pb for all
lakes for about 100 years. The values for Zn increase continuously to the
4-130
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present for lakes with a pH > 6.0 and decrease in younger sediments for those
lakes with pH < 5.5 (see Section 4.4.3.2.2, Figure 4-30), suggesting recent
acidification of those lakes with decreasing In (as well as Ca, Mn, and
possibly Mg).
Field and laboratory soil lysimeter studies by Cronan and Schofield (1979)
and Cronan (1980) indicate that modern soil solutions have a chemistry (e.g.,
Al concentrations) that is inconsistent with the historical soil horizon
development. This is interpreted to be due to more acidic influx to the soil
from acidic deposition, causing Al leaching where before Al was accumulating
(Section 4.6.2.1).
Episodic decreases in the pH of surface waters (linked quantitatively to
meteorological events) are commonly accompanied by increases in dissolved Al
(Schofield and Trojnar 1980) and other elements, suggesting the direction of
changes to be expected in the mobilization of metals from soils, bedrock, and
sediments as precipitation becomes more acidic (Norton 1981).
Data sets for metal concentrations of lake waters versus pH suggest that,
because of solubility relationships, mobility of certain metals (Al, Zn, Mn,
Fe, Cd, Cu) should be relatively greatly increased with increasingly acidic
deposition (Norton et al. 1981b, Schofield 1976b, Wright and Henriksen 1978).
Other metals (K, Na, Ca, Mg), the concentration of which is in the > 0.1 ppm
range, will also be affected but to a lesser degree relative to iniTial con-
centrations.
Accelerated cation release (from aquatic sediments) has also been demon-
strated during experimental acidification of surface waters. In the field,
Hall and Likens (1980) observed increased release of Al, Ca, Mg, K, Mn, Fe,
and Cd due to artificial acidification of streams. In isolated columns in
lakes and in whole lake acidification experiments, Schindler et al. (1980b)
observed increased leaching of Fe, Mn.and Zn from the sediments. Andersson
et al. (1978), Hongve (1978), Davis et al. (1982), and Norton (1981)
demonstrated in laboratory sediment/water core microcosms that accelerated
leaching of metals from sediment occurs during acidification.
4.6.1.3 Secondary Effects of Metal Mobi 1i zati on—Secondary effects of acidic
deposition may lead to increased or decreased metal mobility. For example,
the release of Hg from sediments and soils and production of methyl mercury
may be promoted by more acidic waters (Wood 1980).
Secondary effects may be operative but have not been demonstrated. For
example, increases of Pb (as Pb2+) and S042' may result in immobi-
lization of both Pb2+ and S04 as the insoluble salt, PbS04-
Similarly Nriagu (1973) has suggested that excess Pb2+ may immobilize
P04 . This could cause a reduction in available phosphate for aquatic
ecosystems. Al sulfate minerals (Nordstrom 1982) are now suggested as being
a control on Al and/or S04 . Increased A13+ in acidified soil
waters could also immobilze phosphate (Section 4.6.2.5). Alternatively,
desorption from or solution of FeOOH from "B" soil horizons in well drained
soils could liberate adsorbed phosphate. These potentially important
4-131
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mechanisms have not been thoroughly investigated in the context of acidic
deposition. Very probably P04 availability will be strongly affected
by increased concentrations of Fe3+ and Al3+ in soil and in surface
waters.
4.6.1.4 Effects of Acidification on Aqueous Metal Speciation—The chemical
form of dissolved metals is important in determining the total mobility of a
metal and the biological effects related to acidification of aquatic
ecosystems. In general, most metals are complexed less at lower pH values
because less HC03-, C03 , OH- and other weak acid ligands are
present. Limits for concentrations of metals for toxicity to organisms
(Gough et al. 1979) are generally based on experiments where the water
chemistry is not well characterized, so such limits are probably excessively
high. Some toxicity limits have been defined for "soft" and "hard" water
(e.g., Howarth and Sprague 1978). The upper limits for toxicity for hard
water are generally much higher than for soft water, reflecting the probable
importance of speciation.
4.6.1.5 Indirect Effects on Metals in Surface Waters--The rate of deposition
of several metals from the atmosphere is increased due to anthropogenic
activities. The metals include Pb, Au, Ag, Zn, Cd, Cr, Cu, Sb, and V.
Primary and secondary effects of acidic deposition on metal mobility include
increased solubility of Al, Zn, Mn, Fe, Cd, Cu, K, Na, Ca, and Mg. These
metals are mobilized by acidic deposition both from the terrestrial system
and from lake sediments.
As aquatic systems acidify, speciation of metals changes. The direction of
changes is generally to a more biologically active species.
4.6.2 Aluminum Chemistry in Dilute Acidic Waters (C. T. Driscoll)
This section is intended to be a review of the literature on aluminum
chemistry in dilute acidic waters. While the general literature on the
chemistry of aluminum is applicable to this discussion it is also voluminous
and as a consequence beyond the scope of this document. While some of the
general literature on aluminum is discussed to illustrate principles, this
review largely addresses studies that are directly applicable to the effects
of acidic deposition.
4.6.2.1 Occurrence, Distribution, and Sources of Aluminum—Aluminum is the
third most abundant element within the earth's crust (Garrels et al. 1975).
It occurs primarily in aluminosilicate minerals, most commonly as feldspars
in metamorphic and igneous rocks and as clay minerals in well-weathered
soils. In high elevation, northern temperate regions, the soils encountered
are generally podzols (Buckman and Brady 1961). The podzolization process
involves mobilizing aluminum from upper to lower soil horizons by organic
acids leached from foliage as well as from decomposition in the forest floor
(Bloomfield 1957; Coulson et al. 1960a,b; Johnson and Siccama 1979).
Aluminum largely precipitates in lower soil horizons (Ugolini et al. 1977).
Ugolini et al. (1977) have observed that during podzolization little aluminum
mobilizes from the adjacent watershed to surface waters. Stumm and Morgan
4-132
-------
(1970) report a median aluminum value of 10 yg Al £-1 for terrestrial
waters, while Bowen (1966) gives an average concentration of 240 yg Al
r*- for freshwaters including bogs. It is noteworthy that values of
aluminum reported for circumneutral waters are generally greater than levels
predicted by mineral equilibria (Jones et al. 1974). Because of the tendency
for aluminum-hydroxy cations to polymerize through double OH bridging when
values of solution pH exceed about 4.5 (Smith and Hem 1972), a considerable
fraction of the "dissolved" aluminum reported in many analyses of natural
water having near-neutral or slightly acidic pH may consist of suspended
microcrystals of aluminum hydroxide. Hem and Roberson (1967) have shown that
crystals having a diameter near 0.1 ym were relatively stable chemically.
Filtration of samples through 0.4 ym porosity membranes, a common practice
in clarifying natural water prior to analysis, may fail to remove such
material (Kennedy et al. 1974). However, the concentrations of dissolved
aluminum are generally low in most circumneutral natural waters due to the
relatively low solubility of natural aluminum minerals.
Superimposed on the natural podzolization process is the introduction of
mineral acids from acidic deposition to the soil environment. It has been
hypothesized that these acids remobilize aluminum previously precipitated
within the soil during podzolization or held on soil exchange sites (Cronan
and Schofield 1979). Elevated levels of aluminum have been reported in
acidic waters within regions susceptible to acidic deposition (Table 4-11).
Many investigators have observed an exponential increase in aluminum
concentration with decreasing solution pH (Hutchinson et al. 1978, Dickson
1978a, Wright and Snekvik 1978, Schofield and Trojnar 1980, Vangenechten and
Vanderborght 1980, Hultberg and Johansson 1981, Driscoll et al. 1984). This
phenomenon is characteristic of the theoretical and experimental solubility
of aluminum minerals. Researchers have hypothesized several mechanisms for
the solid phase controlling aluminum concentrations in dilute water systems,
including poorly crystallized 1:1 clays (Hem et al. 1973) kaolinite (Norton
1976), aluminum trihydroxide (May et al. 1979, Johnson et al. 1981, Driscoll
et al. 1984), basic aluminum sulfate (Eriksson 1981) and exchange on soil
organic matter (Bloom et al. 1979). Johnson et al. (1981) and Driscoll et
al. (1984) compare and discuss solution characteristics of New Hampshire and
Adirondack waters, respectively, with the theoretical solubility of a variety
of aluminum minerals. Eriksson (1981) observed that calculated values of
aquo aluminum in soil solutions from Sweden were similar to values predicted
from mineral solubility reported by van Breemen (1973) for Al(OH)S04, at a
given pH. This lead Eriksson (1981) to suggest that atmospheric deposition
of sulfate has acidified and transformed aluminum oxides to basic aluminum
sulfate in Swedish soils. Unfortunately, Eriksson (1981) failed to consider
fluoride, sulfate, and organic complexation reactions when computing aquo
aluminum levels. Therefore, as suggested by Nordstrom (1982), it is doubtful
that aluminum sulfate minerals (e.g., jurbanite, alunite, basaluminite)
control aquo aluminum levels in waters acidified by acidic deposition. In
actuality it is extremely difficult to identify a specific solution con-
trolling phase. Analysis of soils and sediments by x-ray diffraction has
failed to confirm the presence of hypothesized solution controlling minerals
of aluminum (Driscoll et al. 1984).
4-133
-------
TABLE 4-11. ALUMINUM CONCENTRATIONS IN DILUTE ACIDIC WATERS
-P.
i
co
-P.
Location
Lakes
Sweden
Norway
Scotland
Belgium
USA
USA
USA
Canada
Canada
Streams
USA
USA
Description
Swedish West Coast, 1976
Regional Survey, 1974-77
Southwestern Scotland, 1979
Moorland pools Northern
Belgium, 1975 - 1979
Adirondacks, 1977-1978
New England, 1978-1981
New England, 1978-1980
Ontario various locations, 1980
Sudbury, Ontario
Adirondacks, 1977-1978
Adirondacks, 1977
PH
4
4
4
3
3
4
4
4
4
4
4
.0
.2
.4
.5
.9
.0
.2
.1
.3
.0
.4
Range
- 7
- 7
- 6
- 8
- 7
- 8
- 7
- 6
- 7
- 7
- 6
.4
.8
.4
.5
.2
.2
.0
.5
.0
.6
.5
Al
vd
10
0
25
300
4
0
0
6
150
92
100
Range
Al £-!
- 670
- 740
- 310
- 8000
- 850
- 579
- 440
- 856
- 1150
- 1170
- 1000
Reference
Dickson 1978a
Wright et al .
Wright et al .
Vangenechten
Vanderborght
Driscoll 1980
1977
1977
and
1980
Haines and Akielaszek
1983
Norton et al .
Kramer 1981
1981a
Scheider et al . 1975
Driscoll 1980
Schofield and
USA
New England, 1978-1981
4.1 - 7.7
14 - 385
Trojnar 1980
Haines and Akielaszek
1983
-------
TABLE 4-11. CONTINUED
u>
en
Location
Streams (cont.)
USA Hubbard
Description
Brook stream order 1
2
3
2
3
3
4
4
5
average
pH Range
4.73
4.94
5.09
5.19
5.54
5.46
5.51
5.58
5.68
4.90
Al Range Reference
yg Al £-!
710 Johnson et al .
320 1981
210
200
150
190
180
160
150
230
Groundwaters
Sweden
West Coast, 1977-1978
3.8 - 5.7
100 - 2600
Hultberg and
Johansson 1981
USA
Hubbard Brook seepwater, 1979
4.6 - 6.5
0 - 700
Mulder 1980
-------
4.6.2.2 Aluminum Speciatlon—Dissolved monomeric aluminum occurs as aquo
aluminum, as wellas hydroxide, fluoride, sulfate, and organic complexes
(Roberson and Hem 1969, Lind and Hem 1975). Past investigations of aluminum
in dilute natural waters have often ignored non-hydroxide complexes of
aluminum (Cronan and Schofield 1979, N. M. Johnson 1979, Eriksson 1981).
Driscoll and coworkers (Driscoll 1980; Driscoll et al. 1980, 1984) have
fractionated Adirondack waters into inorganic monomeric aluminum, organic
monomeric aluminum, and acid soluble aluminum. They observed that inorganic
monomeric aluminum levels increased exponentially with decreasing solution
pH. Organic monomeric aluminum levels were strongly correlated with total
organic carbon (TOO concentration but not pH. Acid soluble aluminum levels
were relatively constant and not sensitive to changes in either pH or TOC.
Driscoll et al. (1984) reported that organic complexes were the predomi-
nant form of monomeric aluminum in Adirondack waters, on the average account-
ing for 44 percent of monomeric aluminum. Aluminum fluoride complexes were
the second major form of aluminum and the predominant form of inorganic
monomeric aluminum, accounting for an average of 29 percent of the monomeric
aluminum. Aquo aluminum and soluble aluminum hydroxide complexes were less
significant than aluminum fluoride complexes. Aluminum sulfate complexes
were small in magnitude.
4.6.2.3 Aluminum as a pH Buffer—Dilute water systems are characteristically
low in dissolved inorganic carbon (DIG) due to limited contact with soil.
Because dilute waters are inherently low in DIG, they are limited with re-
spect to inorganic carbon buffering capacity. Consequently, non-inorganic
carbon acid/base reactions, such as hydrolysis of aluminum and protonation/
deprotonation of natural organic carbon, may be important in the pH buffering
of dilute waters.
Several researchers have investigated organic carbon, weak acid/base systems
in dilute waters. Dickson (1978a) observed that elevated levels of aluminum
increased the base neutralizing capacity (BNC) of Swedish lakes. Waters were
strongly buffered by aluminum in the pH range 4.5 to 5.5. The BNC of
aluminum was particularly evident when acidified lakes were treated with base
(limed). Aluminum BNC was comparable in magnitude to hydrogen ion and
inorganic carbon BNC; therefore, the presence of aluminum substantially
increased base dose requirements and the cost associated with the restoration
of acidified lakes.
Johannessen (1980) investigated non-hydrogen/inorganic carbon buffering in
Norwegian waters. While reiterating the importance of aluminum as a buffer
in dilute acidified waters, she also evaluated the role of natural organic
acids. Natural organic matter reduced the degree to which aluminum
hydrolyzed in the pH range 5.0 to 5.5, presumably due to complexation re-
actions, and therefore decreased the buffering of aluminum. Natural organic
matter also participated in proton donor/acceptor reactions; the extent to
which total organic carbon (TOC) would dissociate/associate protons was 7.5
yeq per mg organic carbon. Johannessen (1980) concluded that organic
carbon was the most important weak acid/base system in acidic Norwegian
waters because of the high organic carbon concentration relative to aluminum.
4-136
-------
Glover and Webb (1979) evaluated the acid/base chemistry of surface waters in
the Tovdal region of southern Norway. The BNC of hydrogen ion was small
compared to the BNC of weak acid systems. These investigators suggested that
of the total weak acid BNC, 40 to 60 ueq ir1 could be attributed to
dissolved aluminum and silicon, while 20 to 50 yeq £-1 could be
attributed to natural organic acids. Solution titrations were characterized
as having a major proton dissociation constant (Ka) 1 x 10-6 to 5 x 10-?,
in addition to some less well defined ionization at higher pH values.
In a comparable study, Henriksen and Seip (1980) evaluated the strong and
weak acid content of surface waters in southern Norway and southwestern
Scotland. In addition to a titrametric analysis, the aluminum, dissolved
silica, and TOC content of water samples were determined. Weak acid
concentrations, determined by a Gran (1952) calculation, were evaluated by
multiple regression analysis. Most of the variance in the weak acid
concentration could be explained by the aluminum and TOC content of the
waters. Thus, it was concluded that the weak acid content of acidified lakes
in southern Norway and Scotland was largely a mixture of aluminum and natural
organic acids.
Driscoll and Bisogni (1984) quantitatively evaluated weak acid/base systems
buffering dilute acidic waters in the Adirondack region of New York State.
Natural organic acids were fit to a monoprotic proton dissociation constant
model (pKa = 4.41), and the total organic carbon proton dissociation/
association sites were observed to be empirically correlated to TOC
concentration. Aquo-aluminum levels, calculated from field observations,
appeared to fit an aluminum trihydroxide solubility model.
Calculated buffering capacity (B) is plotted against pH in Figure 4-42 for
a hypothetical system that has some properties in common with Adirondack
waters (Driscoll and Bisogni 1984). Buffering capacity is defined as the
quantity of strong acid or base (mols £-1) which would be required to
change the pH of a liter of solution by one unit. Conditions specified for
the construction of Figure 4-42 are indicated in the figure title. Aluminum
species may dominate the buffer system at low pH if these conditions are
fulfilled, suggesting that the lower limit of pH observed in acidic waters
with elevated aluminum levels may be controlled by the dissolution of
aluminum. At higher pH values the buffer system is dominated by inorganic
carbon and would be even more strongly dominated if carbonate solids were
present.
Note that aluminum polymeric cations and particulate species that may occur
in acidic solutions provide some solution buffering (both ANC and BNC),
However, these large units may be slow to equilibrate with the added titrant.
Therefore, ANC and BNC determinations have limitations in acidic waters due
to heterogeneity phase problems.
4.6.2.4 Temporal and Spatial Variations in Aqueous Levels of Aluminum--
Pronounced temporal and spatial variations in levels of aqueous aluminum have
been reported for acidic waters. Schofield and Trojnar (1980) observed that
high aluminum levels occurred during low pH events in streams, particularly
during snowmelt. Driscoll et al. (1980) also observed this phenomenon and
4-137
-------
CQ
Q.
ALUMINUM
WATER
ORGANIC SOLUTES
CARBONATE
PH
Figure 4-42.
Buffer capacity diagram for dilute Adirondack water systems
(Driscoll and Bisogni 1984). Equilibrium with aluminum
trihydroxide (pKso = 8.49), organic solutes (CTorg =
2 x 10"5, pKorg =4.4) and atmospheric carbon dioxide
= 10~3-5 atm) were assumed.
4-138
-------
attributed aluminum increases to inorganic forms of aluminum. During low
flow conditions, neutral pH values were approached in streams (pH 5.5 to 7.0)
and inorganic monomeric aluminum levels were low. During summer months,
levels of TOC in streams increased and organically complexed aluminum levels
increased. As mentioned previously, levels of organic monomeric aluminum
were strongly correlated with surface water TOC (Driscoll et al. 1984).
Johnson et al. (1981) studied temporal and spatial variations in aluminum
chemistry of a first-through-third order stream system in New Hampshire.
Observations of temporal variations in aluminum were similar to those
reported for the Adirondacks (Driscoll et al. 1980, Schofield and Trojnar
1980). Johnson et al. (1981) reported decreases in hydrogen ion and aluminum
levels with increasing stream order. They suggested a two-step process for
the neutralization of acidic deposition. Mineral acidity entering the
ecosystem from atmospheric deposition was converted to a mixture of hydrogen
ion and aluminum BNC (acidity) in headwater streams and was subsequently
neutralized through the dissolution of basic cation (Ca2+, Mg2+, Na+,
K+) containing minerals within the soil environment.
Driscoll (1980) has evaluated temporal and spatial variations in aluminum
levels in acidic lakes. During summer stratification, monomeric aluminum
levels were low in the upper waters and increased in concentration with
depth. Low aluminum levels reported in the upper waters during the summer
coincided with elevated pH and ANC values. The increased pH and ANC values
were attributed to algal assimilation of nitrate (Brewer and Goldman 1976).
During ice cover, pH (and ANC) values were low and aluminum levels high
directly under the ice. The pH values increased and aluminum values
decreased with depth. The clinograde distribution of pH and aluminum
observed during ice cover periods has been attributed to reduction processes
in sediments (e.g. denitrification). These processes generate ANC, which
diffuses into the lower waters. During fall and spring turnover, aluminum is
evenly distributed throughout the water column of acidic lakes. Aluminum
levels were particularly high during the spring season because of inputs of
low pH, high aluminum stream water associated with spring snowmelt.
Few studies have considered temporal and spatial variations in aluminum
chemistry of groundwaters. Hultberg and Johansson (1981) have observed
acidification events in groundwater chemistry in Sweden. They hypothesized
that much of the atmospheric input of sulfur was retained within the
terrestrial ecosystem as reduced sulfur forms. During extremely dry con-
ditions, the water table was lowered and pools of reduced sulfur within the
soil become oxidized by molecular oxygen entering the zone. Very low pH
values (< 4.0) and very high aluminum levels (> 40 mg Al «,-!) have been
reported in groundwater by Hultberg and coworkers (Hultberg and Wenblad 1980,
Hultberg and Johansson 1981) when a prolonged dry period was followed by a
rainfall event. It is difficult to attribute conclusively groundwater acid-
ification to atmospheric deposition of sulfate. A possible source of the
acidity in the groundwater studied by Hultberg and Johansson (1981) was the
oxidation of reduced iron minerals, likely to have been present naturally in
the upper part of the zone of saturation. This oxidation would have occurred
when the water table declined due to dry weather and molecular oxygen entered
the zone. The hydrogen ion produced by iron oxidation with molecular oxygen
4-139
-------
would not be significantly mobilized in the groundwater until the water table
increased again to a more normal level.
4.6.2.5 The Role of Aluminum in Altering Element Cycling Within Acidic
Vlaters--In acidic water systems conditions of supersaturation with respect to
aluminum trihydroxide have been reported (Driscoll et al . 1984). During
conditions of supersaturation, aluminum will hydrolyze, forming particulate
aluminum oxyhydroxide. The acid-soluble aluminum fraction mentioned earlier
would include the microcrystalline hydroxide particles and their polymeric
hydroxycation precursors. Smith and Hem (1972) observed that during the
polymerization process, aluminum hydroxide units displayed metastable ionic
solute behavior until they contained from 100 to 400 aluminum atoms. When
particles developed to that size their behavior was characteristic of a
suspended colloid. Microcrystalline particles have a very large specific
surface area and may adsorb or co-precipitate organic and inorganic solutes.
The cycling of orthophosphate (Huang 1975, Dickson 1978a), trace metals (Hohl
and Stumm 1976) and dissolved organic carbon (Dickson 1978a, Davis and Gloor
1981, Driscoll et al . 1984, Hall et al . 1984) within acidic surface waters
may be altered by adsorption on aluminum oxyhydroxides. However, few studies
have addressed this specific hypothesis.
Huang (1975) studied the adsorption of orthophosphate on T-AleOs- He
observed an adsorption maximum at pH 4.5. While Huang (1975) studied the
adsorption of high levels of orthophosphate (10-4 to 10-3 M) t much higher
than would be observed in natural dilute water systems, his observations of
phosphate aluminum interactions may be generally applicable.
Dickson (1978a) observed that when acidic lake water, elevated in aluminum,
was supplemented with orthophosphate (50 and 100 ug P £-1), dissolved
phosphorus was removed from solution. The removal of phosphorus was most
pronounced at pH 5.5. Dickson (1978a, 1980) suggested that aqueous aluminum
may substantially alter phosphorous cycling within acidic surface waters
through adsorption or precipitation reactions. This hypothesis is noteworthy
because phosphorus is often the nutrient limiting algal growth in dilute
surface waters (Schindler 1977). Any decrease in aqueous phosphorus induced
by adsorption on aluminum oxyhydroxides may result in a decrease in algal
growth and an accompanied decrease in algal generated ANC (see Section
4.7.2). Any decrease in ANC inputs would result in an aquatic ecosystem more
susceptible to further acidification.
Aluminum forms strong complexes with natural organic matter (Lind and Hem
1975). Complexation substantially alters the character of natural organic
acids. Driscoll et al . (1984) observed that DOC was removed from the water
column of an acidic lake after CaCOa addition. They hypothesized that DOC
sorbed to the particulate aluminum that had formed within the water column
shortly after base addition. Driscoll (1980) observed decreases in water
column TOC during conditions of supersaturation with respect to A1(OH)3 in
an acidic lake. He hypothesized that natural organic carbon was scavenged
from solution by particulate aluminum formed in the water column. Davis
(1982) has studied the adsorption of natural dissolved organic matter at the
Y-Al203/water interface. He observed that natural organic matter
adsorbs by complex formation between the surface hydroxyls of alumina and
4-140
-------
acidic functional groups of organic matter. Davis (1982) indicated that DOC
adsorption was maximum at pH 5. Davis and Gloor (1981) reported that DOC
associated with molecular weight fractions greater than 1000 formed strong
complexes with the alumina surface, but low molecular weight fractions were
weakly adsorbed. Davis (1982) suggests that under conditions typical for
natural waters almost complete surface coverage by adsorbed organic matter
can be anticipated for alumina. Organic coatings may be important with
respect to subsequent adsorption of trace metals and anions.
Hall et al. (1984) observed a decrease in DOC levels of a third order stream
in New Hampshire after aluminum chloride (A1C13) addition. In addition, a
reduction in surface tension occurred at the air-stream interface and was
attributed to a decrease in the solubility of DOC due to interactions with
aluminum.
DOC loss to acidic waters is significant in several respects. DOC represents
a weak base that serves as a component of solution ANC (Johannessen 1980,
Driscoll and Bisogni 1984). DOC also serves as an aluminum complexing
ligand. Complexation of aluminum by organic ligands mitigates aluminum
toxicity to fish (Baker and Schofield 1980). Therefore, any loss of DOC may
translate to an environment less hospitable to fish.
4.6.3 Qrganics (C. S. Cronan)
4.6.3.1 Atmospheric Loading of Strong Acids and Associated Organic Micro-
poll utants--Thisfirst subsection deals with £fie association (but h~6~t
necessarily interaction) between anthropogenic strong acids and organic
micropollutants introduced to aquatic systems via long-range transport and
wet/dry deposition processes. Methods for isolating and characterizing
organic micropollutants in natural samples have been described by Gether et
al. (1976) and Heit et al. (1980). These methods were used by Lunde et al.
(1976) to identify a wide range of organic pollutants in rain and snow
samples from Norway, including alkanes, polycyclic aromatic hydrocarbons
(PAH's), phthalic acid esters, fatty acid ethyl esters, and many other
chemicals of industrial origin. Concentrations ranged from one to several
hundred ng £-1, with polychlorinated biphenyl (PCB) concentrations
registering five times higher than freshwater or seawater.
In a related study, Alfheim et al. (1978) examined the access of certain
non-polar organic pollutants to lakes and rivers in Norway. Results
indicated that PCB concentrations in water samples from a lake in southern
Norway were considerably lower than in melted snow from the same area. Two
explanations were offered to account for these observations: (1) the PCB's
in the water column may have been associated with particulate matter,
preventing them from being detected in the dissolved phase, and (2)
terrestrial humic substances may have complexed the PCB's and related
pollutants, thereby reducing their leaching into lakes and rivers.
The studies by Heit et al. (1981) focused on the historical patterns of
organic pollutant deposition to remote Adirondack lakes. Using lake sediment
cores and advanced analytical techniques, they found the following results.
First, all of the nonalkylated 3- to 7-ring parental PAH's, with the
4-141
-------
exception of perylene, decreased in concentration with sediment depth.
Surface concentrations of many of these compounds approached or exceeded
levels reported for sediments from urban and industrialized areas, while
baseline levels lower in the core were similar to those reported for pristine
areas such as in northern Ontario. Overall, the data indicated that all of
the parental PAH compounds except perylene entered these Adirondack lakes
primarily through anthropogenic rather than natural processes.
These investigations by Alfheim et al. (1978, 1980) and Heit et al. (1981)
have shown that a broad range of organic micro-pollutants may originate in
industrial centers and be carried downwind to remote ecosystems by long-range
atmospheric transport. Thus, similar patterns and processes may contribute
to the atmospheric transport and deposition of both anthropogenic strong
acids and organic micro-pollutants.
4.6.3.2 Organic Buffering Systems—Organic and/or aluminum weak acid buffer
systems may dominate the acid-base chemistry of surface waters in watersheds
characterized by the following kinds of features: granitic bedrock, thin or
impermeable surficial deposits, steeper slopes, high water tables, or
extremely permeable siliceous surficial deposits. In such soft water
ecosystems, organic and aluminum weak acids may provide the only buffering
protection against further acidification by anthropogenic strong acids.
Likewise, natural humic materials may themselves have sufficiently low pka
constants that they contribute to the free acidity of surface waters.
Organic weak anions may be particularly significant in providing ANC below pH
5.0, with the greatest buffer intensity for the organics exhibited in the
range of pH 4.5 (Figure 4-42) (Driscoll 1980).
The organic species responsible for contributing to the buffer capacity of
these soft water lakes include a range of hydrophilic and hydrophobic, low
and high molecular weight compounds. These organic solutes may range from
simple carboxylic acids like malic acid to complex polyphenolic compounds
like the model fulvic acid described by Schnitzer (1980). On the average,
these organic acids in natural waters might be expected to contribute 5 to 10
yeq of anionic charge per mg carbon (Driscoll 1980; Cronan, unpub. data),
and perhaps 5 to 20 yeq per mg organic carbon in total acidity (Schnitzer
1978, Henriksen and Seip 1980) (Section 5.2.1, Chapter E-5). Historically,
organic acid buffer systems were probably relatively common in soft water
aquatic systems. However, the relative importance of aluminum buffering
Section 4.6.2.3) may have increased recently in those soft water lakes that
have experienced modern acidification from atmospheric deposition (Henriksen
and Seip 1980).
4.6.3.3 Organo-Metallic Interactions—Acidification of surface waters may
affect metal-organicassociationsand trace metal speciation. Stability
constants for metal-fulvic acid (FA) complexes have been shown to decrease
with decreasing pH. For example, the conditional stability constant for
Pb2+-FA at pH 5.0 is 104-1, whereas it is 10*.6 at pH 3.0; likewise,
the Zn2+-FA stability constant at pH 5.0 is 103-7, but is 1Q2A at pH
3.0 (Schnitzer 1980). Because of this effect of pH on metal-organic
complexation, one might expect lake acidification to result in decreased
concentrations of organically-complexed metals and correspondingly higher
4-142
-------
concentrations of free inorganic trace metals. Simultaneously, the decreases
in pH could lead to increased protonation of organic acid functional groups,
thereby increasing the hydrophobic character of the organic acids. This
process could affect the adsorption of humic materials on mineral surfaces
and could also affect interactions between humic/fulvic monomers. The net
result of this could be to increase clay interlayer adsorption of fulvic
acids (with associated clay degradation) and to increase the polymerization
and settling of aquatic humic materials (Schnitzer 1980).
Along similar lines, there may be very important biological consequences
resulting from acidification of natural waters containing metal-organic
complexes. Driscoll et al. (1980) and others have already shown that free
inorganic species concentrations of trace metals like aluminum are signi-
ficantly more toxic than are the organically-complexed forms. Thus, where
atmospheric deposition leads to a shift from organically complexed to free
inorganic species of trace metals, there may be attendant impacts on aquatic
biota.
4.6.3.4 Photochemistry--Another interaction that has been described is the
effect of decreasing pH on the coloration or light absorption of aquatic
humic materials. For instance, Schindler (1980) and Schindler and Turner
(1982) found that lake coloration and extinction coefficients decreased with
decreasing pH, even though no measurable change in the DOC occurred. This
change in lake transparency resulted in an increase in primary productivity
in the experimental lake. In addition, the acid-induced increases in
transparency accelerated the rates of hypolimnion heating and thermocline
deepening; at the same time, there was no significant effect on the lake's
total heat budget. In terms of processes, the data were interpreted to
indicate that acidification caused a qualitative change in the structure of
aquatic humus and its ability to absorb light. Aimer et al. (1978) also
found evidence of changes in lake transparency associated with lake
acidification in Sweden; however, they observed lower concentrations of DOC
in transparent acidified lakes. According to their data, this scavenging of
organic carbon from the lake water column may have been largly due to the
formation of insoluble organic-aluminum coloids and the subsequent
sedimentation of these particulates to the lake bottom.
4.6.3.5 Carbon-Phosphorus-Aluminum Interactions—The potential impact of
acidic deposition upon aluminum leaching and phosphorus availability has been
discussed in Section 4.6.2.5 and described by Dickson (1980) and Cronan and
Schofield (1979). As Dickson (1980) has shown experimentally, increased
concentrations of inorganic aluminum in freshwaters may cause increased
precipitation of aluminum phosphates from the water column, resulting in
decreased biological availability of phosphorus. However, where humic
materials are present, the organic ligands will tend to bind the aluminum
preferentially, leaving the phosphorus uncomplexed. Therefore, one would
assume that where one finds increased concentrations of aquatic humic
materials these will tend to decrease the toxic potential of aluminum leached
from soils and will tend to preserve the availability of phosphorus in
aluminum-rich waters.
4-143
-------
4.6.3.6 Effects of Acidification on Organic Decomposition In Aquatic
Systems—Lake and stream acidification associated with atmospheric deposition
may also cause reductions in the rate of organic matter turnover and may
ultimately lead to decreased nutrient cycling and availability (Chapter E-5,
Sections 5.3.2.1 and 5-8). Traaen (1980) found that organic matter decom-
position was retarded at pH 4-.0 to 4.5 compared to control streams and
suggested that this effect could be important for lakes dependent upon
allochthonous inputs of carbon. Friberg et al. (1980) observed that leaf
litter decay was much slower in an acid stream (pH 4.3 to 5.9) than in a
paired stream at pH 6.5 to 7.3. This was interpreted to indicate that stream
acidification caused biotic disturbances among the aquatic decomposer pop-
ulations. Finally, Francis and Hendrey (1980) compared the decomposition
rates for leaf litter in three nearby lakes at pH 5.0, 6.0, and 7.0. Results
indicated that decomposition of beech leaves was inhibited considerably and
bacterial populations were approximately an order of magnitude lower in the
most acidic lake. These studies suggest a need to investigate what holistic
import reduced organic matter turnover in acidified aquatic systems will
have.
4.7 MITIGATIVE STRATEGIES FOR IMPROVEMENT OF SURFACE WATER QUALITY
(C. T. Driscoll and G. C. Schafran)
4.7.1 Base Addition
The most effective means of regulating acidification would be to control
hydrogen ion inputs. For atmospheric inputs this involves many political,
social, economic, and energy related considerations. An alternative strategy
is to symptomatically treat acidified waters by chemical addition. Various
substances have been proposed for use as neutralizing agents (Grahn and
Hultberg 1975); however only lime (CaO, Ca(OH)2) and limestone (CaCOa)
have been used to any extent. Two base addition strategies have been
practiced: direct lake addition and watershed/stream addition. While direct
lake addition is the less expensive approach, the relative effectiveness of
the two strategies has not been evaluated. In addition, the positive and
negative consequences of these strategies have not been fully evaluated.
A variety of methods for the treatment of acidic waters associated with mine
drainage have been researched and developed (Hodge 1953, Pearson and
McConnell 1975a,b). Because mine drainage is often extremely acidic and
contains elevated levels of hydrolyzing metals it is extremely difficult to
extrapolate base addition concepts and technology developed for mine drainage
to dilute acidic waters. Therefore, this critical assessment will address
only base addition to dilute water systems. Fraser et a!. (1982) and Fraser
and Britt (1982) compiled a detailed review of base addition technology and
effects that should be referred to for information beyond the scope of this
document.
4.7.1.1 Types of Basic Materials—Several types of basic materials have been
used or proposed for neutralizing acidified surface waters. These materials
include calcium oxide, calcium hydroxide, calcium carbonate, sodium
carbonate, olivine, fly ash, and industrial slags (Grahn and Hultberg 1975).
4-144
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There are many considerations in selecting a base material to be used in
neutralization. Scheider et al. (1975) have summarized these considerations.
1) It must be readily available in large quantities.
2) It should be relatively inexpensive.
3) It must be safe to handle and store using conventional safety
precautions.
4) It should have a high neutralization potential; i.e. a small
quantity of chemical should be capable of neutralizing a large
quantity of water.
5) Adding a known quantity of chemical must produce a predict-
able change in pH. This is critical if pH sensitive organisms
are already living in the lake.
6) It must be amenable to a relatively simple application
technique such that a large quantity of chemical could be
applied in a short period of time with a minimum of labor and
equipment.
7) It must provide for a natural deficiency in the aqueous acid
neutralizing capacity; i.e., it should be a normal component of
the pH buffer system.
8) It should not initiate any significant ion exchange process in
the lake sediment which could impair the quality of the lake
water.
9) It must not add any extraneous contaminants to the lake water.
Calcium oxide (quicklime, CaO) and calcium hydroxide (hydrated lime,
Ca(OH)2) have been used to neutralize acidified surface waters. These
materials are relatively inexpensive and effective. Lime is generally used
in a powdered form and is very soluble when added to water. Because lime is
a soluble strong base, it readily increases the pH of dilute solutions. If
the solution is in contact with atmospheric carbon dioxide after strong base
addition, the pH will slowly decrease. This response occurs because
atmospheric carbon dioxide will dissolve into solution, neutralize the
hydroxide, and eventually form a bicarbonate solution:
C02 introduction
+
Ca2+ + 20H" = Ca2+ + 20H" + 2C02 = Ca2+ + 2HC03"
Acidic waters generally have a low aqueous buffering capacity. As a result,
substantial increases in pH will occur upon addition of typical quantities of
strong base (200 to 400 yeq £-1 yr-1). Lake water pH values which
were below 5.0 prior to neutralization may increase to above 10.0 immediately
4-145
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after strong base addition. This change may result in pH shock to organisms.
These problems are accentuated within certain microenvironments, particularly
if mixing is incomplete. As a result, dosage control must be carefully
monitored.
Calcium oxide is an extremely corrosive material that generates considerable
heat when contacting water, which makes handling and storage very difficult.
Calcium hydroxide is less hazardous and does not generate heat upon contact
with water.
Calcium carbonate, commonly referred to as limestone, is a slightly soluble
base. Dissolution of limestone is slow, and a maximum pH of 8.3 is realized
when an aqueous system is in equilibrium with CaC03 and atmospheric C02
(Stumm and Morgan 1970). The dissolution kinetics of limestone are a
function of solution characteristics, impurities in the stone, and the
surface area of the stone (Pearson and McDonnell 1975a). Limestone commonly
contains a significant amount of magnesium (often called dolomitic
limestone). The greater the magnesium component in the limestone the slower
the dissolution rate. For applications to acidic surface water, enhanced
dissolution rates of slightly soluble bases are generally desirable.
Therefore, it is best to use a high purity stone (e.g., low magnesium
content). Limestone can be obtained in a variety of sizes. Powdered
limestone (agricultural limestone, 0 to 1 mm) is often used in water neu-
tralization efforts. Dissolution is enhanced because of the large surface
area associated with the small particles. Larger stone (0.5 to 2 cm) may be
used for limestone barriers in streams (Section 4.7.1.3.2) or limestone
contactors in springs.
An important consideration with regard to limestone dissolution is solution
characteristics. Dissolution rates are greatest in solutions of low pH, low
dissolved inorganic carbon, and low calcium. This condition is character-
istic of dilute acidified waters. Another important consideration is the
presence of hydrolyzing metals (Al, Fe, Mn) and dissolved organic carbon.
Upon increases in pH, these components may deposit on the surface of the
stone, inhibiting dissolution and therefore decreasing the effectiveness of
the base. Pearson and McDonnell (1975a) observed that the dissolution rate
of CaC03 decreased by up to 80 percent when CaC03 was coated with iron
and aluminum.
Calcium carbonate is generally favored for use as a base because inorganic
carbon is directly supplied upon dissolution and dissolution rates are
relatively slow. Aquatic organisms are less prone to pH shock with CaC03
treatment than with strong base addition.
Sodium carbonate (Na2C03, soda ash) is a soluble base which has been used
as a neutralizing agent (Lindmark 1981). Sodium carbonate is readily soluble
and directly applies dissolved inorganic carbon to solution. Therefore, it
is an effective base because there are minimal losses due to incomplete dis-
solution while fluctuations in pH are less extreme. Sodium carbonate is
generally an expensive base and therefore might not be used in lieu of
calcium base sources [CatOHig, CaC03J (see Section 4.7.1.2.2).
4-146
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Olivine (Mg, Fe)2 5104, is a natural silicate mineral that has been used
in neutralization efforts (Hultberg and Andersson 1982). Olivine is a
continous reaction series in which magnesium and ferrous iron can freely
substitute for each other. Upon dissolution of Fe2$i04» iron will
oxidize and precipitate as Fe(OH)3 and thereby contribute to the acidi-
fication of water. Therefore, the effectiveness of olivine as a neutralizing
agent increases with increasing magnesium content. Olivine is a slightly
soluble mineral; therefore, dissolution characteristics and application
difficulties associated with biological, chemical fouling will be in some
ways similar to those associated with CaCOa-
Fly ash is a material of diverse chemical composition. Western coals have
been found to produce fly ash that is characteristically basic (enriched in
calcium) while combustion of eastern coals generally results in an acidic fly
ash (enriched in iron) (Edzwald and DePinto 1978). Basic fly ash has been
shown to be effective in the neutralization of acidic surface waters.
Neutralization by fly ash is accomplished by the release of hydroxyl ion
rather than inorganic carbon to solution.
Fly ash is a waste byproduct so finding a way to use it is desirable. Waste
deposits of basic fly ash are primarily located in the midwestern United
States while most of the acidic waters are located in the northeast. Costs
of transporting fly ash would probably be prohibitive and certainly less
economical than using alternative neutralizing agents located in the
northeast. Another problem associated with fly ash is trace metal contam-
ination. Edzwald and DePinto (1978) have indicated that release of trace
metals to solution from fly ash is comparable to that released from sediments
upon acidification to pH 4.0.
It has been proposed that industrial slags could be used in the neutrali-
zation of acidic waters (Grahn and Hultberg 1975). One type of slag
formation is the use of calcium carbonate to produce metals from ores. Basic
slags formed in this and other processes vary considerably with respect to
physical and chemical properties (Grahn and Hultberg 1975). Basic slags are
largely composed of calcium (CaO) and silicon (SiOg) oxides. While basic
slags may contain similar calcium (CaO) levels, dissolution rates and
therefore neutralization characteristics can vary considerably. The dis-
solution rate of CaO within a slag is a function of the manner in which CaO
is bound to Si02 (Grahn and Hultberg 1975). Slags that increase solution
pH to the 6.0 to 8.0 range and have long term neutralizing properties are the
most desirable for lake and stream management applications. The determi-
nation of slag dissolution characteristics may be accomplished through
laboratory testing. The trace metal content of slags may be high; therefore,
potential for metal leaching exists.
Costs associated with fly ash or basic slags, should they be found acceptable
for use, would be largely attributed to transportation and handling, as these
materials are waste products. Resistance to the use of these materials may
be encountered if a substance the public perceives as "waste" is recommended
for application to pristine waters.
4-147
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4.7.1.2 Direct Water Addition of Base—Direct water addition of base is the
most common management strategy for acidified lakes. It has been practiced
in Sweden, Norway, Canada, and in the United States. The sources and sinks
of hydroxide within an acidified lake environment are not quantitatively
known; therefore there is no rational means of computing base dose require-
ments. Likewise, there is no accepted method for applying base to acidified
lakes.
4.7.1.2.1 Computing base dose requirements. Addition of base to acidified
waters should not be done arbitrarily. Tor cost effective use, a rational
method for base dose determinations should be used; however, to date none has
been developed. Hydroxyl ion sinks are gaseous, aqueous, and solid in
nature. These sinks include atmospheric carbon dioxide, aqueous hydrogen
ion, aluminum, inorganic carbon, and organic carbon, as well as exchange with
lake sediments.
It is desirable to impart sufficient inorganic carbon ANC to a water so that
future inputs of acid may be neutralized without a drastic decrease in pH.
Consumption of base by base neutralizing components must be realized before
residual ANC can be imparted to water. A description of the aqueous base
neutralizing capacity (BNC aq) can be described by thermodynamic
expressions:
BNC aq = 2[H2C03] + [HC03~] + 3[Al+3]
+ [A1(OH)2+] + 3[A1-F] + 3[A1-S04] + [RCOOH] + [H+]
- [A1(OH)4-] - [OH-]
where Al-F is the sum of all aqueous aluminum- fluoride complexes
(mols £-1) ,
A1-S04 is the sum of all aluminum - sulfate complexes
(mols X,"1) , and
RCOOH is the dissolved organic carbon base neutralizing capacity
(mols i-1) .
Driscoll et al . (1984) found that aquo-aluminum levels in Adirondack waters
appear to follow an aluminum trihydroxide solubility model. The speciation
of aluminum can be calculated with aluminum, fluoride, sulfate, and pH
determinations as well as pertinent thermodynamic equilibrium constants.
Dissolved organic carbon can exert some base neutralizing capacity in dilute
waters. From observations of Adirondack waters, Driscoll and Bisogni (1984)
developed an empirical formulation relating aquatic humus (dissolved organic
carbon, DX) to the mols of proton-dissociable organic acid/base:
where
CT = 2.62 x 10-6 (DOC) + 7.63 x 10'6
DOC is the dissolved organic carbon concentration (mg C £~
and,
4-148
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CT is the total, organic carbon proton dissociation/
association sites (mols £~1).
A monoprotic proton dissociation constant (pKa=4.4) was also developed for
Adirondack surface waters. From these relationships the contribution of BNC
from aquatic humus may be quantified:
[RCOOH] =
CT x
Ka
Other metals (Cu, Mn, In, Ni, Fe) are not included in the BNC equation due to
the low concentrations usually found in natural waters. Collectively, BNC
realized from these metals is not substantial compared to other aqueous
components. However, these metals may exert substantial BNC when concen-
trations are high. High concentrations would most likely be found in acidic
waters located near large industrial areas, where atmospheric deposition of
metals is high. This condition has been observed in the Sudbury region of
Ontario, Canada, where levels of copper and nickel have been observed at
concentrations greater than 1.0 mg A-l (Scheider et al . 1975, Yan and
Dillon 1981).
If equilibrium with atmospheric carbon dioxide is assumed, an upper limit of
the aqueous BNC may be estimated. Driscoll and Bisogni (1984) have made such
an analysis to neutralize a "typical" Adirondack lake (Table 4-12) to pH 6.5
(Table 4-13). It is apparent that a substantial portion of the aqueous BNC
is associated with the hydrolysis of aluminum, and this should not be over-
looked when one computes base dose requirements for acidified waters.
In determining BNC of an aquatic system, one must consider the lake sediment
as well as the overlying water. One of the consequences of lake acidifi-
cation is the accumulation of organic sediments. These sediments have
considerable exchange capacity and contribute significantly to the overall
BNC of the aquatic system. During the acidification process, BNC associated
with sediment exchange sites buffers the overlying water. Upon neutrali-
zation, the sediment exchanges back into the water column, consuming added
base. Neutralization of the water occurs quickly after base addition,
whereas the exchange with the sediment may be slow.
Base additions (CaC03 and/or Ca(OH)o) of 477, 196 and 477 yeq £-1
were applied to Middle, Lohi , and Hannan Lakes, respectively, in the Sudbury
region of Ontario, Canada (Dillon and Scheider 1984). Of these applications
161, 86 and 148 yeq &'1 (or 34, 44 and 31 percent, respectively, of the
base applied) were consumed by reactions with lake sediments. Therefore,
sediment reaction would appear to be a major component of overall-lake base
demand.
Determining sediment base demand of a lake is difficult; no accepted methods
are available. Scheider et al . (1975) determined the base demand of sedi-
ments from Sudbury lakes by titrating sediments with Ca(OH)2 to a pH of 8.0
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TABLE 4-12. COMPONENTS OF BASE NEUTRALIZING CAPACITY IN
TYPICAL ADIRONDACK LAKE WATER
(ADAPTED FROM DRISCOLL AND BISOGNI 1984)
Parameter Value
pH 4.95
Inorganic Monomeric Aluminum 0.2 mg Al &"1
Aluminum Fluoride forms 0.105 mg Al £'
Aluminum Sulfate forms 0.005 mg Al JT
Free Aluminum 0.04 mg Al £"1
A1(OH)2+ 0.03 mg Al jT1
AHOH)2+ 0.02 mg Al JT1
TOC 5.0 mg C JT1
TABLE 4-13. AMOUNT OF BASE REQUIRED TO NEUTRALIZE
BASE NEUTRALIZING CAPACITY OF
TYPICAL ADIRONDACK LAKE WATER TO pH 6.5
(ADAPTED FROM DRISCOLL AND BISOGNI 1984)
Acid component Base required (eq £~
Hydrogen Ion 1.1 x (10-5)
Carbonate 1.3 x (10-5)
Aluminum 2.0 x (10-5)
Organic Carbon 0.4 x (10-5)
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and arbitrarily assuming a reactive layer of 5 cm in the lakes. Intralake
variations in sediment base demand up to a factor of 10 were noted.
Through studies of base application to improve fish production in south-
eastern U.S. lakes, Boyd (1982) has developed a table in which sediment pH
and texture are used to calculate base dose requirements.
Menz and Driscoll (1983) used experimental data obtained from Sudbury,
Ontario and Adirondack, NY, liming experiments to develop a sediment base-
demand model. The base-demand of sediments (meq m~2) as a function of the
increase in ANC (due to base addition) of the water column was fit to a
Langmuir-type model:
ANCa
SD =
K + ANCa
where: SD is the sediment demand of base (meq nr2) ,
SDmax 1S tne maximum demand of base (meq m"2) ,
ANCa 1S tne increase in water column ANC after base addition
(yeq £-1), and
K is the sediment demand constant (yeq £-1).
This sediment demand model was coupled to aqueous thermodynamic calculations
(see above) to determine the overall base demand of a lake. Base dose
calculations using the simple model proposed by Menz and Driscoll (1983)
depend on the volume of water to be treated, the sediment surface area, the
solution water quality, the length of time over which the lake is to be
treated, and the ANC the lake is to be increased to after treatment.
Another element of uncertainty in base dose calculations is base dissolution
efficiency. For soluble bases (e.g., Ca(OH)2, NaeCOa) a dissolution
efficiency of 100 percent may be a reasonable assumption. However, the
dissolution efficiency of sparingly soluble bases (e.g., CaC03, olivine)
will depend on the method of application, the size and the impurity content
of the base, and the extent of base-particle coating (e.g., Al , Fe, organic
matter) that impeded dissolution. Driscoll et al . (1982) observed an
accumulation of CaC03 coated with organic detritus and metals within the
sediments of a limed lake. Conversely, Dillon and Scheider (1984) observed
complete dissolution of CaC03 after application to Sudbury lakes.
To develop a rational means for determining base dose requirements, further
research is needed to enhance our quantitative understanding of components
that exert a base demand in acidic lakes and of base application efficiency.
Base dose application rates have been reported in the literature. In
southern Sweden, direct water addition doses needed for neutralization have
been noted: 200 to 400 yeq £ -1 annually (Bengtsson et al . 1980), which
corresponds to 1000 to 1500 yeq ha-1 of watershed yr*1. Blake (1981)
4-151
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has reported dose requirements of 7340 eq CaC03 ha-1 of lake surface
area for initial treatment of Adirondack lakes. The period of time over
which these levels are effective was not reported.
4.7.1.2.2 Methods of base application. Managing acidic surface waters by
adding chemfcals is a relatively new concept that has been practiced to only
a very limited degree. Chemical addition strategies have generally evolved
through trial and error, and there is no single, accepted method for applying
chemicals to surface waters. The following are some of the reported methods
of chemical application.
It has been suggested that waters amenable to neutralization should be ranked
so lakes used for fishing and recreation are treated first (Blake 1981).
These waters are generally accessible lakes, which are less costly to treat
than remote waters. To determine the benefit derived from neutralization, a
cost-benefit ratio can be used. This cost-benefit ratio (Blake 1981) might
compare the cost of neutralization to the value derived by anglers. Lakes
with a low cost-benefit ratio might be considered for lake neutralization
programs. Lakes having long retention times should be favored over those
with shorter retention times (< 1 yr). Because lakes with short retention
times experience a relatively fast "washout" of base-induced ANC, these
systems are susceptible to reacidification and the effective period of
neutralization is short.
Once lakes to be neutralized are selected, application procedures must be
planned. The method of application and the location of base addition should
be optimized for the maximum dissolution of base, worker safety, and minimum
cost.
Several ideas for the optimum placement of CaCOs have been presented in the
literature. Sverdrup and coworkers (Bjerle et al. 1982, Sverdrup 1983,
Sverdrup and Bjerle 1982) have developed a model to describe CaC03
dissolution after application in acidic lakes. The major parameters
influencing CaC03 dissolution are particle settling depth and solution
characteristics. Sverdrup (1983) indicates that particles larger than 60
ym in diameter dissolve to only a limited extent in dilute acidic systems
and therefore are of little use in lake liming. Calcium carbonate resting on
(or within) lake sediments has very slow dissolution rates. Curtailed
dissolution may be attributed to burial, limited turbulence, or coating of
CaCOs particles by hydrous metal oxides or organic matter. Therefore,
dissolution during water column sedimentation should be maximized for the
most efficient application of base. Sverdrup's (1983) calculations suggest
that CaC03 should be applied in the deepest portion of a lake.
Driscoll et al. (1982), however, indicate that turbulence will enhance
dissolution. They suggest CaCOs should be placed in the littoral zone
where turbulence will enhance the dissolution rate. Within the littoral
zone, areas that are sandy and not laden with organic detritus provide the
best location for CaCOs placement. If CaCOs is placed in organic sed-
iments, particles may become buried or coated with metal and/or organic
matter. If applied to the littoral region, CaCOs should be dispersed so
only a thin layer accumulates on sediments. This type of application
4-152
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will ensure that a large surface area of base directly contacts the water and
increases dissolution efficiency. Driscoll et al. (1982) observed that when
CaC03 was applied in a thick (> 0.5 cm) layer coating by organic detritus
and metals curtailed dissolution; when deposits were spread thin (< 0.5 cm)
the CaCOs dissolved before becoming coated.
The application of base materials has been accomplished in several ways.
Transport and application vehicles include trucks, boats, helicopters, and
airplanes. The accessibility of the water to be neutralized largely deter-
mines the method selected. Two prevalent methods of application are by boat
or helicopter.
Application by boat is usually limited to readily accessible lakes and ponds.
For an efficient operation, base transported by truck must by easily trans-
ferred to a boat. This method necessitates unloading the truck close to the
water. Lime (Ca(OH)2) transported in bags is a commonly used base in boat
application. These bags are loaded onto the boat and then emptied as the
boat moves slowly through the water. Calcium carbonate may also be applied
in this manner. Scheider et al. (1975) mixed lake water and base on board a
boat, water was pumped into a hopper where base was poured from a bag and
mixed, with the resulting slurry discharged into the backwash of the boat.
Using one 5 m boat and a five-man crew, approximately 7.3 x 10^ kg was
applied in an average working day.
Large amounts of powdered CaC03 have been applied to an Adirondack lake by
using a pontoon barge (~ 3 x 103 kg capacity) (Driscoll et al. 1982).
The base was transported to the application site and washed off the barge
with water supplied by a gasoline-powered fire pump. In this manner a
three-man crew can apply 30 x 103 kg of CaC03 in an average working day.
Helicopters have been used to transport large quantities of base to remote
areas. Blake (1981) has discussed different methods of helicopter appli-
cation. Transporting bagged lime by helicoper into lakes in the winter was
not a viable application method due to the considerable labor required,
extremely low temperatures, and swirling snow that made flying difficult.
Another attempted procedure involved mixing water and lime in a fire-fighting
water bucket and spreading the slurry on the lake surface. This technique
proved inadequate because mixing equipment and a large crew were needed. In
addition, transporting large quantities of water was required. The most
practical method was direct lime application with a "bucket" (~ 1 x 10
kg capacity) suspended from a helicopter. Upon flying over a lake the pilot
opened a trap door, thereby dropping the lime to the lake. The most
efficient variation of this operation involved two buckets, with one in use
while the other was being filled.
In Norway, agricultural limestone (CaCOs) nas been applied on a frozen lake
(Hinckley and Wisniewski 1981). After limestone was applied by a manure
spreader in a 2 meter wide strip along the shoreline, a snow blower blew the
limestone and snow mixture into a 10-meter strip. Upon ice melt the base
mixed with the lake water, resulting in neutralization.
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Sodium carbonate (soda ash) is not generally used as a neutralizing material.
However, Lindmark (1981) has hypothesized that the sodium from soda ash will
exchange with cations on sediment exchange sites. Treated sediments
containing sodium may exchange with inputs of base neutralizing capacity
(e.g. H+, Al) and serve to buffer the lake against reacidification.
Lindmark (1981) suggests that calcium binds irreversibly with sediment
exchange sites; therefore, calcium treatments will not introduce the sediment
buffering that sodium treatments may provide. Lindmark (1981) argues that
the effectiveness of soda ash offsets its higher chemical cost (Table 4-14)
and is therefore economically competitive with more conventional basic
materials (e.g., Ca(OH)2, CaCOs). Lindmark1s hypothesis is controversial
because monovalent cations do not compete effectively with polyvalent cations
for sites on an exchanger in dilute solutions.
To neutralize with soda ash a 10 percent solution of sodium carbonate has
been applied to sediments of an acidic lake (Lindmark 1981). The sodium
carbonate was mixed on land and pumped to a moveable raft, and a land-based
compressor supplied air to the raft. From the raft, air and the sodium
carbonate solution were piped to a chemical rake (10 m wide) that moved along
the lake bottom. Bubbles of compressed air were released 15 cm below the
sediment surface, helping to break up the sediment. Sodium carbonate was
injected directly within the sediments. In this manner, good contact with
the base was assured. Unfortunately, data are not currently available to
evaluate the cost-effectiveness of sodium carbonate treatment. Since soda ash
is a relatively expensive base, more research is needed before this
technology can be evaluated as a potential lake management tool.
Neutralizing acidified waters through base addition is a relatively new
strategy that has not yet been extensively practiced. Application methods
must be chosen that will be compatible with the constraints inherent with
each site. If base addition becomes a more widespread procedure to mitigate
acidification, new techniques for application will be developed, along with
the refinement of existing methods.
4.7.1.3 Watershed Addition of Base—Watershed addition of base, including
stream treatment, is a relatively new strategy that has been evaluated to
only a limited degree. Research addressing watershed addition of base has
been conducted largely by Swedish scientists (Bengtsson et al. 1980, Hultberg
and Andersson 1982). This discussion will essentially reflect upon the
Swedish experience, in addition to addressing pertinent biogeochemical
concepts.
4.7.1.3.1 The concept of watershed application of base. The concept of base
addition to watersheds was developed to overcome the potential introduction
of BNC (H+, A13+) to a neutralized lake by ground-water and streams
(Section 4.4.1.4). Watershed/stream base treatment theoretically should
enhance the neutralization of ground and stream waters and result in a more
complete and compatible neutralization.
There is considerable experience to draw upon with respect to neutralization
of soils, since applying lime (agricultural grade CaCO^) is a common
agricultural practice. However, forest ecosystems are considerably different
4-154
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TABLE 4-14. CHEMICAL COST COMPARISON OF NEUTRALIZATION AGENTS*
Chemical
CaC03
Ca(OH)2
Ca(OH)z
CaO
N32C03
(Mg,Fe)2Si04b
H3P04
Form Equivalent Mass
supplied weight basis
(g eq-1) (dollars x
10-3 kg-1)
bags (325 50 20.90
mesh)
bulk 37 35.75
bags 37 46.75
bulk 28 34.38
bulk 53 101.20
bulk (100 86 22.00
mesh)
agricultural 5.75C 6.49
grade (70%
solution)
Costa
Equivalence
basis
(dollars
eq-D
1.04 x 10-3
1.32 x ID'3
1.73 x 10-3
9.62 x 1Q-4
5.36 x 10-3
1.89 x 10-3
3.73 x ID'5
achemical costs as reported in Chemical Marketing Reporter,
August 31, 1981.
bThis analysis assumes the above stoichiometry, however this may vary
from source to source.
cEquiva1ent weight is computed based on the assumption that N03~ is
the nitrogen source, which is assimilated and never reoxidized within the
aquatic system (see Section 4.7.2.1).
4-155
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from agricultural ecosystems, and it is difficult to extrapolate from one to
the other.
The acid/base chemistry of soil systems is extremely complex, with reactions
such as cation exchange, mineral dissolution, and biological uptake all
influencing soil solution acidity. In forest soils derived from silicate
bedrock, the bulk of the cation exchange capacity may be attributed to
natural organic matter and to a lesser extent clay minerals (e.g., kaolinite,
vermiculite, illite). The exchangeable cations are largely basic cations
(Ca2+, Mg2+, Na2+, K2+) and/or acidic cations (Ain+, H+). At near neutral pH
values, the exchangeable cation pool is largely comprised of basic cations.
As soil pH decreases, the exchangeable acidity (Ain+, H+) is thought to
increase. Another reaction of interest is biological uptake of cations.
Forest biomass requires cationic nutrients (e.g., Ca2+, Mg2+, K+) for
growth. An aggrading forest will assimilate basic cations and tend to
deplete soil pools.
Cation exchange and biological uptake reactions are significant consider-
ations With regard to watershed liming. Forest soils are generally nutrient
poor and elevated in exchangeable acidity. Upon addition of base [Ca(OH)2»
CaCOa] to a forest soil, a considerable shift in ionic equilibria would
ensue. The introduction of elevated levels of calcium would result in a
Ca2+-H+ exchange on soil exchange sites. The release of protons would
neutralize the associated hydroxide or carbonate introduced in the liming
process. Biological uptake of calcium may result from calcium addition; this
would generate protons as well and neutralize the associated basic anion.
Terrestrial acid/base reactions are much more complicated and more poorly
understood than aquatic acid/base reactions. It is difficult to evaluate,
much less quantify, perturbations in acid/base chemistry that result from
watershed liming. As a result, assessing the efficiency of a watershed
liming program is difficult.
Stream neutralization techniques have also been attempted. Stream neutral-
ization is of interest because streams are valuable aquatic resources and
maintaining stream water quality is of concern. An important consideration
is the fact that acidic streams may flow into acidic lakes and influence lake
biogepchemistry. When an acidified lake is limed, it will still experience
the introdution of BNC (Ain+, H+) from stream inputs. Aquatic organisms
(particularly fish) that use the stream for feeding or reproduction may be
adversely affected by the extensive aluminum hydrolysis resulting from the
introduction of acidic stream water to a neutralized lake. Stream (and
watershed) liming could help minimize this water quality problem.
4.7.1.3.2 Experience in watershed liming. Experiments with watershed liming
are limited to those conducted in Sweden (Bengtsson et al. 1980). Agricul-
tural lime (powdered CaCOa, 0 to 0.5 mm) has been transported to the water-
shed in large trucks and applied as a slurry with a sprayer truck. In this
manner one man is able to apply 40 x 10* kg of CaCOs per day (Hinckley
and Wisniewski 1981). The CaCOa dose required to achieve adequate neutral-
ization of water systems is generally two orders of magnitude greater than
that of direct water addition (Bengtsson et al. 1980). This high dose is
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undoubtedly due to the many base consuming processes that occur within forest
soil systems. Application rates are generally in the range of 5000 to 7000
kg CaCOa ha*1 yr"*. Powdered olivine (0 to 1 mm), a magnesium iron
silicate, has also been used as a base in watershed application experiments
(Hultberg and Andersson 1981).
Water quality information resulting from watershed application experiments
has not been published; however, authors indicate that watershed liming
efforts have been successful (Bengtsson et al. 1980, Hultberg and Andersson
1982). Hultberg and Andersson reported that some damage to the terrestrial
environment may be associated with liming. Sphagnum moss was severely
damaged as a result of CaCOa addition. Damage to lichens, spruce needles,
and other types of moss was also observed. Similar damage occurred with
olivine application experiments; however, the extent of damage was
considerably less than that from CaCOa addition.
There are problems associated with stream neutralization practices. It is
reasonable to say that no cost effective method of achieving stream
neutralization has been developed. Problems center around the drastic
temporal changes in water flow and water quality that occur in headwater
streams. During spring and autumn, water flow and solution BNC are high.
During summer, water flow and BNC are low. A successful neutralization
scheme must adequately account for the tremendous temporal fluctuation in
base dose requirements of acidic streams.
Four approaches have been attempted to achieve stream neutralization. The
simplest approach is CaCOa addition to the streambed (Hultberg and
Andersson 1982). Stream additions have been attempted with both coarse (5 to
15 mm) and fine (0 to 0.5 mm) CaCOa. Coarse CaCOa will tend to stay in
the stream bed, but neutralization is generally inadequate because of the
rather low surface area of the stone. Fine CaCOa will more readily
dissolve (due to a greater surface area) but is more influenced by stream
turbulence. Powdered CaCOa tends to be transported to pools, where it
settles within organic detritus, or it can be washed into a lake. In these
sites CaCOa is ineffective in supplying BNC to streams.
Another approach to achieve stream neutralization is the limestone barrier.
Driscoll et al. (1982) constructed a limestone barrier of perforated
55-gallon drums filled with CaCOa (5 to 15 mm), in an attempt to neutralize
an acidic stream. The barrels were placed across the width of the stream,
2-barrels high with loose limestone filling spaces between the barrels.
Screens were placed upstream to filter out debris that might clog the pores
of the barrier. Stream neutralization was accomplished for approximately one
week, largely due to fine material associated with the larger stone. The
coating of the stone by hydrolyzed aluminum, iron, and organic detritus
quickly curtailed further neutralization. The coating diminished calcium
carbonate dissolution and rendered the barrier ineffective as a means of
achieving neutralization.
Diversion wells have been used to treat acidic streams in Sweden (Swedish
Ministry of Agriculture Environment Committee 1982). Diversion wells consist
of a cylinder embedded within a stream bank or channel and filled with
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CaC03- A pipe diverts stream water, by gravity, to the cylinder. Water is
introduced through the bottom of the cylinder and flows upward through the
CaC03 bed. Water neutralized by passage through the cylinder bed overflows
back into the stream, increasing the ANC. The upflow velocity results in
particle abrasion, which aids to restrict particle coating. A series of
diversion wells may be placed in a stream such that the inflow pipes will be
located at various levels of stream stage. Thus during high-flow conditions
several diversion wells would be operating and treating a large volume of
flow. As stream flow decreases, the stream depth would decrease; therefore
the number of operating wells and volume of water treated would decrease.
A fourth type of stream neutralization, automated base addition systems, is
the most effective means of supplying ANC to acidic streams. However, they
are extremely expensive in terms of both capital and operating costs.
Swedish scientists have used river silos (cylindrical storage bins) to
accomplish stream neutralization (Hinckley and Wisniewski 1981). These silos
hold 30 x 103 kg of base and can meter up to 1300 kg day"1 of base into a
stream. Each silo costs approximately $20,000 (1981 dollars). The rate at
which base is metered from the silo to the stream is activated by pH or flow
sensing devices. An automated system, like the river silos, would seem to be
the best means of applying an adequate base dose to varying water flow and
quality conditions.
In addition to cost, however, there are several problems associated with
automated stream treatment systems. The silos may be used only in easily
accessible streams, and the automated operation is not entirely reliable and
will malfunction occasionally. Also, stream base addition requirements are
considerable during high flow conditions; silos must be constantly refilled
during spring and autumn (Hinckley and Wisniewski 1981). These problems are
not severe in themselves, but they imply that stream neutralization efforts
may be interrupted periodically. Interruption of base addition will most
likely occur during high flow, low pH conditions (spring, autumn, and winter)
when water quality conditions are most critical. Periodic discontinuities in
base addition have severe implications for aquatic organisms. The response
of water quality and aquatic organisms to acute fluctuations in ANC, from
equipment malfunctions, needs to be evaluated before automated base addition
systems are implemented as part of a stream management program.
4.7.1.4 Water Quality Response to Base Treatment—Lake water neutralization
by base addition may Be accomplished b~ydTrect base addition or by
watershed/stream input neutralization. Few studies of the water quality
response in groundwater or streams as a result of neutralization are reported
in the literature. Likewise, an evaluation of lake neutralization by
watershed/stream input neutralization has not been made. As a result, this
discussion of the water quality response to base treatment reflects only the
results reported from direct base addition studies.
o Transparency increases immediately following base addition especially in
colored waters (Yan and Dillon 1981, Hultberg and Andersson 1982). This
response appears to be due to the removal of dissolved organic matter by
co-precipitation with metals (Yan and Dillon 1981). The long term
consequence, however, has been the reduced transparency in neutralized
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lakes. Decreases in epilimnion thickness and decreased hypolimnetic
temperatures have been associated with these transparency changes (Van
and Dillon 1981). Upon reacidification, transparency has increased.
0 A natural consequence of base addition is the resulting increase in pH.
Response of pH is dependent on the neutralizing agent used. When soluble
base such as Ca(OH)2 is applied, pH increases sharply and a maximum pH
is realized shortly after addition. This increase in pH is followed by a
decline in pH due to atmospheric carbon dioxide influx. When equilibrium
with COa is approached, stabilization of pH results. If acidic inputs
are significant through either streamflow, groundwater infiltration, or
sediment cation exchange, a gradual but steady decrease in pH will occur.
When a slightly soluble base (e.g., CaCOa) is added to an aquatic system,
the pH Increase is less dramatic. With calcium carbonate addition the rate
of pH increase depends on particle size and degree of water contact.
Increases in stone surface area exposed to solution enhance dissolution
rates, resulting in a more rapid pH increase. Acid neutralizing capacity
also increases after base addition (Bengtsson et al. 1980). ANC increases
are initially considerable but may decrease significantly with slight
decreases in pH.
0 Increases in dissolved inorganic carbon result from neutralization.
Increases in pH from less than 5.0 to greater than 6.5 cause dissolved
inorganic carbon equilibria to shift from a H2C03 (C02[aq] +
H2C03) dominated system to a bicarbonate dominated system. If the
environment is open to atmospheric carbon dioxide, increases in dissolved
inorganic carbon will result. When a noncarbonate base (e.g., Ca(OH)2)
is added, the increase in inorganic carbon is due entirely to atmospheric
C02» whereas when a carbonate base (e.g., CaCOa) is added, inorganic
carbon increases are due to the base itself as well as atmospheric C02.
0 Trace metals concentrations generally decrease after base additions to
acidified waters. Metals found in elevated concentrations in acidified
waters include Al, Mn, and Zn. Of these trace metals, aluminum is
probably of the most concern, with concentrations of 0.2 to 1.0 mg Al
SL~I commonly observed (Driscoll 1980). As solution pH increases,
due to base addition, aluminum hydrolyzes and precipitates. It has been
observed that aluminum in hydrolyzed forms is toxic to fish (Driscoll et
al. 1980). In Swedish liming experiments, fish kills were experienced
shortly after base application (Dickson 1978b). Fish stocking should be
attempted only after hydrolyzed aluminum has settled from the water
column.
Addition of base generally results in decreased concentrations of other trace
metals in addition to aluminum (Scheider et al. 1975, Van et al. 1977,
Driscoll et al. 1982). Sediment trap analyses support water column data,
showing a rapid accumulation of metals in traps following an increase in pH.
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Decreases In trace metal levels from the water column may be explained by
hydrolysis and precipitation, or adsorption on hydrous aluminum oxides formed
by base addition. Adsorption on metal precipitates is also considered to be
a mechanism by which dissolved organic carbon and phytoplankton are removed
from the lentic environment.
Sulfate response to neutralization appears to be minimal. Comparing lakes
that had been neutralized to control lakes showed no significant variation in
temporal changes in sulfate (Scheider et al. 1975).
Basic cation chemistry, excluding the cation associated with base addition,
appears to be unaltered by neutralization. Levels of calcium are observed to
increase, as expected, due to dissolution of calcium-based chemical
neutralizing agents. The temporal increase in calcium concentration will
depend on the dissolution rate of base. Calcium levels increase quickly with
soluble bases [Ca(OH)2] and more slowly with slightly soluble bases
(CaC03). Once the initial dissolution has occurred, calcium levels peak in
concentration and then decline due to export from the lake or exchange with
sediments.
A major problem associated with lake neutralization is the potential for
reacidification. Reacidification results in the resolubilization of trace
metals (Al, Mn, and Zn) which are presumably introduced from the lake
sediments (Driscoll et al. 1982). Reacidification does not result in an
immediate reintroduction of dissolved organic carbon (DOC). It appears that
DOC must be reintroduced to the water column from terrestrial inputs (e.g.,
stream and groundwater inflows) and therefore takes considerable time to
appear. The loss of DOC implies that there are few available organic ligands
to complex trace metals, particularly aluminum, that enter the water column
during reacidification. This response translates to a decrease in water
quality, particularly with respect to potential for aluminum toxicity to
fish.
Another consideration is input of stream water (and groundwater) to
neutralized lakes. The introduction of acidic water to a neutralized lake
results in a localized metal hydrolysis region at the stream (and
groundwater)--!ake interface. These chemical transformations may have
implications with respect to aluminum toxicity to fish, particularly those
fish that associate with stream systems for reproduction and feeding. If
aluminum is hydrolyzing in this environment it may be unsuitable for
habitation by fish. Any program to stock fish in a neutralized lake must
consider problems associated with acidic stream/groundwater quality entering
the lake environment.
4.7.1.5 Cost Analysis, Conclusions and Assessment of Base Addition--
4.7.1.5.1 Cost analysis. It is extremely difficult to make a cost
comparison of different acidic lake management strategies. It is relatively
easy to tabulate capital, chemical, labor, and operating costs, but any
economic evaluation must be based on the effectiveness of the treatment.
Little is known of the effectiveness and efficiency of various treatment
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strategies. As a result, any economic evaluation of management strategies
for acidified waters should be viewed with caution.
Costs of chemicals that have been proposed for use in neutralization efforts
are listed in Table 4-14, which shows the considerable range in chemical
costs. This tabulation is somewhat misleading because it does not
incorporate application efficiency into the analysis. Soluble bases
(Ca(OH)2, CaO, Na?C03) are undoubtedly the most efficient means to add
base, while slightly soluble bases (CaCOa, MgFeSi04) and phosphorus
(Section 4.7.2.2) are potentially less efficient. Very little is known about
the relative efficiency of neutralization strategies, and without such an
understanding chemical cost comparisons are difficult.
Costs involved in neutralization efforts will vary greatly with lake location
and accessibility. Blake (1981) determined costs for six accessible ponds
treated (by boat) in 1977-78 and four remote ponds treated (by helicopter) in
1978-79, totaling 79 ha and 39 ha, respectively. Neutralization cost for
accessible ponds was $131 ha-1 while cost for the remote ponds was $341
ha-1. These were experimental efforts, so costs may be substantially
reduced if base addition is implemented on a routine basis. Costs for liming
remote ponds by helicopter on a routine basis were estimated to average $247
ha'1. This was based on the following costs: helicopter - $250 hr'1,
lime - $44 x 10-3 kg-l delivered onsite, travel expenses - $100 day-1,
the ability to apply 4.5 x 103 kg of lime hr-1, and the use of an
eight-man ground crew at $35 day-1 person'1 (Blake 1981). Neutralization
of a series of lakes has been shown to be the most efficient operation. Four
ponds treated in 1977 by a three-man crew cost approximately $74 ha'1
(Blake 1981).
Costs associated with application by boat are not detailed in the method
described by Scheider et al. (1975). However, a comparative cost analysis
may be determined. A five-man crew using a 5-meter boat was able to apply
7.3 x 103 kg day1 of hydrated lime. Since the major costs of base
addition are associated with labor and the cost of base, a reasonable
comparative estimate can be formulated.
Using chemical, labor, and transportation cost data obtained by the above and
other investigators, Menz and Driscoll (1983) estimated the costs of
neutralizing acidic Adirondack lakes through a program of base addition. In
this analysis lakes were subdivided as accessible (those lakes with access by
road so they can be treated by boat) and inaccessible (those lakes with no
road access and requiring helicopter treatment). Costs to treat accessible
lakes for a 5-year treatment period were approximately $50.75 ha'1.
Chemical transportation cost to the site represented the major component of
cost for the treatment of accessible lakes. The cost to treat inaccessible
lakes for a 5-year treatment period was approximately $500 per surface
hectare. Most of this cost was associated with the cost of applying the
chemical. It is noteworthy that costs vary, from lake to lake, with the
desired target pH (and ANC), and with the treatment period. Overall results
were derived from water quality data from 777 of the 2,877 Adirondack lakes
sampled to date (Pfeiffer and Festa 1980). The estimated annual cost for a
5-year base addition program for the lakes in this sample would be in the
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range of 2 to 4 million dollars, depending on the specific target pH of the
water.
Presently, the support for most lake neutralization programs comes directly
from government agencies. Sweden has been the most active, with over 300
individual projects involving 1000 lakes and waterways (Bengtsson et al.
1980). As concern for the problem increases, private groups (i.e., sportsman
clubs, lake associations) may become actively involved in neutralization
programs. However, limited resources will probably prevent the
neutralization and management of all acidified lakes.
4.7.1.5.2 Summary—base additions. Base addition is currently the most
viable strategy available for managing acidic lakes. Methods used to compute
base application requirements are crude due to our lack of understanding of
the efficiency of treatment techniques and sediment interactions. A benefit
associated with base addition is the alteration of the chemical environment
(e.g., increases in pH and calcium, decreases in trace metal levels). Such a
chemical alteration might result in an environment more hospitable to
desirable aquatic biota (e.g., decreases in Sphagnum, increases in fish
populations). However, in addition to the benefits associated with base
addition, there are costs. These costs include financial as well as
environmental costs. Environmental costs include pH shock associated with
dramatic increases in pH, the problems associated with aluminum hydrolysis at
the stream-neutralized lake interface, and the potential for lake
reacidification. These and other environmental costs have not been fully
evaluated.
Base addition has become a popular strategy to mitigate water quality
problems associated with acidification. However, before base addition is
implemented as a regional, acidic lake management alternative it should be
more thoroughly evaluated.
4.7.2 Surface Water Fertilization
Soft water lakes are generally thought to be phosphorus growth limited (N/P >
12). As a result, fertilization by phosphorus addition might serve as a
means of restoring acidified lakes. However, this hypothesis has been
researched and evaluated only to a limited degree. This analysis is a
summary of the limited studies on nutrient addition to acidified waters, as
well as an extrapolation of some concepts pertinent to natural waters.
Further research is needed to evaluate effectively lake response and
consequences associated with nutrient addition.
4.7.2.1 The Fertilization Concept—The concept of phosphorus addition as a
strategy for the management of acidified lakes is twofold:
1. To supply ANC through biological uptake of nitrate; and
2. To increase aquatic biomass and species diversity.
The idea of supplying ANC through biological uptake of nutrients, is
summarized by the following stoichiometric expression (Stumm and Morgan
1970).
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106 C02 + 16 N03- + H2P04" + 122 HgO + 17 H+ (+ Trace Elements, Energy)
photosynthesisf -t-respiration
C106H262°110N16P (alga"1 protoplasm) + 138 02
The above equation is a generalized relationship and may vary significantly
from ecosystem to ecosystem, as well as temporally within a given aquatic
system. Regardless of the inadequacies of the stoichiometric expression, it
provides a framework through which microbially mediated changes in solution
acid/base chemistry might be understood.
The stoichiometric expression suggests that uptake of nutrients by algae will
result in the consumption of protons or the generation of ANC within the
aquatic environment. This response results from the assimilation of nitrate
as a nitrogen source. For the organism to maintain an electroneutrality
balance, the uptake of nitrate must be countered by an equivalent cation
uptake (or anion release). In the above expression this is realized through
hydrogen ion uptake.
This expression is somewhat simplistic, for in actuality a number of
additional factors should be considered.
1) Although nitrate nitrogen is generally the predominant nitrogen source in
aerobic waters, uptake of ammonium or organic nitrogen could occur. Under
these circumstances the stoichiometry would significantly change. In
fact, assimilation of ammonium as a nitrogen source would result in
consumption of ANC (Brewer and Goldman 1976).
2) Plants require certain cations as nutrients (e.g., Ca2+9 Mg2+, Fe).
The uptake of cations by algal protoplasm would diminish the quantity of
ANC generated through photosynthesis.
3) Although carbon fixation through photosynthesis results in generation of
ANC, respiration will result in consumption of ANC. This process may
partially account for why acidic lakes have a higher ANC in summer months
than in winter months. Therefore, only net removal of reduced nitrogen
associated with algal material through lake outflow or permanent burial in
sediments will result in a net production of microbially mediated ANC.
The concept of acid neutralizing changes generated by phytoplankton growth
has been studied by Brewer and Goldman (1976). Such processes may be
important in dilute water acid/base chemistry. According to the above
stoichiometric expression, 4.8 x 10-3 yeq of ANC would be generated per
microgram of net algal biomass produced, or 5.5 x 10'1 yeq of ANC would
be generated per g of net phosphorus fixed by algal uptake.
The second reason for nutrient addition is to replenish the biomass of
acidified lakes. Hendrey et al. (1976) have suggested that phytoplankton
biomass is reduced by lake acidification. Dillon et al. (1979) suggest that
phytoplankton biomass is better correlated with total phosphorus levels than
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with pH. However, acidification may alter phosphorus cycling (Section
4.7.2.2). Nutrient addition may help replenish phosphorus lost (possibly) by
acidification and increase the productivity and species composition of these
lakes.
4.7.2.2 Phosphorus Cycling in Acidified Water—Phosphorus cycling is reason-
ably wellunderstood in circumneutrallakes (Hutchinson 1975). Generally
phosphorus will enter a lake through direct atmospheric precipitation,
groundwater, or stream flow. It may be exported from the lake by groundwater
or stream flow. Within the lake, phosphorus may be assimilated by phyto-
plankton or macrophytes. Once in the form of particulate phosphorus, it may
be consumed by organisms, released to the water by oxidation reactions, or
lost to the sediments. Within the sediments, phosphorus may be released by
decomposition processes. This released phosphorus may bind with aluminum,
calcium, or iron or diffuse vertically back into the water column.
In acidified waters aluminum might alter phosphorus cycling through precip-
itation or adsorption reactions. Aluminum can directly precipitate with
orthophosphate to form AlPCty (varascite). A more plausible mechanism by
which aqueous phosphorus levels might be regulated is adsorption on hydrous
aluminum oxides (Huang 1975). The adsorption is pH dependent with a maximum
near pH 4.5. It is likely that increases in pH of acidic water result in the
formation of hydrous aluminum oxides. These oxides would serve as an
adsorbent that could effectively scavenge phosphorus from the water column.
Upon nutrient addition to an acidic lake, competition between algae and
aluminum for a given phosphorus molecule will ensue. It is difficult to
state how phytoplankton uptake of phosphorus is altered by the presence of
aqueous aluminum. This competition is undoubtedly complicated and altered by
environmental conditions such as pH, general water chemistry, light, and
temperature.
Although changes in water quality may result on a short term basis, most of
the added phosphorus will be lost to the sediments (Schindler et al. 1973,
Scheider et al. 1976). The degree to which sedimented phosphorus diffuses
back to the water column is virtually unknown for acidic lakes. However,
because these systems are generally aerobic, have reduced decomposition
rates, and undoubtedly contain significant levels of amorphous iron and
aluminum oxides that potentially bind phosphorus, it is doubtful that sig-
nificant vertical diffusion of phosphorus occurs. If fixed nitrate asso-
ciated with algal uptake of phosphorus is lost from the system, applying
phosphorus has been efficient from the standpoint that ANC was produced in
the water column. However, if fixed nitrate reaches the sediment, is
oxidized, and diffuses to the water column while the associated phosphorus
remains in the sediment, phosphorus application would be inefficient (no net
generation of ANC to the water column resulting). Schindler et al. (1973)
have indicated that nitrogen sedimentation and removal are less efficient
than phosphorus sedimentation and removal.
4.7.2.3 Fertilization Experience and Water Quality Response to Fertilization
--As mentioned previously there has been limited experience with fertili-
zation of acidic lakes. Most of the work has been accomplished by Canadian
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scientists (Scheider et al. 1975, 1976; Scheider and Dillon 1976; Dillon et
al. 1977, 1979). Generally a desirable water column phosphorus level is
chosen for a particular lake, and a model such as that of Dillon and Rigler
(1974) is used to calculate the required phosphorus dose. Usually H3P04
is applied because of its low cost, ease of handling, and solubility (Table
4-14). Application is usually made in the late spring or early summer;
periodic additions may be made throughout the summer to enhance assimilation
efficiency.
Nutrient addition has generally been used to increase the standing crop of
food chain components within a lake. To accomplish this, phosphorus addition
has generally been practiced after liming. Phosphorus consuming reactions
are minimized by precipitating aluminum with base and allowing aluminum to
settle out of the water column prior to any phosphorus addition.
Few data have been reported on ANC changes as a result of phosphorus addi-
tion. However, Dillon and Scheider (1984) observed decreases in inorganic
nitrogen (largely nitrate) and increases in total organic nitrogen folowing
nominal orthophosphate additions of 10 to 15 yg P £"*• to neutralized
lakes (Hannah and Middle) in the Sudbury region of Ontario, Canada. They
calculated the theoretical increase in ANC resulting from observed changes in
nitrogen chemistry for fertilized lakes (Hannah and Middle) in comparison to
a neutralized lake that received no phorphorus addition (Lohi Lake). The ANC
generated from nitrogen transformations for the fertilized lakes was 2 to 8
neq £"i greater than the control lake. In addition, the ANC generated
from nitrogen transformations declined dramatically after phosphorus addi-
tions were terminated.
Observed changes in aquatic biota have been more significant. Small addi-
tions of total phosphorus resulted in significant increases in phytoplankton
biomass of neutralized Canadian lakes (Dillon et al. 1979). No single obser-
vation in phytoplankton species composition has been reported. Shifts to
communities dominated by Chrysophytes (Langford 1948), by blue greens (Smith
1969), and by different groups in different years of fertilization (Schindler
et al. 1973) have been reported. Shifts in green or bluegreen algae domi-
nance can generally be attributed to the nitrogen to phosphorus ratio within
the lake (Schindler 1977).
Dillon et al. (1979) observed changes in phytoplankton resulting from small
levels of phosphorus added to a limed lake (Middle Lake, Ontario). In the
first year after addition blue-green algae biomass increased significantly.
The second year after fertilization, green algae were generally dominant.
Fertilization of a second lake (Hannah Lake, Ontario) resulted in an increase
in biomass but no change in the structure of the phytoplankton community.
Although increases 1n phytoplankton biomass were evident, no conclusions with
regard to changes in zooplankton population could be made from this study.
In enclosure experiments within a limed lake, Scheider et al. (1975) observed
that fertilization with phosphorus and wastewater effluent resulted in an
increase in the standing stock of bacteria, phytoplankton, and zooplankton.
Hultberg and Andersson (1982) investigated nutrient addition as a means of
supplementing liming efforts in Sweden. They reported few results except for
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a shift in lake phytoplankton from Peridineans to primarily chlorophyceans,
which they attributed in part to fertilization.
Little work has been done with water chemistry response to phosphorus
addition. Dickson (1978b) has observed the precipitation of phosphorus added
to acidic lake water; precipitation was most dramatic at pH 5.5. The
presence of DOC inhibited the precipitation of phosphorus by aluminum.
Scheider et al. (1975) observed decreases in phosphorus added to enclosure
experiments. They attributed this to precipitation of the phosphorus by
metals.
4.7.2.4 Summary—Surface Water Fertilization—It is difficult to assess
critically phosphorus addition as a management strategy to improve the water
quality of acidic lakes because the general process has not been effectively
evaluated. While the chemical costs associated with phosphorus addition are
low (Table 4-14) applications may not be efficient, particularly in view of
potential interactions with aluminum (Schindler et al. 1973, Scheider et al.
1976). In the few studies conducted, the benefits accrued to the ecosystem
have not been evaluated.
4.8 CONCLUSIONS
Acidification of lakes and streams, with resultant biological damage, has
been widely acknowledged in the last decade (WAS 1981, NRCC 1981, U.S./Canada
1982). Assessing causal relationships remains difficult, however, because
effects of acidic deposition on any one component of the terrestrial-
wetland-aquatic systems depend on not only the composition of the atmospheric
deposition but also on the effect of the atmospheric deposition on every
system upstream from the component of interest. Composition of aquatic
systems results, moreover, from biological processes in addition to chemical
and physical processes; thus, assessing results of acidification on all three
processes is required. Our knowledge of past, current, and future acidifi-
cation trends, of critical processes that control acidification, and of the
degree of permanency of chemical and biological effects remains incomplete
and subject to debate.
This chapter has critically reviewed how aquatic chemistry responds to acidic
deposition. After defining concepts involved in discussions of aquatic
chemistry and acidic deposition, the chapter listed those characteristics of
terrestrial and aquatic systems that ameliorate or enhance the effect of
acidic deposition. It then discussed aquatic systems' theoretical and
practical sensitivity to acidic deposition and identified locations of
sensitive and affected systems. The chapter also considered the interaction
of aquatic acidification with the metal and organic biochemical cycles and
then concluded by discussing alternative methods for improving water quality
where acidification has occurred.
The following statements summarize the content of this chapter.
o Each of several components of aquatic or terrestrial systems may
assimilate some or all acidic deposition falling in a watershed. These
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components are vegetative canopy, soils, bedrock, wetlands, or an aquatic
system itself (Section 4.3.2).
Soils assimilate acidic deposition through dissolution, cation exchange,
and biologic processes. Generally, soils containing carbonate materials
have abundant exchangeable bases and can assimilate acidic deposition to
an almost unlimited extent. Soils that contain no carbonate materials
can assimilate acidic deposition because of cation exchange reactions,
silicate-mineral dissolution reactions, and in some cases Fe and Al oxide
dissolution. Assimilation ability is affected by soil chemical nature
(especially CEC and BS), the permeability at each layer, the surface area
of the soil particles, and the amount of soil in the watershed (Section
4.3.2.2).
Hydrology, specifically flow paths and residence times, can determine the
extent of reactions between strong acid components of deposition and each
component the water contacts. Flow paths and residence times are
controlled by many factors, including topography and climate (Section
4.3.2.4).
Alkalinity or acid neutralizing capacity (ANC) determines the
instantaneous ability of a lake to assimilate acidic deposition, but the
ANC renewal rate depends upon the ANC supply rate from the watershed. In
addition, internal production of alkalinity is important, especially in
lakes with low alkalinity. Because biological processes can alter the
relative amounts of acidity and alkalinity within the body of water,
nutrient status is important in determining the sensitivity of a lake to
acidification (Section 4.3.2.6).
Aquatic systems sensitive to acidification by acidic deposition are
commonly waters of low pH and alkalinity. An approximate boundary
between sensitive and insensitive systems in North America is 200 yeq
&~l of alkalinity (prior to the onset of acidification) (Section
4.3.2.6.1). This concentration is chosen because: (a) in North America
acidic deposition has resulted in about 100 peq £-1 of potential
long-term acidification of surface water (Section 4.3.1.5.2, NRCC 1981);
(b) during spring snowmelt or heavy rainfall, short-term alkalinity of >
100 yeq £-1 occurs, and (c) biological effects due to acidification
begin when aquatic systems reach alkalinities of about 40 y eq £-1
(range of 10 to 90 yeq £-1).
Regions in North America contain aquatic systems sensitive to
acidification. These regions are found throughout much of eastern Canada
and New England; and parts of the Allegheny, Smoky, and Rocky Mountains
and the Northwest and North Central United States (Figures 4-5 to 4-8;
Galloway and Cowling 1978, Omernik and Power 1982, NAS 1981, NRCC 1981).
However, a large amount of more detailed survey work is required to
determine the levels of alkalinity and degree of sensitivity (Section
4.4.3).
Although there can be significant problems with comparing old and new
data, overall, the analysis of temporal records shows recent decreases in
4-167
-------
alkalinity and pH in some otherwise undisturbed streams and lakes in
areas receiving acidic deposition (Section 4.4.3.1.2). As yet, no body
of evidence exists suggesting that changes of such magnitudes, and at
such rates, occur in otherwise undisturbed areas not receiving acidic
deposition.
The limited application of paleolimnologic indicators shows decreases in
pH in northeastern United States over the last 10 to 80 years (depending
on the lake) for most (9 of 15) acidic lakes studied (Section 4.4.3.2).
For at least 3 of these acidified lakes, the recent decline in pH may
reflect in part a recovery from an earlier higher pH due to a temporary
period of mild eutrophication (Davis et al. 1983). For 4 of the 9
acidified lakes, however, no such pattern of pH increase followed by pH
decrease has been noted (Del Prete and Schofield 1981, Charles 1984).
Although acidic waters do occur naturally, in some cases, and changing
land use may locally alter the pH regime of surface waters, it appears
that regional acidification and episodic pH depressions (pH < 5.0) in
clearwater oligotrophic lakes and streams occur only in response to
increased atmospheric deposition of strong acid (Section 4.4.3.3). In
areas not receiving acidic deposition, but with identical changes in land
use, regional acidification of clearwater oligotrophic surface waters has
not occurred.
Predictive modeling of the effects of acidic deposition on surface water
chemistry is a complicated task. Some steady-state approaches exist and
some time-variable models are in development. Interpretation of
predictions from such models requires care, with full cognizance of their
assumptions and limitations (Section 4.5).
Addition of acidic deposition to terrestrial and aquatic systems can
disrupt the natural biogeochemical cycles of some metal and organic
compounds to such a degree that they can cause biological effects
(Section 4.6). The chemical form of dissolved metals is important in
determining the total mobility of a metal and the biological effects
related to acidification of aquatic ecosystems. Acidification increases
the concentration of many metals in surface waters and changes speciation
toward more biologically active forms.
Waters may be treated with base substances to neutralize the effects of
acidic deposition. Only lime and limestone have been used to any extent
in either direct lake additions or watershed/stream additions. Several
other materials have been proposed, but tests for effectiveness and
operability must be conducted. Organic carbon addition and surface water
fertilization have also been proposed but also must be tested (Section
4.7).
4-168
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS
E-5. EFFECTS ON AQUATIC BIOLOGY
5.1 INTRODUCTION (J. J. Magnuson)
The loss of fish populations from seemingly pristine oligotrophic waters was
the first and most obvious indication that atmospheric deposition was affect-
ing aquatic ecosystems (Dannevig 1959, Beamish and Harvey 1972, Cowling
1980). Changes in water chemistry, particularly increases in acidity, were
found to be associated with these local fish extinctions. Later studies have
included the effects of acidification on other aquatic organisms, such as
those associated with bottom substrates (the benthos), tiny plants and ani-
mals floating freely in the water column (the plankton), and rooted aquatic
plants (macrophytes). The resultant literature is large, widely scattered,
and varies considerably in its scientific merit. The purpose of this chapter
is to review and evaluate this literature critically, and to summarize the
effects of acidification on aquatic organisms.
The chapter begins with a section on naturally acidic waters, including a
discussion of what organisms occur in such habitats and how their distri-
butions relate to distributions in habitats recently acidified by man's
activities. Subsequent sections critically evaluate the literature regarding
the response of benthic organisms, macrophytes and wetland plants, plankton,
fishes and other aquatic biota to acidification. These are followed by a
discussion of ecosystem-level responses to acidification and a section on
mitigative options. The final section summarizes the known effects of acidi-
fication on aquatic biota and indicates potential effects that need to be
addressed.
It should be kept in mind that acidification of freshwaters is a complex
process that involves more than merely increases in acidity. Other well-
documented changes include increased concentrations of metal ions, increased
water clarity, the accumulation of periphyton (microflora attached to bottom
substrates) and detritus, and changes in trophic interactions (e.g., loss of
fish as top predators). The response of aquatic systems to acidic deposition
must be viewed in terms of all these changes that together constitute the
acidification process.
Evidence linking changes in aquatic communities to acidification can be
divided into three types. The first type consists of field observations,
which are 1) descriptions of conditions before and after acidification is
suspected to have occurred or 2) contemporary comparisons of water bodies
thought to exhibit different degrees of acidification. Problems exist with
5-1
-------
this type of correlation approach. For example, before and after studies may
be difficult to interpret if methodologies have changed in the interim (see
Chapter E-4, Section 4.4.3.1), or if other factors such as land-use practices
have also changed. In comparative studies, pH is frequently correlated with
other limnological parameters (e.g., lake size, nutrient concentrations),
making it difficult to attribute inter-lake biotic differences solely to
differences in pH. Despite these problems, field observations provide the
earliest indications of changes in biotic communities and provide a basis for
forming hypotheses that can be further evaluated when consistent trends are
observed in repeated studies.
The second type of evidence consists of field experiments, which range from
modifying the conditions of enclosures in a lake (Muller 1980) to inten-
tionally acidifying an entire lake or stream (Schindler et al. 1980b, Hall et
al. 1980). These studies generally minimize the problem of confounding
factors, which plague field observation studies, and have contributed much to
our understanding of how organisms are affected by the acidification process.
However, experimental manipulations that focus on one variable may miss
effects which are due to the interaction of several variables. For example,
acidifying an entire lake may not reveal a major reason for fish kills in
waters acidified by acidic deposition, namely aluminum released when the
surrounding watershed is also acidified. A great difference also exists
between the time scale of experimental acidifications (which typically occur
over a period of months or a few years) and of regional acidification (which
occurs over many years).
The third type of evidence consists of laboratory experiments, whereby the
effect of a particular stress (low pH, aluminum) is evaluated after all other
variables are carefully controlled. These experiments typically consist of
bioassays involving one species and one or a small number of stresses. Most
of our understanding of the physiological effects of low pH on aquatic
organisms is due to such studies. As with field experiments, these studies
are time consuming, expensive and have yielded data on only a few species.
Predicting community-level changes from laboratory bioassays on a few species
is difficult. A species may experience reduced growth or reproduction in the
laboratory at a low pH, but may prosper in an acidified lake at the same pH
if its competitors suffer even greater reductions in growth and reproduction.
It is obvious that all three types of evidence provide certain kinds of
information yet have certain drawbacks. The strongest conclusions regarding
the effects of acidification on aquatic organisms will be reached when all
three types of evidence yield consistent results. Examples of such cases are
given in the conclusions section (Section 5.10.1).
The significance of changes in species abundances or community composition
lies in how these changes affect important ecosystem processes. These
processes include primary production (the production of new plant tissue
through photosynthesis), nutrient recycling (re-use of nutrients released
through decomposition of organic material), and trophic interactions
(transfer of energy from plants to herbivores to carnivores). A schematic
presentation of these processes and how they may be affected by acidification
is given in Section 5.8 (Figure 5-17). While direct toxic effects of
5-2
-------
acidification on organisms have been relatively easy to document, assessing
effects on ecosystem processes has proven more difficult. We know, for
example, that certain species of algae become dominant under acidic
conditions, yet how this affects the food supply to higher trophic levels or
how total primary productivity is affected has not been well studied. The
growth of algal mats in acidified lakes has been observed, yet how this seal
over the bottom sediments will affect nutrient cycling has not been measured.
Most effort to date has involved describing responses of various taxa to the
acidification process. Future work will need to consider how these changes
affect ecosystem processes.
5.2 BIOTA OF NATURALLY ACIDIC WATERS (F. J. Rahel)
Naturally acidic lakes and streams occur throughout the world and have been
known in the United States since at least the 1860's (Hutchinson 1957,
Patrick et al. 1981). These naturally acidic waters provide insight into the
pH range normally tolerated by aquatic organisms. Such information is
useful in assessing how recent pH declines attributed to cultural
acidification might affect aquatic life. This section's purpose is to
summarize the literature on naturally acidic waters and to examine the
influence of low pH on plants and animals found in such habitats. North
American waters are emphasized, but reference to other geographic areas is
made when cosmopolitan taxa are involved.
5.2.1 Types of Naturally Acid, ;ters
Naturally occurring acidic wat rail into three groups. In the first group
are inorganic acidotrophic waters associated with geothermal areas or lignite
burns, where pH values between 2.0 and 3.0 are not uncommon (Waring 1965,
Brock 1978, Hutchinson et al. 1978). Among the most extreme values recorded
are pH 0.9 for Mount Ruapehu Crater Lake, New Zealand (Bayly and Williams
1973), pH 1.7 from Kata-numa, a volcanic lake in Japan (Hutchinson 1957), and
pH's below 2.0 for several springs in Wyoming (Brock 1978). The high acidity
is due to sulfuric acid, which arises from the oxidation of sulfides such as
hydrogen sulfide (H2$) and pyrite (FeSg). In addition to being extremely
acidic, these waters frequently contain elevated metal concentrations and are
often heated. Assessing the biological effects of low pH under these condi-
tions is difficult, but such sites have provided insight into the lower pH
limit for various taxa (Brock 1973, 1978). This type of naturally acidic
aquatic habitat occurs in North America mainly in the west, and has been most
extensively studied in the Yellowstone Park region of Wyoming (Van Everdingen
1970, Brock 1978).
The second group of naturally acidic waters consists of brownwater lakes and
streams associated with peatlands, cypress swamps, or rainforests, depending
on latitude (Janzen 1974, Moore and Bellamy 1974). Their acidity is derived
from organic acids leached from decayed plant material and from hydrogen ions
released by plants such as Sphagnum mosses in exchange for nutrient ions
(Clymo 1967). These waters commonly have pH's in the range of 3.5 to 5.0 and
owe their dark color to large amounts of dissolved organic matter. As with
acidic geothermal waters, brownwaters have other qualities besides low pH
that may limit aquatic life. For example, they are characterized by low
5-3
-------
concentrations of many of the inorganic ions necessary for plant growth and
osmotic balance in animals (Clymo 1967). There is some evidence that the
dissolved humic compounds may be toxic to amphibians, even at neutral pH
(Gosner and Black 1957, Saber and Dunson 1978). Low oxygen and high carbon
dioxide concentrations are also present in some brownwater habitats (Welch
1952, Kramer et al. 1978). Finally, the low primary productivity of brown-
waters may mean that even physiologically tolerant species may be excluded
due to food scarcity (Janzen 1974, Bricker and Gannon 1976). Brownwater
habitats in North America are associated with either northern peatlands
(Jewell and Brown 1929, Cole 1979, Johnson 1981) or with southeastern
swamplands (Beck et al. 1974, Forman 1979, Kirk 1979).
The third type of naturally acidic habitat consists of ultra-oligotrophic
waters. They are especially common where glaciation has removed younger
calcareous deposits and exposed weather-resistant granitic and siliceous
bedrock. The absence of carbonate rocks in the drainage basin results in
lakes with little carbonate-bicarbonate buffering capacity; hence such lakes
are very vulnerable to pH changes. They often have pH's in the 5.5 to 6.5
range, and most of the acidity appears due to carbonic acid (HgCC^).
These lakes tend to be small and have low concentrations of dissolved ions.
In North America, this type of naturally acidic lake occurs in large areas of
eastern Canada and the northeastern United States, as well as in sections of
western United States and northern Florida (Shannon and Brezonik 1972,
Galloway and Cowling 1978). Many of the lakes which have been, or will be,
affected by acidic deposition belong in this category (see Chapter E-4,
Section 4.3.2).
5.2.2 Biota of Inorganic Acidotrophic Waters
In North America, the most extensively studied inorganic acidotrophic waters
are those of the Yellowstone Park region in Wyoming. Certain species of
eucaryotic algae, fungi, and bacteria have demonstrated remarkable adaptation
to this acidic environment and often form extensive mats (Brock 1978). For
example, the alga Cynani diurn caldar iurn was found at pH 0.05, while the bacte-
rium Sulfolobus acidocaldarius thrived in a thermal spring at pH 0.9 and 60
C. Lower pH limits for other taxa in this environment are summarized by
Brock (1978) and include a pH near 0.0 for fungi, pH 3.0 for Sphagnum mosses,
and pH 2.5 to 3.0 for vascular plants such as sedges (Carex and Eleocharis
spp.) and ericacid shrubs (blueberries, cranberries). Although generally
considered eurytropic, blue-green algae are conspicuously absent from these
acidic environments. Brock (1973, 1978) has assembled data showing that
these algae are intolerant of pH's below 4.0. The inability to survive under
acidic conditions may be due to their lack of membrane-bound chloroplasts
that, in eucaryotic algae, prevent the acid-labile chlorophyll from being
decomposed at low pH.
In ponds exposed to sulfur fumigations from burning bituminous shales, the
euglenoid Euglena mutablis was present at pH 1.8 (Hutchinson et al. 1978,
Havas and Hutchinson 1982). The red chironomid, C h i ronomu s ri pa r i u s, and the
rotifer, Brachionus urceolaris, were abundant at pH 2.8, but no copepods or
cladocerans were present.
5-4
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Among the few Insects reported from acidic thermal waters is the ephydrid fly
Ephydra thermophila (Brock 1978). This fly breeds in streams at pH 2.0 and
is the basis ofa food chain involving several invertebrate predators.
Extensive surveys of invertebrates in the acidic geothermal waters of North
America have not been done, but it seems reasonable that other invertebrate
taxa might tolerate such low pH. For example, in streams polluted by acidic
mine wastes, species of rotifers, midges, alderflies and dytisscids have been
found at pH's near 3.0 (Roback 1974, Harp and Campbell 1967, Parsons 1968).
Vertebrates such as amphibians and fish appear unable to survive in inorganic
acidotrophic habitats, but again no extensive surveys have been undertaken.
Surprisingly, waterfowl do not avoid these lakes, and Canadian geese have
been reported to nest on Turbid Lake in Yellowstone Park (pH - 3.0) (Brock
1978).
Another group of inorganic acidotrophic lakes that have been well studied are
the volcanic lakes of Japan (Ueno 1958). Some of the organisms present in
these lakes belong to cosmopolitan genera and hence provide insight into the
lowest pH that may be tolerated by North American genera. Aquatic mosses
(e.g., Rhynchostegium aplozia) dominate the plant community, although reeds
(Phragmites) occur "along the margins of most lakes, even at pH's below 3.0.
Diatoms (Pinnularia) and rotifers (Rotaria) have been observed at pH 2.7. A
small caldera lake filled with water at pH 3.0 but fertile enough to support
moderate phytoplankton production contained several genera of Crustacea
(Simocephalus, Chydorus, Macrocyclops) and a rotifer (Brachionus). The
teleost Tnbolodon hakonensis from Lake Osoresan-ko (pH 3.5) occurs at the
lowest pH reported for any fish species (Mashiko et al. 1973).
While the work done on inorganic acidotrophic waters has revealed some out-
standing examples of extreme pH tolerance, in general, these waters have very
low species diversity and monocultures of tolerant species are common.
5.2.3 Biota in Acidic Brownwater Habitats
Brownwater habitats do not experience the extremes of temperature, pH, and
metal concentrations common to inorganic acidotrophic waters; consequently
they contain a greater diversity of organisms. They are, however, charac-
terized by low ion concentrations, reduced light penetration and, frequently,
low dissolved oxygen concentrations. These variables interact with the
acidic pH (3.5 to 5.0) to determine species richness and biological
production.
Among the genera of macrophytes reported from acidic brownwater lakes are
Alternanthera, CeratophyHum, Isoetes, Juncus, Limnobium, Nuphar, Potamogeton
and ytricuTaria (Jewell and Brown 1924, Griffiths 1973. Stoneburner and Smock
1980TMany brownwater lakes, however, are characterized by the absence of
macrophytes, which is generally attributed to the stained water and the lack
of a firm substrate on the lake bottom (Welch 1952, McLachlan and McLachlan
1975, Marshall 1979). The shoreline plant community has been well described
for northern bogs and includes sedges (Carex), ericacid shrubs (Vaccim'um
chamaedaphe) and mosses (Sphagnum) (Gates 1942, Heinselman 1970, Vitt and
5-5
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Slack 1975). The characteristic tree along the shore of southeastern
brownwater lakes is the cypress (Taxodium) (King et al. 1981).
Phytoplankton have classically been described as present at low densities
(Birge and Juday 1927, Welch 1952, Stoneburner and Smock 1980). Recent work
has emphasized the predominance of small-bodied algae (the nanoplankton) in
these waters (Bricker and Gannon 1976). Although species from most phyto-
plankton phyla have been reported, certain genera of desmids (Xanthidium,
Euastrum, Hyalotheca) and diatoms (Asterionella, Eunotia, Actinella,
Anomoeoneis, Pinnularia, Melosira) are especially characteristic (Woelkerling
and Gough 1976, Marshall 1979, Patrick et al. 1979, Stoneburner and Smock
1980). As with the phytoplankton, the zooplankton in acidic dystrophic lakes
are frequently dominated by small-bodied forms, particularly rotifers
(Brachionus, Keratell a, Monostyla, Polyarthra) and copepods (Diaptomus,
Cyclops) (Welch 1952, Smith 1957. Bricker and Gannon 1976, Marshall 1979).
Relatively few cladocerans have adapted to this environment although species
from the following genera have been reported: Alona, Bosmina, Chydprus,
Daphnia, Diaphanosoma, Eubosmina, Leptodpra, and PTeuroxus (Marshall 1979,
Von Ende 19/9, Stoneburner and Smock 1980). In lakes where fish are absent
or where darkly stained water and low hypolimnetic oxygen offer some protec-
tion from fish predation, dipteran larvae of the genus Chaoborus are an
important part of the zooplankton community (Von Ende 1979).
A peculiar phenomenon in many acidic brownwater lakes is the large standing
crop of zooplankton relative to phytoplankton. This paradox has lead to
suggestions that bacteria and suspended organic matter (tripton) may be
important food sources for zooplankton in these lakes (Bayly 1964, Bricker
and Gannon 1976, Stoneburner and Smock 1980).
The benthic community in acidic dystrophic lakes is typically impoverished.
This is particularly true of small bogs where a deep layer of decaying peat
obliterates any sand or gravel substrate and prevents macrophyte growth.
Such lakes have dipteran larvae (Chaoboridae and Chironomidae), dragonflies
and damselflies (Odonata), and alderflies (Sialidae) as their main benthic
invertebrates (Welch 1952, McLachlan and McLachlan 1975). Even habitats with
more diverse substrates still have few benthic species although caddisflies
(Trichoptera), whirligig beetles (Gyrinidae), and cranefly larvae (Tipulidae)
are sometimes present (Smith 1961, Patrick et al. 1979). Jewell and Brown
(1929) described an interesting invertebrate community living in pools in the
sphagnum mat of a Michigan bog at pH 3.5 to 4.0. Air-breathing forms like
beetles (Dytisicidae, Haliplidae, Helodidae, Hydrophilidae) and mosquito
larvae (Culex) predominated in these low-oxygen pools, although several
dragonfly species (Odonata) and the cladoceran, Acantholebris curvirostri,
were also present.
Notably absent from acidic bog waters are mayflies (Ephemeroptera);
crustaceans such as amphipods, ostrocods and crayfish; molluscs (snails,
clams); sponges; and annelids (oligochaetes, leeches) (Pennak 1953, Wetzel
1975). The absence of organisms that have a calcified exoskeleton is not
unexpected in brownwater habitats due to the low pH and the extremely low
concentration of calcium. An exception to this generalization is the
5-6
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occurrence of the fingernail clam (Pisidium) in bog lakes at pH's below 5.0
(Griffiths 1973).
Summaries of fish species distribution in relation to pH exist for both
northern and southern brownwater habitats (Frey 1951, Hastings 1979, Rahel
and Magnuson 1983). Slow growth and low species diversity characterize the
fish assemblages in these waters (Smith 1957, Garton and Ball 1969). In
northern midwestern lakes where ice cover occurs, winter anoxia interacts
with pH to determine the structure of fish assemblages (Rahel 1982). Lakes
with adequate winter oxygen concentrations are dominated by yellow perch
(Perca flavescens), sunfish (family Centrarchidae), and bullheads (Ictalurus
spp.), even down to pH 4.5. If winter oxygen concentrations are low enougTi
to exclude predators, minnows (family Cyprinidae) dominate the fish fauna,
but only if the pH is above 5.2 to 5.4. Lakes that are both very acidic (pH
below 5.2) and experience winter anoxia contain only yellow perch and the
central mudminnow (Umbra limi). Other species that can survive in acidic
northern brownwaters but are probably excluded because suitable habitat or
spawning areas are missing are the northern pike (Esox lucius) and brook
trout (Salvelinus fontinalis) (Jewell and Brown 1924, Smith 1961, Dunson and
Martin 1973).
Southeastern brownwater lakes and streams (pH 4.0 to 5.0) have a more diverse
fish fauna than do similar northern waters (Wiener and Giesy 1979, Frey 1951,
Laerm et. al 1980). Among the more common taxa are various species of
sunfish, pickerel (family Esocidae), catfish (family Ictaluridae), and
killifish (family Cyprinidontidae), along with the American eel (Anguilla
rostrata), lake chubsucker (Erimyzon sucetta), eastern mudminnow(Umbra
pygmaea), pirate perch (Aphredoderus sayanusl, and the yellow perch.
With the exception of the golden shiner (Notemigonus crysoleucas), ironcolor
shiner (Notropis chalybaeus), and the swamp darter (Etheostoma fusi forme),
minnows and darters are conspicuously absent from acidic brownwaters, even
though they may be abundant in nearby neutral waters (Frey 1951, Laerm et al.
1980, Rahel and Magnuson 1983). Predation from bass and pike may exclude
these small-bodied fishes from many habitats, but even when predators are
absent, minnows and darters are rarely found below pH 5.2. Other acid-
sensitive species are the smallmouth bass (Micropterus dolomieui) and walleye
(Stizostedion vitreutn).
5.2.4 Biota in Ultra-01igotrophic Waters
The third category of naturally acidic waters consists of ultra-oligotrophic
lakes and streams. Hydrogen ion concentrations fluctuate in these waters as
a function of photosynthetic activity and carbon dioxide concentrations, with
pH typically varying between 5.5 and 7.0. Low nutrient concentrations result
in low biological productivity at all trophic levels. Most aquatic taxa are
able to tolerate the hydrogen ion concentration of these lakes and thus other
physical/chemical factors (e.g., thermal conditions) or biotic interactions
(predation and competition) are important in determining species composition.
A great diversity of taxa has been reported from ultra-oligotrophic lakes,
but certain groups are characteristic of this lake type. In the
5-7
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phytoplankton, for example, chrysophytes and diatoms (Chrysophyta) along with
desmids and other green algae (Chlorophyta) are diagnostic of oligotrophic
conditions (Hutchinson 1967). Numerous other algae are usually present at
low densities (Schindler and Holmgren 1971, Baker and Magnuson 1976).
Copepods appear to dominate the zooplankton community, but numerous other
taxa have been recorded in surveys of oligotrophic waters (Patalas 1971,
Torke 1979). Factors such as lake depth and size, thermal regimes,
phytoplankton abundance, and fish predation appear to be more important than
pH in determining zooplankton community structure in these lakes (Anderson
1974, Green and Vascotto 1978).
Benthic communities are diverse, although certain genera of midge larvae
(Tanytarsus, Chapborus) along with fingernail clams (Pisidium), the amphipod
Pontopqrela, and the mysid Mysis relicta have classically been associated
with oligotrophy (Hamilton 1971, Brinkhurst 1974, Wetzel 1975). In acidic
streams (pH less than 5.7), mayflies (Ephemeroptera), molluscs, some
caddisfly genera (Hydropsyche), and the amphipod Gammarus are rare, even
though they are abundant in downstream sections having a higher pH (Sutcliffe
and Carrick 1973). These taxa are also missing from streams affected by
acidic mine drainage (Roback 1974). Shell-forming molluscs and crustaceans
may be excluded from oligotrophic waters because of low calcium concen-
trations, even though the pH is circumneutral. Crayfish, for example, were
absent from softwater Wisconsin lakes having calcium concentrations below 2
mg r1 regardless of lake pH (Capelli 1975).
Aquatic macrophytes typical of oligotrophic waters have been summarized by
Hutchinson (1967) and Seddon (1972). Among the representative genera are
Bi dens, El a tine, EHocaulon, I soetes, Juncus, Lobelia, and Sparganium. Most
of these liave a disltfncYl)hysical form, consisting of stiff leaves placed in
a close rosette or on short, unbranched stems as opposed to the long-stemmed,
branched leaf typical of hardwater macrophytes (Fasset 1930). Species
occurring in oligotrophic waters are probably not restricted to the low
nutrient conditions present there but are likely excluded from more fertile
waters by competition from other macrophyte species (Hutchinson 1967).
Identifying fish assemblages typical of oligotrophic waters is complicated by
human activities that affect community composition, such as stocking,
over-exploitation, and eutrophication (Regier and Applegate 1972). Many
high-elevation Palearctic lakes were probably barren of fish following
deglaciation, although the very long and poorly-documented history of fish
introductions by humans makes it impossible to know what percent were
fishless (Nilsson 1972, Donald et al. 1980). These coldwater lakes today are
dominated by salmonids (trout and salmon) and coregonids (whitefish and
ciscoes). Oligotrophic lakes with slightly warmer thermal regimes (because
they are shallower or are located at lower altitudes or farther south than
the salmonid lakes) are dominated by percids (yellow perch) and certain
centrarchids (typically the smallrnouth bass, Micropterus dolomieui) and rock
bass (Ambloplites rupestn's) (Adams and Olver 1977, Rahel and Magnuson 1983).
As with the other faunal groups, the low productivity and biotic interactions
(predation/competition) of these lakes probably have a greater influence on
5-8
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the fish species composition than pH per se. For example, many small-bodied
fish species (e.g., minnows and darters) are commonly absent from oligotro-
phic lakes even though they can tolerate the pH's typical of these waters
(Rahel and Magnuson 1983). Competition, or more likely predation by larger
species, may exclude these fish from biologically unproductive lakes where
there are few macrophytes to provide refuges. Another example involves
yellow perch and whitefish (Coregonus spp.) which successfully coexist only
in large, cold lakes where the pelagic whitefish can avoid competition from
the more littoral-based yellow perch (Svardson 1976).
5.2.5 Summary'
Naturally acidic waters provide insight into the lowest pH tolerated by
various groups of aquatic organisms (Table 5.1). While life has been found
in the most acidic environments sampled, the general observation is that
species diversity declines as pH decreases. The most tolerant organisms are
from the lower trophic levels, with some bacteria and algae able to flourish
at pH's below 1.0. Invertebrates are rarely found below pH 3.0, and fish are
generally limited to pH's above 4.0. Some organisms (especially certain
genera of bacteria) are true acidophiles, unable to grow and reproduce at
neutral pH (Brock 1978). However, most organisms occurring in acidic
environments survive quite well at neutral pH but are excluded from such
environments by competitively superior species.
Species distributions in natural pH gradients provide a means of assessing
the long-term effects of low pH exposure, integrated over all life history
stages and all physiological functions. Such information is seldom obtained
in laboratory bioassays, which are generally short-term, focused on one or
two physiological responses, and ignore the potential for genetic adaptation
to acid stress. Species' acid sensitivity inferred from distributions in
naturally acidic waters may be useful in selecting species to monitor in
waters undergoing cultural acidification. For example, acid-tolerance rank-
ings of fish species, based on distributions among naturally acidic Wisconsin
lakes (Figure 5-1), were correlated with acid-tolerance rankings from cultur-
ally acidified Canadian lakes (Figure 5-2). This allowed predictions of
which fish species should be monitored in Wisconsin lakes susceptible to
acidification (Rahel and Magnuson 1983).
Studies of species distributions relative to pH are subject to misinterpre-
tation if other correlated factors are not adequately considered. Among the
factors that can interact to influence species distributions are pH, metal
concentrations, and temperature in geothermal waters; pH, oxygen concentra-
tions, and substrate composition in dystrophic waters; and pH, low nutrient
concentrations, and predation in ultra-oligotrophic waters. The problem of
separating out the effects of confounded factors is illustrated by work on
the distribution of rotifers in Wisconsin lakes. Alkaline waters (above pH
7.0) contained relatively few species of rotifers but large numbers of
individuals. In contrast, waters with pH below 7.0 contained large numbers
of species but few individuals (Pennak 1978). Hence, rotifer species
diversity increased with decreasing pH. However, this was probably because
competitive interactions were influenced by factors correlated with pH, not
because most species of rotifers could not tolerate neutral pH. In another
5-9
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TABLE 5-1. LOWER pH LIMITS FOR DIFFERENT GROUPS OF ORGANISMS IN
NATURALLY ACIDIC WATERS
Approx.
Group Lower
PH
limit
Bacteria
Plants
Fungi
Eucaryotic
algae
Blue-green
algae
Vascular
plants
Mosses
Animal s
Protozoa
Rotifers
Cladocera
Copepods
Insects
Amphipods
Clams
Snails
Fish
0.8
2-3
0
0
1-2
4.0
2.5-3
3.0
2.0
3.0
3.5
3.0
3.0
3.6
2.0
3.0
5.8
5.8
4.5
6.0
5.8
6.2
3.5
4.0
4.5
Examples of Species
Occurring at Lower pH Limit
Thiobacillus thiooxidans,
Suifolobus acidocaldarius
Bad 1 1 us , Streptomyces
Acontium velatum
Cyam'dium caldarium
Euglena mutabilis,
Chi amydomonas acidophila,
Chi orel la
Mastigocladus, Synechococcus
Eleocharis, Carex,
Ericacean plants,
Phragmites
Sphagnum
Amoebae, Heliozoans
Brachionus, Lecane, Bdelloid
Collotheca, Ptygura
Simocephalus, Chydorus
Macrocyclops
Cycl ops
Ephydra thermophila
Chironomus riparius
Mayflies
Gammarus
Pi si di urn
Most other species
Amnicola
Most other species
Tribolodon hakonensis
Umbra 1 imi
Sunfishes
(Centrarchidae)
Reference
Brock 1978
Brock 1978
Brock 1978
Brock 1978
Brock 1978
Brock 1978
Brock 1978
Hargreaves et al . 1975
Ueno 1958
Brock 1978
Brock 1978
Hutchinson et al . 1978
Edmondson 1944
Ueno 1958
Ueno 1958
Hutchinson et al . 1978
Brock 1978
Hutchinson et al . 1978
Sutcliffe and Carrick
1973
Sutcliffe and Carrick
1973
Griffiths 1973
Pennak 1978
Pennak 1978
Pennak 1978
Mashiko et al . 1973
Rahel and Magnuson 1983
Rahel and Magnuson 1983
5-10
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COMMON NAME
CENTRAL MUDMINNOW
YELLOW PERCH
BLACK BULLHEAD
BLUEGILL
LARGEMOUTH BASS
WHITE SUCKER
YELLOW BULLHEAD
PUMPKINSEED
GOLDEN SHINER
NORTHERN REDBELLY
BROOK STICKLEBACK
PIKE
ROCK BASS
MOTTLED SCULPIN
SMALLMOUTH BASS
MUSKELLUNGE
BLACK CRAPPIE
BURBOT
CREEK CHUB
CISCO
IOWA DARTER
JOHNNY DARTER
REDHORSE
COMMON SHINER
MIMIC SHINER
TROUT-PERCH
BLUNTNOSE MINNOW
LOGPERCH
BLACKNOSE SHINER
FATHEAD MINNOW
NUMBER OF LAKES IN A GIVEN pH RANGE
(50 lakes pH > 7.0)
FAMILY 7
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pH RANGE
.0 6.0 5.0 4.
1 i i I »
i i i i i
NUMBER OF
0 LAKES
en
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84
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oc
-------
4.0 5.0 6.0
pH
(Naturally Acidic Lakes)
7.0
1. YELLOW PERCH
2. BLUEGILL
3. LARGEMOUTH BASS
4. PUMPKINSEED
5. WHITE SUCKER
6. GOLDEN SHINER
7. NORTHERN PIKE
8. ROCK BASS
9. SMALLMOUTH BASS
10. IOWA DARTER
11. JOHNNY DARTER
12. BLUNTNOSE MINNOW
13. COMMON SHINER
Figure 5-2. Lowest pH at which fish species appeared unaffected in
culturally acidified lakes (Harvey 1980) compared to the
lowest pH at which they occurred in naturally acidic lakes,
Diagonal line separates species occurring at lower pH .in
naturally acidic lakes (upper) from those that occur at a
lower pH in culturally acidified lakes (lower). Adapted
from Rahel and Magnuson (1983).
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example, Weiner and Hanneman (1982) failed to find a relationship between
reduced fish growth and low pH in a set of naturally acidic Wisconsin lakes,
even though growth reductions at low pH are consistently observed in labora-
tory bioassays (Section 5.6.4.1.3). They attributed the lack of correlation
between fish growth and pH to the overriding effects of population density.
Experimental manipulations offer potential for separating the effects of
these confounding factors from the effects of pH. A good example is the
alkalinization of an acidic brownwater lake (Smith 1957). When the pH was
raised by adding lime, several stocked fish species reproduced successfully
for the first time. However, as the pH returned to its former level,
reproduction stopped, suggesting that hydrogen ion concentration was the
limiting factor.
In some cases, naturally acidic environments are free of the confounding
stresses associated with culturally acidified environments. This is espe-
cially true of metal toxicants, which are common in waters affected by acidic
mine drainage or acidic deposition (Parsons 1977, Cronan and Schofield 1979)
but rare in acidic brownwater and ultra-oligotrophic lakes. As a result of
organic complexation, comparison of fish species distributions relative to pH
in these different water types has helped to identify aluminum toxicity, not
pH, as the major reason for fish kills in lakes affected by acidic deposition
(Muniz and Leivestad 1980a).
Data on the biota of naturally acidic environments will continue to be
instructive in studies of culturally acidified waters and should be especial-
ly useful in evaluating the long-term effects of chronic acid stress.
This section is summarized as follows:
1. Naturally acidic lakes fall into three major groups:
0 inorganic acidotrophic waters (pH commonly less than 4.0)
0 dystrophic waters (pH commonly 3.5 to 5.0)
0 ultra-oligotrophic waters (pH commonly 5.5 to 7.0)
2. In naturally acidic waters, hydrogen ion concentration can be
strongly implicated as limiting the occurrence of:
° invertebrates with calcified exoskeletons below pH 5.5
(mayflies, Gammarus. snails, clams)
0 blue-green algae below pH 4.0
0 some species of minnows (Cyprinidae) and darters (Percidae)
below pH 6.0
0 several species of sunfish (Centrarchidae) below pH 4.5
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These pH limits for survival and reproduction are similar to those
observed in culturally acidified waters.
3. Lower safe pH limits inferred from a species distribution among
naturally acidic waters may not always be valid for culturally
acidified waters. For example, these limits may be:
o too low if other stresses (e.g., aluminum) are present in
culturally acidified lakes, or
0 too high if species are absent from naturally acidic lakes
because of factors other than low pH: e.g., high temperature
or metals in inorganic acidotrophic waters; low sodium, and
calcium concentrations or unsuitable habitat in dystrophic
waters.
5.3 BENTHIC ORGANISMS (R. Singer)
5.3.1 Importance of the Benthic Community
The term benthos refers to the community of organisms which live in and on-
bottom sediments of lakes and streams. The following groups are important
components of the benthos: microbes, periphyton, microinvertebrates,
Crustacea, Insecta, Mollusca, and Annelida. These organisms interact with
biological and chemical components of the water column by processing
detritus, recycling inorganic nutrients, mixing sediments, and serving as a
principal food source for fish, waterfowl, and riparian mammals. Most of the
energy and nutrients in lakes and streams ultimately passes through the
benthos, so any alteration of this community is likely to affect plankton,
fish, and water chemistry. Studies of the effects of acidic deposition on
this community have begun only recently (Singer 1981a), and not all benthic
components have received equal treatment.
Microbes rapidly colonize the surfaces of leaf litter and other organic
debris. Many benthic macroinvertebrates then process the debris, further
facilitating its decomposition by microorganisms. Macroinvertebrate
"shredders" rip and chew leaves, vastly increasing surface area, and partial-
ly digest material as it passes through their guts. Without these
invertebrates, organic detritus decomposes very slowly (Brinkhurst 1974).
After the macroinvertebrates have broken up the detritus, fungi, bacteria,
and protozoans complete the digestion and release inorganic nutrients into
the water. The pH of the water in part controls the solubility equilibria of
these inorganic constitutents and largely determines whether they will be
available for recycling by plants. In addition, the rate of decay depends on
the metabolic efficiency of this rnicrobial community, which is also pH
dependent (Laake 1976, Gahnstrom et al. 1980).
Macroinvertebrates aerate sediments by their burrowing movements. The top
few centimeters of sediments generally demonstrate large gradients of pH, Eh
(oxidation-reduction potential — the concentration of free electrons), dis-
solved 02, and other constituents (Hutchinson 1957). Losses or alterations
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of plant and animal communities have profound effects on the chemistry of
this top layer of sediments (Mortimer's "oxidized microzone" 1941, 1942), yet
little work has centered on this habitat in acidified lakes. Mitchell et al.
(19815) found that the presence of burrowing mayflies (Hexagenia) affected
sulfur dynamics in sediment cores taken from acidic lakes. Sediment/water
column biological and chemical interactions are difficult to study because
events occur across strong chemical gradients over short distances (Mitchell
et al. 1981a). These gradients are easily perturbed by experimental
procedures, including in situ measurements. Despite these procedural
difficulties, it is important to determine the influence of pH-related
alterations of the sediment community on the chemistry and biota of the water
column.
Benthic animals are at the base of most food chains that lead to game fish.
It has been suggested that eliminating the amphipod Gammarus lacustris
(Section 5.3.2.4) and most molluscs (Section 5.4.2.6) might reduce trout
production by 10 to 30 percent (0kland and 0k1and 1980); however this
prediction has not been verified. Rosseland et al. (1980) reported that
trout in acidified waters shifted their diet from acid-sensitive inverte-
brates such as mayflies and bivalves to acid-tolerant forms such as corixid
bugs and beetles. Although decline of fish populations due to alteration of
the benthic community has not been studied, stress on fish populations as a
result of nutrient changes should be considered. Fish fry, which are more
dependent on smaller invertebrate prey than are adults, might be more sensi-
tive to changes in the benthic community. These effects have not been
considered experimentally, however.
Finally, changes in the benthic plant community (Section 5.5) affect
macroinvertebrate distribution. The littoral habitat is an important area
for benthos, and alterations in plant community structures are likely to
affect all other trophic levels. These interactions remain to be investi-
gated, but Eriksson et al. (1980a) have suggested that many of the observed
changes in water chemistry and plankton communities are due to biological
alterations, not direct chemical toxicology. They reported an increase in
water clarity, alteration of planktonic communities, and even a drop in pH
(by 0.5 units) when fish were eliminated from a neutral lake by poisoning.
The results extend and verify similar work reported by Stenson et al. (1978).
Sources of energy to benthos include primary production by higher plants
(macrophytes) and attached algae (periphyton), and energy derived from
detritus raining from the water column above (autochthonous inputs) and from
detritus washed into the basin (allochthonous inputs). Lakes (lentic
systems) receive most of their energy from autochthonous sources, whereas
streams (lotic systems) derive their energy from primarily outside,
allochthonous sources (e.g., Wetzel 1975). Consequently, shredding and
scraping benthic insects and crustaceans are relatively more important in
streams than lakes, while detritus-consuming worms and midges are more
abundant in lakes.
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5.3.2 Effects of Acidification on Components of the Benthos
The diversity of benthic organisms is often confusing to non-specialists. It
must be emphasized that the loss of fish populations, although one of the
most observable effects, is neither one of the earliest nor only biological
effect(s) of acidification; alterations in the benthic community integrate
annual loadings at levels of stress which are not observable in fish
populations. The ultra-oligotrophic lakes characteristic of sensitive areas
harbor ecosystems which are unique. These ecosystems may be damaged at
levels of acidification that may not affect fish at all. The concept of an
endangered ecosystem is as viable as the more generally accepted view of the
endangered species.
Using historical collections and known water quality requirements of organ-
isms allows specialists to generalize about past water chemistry parameters.
Moreover, the low mobility and long life cycles of many benthic organisms
allow one to make conclusions about the extremes of water quality fluctua-
tions in past years. However, experimentation on benthic communities is
difficult.
5.3.2.1 Microbial Community--Studies of the effects of acidification on
benthic protozoans have not been conducted. Other members of this community
include bacteria and fungi. It was reported that acidification of lakes
causes bacterial decomposers to be replaced by fungi (Hendrey et al. 1976,
Hendrey and Barvenik 1978) and proposed (Grahn 1976, 1977; Hultberg and Grahn
1976) that the shift to fungi accounts for the observed {Leivestad et al.
1976) accumulation of detritus in acidic lakes. Liming of lakes to increase
the pH brings a rapid restoration of normal microbial activity (Scheider et
al. 1975, 1976; Gahnstrom et al. 1980).
Traaen (1976, 1977) showed that leaf packs in lakes were processed much more
slowly at lower pH (5.0) than at higher pH (6.0) values, but he also cau-
tioned (1977) that many factors other than acidity can affect leaf
processing. Burton (1982) has confirmed the impact of low pH on processing
of organic matter. Friberg et al. (1980) reported an increased accumulation
of detritus and a reduction in numbers of scraping insects in an acidic (pH
4.3 to 5.9) stream as compared to a neutral (pH 6.5 to 7.3) stream. Hall et
al. (1980) and Hall and Likens (1980a,b) artificially acidified a stream in
Hubbard Brook, NH, and showed that scrapers were largely lost. In addition,
they reported that insects that feed by collecting debris were inhibited.
Hall et al. (1980) observed a growth of basidiomycete fungus on birch leaves
in an artificially acidified portion of a stream; such fungal growth was
lacking in the non-acidified control section. Hultberg and Grahn (1976) and
Grahn et al. (1974) described an accumulation of a "fungal mat" on the bottom
of many acidified Scandinavian lakes. It is now understood that this coarse
particulate material is a mixture of detritus, some fungi, and mostly algae
(Stokes 1981) (Section 5.3.2.2). The original description of this layer of
material as a "fungal mat" (Hendrey et al. 1976) was erroneous (Hendrey and
Vertucci 1980) due to the senescent, colorless state of the common blue-green
algal (Phormidium spp.) component of the mat.
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Some controversy exists regarding the effects on microbial metabolism brought
about by acidification (Baath et al. 1979). The accumulation of detritus in
acidic lakes suggests a reduction in decomposition by bacteria (Leivestad et
al. 1976). The reduction of oxygen utilization by acidified cores (Hendrey
et al. 1976) supports this view. Furthermore, liming increased oxygen
consumption of previously acidic cores (Gahnstrom et al. 1980). At pH levels
below 5.0, oxygen consumption, ammonia oxidation, peptone decomposition, and
total bacterial numbers all declined (Bick and Drews 1973). In contrast,
Schindler (1980) reported no change in decomposition rates in an artificially
acidified lake, and Traaen (1978) observed no clear differences in the
planktonic bacterial populations from seven lakes of pH < 5.0 as compared to
seven lakes of pH > 5.0 (see also Boylen et al. 1983). Traaen argued that
acidic inputs should affect the plankton populations prior to affecting
benthic algae. His results showed that the distribution of bacterial popula-
tions was more strongly influenced by organic inputs and temporal and spatial
(depth) patchiness than by pH. Gahnstrom et al. (1980) reported that inhibi-
tion of oxygen uptake by sediments increased in acidic lakes as compared to
reference lakes, only in the littoral sediments. They argued that the inhi-
bition of microbial activity in the littoral zone might be due to the inflow
of acidic runoff, which is restricted to the epilimnion during snowmelt and
autumn rains (Hendrey et al. 1980a). All these studies demonstrate that
decomposition of organic material is inhibited below pH 5.0 but not necessar-
ily by a reduction in standing crop of bacteria. The resulting accumulation
of organic matter undoubtedly affects water chemistry, fish habitats,
nutrient cycling, and primary productivity.
Microbial effects on other trophic systems probably involve alterations of
sulfur, nitrogen, and phosphorus dynamics. Methylation of mercury (Tomlinson
1978, Jernelov 1980) and other heavy metals may have profound effects on
higher trophic levels (Galloway and Likens 1979; refer also to Chapter E-6) .
The release of aluminum from sediments below pH 5.0 (Driscoll 1980) is
another potentially serious impact that has not been adequately studied.
5.3.2.2 Peri phytpn—The periphytic community of algae lives attached to
macrophytes and directly on sediments and makes important contributions to
primary production and nutrient cycling, particularly in lotic (stream)
systems. Changes in the species composition of this community reflect
changes in the chemistry of both the water column and the sediments. These
algae are an important food source for the grazing macroinvertebrates that
are a principal source of food for fish. Algal seasonal growth and decompo-
sition store and periodically release nutrients and other ions.
5.3.2.2.1 Field surveys. Acidic lakes develop periphytic communities domi-
nated by species known to prefer acidic water, and dramatic decreases in
species diversity below pH 5.5 have been observed (Aimer et al. 1974; see
Section 5.2). One of the most striking aspects of many acidified lakes is
the presence of a thick mat of algae which overlies the substrate. This mat
overgrows all the rooted plants and, to a large degree, physically and chemi-
cally isolates the lake bottom from the overlying water. The mats vary in
shape, texture, and species composition from lake to lake, seemingly
irrespective of water chemistry parameters. Three types of mats were
described by Stokes (1981):
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1) Cyanophycean mats, dominated by the blue-green algae
Osci'llatoria sp., Lyngbya sp., and Pseudoanabaena sp. in Sweden
at pH 4.3 to 4.7 (Lazarek 1980) and Phormidium sp. in New York
at pH 4.8 to 5.1 (Hendrey and Vertuccl 1980). These mats are
dark blue-green with occasional flecks of orange-colored
carotene-rich material. They are thick, felt-like, and encrust-
ing. Stokes reported cyanophycean mats at depths of 2 to 3 m,
but they have been observed as deep as 5 m in an acidic (pH <
4.9) Adirondack lake (Singer et al. 1983).
2) Chlorophycean mats, dominated by green algae such as Mougeotea
sp. and Pleurodiscus sp. at pH 3.9 to 5.0 in Canadian lakes
(Stokes 1981).These mats are coarser than cyanophycean mats.
They tend to be loosely packed, green to reddish purple, and may
extend to 4 m deep. Unlike cyanophycean mats, chlorophycean
mats are not compacted and do not retain their structural
integrity when lifted. A chlorophycean mat developed after the
experimental acidification of a whole lake was completed
(Schindler and Turner 1982).
3) Chlorophycean epiphytic or periphytic algae, dominated by green
algae such as Spirogyra sp., Zygnema sp., Pleurodiscus sp., and
Mougeotea sp., Oedogonium, and Bulbochaete. This community
appears as bright grass-green clouds hanging from macrophytes
and resting lightly on the bottom. They appear around pH 5.0
and have been reported in Canada (Stokes 1981), the Adirondacks
(Hendrey and Vertucci 1980), and Sweden (Lazarek 1982). They
also appeared in artificially acidified channels (Hendrey 1976),
in artificially acidified cylinders (Muller 1980, Van and Stokes
1978), and in an artificially acidified lake at pH 5.6
(Schindler 1980).
I have observed all three types of mat communities in a survey of five
Adirondack lakes below pH 4.9. These lakes were all about the same size
( -30 ha), low in nutrients, located near each other, and similar in
morphometry. Why one community dominates one lake but is not found in
another is unknown. Part of the explanation may be that the three types of
mats may represent stages in a pattern of seasonal succession. Lazarek
(1982) has reported seasonal succession among epiphytes from one acidic (pH
4.3 to 4.7) Swedish lake. As these mats are the most conspicuously visible
characteristics of acidified lakes, their significance and effects on other
physical and chemical components deserve more attention.
5.3.2.2.2 Temporal trends. The shells of diatoms (Bacillariophyceae) are
made of Si02 and are very resistant to weathering. Deposition of plank-
tonic and benthic diatoms to sediments produces a record of the past popula-
tions in the lake once the cores are dated by radioactive decay (Norton and
Hess 1980). The pH tolerance of many diatoms has been tabulated elsewhere
(e.g., Lowe 1974). Thus the ancestral pH may be inferred from the strati-
graphic record. This technique is subject to variances caused by macro-
invertebrate mixing, local changes in pH sensitivity of species, and the
numerous other factors besides pH that determine the distribution of species
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(Norton et al. 1981). Nonetheless, inferred pH generates a value that
reflects the real water column value to within < 1.0 pH unit, which often
lies within the range of normal seasonal pH variation. The method's accuracy
is even better for comparing groups of lakes with similar current pH values
along a temporal gradient of past pH levels by regression analysis (Norton et
al. 1981). The inferred pH values calculated from diatom stratigraphy relat-
ed very well to the values estimated from using the shells of cladoceran
remains (Norton et al. 1981).
Berge (1976) compared the diatom assemblages in sediments from seven
Norwegian sites with the communities from the same sites as reported in 1949
and found no quantitative change in the diatoms in the 26-year period.
However, he noted a marked shift towards species that required or preferred
low pH. In an even longer period (ca. 1920-1978) Aimer et al. (1974)
reported a reduction in diatoms from cores taken from Scandinavian lakes
which have become more acidic. Dam et al. (1980) reported a more obvious
shift towards acid-tolerant diatoms in sediments from acidic Dutch lakes.
Three hundred years of diatom deposition in sediments was used to calculate
pH values in two Norwegian lakes (Davis and Berge 1980). The pH tolerance of
diatoms was determined from present-day distributions, and the pH in the past
was inferred from the species composition in the dated sediment layers. One
lake has remained constant at ~ pH 5.0 while the other went from pH 5.1 to
4.4 since 1918 (Davis et al. 1983).
More recently (Davis et al. 1983), results of sediment core analyses from
nine Norwegian lakes and six New England lakes were compared (also see
Chapter E-4, Section 4.4.3.2). The range of pH tolerance of the diatoms was
determined by studying current distributions in 36 Norwegian and 31 New
England lakes. The three Norwegian Lakes currently with pH < 5.0 have de-
creased in pH by 0.6 to 0.8 units since 1890-1927. The lakes currently above
pH 5.0 have decreased 0 to 0.3 pH units since 1850. All six of the New
England lakes decreased 0.2 to 0.4 units, but some of these changes might be
due to land use changes (reforestation) which are in the historical record.
Another anomaly was the record of heavy metal pollutants in the sediments
several decades prior to the changes in the diatom communities. This was
ascribed to the buffering of the watershed, which released metals while
retaining protons for many years, thus keeping the lake pH stable, or alter-
natively, to the former high emissions of neutralizing particulates like fly
ash.
An interesting change in the diatom community structure is also apparent from
an analysis of the data (Berge 1976, Dam et al. 1980, Davis and Berge 1980,
Norton et al. 1981, Davis et al. 1983). The species of diatoms which indi-
cate acidic (pH < 5.0) conditions are primarily benthic, whereas those from
circumneutral (pH 6.0 to 7.2) are planktonic. This implies that the diatom
community shifts to benthic production in acidic lakes. Diatoms are common
but not dominant members of the algal mats of present-day acidic lakes
(Stokes 1981).
Del Prete and Schofield (1981) used sediment cores to study the succession of
diatom species in three Adirondack lakes. They observed an increase in
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dominance by acid-tolerant species in the most acid-impacted lakes. A trend
towards species tolerant of low nutrient waters was also reported.
label!an'a fenestrata and Cyclotella stelligera increased in numbers most
directlywfthincreasing acidity, althoughsome of the results were
equivocal.
Coesel et al. (1978) have compared the desmid populations from a group of
lakes in the Netherlands with community compositions reported in studies done
in 1916-25, 1950-55, and with their own survey in 1977. Many of the species
from the rich flora in the earliest survey were lost due to cultural eutro-
phication. In the most recent survey, those ponds that were not impacted by
nutrient additions were affected by acidic deposition, as reflected by the
paucity of desmid species. These ponds appeared to have undergone oligo-
trophication. The eutrophic ponds remained well-buffered and unchanged.
Thus, the effects on community composition brought on by cultural eutrophi-
cation can be separated from the changes caused by acidification.
These studies of temporal trends demonstrate that many acidic lakes have
become acidic in historic times, but they do not prove that this acidifica-
tion is universally a consequence of atmospheric deposition. Deforestation,
followed by eutrophication and reforestation, can cause the pH of a lake to
rise and then fall. Even so, pH's of some lakes have fallen about 0.5 units
in locally unperturbed watersheds in historic times.
5.3.2.2.3 Experimental studies. Muller (1980) studied the succession of
periphyton in artificiallyacidified chambers held in situ in Lake 223,
Experimental Lakes Area, in northwestern Ontario (Schindler et al. 1980b).
At the control pH of 6.25, a succession occurred in the chambers from domi-
nance by diatoms in the spring to dominance by green algae (Chlorophyta) in
mid-July. In enclosures at pH < 6.0, Chlorophyta dominated the periphyton
throughout the sampling period. Blue-green algae (Cyanophyta) were reduced
and almost eliminated under the most acidic conditions. Muller observed no
trend with respect to changes in biomass but noted a sharp decrease in
species diversity (as measured by Hill's index) in the acidified (pH 4.0)
chambers. Changes in primary production (l^C) showed no trend with pH.
The dominance of the periphyton by Chlorophyta in the acidified samples was
due almost entirely to the growth of Mougeotea sp., which by June represented
96 percent of the biomass and cell numbers at pH 4.0. This taxon was re-
sponsible for less than 4 percent of the biomass and cell numbers in the
natural lake water. Interestingly, during May, the blue-green alga, Anabaena
sp., rose from 3.4 percent of the biomass in the lake water (pH 6.2) to 4.3
percent at pH 4.0, but this species was almost absent by June. In spite of
its low biomass this alga accounted for 25 percent and 41 percent of the
total cell numbers in these two samples. Muller's (1980) work demonstrates
the need to consider natural seasonal patterns of succession when we super-
impose the effects of acidification on aquatic ecosystems. The only other
report of seasonal changes in periphyton (Lazarek 1982) dealt with algae
living attached in Lobelia dortmanna and verified the succession from diatoms
to green algae (Mougeotea spp.) during the growing season.
Higher standing crops but lower rates of C-fixation per unit chlorophyll
occurred in periphyton growing in artificial stream channels at reduced pH
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(Hendrey 1976). The total rate of 14C-uptake was similar over a wide range
of [H+]. Increased standing crop was attributed to a combination of three
mechanisms: 1) enhanced growth by acid-tolerant taxa, 2) reduction in graz-
ing by the reduced macroinvertebrate population, and 3) inhibition of micro-
bial decomposition (Hendrey 1976).
In an artificially acidified section of a softwater stream in New Hampshire,
Hall et al. (1980) reported an increase in periphyton numbers and substrate
chlorophyll a^ concentration. They did not perform a taxonomic analysis of
the periphyton community.
Periphyton communities respond to acidification by alterations in species
composition, increases in the standing crop, decreases in the amount of
growth per unit of biomass, and formation of atypical mats which cover the
substrate. These changes produce dramatic, visually obvious changes in lakes
and streams at pH < 5.0.
5.3.2.3 Microinvertebrates—The responses of several minor groups of inver-
tebrates to acidification have been studied. The Nematoda and Gastrotricha
are both common but poorly studied inhabitants of interstitial water in
sediments (meiofauna). They feed on detritus and other organic material
lying between the grains of sand in sediments. The ubiquitous meiobenthic
gastrotrich, Lepidodermella squammata, was almost totally eliminated under
laboratory conditions below pH 6.4 (Faucon and Hummon 1976). Unfortunately,
the pH gradient was achieved by mixing unpolluted creek water with water from
a stream receiving acidic strip mine drainage, so it is not easy to
generalize to streams receiving acidic deposition. Hummon and Hummon (1979)
added CaC03 to the acidic mine drainage and showed that at the same pH,
water with more carbonate (C032~) ameliorated the deleterious effects of
acid stress. The extreme sensitivity of these animals to some component of
the acidic water, possibly low 003^- or high concentrations of metal
ions, bears further investigation. Roundworms (Nematoda) normally have a
ubiquitous distribution (Ferris et al. 1976). However, in an extensive
survey of Norwegian lakes, sub-littoral sediments of acidic lakes had a
scarcity of roundworms when compared to shallow sediments from the same lakes
(Raddum 1976). No other mention is made of the Nematoda in the literature
pertaining to the acidification of aquatic systems.
Freshwater sponges (Porifera) are epifaunal and directly exposed to changes
in water chemistry alterations. However, their response to acidic deposition
has not been studied. Jewell (1939) studied the distribution of Spongillidae
from 63 lakes, bogs and rivers in Wisconsin with various levels of hardness
and pH. She found that most of the species did have limited ranges of Ca2+
concentrations in which they flourished. Six common species were exposed to
chemically modified water, and growth was observed. The lowest pH in this
experiment was 5.9, but there were indications that the most important
parameter was the availability of Ca(HC03)2« As filter feeders, sponges
are important reprocessers of suspended organic matter and are particularly
useful indicators of water quality because of the large volume of water which
passes through their tissues.
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Aquatic mites (Acarina) are not generally collected in surveys of benthic
fauna, but Raddum (1976) noted that rnites occurred in great abundance in the
shallow water of an acid-impacted lake. At a depth of 0.5 m, mites were
third in abundance after nematodes and midges (Chironomidae). At depths > 2
m almost no mites were observed. The shallow mites probably receive their
nutrition from the shore or the water surface, rather than the lake
substrate. In contrast, Wiederholm and Eriksson (1977) observed mites in
deep water (> 10 m) in an acidic lake in Sweden, and Collins et al. (1981)
reported no differences between the distribution of mites in acidic and
control lakes. Clearly, much work needs to be performed on the distribution
of this group to obtain a more complete understanding of how acidic
precipitation affects their distribution.
5.3.2.4 Crustacea—Benthic crustaceans include familiar large forms such as
crayfish (Decapoda), sow bugs (Isopoda), and scuds (Amphipoda), but also
smaller forms such as benthic copepods, mysids, claducerans, and other
branchiopods (e.g., Lepidurus). All these forms, whether large or small,
contribute to the ecosystem dynamics by feeding on detritus or on smaller
detritivores and thus converting the organic material into a form palatable
to fish and other carnivores.
The distribution and characteristics of habitats containing the isopod
Asellus aquaticus (aquatic sow bug) and the amphipod Gammarus lacustris
(scud) were summarized by K. 0kland (1979a, 1980a). Both of these species
are important as food for fishes and as detritus processors. A. aquaticus
populations were reduced below pH 5.2 and absent below pH of 4.8~. While G.
lacustris was able to out-compete A. aquaticus at pH 7.0, Asellus ouT-
competed Gammarus at sites stressed" by either acidic inputs or organic
enrichment. A_. aquaticus was widely distributed in acid-stressed lakes at pH
5.0 (K. 0kland 1980b) but JS. lacustris was inhibited below pH 6.0 (K.
0kland 1980c) probably due to the low calcium concentration in the acidic
water.
In the laboratory, Gammarus pulex demonstrated no avoidance of pH 6.4 to 9.6
(Costa 1967). However, within 12 to 15 minutes after the pH was lowered to
6.2 in one part of the tank, the amphipods began to stay near the alkaline
side. Immature Gammarus performed this avoidance behavior faster than did
adults.
Sutcliffe and Carrick (1973) verified that in England £. pulex is not
normally found below pH 6.0, but they pointed out that it was found in France
at pH 4.5 to 6.0. They suggested that the avoidance response (Costa 1967)
might explain its limitation to near-neutral water, instead of direct
mortality due to low pH. Laboratory studies (Borgstrom and Hendrey 1976)
suggest, however, that direct mortality is important at pH < 5.0. £.
lacustris achieved 96 hr TLso at pH 7.26 in Montana, but populations from
Utah withstood pH 5.7 in similar laboratory bioassays in hard (135 mg £-1
CaCOa in Montana, 200 mg £-1 CaC03 in Utah) water (Gaufin 1973). A
different species, £. fossarum, from Germany, showed no mortality at pH 6.0,
and had a 96 hr TLso of - 4.7. At pH 5.0, 30 percent of the laboratory
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population survived for 10 days (Matthias 1982). K. jOkland (1980a)
ascribed these differences to the variable sensitivity of different
populations.
Steigen and Raddum (1981) noted that A^ aquaticus responded to acidification
by leaving the water, so they confined some of the animals in wire-enclosed
tubes. The confined individuals resorted to cannibalism, but the increased
energetic demands Steigen and Raddum measured caused by the H+ stress
resulted in losses of total caloric value in the confined animals. The
unconfined specimens left the water but returned to feed, sometimes canni-
balistically, and the survivors gained in caloric content. This behavioral
response may be the mechanism by which Asellus can tolerate more acidity than
can Gammarus.
The opossum shrimp, Mysis relicta, is a bottom-dwell ing crustacean character-
istic of deep water. It enters the water column at night to feed on plankton
and, in turn, provides food for fish (Pennak 1978). When Experimental Lake
223 was artificially acidified from pH 6.6 to 5.3, Mysis populations were
eliminated at ~ pH 5.9 (Schindler and Turner 1982).
Eggs of the tadpole shrimp, Lepidurus arcticus (Eubranchiopoda, Notostraca)
took longer to hatch and the larvae matured more slowly than normal at pH <
5.5 than at pH values > 5.5 (Borgstrom and Hendrey 1976). At pH < 4.5,
larvae of L_. arcticus died in two days and eggs never hatched. A survey from
Sweden (Borgstrom et al. 1976) reported that L, arcticus was not found below
pH 6.1.
Laboratory bioassays of the crustaceans Daphm'a middendorffiana, Diaptomus
arcticus, Lepidurus arcticus and Branch! neeta paludosa have provided
additional evidence (Havas and Hutchinson T5BT) of tfie~ sensitivity of
crustaceans to acid stress. Animals collected from an alkaline (pH 8.2) pond
were exposed to naturally acidic water (pH 2.8) from a nearby pond which
received aerial deposition from the Smoking Hills of the Canadian Northwest
Territories. The acidic water was amended with NaOH to provide a range of pH
treatments. A critical pH was 4.5, at which mortality drastically increased
for all of the individuals. Mortality did not occur in control water lacking
heavy metal contamination (Al, Ni, Zn). These authors suggested that their
critical pH of 4.5 was lower than that reported in other studies because the
water in the Smoking Hills area is higher in total conductivity (1.3 mho
cm"1) than that of other acidic clear water systems (Havas and Hutchinson
1982).
An increased abundance of benthic cladocerans has been reported (Collins et
al. 1981) from two of three acidic lakes studied in Ontario.
Crayfish are very important components of the benthos as detrital processors
and as food for larger game fish. Species of crayfish show some variation in
sensitivity to pH. Malley (1980) indicated that Orconectes virilis, in soft-
water of - 22 ymhos cnrl conductivity and Ca*+ of 2.8 mg £-1, was
stressed by pH < 5.5. However, Cambarus sp. was reported (Warner 1971) in a
stream receiving acidic mine drainage at pH 4.6, Ca2+ of 12 mg jr1, and
conductivity of 96 ymhos cm'1. Cambarus bartoni was found in three
5-23
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acidic lakes (pH 4.6 to 4.9, - 3 mg £-1 Ca2+) and Orconectes
propinquis was collected in one of three acidic lakes (Collins et al. 1981) .
"I naVe seen Orconectes spp. in two lakes of pH 4.8 and 5.0 in the
Adirondacks.
This apparent discrepancy in pH tolerances of various crayfish may not be
entirely due to interspecific or inter-population differences. The crayfish
Orconectes virilis has difficulty recalcifying its exoskeleton after molting
at pH < 57!T Uptake of 45ca2+ by crayfish stopped at pH 4.0 and was
inhibited at pH 5.7 (Malley 1980). Infestation of this species by the para-
sitic protozoan Thelahom'a sp. and reduction in recruitment of young at pH
5.7 was also reported (Schindler and Turner 1982). Hence the tolerance of
Cambarus to pH 4.6 from an acidic mine drainage stream may be due to the
higher Ca2+ concentration in the stream compared to habitats affected by
acidic deposition. The ameliorative effect of cations is suggested by the
inability of the crayfish Astacus pallipes to transport 22Na+ below pH
5.5 (Shaw 1960). Stress is a function of both low pH levels and low calcium
levels, and the responses to these stresses undoubtedly vary between life
cycle stages and species.
5.3.2.5 Insecta--The importance of insects in lakes and streams is discussed
in Section 5.2. These animals are important ecologically but also, because
their tolerance to various stresses is well known, they are important as
water quality indicators.
Studies of benthic insects exposed to acid stress include surveys, mostly
from Europe and Canada, and some experimental manipulations. Survey work
involves presence-absence data from which tolerances have been assumed. The
general conclusion drawn from surveys of lakes and streams (Sutcliffe and
Carrick 1973; Conroy et al. 1976; Wright et al. 1975, 1976; Hendrey and
Wright 1976; Leivestad et al. 1976; Wiederholm and Eriksson 1977; Raddum
1979; Friberg et al. 1980; Overrein et al. 1980) is that species richness,
diversity, and biomass are reduced with increasing acidity. Because preda-
tion by fish is eliminated in some waters and food should be abundant due to
the accumulation of detritus (Grahn et al. 1974), one might suppose that
insect biomass would increase. However, acidity imposes stresses that are as
severe as predation (Henrikson et al. 1980b), and the lack of bacterial
decomposition of detritus (Traaen 1976, 1977) may render the detritus
unpalatable to insects (Hendrey 1976, Hendrey et al. 1976).
5.3.2.5.1 Sensitivity of different groups. The sensitivity of benthic
insects to pH stress varies considerably among taxa and among different life
cycle stages (Gaufin 1973, Raddum and Steigen 1981). Responses are
physiological and behavioral.
Mayflies seem to be particularly sensitive to acidic conditions. Female
mayfly adults (Baetis) did not lay eggs on otherwise suitable substrates in
water with pH < 6.0, although three different species were found within 200
to 300 m in neutral brooks with similar substrates (Sutcliffe and Carrick
1973). The adult presumably can detect high levels of acidity by dipping her
abdomen into the water as she flies. Besides Baetis, the common mayflies
Ephemerella igm'ta and Heptagem'a 1 ateral is were absent only from the acidic
5-24
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region of the River Duddon, England (Sutcliffe and Carrick 1973). A Swedish
survey (Nilssen 1980) also found mayflies to be sensitive to pH stress. A
plot of the number of mayfly species vs pH of 35 lakes and 25 rivers indicat-
ed that the number of species decreased logarithmically with decreasing pH.
Species were lost in two groups; one group did not appear below pH 6.5, and
another decline in species numbers occurred below pH 4.5 (Borgstrom et al.
1976, Leivestad et al. 1976). In another survey (Fiance 1978) the distribu-
tional pattern of the mayfly Ephemerella funeral is was studied in the Hubbard
Brook, NH, watershed during a Z-year period.Nymphs were absent from waters
of pH < 5.5. The 2-year life cycle of this mayfly makes it particularly
sensitive to irregular episodic stresses because a single drop in pH may
eliminate the insects for several years. In an experimentally acidified
section of a New Hampshire stream (pH 4.0), mayfly (Epeorus) emergence was
inhibited and drift of nymphs increased (Hall et al. 1980; Hall and Likens
1980a,b; Pratt and Hall 1981). These responses suggest that mayflies exhibit
both behavioral and physiological responses to acidity.
Laboratory bioassays verified that mayflies were the most acid-sensitive
order of insects (Bell and Nebeker 1969, Bell 1971, Harriman and Morrison
1980; Table 5.2). Exposing caged transplanted insects to acidified river
water showed that mayflies could not survive and would try to leave in the
drift (Raddum 1979).
In contrast, dragonflies and damselflies (Odonata) (Table 5-2) are much more
resistant to low pH (Bell and Nebeker 1969, Bell 1971, Borgstrom et al.
1976). The dragonfly nymph, Li be!Tula pulchella. tolerated pH 1.0 for sever-
al hours (Stickney 1922). Dragonfly nymphs (Anisoptera, Odonata) may be able
to endure episodic acidic stress by closing their anus, through which they
respire, but this behavior has not been investigated. Dragonflies burrow
into sand and mud, turning over material and changing the structure of the
habitat. They are also major predators on oligochaete worms, midges
(Chironomidae) , and small insects; they are even known to feed on tadpoles
and small fish (Needham and Lloyd 1916).
Tolerance to acidification within the Plecoptera (stoneflies) is variable
according to surveys (Sutcliffe and Carrick 1973, Leivestad et al. 1976),
field manipulations (Raddum 1979; Hall and Likens 1980a,b), and laboratory
studies (Bell and Nebeker 1969, Bell 1971). Stoneflies and mayflies are
preferred trout food in streams, as evidenced by the attempts of fishermen to
mimic these body forms with their flies (Schweibert 1974). Plecoptera are
ecologically very important components of streams, where smaller forms cling
to rocks, feeding on the drift of detritus, and algae and larger forms seek
smaller invertebrates as prey. Critical sensitivity of this group begins
between pH 4.5 to 5.5, and their distribution generally follows that of
mayflies, except for some tolerant forms like Taem'opteryx. Nemoura,
Nemurella. and Protonemura (Raddum 1979).
Caddisflies (Trichoptera) include burrowers, sprawlers, filter feeders,
predators, detritivores, and forms found specifically in running or standing
water. They occupy many niches and are difficult to lump into
generalizations. Most of the larvae live in cases made from local materials.
Caddisflies have been found in water near pH 4.5 in field surveys (Sutcliffe
5-25
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TABLE 5-2. RESULTS OF LABORATORY STUDIES ON pH TOLERANCE OF SELECTED
INSECT NYMPHS. TL50 IS THE pH WHICH IS LETHAL TO 50% OF THE
ORGANISMS. RESULTS OF DIFFERENT STUDIES ARE REPORTED HERE AS THE
NEGATIVE LOGARITHM OF THE AVERAGE HYDROGEN ION CONCENTRATIONS
Organisms
Ephemeroptera
Baetis sp.
Cinygmula par
Ephemerella doddsi
Ephemerella grandis
Ephemerella subvaria
Heptagenia sp.
Hexagenia 1 imbata
Leptppnlebia sp.
Rhfthrogena" robusta
Stenonema rubrum
Odonata
Boyeria vinosa
Ophiogomphus ripinsulensis
Plecoptera
Acroneuria lycorias
Acroneuria pacifica
Arcynopteryx parallel a
Isogenus aestivalis
Isogenus frontal is
Isoperla fulva
Nemoura cinerea
Pteronarcella badia
Pteronarcys californica
Pteronarcys dorsata
Taemopteryx maura
Trichoptera
Hydropsyche betteni
Hydropsyche sp.
Arctopsyche grandis
Limnephilus ornatus
Brachycentrus americanus
Brachycentrus occidental is
Cheumatopsyche sp.
96 hr
4.5
6.11
4.10
3.6
4.65
6.17
5.66
5.20
4.60
3.32
3.25
3.50
3.32
3.8
4.37
5.08
3.68
4.5
2.6
3.92
4.44
4.25
3.25
3.15
3.28
3.4
2.82
1.50
PH
Long-term 50% successful
TLso (days) emergence References3
5.8(48)
5.38(30) 5.9
5.5(33)b
4.42(30) 5.2
4.30(30) 5.2
3.85(30) 5.0
5.8(90)
4.50(30) 6.6
4.52(90)
4.95(90)
5.00(30) 5.8
3.71(30) 4.0
3.38(30) 4.7
2.45(30) 4.0
4.3(90)
4.52(90)
M
G
G
G
B,
G
G
G
G
B
B,
B,
B,
G
G
G
B,
G
M
G
G
B,
B,
B,
G
G
G
B,
G
G
BN
BN
BN
BN
BN
BN
BN
BN
BN
5-26
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TABLE 5-2. CONTINUED
PH
96 hr Long-term 50% successful
Organisms TLso TLso (days) emergence References3
Diptera
Atherix variegata G
Holorusia sp. 2.8 G
Simulium vittatum 3.63 4.2(68) G
^References: B = Bell 1971, BN = Bell and Nebeker 1969, G = Gaufin 1973,
M = Matthias 1982.
^Seventy of 90 survived.
5-27
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and Carrick 1973, Leivestad et al. 1976, Raddum 1976) but not at pH 4.0
{Raddum 1979; Hall and Likens 198pa,b). Raddum (1979) observed that the
running water caddisflies Rhyacophila nubila, Hydropsyche sp., Polycentropus
flavomaculus, and Plectorenemia conspersa all survived pH 4.0 in the labora-
tory, but only P. conspersa did well in situ at pH 4.8. Raddum explained the
loss of RhyacopTvna and Hydropsyche in the field by alterations in their food
supply. P. flavomaculatus became~canm'balistic at pH 4.0, which may explain
its absence in the stream but its survival when isolated during laboratory
experiments. The problem of cannibalism points out the difficulties in
relating laboratory studies to field observations. Another caddisfly,
Limnepnilus pallens, was collected from an alkaline (pH 8.2) pond and
subjected to more acidic water both in the laboratory and in situ (Havas and
Hutchinson 1982). The larvae survived in pH 3.5 water, and actually did
better in metal-contaminated, sulfate-fumigated water. This acidic water was
near the alkaline pond from which the caddisflies were collected, but no
larvae lived in the acid pond. Possible explanations for the absence of the
caddisflies from water in which they could survive were: 1) absence of
suitable food, 2) sensitivity to the acidity during emergence, 3) absence of
suitable case building material in the acidic pond.
Most other insects are largely unaffected or slightly favored in acidic lakes
and streams. The alderfly, S1alis (Megaloptera) , increased its emergence
rates in an artificially acidified stream (Hall and Likens 1980a,b). It was
found commonly in shallow water in an acidic (pH 3.9 to 4.6) Swedish lake
(Wiederholm and Eriksson 1977) and in a highly variable (pH 6.2 to 4.2)
Norwegian lake (Hagen and Langeland 1973).
Several true flies (Diptera) increase in relative abundance at low pH (Hagen
and Langeland 1973, Wiederholm and Eriksson 1977, Raddum 1979, Collins et al.
1981, Raddum and Saether 1981). The most successful dipterans are the midges
(Chironomidae), the predacious phantom midge (Chaoborus, Chaoboridae) and in
streams, the black fly (Simulidae). Black fly adults are notorious as biting
pests when they emerge in the spring. Often, the principal insects in acidic
lakes are the midges (Chironomidae) Chironomus riparius (Havas and Hutchinson
1982), Procladius sp., Limnochironomous sp., Stichtochironomus sp., Sergentia
coracina, and phantom midges (Chaoborus) (Leivestad et al. 1976, Raddum and
Saether 1981). These insects comprised 56 and 41 percent of the benthos of a
Swedish acidic lake (pH 3.9 to 4.6) (Wiederholm and Eriksson 1977).
Chironomids appear to be preadapted for acidification, because the same
species are found in clearwater acidic lakes as in humic acid lakes (Raddum
and Saether 1981). Uutala (1981) reported that the chironomid fauna of two
acidic Adirondack lakes were reduced in biomass as compared to fauna in
nearby control lakes. The different life cycle stages have variable
responses to pH stress, but the molting period is the most sensitive (Bell
1970).
The dominance of the benthos of acidic lakes by midge larvae is not
surprising, as these insects are abundant in almost all lakes, but the
observed shift in dominant species does suggest that benthic community
structure is altered. Direct toxicity is probably not the explanation for
the absence of certain species. For example, some Orthocladius consobrinus
tolerate pH 2.8 in the laboratory, but this species was not found in acidic
5-28
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pools (pH 2.8) in the Smoking Hills, even though it was found in nearby
alkaline (pH 8.2) pools (Havas and Hutchinson 1982).
Other insects abundant in acidic waters are the true bugs (Hemiptera) such as
water striders (Gerridae), backswimmers (Notonectidae), and water boatmen
(Corixidae) and beetles (Coleoptera) of the families Dytiscidae and Gyrinidae
(Raddum 1976, Raddum et al. 1979, Nilssen 1980). These insects prey on other
insects and small crustaceans, both benthic and planktonic. They are meta-
bolically very active and receive most of their Og from the atmosphere,
thus reducing the amount of soft body tissue exposed directly to the water,
in contrast to gilled insects and crustaceans.
5.3.2.5.2 Sensitivity of insects from different micrphabitats. Important
generalizations are better made by analyzing the data after grouping the taxa
by functional guilds and microhabitats rather than by phylogenetic associa-
tions (Merritt and Cummins 1978). Collins et al. (1981) compared three
acidic softwater lakes (4.6 to 4.9) with 11 neutral softwater lakes in
central Ontario and reported no significant differences in populations of
animals living in sediments (infauna). Observations of epifauna by scuba
divers concurred with the general observation that acidic lakes have
depauperate populations of mollusc and insect.
It is hardly surprising that infaunal communities, which are protected by the
buffering capacity of the substrate, are less affected than epifaunal
communities. Still, few studies have organized data in such a manner as to
verify that epifaunal insects are indeed the targets of acid stress. Also, a
perusal of the data presented above suggests that it is epifaunal forms with
filamentous gills that are most sensitive to low pH. Air-breathing beetles
and bugs survive low pH stress well as do infaunal forms with filamentous
gills, such as the burrowing mayfly, Hexagem'a. Metabolic and physical
actions of Hexagenia nymphs increased the Eh, NH3, inorganic S, S04, and
decreased the pH as compared to control microcosms lacking nymphs or with
dead nymphs (Mitchell et al. 1981b). Thus, not only does the chemistry
affect the biota, but conversely the biota alters the chemistry.
5.3.2.5.3 Acid sensitivity of insects based on food sources. Total inverte-
brate biomass in an acidic (pH 4.3 to 5.9) stream was -2.6 times less than
that of a neutral stream (pH 6.5 to 7.3) 6 km away in southern Sweden
(Friberg et al. 1980). Organizing species lists into guilds based on eating
methods shows that in the acidic water, shredders increased in relative
abundance at the expense of scrapers. These data differ from those reported
by Hall and Likens (1980a,b) from an artificially acidified stream in New
Hampshire, where shredders and predators were not affected. The tolerance of
predators, mostly predacious diving beetles (Dytiscidae) , water striders
(Gerridae), and water boatmen (Corixidae), has been noted in numerous
correlative surveys (Leivestad et al. 1976, Raddum et al. 1979). Shifts in
the activities of these different functional guilds affect detritus
processing and may be either a cause or a result of the inhibition of
microbial detritus processing (Section 5.3.2.1).
5.3.2.5.4 Mechanisms of effects and trophic interactions. It is likely that
factors other than HT concentration stress organisms in acidified waters
5-29
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(Overrein et al. 1980). Malley's (1980) work (see Section 5.3.2.4) suggests
that reduced calcium deposition may limit insects as well as crustaceans.
Havas (1981) suggested that Na+ transport may be affected. Effects of
increased Al concentrations on invertebrates have not been studied as
intensively as they have with fish (Baker and Schofield 1980). Other metals,
such as Hg (Tomlinson 1978) may also be important. Nutrient depletion,
inefficient microbial digestion, substrate alteration, dissolved oxygen
stress, and changes in other populations (e.g., fish predation) all may act
on insect populations. The water boatman, Glaenocorisa propinqua propinqua,
a predator on zooplankton and other smallinvertebrates,is tolerant of
acidity and is common in acidic lakes. The addition of perch to one-half of
a lake divided by a net vastly reduced numbers of Glaenocorisa on the side
with fish. The only change in water chemistry was a decrease in total
phosphorus from 3-8 to 2-6 ug £~1 when fish were added (Henrikson and
Oscarson 1978). Different taxa respond in various ways. Some may make
behavioral adaptations; others, like the water boatmen (Corixidae), can alter
rates of Na+ pumping (Vangenechten et al. 1979, Vangenechten and
Vanderborght 1980).
For reasons which are not clear, a shift towards larger species within a
higher taxon occurs (Raddum 1980). This may be due to reduced predation
pressure on larger insects in the absence of fish or because larger species
have less surface/volume and can cope better with chemical and osmotic
stress. Increased abundance of insect predators may be due to the opening of
this niche as a result of fish loss (Henriksen et al. 1980b) or due to the
larger size of these predators. Community alterations, and even modifica-
tions of water column chemistry, have been traced to fish removal (Stenson et
al. 1978) independent of pH changes. Thus, it is dangerously simplistic to
ascribe changes in community composition to merely the physical-chemical
alterations of acidification without also considering the varied biological
interactions.
Alterations in insect populations are likely to affect fish populations
(0kland and 0kland 1980). Rosseland et al. (1980) reported that corixids
composed 15 percent of the gut content, by volume, of trout from neutral
waters but 44 percent in a declining population from an acidic (pH < 5.5)
stream. However, no causal relationship between shifts in diet and
population decline can be made at this time.
5.3.2.6 Mol1usca--Mol1 uses provide food for vertebrates (fish, ducks,
muskrats, etc.).Clams are filter feeders and are important bioindicators of
water quality conditions. Snails scrape the substrate and the surfaces of
aquatic plants, controlling the periphyton in waters in which they live. The
impact of acidity on molluscan populations is dramatic. The calcareous shell
of these animals is highly soluble at pH < 7.0 and acidic conditions require
that the animals precipitate fresh CaC03 faster than it can dissolve.
The only thorough survey of clams and snails in acid-impacted waters was done
in Norway (J. 0kland 1969a,b, 1976, 1979, 1980; K. 0kland 1971, 1979b,c,
1980b; 0kland and 0kland 1978, 1980; 0kland and Kuiper 1980). About
1500 localities, mostly lakes in Norway, were surveyed between 1953 and 1973.
Fingernail clams (Sphaeriidae) and snails (Gastropoda) were sampled.
5-30
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Sphaeriidae live in sediments (infaunal) and no surveys of the more epifaunal
unionid mussels have been conducted. Norway has 17 species of Pisidium and
three of Sphaerium. None of these clams normally occurred below pH 5.0. The
six most common sphaeriids were eliminated below pH 6.0. These common
species were found in lakes with low alkalinities but with pH values -6.0.
Thus, their absence from these poorly buffered lakes serves as an indication
of acidification, not just low CaC03 stress (0kland and Kuiper 1980).
Freshwater snails (Gastropoda) were reported to be stressed much like the
clams from the Norwegian survey. Of the 27 species of snails reported in
Norway, only five were found below pH 6.0 (J. 0kland 1980). Snails could
tolerate higher H+ concentrations if the total hardness were higher, indi-
cating that pH may stress snails by reducing the CaC03 availability. The
authors (0kland and 0kland 1980) estimated that the crustacean Gammarus
lacustris and the molluscs accounted for 45 percent of the caloric input of
trout, and they predicted that trout production could be reduced by 10 to 30
percent below pH 6.0 due to the loss of food resources. This prediction has
not been supported by the fish surveys (Section 5.6.2.3).
Some additional distributional data, which corroborate the 0klands'
conclusions cited above, have been reported from Sweden (Wiederholm and
Eriksson 1977), Norway (Hagen and Langeland 1973, Nilssen 1980), and from a
river in England (Sutcliffe and Carrick 1973). These later authors emphasiz-
ed the absence of the freshwater limpet snail Ancylus fluviatilis as an
indicator of pH levels which frequently fall below 5.7. They also concluded
that pH served to limit the distribution of molluscs by reducing the
availability of CaC03, as measured by water hardness.
The physiological response of molluscs to pH stress was studied by Singer
(1981b). In Anodonta grandis (Unionidae) from six lakes in New York and
Ontario with various levels of pH and hardness, marked differences in shell
morphometry and ultrastructure were observed. The clams from alkaline lakes
(pH > 7.2) had thick shells with fine layers of organic conchiolin
interspersed. The clams from softwater neutral lakes had thinner shells,
with relatively thick prismatic layers. Clams from a slightly acidic lake
(pH 6.6) had thin shells with heavy plates of organic material substituting
for the normal CaCOs matrix. Using unionid shells from museum collections
as indicators of pre-acidification water quality was suggested.
5.3.2.7 Annelida—Aquatic worms have been used extensively as indicators of
organic (Goodnight 1973, Brinkhurst 1974) and inorganic (Hart and Fuller
1974) pollution. With an increase in organic detritus and a decrease in 02
concentrations, the benthic community is typically dominated by Tubifex spp.
and Limnodrilus hoffmeisteri (Brinkhurst 1965, Howmiller 1977). Considering
their tolerance of otherstresses and the abundance of detritus, it is
surprising that oligochaetes are reduced in biomass in acidic lakes. Raddum
(1976, 1980) found few oligochaetes in water deeper than 20 m in 18 acidic
lakes (pH < 5.5) and normal fauna in 16 other more neutral Norwegian oligo-
trophic lakes. The acidic lakes, however, had more oligochaetes than the
non-acidic lake at a depth of 0.5 m. The difference in numbers at greater
depths was more pronounced in the spring and autumn. Neutral lakes had three
to four times the total number of oligochaetes per square meter. Raddum
5-31
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(1980) attributed the reduction in numbers of oligochaetes in acidic lakes to
pollutants associated with acidic deposition (e.g., heavy metals and
aluminum). These worms, however, are routinely collected in vast numbers
directly below sewage and industrial effluents with far greater concentra-
tions of pollutants (Hart and Fuller 1974, Chapman et al. 1980). An alterna-
tive explanation for their reduction in numbers might be the unpalatability
of their detrital food due to the slower decomposition rates in acidic lakes
(Traaen 1977). Oligochaetes are not normally abundant in nutrient-poor
waters, and their low numbers in acidic lakes may be as much a function of
the low nutrient level characteristic of acidic lakes as of pH.
One study that mentioned the distribution of leeches (Hirudinea) in acidic
lakes (Nilssen 1980) reported that these worms disappeared below pH 5.5.
Leeches characteristic of eutrophic waters (Hirudo medicinal is, Glossiphonia
heteroclita) were absent from even mildly acidiclakes.Raddum(i960)
reported that Hirudinea were restricted to waters above pH 5.5, largely
because of the loss of prey below this pH, even though many leeches are
detritivores and scavengers, not obligate carnivores (Pennak 1978). These
anecdotal observations should be viewed with caution because leeches are not
always common in neutral oligotrophic lakes, and I saw an unidentified leech
on the bottom of acidic Woods Lake (Herkimer Co., NY) while diving in 6 m of
water.
5.3.2.8 Summary of Effects of Acidification on Benthos—Table 5-3 summarizes
some of the expected consequences of acidifying a lake or stream to pH 4.5.
The following generalizations may also be made, based on the best available
current evidence.
1. Bacterial decomposition of litter in bags in situ and debris in vitro is
reduced significantly (p < .001), as measured by respiratory rates and
weight loss, between pH 6.0 and 4.0. Planktonic bacterial standing
crops do not change significantly, although metabolic rates are
depressed. Insects and crustaceans responsible for shredding and
processing detritus are almost completely eliminated between pH 6.0 and
4.0.
2. In most acidified lakes below pH 5.0, a mat of algae covers most of the
substrate from ~ 1 to 5 m to the limit of light penetration. These
mats are of 3 types: a) an encrusting, felt-like, black to blue-green
mat composed of blue-green algae (Cyanophyta) 0.5 to 2 cm thick; b)
coarse, loosely compacted dark green mats composed of green algae
(Chlorophyta) 1 to 4 cm thick; c) cloud-like layers of green filamentous
algae (Chlorophyta) which rest on the bottom in depths as thick as 1.5
m. All three types of mats include debris, diatoms, fungi and bacteria.
These mats are often the most visible aspect of acidified lakes. They
may have profound effects on fish spawning habitats, nutrient cycling,
and sediment chemistry, but their origin, differentiation into types,
and chemical interactions have not been studied. They have been exten-
sively noted in field surveys and have developed in artificially acidi-
acidified chambers and stream channels below pH 5.0.
5-32
-------
TABLE 5-3. SUMMARY OF THE EFFECTS OF pH 4.5 WATER ON BENTHOS
Taxon
Common name
Microhabltats
Sensitivity to acid
(pH 4.5) stress
References
Bacteria
Perlphyton
Algae
Blue-greens (Cyanophyta)
Greens (Chlorophyta)
Diatoms (Bacillariophyceae)
Dlnoflagellates (Pyrrophyta)
en
co
co
Crustacea
Decapoda
Crayfish
Isopoda
Anphlpoda
Aquatic sowbug
Scud
All substrates.
On plants (epiphytic)
On rocks (epipllthlc)
On mud (eplpellc)
On "encrusting mat".
Burrowers, deposit
feeders and grazers
In lakes and streams.
Deposit feeder In
lakes and streams,
under rocks and in
littoral vegetation.
Detritivore In lakes
and slow-flowing
areas of streams.
Found among plant
stems.
Growth rate or 02
uptake Inhibited.
Increasing standing
crop in lakes and
streams.
Development of dis-
tinct types of perl-
phyton communities.
Sensitivity variable
and highly species
specific. Effects
vary depending on
other cation concen-
trations.
Asellus aquatlcus
tolerant to ~ pH
5.0.
Not generally found
below pH 6.0. Sen-
sitivity differences
between species have
been described.
Blck and Drews
1973, Baath et al.
1979, Gahnstrom et
al. 1980
Hendrey 1976, Hall
et al. 1980, Yan
and Stokes 1978,
Hendrey and
Vertucci 1980,
Muller 1980
Singer et al.
1983, Stokes 1981
Mai ley 1980,
Collins et al.
1981, Shaw 1960
K. 0kland 1979a,
1980b.
K. 0kland
1980a,b,c;
Sutcliffe and
Carrlck 1973;
-------
TABLE 5-3. CONTINUED
Taxon
Common name
Mlcrohabitats
Sensitivity to add
(pH 4.5) stress
References
Eubranchlopoda Tadpole shrimp
Smal 1, often tempo-
rary, ponds or back-
waters of streams;
often only abundant
seasonally.
Not found in Sweden
pH 6.1. Growth re-
duction and hatching
failure below pH
5.0.
Borgstrom et al.
1976, Borgtrom and
Hendrey 1976
Insecta
Ephemeroptera Mayflies
en
co
Odonata DragonfUes (Anlsoptera)
Damsel flies (Zygoptera)
Plecoptera Stoneflies
Trlchoptera Caddlsflies
Include burrowing and
surface dwelling
forms. Found In
lakes and streams.
Predators, detriti-
vores, herbivores.
Predators In mud,
littoral debris, and
rock substrates in
lakes and streams.
Predators, detriti-
vores and herbivores
in flowing streams.
All benthic habitats.
Sensitivity varies
between groups but
generally not
tol erant
Tolerant to pro-
longed severe acid
stress.
Most genera are
sensitive but some
are tolerant (see
text).
Some genera are very
tolerant, but others
are sensitive (see
text)
Sutcliffe and
Carrlck 1973,
Nllssen 1980,
Pratt and Hall
1981, Leivestad
et al. 1976,
Borgstrom et al.
1976. Raddum 1976
Stlckney 1922,
Borgstrom et al.
1976, Bell and
Nebeker 1969, Bell
1971
Sutcliffe and
Carrlck 1973;
Leivestad et al.
1976; Hall and
Likens 1980a,b;
Raddum 1979
Sutcliffe and
Carrlck 1973;
Leivestad et al.
1976; Hall and
Likens 1980a,b;
Raddum 1976, 1979
-------
TABLE 5-3. CONTINUED
Taxon
Common name
Mlcrohabltats
Sensitivity to acid
(pH 4.5) stress
References
Olptera
en
i
CO
en
Hemlptera
Coleoptera
Nollusca
Pelecypoda
True flies
Midges (Chironomidae)
Phantom ghost midge
(Chaobotidae)
Black flies (Simulidae)
True bugs
Water striders (Gerridae)
Backswimmer (Notonectidae)
Water boatman (Corixidae)
Beetles
Predacious diving beetle
(Dytiscidae)
Whirligig beetle (GyHnidae)
Clams
Major detritlvores in
lakes and enriched
streams, living in
mud or on substrates
In tubes.
Predator on sub-
strates and 1n water
column of lakes
Predatory in streams
on rock substrates
Predators on water
surface, in water
column, and over sub-
strates.
Predators on water
surface, in water
column, and over sub-
strates.
Filter feeders in
substrates, detriti-
vores in lakes and
streams.
Most reports show
increase in numbers,
but some report
decreases.
Tolerant of acid
stress.
Tolerant of acid
stress
Tolerant of acid
stress.
Tolerant of acid
stress.
All mollusca are
highly sensitive to
pH stress. The most
tolerant are finger-
nail clams which are
rarely found as low
as pH 5.0.
Raddum and Saether
1981, Uutala 1981
Wiederholm and
Eriksson 1977,
Leivestad et al.
1976
Leivestad et al.
1976
Raddum 1976,
Raddum et al.
1979, Nllssen
1980
Raddum 1976,
Raddum et al.
1979, Nllssen 1980
J. Dkland 1976,
1980; K. 0kland
1971, 1979b,c,
1980b; flkland
and flkland
1978, 1980;
flkland and
Kulper 1980;
Singer 1981b
-------
en
i
oo
cn
TABLE 5-3. CONTINUED
Taxon
Common name
Mlcrohabitats
Sensitivity to acid
(pH 4.5) stress
References
Annelida
Oligochaeta Aquatic earthworms
Detritivores in lakes Standing crops low
and streams with soft In acidic waters
substrates.
Raddum 1976, 1980
Hlrudlnea Leeches
Predators, detrlti-
vores.
Anecdotal observa-
tions report no
leeches below pH
5.5.
Nllssen 1980,
Raddum 1980
-------
3. Many invertebrates are very sensitive to pH. Amphipods, which are an
important fish food in rivers and some lakes, cannot tolerate pH < 6.0,
based on field observations, laboratory bioassays, and field enclosure
experiments. Snail populations are stressed below pH 6.0 and absent
from the field below pH 5.2. Large mussels cannot survive below pH 6.6,
but fingernail clams can survive in sediments with overlying water with
pH values as low as 4.8. The crustacean water louse (Isopoda) and many
species of stoneflies (Plecoptera), mayflies (Ephemeroptera), and
caddisflies (Trichoptera) die at pH < 5.0, as determined by field obser-
vations and laboratory bioassays. Insects are often limited by mecha-
nisms not related to direct toxicity. Some dragonflies and many pre-
dacious beetles (Coleoptera), and true bugs (Hemiptera) occur commonly
in acidified (pH < 5.0) lakes. They fill the niche normally occupied by
planktivorous fish and represent a major alteration of food chains.
Most of these active predacious insects receive their air supply from
the surface.
4. Forms which live CM the substrate (snails, stoneflies, mussels, etc.)
are more sensitive to pH drops than those which live in the substrate
(e.g., fingernail clams, midge larvae, burrowing mayfTTes). In those
groups that have been studied in the laboratory (crayfish, backswirraners,
molluscs), high calcium concentrations (> 2 mg jr1) can ameliorate
the effects of low pH.
Fish shift their food to available prey, but the nutritional effects of
switching from a diet of largely amphipods, mayflies, and stoneflies to one
of water boatmen, beetles, and water striders are not known. Effects on
different age classes of fish are likely to vary. Changes in the rates of
detrital processing and decomposition rates affect primary productivity and
hence the whole ecosystem.
5.4 MACROPHYTES AND WETLAND PLANTS (J. H. Peverly)
5.4.1 Introduction
The softwater, low alkalinity, oligotrophic lakes in temperate regions
susceptible to acidic deposition support a flora characterized by the isoetid
or rosette plants. This contrasts with hardwaters which support vittate
species, having elongated stems with leaf nodes. Plants commonly observed
in softwater lakes are listed in Table 5-4.
In general, emergent plants in these lakes grow only in a narrow band along
the shore. The submerged, three-inch high isoetids extend from shore to the
3 to 4 m depth and coexist with some lilies and bladderwort. Beyond 4 m,
Nitella spp., bladderwort, and mosses dominate.
Life in the water depends on the presence and growth of aquatic plants as
well as other inputs from the basin (Section 5.3.1). Macrophytes stabilize
the sediments; clear, cool and oxygenate the water; and provide colonization
sites for insects, small plants and animals, and bacteria. These organisms
in turn are a major food source for the larger aquatic animals, such as
5-37
-------
TABLE 5-4. PLANTS COMMONLY OBSERVED IN SOFTWATER (LOW ALKALINITY)
OLIGOTROPHIC LAKES
Species
Common name
Type
Spargam'um spp.
Brasenia
schreberi Gmel.
Nuphar
advena Ait.
Nymphaea
odorat'a Ait.
Isoetes spp.
Lobelia
dortmanna L.
Eriocaulon
septangu'lare With
Myriophyllum
TTi
tenel1 urn BTgel
Potatnogeton spp.
Burreed
Water
shield
Yellow
lily
White
lily
Quillwort
Pipewort
Pond weeds
Emergent
Floating leaves,
rooted
Floating leaves,
rooted
Floating leaves,
rooted
Submerged,
rooted
(iosetids)
Submerged,
rooted
(iosetids)
Submerged,
rooted
(iosetids)
Submerged,
rooted
(iosetids)
Submerged,
rooted
Response to
acidification
Pontederia
cordata L.
Juncus sp.
Pickerel
weed
Rush
Emergent
Emergent
Unknown
Stimulated
growth
(Hultberg and
Grahn 1976)
Unknown
Unknown
Unknown
Unknown
Overgrown
(Hultberg and
Grahn 1976)
Overgrown
(Hultberg and
Grahn 1976)
Oxygen evolu-
tion falls
(Laake 1976)
Unknown
Unknown
Decreased
growth
(Roberts et
al. 1982)
5-38
-------
TABLE 5-4. (CONTINUED)
Species
Common name
Type
Response to
acidification
Eleocharis spp.
Utricularia spp.
Sphagnum spp.
Drepanocladus spp.
Fontinail's spp.
Nitella spp.
Spike rush Submerged,
rooted
Bladderwort Submerged,
unrooted
Moss
Moss
Moss
Stonewort
Submerged,
attached
Submerged,
attached
Submerged,
attached
Submerged,
attached
Unknown
Unknown
Stimulated
growth (Grahn
1977)
Unknown
Unknown
Unknown
5-39
-------
fishes, amphibians, aquatic mammals, and waterfowl. Thus, aquatic plants
fill an important role in the entire aquatic ecosystem.
Macrophyte growth in softwater lakes can be a major part of total lake
production and is largely attributable to growth by isoetids (Hutchinson
1975, Hendrey et al. 1980b). Because isoetids are perennial and evergreen,
they can continue to photosynthesize and produce oxygen under winter ice
cover, and provide a stable, constant source of grazing material. Standing
crop varies from < 5 to 500 g dry wt m~2 in August, but annual productivity
is only about 50 percent of standing crops (Moeller 1978, Sand-Jensen and
Sondergaard 1979).
Plant productivity in softwater lakes is not high because the carbon dioxide
(003) level in the water is low (0.02 mM C02 at pH 5.0) and major nutri-
ent minerals such as P, K, N, and Ca are in limited supply (Hutchinson 1975).
However, these aquatic macrophytes have several means of overcoming such
difficulties and producing enough tissue to support an aquatic animal
community. First, aquatic macrophytes recapture up to 50 percent of their
own respiratory COg and store it in an internal gas chamber system for
reuse in photosynthesis (Sondergaard 1979). Secondly, the isoetids are able
to exist and grow in oligotrophic water, where other aquatic macrophytes
cannot, by more efficient use of nutrients in the sediments. This is ac-
complished by the root systems, which are efficient sites for absorption of
carbon, nitrogen, phosphorus, and potassium. The relative root-to-shoot
ratio is large in these plants (0.5 to 0.6, Sondergaard and Sand-Jensen
1979), indicative of the greater role of roots in nutrient absorption. In
addition, water in the sediments where the roots grow often has a carbon
level of 1 to 5 mM (Wium-Andersen and Andersen 1972), 50 to 100 times that in
the overlying water column. Vittate plants, which depend more on leaf
absorption for carbon supply, cannot grow in these low carbon waters.
The accumulation of nutrients in plant tissues, acquired through the roots
from the sediments, recirculates sediment nutrients back into the overlying
water, where they can be used by other organisms. For instance, in a 200 g
dry wt m~2 crop of Eriocaulon septangulare, there would be about 50 g
carbon, 2 g nitrogen, 0.1 g phosphorus, and 1.5 g potassium. About 0.24 g
carbon, 0.2 g nitrogen, 0.01 g phosphorus, and 0.4 g potassium (Moeller 1975)
would be dissolved in water 1 m deep over this meter square area. Clearly,
nutrient release from such plant beds could increase the concentration of
available nutrients in the water column.
Lilies and emergent plants can also obtain carbon by absorption of C02 from
the atmosphere and translocation to carbon reserves in rhizomes under the
water surface. Mosses and algae are not as involved in processes that
transfer nutrients from sediments or air into the water column.
In addition to major nutrients, rooted aquatic macrophytes (including
isoetids) are exposed to elevated levels of metals in the sediments (e.g.,
iron, manganese, copper, zinc, aluminum). These elements can also be
absorbed by roots and transported to the shoots, where they are able to enter
biological cycles slowly as the plants senesce and decay. However, concen-
tration differences between sediment and water column levels of metals
5-40
-------
available for absorption are not always as great as for the major nutrients.
This is especially the case where rooted plant activity is high, as oxygen
release at the root surfaces (Wiurn-Andersen and Andersen 1972) raises the
redox potential. This tends to precipitate iron and manganese compounds
(Tessenow and Baynes 1978) and remove phosphorus from solution. Metals not
affected by redox potential, like aluminum, would remain in solution in the
rhizosphere, and still be available for uptake by the roots. Indeed, the
aluminum contents of plant tissues (0.4 to 22 g kg"1) from both neutral and
acidified lakes (Al 0.03 to 0.2 mg £-1) in the Adirondacks and Ontario
were elevated above Hutchinson's (1975) mean value of 0.36 g kg~* (Best and
Peverly 1981, Miller et al. 1982).
Lilies interact much more with sediments than with water and generally tend
to accumulate less of the above metals. The mosses and algae interact almost
exclusivey with the water column and accumulate metals (Ca, K, Fe, Al) under
certain water conditions. Aquatic macrophytes can recycle Fe, Mn, Cu, Zn,
and Al metals from sediments, but they can also restrict exchange of Fe and
Mn between water and sediment by oxidizing the top 15 to 20 cm of sediments
(Tessenow and Baynes 1978).
Mosses and algae that grow close to the bottom not only absorb metals meta-
bolically, but also physically adsorb them onto tissue surfaces. Sphagnum
spp. are known to have especially high adsorption capacities for metals,
including calcium, iron, aluminum, and potassium (Clymo 1963, Hendrey and
Vertucci 1980). Metals adsorbed in this fashion are effectively removed from
biological cycles for long periods, as the elements remain bound to dead
tissues, which often persist for years. Mats of Sphagnum spp. and algae have
formed on the bottom of some softwater lakes. Hultberg and Grahn (1976)
suggested that mats of this nature decrease productivity by restricting
exchange of nutrients between sediments and water.
The tissues produced by growing plants eventually die, releasing nutrient
elements and metals back to the water by a variety of decay processes.
Carbon dioxide is produced by plankton and microorganisms from this dead
plant material, along with dissolved phosphorus, potassium, ammonia and
calcium. The metals are released, often in a form complexed with organic
acids that keeps them in solution, thus readily available for uptake.
5.4.2 Effects of Acidification on Aquatic Macrophytes
Direct effects of acidification on aquatic macrophytes have not been well-
documented. However, in two reports of laboratory results, oxygen evolution
was reduced up to 75 percent by a pH decrease from 7.0 to 4.0 in both
softwater (Laake 1976) and hardwater plants (Roberts et al. 1982). In the
field, nutrient ions and metals (such as calcium, magnesium, sodium,
potassium, manganese, and iron) may be leached out of the tissues, especially
during the episodic pH drops associated with snowmelt. This could have a
negative effect on plants in the spring when new growth is quite susceptible
to nutrient imbalances.
Most effects of acidification on aquatic plant distribution and growth are
indirect. Specifically, these would include decreased carbon supply for
5-41
-------
photosynthesis, nutrient depletion, increased metal concentrations, and
decreased rates of nutrient recycling (Grahn et al. 1974, Andersson et al.
19785, Schindler et al. 1980a). The dominance of isoetid species in soft-
water lakes of pH 5.5 to 6.5 is a response in part to low carbon and major
nutrient availability in the water column. As acidic deposition causes the
pH to decline, these factors become even more limiting. For instance,
Lobe!la dortmanna rooted in sediment cores showed a 75 percent reduction in
oxygen production at pH 4.0 compared to the control (pH 4.3 to 5.5), and the
period of flowering was delayed 10 days at the low pH (Laake 1976). As a
result, species more tolerant of low nutrient supplies and higher metal
concentrations may become dominant.
Measurements over 15 years in one acidified Swedish lake with a pH drop of
0.8 units between 1967 and 1973 showed that isoetid species were replaced by
Sphagnum sp. and blue-green filamentous algae, which grew over the bottom in
that time span, smothering the low-growing isoetids (Grahn 1977). This is
viewed as detrimental to overall lake quality because Sphagnum beds are not a
good habitat for most aquatic animals. In addition, Sphagnum tends to
perpetuate the conditions that exclude other speciesbyexchanging
metabolically-produced hydrogen ions for nutrients and metals in the water
via adsorption processes. Thus, acidification and oligotrophication
continue. As the Sphagnum grows, it forms a mat of increasing area. The
dead stems decay slowly and continue to hold adsorbed elements. As a result
of this mat barrier and because Sphagnum has no roots to exploit the
sediment, interchange of dissolved nutrients between overlying water and
sediments is minimized following Sphagnum invasion. With the exception of
dense Sphagnum beds observed in Lake Golden (pH 4.9) in the Adirondack
Mountains of New York State (Hendrey and Vertucci 1980), large expanding beds
have not been observed in acidified waters of the northeast United States or
Canada (Best and Peverly 1981, Wile 1981).
The effect of acidification on nutrient availability is unclear. Generally,
slower breakdown of organic matter (including Sphagnum tissues) in acidic
waters (see Section 5.3.2.1) would tend to decrease the amount of major
nutrients available for plant growth. In addition, softwater lakes are
inherently low in nutrients. In the Adirondacks, plant tissue concentrations
of the major nutrients indicated that phosphorus was limiting in both acidic
and non-acidic lakes (Best and Peverly 1981).
Other possible indirect effects of acidity on macrophytes are those
associated with increased metal (aluminum, cadmium, iron, manganese, copper,
lead, zinc) concentrations in water and sediments. Tissue analysis of
isoetid plants from both acidified and non-acidified lakes in the Adirondacks
and Ontario have shown elevated levels of aluminum, copper, iron, and lead in
roots and shoots from acidic waters (Best and Peverly 1981, Miller et al.
1983). Concentrations of manganese, cadmium, and zinc were lower in plants
from acidic waters, corresponding to one report of lower measured metal
levels in sediment of an acidified lake (Troutman and Peters 1982).
Toxic tissue levels of metals discussed above are not presently known.
Effects of increased metal accumulation on isoetid productivity are not
clear, but these metals have been shown to be toxic to aquatic plants.
5-42
-------
Concentrations of Al, Zn, and Cu in sediments measured by Stanley (1974)
produced 50 percent reduction in Myriophyllum spicatum root weight. However,
these concentrations were greater than those reported to occur in acidified
lake sediments, at least for Adirondack lakes (Best and Peverly 1981).
If metal concentrations increase in tissues, but do not inhibit growth, there
is a potential for increased cycling of metals. However, Sphagnum spp.
qrowth may be a positive factor, removing metals from the water Dy aasorption
(clymo 1963) and by barrier formation between the sediments and water.
Acidification of brown waters that contain organic acids causes clearing of
the water column by organic precipitation with metals especially aluminum
(Aimer et al. 1978) (Chapter E-4, Section 4.6.3.4). The result is increased
light penetration to greater depths, with plant growth perhaps increased over
a larger area. This could lead to a larger food base for aquatic animals and
could be a positive factor if the increased growth is not represented solely
by Sphagnum spp. and blue-green algae.
5.4.3 Summary
0 There is currently no trend towards dominance of macrophyte communi-
ties by Sphagnum sp. in 50 oligotrophic, softwater lakes surveyed in
North America!In fact, dominant species are the same in both acidi-
fied (pH less than 5.6) and non-acidified (pH 5.6 to 7.5) lakes.
o With continued acidification, shifts to Sphagnum spp.-dominated
macrophyte communities have been documented in six Swedish lakes
acidified for at least 15 years. This does not seem to be a general
property of acidified lakes.
0 Standing crops of macrophytes vary widely (5 to 500 g dry wt m~2)
in softwater, oligotrophic lakes and acidification produces no
definite trend in standing crop changes. Based on one report, annual
productivity is equal to one-half the summer standing crop in a
non-acidified lake. Oxygen production was reduced 75 percent at pH
4.0 versus pH 4.3 to 5.5 in one flow-through experiment.
o The only known effect of acidification on macrophytes in the field is
that of increased metal content in the tissues, especially Al. In
acidified lakes, mean aluminum concentration in plant tissue (dry wt
basis) is 3.0 to 5.0 g kg-1 (about ten times higher than normal)
while mean manganese concentration is 0.02 to 4.0 g kg-1 (about
one-fifth of normal). In general, concentrations of iron, lead and
copper are higher, while cadirnium and zinc are lower in the tissues
of plants from acidified lakes.
5-43
-------
5.5 PLANKTON (J. P. Baker)
5.5.1 Introduction
The term plankton refers to organisms that live suspended within the water
column, are generally small to microscopic in size, have limited or no powers
of locomotion, and are more or less subject to distribution by water
movements (Wetzel 1975). The plankton community consists of animals
(zooplankton), plants (phytoplankton), and microbes. Effects of acidifica-
tion on zooplankton and phytoplankton will be considered within this section;
effects of acidification on the microbial community were included in Section
5.3. Discussions focus on plankton communities within the open-water zone.
Interactions with populations in littoral and benthic regions are important,
but poorly understood with regard to potential effects of acidification.
Zooplankton and phytoplankton communities are usually quite complex, composed
of a large number of species, and subject to significant spatial and temporal
variations. These variations in occurrence and importance of species of
phytoplankton and zooplankton make it difficult to obtain a representative
sampling of the plankton community. Attempts at relating differences in
plankton communities between lakes or within a given lake to acidity or other
environmental parameters are hindered by this natural diversity and
variability.
Six phyla of algae typically contribute to phytoplankton communities of
freshwater ecosystems: Cyanophyta (blue-green algae), Chlorophyta (green
algae), Pyrrophyta (primarily dinoflagellates), Chrysophyta (yellow-green
algae; includes the chrysomonads and diatoms), Euglenophyta (euglenoids), and
Cryptophyta (primarily cryptomonads). Photosynthesis by phytoplankton plays
a significant role in the metabolism of lakes (Schindler et al. 1971, Jordan
and Likens 1975, Wetzel 1975), and in determining the quantity of secondary
or tertiary (e.g., fish) production within a lake (Smith and Swingle 1939,
Hall et al. 1970, Makarewicz and Likens 1979).
The animal components of freshwater plankton communities also constitute a
diverse collection of organisms from many phyla. The most important taxo-
nomic groups are protists (Phylum Protozoa), rotifers (Phylum Aschelminthes,
Class Rotifera, or as a separate Phylum Rotifera), insects (Phylum
Arthropoda, Class Insecta), and two subgroups of the Class Crustacea (Phylum
Arthropoda), the Subclass Copepoda and the Order Cladocera (Subclass
Branchiopoda) (Edmondson 1959). A large number of trophic levels are also
represented—herbivores, omnivores, and carnivores. Thus, both the structure
(variety in types of organisms represented) and function (energetic inter-
actions among individual organisms) of the plankton community are complex.
Data on acidification and effects on plankton communities are limited almost
entirely to field observations and correlations. Experiments designed to
elucidate causal mechanisms for observed changes are, for the most part,
lacking, at both the physiological and ecological level. The large number of
interacting factors potentially involved in the reaction of plankton to
acidification makes a critical analysis of currently available data very
5-44
-------
difficult. In some cases, results appear contradictory. With an increased
understanding of causal mechanisms, many of these apparent contradictions
should be resolved.
5.5.2 Effects of Acidification on Phytoplankton
5.5.2.1 Changes in Species Composition--In extensive surveys of acidic lakes
in Norway, Sweden, eastern Canada, and the United States, altered species
composition and reduced species richness (number of species) in the
phytoplankton community were consistently correlated with low pH levels.
Results from 18 field studies that support this conclusion are summarized in
Table 5-5. Decreases in species richness appear most rapidly in the pH
interval 5.0 to 6.0 (Aimer et al. 1974, 1978; Leivestad et al. 1976;
Kwiatkowski and Roff 1976). For example, in a survey of lakes in the west
coast region of Sweden, lakes with pH values of 6.0 to 8.0 generally
contained 30 to 80 species of phytoplankton per 100 ml sample. Lakes with pH
levels below 5.0 had only about a dozen species. In some very acidic lakes
(pH 4.0), only three species were collected (Aimer et al. 1978).
In general, species are lost from all classes of algae as pH declines.
However, proportionally larger losses occur within some groups than in
others. As a result, the dominant algae in acidic lakes are often different
from those characteristic of circumneutral lakes.
In six out of nine investigations (Table 5-5), dinoflagellates (Phylum
Pyrrophyta), and often the same species of dinoflagellates, were reported to
dominate in acidic lakes. Aimer et al. (1974, 1978) reported that the
dominant species in acidic waters sampled in the west coast region of Sweden
were Peri dim'urn inconspicuum and Gymnodinium cf. ubem'mum (both
dinoflagellates). Stokes (1980) and Van (1979) noted that, in lakes in the
Sudbury Region of Ontario with pH values below 5.0, up to 50 percent of the
biomass consisted of dinoflagellates, especially Peridiniurn 1imbatum and
Peridinium inconspicuum. In Carlyle Lake (pH 4.8 to 5.1) near Sudbury,
acidification experiments within limnocorrals resulted in the proliferation
of Peridinium 1imbatum (a 75 percent increase in biomass). At pH 4.0 this
single species accounted for 60 percent of the total phytoplankton biomass
(Van and Stokes 1978). Hendrey (1980) investigated 3 lakes in the Adirondack
Region of New York State. In the most acidic lake (pH 4.9), Peridinium
inconspicuum comprised a significant fraction of the biomass in the ice-free
season.Species of chrysophyceans (Phylum Chrysophyta) were also important.
The dominance of dinoflagellates in many acidic waters has not been
adequately explained (NRCC 1981).
Dinoflagellates are not always reported as the dominant algal group in acidic
environments. In a survey of Florida lakes, Crisman et al. (1980) reported
that in the most acidic lakes (pH 4.5 to 5.0) green algae (Phylum
Chlorophyta) accounted for about 60 percent of the total phytoplankton
abundance. However, the genus Peridinium was also reported as a dominant
taxon in these lakes. In Wavy Lake (pH 4.3 to 4.8) near Sudbury, Ontario,
Conroy et al. (1976) noted that chrysophyceans (Phylum Chrysophyta) of the
genus Dinobryon dominated. Together chrysophyceans and green algae
constituted an average of 90 percent of the standing crop. In two non-acidic
5-45
-------
TABLE 5-5. SUMMARY OF OBSERVATIONS RELATING SPECIES DIVERSITY AND SPECIES COMPOSITION
OF THE PHYTOPLANKTON COMMUNITY TO ACIDITY
Location
(reference)
Reductions In
species diversity
Dominant species
in acid water
Species missing
In acid water
General
comments
Swedish West
Coast (Aimer et
al. 1974,
1978)
Numbers of species per 100
sample:
pH 6-8: 30 to 80 species
pH < 5: about 12
pH < 4: 3
cn
-is.
CT>
In most acid waters:
dlnoflagellates (Pyrrophyta)
Perldlnlun Inconsplcuum
GynmodlnluB cf. uberrlnum
In a few lakes with pH about 4:
9reen algae (Chlorophyta) -
AnklstrodesBus convolutus
Oocystis submarlna
OocystlT lacustrf?
Other common species:
chrysophyceans (Chrysophyta)
Dlnobryon crenulatum
Dlnobryon sertularla
green algae (Chlorophyta)
Chlamydomonas sp.
The classes Chlorophyceae
(Chlorophyta) and
Chrysophyceae (Chrysophyta)
had greatly reduced numbers
of species
Absence of diatoms (class
Baclllarlophyceae, Phylum
Chrysophyta) and bluegreen
algae (Cyanophyta) at pH <5:
Chroococcus limneticus
Men smopedl a~tenm ssiroa
Species common In ollgotrophic
lakes, but absent at pH <6:
bluegreen algae (Cyanophyta)
Gomphosphaeria lacustris
green algae (cniorophyta) -
Scenedesmus serratus
chrysophyceans IChrysophyta)
Dlnobryon divergens
Dlnobryon bavaricum
Olnobryon borgei
TTTnobryon sucecicum
Kephyrton spirale
sticnogioeaaoeaerl e1nl1
diatoms (Chrysophyta) -
RMzosolenla longlseta
Cyclotella bodanlca
cryptophytes ICryptophyta) -
Rhodomonas mlnuta
dlnoflagellates (Pyrrophyta)
Ceratlum hirundlnella
One stop survey of
115 lakes In August
1972 and 60 lakes In
August 1976
Greatest change In
species composition
occurred In the pH
Interval 5 to 6
2.
Swedish West
Coast (Hultberg
and Andersson
1982)
Following Umlng, number of
species generally Increased
Dominant species In acid,
ollgotrophic lakes:
dlnoflagellates (Pyrrophyta) -
Perldlfllum Inconsplcuum
Gymnodlnfuro sp.
Following Hmlng, the Importance
of genus Perldlnlum declined
Prevalent (30 to 401 of the
blomass) 1n humlc lake:
bluegreen algae (Cyanophyta) -
Herlsmopedla sp.
green algae IChlorophyta) -
Oocystis sp.
Diatoms Insignificant In all
add lakes
Following 1 lining, the importance
of species of green algae
(Chlorophyta) and
chrysophyceans (Chrysophyta),
and, In some cases, diatoms
(Chrysophyta) Increased
pre- and post-11mtng
studies; long-term
monitoring of four
lakes
-------
TABLE 5-5. CONTINUED
Location
(reference)
3. Southern Norway
(Hendrey and
Wright 1976,
Lei vest ad et
al. 1976)
Reductions in Dominant species
species diversity in acid water
Number of species identified
per lake:
pH > 4.5: 10 to 25 species
pH 4 to 4.5: < 10
Species missing
in acid water
Decrease in importance of
species of green algae (Class
Chlorophyceae, Phylum
Chlorophyta)
General
comments
One-stop survey of
lakes in October
55
1974
No consistent trend relating pH
to numbers of species of
diatoms (Chrysophyta) or
bluegreen algae (Cyanophyta)
4.
Southern Norway
(Raddum et al.
1980)
The number of algal species
collected at any one time
was generally lower in clear
water acid lakes
Periodic sampling of 13
lakes throughout an
entire growing season
5.
Canadian
Shield-Sudbury
Ontario (Stokes
1980)
Indices of both diversity and
species richness declined
with decreasing pH level
At pH < 5, up to 50% of the
biomass consisted of
dinoflagellates (Pyrrophyta)
especially-
Peridinium limbatum
Peridinium Tnconspicuum
However, this was not the case
in a naturally acidic
dystrophic lake
In oligotrophic lakes with pH
< 5, importance of species of
green algae (Chlorophyta) and
chrysophyceans (Chrysophyta)
decreased
9 lakes (pH 3.9 to 7.0)
sampled at monthly
intervals for 2 summer
seasons
Acidic lakes near
Sudbury, Ontario have
high concentrations of
metals that may
influence phytopiankton
6.
Sudbury Region
of Ontario
(Van 1979)
Number of taxa observed in
acidic lakes was less than
in non-acidic lakes
Biomass in acid lakes dominated
by a dinoflagellate
(Pyrrophyta) Peridinium
inconspicuum
Proportion of the biomass
contributed by dinoflagellates
was correlated with hydrogen
ion activity, but not with
phosphorus concentration
Host common genera in acid
lakes:
dinoflagellates (Pyrrophyta) -
feridinium
cryptophytes (Cryptophyta) -
Cryptpmonas
chrysophyceans (Chrysophyta) -
Dlnobryon
green aigae (Chlorophyta) -
Chlamydomonas. Oocystis
Non-acidic oligotrophic lakes
typically dominated by
chrysophyceans (Chrysophyta)
and diatoms (Chrysophyta),
but in acidic lakes sampled
a dinoglagellate (Pyrrophyta)
dominated
Comparison of 4 acidic
lakes with 10 non-acidic
lakes. Intensive
sampling. Samples
collected at a weekly or
bi-weekly frequency at
2 m depth intervals at
the deepest spot in each
lake for one or two
summer seasons
The change in community
structure apparently
occurs over a pH range
of 4.7 to 5.6
-------
TABLE 5-5. CONTINUED
Location Reductions 1n
(reference) species diversity
7. Sudbury Region
of Ontario
(Dillon et al
1979)
Dominant species Species missing
In acid water In acid water
Following liming, dominance
shifted from dinoflagellates
(Pyrrophyta) and cryptophytes
(Cryptophyta) to the
chrysophyceans (Chrysophyta)
more typically observed In
clrcumneutral waters
General
comments
Three of the acidic lakes
sampled by Van (1979)
were limed 1973-1975;
pH levels were raised
from <4.7 to above 6
8.
Sudbury Region
of Ontario
(Conroy et al.
1976)
tn
i
-P.
CO
LaCloche
Mountain Region
of Ontario
(Kw1atkowsk1
and Roff 1976)
In the two acidic lakes, a few
genera usually dominated the
biomass, resulting in a low
diversity Index. In the
non-acidic lakes, the
biomass was more evenly
distributed throughout a
large number of genera
Strong relationship between
diversity of phytoplankton
and pH level, with the
diversity Index dropping
off sharply below pH 5.6
All of the major groups of
phytoplankton decreased
markedly in their numbers of
species with Increasing
acid conditions. Comparing
the highest pH lake sampled
(pH about 6.7) with the
most acid lake (pH about
4.4), the numbers of species
of green algae (Chlorophyta)
were reduced from 26 to 5;
Chrysophyta from 22 to 5;
bluegreen algae (Cyanophyta)
from 22 to 10. Numbers of
species of diatoms
In acidic Wavy Lake, the
dominant genus was a
chrysophycean (Chrysophyta)
Dinobryon. Most of the
species identified in Wavy
Lake belonged to the green
algae (Chlorophyta) and
chrysophyceans (Chrysophyta).
Together these two groups
represented on the average
90S of the standing crop.
In the 2 non-acidic lakes,
these 2 groups accounted for
only 211 and 23% of the
standing crop
In acidic Florence Lake, a
considerable biomass of the
bluegreen algae (Cyanophyta)
Merismopedia sp. developed in
August
Few or no diatoms (Chrysophyta)
present in acidic waters while
they dominated in non-acidic
Mi Herd Lake and were
significant In non-acidic
Flack Lake
Acidic Wavy Lake had few blue-
green algae (Cyanophyta)
In acidic Florence Lake,
however, a considerable bloom
of the bluegreen algae
Merismopedia sp. developed in
August
Both of the non-acidic lakes
also had substantial
populations of bluegreen
algae, although of different
species
Species common in acidic waters:
Bluegreen algae (Cyanophyta) -
Aphanocapsa sp
Chi
roococcus Prescottil
OsclllatorTa sp.
Rhabdpderma lineare
Green algae (Chlorophyta) -
Carterla sp.
Chlamydomonas sp.
Chi ore II a ellipsoidea
Closterium sp.
Species occuring in lower
density In acidic waters:
Aphanocapsa sp.
Cnroococcus dispersus
Chropcoccus TTineticus
OsclllatorTa sp.
Ankistrodesmus sp.
Carteria sp.
Chlamydomonas sp.
Oocystls sp.
Scenedesmus sp.
Two acidic lakes, Wavy (pH
4.3 to 4.8) and Florence
(pH 4.4 to 4.9), a
non-acidic oligotrophic
lake, Flack Lake (pH 6.1
to 7.2), and a non-
acidic mesotrophic lake,
Millerd (pH 6.0 to 6.9),
were sampled 7 to 9
times.
The Important species in each lake shifted according to pH level.
In the more neutral lakes, the green algae (Chlorophyta)
comprised between 40 and 501 of the total algal flora, with
bluegreen algae (Cyanophyta) accounting for only 301. In acidic
lakes, however, bluegreen algae constituted about 60S, and green
algae only about 251 of the algal flora.
6 lakes sampled weekly
for 2 months in 1972
and 1973 (lake pH
range of 4.4 to 6.7)
Lakes with similar pH
values had similar
species composition
as evaluated by the
coefficient of
community and
percentage similarity
of community. Thus,
community structure
in these lakes
reflected the pH
gradient
-------
TABLE 5-5. CONTINUED
Location
(reference)
Reductions In
species diversity
Dominant species
1n acid water
Species missing
In acid water
General
comments
9. Cont.
(Chrysophyta) collected In
samples were also greatly
reduced In the two most
acidic lakes relative to the
other lakes sampled
Cryptmonads (Cryptophyta) -
(considered by the authors as In the Phylum Pyrrophyta)
Cryptomonas erosa
(.ryptomonas ovata
Dlnoflagellates (Pyrrophyta) -
Species of the genera Perldlnlum
and Slenodlnlum. although present
In some lakes, never reached
significant proportions 1n either
acidic or non-acidic lakes
Many of the species of diatoms
(Chrysophyta) common to the
more neutral lakes were
absent from acidic lakes
10.
UCloche
Mountain Region
of Ontario
(Yan and Stokes
1978)
Oi
I
vo
Phytoplankton community
dominated by Perldlnlum
llmbatum (a dlnoflagellate.
Phylum pyrrophyta), and
Cryptomonas ovata
(a cryptomonad, Phylurn
Cryptophyta, but considered by
the authors 1n the Phylum
Pyrrophyta)
These 2 groups formed between
50-901 of the blomass In all
collections
Intensively sampled one
acid lake, Carlyle
Lake (pH about 5.0).
also studied by
Kw1atkowks1 and Roff,
1976. Samples
collected at weekly
Intervals late June to
late July, 1974
11. Ontario, North
of Lake Huron
(Johnson et al.
1970)
Species diversity lower 1n 2
acid contaminated lakes than
In the drcumneutral lake
Many species of the Class
Chrysophyceae (Chrysophyta),
the class Myxophyceae
(Cyanophyta; bluegreen algae),
and diatoms (Class Bacillario-
phyceae. Phylum Chrysophyta)
developed in the clrcumneutral
lake that were absent or
occurred In only small numbers
In the 2 acidic lakes
Three lakes - one
clrcumneutral and two
acidic lakes,
acidified as a result
of contamination by
acid leachate from
processing of local
uranium ores
Associated with low pH
levels were high
levels of calcium,
sulfate, and nitrate,
and, to a lesser
extent, elevated heavy
metals concentrations
-------
TABLE 5-5. CONTINUED
Location
(reference)
Reductions in
species diversity
Dominant species
in acid water
Species missing
in acid water
General
comments
12. Adirondack
Region of New
York State
(Hendrey 1980,
Hendrey et al.
1980b)
Total number of species
identified in each lake
decreased with increasing
acidity:
circumneutral lake - 64
intermediate - 38
acidic - 27
Species of the Class
Chrysophyceae (Chrysophyta)
dominated the biomass of the
most acidic lake, although
dinoflagellates (Pyrrophyta),
especially Peridinium
inconspicuum. comprised a
significant fraction of the
biomass in the Ice-free
season
Numbers of species of green
algae (Chlorophyta) and blue-
green algae (Cyanophyta)
decreased most markedly
Dinpbryon spp. (Chrysophyta) are
the typical dominant phyto-
plankters during the summer in
Adirondack lakes
Intensive sampling of
three lakes - one
acidic (pH about 4.9),
one intermediate (pH
about 5.5), and one
circumneutral (pH
about 7.0)
13. Adirondack
Region of New
York State
(Charles 1982)
All lakes with pH > about 5.8
had euplanktonic diatoms
(class Bacillariophyceae,
Phylum Chrysophyta) present in
their surface sediments.
Lakes with a lower pH had
none.
Survey of sediment
diatom assemblages and
lake water
characteristics for
39 lakes
en
i
tn
O
14. Florida Mean number of taxa in acidic
(Crlsman et al. lakes was 10.8 vs. 16.5 for
1980) non-acidic lakes
In most acidic lakes (pH 4.5 to
5.0), green algae (Chloro-
phyta) accounted for 60t of
the total phytoplankton
abundance; blue-green algae
(Cyanophyta) only 25%.
Opposite pattern in circum-
neutral lakes.
Highly acidic lakes were
dominated by:
Green algae (Chlorophyta) -
Scenedesmus
flnklstrodesmus
Staurastrum
and several species of small
coccoid green algae
Dinoflagellates (Pyrrophyta) -
Peridinium
In lakes of pH 6.5 to 7.0,
bluegreen algae (Cyanophyta)
made up 631 of total phyto-
plankton abundance, while
green algae (Chlorophyta) were
responsible for only 31%.
Opposite pattern in acidic
lakes
Survey of 13 poorly
buffered lakes in
northern Florida with
pH levels below 5.6,
and 7 comparable lakes
in southern Florida
but with pH levels
above 5.6
15. Missouri (Lind
and Campbell
1970)
Reduced species diversity in
acid lake
Study of a very acid
lake (pH 3.2 to 4.1)
affected by strip
mining)
-------
TABLE 5-5. CONTINUED
Location
(reference)
Reductions In
species diversity
Dominant species
in acid water
Species missing
in acid water
General
comments
16. England
(Haryreaves et
al. 1975)
Number of algal species
present per water was
negatively correlated with
total acidity
Survey of 15 waters with
pH levels of 3.0 or
less; most affected by
strip mining
activities
17. Smoking Hills
Region, North-
west Terr.,
Canada
(Hutchinson et
al. 1970)
In these very acidic ponds.
phytoplankton populations
were depleted
Even at these extremely low pH
levels, some species of algae
still commonly occurred:
Euglenoids (Euglenophyta) -
Euglena mutabilis
Diatom (Chrysophyta) -
Nitzshia sp.
Oinoflagellate (Pyrrophyta) -
Gymnodinium ordinal:urn
Ponds affected by
spontaneous burning
of bituminous shale
deposits. pH values
as low as 1.8
en
t
en
18. New Zealand
(Brock and
Brock 1970)
Lower pH limit below which blue-
green algae (Cyanophyta) were
unable to grow is about 4.8 to
5.0. However, at lower
temperatures (<56 C) then in
the study waters, bluegreen
algae may be able to tolerate
more acid pH values
Analysis of algal
populations along the
pH gradient as acidic
(pH about 3.8) thermal
waters and alkaline
(pH 8.2 to 8.7) hot
springs flow into a
lake, Uaimangu
Cauldron
-------
lakes sampled, these algae accounted for only 21 and 23 percent of the
standing crop. On the other hand, during the experimental acidification of
Lake 223, Ontario, from pH 6.7 to 7.0 in 1976 to pH 5.4 in 1980, the
importance of chrysophyceans gradually decreased, with a corresponding
increase in green algae (Phylum Chlorophyta) (Schindler and Turner 1982).
Blooms of Chlorella, a green alga, within the hypolimnion (associated with
increased water clarity) for the most part accounted for the increase in
importance of green algae.
A dominance of blue-green algae in acidic waters has also been reported.
Conroy et al. (1976) observed a bloom of blue-green algae (Merismopedia sp.)
in acidic Florence Lake (pH 4.4 to 4.9) in Ontario. Hultberg and Andersson
(1982) noted that blue-greens (again Men'smopedia sp.) were prevalent in
humic acid lakes in Sweden. Stokes (19au) noted tnat the typical dominance
of dinoflagellates in acidic waters near Sudbury, Ontario did not apply to
naturally acidic, dystrophic lakes. Thus, various circumstances, such as the
presence of high concentrations of humic organic materials in the water, may
be conducive to developing populations of blue-green algae under acidic con-
ditions.
Another approach to assessing the effect of acidification on phytoplankton is
to determine which taxa common in circumneutral lakes are missing or reduced
in waters at low pH levels. Again, it is difficult to generalize. Of 11
papers dealing with this question (Table 5-5), in seven papers, diatoms
(Class Bacillariophyceae, Phylum Chrysophyta) were reported to be reduced in
importance in acidic waters; green algae (Phylum Chlorophyta) in six papers,
blue-green algae (Phylum Cyanophyta) in five papers, and chrysophyceans
(Class Chrysophyceae, Phylum Chrysophyta) in four papers. In many cases,
shifts in acidity were also associated with a shift in major species within a
given group of algae.
The observation that different species of algae are characteristic of waters
with different pH levels has also been used to predict an approximate lake pH
level based upon the composition of the algal flora within the lake. Because
the siliceous cell walls of diatoms (both planktonic and benthic) are well
preserved in lake sediments, this group of algae has most frequently been
used in these analyses. Use of this technique for estimation of historic
changes in pH is discussed in greater detail in Section 5.3.2.2.2 and Chapter
E-4, Section 4.4.3.2.
5.5.2.2 Changes in Phytoplankton Biomass and Productivity—Available data on
acidification and primary productivity in acidic lakes yield no clear corre-
lation between pH level and algal biomass or productivity. Relative to
primary productivity and/or phytoplankton biomass in circumneutral lakes,
levels in acidic lakes in some cases are reduced, in others unchanged or even
increased (Table 5-6).
Field correlations must be interpreted with care. For example, lakes low in
nutrients may be particularly sensitive to acidification. At the same time,
low nutrient levels limit primary productivity. As a result, any correlation
between lake pH level and phytoplankton biomass or productivity may reflect
only their common association with nutrient status and not a cause-and-effect
relationship between pH and phytoplankton response.
5-52
-------
TABLE 5-6. THE RELATIONSHIP BETWEEN LAKE ACIDITY AND PHYTOPLANKTON BIOMASS AND/OR
PRODUCTIVITY—OBSERVED RESPONSE TO LOW pH
en
01
Co
Significant Decrease
In six lakes near Sudbury, Ontario,
concentrations of chlorophyll £ were
positively correlated (p < 0.01) with pH;
primary productivity (on a volumetric
basis) was lowest In the most acidic lake
Uwfatkowskl and Roff 1976).
In three Adirondack lakes, the most acidic
lake (pH 4.7 to 5.1) had the lowest level
of chlorophyll ^, the least acidic lake had
the highest level of primary productivity
(on an areal basis) (Hendrey 1980).
In a survey of Florida lakes, mean
chlorophyll a concentrations were
s1gnficantly~lower In acidic lakes (1.88
mg nr3) than in non-acidic lakes (7.53 mg
m-3) (Crlsman et al. 1980).
Significant Increase
In 58 lakes along the west coast of Sweden,
the largest biomass of phytoplankton occurred
In the most acidic lakes (pH 4.5). and the
lowest bionass at Intermediate pH levels
(pH 5.1 to 5.6) (Aimer et al. 1978).
In acidification experiments within llmno-
corrals in Carlyle Lake (pH 4.8 to 5.1), near
Sudbury, Ontario, after 28 days the biomass
of phytoplankton was highest at the lowest pH
tested (pH 4.0), and lowest at pH 6.0 and 6.5
(Yan and Stokes 1978).
During experimental acidification of Lake 223
(Experimental Lakes Area in western Ontario),
the pH decreased gradually from pH 6.7 to 7.0
in 1976 to pH 5.4 In 1980. Over that time
period, chlorophyll and algal biomass Increased
significantly, associated with hypollmnetlc
algal blooms of Chlorella, and apparently in
response to Increased water clarity (Schlndler
and Turner 1982).
Mo Change
The National Research Council of Canada (1981) collated
measurements of algal biomass and productivity for
oligotrophic lakes In the Canadian Shield Region of
Ontario. Neither biomass nor production were significantly
correlated with pH. Algal biomass was significantly (p <
0.01) correlated with total phosphorus concentration.
In the fall of 1973, the pH of one Ontario lake. Middle
Lake (pH about 4.4) was raised to around 7.0 by additions
of base. Total phosphorus levels did not Increase, nor did
phytoplankton biomass (Van 1979). Experimental Increases
In phosphorus levels In acidic lakes (with or without
neutralization) have, however, induced significant
Increases In phytoplankton biomass (Dillon et al. 1978.
Hendrey et al. 19800).
Within eight plastic enclosures In Eunice Lake, an
oligotrophic lake with pH 6.5 In British Columbia, acid
addition (minimum pH 5.5) resulted In no significant change
in chlorophyll content. Additions of acid plus nutrients
(minimum pH 5.0) Increased algal biomass (Maroorek 1984).
In three Swedish lakes sampled from 13 to 15 Hay 1975,
rates of phytoplankton production per volume of water were
somewhat lower in the most acidic lake (pH 4.6). However,
because of greater water transparency in this acidic lake,
measurable primary productivity was maintained to a greater
depth. Levels of primary productivity on an areal basis,
per square meter of lake surface, were similar in all three
lakes (Aimer et al. 1978).
In 13 lakes In southern Norway, chlorophyll ^ content was
not significantly correlated with lake pH (Radduo et al.
1980).
-------
Three investigators have reported lower levels of phytoplankton biomass
and/or productivity in acidic lakes than in circumneutral lakes, based on
measurements from six lakes near Sudbury, Ontario (Kwiatkowski and Roff
1976), three lakes in the Adirondacks, New York (Hendrey 1980), and a survey
of Florida lakes (Crisman et al. 1980). None of these studies included a
simultaneous analysis of nutrient availability. In addition, careful examina-
tion of data on primary productivity collected by Kwiatkowski and Roff (1976)
indicates that, with the exception of two lakes, no clear relationship exists
between productivity and lake pH. The productivity reported for the most
acidic lake (pH 4.0 to 4.6, about 3 mg C nr3 hr~l) is well within the
range normally observed in non-acidic lakes in the region (0.3 to 6.9 mg
nr3 hr1) (NRCC 1981). Values Kwiatkowski and Roff measured in the five
remaining lakes were well above the norm. Thus, no conclusive data are
available to support the hypothesis that acidification results in decreased
algal biomass and productivity.
In contrast, three field surveys and four field experiments suggest that
acidification causes no change, or perhaps even an increase, in phytoplankton
biomass (Table 5-6). Surveys in Ontario (compiled in NRCC 1981) and Norway
(Raddum et al. 1980) found no correlation between lake pH and algal biomass;
in Sweden (Aimer et al. 1978), the largest biomass occurred in the most acid-
ic lakes. Acidification experiments within limnocorrals yielded no change
(Marmorek 1984) or an increase (Yan and Stokes 1978) in algal biomass.
Experimental acidification of an entire lake (Lake 223 in the Experimental
Lakes Area, Ontario) also was associated with a significant increase in
phytoplankton biomass (Schindler and Turner 1982).
Increased accumulations of phytoplankton in acidic waters may reflect either
an associated increase in the rate of production or a decrease in the rate of
loss (e.g., decreased predation). No studies report an increase in phyto-
plankton productivity with acidification or in acidic lakes, although data
are not abundant. Two field surveys suggest no relationship between lake pH
and primary productivity (Table 5-6). Predator-prey interactions within the
plankton community are complex. Detailed studies related to effects of
acidification on phytoplankton mortality are not available. Potential
changes, based on ecological theory, are discussed in Section 5.5.4.
In a number of laboratory studies, primary productivity in algal cultures has
been shown to be a function of pH level (e.g., Hopkins and Wann 1926, Bold
1942, Sorokin 1962, Brock 1973, Goldman 1973, Moss 1973, Cassin 1974). For
each species, growth responses to pH form an inverted U-curve, with an opti-
mum pH level for maximum growth, and significantly lower growth rates at
lower and higher pH levels. The optimum pH for growth varies significantly
between species. Moss (1973) found a lower limit for growth of most algal
species at pH levels above 4.5 to 5.1. However, three of 33 species tested
grew well at pH levels below 4.0. Sixteen of 33 species were capable of
significant growth below pH 5.0. No distinct differences were found between
groups or types of algae with regard to minimum pH tolerated (Moss 1973).
Blue-green algae in general (Phylum Cyanophyta), however, may be less
tolerant of pH levels below 5.0 (Bold 1942, Brock 1973, Moss 1973).
5-54
-------
The presence of an alga at a low pH level does not necessarily imply a pref-
erence for acidic conditions or that photosynthesis and growth are optimal
(Hendrey et al. 1980b). The proliferation of Peri dim'urn species at pH levels
4.0 to 5.0 does not mean that these organisms do best at pH levels 4.0 to
5.0, only that its competitors do less well.
The growth of algae in acidic waters indicates a physiological ability to
tolerate low pH levels, and conditions associated with low pH, e.g., a shift
in the form and availability of aqueous inorganic carbon and other necessary
plant nutrients, and increased concentrations of some metals, especially
aluminum (Chapter E-4, Section 4.6.2). Research has not yet clearly defined
physiological responses of algae to acidic conditions, or why some species
can tolerate higher acidity than others.
5.5.3 Effects of Acidification on Zooplankton
Results from 14 field surveys of zooplankton communites are summarized in
Table 5-7. In each study, acidic lakes had fewer zooplankton species (e.g.,
Figure 5-3). In Norway, clearwater lakes with pH levels below 5.0 contained
7.1 species on the average as compared to 16.1 species in less acid lakes (pH
> 5.5) (Overrein et al. 1980). Sprules (1975a,b) found nine to 16 species of
crustacean zooplankton in lakes with pH levels above 5.0 in the LaCloche
Mountain Region of Ontario, but only one to seven species in acidic lakes, pH
< 5.0. In the northeastern United States, lakes with pH below 5.0 contained
three to four species of planktonic crustaceans; lakes with pH above 5.5
contained six to 10 species (Confer et al. 1983). The greatest change in
species number and types of dominant species occurred between pH 5.0 to 5.3
(Sprules 1975a, Roff and Kwiatkowski 1977).
Likewise, experimental acidification of Lake 223, Ontario, from pH 6.7 to 7.0
in 1976 to pH 5.4 in 1980, resulted in a decline in the number of zooplankton
species present in the lake. A decrease in the mean epilimnetic pH from 6.1
to 5.8 was associated with the disappearance of one species; decrease to pH
5.6 led to the loss of two more species (Malley et al. 1982).
For the most part, species dominant in acidic lakes are also important
components of zooplankton communities in non-acidic lakes in the same region.
There is no invasion by new species.
Certain species of planktonic rotifers of the genera Keratella, Kellicottia,
and Polyarthra tolerate acidic conditions and can be found in the pH range
4.4 to 7.9. In Scandinavia, species common in acidic lakes were Keratella
cochlearis, Keratella serrulata, Kellicottia longispina, Polyarthra remata.
and Polyarthra vulgariTISpecies reduced Tn abundance with acidiflcation
incl uded Asp!anchna pri'odonta, Conpchilus unjcornis, Conpchilus mincornis,
and KerateTTaThiemalis (Aimer et al. 1974, 1978; Raddum 1978; Hultberg and
Andersson 1982TTIn Ontario, species of Keratella and Kelli'cottia were also
important in acidic lakes (Keratella taurocephala, Keratella cochlearis,
Kellicottia bostorn'ensis, Kellicottia longispinia) (Roff and Kwiatkowski
1977).Experimentalacidification of Lake 223,to pH 5.4, resulted in
increased numbers of Polyarthra vulgaris, Polyarthra remata, Keratella
taurocephala, and Kellicottia longispina (Malley et al.
5-55
-------
TABLE 5-7. SUMMARY OF OBSERVATIONS RELATING SPECIES COMPOSITION, SPECIES DIVERSITY, AND
BIOMASS OF THE ZOOPLANKTON COMMUNITY TO ACIDITY
en
Changes In species composition and abundance of:
Location
(reference)
1. Southern
Sweden
(Aimer et
al. 1974.
1978)
General
observations
Number of species
lower In acid lakes
In acid lakes, often
just a few species
occur but the number
of Individuals can be
rather great
In highly acidic
lakes (pH < 5)
Polyarthra remata,
Bosalna coreqoni. and
Olaptomus gracllis
often dominate
Rotifers
Polyarthra remata,
Polyarthra vulgaris.
Keratella cochlearis. and
Kellicottia longispina
common at most pH levels,
4.4 to 7.9
Polyarthra remata dominant
in several lakes wi th pH
< 5.5
Conochllus unicorn 1s
present in many lakes but
less prevalent In acid
waters
Many of the other rotifers
Cladocerans Copepods
Bosmina coregoni common Dlaptomus gracilis and
and occurred at all pH Cyclops spp. common at all
levels pH levels
All Daphnia species Heterocope append iculata
were sensitive to low occurred mostly at pH>5.5
pH levels. Only a few
Individuals found at
pH < 6
Diaphanosoma
brachyurum, Holopedium
gibber urn, and leptodora
kindti common but
mainly at pH > 4.9
Bythotrephes longimanus
Others Comments
One-stop survey
of 84 lakes in
August 1971
Samples col-
lected with 75
(i mesh net
appear to have preferences
above 5.5
found more frequently
1n lakes with pH < 5.4.
At higher pH levels,
fish predatlon may keep
the population at low
levels
Common in non-acidic
lakes:
Diaphanosoma
Hoi ppedi uro
uapnnja cristata
Bosmina
2. Southern
Sweden
(Hultberg
and
Andersson
1982)
In acidic lakes,
zooplankton community
dominated by a few
species
Dominants in acid lakes:
Polyarthra spp.
Keratel la cochlearis
Kellicottia longispina
Common after liming:
Polyarthra spp.
Keratella cochlearis
Asplanchna priodonta
Conochllus mlncornis
Dominant in acid lakes:
Bosmina coregoni
Common after liming:
Bosmina coregoni
Diaphanosoma sp.
baphnia cristata
Limnoslda froutosa
Hoi oped 1 urn gibber urn
Cer1odaj>hnia
Dominants in acid lakes:
Eudiaptomus gracilis
Cyclops spp.
Common after liming:
Eudiaptomus gracilis
Pre- and post-
liming studies.
Effects of lim-
ing on zooplank-
ton are difficult
to evaluate due
to simultaneous
rotenone treat-
ments
quadrangula
-------
TABLE 5-7. CONTINUED
in
i
in
Location
(reference)
3. Southern
Norway
(Hendrey
and Wright,
1976)
4. Norway
(Raddun et
al. 1980,
Raddun
1978,
Hoboek and
Raddun
1980)
Changes in
General Rotifers
observations
Total number of
species collected
decreased with
decreasing pH
Number of species Species occurring with
lower in acid lakes. equal frequency in acid
In southern Norway, and non-acid clearwater
clearwater lakes with lakes:
pH < 5 held on the Kellicottia longispina
average 7.1 species; Keratella serrulata
equally acid nuraic
lakes, 11.7 species;
less acid (pH > 5.5)
clearwater lakes,
16.1 species on
average
species composition and abundance of:
Cladocerans
Daphnla galeata absent
at pH < 579
Eubosmina longispina
common at all pH
levels, 4.1 to 7.7
Hoi oped 1 urn gibberum
occurred frequently at
pH levels 4.2 to 7.2
Daphnia longispina
appeared in samples pH
4.6 to 6.8
Species occur 1ng with
equal frequency in acid
(pH < 5) and non-acid
(pH > 5.5) clearwater
lakes:
Bosmina (Eubosmina)
longispina
Copepods Others
Eudiaptomus gracilis
cannon over wide range of
pH, 4.1 to 6.6. Host
frequently dominant at low
pH levels; rarely dominant
at pH>5.5
Heterocope sal lens
occurred pH 4.1 to 6.6
Ancanthod 1 aptomus
denticornis and
Mixodiaptomus laciniatus
did not occur at pH < 5
Cycl ops scutifer appeared
at pH 4.6 to 7.7
Species occurring with Chaoborus
equal frequency in acid flavicans
and non-acid lakes: absent In
Eudiaptomus gracilis clearwater
Heterocope~sal iens acid lakes
Species more frequent in
non-acid lakes:
Cyclops scutifer
Cycl ops aby ssorum
Mesocycl ops l eucTarti
Comments
One-stop survey
of 57 lakes
during fall 1974
Samples collected
with single
verticle haul of
a 75 11 mesh
net
Survey of 27
lakes; sampled
1 to 5 times (3
vertical net
hauls, with 90
v mesh net, per
visit) from June
to September 1977
- 1979
All major groups
contributed to the
lowered number of
species, but
cladocerans
apparently most
affected
Species more frequent in
non-acidic lakes:
Conochl1 us spp.
Asplanchna sp.
Keratella"
cochlearls
Keratella
Species more frequent In
acidic lakes:
Polyarthra spp.
Species more frequent
in non-acid lakes:
Holopedium gibberum
Diaphanosoma
brachyurum
Cerlodaphnia
quadranguTa
Daphnla longispina
Daphnia 'galeata
Bythotrephes
longimanus
Polyphemus pedlculus
Leptodora kinati
-------
TABLE 5-7. CONTINUED
Changes in species composition and abundance of:
Location
(reference)
General Rotifers Cladocerans
observations
Copepods Others Comments
4. cont. Some species tolerate
acid conditions 1n
the presence of
humus, but are absent
from add clearwater
lakes
The species number of
filter-feeders
reduced In clear-
water acid lakes.
Changes for
raptorial species
not as obvious
en
CO
5. Southern Lower abundance of
Norway Daphnla longispina and
(Nllssen, Daphnla longlremus at
1980) pH
-------
TABLE 5-7. CONTINUED
en
i
en
Changes In species composition and abundance of:
Location
(reference)
6. LaCloche
Mountain
Region of
Ontario
(Roff and
Kwiatkow-
ski 1977)
General
observations
Significant reduction
in numbers of species
and numbers of
individuals at lower
pH levels (pH 4.4 to
4.8)
Diversity index
declined sharply
below pH 5.3
Mean size of
crustacean
zooplankters
identical in acid vs.
non-acid lakes
Rotifers
Standing crop of rotifers
reduced at pH levels 4.4
to 4.8
In all lakes with pH>5.8,
rotifers represented by a
variety of species with no
one species being dominant
In highly acidic waters
(pH about 4.4), Keratella
taurocephala dominated.
As the pH increased,
Keratella cochlearis,
Kellicottia bostonienis,
and Kellicottia longlsplna
increased in occurrence
Polyarthra euryptera and
Polyarthra dollchoptera
Cladocerans
Standing crop of
cladocerans reduced at
pH levels below 5;
maximum at pH 5 to 6
Leptodora kindti found
only at pH>5.0
Daphnia gal eat a
menaotae, uaphnia
retrocurva, and
Dlaphanosoma
leuchtenbergianum found
in all lakes but rare
at pH 4.4 to 4.8
Bosmina longirostris,
buDosmlna tublcen, and
Hoi oped i urn gibberum
common in all lakes
Copepods
Standing crops of
cyclopoid copepods but not
calanoid copepods reduced
at pH levels 4.4 to 4.8
Diaptomus ml nut us occured
abundantly in all lakes at
all pH levels, 4.4 to 6.0
Diaptomus oregonensis and
Epischura lacustris only
encountered in lakes with
pH>5.6
Cyclops bicuspidatus
tnomasi and nesocyclops
edax found in all lakes,
pHT.4 to 6.8
Others Comments
Six lakes with
pH levels 4.0 to
7.1 sampled at
weekly intervals
June and August
1972 and Hay and
July 1973
Vertical haul
with 60 la mesh
net; and
Schindler-Patalas
trap at various
depths.
rare at pH 4.7 to 5.0;
absent pH<4.4
-------
TABLE 5-7. CONTINUED
tn
i
CTl
o
Changes In species composition and abundance of:
Location
(reference)
7. LaCloche
Mountain
Region of
Ontario
(Sp rules
1975a,b
General
observations
Above pH 5.0,
communities with 9-16
species, 3-4
dominants; In lakes
with pH < 5.0, 1 to 7
species with only 1
or 2 dominants
Discontinuity in
species distribution
at pH 5.0 to 5.2.
641 of all species
identified occurred
never or rarely at
pH < 5.0. In some
lakes, only Diaptomus
minutus remains.
Above pH 5.0, pH had
little effect on
tolerant species and
Rotifers Cladocerans
Tolerant species
distributed independent
of pH:
Bosmina
Diaphanosoma
1 euchtenbergl anum
Holopedium glbPerun?
Never occur pH < 5.0:
Leptodora kindti
Daphnia galeata
mendotae
Dapnnia retrocurva
Daphnia ambiqua
Daphnia lonqiremis
Occur primarily in
lakes with pH < 6.0:
Polyphemus pediculus
Daphnia catawba
Daphnia pulicaria
Copepods Others
Tolerant species
distributed independent of
pH:
Mesocyclops edax
Cyclops bicuspidatus
thomasi
Diaptomus minutus
Never occur pH < 5.0:
Tropocyclops prasinus
mexicanus
Epischura lacustri s
maptomus oregonesis
Diaptomus minutus dominant
in most lakes pH < 5.0; in
some cases the only
species present
Comments
One-time sampling
of 47 lakes from
July to early
September 1972 -
1973
Vertical hauls
with either
75 M or 110 u
mesh net
pH ranged from
3.8 to 7.0
only a slight effect
on the total number
of species
In regression
analyses, pH alone
accounted for 53% of
the variance in
number of species
-------
TABLE 5-7. CONTINUED
en
Changes
Location General Rotifers
(reference) observations
8. Sudbury Numbers of species
Region of reduced In acid lakes
Ontario (pH 4.1 to 4.4) with
(Yan and an average of only
Strus 1980} 3.7 species per
sample vs 10.6 In
non-acid lakes
Total community biomass
lower In acid lakes
than In nonacid lakes.
Decreased blomass
resulted from both a
decrease 1n numbers
(except In one lake)
and the small size of
the community domi-
nants (primarily
Bosmina longirostris)
In acid lakes
The greatest reductions
were observed In the
lake with the highest
metal concentrations
Contamination with
copper and nickel
appeared to have some
effect on the zoo-
plankton community over
and above effects of
low pH
In species composition and abundance of:
Cladocerans
Major species In non-
acid lakes:
Bosmina longirostrls
Holopedium qibberum
Diaphanosoma
leuchtenberglanum
Daphnla galeata
mendotae
In add waters, Bosmina
longirostrls accounted
for an average of 79%
of the total crustacean
blomass vs 3% 1n
non-acid lakes
In acidic Clearwater
Lake zooplankton
community characterized
by the Importance of
Bosmina longirostrls.
and the absence of
Daphnla sp. and the
other common
cladocerans, Holopedium
qiDDerurn and
Diaphanosoma
leuchtenberglanum
Copepods Others
Major species In non-acid
lakes:
Cyclops blcuspidatus
thomasl
Tropocyclops praslnus
mexlcanus
Dlaptomus ml nut us
Copepods contributed an
average of 65S of the
total blomass and 851 of
the total Individuals 1n
non-acid lakes
Dlaptomus mi nut us formed
between 44 and 73% of all
crustacean zooplankton,
and dominant 1n all
non-acid lakes
In acidic Clearwater Lake,
zooplankton community
characterized by the
absence of Tropocyclops
praslnus mexlcanus and
nesocyclops edax and by
the scarcity of Cycl ops
blcuspidatus thomasl and
uiaptomus minutus
Cyclops vernal Is often a
codomlnant with Bosmina
longirostrls In early
Comments
Sampled 4 acidic
lakes (pH 4.1 to
4.4) and one less
acidic lake (pH
5.7) In the
vicinity of
Sudbury plus 6
non-acidic lakes
(pH 5.7 to 6.6)
in Muskoka-Hali-
burton Region of
Ontario
Acidic lakes also
have high levels
of copper and
nickel which may
adversely effect
zooplankton
Samples collected
summers 1973-1977
as vertical hauls
with 80 v mesh
tow net and at 2-
to 3-m intervals
with a plastic
trap
spring
-------
TABLE 5-7. CONTINUED
en
01
ro
Changes In species composition and abundance of:
Location
(reference)
9. Sudbury
Region of
Ontario
(Van et
al. 1982)
10. Georgian
Bay
Region of
Ontario
(Carter
1971)
General Rotifers
observations
In the non-acid lake, Rotifers generally form
collections included only about 1% of total
7 species on the zooplankton biomass in
average vs 3.7 from non-acidic oligotrophic
the acid lake lakes in the Sudbury area
Standing crop
generally greater in
very acid (pH 4.7 to
5.2) than in slightly
acid or alkaline
ponds
About 14 species
present in non-acid
waters were absent
from the very acid
lakes
Cladocerans
Cladocerans unimportant
in non-acid lake,
forming <5% of the
average biomass
In the acid lake,
Bosmina lonqirostris
comprised 40% of the
crustacean zooplankton
biomass
Species occurring in
all ponds independent
of pH:
Bosmina longirostris
Ceriodaphnia
quadranguTa
Piaphanosoma
leuchtenEerglanum
Species occurring only
in less acid and
alkaline ponds:
Leptodora kindti
Daphnia ambiqua
Daphnia retrocurva
Ceriodaphnia
lacustris
Bosmina coregon i
coregoni
Hoi oped i urn gibber urn
Copepods
Diaptomus minutus major
contributor to total
zooplankton biomass in the
non-acid lake. Cyclops
scutifer, Mesocyclops
edax, and I ropocyc I ops
prasinus mexicanus also
Important
In the acid lake,
Diaptomus minutus
comprised 32% of the
crustacean zooplankton
biomass. Chydorus
sphaericus and Cyclops
vernal is also common
Species occuring in all
ponds independent of pH:
Diaptomus reighardi
Cyclops vernal is
Cyclops bicuspidatus
thomasi
Mesocyclops edax
Species occuring only in
less acid and alkaline
ponds:
Epischura lacustris
Epischura lacustris
Diaptomus minutus
Diaptomus oregonensis
Tropocyclops prasinus
mexicanus
Others
2 individuals
of Chaoborus
flavicans
collected in
non-acid
lake.
In the acidic
lake,
Chaoborus
flavicans,
Chaoborus
albatus, and
cnaoborus
americanus
occurred
Comments
Pre- and post-
fertilization
study of one
acidic (pH about
4.6) and one non-
acidic (pH about
6.0) lake; only
pre-fertilization
data included
here
Samples collected
in Plexiglas
trap at 1-, 4-,
and 7-m; 76 u
mesh net
32 ponds sampled
up to 10 times
over a 3-year
period. Samples
collected with a
Clarke-Bumpus
or transparent
zooplankton trap
The acidity in
these waters is
attributed mainly
to large amounts
of organic (humic)
acids
-------
TABLE 5-7. CONTINUED
Location
(reference)
Changes in species composition and abundance of:
General
observations
Rotifers
Cladocerans
Copepods
Others
Comments
10. cont.
in
I
GO
Bosmlna longirostris
was the most consist-
ently abundant crusta-
cean in all ponds. Its
greatest numbers were
usually found in the
very acid ponds.
11. Smoking
Hills area
in North-
west Terr.,
Canada
(Hutchinson
et al. 1978)
12. Adirondack
Region of
New York
State and
White
Mountain
region of
New Hamp-
shire
(Confer et
al. 1983)
The only zooplankton
present in these very acid
waters (pH 2.8 to 3.6)
were rotifers; Branchionus
urceolaris the dominant
form
Number of zooplankton
species and
zooplankton biomass
strongly related to
pH (p < 0.01). For
each unit decrease in
pH lakes contained on
the average 2.4 fewer
species and 22.6 mg
dry wt m2" less
zooplankton biomass
Identified pH range for
distribution of
species:
Bosmina longirostris,
b.Z-6.7
Bosmina coregonl ,
».&-/.
-------
TABLE 5-7. CONTINUED
Changes In species composition and abundance of:
Location General
(reference) observations
13. Great
Britain
(Lowndes
1952)
14. Great Number of species
Britain lower in low pH
(Fryer waters In pH range 3
1980) to 7.
Rotifers Cladocerans
Identified pH range for
distribution of
species:
Dlaphanosoma
brachyurJJi,
4.3-9.8
Daphnla pulex,
b.S-'S.Z
Daphnla 1ong1sp1na,
7.0-9.2
Cer1odaphn1a
reticuiata, 6.2-9.2
Cerlodaphnla
quaarangula.
4.2-9.2
Bosmlna longlrostrls,
6.9-9.2
Polyphemus pedlculus,
4.6-9.2
Bythotrephes.
longlmanus, 6.7-7.2
Leptbdora klndtl ,
6.7-8.4
Found In waters with
pH < 5.0:
Dlaphanosoma
brachyurui
Cerioaapnma
guadranguTa
Bosmlna coregonl
Polyphemus pedlculus
Cope pods
Identified pH range for
distribution of species:
Dlaptonus graclHs.
4.7-S.Z
Cyclops abyssorun,
6.2-7.3
Cyclops vernalls,
4.4-S.Z
Cyclops bicuspldatus.
4.1-9.2
Found 1n waters with
pH < 5.0:
Cyclops abyssorum
Tropocyclops praslnus
Others Comments
One- time sampling
of 70 water
bodies
Acidity attrib-
utable primarily
to high levels of
organic (humlc)
adds
-------
NUMBER OF SPECIES PER COLLECTION
C7I
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-------
Among crustacean zooplankters, several species of cladocerans appear
sensitive to acidity. In particular, field surveys (Table 5-7) indicate that
many species of the genus Daphnia are absent or uncommon below pH 5.5 to 7.0
(Lowndes 1952, Carter 1971, Aimer et al. 1974, Sprules 1975a, Hendrey and
Wright 1976, Hobaek and Raddum 1980, Yan and Strus 1980, Nilssen 1980). In
addition, in laboratory experiments with Daphnia magna and Daphnia pulex,
reductions in survival and reproduction, and physiologicalImbalances
occurred at pH levels below 5.0 to 6.0 (Davis and Ozburn 1969, Potts and
Fryer 1979).
Counterbalancing the scarcity of daphnids in acidic lakes is an increase in
the abundance of species of the cladoceran genus Bosmina. In Scandinavia,
Bosmina coregpni and Bosmina longispina were common at all pH levels greater
than 4.1 to 4.5 (Aimer et al. 1974, Hendrey and Wright 1976, Raddum 1978,
Hultberg and Andersson 1982). In Ontario, Bosmina longirostris accounted for
a large fraction of the zooplankton biomass in acidic lakes (pH < 5) (Carter
1971, Roff and Kwiatkowski 1977, Yan and Strus 1980, Yan et al. 1982). Other
cladocerans common in temperate, oligotrophic lakes (e.g., Diaphanosoma
brachyurum, Diaphanosoma leuchtenbergianum, Leptodora kindti, Holopedium
gibberum, Polyphemus pediculus, Ceriodaphnia quadrangula, and Bythotrephes
longimanus) often are less abundant in waters with pH levels below 4.7 to 5.0
(Table 5-7). Acidification of Lake 223 down to pH 5.4, however, resulted in
no consistent trends in the numbers of Bosmina longirostris, Daphnia galeata
mendotae, and Diaphanosoma brachyurum, and a possible increase in tne numbers
of Holopedium gibberum rWalley et al. 1982).
Copepods prevalent in acidic waters (pH 4.1 to 5.0) are Diaptomus gracilis in
Scandinavia and Diaptomus minutus in North America (Table b-/).In addition,
frequently reported as common in acidic waters are Heterocope saliens in
Scandinavia, and Cyclops vernal is, Cyclops bicuspldatus thomasi, and
Mesocyclops edax in North America. Species noted as being more frequent in
non-acidic lakes include Epischura lacustris, Diaptomus oregonensis,
Tropocyclops prasinus mexicanus, Heterocope appendiculata, Ancanthodiaptomus
denticornis, and Mixodiaptomus lacTm'atus. Similarly, experimental acidifi-
cation of Lake 223 to pH 5.4 resulted in no consistent change in populations
of D i aptomus minutus, Cyclops bicuspidatus thomasi, and Mesocyclops edax, but
a decline in numbers of Tropocyclops prasinus mexicanus, and extinction of
Epischura lacustris below pH 5.8 and Diaptomus sici1is~~5elow pH 6.1 (Malley
et al. 1987H
Experimental acidification of Lake 223, Ontario also resulted in the
extinction of the opposum shrimp, Mysis relict
predator, below about pH 5.6 (Malley et al. 1982)
Of the insects, midge larvae Chaoborus spp. are important zooplankters in
many lakes. Little is known about effects of acidification on Chaoborus,
although it appears to persist in some acidic environments down to pH 4.2 to
4.5 (Scheider et al. 1975, Yan et al. 1982, Confer et al. 1983, Marmorek
1984). On the other hand, Hobaek and Raddum (1980) observed that Chapbprus
flavicans was absent in clearwater acid lakes (pH < 5.0). Nilssen (1980)
reported the extinction of Chaoborus in an acidic lake (pH 4.2 to 5.0), where
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carapace remnants in bottom sediments verified its presence in earlier
years.
No data on impacts of acidification on the productivity of the zooplankton
community are available. Studies on changes in community biomass are also
limited. Thus, the functional response of the zooplankton community to
increasing levels of acidity is still largely unknown.
Three surveys of abundance of zooplankton in acidic lakes have been
conducted, involving lakes near Sudbury, Ontario contaminated with both acid
and metals (Van and Strus 1980), lakes in the LaCloche Mountain Region of
Ontario (Roff and Kwiatkowski 1977), and headwater lakes in the Adirondacks,
New York, and White Mountain Region of New Hampshire (Confer et al . 1983).
In each case, the biomass and/or numbers of zooplankton in acidic lakes were
reduced relative to that in circumneutral lakes in the same region. Confer
et al. (1983) reported an average decrease of 22.6 mg dry wt nr2 per unit
drop in pH. Roff and Kwiatkowski (1977) concluded that standing crops of
rotifers, cladocerans, and cyclopoid copepods (but not calanoid copepods)
were reduced at pH levels below 5.0. The mean size of crustacean zooplank-
ters was, however, identical in acidic vs non-acidic waters. Van and Strus
(1980) found total community biomass to be markedly lower (by almost 80
percent, on the average) in acidic lakes (pH 4.1 to 4.4) than in non-acidic
lakes (pH > 5.7). Decreased biomass resulted from both a decrease in numbers
of individuals (except in one acidic lake) and the small size of the dominant
species (primarily Bosmina longirostris).
In contrast, experimental acidification of Lake 223, Ontario, and limno-
corrals within Lake Eunice, British Columbia, resulted in no change, or even
a slight increase, in zooplankton standing crops (Malley et al. 1982,
Marmorek 1984). The lowest pH level attained in both these cases, however,
was pH 5.4.
Although more data are necessary, particularly for regions outside Ontario,
the tentative conclusion is that acidification to pH < 5.0 results in not
only fewer species but also decreased biomass of zooplankton.
5.5.4 Explanations and Significance
5.5.4.1 Changes in Species Composition—The most discrete and identifiable
changes that occur in plankton communities with acidification are a decline
in the number of species and a shift in species composition. It is possible
to speculate on why these changes occur and what they may mean to the system.
The species that predominate in an environment are those best adapted to
survive and reproduce in that environment. Acidification changes the
environment; thus, it is not surprising that the composition of the plankton
community also changes.
Adaptation to acidic conditions, however, involves more than just an ability
to tolerate low pH levels. Numerous other chemical, physical, and biological
changes associated with acidification require organisms to make adjustments.
Chemical changes associated with low pH include elevated concentrations of
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metals and alterations in the form and availability of plant nutrients,
particularly inorganic carbon and phosphorus (Chapter E-4, Section 4.6.3.5).
With increased acidity, lake transparency typically increases (Chapter E-4,
Section 4.6.3.4), potentially altering physical mixing and thermal regimes.
Finally, as the increased acidity directly and indirectly affects other
organisms in the water, predator-prey and competitive interactions will
shift. All these factors influence which (and how many) species will be
important within an ecosystem. Unfortunately, at this time we do not know
enough about tolerances and preferences of species for pH levels, concentra-
tions of metals, etc. to elucidate which factors result in observed changes
in species composition.
One factor that has received some attention is the possible importance of
predator-prey interactions. Acidification results in a decline in abundance
of fish (Section 5.6), important zooplankton predators. Changes in plankton
communities in response to changes in fish populations have been clearly
demonstrated in numerous studies (e.g., Brooks and Dodson 1965, Hall et al .
1970, Nilssen and Pejler 1973, Zaret and Kerfoot 1975, Andersson et al .
1978a, Lynch 1979, McCauley and Briand 1979, Henrikson et al . 1980a, b, and
Lynch and Shapiro 1981). In general, in the absence of planktivorous fish,
the zooplankton community is typically dominated by large-bodied species.
Fish prey preferentially on larger, more-visible zooplankton (O'Brien 1979).
With the elimination of fish, increased populations of relatively large-
bodied carnivorous and omnivorous zooplankton (e.g., Chaoborus spp.,
Leptodora kindti , Epischura lacustris, and My sis relicta) consume smaller
zooplankton species and reduce standing crops of small-bodied zooplankton to
low levels (Dodson 1974). Often, as a result of increased zooplankton
grazing on phytoplankton, inedible algal species constitute a greater
proportion of the total phytoplankton biomass.
In acidic waters, however, the species of zooplankton that frequently
dominate are relatively small. Bosmina coregoni , Bosmina longispina, and
Bosmina longirostris are all small (maximum length about 0.5 to 0.7 mm)
compared to other species of cladocerans common in non-acidic, temperate,
oligotrophic lakes, e.g., Daphina longispina (2.2 mm), Daphm'a galeata
mendotae (2.3 mm) , Daphm'a ambigua (1.7 mm), 'Holopedium gibberum (1.2 mm) ,
Diaphano'soma brachyurum TlTI mm) , and Ceriodaphnia quadrangul a (0.9 mm)
l9
(Nilssen and Pejler 19/3, Makarewicz and Likens l9/y, Lyncn lybu) . Diaptomus
minutus, a common copepod in acidic lakes in North America, has a maximum
length of about 1.0 mm as compared to 1.2 mm for Cyclops scutifer and
Mesocyclops edax (Makarewicz and Likens 1979).
Lynch (1979), in an experimental investigation of predator- prey relationships
in a Minnesota pond, concluded that zooplankton community structure was
controlled not only by the abundance of vertebrate predators, but also by the
abundance of invertebrate predators and the relative competitive abilities of
herbivorous zooplankters. Small-bodied zooplankton are presumably less
susceptible to vertebrate predators but also more susceptible to invertebrate
predators. Small-bodied zooplankton (including Bosmina longirostris)
dominated in vertebrate- free environments when invertebrate predators (e.g.,
Chaoborus) were rare and the competitive dominant was of intermediate or
small size. Janicki and DeCosta (1979) suggested that Bosmina longirostris
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dominates in acidic Cheat Lake (impacted by acid mine drainage) because of
its high reproductive potential and the intolerance of its major predator,
Mesocyclops edax, to acidic conditions. Populations of a number of crusta-
cean pTanktonic predators including Epischura lacustris, Leptodora kindti,
and Mysis relicta do seem to be reduced Tn acidic fakes (Nilssen 1980,
Schindler and Turner 1982; Table 5-7). Data on abundance of Chaoborus are
scarce and somewhat contradictory (Section 5.5.3). The characteristic
abundance of small-bodied zooplankton in acidic lakes may, however, be
related to a reduced abundance of invertebrate predators. Data are
insufficient for a detailed analysis of this hypothesis.
The elimination of fish and the reduced importance of predacious zooplankton
in acidic lakes are probably direct consequences of acidification. Changes
in these populations occur while their food supplies are still abundant (NRCC
1981, Malley et al. 1982). The persistence of small-bodied herbivores is
indicative of their tolerance of low pH and elevated metal concentrations.
The dominance of small-bodied herbivores may, however, be the result of a
complex interaction between declining fish populations, reduced invertebrate
predation, increased water clarity, and the relative survival, growth, and
reproductive capabilities of zooplankton species in acidic environments.
In addition to changes in zooplankton communities, associated with acidic
conditions are marked shifts in the species composition of the phytoplankton
community (Section 5.5.2), an important food source for zooplankton. Some
algae are more edible than others (Porter 1977). A high proportion of the
phytoplankton in many acidic lakes are dinoflagellates, relatively large
phytoplankters that may be less readily consumed and digested by many
herbivorous zooplankters. Van and Strus (1980) found that the average
diameter of the alga Peri dim'urn inconspicuum, the dominant phytoplankter in
acidic Clearwater Lake, was 14 iTnT. Yet, the maximum size of a particle
likely to be ingested by Bosmina longirostris, the dominant zooplankter, was
10 to 14 pm, with 85 percent of the particles ingested usually less than 5
ym in diameter. The dominant phytoplankter, comprising almost one-half of
the phytoplankton biomass in Clearwater Lake, may therefore be relatively
unavailable as an energy source to the dominant zooplankter in the lake.
It is possible that the dominance of dinoflagellates in acidic waters
reflects primarily the change in zooplankton community structure. The
abundance of relatively small-bodied, herbivorous zooplankton may result in
selective removal of edible algal taxa, and the subsequent dominance of the
phytoplankton by larger, inedible forms. Van and Strus (1980), however,
discount this hypothesis. Based on observations in acidic Clearwater Lake,
Ontario, Van and Strus (1980) calculated that the filtering rate for zoo-
plankton in this acidic lake was 5 to 18 times lower than estimated rates
for non-acidic oligotrophic lakes in the same region. Assuming these
calculations are correct, herbivore grazing should exert little control over
phytoplankton community structure.
Alternatively the shift in the phytoplankton community may reflect relative
tolerance to low pH and elevated metal levels. If the tolerant species of
algae are also less edible, then transfer of energy from phytoplankton to
herbivorous zooplankton may be reduced. This may occur even though the total
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biomass and productivity of these primary producers are comparable to those
in circumneutral waters. Repercussions at higher trophic levels (e.g., fish)
are possible, but the current level of understanding suggests that changes in
phytoplankton community structure are relatively insignificant for the
ecosystem as a whole compared to other documented ecological changes
associated with acidification.
5.5.4.2 Changes in Productivity—Available data on acidification and primary
productivity in acidic lakes yield no clear correlation between pH level and
algal biomass or productivity. Primary productivity and/or phytoplankton
biomass in a few cases were lower in acidic lakes relative to circumneutral
waters, in other cases equal or even greater (Section 5.5.2.2, Table 5-6).
Changes in phytoplankton community biomass and productivity with increased
acidity may reflect a balance between positive and negative factors.
Differences in the importance of these factors between systems may account
for inconsistencies in the response of different aquatic systems to acidic
deposition.
The biomass of phytoplankton at any given time is a function of its rate of
production vs its rate of loss. In some acidic systems phytoplankton biomass
accumulates (Aimer et al. 1978, Van and Stokes 1978), suggesting either an
increase in primary productivity per unit biomass or a decrease in the loss
function. No studies have indicated increased productivity per unit biomass
with increased acidity (Section 5.5.2.2); thus, most authors (Hendrey 1976,
Hall et al. 1980) have concluded that any accumulation of algal biomass in
acidic waters results from a decreased rate of loss or depletion, i.e.,
decreased grazing or decreased decomposition. Lower zooplankton biomass or
shifts in zooplankton community structure (Section 5.5.3) may decrease
grazing pressure on phytoplankton. Such a conclusion, however, is purely
speculative.
As common as increased standing crops of phytoplankton are observations of
decreased biomass associated with acidic conditions. Conclusions that
phytoplankton biomass decreased with increasing acidity imply that either
rates of production have decreased or rates of loss have increased, or that
both have occurred. Although good evidence for lower primary productivity in
acidic waters is lacking, there is a theoretical basis suggesting that a
number of changes associated with acidification and acidic deposition could
reduce productivity. Factors that could decrease primary productivity with
declining pH levels include: (1) a shift in pH level below that optimal for
algal growth; (2) an increase in metal concentrations above those optimal for
growth; (3) decreased nutrient availability; and (4) a shift in species
composition within the phytoplankton community to species with lower
photosynthetic efficiencies.
Three primary mechanisms have been proposed whereby nutrient availability may
be reduced in acidic environments: inhibition of nutrient recycling,
decreased availability of inorganic carbon, and/or decreased availability of
phosphorus. Grahn et al. (1974) suggested that a decreased rate of decompo-
sition and the accumulation of coarse detritus, benthic algae, and macro-
phytes (especially Sphagnum) on the bottom of acidic lakes decreased
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recycling of nutrients and prevented exchange of nutrients and other ions
between sediments and the overlying water (Sections 5.3 and 5.4). A
reduction in these processes could significantly reduce quantities of
nutrients available to primary producers (e.g., Kortmann 1980) and induce
what Grahn et al. (1974) termed oligotrophication of the lake system. No
data are available to confirm this hypothesis, however.
Besides this decrease in nutrient cycling resulting from a biological
perturbation, increased acidity may also decrease nutrient availability via
chemical interactions. Potential effects on inorganic carbon and phosphorus
have received the most attention.
At lower pH levels, the total quantity of inorganic carbon available for
algal uptake is reduced and a greater proportion of it occurs as aqueous
COg rather than as bicarbonates or carbonates. The National Research
Council Canada (1981) calculated that for a typical softwater lake at pH 4.2
in equilibrium with the atmosphere, the quantity of inorganic carbon consumed
by phytoplankton per day amounted to about 14 percent of the total quantity
of dissolved inorganic carbon available in the lake. Thus, it is possible
that during periods of peak photosynthesis, phytoplankton may take up
inorganic carbon from the water at a rate faster than it can be replaced by
diffusion from the atmosphere. Phytoplankton productivity at these times may
be carbon limited. The significance of these occasional limitations during
periods of peak photosynthesis to annual levels of production has not been
evaluated.
In oligotrophic lakes, phosphorus availability often limits primary produc-
tion (Wetzel 1975, Schindler 1975). Chemical interactions between aluminum
and phosphorus (Chapter E-4, Section 4.6.3.5) in acidic waters or within
watersheds receiving acidic depositions may decrease phosphorus availability
with decreasing pH level and, as a result, decrease primary productivity.
Despite considerable research on the chemical nature of aluminum-phosphorus
interactions, no field studies regarding acidification of surface waters have
been completed to confirm or reject this hypothesis.
Shifts in species composition within the phytoplankton community with
increased acidity were discussed in preceding sections. It is possible that
species of algae predominating in acidic waters have inherently lower levels
of photosynthetic efficiency than do species dominant in similar but non-
acidic waters. In this case, a reduced level of primary productivity may be
an indirect effect of the shift in species composition. Following removal of
the fish population from an oligotrophic, circumneutral lake in Sweden, not
only did the species composition and diversity of the phytoplankton community
change, but limnetic primary production was reduced (Henrikson et al.
1980a,b). It was hypothesized that, with the removal of fish, increased
grazing pressure by zooplankton selected for relatively inedible forms of
algae, and the inedible forms were less productive and less efficient users
of available nutrients, in part because of their larger size. None of these
hypotheses has been tested. Andersson et al. (1978a) also found decreased
primary productivity in the absence of fish, and Redfield (1980) varied
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zooplankton grazing intensity and determined that concentrated grazing
decreased algal productivity.
Despite these apparently good reasons for why acidification should decrease
primary productivity, the available evidence suggests that there is no
consistent decrease. In part, this may reflect counterbalancing factors
working to increase productivity with acidification, e.g., increased lake
transparency or, to a lesser extent, increased nutrient availability
resulting from plant nutrients associated with acidic deposition.
A notable feature of many acidic lakes is their remarkable clarity. Water
chemistry changes with acidification that may contribute to increased water
clarity are discussed in Chapter E-4, Section 4.6.3.4. As the absorption and
scattering of light in the water decreases with acidification: (1) a greater
amount of light may be available for photosynthesis; (2) light may penetrate
to greater depths, increasing the size of the euphotic zone; and (3) adequate
light for photosynthesis may extend down into the thermocline and hypolimnion
where nutrient levels are generally higher (Johnson et al. 1970). Thus,
photosynthesis per unit area of lake surface may increase.
Associated with acidic deposition are relatively large inputs of sulfate and
nitrate (Chapter E-4, Section 4.3.1.1). Both are nutrients required for
plant growth. Productivity in most oligotrophic lakes, however, is
phosphorus-limited. Thus, nutrients associated with acidic deposition
probably stimulate primary productivity very little. In the few lakes that
are nitrogen-limited, the response may be more significant, but no studies
are available to confirm this.
It is obvious that transformations in the structure and function of the
plankton community with increased acidity are the result of a complex series
of reactions. There is no simple explanation for why observed differences or
changes occur, nor is there any reason to expect responses to be identical in
different aquatic systems. Photosynthesis by phytoplankton plays a signifi-
cant role in driving and controlling the metabolism of lakes (Section 5.5.1).
Any decrease in productivity could have repercussions at all trophic levels,
including reduced fish production. The limited evidence available (Section
5.6), however, indicates that direct effects of acidification on fish appear
more important than indirect food chain effects. Thus, although acidifi-
cation affects the quality and may, to a lesser extent, affect the quantity
of plankton production, the significance of these changes to the aquatic
ecosystem as a whole has yet to be established.
5.5.5 Summary
0 Acidification results in a marked shift in the structure of the
plankton community. For both phytoplankton and zooplankton, the
total number of species represented decreases with increasing
acidity. For zooplankton, the greatest change in species
composition occurs in the pH range 5.0 to 5.3; for phytoplankton,
in the pH interval 5.0 to 6.0.
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Zooplankton communities in acidic lakes are simplifications of
communities typical of circumneutral lakes in the region.
Species dominant in acidic lakes are also important components of
zooplankton communities in non-acidic lakes. In Scandinavia,
acidic lakes (pH < 5.0) are characterized by the prevalence of
Diaptomus gracilis, and Bosmina coregoni or Bosmina longispina.
In North America the typical dominant association in acidic
waters is Diaptomus minutus and/or Bosmina longirostris.
Generalizations about changes in community structure for phyto-
plankton populations with acidification are more difficult to
make. In many acidic waters (but certainly not all), dinoflag-
el lates (Phylum Pyrrophyta) predominate. Dinoflagellate species
Peridinium inconspicuum and Peridinium limbatum in particular are
reported as dominants, often constituting large proportions of
the total biomass. Dinoflagellates also occur in circumneutral
lakes. Their abundance in acidic lakes is often counterbalanced
by the absence of most planktonic species of diatoms and some
common species of green algae, blue-green algae, and chryso-
phyceans.
Despite the altered structure of the plankton community, produc-
tivity may remain unaffected. Relative to levels of primary
(phytoplankton) productivity in circumneutral lakes, primary
productivity in acidic lakes in some cases is lower, in others
equal. A cause-and-effect relationship between primary produc-
tivity and acidification has not yet been established. In two
field experiments, increased acidity resulted in increased
phytoplankton biomass. In one field experiment, acidification
had no effect on phytoplankton biomass.
Data on zooplankton productivity in acidic lakes are non-
existent. In three lake surveys, zooplankton biomass was lower
in acidic lakes than in circumneutral lakes in the same region.
In contrast, in two field acidification experiments, zooplankton
standing crop was unchanged, or even slightly increased.
Shifts in the structure and function of the plankton community
with acidification may represent both direct and indirect
reactions to the decrease in pH level. Associated with the
increased acidity are modifications in a large number of other
chemical, biological, and physical aspects of the environment
that may affect the plankton community. Because of the complexi-
ties of these interactions, little is known about what controls
potential changes in phytoplankton and zooplankton communities,
why responses differ in different lakes, and the significance of
these changes to other trophic levels. Loss of fish populations
seems to occur independently of effects of acidification on lower
trophic levels. However, phytoplankton and zooplankton do play
significant roles in nutrient and energy cycling.
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5.6 FISH (J. P. Baker)
5.6.1 Introduction
The clearest evidence for impacts of acidification on aquatic biota is the
documentation of adverse effects on fish populations. The literature is
extensive and varied. Available data on effects of acidification on fish are
of at least seven types:
1) historic records of declining fish populations in lakes and rivers,
coupled with historic records of increasing acidity;
2) historic records of declining fish populations in lakes and rivers
currently acidic but with no historic records on levels of acidity;
3) regional lake survey data and correlations of present-day fish
status with present-day acidity levels in lakes and rivers;
4) data on success/failure of fish stocking efforts related to acidity
of the surface water;
5) experimental acidification of aquatic ecosystems and observations of
biological responses;
6) results of in situ exposures of fish to acidic waters; and
7) laboratory bioassay data on survival, growth, behavior and physio-
logical responses of fish to low pH, elevated aluminum concentra-
tions, and other water quality conditions associated with acidifica-
tion.
Each of these data sets is reviewed: numbers (1) through (4) in Section
5.6.2, Field Observations; numbers (5) and (6) in Section 5.6.3, Field
Experiments; and number (7) in Section 5.6.4, Laboratory Experiments.
Combined, they provide strong evidence that acidification of surface waters
has adverse effects on fish, in some cases resulting in the gradual
extinction of fish populations from acidified lakes and rivers.
Loss of fish populations from acidified surface waters is not, however, a
simple process and cannot be accurately summarized as "X pH results in the
disappearance of "Y" species of fish. At the very least, biological and
chemical variation within and between aquatic ecosystems must be taken into
account. For example, tolerance of fish to acidic conditions varies
markedly, not only between different species but also between different
strains or populations of the same species and among individuals within the
same population. In addition, the water chemistry within an acidified
aquatic system typically undergoes substantial temporal and spatial
fluctuations. The survival of a population of fish may be more closely keyed
to the timing and duration of acid episodes in relation to the presence of
particularly sensitive life history stages, or to the availability of "refuge
areas" during acid episodes, or to the availability of spawning areas with
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suitable water quality, than to any expression of the annual average water
quality. Because of these complexities, summary of effects of acidification
on fish in one or a few simple concluding tables can be misleading. In
addition, our understanding of functional relationships between acidification
and fish responses is still incomplete.
5.6.2 Field Observations
By themselves, field observations often fail to establish cause-and-effect
responses definitively. Most extensive field observations are simply
correlations between acidity of surface waters and absence of various fish
species. Potential confounding factors, such as lake size, stream order, or
dissolved oxygen levels, are difficult to evaluate. Unfortunately, only in a
few instances are historic records available that provide concurrent docu-
mentation of the decline of the fish population and the gradual increase in
water acidity. Clear demonstration that the absence of fish resulted from
high acidity requires supporting evidence from experiments conducted in the
field or laboratory. A review of observed fish population changes apparently
related to acidification does, however, serve to establish the nature and
extent of the potential impact of acidification on fish.
5.6.2.1 Loss of Populations
5.6.2.1.1 United States
5.6.2.1.1.1 Adirondack Region of New York State. The Adirondack region
of New York State is the largest sensitive (low alkalinity) lake district in
the eastern United States where extensive acidification has been reported
(Chapter E-4, Section 4.4.3.1.2.3). The region encompasses approximately
2877 individual lakes and ponds (114,000 surface ha) (Pfeiffer and Festa
1980), and an estimated 9350 km (6700 ha) of significant fishing streams
(Colquhoun et al. 1981). Twenty-two fish species are native to the region,
including brook trout (Salvelinus fontinalis), lake trout (Salvelinus
namaycush), brown bullhead (Ictalurus nebulosus), white sucker (Catostomus
commersoni), creek chub (Semoti1 us atromaculatus), lake chub (Couesius
piumbeusT, and common shiner (Notropis cornutuDTSreeley and Bishop 1932).
In addition, a variety of other species (e.g., smallmouth bass, Micropterus
dolomieui; yellow perch, Perca flavescens) have been introduced into
Adirondack waters, especially Into the larger, more accessible lakes. Brook
trout are frequently the only game fish species resident in the many small
headwater ponds that are located at high elevations and are particularly
susceptible to acidification (Pfeiffer and Festa 1980). Although native to
the Adirondacks, in some waters brook trout populations were introduced and
must be maintained by stocking due to a lack of suitable spawning area.
Information relevant to effects of acidification on Adirondack fish popula-
tions evolves primarily from three sources: (1) a comprehensive survey of
water quality and fish populations in many Adirondack surface waters
conducted by the New York State Conservation Department in the 1920's and
1930's (Greeley and Bishop 1932), followed by sporadic sampling of lakes and
rivers up until the 1970's (data maintained on file by the state); (2) in
1975, a complete survey of all lakes (214) located above an elevation of
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610 m (Schofield 1976c); and (3) from 1978 to the present, accelerated
sampling by the New York State Department of Environmental Conservation (DEC)
of low alkalinity lakes or lakes that contain particularly valuable fisheries
resources (Pfeiffer and Festa 1980). In addition, a preliminary survey of
fish populations and water quality for 42 Adirondack streams was completed by
the DEC in 1980 (Colquhoun et al. 1981). None of these efforts has involved
intensive studies of individual aquatic systems.
Evaluations of Adirondack data to date are limited to correlations of
present-day fish status with present-day pH levels and, for a limited number
of lakes, a comparison of current data with historic data on pH and fish
population status. Each of the studies concluded that the geographic
distribution of fish is strongly correlated with pH level, and that the
disappearance of fish populations appears to have been associated with
declines in pH. Indices of fish populations in Adirondack streams were
statistically (p < 0.05) correlated with pH measurements (taken in the spring
1980) (Colquhoun et al. 1981). Schofield (1976c, 1981, 1982) noted fewer
fish species in lakes with pH levels below 5.U (Figure 5-4). Schofield and
Trojnar (1980) also observed that poor stocking success for brook trout
stocked into 53 Adirondack lakes was significantly (p < 0.01) correlated with
low pH and elevated aluminum levels.
In many of the acid waters surveyed in the 1970's, no fish species were
found. In high elevation lakes, about 50 percent of the lakes had pH less
than 5.0 and 82 percent of these acidic lakes were devoid of fish. Thus, of
the total lakes surveyed, 48 percent had no fish. High elevation lakes,
however, constitute a particularly sensitive subset of Adirondack lakes, and
these percentages do not apply to the entire Adirondack region.
Unfortunately, neither a complete survey nor a random subsampling of all
Adirondack lakes and streams has yet been attempted.
All lakes now devoid of fish need not, however, have lost their fish popula-
tions as a result of acidification or acidic deposition. A portion of these
lakes never sustained fish populations. In addition, if earlier fish
populations have disappeared, it must be demonstrated that acidification was
the cause.
For 40 of the 214 high elevation lakes, historic records are available for
the 1930's (Chapter E-4, Figure 4-22) (Schofield 1976a). In 1975, 19 of
these 40 lakes had pH levels below 5.0, and also had no fish. An additional
two lakes with pH 5.0 to 5.5 also had no fish. Thus, 52 percent had no fish.
In the 1930's, only three lakes had pH levels below 5.t) and, again, none of
these had fish at that time. One additional lake with a pH 6.0 to 6.5 also
had no fish. Thus, in the 1930's, only 10 percent of the 40 lakes had no
fish. This implies that 17 lakes (or 42 percent) have lost their fish
populations over the 40-year period. If this holds true for high elevation
lakes in general, then 39 percent (83 lakes) of the high elevation Adirondack
lakes may have actually lost fish populations. However, this assumes that
the subset of 40 lakes represents an unbiased subsample of the 214 high
elevation Adirondack lakes.
5-76
-------
cr
NORWAY
WRIGHT et al
1975
LA CLOCHE MOUNTAINS, ONTARIO
HARVEY 1975
mm
50
40
30
20
ADIRONDACK MOUNTAINS, NEW YORK
SCHOFIELD 1976c
LEGEND
FISH PRESENT
NO FISH PRESENT
4.0 4.5 5.0 5.5 6.0 6.5 7.0 7.5
Figure 5-4. Distribution of fish in relation to lake pH.
5-77
-------
For Adirondack lakes in general, the DEC reports that about 180 lakes (6
percent of the total), representing some 2900 ha (3 percent of the total),
have lost their fish populations (Pfeiffer and Festa 1980, Schofield 1981).
The basis for this estimation has not, however, been clearly delineated.
Presumably, there are 180 lakes for which recent (1970's) fish sampling
efforts have yielded no fish and for which historic records of fish surveys
(1930's to 1960's) are available that indicate the presence of fish in
earlier years. All are listed as former brook trout ponds (Pfeiffer and
Festa 1980). Because the names of these lakes have not been published and
the data are available only in DEC files, this important conclusion cannot be
critiqued or validated.
It is also necessary to demonstrate that the loss of fish from Adirondack
lakes has occurred as a result of acidic deposition and/or acidification of
surface waters.
Retzsch et al. (1982) argued that "although precipitation acidity cannot be
excluded as a possible cause, it represents only one of a number of factors
that may alter fish populations in the Adirondacks." They consider that loss
of fish populations in the Adirondacks may also be a result of (1) natural
acidification with the development of naturally acidic wetlands adjacent to
lakes (see Chapter E-4, Section 4.4.3.3.); (2) declines in the number of fish
stocked into Adirondack lakes and changes in management practices; (3)
introductions of non-native fish species; (4) increased recreational use and
fishing pressure; and (5) construction of dams (manmade or beaver) and
manipulations of lake levels and stream flow.
All of these reasons sound feasible, yet the DEC argues in return that loss
of fish has occurred in the absence of alternative explanations other than
acidification of surface waters (N.Y. DEC 1982). For example, inadvertent
introductions of non-native fish species occur primarily in accessible low
elevation waters that are generally not, at present, impacted critically by
acidification. Non-game fish species, not subject to stocking, management,
or fishing pressure, have also been reduced or eliminated. In addition,
numerous waters located in the immediate proximity of high-use public camp-
grounds in the Adirondacks have maintained excellent trout populations
throughout the years despite heavy fishing pressure (N.Y. DEC 1982). Dean et
al. (1979) evaluated the impact of black fly larvacide on 42 Adirondack
stream fish populations and found no significant differences in occurrence
and density of fish in treated versus untreated streams. By default,
acidification has been implicated as a factor causing the loss of fish in a
number of lakes and streams.
A detailed analysis of the raw data set has not, however, been published that
examines, for individual lakes, evidence for loss of fish populations and
potential explanations for these losses, including acidification. Still, the
data set in total is sufficient to conclude that loss of fish in the
Adirondacks, at least for some surface waters, was associated with
acidification. The number of fish populations adversely impacted, and the
significance of these losses relative to the total resource available in the
Adirondacks is, however, inadequately quantified at the present time.
5-78
-------
5.6.2.1.1.2 Other regions of the eastern United States. Schofield
(1982) summarized available data relating water acidity and fish population
status for areas in the eastern United States with waters potentially acidi-
fied by acidic deposition (Chapter E-4, Section 4.4.3.1.2.3). Very few of
these studies, with the exception of studies in the Adirondack region,
included comprehensive inventories of fish populations or historic changes in
fish population status with time. Davis et al. (1978) noted that in Maine
lakes biological effects have not yet been detected. Haines (1981a) dis-
cussed the potential for adverse effects of acidification on Atlantic salmon
(Salmo salar) rivers of the eastern United States. Although the rivers were
defined as "vulnerable," no discernable effect on salmon returns was
reported. Crisman et al. (1980) sampled gamefish populations in the two most
acidic lakes (pH 4.7 and 4.9) in the Trail Ridge area of northern Florida.
Populations of largemouth bass (Micrppterus salmoides) and bluegill sunfish
(Lepomis macrochirus) exhibited no clear evidence of stress directly related
to low pH values or elevated aluminum concentrations. In Pennsylvania, some
fish species have disappeared from a few headwater stream systems (Arnold et
al. 1980), but no consistent trends in the data set conclusively demonstrated
acidification impacts (Schofield 1982). Jones et al. (1983) investigated
fish kills in fish-rearing facilities in the Raven Fork watershed at
Cherokee, North Carolina. Episodic pulses of low pH and elevated aluminum
levels were identified as the cause of death, but the forest-soils complex,
rather than acidic deposition, appeared to be the primary factor controlling
H+ and aluminum in the stream following storm events. Section 5.2 reviews
the distribution of fish in naturally acidic waters of the United States.
In regions of the United States, other than the Adirondack Mountain area of
New York State, no adverse effects of acidic deposition and/or acidification
on fish have been definitely identified. Discussions generally refer only to
"potential impact."
5.6.2.1.2 Canada
5.6.2.1.2.1 LaCloche Mountain Region of Ontario. Information collected
on fish populations in the LaCloche Mountain region of Ontario provides some
of the best evidence of adverse effects of acidification on fish. The
principal source of acid entering the LaCloche area is sulfur dioxide emitted
from the Sudbury smelters located about 65 km northeast (Beamish 1976).
Large acidic inputs have resulted in relatively rapid acidification of many
of the region's lakes—acidification rapid enough that fish population
declines, and in some cases extinctions, have occurred over the course of the
15 years that the lakes have been monitored by researchers from the
University of Toronto (H. Harvey, R. Beamish, and other associates).
Metal concentrations measured in acidic waters in the LaCloche area ranged
from 2 to 5 yg Cu i'1, 8 to 12 y g Ni £-1, 24 to 36 pg Zn
jf1, and 1 to 4 wg Pb JT1 (Beamish 1976). Because of atmospheric
transport of metals from the relatively nearby Sudbury smelters, these values
may be slightly greater than levels typical of acidic waters in other regions
discussed in Section 5.6.2.
5-79
-------
The LaCloche Mountains cover 1300 km? along the north shore of Lake Huron.
Contained within this area are 212 lakes, approximately 150 of which have
been surveyed for chemical characteristics, 68 for fish populations. Fish
populations in several of the lakes have been studied in detail since the
late 1960's and early 1970's (Beamish and Harvey 1972, Beamish 1974a,b;
Beamish et al. 1975, Harvey 1975). Major sport fishes common in these lakes
include lake trout, smallmouth bass, and walleye (Stizostedion vvtreum).
Other fish occuring very frequently are yellow percff^pumpkinseedsunfish
(Lepomis gibbosus), rock bass (Ambloplites rupestris), brown bullhead, lake
herring (Coregorus artedi i), and white sucker. LaCloche Mountain lakes in
general have waters with low ionic content and are quite clear, indicative of
low organic acid content (Harvey 1975). Of 150 lakes surveyed in 1971, 22
percent had pH levels below 4.5 and 25 percent were in the pH range of 4.5 to
5.5 (Beamish and Harvey 1972).
Harvey (1975) noted that the number of species of fish in 68 LaCloche
Mountain lakes was significantly (p < 0.005) correlated with lake pH (Figure
5-4). In addition, however, number of species of fish per lake was also
significantly correlated with lake area and other physical features. Because
small lakes tend to have low pH values, the effects of these two independent
variables on fish may be confounded. A covariate analysis based on data
presented in Harvey (1975) indicated, however, that the correlation with lake
pH was still significant (p < 0.005) even after adjustment for differences in
lake area. Of the 31 lakes with pH < 5.0, 14 had no fish. Fourteen lakes
had pH values of 6.0 or greater, and all of these had at least one species of
fish present with usually seven or more species occurring.
For the 68 LaCloche Mountain lakes surveyed during 1972-73, 38 lakes are
known or are suspected to have had reductions in fish species composition
(Harvey 1975). Based on historic fisheries information, some 54 fish
populations are known to have been lost, including lake trout populations
from 17 lakes, small mouth bass from 12 lakes, largemouth bass from four
lakes, wallyeye from four lakes, and yellow perch and rock bass from two
lakes each. Assuming that lakes with current pH < 6.0 originally contained
the same number of species as lakes with an equal surface area and pH > 6.0,
an estimated 388 fish populations have been lost from the 50 lakes surveyed
with pH < 6.0 (Harvey and Lee 1982).
The gradual disappearance of fish populations with time and with increased
acidity has been described in detail for Lumsden Lake, George Lake, and
O.S.A. Lake (Table 5-8; Beamish and Harvey 1972, Beamish 1974b, Beamish et
al. 1975). Lake pH levels measured in 1961 by Hellige color comparator were
6.8, 6.5, and 5.5 in Lumsden, George, and O.S.A Lakes, respectively. In
1971-73, pH levels measured in the three lakes with a portable pH meter were
4.4, 4.8 to 5.3, and 4.4 to 4.9, respectively. In the 1950's, eight species
of fish were reported in Lumsden Lake. Over the period 1961-71, a drastic
decline in the abundance of both game and non-game fish occurred. In George
Lake, during the interval 1961-73, lake trout, walleye, burbot, and small -
mouth bass disappeared from the lake, and from 1967 to 1972 the white sucker
population decreased in number by 75 percent and in biomass by 90 percent.
For O.S.A. Lake in 1961, local residents reported good catches of lake trout
and smallmouth bass. In 1972, intensive fish sampling yielded only four
5-80
-------
TABLE 5-8. LOSS OF FISH SPECIES FROM LUMSDEN LAKE AND GEORGE
LAKE, ONTARIO (FROM BEAMISH AND HARVEY 1972, BEAMISH ET AL. 1975,
HARVEY AND LEE 1982)
Date
Species information
Lumsden Lake
1950's
1960
1960-65
1967
1968
1969
1969
1970
Eight species present
Last report of yellow perch
Last report of burbot
Sport fishery fails
Last capture of lake trout
Last capture of slimy sculpin
White sucker suddenly rare
Last capture of trout-perch
Last capture of lake herring
Last capture of white sucker
Last capture of lake chub
George Lake
1961
1965
1966
1970
1971
1972
1973
Last spawning of walleye
Last capture of smallmouth bass
Last spawning of lake trout
Last capture of trout-perch
Last capture of burbot
Most white suckers fail to spawn
Last capture of walleye
Brown bullhead fail to spawn
Northern pike, pumpkinseed sunfish, rock bass,
brown bullhead, and white sucker fail to spawn
Last capture of lake whitefish
Lake trout rare
5-81
-------
TABLE 5-8. CONTINUED
Date Species information
George Lake (continued)
1974 Northern pike and pumpkinseed sunfish rare
1978 Few age classes of white suckers remain
1979 Brook trout and muskeTlunge rare
White sucker, brown bullhead, rock bass, lake
herring, and yellow perch present
5-82
-------
yellow perch, two rock bass, and eight lake herring. By 1980, no fish
remained (Harvey and Lee 1980).
Harvey (1979) summarized the apparent tolerance of fish in the LaCloche
Mountain region to pH, based on the occurrence of species in lake surveys and
their disappearance with acidification (Figure 5-5). Beamish (1976) conclud-
ed that increased acidity was the principal factor resulting in the loss of
fish populations.
5.6.2.1.2.2 Other areas of Ontario. Harvey (1980) estimated that
approximately 200 lakes in Ontario have lost their fish populations. For the
most part, however, these lakes are in the vicinity of Sudbury, Ontario.
Studies that suggest fish loss in response to acidification for other areas
of Ontario are very limited. Although the Muskoka-Haliburton region of
Ontario receives large inputs of acidic deposition, and decreases in alkali-
nity have been suggested for some lakes (Chapter E-4, Section 4.4.3.1.2.2),
no adverse effects on fish populations have been documented; pH values
apparently have not decreased to levels harmful to fish.
5.6.2.1.2.3 Nova Scotia. In Nova Scotia, rivers with pH < 5.4 occur
only in areas underlain by granitic and metamorphic rock; all flow in a
southerly direction to the Atlantic Coast (Watt et al. 1983). Thirty-seven
rivers within this region have historic records indicating that they sus-
tained anadromous runs of Atlantic salmon. For 27 of these rivers (Table
5-9), almost complete angling catch records are available from annual reports
of Federal Fishery Offices for the period 1936 to 1980. Of these 27, five
rivers have undergone major alterations since 1936 that potentially could
have impacted salmon stocks. For the 22 remaining rivers, 12 presently have
pH > 5. Statistical analysis of angling catch from 1936 to 1980 indicated
that only one of these 12 rivers had experienced a significant (p < 0.01)
decline in salmon catch since 1936, one river a significant (p < 0.05)
increase, and 10 no significant trend in angling catch with time. In
contrast, of the 10 rivers with current pH < 5.0, nine have had significant
(p < 0.02) declines in success since 1936, and one, no significant trend.
Salmon angling records for rivers with pH < 5.0 vs pH > 5.0 are compared in
Figure 5-6. From 1936 through the early 1950's, angling catch in the two
groups of rivers were similar. After the 1950's, angling catch in rivers
with pH < 5.0 declined, while salmon catch in rivers with pH > 5.0 continued
to show no significant trend with time.
Year-to-year variations in salmon catch are considerable, reflecting the many
factors affecting angling success and reporting accuracy. Between the two
groups of rivers (pH < 5.0 and pH > 5.0), however, occurrence of high and low
success years generally correspond. Both groups of rivers are well distrib-
uted along the 500 km Atlantic coastline of Nova Scotia. Tag return data
suggest that salmon stocks in this area all share a common marine migratory
pattern. Biological and physical factors leading to greater or lesser angler
success (e.g., sea survival, river discharge rates, or juvenile year-class
survival) probably act uniformly over the entire area (Watt et al. 1983).
5-83
-------
SPECIES
YELLOW PERCH
PUMPKINSEED
ROCK BASS
WHITE SUCKER
NORTHERN PIKE
LAKE HERRING
BLUEGILL
LAKE WHITEFISH
SMALLMOUTH BASS
LARGEMOUTH BASS
LAKE TROUT
BROWN BULLHEAD
GOLDEN SHINER
IOWA DARTER
JOHNNY DARTER
COMMON SHINER
BLUNTNOSE MINNOW
4
NUMBER OF LAKES
CONTAINING SPECIES
h, , , „
1 1 1 1 t
an
*rU
07
O/
29
oc
C.J
?n
f-\j
91
to
f.
19
7
/
7
/
10
?n
£
4.5 5.0 5.5 6.0 6.5 7.0
11
19 I 13 I 8 | 10
NUMBER OF LAKES
Figure 5-5. Frequency of occurrence of fish species in six or more
La Cloche Mountain lakes in relation to pH. Vertical bar,
lowest pH recorded; dashed line, stressed populations, e.g.,
missing year classes; solid line, populations which appear
unaffected (Harvey 1979).
5-84
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TABLE 5-9. MAJOR RIVERS IN NOVA SCOTIA ON THE ATLANTIC COAST,
pH LEVELS AND STATUS OF ATLANTIC SALMON STOCKS
River
Musquodoboit6
St. Mary's
LeHavee
Ecum Secum
Petit
Ship Harbour
Gold
Salmon (Digby)
East Ship Harbour
West Ship Harbour
Moser
Quoddy
Kirby
Medway6
Salmon (Port Dufferin)
Gaspereau
Mersey6
Middle
Liscomb
Ingram
Tangier
East
Tusket
Issacs Harbour
Nine Mile
Salmon (Lawrencetown)
Clyde
Barrington
Jordan
Sable
Broad
Roseway6
Larry' s"
Mean3
PH
1980-81
6.7
6.1
5.7
5.6
5.6
5.5
5.4
5.4
5.4
5.4
5.3
5.2
5.0
5.0
5.0
4.9
4.8
4.8
4.8
4 '.7
4.6
Rangeb
PH
1979-80
6.6-6.9
6.1-6.8
6.0-6.1
5.6-5.9
5.6-6.0
5.1-5.7
5.3-5.4
5.0-5.4
5.5-6.2
5.2-5.8
4.9-5.4
5.0-5.3
5.0-5.5
4.9-5.1
4.5-4.8
4.6-4.6
4.5-4.7
4.4-4.6
4.3-4.6
4.3-4.5
4.3-4.5
Recorded
Presence (+)
or Absence (-) Regression of
of Salmon0 Angling Catch
pre-19bU 1980-82 on Yeard
+
+ + NS
D
+ NS
+ NS
+ NS
+ + +
D
D
D
+ NS
+ NS
+ NS
+ + NS
+ NS
+ NS
D
+ +
+ NS
+ +
+
+ +
+
+
t - :
+
+
+
+
+
+
5-85
-------
TABLE 5-9. CONTINUED
awatt et al. 1983, Rivers with 1980-81 mean pH recorded have angling
data available over the past 45 years and are represented in Figure
5-11.
bFarmer et al. 1981; pH range from three pH measurements per
river—April or May 1979, September or November 1979, and February or
March 1980.
cWatt et al. 1983; pre-1960 presence/absence based on catch records;
1980-82 based on electrofishing for juvenile salmon and/or catch data.
dWatt et al. 1983; 27 rivers with angling records 1936 to 1980—no
significant trend (NS), significant increase in catch with time (+)
decrease in catch with time (-), major disturbance in watershed (D).
Historical pH records available.
fpH level reported as < 4.7 in Watt et al. 1983.
5-86
-------
on
i
00
MEAN FOR 12 RIVERS WITH pH >5.0 (1980)
©—MEAN FOR 10 RIVERS WITH'pH< 5.6 (1980)
1935
1940 1945 1950 1955 I960
YEAR
1965
1970
1975
1980
Figure 5-6. Average angling success for Atlantic salmon in 22 Nova Scotia rivers since 1936. Data
were collected from reports of federal fishery offices and normalized by expressing each
river's angling catch as a percentage of the average catch in that river during the first 5
years of record (1936-40) (Watt et al. 1983).
-------
Decreases in salmon catch over time are, on the other hand, clearly
correlated with present-day pH values 5.0 and below.
Watt et al. (1983), concluded that at present in Nova Scotia, seven former
salmon rivers with mean annual pH < 4.7 no longer support salmon runs (Table
5-9). An electrofishing survey in the summer of 1980 failed to find any
signs of Atlantic salmon reproduction in any of these seven rivers. Farmer
et al. (1980), however, observed that for the most part these rivers are all
also naturally somewhat acidic (highly colored waters, indicating the
presence of organic acids), and historically had relatively low fish
production. Peat deposits and bogs are common to much of this area. Inputs
from these materials probably contribute to the low pH levels and have some
impact on salmon production. Historical records of pH for a few rivers
within this area (Chapter E-4, Section 4.4.3.1.2.2) do, however, indicate
that acidity increased from the mid-19501s to early 1970's. Acidic condi-
tions and acidification, therefore, probably contribute to the loss of
Atlantic salmon populations in Nova Scotia.
The estimated lost (rivers with pH < 4.7) or threatened (rivers with pH 4.7
to 5.0) Atlantic salmon production potential represents 30 percent of the
Nova Scotia resource, but only 2 percent of the total Canadian potential.
Atlantic salmon rivers in New Brunswick, Prince Edward Island, and other
areas of Nova Scotia generally have pH levels above 5.4 and are not under any
immediate acid threat (Watt 1981).
5.6.2.1.3 Scandinavia and Great Britain
5.6.2.1.3.1 Norway. Extensive information on acidification and loss
of fish populations in Norwegian waters has been collected under the auspices
of the joint research project SNSF—"Acid Precipitation-Effects on Forest and
Fish," 1972-1980. Documentation of the effects of acidification on fish is
derived principally from (1) yearly records of catch of Atlantic salmon in 75
Norwegian rivers from 1876 to the present; (2) a survey of water chemistry
and fish population status in 700 small lakes in southern Norway in 1974-75;
(3) collation of information on fish population status (current and historic)
for some 5000 lakes in southern Norway, validated with testfishing in 93
lakes during 1976-79; and (4) detailed analyses of historic changes in fish
population status related to land use changes with time in selected
watersheds. Together these data provide strong evidence that acidification
has had profound impacts on fish.
Statistical data for the yearly salmon catch from major salmon rivers in
Norway have been recorded since 1876 (Figure 5-7) (Jensen and Snekvik 1972,
Leivestad et al. 1976, Muniz 1981). While catch in all rivers declined
slightly from 1900 until the 1940's, in 68 northern rivers the decline was
followed by a marked increase, and catch in the 1970's equalled or exceeded
that around 1900. In contrast, in seven southern rivers, annual catch
dropped sharply over the years 1910-17, has declined steadily since then, and
is now near zero. This decrease is reflected in all seven rivers and cannot
be explained by known changes in exploitation practices. Massive fish kills
of Atlantic salmon (Section 5.6.2.4) were reported in these rivers as early
as 1911. Efforts over the last 50 years to restock with hatchery-reared fry
5-88
-------
250
200
150
20
10
1900
1920
1940
YEAR
1960
1980
Figure 5-7. Yearly yield for Atlantic salmon fisheries in seven rivers
from the southernmost part of Norway (botton curve) compared
with 68 rivers from the rest of the country (top curve).
(Leivestad et al. 1976).
5-89
-------
and finger!ings have been unsuccessful. In the seven southern rivers, pH
levels averaged 5.12 in 1975, as compared to an average pH of 6.57 for 20 of
the 68 northern rivers. Leivestad et al. (1976) reported that acidity in
southern rivers has been steadily increasing; from 1966 to 1976 hydrogen ion
concentration increased by 99 percent.
In 1974-75, the SNSF project completed a synoptic (nonrandom) survey of water
chemistry and fish population status in 700 small to medium-si zed lakes in
Stfrlandet (the four southernmost counties of Norway) (Wright and Snekvik
1978). Based on interviews with local residents, fish populations in lakes
were classified as barren, sparse population, good population, and
overpopulated. The principal species of fish was brown trout (Salmo trutta).
Other important species were perch (Perca fluviatilis), char (Salvelinus
alpinus), pike (Esox lucius), rainbow trout (Salmo gairdneri), ana orooK
trout. About 40 percent of the 700 lakes were reported as barren of fish,
and an additional 40 percent had sparse populations. Fish status was clearly
related to water chemistry; most low pH, low conductivity lakes were either
barren or had only sparse populations. Above pH 5.5, few lakes were barren.
A stepwise multiple regression of fish status against chemical variables pH,
N03~, S042', C1-, Na+, K+, Ca2+, Mg2+, A13+, and HC03" indicated that pH and
Ca2+ were the two most important chemical variables (r = 0.53).
The original data base on fish populations in Srfrlandet collected by Jensen
and Snekvik (1972) and Wright and Snekvik (1978) has gradually been extended
to the whole country. By 1980, data on fish in more than 5000 lakes in the
southern half of Norway had been collected by interviewing fisheries
authorities, local landowners, local fishermen's associations, and other
local experts (Sevaldrud et al. 1980, Overrein et al. 1980, Muniz and
Leivestad 1980a). Interview data ware validated for 93 lakes by comparison
with results from a standardized testfishing program. Interview data provid-
ed an accurate assessment of actual fish stocks for over 90 percent of the
lakes (Rosseland et al. 1980).
At present, fish population damage has apparently occurred in an area of
33,000 km2 -jn southern Norway. Twenty-two percent of the lakes at low
elevations below 200 m have lost their brown trout populations; 68 percent of
the trout populations in high altitude lakes above 800 m are now extinct. In
13,000 km2 of this area, fish populations in all lakes are extinct, or near
extinction. Water chemistry data are available for a subset of these 5000
lakes, and again fish population status is clearly correlated with pH (Figure
5-8).
Besides information on the current status of fish populations in these 5000
lakes, the SNSF project has also compiled available historic information on
changes in fish populations with time. For almost 3000 lakes in Srfrlandet,
the population status of brown trout has been recorded by local fishermen
since about 1940. The time trend for loss of populations is diagrammed in
Figure 5-9. The rate of disappearance of brown trout from lakes in
S)6rlandet has been particularly rapid since 1960. Today, more than 50
percent of the original populations have been lost, and approximately 60
percent of the remaining are in rapid decline (Sevaldrud et al. 1980).
5-90
-------
Kc < 10 MS • cm"1 n = 203
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80
60
40
20
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80
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40
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•
•
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y
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-
p
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co
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Kr 10-20 uS • cm"1 n = 320
V*
^
1
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Kc > 30
oo
cm" n = 227
oo
in
PH
LEGEND
LOST POPULATIONS
TROUT PRESENT
Figure 5-8. Status of brown trout populations from the affected areas in
the four southernmost counties (Rogaland, Vest-agder, Aust-
Agder, and Telemark) in Norway grouped according to lake pH
and conductivity. The data are given as percentage of lakes
with or without trout within each class of pH and conductivity
(Muniz and Leivestad 1980a).
5-91
-------
3000
2500
2000
1500
o
1000
500
LOST POPULATIONS
REMAINING POPULATIONS
RECENT
STATUS
Sparse
Good
1940 1950 1960 1970
YEAR
•v
STATUS OF
REMAINING
POPULATIONS
No data
on changes
Decrease
•Unaffected
• Decrease
Unaffected
No data
on changes
Figure 5-9. Time trend for population losses of brown trout in the
affected areas in the four southernmost counties (Rogaland,
Vest-Agder, Aust-Agder, and Telemark) in Norway (Sevaldrud
et al. 1980).
5-92
-------
Attempts at restocking acidified lakes containing reduced populations have
largely failed (Overrein et al. 1980).
A relationship between water acidity and fish population status or even water
acidification and concurrent loss of fish populations does not necessarily
implicate acidic deposition as the primary cause for adverse effects on fish.
Evidence for the association between acidic deposition and acidification of
surface waters is considered in Chapter E-4. However, several studies have
been completed in Norway that examine alternate explanations for acidifica-
tion, e.g., changes in land use, specifically as they relate to historic
changes in fish populations (Drabl/6s and Sevaldrud 1980, Drabltfs et al.
1980). In each of three study areas, no correlation between shifts in land
use and human activities and changes in fish status was found. Areas that
have experienced changes in land use (e.g., abandonment of pasture farms or
discontinuance of lichen harvests) do not have any higher proportion of lakes
with declines in fish population than do areas without such land use changes.
In contrast, fish population declines are correlated with inputs of acidic
deposition.
5.6.2.1.3.2 Sweden. Sweden has about 90,000 lakes, many of which have
low alkalinity and are potentially sensitive to acidic deposition. Extensive
surveys of acidification and fish population status have not, however, been
completed. In southern Sweden, 100 lakes with pH 4.3 to 7.5 were sampled in
the 1970's (Aimer et al. 1978). Apparently as a result of acidification
(i.e., disappearance of fish was associated with current low pH levels in
lakes), 43 percent of the minnow (Phoxinus phoxinus) populations, 32 percent
of the roach (Rutilus rutilus), 19 percent of the artic char, and 14 percent
of the brown trout populations had been lost. In a study of six lakes in
southern Sweden, Grahn et al. (1974) cited historic pH data suggesting a pH
decline of 1.4 to 1.7 units since the 1930-40's and the simultaneous elimi-
nation of minnows, roach, pike and brown trout from two or more of these six
lakes. Disappearances of populations of roach in lakes in southwestern
Sweden were recorded as early as the 1920's and 1930's (although not defi-
nitely correlated with acidification) (Dickson 1975). In eastern Sweden,
loss of roach from Lake Arsjon near Stockholm occurred in association with a
decrease in pH readings: pH 5.1 to 5.3 in 1974 as compared to pH 6.0
measured colometrically in the 1940's (Milbrink and Johansson 1975).
5.6.2.1.3.3 Scotland. Investigations in Scotland (Harriman and
Morrison 1980, 1982) indicated that intensive afforestation can result in
acidification of streams and subsequent reduction and loss of fish
populations. The role of acidic deposition in this acidification process has
not yet been clearly established. In a study of 12 streams draining forested
and nonforested catchments, an electrofishing survey failed to yield any
trout in most streams draining forested catchments (mean pH 4.34), while
moorland streams (mean pH 5.40) invaribly had resident trout populations.
5.6.2.2 Population Structure—The well-being of a population can be judged
in part by examination of its age composition (NRCC 1981). Theoretically,
age one fish should be more numerous than age two fish; age two fish more
numerous than age three fish; age three fish more numerous than age four
fish, etc. Two factors commonly alter this theoretical distribution: gear
5-93
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selectivity and large natural variations in year class strength. Almost all
procedures for sampling fish populations are size selective. Often, small,
young fish are poorly sampled. In addition, relative numbers of fish in each
age group may fluctuate greatly from year to year as a consequence of natural
environmental and biological factors (e.g., year-to-year temperature varia-
tions, competition between age groups). The frequent absence of one or
several age groups within a population may, however, be indicative of a
population under stress or undergoing change. Studies of fish populations in
acidic waters frequently reveal reduced or missing age groups.
Deviations from the expected age class distribution in acidic lakes result in
some cases from the absence of young fish, in others from the absence of
older fish. A population with only fairly large, fairly old individuals
suggests that recruitment and/or reproduction have failed. A population with
only young fish may imply the occurrence of a mortality factor acting only on
fish after a certain age (e.g., after sexual maturity), or an earlier
recruitment failure. Both types of distributions have been observed in
acidic waters, although the absence of young fish occurs much more
frequently. Decreased recruitment of young fish has been cited as a primary
factor leading to the gradual extinction of fish populations in acidic waters
(Schofield 1976a, Overrein et al. 1980, Haines 1981b).
Studies of lakes in the LaCloche Mountain region of Ontario by Beamish,
Harvey, and others provide detailed observations of the structure of fish
populations in acidic and acidifying lakes. White suckers were last reported
in Lumsden Lake in 1969 (Table 5-8) at a pH of 5.0 to 5.2 (Beamish and Harvey
1972) (Section 5.6.2.1.2.1). Intensive sampling in 1967 yielded no young-of-
the-year and very few age one fish, suggesting poor recruitment of white
suckers in both 1967 and 1966. In contrast, in George Lake examination of
the age distribution of white suckers in 1972 indicated no obviously missing
year classes and thus no major reproductive failures prior to 1972 (pH 4.8 to
5.3) (Beamish et al. 1975). Although reduced in number, white suckers were
still present in George Lake in 1979 (Harvey and Lee 1980). In 1972, O.S.A.
Lake had a pH of about 4.5. Intensive sampling yielded only a small number
of very old fish—eight lake herring aged 6 to 8 years, four yellow perch
aged 8 years, and two rock bass aged 13 years (Beamish 1974b). By 1980, no
fish remained in O.S.A. Lake (Section 5.6.2.1.2.1).
In addition to these intensive studies of individual lakes in the LaCloche
Mountain region, Ryan and Harvey (1977, 1980) surveyed (through rotenone
applications) the age distribution of populations of yellow perch and rock
bass in 32 and 20 LaCloche Mountain lakes, respectively. For both species,
lakes with lower pH levels had a higher frequency of populations missing the
age 0 group (young-of-the-year). The most acidic lakes yielding young-of-
the-year yellow perch and rock bass were characterized by a pH of 4.4 and
4.8, respectively.
Absence of young age groups in fish populations from acidic and acidifying
lakes has also been documented for a few lakes in the Adirondack region and
in Scandinavia. In South Lake in the Adirondacks, white suckers netted in
1957-68 (pH 5.3 in 1968) ranged in length from 15 to 51 cm, suggesting a wide
range of age classes. By 1973-75 (pH 4.9 in 1975), however, recruitment of
5-94
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young fish appears to have ceased. White suckers collected ranged from 30 to
49 cm in length. Five suckers captured in 1975 were aged 6 to 8 years
(Schofield 1976b, Baker 1981). In Lake Skarsjon in Sweden, prior to lake
liming (pH 4.5-5.5) only very large, old perch remained in the lake (Figure
5-10). One year after liming (pH ~ 6.0), reproduction was reestablished
and two size classes of perch were present, both very large, old fish and a
new group of small, one-year-old perch (Muniz and Leivestad 1980a).
Recruitment failure may result either from acid-induced mortality of fish
eggs and/or larvae or because of a reduction in numbers of eggs spawned.
Beamish and Harvey (1972) attributed the lack of reproduction in fish
populations in LaCloche Mountain lakes to a failure of adult fish to spawn.
In Lumsden Lake in 1967, no spawning activity was observed in the lake or in
the inlet or outlet streams during the normal spawning period. Mature female
white suckers were found to be resorbing their eggs in June. In George Lake,
in 1972 and 1973 about 65 to 75 percent of the population of female white
suckers failed to release their ova to be fertilized. In 1973, most brown
bullheads, rock bass, pumpkinseed sunfish, and northern pike had also not
spawned when examined after their normal spawning period (Beamish et al.
1975). Biochemical analyses of fish from George Lake indicated that females
exhibited abnormally low levels of serum calcium during the period of ovarian
maturation. Lockhart and Lutz (1977) hypothesized that a disruption in
normal calcium metabolism, induced by low pH, affected female reproductive
physiology. In these lakes, therefore, failure of female fish to spawn was
an important contributing factor to reproductive failures.
This response, failure of female fish to spawn, has not, however, been
reported elsewhere. From a survey of 88 lakes in Norway, Rosseland et al.
(1980) noted that female fish remaining in acidic lakes had normal gonads,
and indications of unshed or residual eggs were rare. Studies conducted in
Scandinavia and the United States (Schofield 1976a, Muniz and Leivestad
1980a) suggest that increased mortality of eggs and larvae in acidic waters
is the primary cause of recruitment failures. In Norway, total mortality of
naturally spawned trout eggs was observed in an acidic stream a few weeks
after spawning (Leivestad et al. 1976).
In addition to the lack of young fish in a population, associated with
recruitment failure as described above, loss of older fish has been observed
in acidic waters. Three lakes in the Tovdal River, Norway, were testfished
from 1976 to 1979 (Figure 5-11) (Rosseland et al. 1980). Before 1975, brown
trout populations in these lakes were stunted and grew to 8 to 10 years of
age. In 1975, the Tovdal River had a severe fish kill. Since 1976, no
post-spawning brown trout (age 5 and up) have been found, and the population
is dominated by young fish. Testfishing in autumn indicated the presence of
maturing recruit-spawners. By each subsequent year, however, this age group
had disappeared while their offspring survived. Researchers speculated that
stress associated with spawning activities, coupled with acid-induced stress,
resulted in significant post-spawning mortality (Muniz and Leivestad 1980a).
Harvey (1980) proposed that loss of older fish with acidification was also
occurring in George Lake (LaCloche Mountain region) (colorimetric pH 6.5 in
1960; pH 5.4 in 1979). In 1967, white suckers up to 14 years of age occurred
5-95
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o:
LU
CO
0
RECRUITMENT FAILURE
PERCH POPULATION LAKE ST. SKARSJ0N 1976
8
6
4
Approximately
15 years old
ONE YEAR AFTER LIMING
^•1975 Reproduction
1 year old
10 20
LENGTH (cm)
30
Figure 5-10. Liming of Lake St. Skarsjrfn, Sweden, in 1975 reestablished
reproduction of perch population (Muniz and Leivestad
1980a).
5-96
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CATCH YEAR
1975
1976
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J
8G
60
40
20
80
60
40
20
80
1977 S 60
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1978
inn
80
1979 60
40
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1979
-
1978
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'//t
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///
1974
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1973
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JUVEN
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1971
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1970
733{
1969
3&XL
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LEGEND
ILE FISH(I-II)
IT SPANNERS (I
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1968
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)
1967
)
|l967
Figure 5-11. Age distribution of brown trout in Lake Tveitvatn,
Tovdal, Norway (Rosseland et al. 1980)
5-97
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in the lake. By 1972, few fish were older than 6 years. Sampling in 1979
revealed a population with 90 percent of the white suckers aged 2 and 3
years.
It is unlikely that loss of older fish in either of these cases resulted from
over-fishing.
5.6.2.3 Growth—Observations on fish growth in acidic waters and changes in
growth rate over time with acidification suggest that indirect effects of
acidification, via changes in food availability, are generally insignificant
for adult fish. In very few cases have reduced growth rates been reported.
For the most part, fish in acidic and/or acidified waters grow at the same
rate or faster than fish in circumneutral waters in the same region.
Decreases in fish growth rate associated with acidification have been docu-
mented only for acidic lakes in the LaCloche Mountain region, Ontario. In
1975, Beamish et al. (1975) reported that growth rates for white suckers in
acidic George Lake (pH 4.8 to 5.3, 1972-73) had declined over the period 1967
to 1973, and this was apparently associated with lake acidification. In more
recent surveys, however, this trend appears to have reversed. Fish collected
in 1978 and 1979 were larger (at a given age) than fish in 1972, and similar
in size to fish collected in 1967 to 1968 (Harvey and Lee 1980). Therefore,
even in this instance, consistent decreases in growth over time with
increased water acidity have not occurred.
On the other hand, several studies suggest increased fish growth in acidic
waters and/or with acidification. For two acidic lakes in the Adirondacks
sampled in the 1950's and 1970's, numbers of brook trout caught decreased
over the 20-year period, and significant increases in fish growth were
observed (Schofield 1981). Roach in acidic lakes (pH 4.6 to 5.5) in Sweden
grew at substantially faster rates than roach in circumneutral lakes (pH 6.3
to 6.8) (Aimer et al. 1974, 1978). Growth of rock bass in 25 LaCloche
Mountain lakes was also significantly (p < 0.05) faster in lakes with greater
acidity, even after adjustment for effects of lake depth on fish growth (Ryan
and Harvey 1977, 1981). Jensen and Snekvik (1972) described a common pattern
of change in lakes in Stfrlandet, Norway over the last 50 years. Densities
of fish in lakes declined, presumably associated with acidification and the
onset of increased recruitment failure. Simultaneously, fishing quality
increased, with a greater number of large trout available. Eventually,
however, with continued recruitment failures, in many lakes populations
disappeared entirely.
Rosseland et al. (1980), on the other hand, in a survey of 88 lakes in
southern Norway, found no obvious tendency for increase in growth in sparse
populations in acidic lakes, despite the fact that fish from acidic lakes had
higher proportions of full stomachs and were in better condition (i.e.,
weighed more for a given length). Ryan and Harvey (1980, 1981) observed that
yellow perch in 39 LaCloche Mountain lakes grew more quickly in more acidic
waters up to age three years, but thereafter grew more slowly. In addition,
yellow perch collected from George Lake in 1973 and 1974 (pH 4.6) at age one
to four years were significantly larger than perch of the same age collected
in 1966 (pH 5.8); this trend was reversed for age groups five years and
5-98
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older. Up to age four, yellow perch feed primarily on plankton and benthic
invertebrates. Large perch feed preferentially on small fish.
Fish growth reponse to acidification may be a complex function of two
factors: acid-induced metabolic stress and food availability. Reduced
growth in acidic waters as a result of physiological stress has been noted
frequently in laboratory experiments (Section 5.6.4.1.3). Presumably,
similar responses occur in acidic lakes and streams. Observations of
increased or unchanged growth in acidified surface waters, however, suggest
that adverse effects of acidity on fish metabolism and physiology are
counterbalanced, in part or totally, by changes in food availability.
Acidification is associated with substantial changes in the structure and, in
some cases, the function of lower trophic levels (Sections 5.3 and 5.5).
Despite the fact that some important prey organisms are sensitive to acidic
conditions and, as a result, fish may be required to shift their predation
patterns, still in most acidic lakes food does not seem to be a significant
limiting factor for adult fish (Beamish et al. 1975, Hendrey and Wright
1976). Possibly, with decreased fish density resulting from recruitment
failures or fish kills, decreased interspecific and/or intraspecific competi-
tion for food supplies may lead to increased food availability for the fish
remaining. Increased food availability may balance any negative effects of
acid-induced metabolic stress.
Detailed studies of effects of food availability on fish at all life history
stages in acidic waters are not, however, available. Therefore, the conclu-
sion that shifts in food availability with acidification have no adverse
effects on fish survival or production is preliminary. The growth response
for any particular species may depend on its sensitivity to acidic conditions
relative to the sensitivity of desirable prey items. As a group, aquatic
invertebrates appear more tolerant than fish. Therefore, fish that feed
primarily on invertebrates often experience increases in growth with
acidification. However, fish that require or prefer prey intolerant of
acidification may be adversely affected by reduced food supplies.
5.6.2.4 Episodic Fish Kills—Observations of dead and dying fish in acidify-
ing waters are not common. Mechanisms of population extinction (e.g.,
recruitment failure) are often too subtle to be easily detected. However,
instances of massive acute mortalities of adult and young fish have occurred,
typically associated with rapid decreases in pH resulting from large influxes
of acid into the system during spring snowmelt or heavy autumn rains.
Chemical characteristics and occurrence of these short-term acid episodes are
described in Chapter E-4, Section 4.4.2. In general, organisms are less
tolerant of rapid increases in toxic substances than they are of chronic
exposure and gradual changes in concentration. As a result, the rapid
fluctuations in acidity associated with short-term acidification (defined in
Chapter E-4, Section 4.2.3) may be particularly lethal to fish and may play
an important role in the disappearance of fish from acidified lakes and
streams.
Fish kills apparently associated with acid episodes have been reported
numerous times in the streams and rivers of southern Norway (Jensen and
5-99
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Snekvik 1972, Muniz 1981). The first records of mass mortality of Atlantic
salmon date from 1911 and 1914, and coincide closely with the sharp drop in
salmon catch recorded for rivers in southern Norway over the years 1910-17
(Figure 5-7). Additional observations of mass mortality were reported in
1920, 1922, 1925, 1948, and 1969, in each case following either heavy autumn
rains or rapid snowmelt, particularly in May to June. In 1948, a massive
mortality of salmon and sea trout (Salmo trutta) occurred in the Frafjord
River. At least 200 dead salmon and sea trout were collected, some of the
salmon weighing more than 20 kg. The pH measurements (colorimetric) taken
when dead fish first appeared were 3.9 to 4.2. One month later the pH was
4.7 to 4.8.
A similar episode occurred in the Tovdal River (Norway) in the spring of 1975
(Leivestad et al. 1976). Dead fish were first observed at the end of March.
During the first weeks of April thousands of dead trout covered a 30 km
stretch of the river. The Tovdal River valley is sparsely populated and has
no industry. Veterinary tests failed to find signs of any known fish
diseases. The pH of the river was about 5.0. In March, at two stations
downstream, a drop in water pH was recorded apparently associated with a
period of snowmelt at altitudes below 400 m. At higher altitudes, no dead
fish were found, and temperatures probably never rose above freezing.
Leivestad and Muniz (1976) observed the physiological response of fish to
this acid episode in the Tovdal River. Trout surviving within the affected
30 km area of river had substantially lower levels of plasma chloride and
plasma sodium than did fish from apparently unimpacted reaches of the river.
In the upper reaches of the river, the snow started to melt on April 21 and
continued at a moderate rate until May 6. The pH dropped from 5.2 to a
minimum of 4.65. Blood samples from fish collected in this area on May 15
had significantly lower plasma sodium and/or chloride compared to samples
from fish from the same area taken before and after snowmelt. Leivestad and
Muniz (1976) proposed that increased acidity interfered with osmoregulation
perhaps via impairment of the active transport mechanism for sodium and/or
chloride ions through the gill epithelium. Additional evidence for the
adverse effects of acidity on ionic balance in fish is available from
laboratory bioassays (Section 5.6.4.1.5).
Fish kills attributed to short-term acidification have been reported for only
one water outside of Norway. During each spring 1978 to 1981, coincident
with spring run-off, dead and dying fish, especially pumpkinseed sunfish,
were observed in Plastic Lake, LaCloche Mountain region, Ontario (Harvey
1979, Harvey and Lee 1982). Measured pH levels were 5.5 at the lake surface
and 3.8 in the major inlet. Field experiments to verify these toxic
conditions in Plastic Lake were completed in 1981 and are described in
Section 5.6.3.3.
In addition to these observations of mass mortalities of fish attributed to
acid episodes under natural field conditions, several instances of unusually
heavy fish mortality have been reported within fish hatcheries receiving
water directly from lakes or rivers. In Norway, poor survival of eggs and
newly-hatched larvae of Atlantic salmon, attributed to water acidity, were
reported as early as 1926 in hatcheries on rivers in Stfrlandet (Muniz
5-100
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1981). In Nova Scotia, 19 to 38 percent mortality of Atlantic salmon fry
occurred in 1975 to 1978 at the Mersey River hatchery (Farmer et al. 1981).
In Norway and Nova Scotia, neutralization of the water by passage through
limestone alleviated the problem. In the Adirondacks, adult, yearling, and
larval brook trout, which had been maintained without incident over the
winter 1976-77 in water from Little Moose Lake, experienced distress and
mortality during the first major winter thaw in early March (Schofield and
Trojnar 1980). The minimum pH measured was 5.9 on March 13 (with 0.39 mg Al
JT1). Mortalities occurred over a 5-day period from March 13 to 17.
Deaths included three adult brook trout, 25 yearlings (132 to 167 mm), and an
undetermined number of recently hatched fry. Eyed brook trout eggs exposed
to the same water did not experience significant mortality.
All of the above observations of fish kills were associated with episodic
increases in acidity. Grahn (1980), however, recorded fish kills in two
lakes in Sweden associated with decreases in acidity. In June 1978 in Lake
Ransjon and in June 1979 in Lake Amten, large numbers of dead ciscoe
(Coregonus albula) were discovered. A weather pattern of heavy rainfall,
decreasing pH levels, and increasing aluminum concentrations in the lakes,
followed by a long period of dry, sunny weather preceded fish kills in both
lakes. The pH levels in the lake epilimnion during this long, dry period
increased from approximately 4.9 and 5.4 to 5.4 and 6.0, respectively. Grahn
(1980) hypothesized that the increase in pH level precipitated aluminum
hydroxide and that ciscoe, migrating into the epilimnion to feed, were
exposed to these lethal conditions. Laboratory experiments (Section 5.6.4.2)
have also noted that aluminum is particularly toxic to fish as it precipi-
tates out of solution. Dickson (1978) reported that acidic lake waters
immediately after liming (pH values increased to 5.5 and above), were toxic
to trout. Concentrations of aluminum were still high and, presumably,
aluminum would be actively precipitating out of solution.
5.6.2.5 Accumulation of Metals in Fish—An indirect result of acidification
of surface waters may be accumulation of metals in fish. Evidence for this
relationship is derived from correlations between metal concentrations in
fish and lake and stream pH levels, and evaluations of metal chemistry and
availability in oligotrophic, acidic waters. Data are presented in Chapter
E-6, Section 6.2.3. Elevated levels of mercury in fish from acidic waters
have been measured in Sweden, Ontario, and the Adirondack region of New York
(Aimer et al. 1978, Schofield 1978, Bloomfield et al. 1980, Hakanson
1980, Jernelov 1980, Suns et al. 1980). There is no evidence that this
bioaccumulation has adverse effects on the fish, although it may represent a
hazard for human health. Other metals in addition to mercury occur at
elevated concentrations in acidified waters and potentially may accumulate in
fish and other biota. Data on these accumulations and their effects on fish
are, however, very limited.
5.6.3 Field Experiments
Correlations between fish population status and acidity of surface waters,
and field observations of declines in fish populations concurrent with
acidification of a lake, river, or stream, strongly imply that acidification
has serious detrimental effects on fish. Such observations, however, rarely
5-101
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prove cause and effect. In experiments, one variable is changed, and the
response to that change is recorded. Thus, the cause and its effect are
clearly delineated.
Whole-ecosystem acidification experiments have been carried out at two
locations: Lake 223 in the Experimental Lakes Area, Ontario and Norn's Brook
in the Hubbard Brook Experimental Forest, New Hampshire. In both cases, acid
was added directly to the water and pH levels were held fairly constant.
Despite these deviations from the process of acidification in nature, results
from these two experiments demonstrate important biological changes
associated with increased water acidity.
5.6.3.1 Experimental acidification of Lake 223, Ontario—Lake 223 is a
small, oligotrophic lake on the Precambrian Shield of western Ontario. Prior
to acidification, surface waters had an average alkalinity of about 80 yeq
£-1 and pH of 6.5 to 6.9. Five species of fish were present: lake trout,
white sucker, fathead minnow (Pimephales promelas) , pearl dace (Semotolus
margarita) and slimy sculpin (Cottus cognatus). Beginning in 1976, additions
ot sulfuric acid to the lake epilimnion gradually reduced lake pH. Early in
each ice-free season, lake pH was decreased to a predetermined value and then
maintained at that value through the following spring, at which time pH was
again reduced. Mean pH values were 6.8 in 1976, 6.1 in 1977, 5.8 in 1978,
5.6 in 1979, 5.4 in 1980, and 5.1 in 1981. Biological responses to this
acidification have been described in Schindler et al. 1980b, Schindler 1980,
Mai ley et al. 1982, Schindler and Turner 1982, Mills 1984, NRCC 1981, and
U.S./Canada MOI 1982, and are summarized in Table 5-10.
A number of important biological changes occurred at pH values of 5.8 to 6.0,
notably the disappearance of the opossum shrimp (Mysis re!ieta), a benthic/
planktonic crustacean (Section 5.5.3), and the collapse of the fathead minnow
population. Although both these species were important prey for lake trout,
no effects on trout populations were detected. Lake trout density and
population structure remained stable, and year-class recruitment failures
were not detected until 1981 at a pH of 5.1. At the onset of acidification
(1976), fathead minnows were abundant while pearl dace were rare. With the
collapse and eventual extinction of the fathead minnow population as the pH
declined to 5.5, pearl dace abundance increased dramatically (perhaps in
response to the loss of its closest competitor). The increased abundance of
pearl dace and a succession of strong year classes of white suckers in 1978
to 1980 apparently provided adequate food alternatives for the lake trout.
Despite many changes in lower trophic levels, lake trout and white sucker
populations showed no definite indications of stress until 1981, pH about
5.1, when reproductive failures occurred. During the early years of
acidification, population numbers of both species increased and growth rates
were relatively unchanged. The primary food source for white suckers,
benthic dipterans, increased in abundance. Although types of prey available
to lake trout changed dramatically, suitable food remained abundant. Both
species spawned successfully all years of study prior to 1981, and there were
no indications of egg resorption or skeletal malformations.
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TABLE 5-10. BIOLOGICAL CHANGES IN LAKE 223 IN RESPONSE TO
EXPERIMENTAL ACIDIFICATION (MILLS 1984, SCHINDLER AND TURNER 1982)
PH
Recorded change
Below 6.5 Increased bacterial sulfate reduction partially neutralize
acid additions
Increased abundance of Chlorophyta (green algae)
Decreased abundance of Chrysophyceans (golden brown algae)
Increased abundance of rotifers
Increased dipteran emergence
5.8-6.0 Disappearance of the opossum shrimp (Mysis re!ieta)
Reproductive impairment of the fathead minnow (Plmephales
pronnel as)
Possible increased embryonic mortality of lake trout
(Salve!inus namaycush)
Inhibition of calcification of exoskeleton of crayfish
(Orconectes virilis)
Disappearance of the copepod Diaptomus sicilis
5.3-5.8 Increased hypolimnetic primary production
Development of Mougeotea algal mats along shoreline
Increased infestation of crayfish with a parasite Theloham'a
sp.
Collapse of the fathead minnow population
Increased abundance of the pearl dace minnow (Semotilus
margarita)
Decreased abundance of the slimy sculpin (Cottus cognatus)
Decreased abundance of crayfish
Increased abundance of white sucker (Catostomus commersoni)
Increased abundance of lake trout
Disappearance of copepod Epischura lacustris
First appearance of the cladoceran Daphma catawba x
schoedleri
Below 5.3 Recruitment failure of lake trout
Recruitment failure of white sucker
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The population of bottom-dwell ing slimy sculpin gradually declined throughout
the acidification 1976 to 1981. Potential reasons for the decline include
direct adverse effects of increased acidity and/or increased trout predation,
associated with an increase in water clarity.
Among the fish, fathead minnow seemed to be most sensitive to acidification.
Fathead minnows are ubiquitous in lakes in northern North America and form an
important part of aquatic food chains. The population in Lake 223 disap-
peared extremely quickly, probably as a result of two factors: its
particular sensitivity to acidity and its short life span. Recruitment
failure occurred initially at pH 5.8 in 1978. Prior to acidification,
fathead minnow in Lake 223 typically lived only three years. Natural
mortality rates during their second and third years of life were extremely
high, over 50 percent per year, presumably as a result of heavy trout
predation. Few individuals remained after the second year of life.
Year-class failure in 1978, therefore, left few spawning adults (age 2 and 3)
the following year. Successive year-class failures in 1978 and 1979 assured
the rapid disappearance of this species from Lake 223.
In summary, experimental acidification of Lake 223 resulted in several
changes in fish populations at pH values as high as 5.8 to 6.0. Adverse
effects on fish and loss of populations occurred primarily as a result of
recruitment failures rather than as a result of increased mortality of adult
fish or reductions in food supplies.
5.6.3.2 Experimental Acidification of Morris Brook, New Hampshire--Norris
Brook, a third order stream inthe Hubbard Brook ExperimentalForest, New
Hampshire, was experimentally acidified to pH 4.0 from April to September
1977 (Hall et al. 1980, Hall and Likens 1980a,b). Brook trout were observed
in the study section before and after acid addition. Small numbers of trout
confined in the study section during low water in June, July, and August were
exposed continuously to water at pH 4.0 to 5.0 and total aluminum levels up
to about 0.23 mg £-1. Trout captured at pH 4.0, 5.0 and 6.4 in August
showed no evidence of pathological changes in gill structure. Most of the
trout, however, moved downstream to areas of higher pH at the onset of acid
addition in the spring. No mortality was observed, only a general avoidance
reaction. Potential effects on young-of-the-year trout and reproductive
success were not included in this study.
5.6.3.3 Exposure of Fish to Acidic Surface Waters—In addition to the above
field experiments involving acidification of an entire ecosystem, smaller
scale field experiments have been conducted involving the transfer of fish
into acidic lakes and streams. It is important to distinguish these small-
scale field experiments from similar exposures of fish to acid waters in
laboratory experiments for two reasons: (1) water quality conditions in
field experiments may undergo substantial natural fluctuations while
conditions are usually held rather constant in laboratory experiments, and
(2) many laboratory experiments create acidic water by diluting strong acids
(H2S04, HN03, HC1) into nonacidic background water. These artificially acid-
ic waters may not precisely mimic acidified surface waters and, as a result,
fish responses recorded in laboratory bioassays may not always accurately
5-104
-------
represent what would occur in the field. In this section, in situ exposures
of fish of acidic surface waters are reviewed in addition to experiments
that, although conducted in a laboratory or hatchery, used unmodified acidic
water taken directly from an acidic lake(s) and/or stream(s).
Excessive mortality of adult fish has been observed in a number of in situ
experiments with fish held in cages in acidic waters. Following observation
of fish kills in Plastic Lake (LaCloche Mountain region, Section 5.6.2.4) in
1979 and 1980, during the spring of 1981 rainbow trout (Salmo gairdneri) were
held in cages at four locations in Plastic Lake and at tour locations in a
control, non-acidic lake (Harvey et al. 1982). No mortality occurred at any
of the cage sites in the control lake (pH 6.09 to 7.34). In Plastic Lake,
however, mortality ranged from 12 percent at the lake outlet (pH 5.0 to 5.85)
to 100 percent at the inlet (pH 4.03 to 4.09). At the inlet, mortalities
commenced on the first day and all fish were dead within 48 hr. Aluminum
accumulated rapidly on the gills of fish tested in Plastic Lake.
During the winter (December to April) 1971-72, Hultberg (1977) placed
seatrout and minnows (Phoxinus phoxinus), both with a mean length of 6.5 cm,
at ten test stations ranging in pH from 4.3 to 6.0 within the watershed of
Lake Alevatten, Sweden. At all but three of the test stations native minnow
populations had disappeared within the ten years preceding the experiment.
Fifty-three percent of the seatrout and 91 percent of the minnows died during
the four-month test. Most of the mortalities (68 percent of the seatrout
total mortality; 59 percent of minnows) coincided with periodic drops in pH
level.
Several Norwegian laboratory experiments with adult fish have used acidic
stream waters (Leivestad et al. 1976, Grande et al. 1978). During simultane-
ous exposure to water from an acidic brook, pH 4.4 to 4.7, all yearling
rainbow trout, Atlantic salmon, and brown trout died within 32 days. Brook
trout were more tolerant, with 30 percent survival of one-year-old trout
after 80 days. Similarly, in tests with fingerling (age 0+) fish in acidic
stream water, rainbow trout and Atlantic salmon were least tolerant (all dead
within 12 days), brown trout intermediate (all dead within 32 days), and
brook trout substantially more tolerant (50 percent survival after 42 days).
By comparison, in stocking experiments at Lake Langtjern, Norway (mean pH
4.95), 24 and 61 percent (age 0+ and age 1+ fish, respectively) of brook
trout stocked were recaptured, as compared to 0.6 and 19 percent of the brown
trout and none of the rainbow trout (Grande et al. 1978). Long-term exposure
of brook trout to acidic stream water (mean pH 4.6, range 4.2 to 5.0)
resulted in decreased growth and reductions in plasma sodium and chloride
levels.
A number of studies have also examined survival of fish eggs incubated in
waters from acidic lakes and streams (Table 5-11). Hatching success and egg
survival of brook trout ova decreased sharply between pH levels 5.0 and 4.6.
For brown trout, hatching was near 100 percent at pH levels 6.2 and 6.5, but
0 percent at pH 4.8 and 5.1. The critical pH for hatching of Atlantic salmon
eggs appears to be 5.0 to 5.6; for walleye about pH 5.4; for roach, something
above pH 5.7.
5-105
-------
TABLE 5-11. SUMMARY OF FIELD EXPERIMENTS WITH FISH EGGS
EXPOSED TO ACIDIC SURFACE WATERS
Species
Brook trout
Brown trout
Brown trout
Atlantic
salmon
Location
Hatchery wi th
water from
Honnedaga Lake
plus 6 tribu-
tary streams
In situ in 2
Norwegian
streams
In situ in 2
Norwegian
streams
In situ in
acidic Mandal
pH %
4.5
4.6
5.0
5.1
5.3
5.4
5.6
4.8
~ 7
5.13
6.55
4.9
- 7
Survival
25
60
90
95
80
85
85
0
100
0
90
< 1
80
Comments
0.10 mg Zn £-1
0.05
0.002
0.002
0.04
0.03
0.02
Exposure from
eyed stage
Spawning observed
in acidic brook
Reference
g
d
f
d
Atlantic
salmon
Atlantic
salmon
River and a
near-neutral
tributary,
Norway
In situ at
several rivers
in Si6rlandet
Norway
5.0
5.5
In situ in
streams,
Scotland
4.2
4.4
4.9
5.8
0
0
54
30
Critical pH
for hatching
Comparison of
forested vs non-
forested catch-
ments
5-106
-------
TABLE 5-11. CONTINUED
Species
Perch
Location
In situ in
Lakes
Stensjon,
PH
4.7
5.7
7.5
% Survival Comments
28
50
89
Reference
e
Roach
Walleye
Trehorningen,
and Malaren,
Sweden
As above
In situ in
series of
small streams
in LaCloche
Mt. area,
Ontario
4.7
5.7
7.5
4.6-
6.7
0
14
100
Hatching success
significantly
reduced at pH
less than 5.4
References
aHarriman and Morrison 1982
bHendrey and Wright 1976; Muniz
cHulsman and Powles 1981
dLeivestad et al. 1976
eMilbrink and Johansson 1975
fyluniz and Leivestad 1980a
9Schofield 1965
and Leivestad 1980a
5-107
-------
In three studies, results from in situ incubation experiments were compared
with concurrent surveys of occurrence of fish species within the same waters.
Leivestad et al. (1976) reported that no brown trout eggs hatched and few
trout fry were found (by electrofishing) in an acidic tributary (pH 4.8),
formerly an important spawning ground. By contrast, in a second tributary
with inferior spawning conditions but pH 6.2, numerous trout fry were
collected. Harriman and Morrison (1982) reported no survival of Atlantic
salmon eggs incubated in acidic streams (pH 4.2 to 4.4) draining forested
catchments in Scotland and the absence of fish from the same streams in an
electrofishing survey. Finally, Milbrink and Johansson (1975) incubated
perch (Perca fluyiatilis) and roach eggs in situ in Lakes Malaren (pH 7.5),
Stensjon (pH ~ 5.7), and Trehorningen (pH ~ 4.7) in Sweden. While some
perch eggs hatched in all three lakes (89, 50, and 28 percent, respectively),
very few or no roach eggs hatched in the two acidic lakes (14 percent in Lake
Stensjon, 0 percent in Trehorningen). Likewise, perch populations occurred
in all three lakes, although extremely few perch were collected in the most
acidic lake, Trehorningen. Roach, on the other hand, have apparently dis-
appeared from Lake Trehorningen. Roach are still prevalent in both Lake
Stensjon and Malaren.
5.6.4 Laboratory Experiments
One of the best ways to prove cause and effect is to conduct experiments in a
carefully controlled environment, i.e., the laboratory. Experimental condi-
tions and fish response can be clearly quantified and dose-response relation-
ships developed with a minimum of time and effort. Unfortunately, laboratory
experiments have several drawbacks. For one, the simplified, controlled
environment of the laboratory may differ from the natural environment in
essential attributes. Factors that cannot be easily incorporated into labo-
ratory experiments include: (1) the temporal and spatial variability in the
field environment; and (2) the potential for compensatory mortality, i.e.,
shifts in the efficacy of natural mortality factors (e.g., predation,
starvation) resulting from the addition of acid-induced mortality and/or
stress. Consequently, results from laboratory experiments cannot be
translated automatically into an expected response in the field.
Serious gaps exist in the understanding of how to use laboratory results in a
quantitative assessment of field observations. It has never been definitely
demonstrated that "X" conditions that yield "Y" response in the laboratory
(e.g., 40 percent mortality) will also yield "Y" response in the field.
Laboratory results are, however, useful in firmly establishing cause and
effect, that increasing acidity has adverse effects on fish, and a qualita-
tive estimate of the levels of acidity of concern.
The more closely the laboratory environment simulates the field experience,
the more realistic the observed response. Laboratory bioassays conducted to
date vary substantially in their use of conditions appropriate to the problem
of acidification of surface waters. Most laboratory experiments concerned
with acidification have focused on the effects of low pH on fish. With
acidification, however, other factors also change in association with
decreasing pH (Chapter E-4, Section 4.6). Increased aluminum concentrations
5-108
-------
in acidic waters, in particular, have been shown to affect fish adversely
(Section 5.6.4.2). Unfortunately, most of the bioassay results to date have
failed to include aluminum. Thus, these results must be interpreted with
caution. In addition to aluminum concentration, other factors change with
acidification, e.g., increased manganese and zinc concentrations and perhaps
a decrease in dissolved organic carbon (Chapter E-4, Section 4.6). The
importance of these other changes to fish populations in acidified waters has
yet to be delineated in either laboratory or field experiments.
Within the discussion of laboratory experiments, Section 5.6.4.1 considers
effects of low pH on fish. Section 5.6.4.2 examines combined effects of both
low pH and elevated aluminum (and other metals). Because of the large number
of experiments dealing with low pH, Section 5.6.4.1 is subdivided into
experiments dealing with survival, reproduction, growth, behavior, and
physiological responses. Reproduction is arbitrarily defined as including
data on survival of fish larvae and fry in acidic water. Section 5.6.4.1.1
(Survival) therefore considers only data for fish approximately aged four
months (fingerlings) and older. Questions related to acclimation to acidic
waters and differences in tolerances among fish strains, as related to
possible mitigation of effects of acidification, are discussed in Section
5.9. Interpretation of laboratory results must also consider that fish
response in a bioassay is a function of testing conditions (e.g., tempera-
ture, flow-through or static water supply), background water quality (e.g.,
water hardness, concentrations of dissolved gases), and characteristics of
the fish tested (e.g., prior exposures and stress, size, age, condition).
5.6.4.1 Effects of Low pH
5.6.4.1.1 Survival. The majority of laboratory experiments designed to
determine the direct toxicity of elevated hydrogen ion concentrations to fish
have been short-term, acute bioassays involving principally pH levels 4.0 and
below (Table 5-12). If two days is arbitrarily selected as the length of an
acid episode, laboratory experiments suggest that a 50 percent fish kill
would occur at approximately pH 3.5 for brook trout, pH 3.8 for brown trout,
pH 3.8 to 3.9 for white suckers, and pH 4.0 for rainbow trout. In contrast,
field observations of fish kills (Section 5.6.2.4 and 5.6.3.3) indicate
mortality of: (1) Atlantic salmon and sea-run brown trout in Frafjord River,
Norway in 1948 at pH 3.9 to 4.2; (2) brown trout in the Tovdal River, Norway
in 1975 at pH 5.0; (3) rainbow trout in Plastic Lake, Ontario at pH 4.0 to
4.1; (4) brook trout in Little Moose hatchery, Adirondacks, NY, at pH 5.9;
and (5) brook trout in Sinking Creek, PA, at pH 4.4 and below.
A few experiments have considered survival of fish following longer-term
exposure to low pH levels (Table 5-13). Apparently, adult fish can survive
quite low pH levels for fairly long time periods. For periods up to 11 days,
brook trout were able to withstand pH levels as low as 4.2 with only small
reductions in survival. During even longer periods of exposure (65 to 150
days), however, a pH level of 4.4 to 4.5 was severely toxic, and only at pH
levels of 5.0 and above was brook trout survival unaffected. Long-term
experiments (> 100 days) with adult rainbow trout, brown trout, arctic char,
5-109
-------
TABLE 5-12. MEDIAN SURVIVAL TIME (HR) FOR FISH EXPOSED TO pH LEVELS
Species
Brook trout *
Rainbow trout *
*
*
*
*
Brown trout *
*
*
Arctic char *
White sucker
Roach
Age/ 2.0-
slze 2.5
10-60 g < 1
fnglt
2 g 1
90 g 1
60-130 g < 1
50 g
50-90 g
1 9
130 g
200-300 g 2
2-5 g
2-5 g
4.5-15 cm
4.5-15 cm
1-5 g
6 g 1-2
60-80 g 3
100-170 g 3
7 mo
7-13 cm
2.6- 3.0-
2.8 3.1
2-3
2 3-6
4 9
1
1-4 4
2 5
1
< 1
1
2
3-7
4 9
3 4
1
< 1
3.2-
3.3
7
3-6
12-14
18
8
2
1
2
1
pH
3.4-
3.5
6-18
45
61-66
5-9
25
10-32
18
3
2
2
3
25
5
level
3.6-
3.7
10-38
334
66-70
8
6
6
3
7
40
10
3
3.8-
3.9
14-51
23
17
27
8
18
30-200
12
4.0- 4.2- 4.4-
4.1 4.3 4.5
20-270
37
83 117 133
133
22 70
55
120
2-4
350 1000
Reference
a
b
c
c
d
e
f
g
g
h
1
i
j
j
k
1
h
h
m
j
^Experiments using low alkalinity water.
*fngl = fingerllng, age 0+, weight usually < 50 g.
References -
a. Daye and Garside 1976
b. D. W. Johnson 1975
c. Robinson et al. 1976
d. Packer and Dunson 1972
e. Swarts et al. 1978
f. Falk and Dunson 1977
g. Kwain 1975
h. Edwards and Hjeldnes 1977
i. McDonald et al. 1980
j. Lloyd and Jordan 1964
k. Brown 1981 with 0.1 mM Ca
1. Edwards and Gjedrem 1979
m. Beamish 1972
-------
TABLE 5-13. PERCENT SURVIVAL OF FISH FOLLOWING CHRONIC EXPOSURE TO LOW PH LEVELS
in
i
Species
Brook trout
*
*
Rainbow trout*
Brown trout*
Arctic char*
Fathead Minnow
Flagfish*
Age/Size
100-300g
10-60g
5g
50g
150-360g
200-300g
60-80g
100-170g
1 yr
Mature
Adult
Length
of
Exposure
(days) 3.2 3.6
5
7 0 85
11
65
150
100
100
100
400
20
4.2- 4.5- 4.8-
4.4 4.6 5.0
60-90
100 100
100 100
0-36
0 75
93
94
90
80
36 86
5.2-
5.6
100
96
98
100
75
79
5.9- 6.5-
6.2 6.8
100
75
97
95
100
85 75
100 93
7.0-
7 .5 Reference
100 a
b
c
d
100 e
f
f
f
85 g
h
*Experiments using low alkalinity water.
References
et al. 1977
bDaye and Garside 1975
°Baker 1981
dSwarts et al . 1978
eMenendez 1976
^Edwards and Hjeldnes 1977
9Mount 1973
"Craig and Baksi 1977
-------
and fathead minnow indicated no substantial reductions in survival at the
lowest pH levels tested, 5.0, 4.8 and 4.6, respectively.
An important objective of many of these experiments was not solely to
determine fish mortality at low pH levels but also to evaluate factors that
influence fish tolerance to low pH. For example, Lloyd and Jordan (1964) and
Kwain (1975) concluded that as fish grow older they became more acid
tolerant. Higher temperatures (5 to 20 C) tended to decrease fish survival
at low pH (Kwain 1975, Edwards and Gjedrem 1979, Robinson et al. 1976).
Water hardness also affected fish tolerance. Lloyd and Jordan (1964) and
McDonald et al. (1980) noted that at low pH levels (pH < 4.0) the resistance
of rainbow trout to acids increased with increasing hardness of water. As a
result, experiments conducted in high alkalinity, hard water (see Tables 5-12
and 5-13) are relatively inappropriate for assessing effects of acidic depo-
sition on fish, a phenomenon confined to dilute, poorly-buffered surface
waters. Brown (1981) suggested that higher calcium levels (more so than
higher sodium, potassium, or magnesium levels) in harder water may be re-
sponsible for the increase in resistance. Within even dilute, low alkalinity
waters, small changes in calcium concentration (0 to 2 mg £-1) have been
shown to have a significant influence on survival times of fish (Brown 1982).
Similarly, in the field (in Norway) the number of fishless lakes was corre-
lated with both pH level and calcium level, with the greatest number of
fishless lakes having both low pH and low calcium (Wright and Snekyik 1978;
Section 5.6.2.1.3.1). The sensitivity of fish to low pH obviously interacts
with a number of other stress and condition factors.
5.6.4.1.2 Reproduction. As discussed in Section 5.6.2.2, loss of fish
populations with acidification is in many lakes and rivers preceded by
successive recruitment failures. These field observations suggest that fish
reproductive processes are particularly sensitive to acidic conditions. This
conclusion is supported by laboratory experiments on effects of low pH on
spawning behavior, egg production, and egg and fry survival. Tolerance to
low pH varies considerably among the early developmental stages and repro-
ductive processes. At the same time, many fish reproduce during the spring
season, a period of large fluctuations in water chemistry. Information on
the timing of these fluctuations in water quality and the occurrence and
sensitivity of various reproductive processes and stages has yet to be tied
together in an analysis of which reproductive process(es) and/or stage(s) may
play key roles in the success or failure of recruitment and survival of the
population.
Studies on the effect of low pH on the entire reproductive cycle have been
completed only for brook trout (Menendez 1976), fathead minnow (Mount 1973),
flagfish (Jordanella floridae) (Craig and Baksi 1977), and desert pupfish
(Cyprinodon £. nevadensis](Lee and Gerking 1981) (Figure 5-12). The pH
level had some effect on all stages (processes) tested, with the exception of
number of eggs spawned by brook trout. However, sensitivity varied among
both life history stages (processes) and species. For brook trout, survival
of eggs and fry appeared to be the phase most sensitive to low pH levels,
with survival significantly (p < 0.05) reduced at pH 6.1 and below. For
fathead minnow, flagfish and desert pupfish, on the other hand, egg produc-
tion appeared particularly sensitive to low pH, with reductions in eggs
5-112
-------
Q. O
OO O
CD
C3 fc«
LU
I/)
U. 10
QC US
LU X3
=> O)
Z Q.
CO
100
80
60
it 40
LU
CD
2 20
0
O
I— '•—
< T3
3= 0)
O §
C3
C3 i-
LU 0)
^3
U. E
O 3
> o
i—i
oe *-*
to
100
80
60
40
20
4.5 5.5 6.5
pH
7.5 8.5
100
80
60
£~ 40
lo o
~ 20
0
4.5 5.5
LEGEND
• BROOK TROUT
P FATHEAD MINNOW
* FLAGFISH
a DESERT PUPFISH
6.5 7.5 8.5
PH
Figure 5-12. Effect of low pH on the reproductive cycle of fish (Menendez
1976, Mount 1973, Craig and Baksi 1977, Lee and Gerking 1981)
5-113
-------
produced per female at pH levels between 6.0 and 7.0. Lee and Gerking (1981)
concluded that reduced egg production at low pH levels resulted primarily
from inhibition of oogenesis (rather than interference with normal spawning
activity). Ruby et al. (1977, 1978) also observed retarded oocyte growth
(and reduced sperm production) for flagfish exposed to pH 6.0 relative to the
control of pH 6.8.
Unfortunately three of these four experiments (all except Craig and Baksi
1977) were conducted in hard water (alkalinity > 500 peq jr1) and two
used fish species that do not occur in surface waters sensitive to acidic
deposition. Conclusions, therefore, must be interpreted cautiously. Results
for brook trout (Menendez 1976), in particular, differ markedly from results
from other researchers using low alkalinity water (Figures 5-13 and 5-14)
and/or naturally acidic surface waters (Section 5.6.3.3). Life cycle
experiments with both fish species and conditions appropriate to acidifica-
tion of dilute surface waters are not yet available. Thus, the relative
sensitivities of reproductive stages to low pH cannot be accurately assessed
at this time.
Data on survival of fish embryos at low pH levels in laboratory experiments
are summarized in Figure 5-13. In each case, hatching was reduced at low pH
levels. Among North American freshwater species, brook trout was the most
tolerant. Excluding results from Menendez (1976), numbers of brook trout
embryos surviving through hatching were reduced substantially (< 50 percent
hatching) only at pH levels below 4.5. Hatchability of white sucker eggs, on
the other hand, dropped off sharply at pH levels 5.0 to 5.2. Number of
fathead minnow embryos hatching declined at pH 5.9. In experiments conducted
in Scandinavia and Great Britain, survival through hatching was reduced below
approximately pH 4.4 for sea-run brown trout and below pH 4.6 for roach.
Experiments with perch and Atlantic salmon yielded inconsistent results.
These pH values for effects on egg survival are distinctly higher than values
noted as acutely toxic to adults (pH 3.5 for brook trout; pH 3.8 to 3.9 for
white suckers; pH 3.8 for brown trout) (Section 5.6.4.1.1).
A number of studies have noted that the hatching process itself appears pH
sensitive (Runn et al. 1977; Peterson et al. 1980a,b; Baker 1981). For eggs
exposed to low pH either throughout their development or just during hatch-
ing, a large proportion of embryos hatch incompletely, with fry remaining
partially encapsulated for days following hatching. Delay or prevention of
hatching can be induced by transfer of eggs into low pH water just prior to
hatching, and normal hatching may occur if eggs are transferred just prior to
hatching from low pH water into control water. Thus, mechanisms involved in
the hatching process especially may be key factors limiting embryo survival
in low pH water (disintegration of the chorion, facilitating mechanical
rupture of the chorion by embryo trunk movements at hatching; Bell et al.
1969, Yamagami 1973, 1981). Mechanisms proposed involved: (1) the relation-
ship between pH and activity of the hatching enzyme (Yamagami 1973), (2)
thicker, more rigid egg capsules at lower pH, with increased resistance to
degradation (Runn et al. 1977, Peterson et al. 1980b), and (3) reduction in
body movements inside eggs at low pH (Peterson et al. 1980b).
5-114
-------
NORTH
AMERICAN
SPECIES
EUROPEAN
SPECIES
ATLANTIC
SALMON
4.0 4.5 5.0 5.5 6.0 6.5 7.0 7.5 8.0 8.5
PH
LEGEND
• BROOK TROUT
o FATHEAD MINNOW
» PERCH
• BROWN TROUT
a WHITE SUCKER
• ROACH
o ATLANTIC SALMON
Figure 5-13. Effect of low pH on survival of fish through hatching.
References:
a Baker and Schofleld 1982 g
b Swarts et al. 1978 h
c Trojnar 1977a 1
d Trojnar 1977b j
e Johansson et al. 1977 k
f Mount 1973
Carrlck 1979
Runn et al. (1n 1975) 1977
Johansson and Mil brink 1976
Peterson et al. 1980a
Peterson et al. 1980b
5-115
-------
Exposure of embryos to low pH levels during early stages of development
(particularly within the first day after fertilization or during water
hardening) also adversely affected survival, although to a lesser extent than
did exposure during hatching (Johansson et al. 1973, Johansson and Milbrink
1976, Daye and Garside 1977, Lee and Gerking 1981, Baker 1981). For roach
eggs exposed to pH 7.7 throughout their development, 89 percent hatched
successfully. After exposure to pH 4.7 for the first 24 hr and then to pH
7.7 from 24 hr to hatch, 52 percent hatched. With exposure to pH 7.7 for 24
hr followed by pH 4.7 to hatch, 20 percent hatched. Finally with exposure to
pH 4.7 throughout development, only 6 percent hatched successfully (Johansson
and Mil brink 1976).
The egg changes its character rapidly after being spawned. Permeability
decreases and the chorion hardens during the first few hours after release,
allowing the egg to become more resistant with time (Lee and Gerking 1981).
Zotkin (1965) noted that teleost eggs exchange water with the surrounding
solution primarily immediately after fertilization and just before hatching.
Exchange of water and ions between the egg and external medium during inter-
mediate periods of development occurs but is limited (Kalman 1959, Zotkin
1965).
Given the evidence that timing of exposure substantially affects the sensi-
tivity of embryos to low pH, it is obvious that to determine the impact of
acidification on embryo survival, the occurrence of particularly susceptible
stages must be evaluated in relation to the timing of fluctuations in pH
level in acidified surface waters. As with the toxicity of low pH to adult
fish, the effect of low pH on fish embryos was also found to be a function of
temperature (Kwain 1975).
At intermediate pH levels, between those recorded to have no consistent
adverse effect on embryo survival and pH levels that result in near 100
percent mortality, some researchers (Mount 1973, Runn et al. 1977, Trojnar
1977b) have observed increased incidence of deformities in larvae after
hatching. Runn et al. (1977) suggest that these malformations result, at
least in part, from the prolongation of the non-hatching period, Peterson et
al. (1980a), in contrast, reported no increase in deformities of Atlantic
salmon fry hatched at low pH levels (5.5 to 4.5).
Finally, pH may determine recruitment success for fish populations in acidic
waters by influencing the survival of young fish larvae (or fry) after
hatching. The direct effect of low pH on fry survival has been examined in
laboratory experiments. Fry survival in field situations would also be
strongly influenced by food availability, predation, temperature, and a large
number of other environmental factors. In general, survival of fry in labo-
ratory bioassays decreased below pH 4.0 to 4.5 for Atlantic salmon; pH 4.2 to
4.4 for brook trout; pH 4.8 for brown trout; pH 5.0 to 5.5 for white suckers;
and pH 5.2 for pike (Figures 5-14 and 5-15).
Evaluations of the relative sensitivities of eggs, sac fry (fish larvae after
hatching but prior to initiation of feeding and swim-up), and fry (after
initiation of feeding) have been inconsistent among experiments, perhaps
reflecting differences in species response. Baker and Schofield (1982)
5-116
-------
100
80
60
40
20 -
3.5
o d
f
LEGEND
/ • BROOK TROUT
/ » WHITE SUCKER
I /
• ATLANTIC SALMON
oBROWN TROUT
xPIKE
pH
Figure 5-14.
Effect of low pH on survival of fish as sac fry. Solid
line, sac fry survival through sw1m-up following
development of eggs and hatching of larvae 1n low pH water
(expressed as percent normal hatch); Dashed line, sac fry
survival without previous exposure to low pH.
PART (A)
a Baker and Schofield 1982
b Swarts et al. 1978
c Johansson et al. 1977
d Trojnar 1977b
PART (B)
a Daye and Garslde 1975
b Johansson and Klhlstrom 1975
c Johansson et al. 1977
5-117
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§
I—I
OS
O
c:
LEGEND
— BROOK TROUT
WHITE SUCKER
>7
PH
Figure 5-15.
Effect of pH on survival of fry exposed for 14 days
after swim-up and initiation of feeding.
fjBaker and Schofield 1982
"Trojnar 1977a; previous exposure during development at
pH 8.0 (o); previous exposure at pH 4.6 to 5.6 (•).
5-118
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and Swarts et a!. (1978) found in successive experiments with brook trout
and/or white sucker that sensitivity to low pH decreased with age. Also, a
high proportion (> 75 percent) of embryos alive at hatching survived through
swim-up with continued exposure to the same low pH level (Trojnar 1977a,
Craig and Baksi 1977, Baker and Schofield 1982). Daye and Garside (1977,
1979), on the other hand, concluded that Atlantic salmon fry were more sensi-
tive to low pH than were eggs. Likewise, Johansson et al. (1977) observed
that Atlantic salmon and brown trout (and to a lesser extent brook trout)
that survived through hatching at low pH levels (pH 4.1 to 5.0) subsequently
suffered substantial mortality (10 to 100 percent) during the four weeks
after hatching until just prior to full resorption of the yolk sac.
Therefore, while some researchers have concluded that fry are relatively (as
compared with fish eggs) tolerant of low pH, other researchers considered fry
to be a particularly sensitive stage in the reproductive cycle of fish.
Because as fry emerge from the nest, redd," or spawning tributary upon swim-
up they may be subjected to an environment and water quality distinctly
different from that to which the eggs (and sac fry) were previously exposed,
an understanding of these relative tolerances is important.
5.6.4.1.3 Growth. The direct effect of low pH on fish growth has been
examined in several laboratory experiments. Although field observations of
changes in growth with acidification indicate a variable response to in-
creased acidity (Section 5.6.2.3), reflecting the large number of variables
determining growth in natural situations, in the laboratory low pH has
consistently resulted in decreased growth. These decreases in growth often
occur at pH levels above those producing substantial fish mortality. Edwards
and Hjeldnes (1977) observed a significant (p < 0.001) decrease in growth
(relative to the control at pH 6.0) of yearling rainbow trout, brown trout,
and arctic char held at pH 4.8 for 3.5 months; mortality levels were less
than 10 percent. Jacobsen (1977) found no significant decrease in growth of
18 month old brown trout after 48 days, but tested pH levels only down to
5.0. Swarts et al. (1978) and Baker (1981) noted delayed development of
brook trout sac fry hatched at pH 4.6 and below. For brook trout embryos
reared at pH 6.5, 6.0 and 5.5, fry were significantly (p < 0.05) shorter
after 3 months than were fry in control water at pH 7.1 (Menendez 1976).
Likewise, flagfish surviving through embryo development and 45 days after
hatching weighed significantly less at pH 6.0, 5.5, and 5.0 than did fry at
pH 6.8 (Craig and Baksi 1977) and rainbow trout reared at pH 4.3 to 4.8 were
shorter (p < 0.001) than controls at pH 7.1 to 7.3 (Nelson 1982).
The decrease in growth at low pH represents a sublethal response to elevated
hydrogen ion concentrations and suggests that fish are physiologically
stressed at pH levels above those that produce acute or chronic mortality.
5.6.4.1.4 Behavior. Behavioral responses of fish to low pH probably play an
important role in determining the effect of surface water acidification on
fish populations. Within a given aquatic system at any time, water quality
may vary substantially (Driscoll 1980). If fish can detect regions of low pH
and by behavioral adaptation avoid exposure to these toxic conditions, the
impact of acidification may be, in part, mitigated. Muniz and Leivestad
(1980a) reported observations of trout concentrated into "refuge areas"
5-119
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during acid incidents. In the acidic river Gjor in Norway during spring snow
melt, hundreds of brown trout from the river crowded into a tiny tributary
with a higher pH. If experimentally restrained within the river, fish died
within a week. Information on the presence of such "refuge" areas and on the
ability of fish to detect and use these areas is necessary for a complete
assessment of the impact of acidification.
Unfortunately, laboratory (and field) data on behavioral responses of fish to
low pH are very limited. Jones (1948) tested sticklebacks (Gasterosteus
aculeatus) in a sharp concentration gradient in a laboratory apparatus.FTsTT
were able to detect and avoid waters with pH £ 5.4, a value slightly above
the lethal level of pH 5.0. Hoglund (1961) concluded that Atlantic salmon
fingerlings avoid water at pH 5.3 and below, roach at pH 5.6 and below.
Johnson and Webster (1977) investigated the effect of low pH on spawning site
selection of brook trout. Female trout clearly avoided areas of water
upwelling at pH 4.0 and 4.5. Discrimination was not evident at pH 5.0.
Preference by adult brook trout for spawning in areas receiving neutral or
alkaline aquifer water may protect eggs and sac fry from adverse water
quality conditions. Decreased spawning activity at low pH (discussed in
Section 5.6.4.1.2) may therefore partially reflect a behavioral response
rather than an adverse effect on reproductive physiology.
5.6.4.1.5 Physiological responses. In the laboratory a decrease in pH level
has been demonstrated to result in a wide diversity of physiological
responses in fish. Some of these observed responses may reflect only a
general response of fish to stress; others appear to be specifically related
to low pH. The following does not represent a complete review of the
extensive and varied literature available on fish responses to acidity; only
major topics are summarized. Fromm (1980) and Wood and McDonald (1982) have
provided a thorough critique of the literature on physiological and
toxicological responses of freshwater fish to acid stress.
The best documented physiological response, and probably the response most
widely accepted as the physiological basis for the toxicity of low pH,
involves interference of elevated hydrogen ion levels with osmoregulatory
mechanisms and impaired body salt regulation. Freshwater fish maintain a
higher salt concentration in their tissues than is in the water that
surrounds them, and must actively take up ions from the surrounding water
through the gill epithelium. Sodium in the water is exchanged for hydrogen
ions or ammonium ions, and'chloride for bicarbonate (Maetz 1973, Evans 1975).
Increased hydrogen ion activity in the surrounding medium may impede the
active uptake of sodium. Brown trout surviving in the Tovdal River, Norway,
collected immediately following a fish kill (apparently resulting from an
acid episode), had significantly reduced plasma chloride and sodium levels
(Leivestad and Muniz 1976, Section 5.6.2.4). The plasma content of
potassium, calcium and magnesium was not affected. Therefore, impairment of
the active transport mechanism for sodium and/or chloride ions through the
gill epithelium was suggested as the primary cause of fish death. Severe
internal ionic imbalance would affect fundamental physiological processes
such as nervous conductions and enzymatic reactions.
5-120
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Laboratory experiments have also found decreased plasma (or whole body)
sodium and/or chloride levels as a result of exposure of organisms to low pH
levels (Packer and Dunson 1970, 1972; Leivestad and Muniz 1976; McWilliams
and Potts 1978; Jozuka and Adachi 1979; Leivestad et al. 1980; McWilliams et
al. 1980; McDonald et al. 1980; McDonald and Wood 1982; Ultsch et al. 1981).
The exact mechanisms behind these effects are not, however, fully understood.
A major influence on branchial ion fluxes is the transepithelial potential
(TEP) across the gills. The TEP of brown trout has been shown to be strongly
dependent on the pH of the external medium, being negative in neutral solu-
tions but positive in acid solutions (McWilliams and Potts 1978). At near
neutral pH, the influx and efflux of sodium were similar, indicating that
trout were in sodium balance. As the pH in the external medium declined,
sodium influx decreased and sodium efflux increased until, at pH 4.0, the
rate of loss of sodium amounted to about 1 percent of the total body sodium
per hr.
These processes are influenced by the content of dissolved salts in the
water, particularly calcium and sodium (McDonald et al. 1980; Brown 1981,
1982). Calcium is essential in the maintenance of ionic balance in fresh-
water fish, probably as a result of its influence on the permeability of
gills to certain ions (McWilliams and Potts 1978, McWilliams 1980a).
Increased calcium concentrations (from near zero to about 40 mg &-1)
decreased membrane permeability and thus decreased the rate of passive sodium
efflux from fish. At the same time, calcium appeared to have no significant
effect on sodium influx (McWilliams 1980a, 1982). The result was a decrease
in the overall rate of sodium loss from fish exposed to low pH in waters with
higher calcium content. Gill permeability also varied between species and
populations of fish (McWilliams 1982), and sodium loss rates declined with
acclimation of fish to acid waters (McWilliams 1980b). These results help
explain the observed correlation between low calcium levels and loss of fish
populations in Norwegian Lakes (Section 5.6.2.1.3; Wright and Snekvik 1978)
and imply that small changes in calcium availability in natural waters (e.g.,
during spring snowmelt; see Chapter E-4, Section 4.4.2) and previous exposure
of fish to high acidity are crucial factors in determining the response of
fish exposed to sudden acid episodes.
A decrease in blood pH levels (by 0.2 to 0.5 pH units) is often associated
with the drop in plasma sodium levels in fish exposed to low pH waters (Lloyd
and Jordan 1964, Packer and Dunson 1970, Packer 1979, Jozuka and Adachi 1979,
Neville 1979a, McDonald et al. 1980, McDonald and Wood 1982, Ultsch et al.
1981) and is possibly a result of hydrogen ion flux across gill membranes
into the blood. McDonald et al. (1980) noted that in moderately high alka-
linity waters (calcium 30 to 50 mg £-!), fish exposed to a pH of 4.3
developed a major blood acidosis (drop in blood pH) but exhibited only a
minor depression in plasma ion levels. In acidified, low alkalinity water
(calcium 6 mg £~M, only a minor acidosis occurred, but plasma ion levels
fell and mortality was substantially greater. Possibly the nature of the
mechanism of acid toxicity varies with the nature of the ionic environment.
A drop in blood pH level would affect a large number of pH-sensitive metabol-
ic reactions. The oxygen-carrying capacity of fish blood drops sharply below
a blood pH level of 7.0 (Green and Root 1933, Prosser and Brown 1961).
5-121
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Decreased oxygen consumption by fish exposed to acid waters has been found by
Packer and Dunson (1970, 1972), Packer (1979), and Ultsch (1978). Carrick
(1981), however, observed no significant differences in oxygen uptake by
brown trout fry at pH 7.0 vs pH 4.0. Neville (1979b) concluded that an
observed increase in serum erythrocyte concentration offset the reduced
capacity of the hemoglobin to transport oxygen brought about by acidosis.
The increase in hemoglobin level, maintenance of arterial oxygen tension
(p02), and constancy of blood lactate levels in rainbow trout exposed to pH
4.0 suggested that there was no oxygen stress despite the acidosis.
At critically low pH levels (<_ 3.5), where death occurs within hours rather
than days, a failure of oxygen delivery to the tissues may be of primary
importance. A marked reduction in blood oxygen capacity due to massive
acidosis, combined with impaired branchial oxygen diffusion as a result of
accumulation of mucous on the gills and a sloughing of gill epithelial tissue
(e.g., Plonka and Neff 1969, Daye and Garside 1976, Ultsch and Gros 1979),
may result in eventual cellular anoxia. However, such low pH levels are
rarely encountered by fish in natural situations. At more moderate pH
levels, mucous accumulation on the gills has not been observed and blood gas
levels remain normal, indicating acid-base and/or ion regulatory failure are
more probable mechanisms of toxicity (McDonald et al. 1980, Fromm 1980).
Finally, Nelson (1982) reported that ossification, amount of calcium deposit-
ed in bone, in rainbow trout fry varied significantly (p < 0.005) as a
function of pH of the medium (pH 4.3, 4.8, and 7.3). After calcium stores
from the yolk sac are exhausted, fry must accumulate calcium from the sur-
rounding water and from food intake. A decrease in skeletal ossification at
low pH level could be partially responsible for increased incidence of skele-
tal deformities observed in some laboratory bioassays at low pH (e.g.,
Beamish 1972, Mount 1973, Trojnar 1977b) and in white suckers from acidic
George Lake, LaCloche Mountain region, Ontario (Beamish et al. 1975). Nelson
(1982), however, noted no increase in deformities despite decreased
ossification.
5.6.4.2 Effects of Aluminum and Other Metals in Acidic Waters—Increases in
certain metal concentrations can be associated with decreasing pH levels in
acidified surface waters (Chapter E-4, Section 4.6.2). Declines in fish
populations as a result of acidification may, therefore, be a function of
both low pH levels and elevated concentrations of some metals. Critical
values for survival of fish populations developed only on the basis of pH
level may therefore be misleading.
Muniz and Leivestad (1980a) noted that naturally acidified water is generally
more toxic to fish than are dilute sulfuric acid solutions of the same pH.
Brown trout exposed to soft waters acidified by additions of sulfuric acid (a
pure hydrogen ion stress) demonstrated physiological stress (impaired regula-
tion of body salts) only at pH levels below 4.6 (Leivestad et al. 1980).
When tests were performed in water from acidified brooks and rivers in south-
ern Norway, water with a pH of 4.6 resulted in significant physiological
stress, rapid salt depletion, and mortality after 11 days (Leivestad et al.
1976; Section 5.6.3.3). For Atlantic salmon, Daye and Garside (1977) found
lower limits for survival of fry to be around pH 4.3 and pH 3.9 for eggs
5-122
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exposed from fertilization through hatching (Section 5.6.4.1.2). Bua and
Snekvik (1972) , on the other hand, used water from the acidic Mandal River,
Norway and found lower limits for survival to be pH 5.0 to 5.5. Schofield
observed stress and heavy mortality among adult, yearling, and sac fry of
brook trout held in an Adirondack hatchery receiving lake water from Little
Moose Lake at pH 5.9 during spring snowmelt in 1977 (Schofield and Trojnar
1980; Section 5.6.2.4). In contrast, in laboratory experiments (Sections
5.6.4.1.1 and 5.6.4.1.2) critical pH levels for brook trout were generally
between pH 3.5 and 4.5. These and other comparisons strongly imply that
acidified lake and river water must contain toxic agents in addition to
hydrogen ions (Muniz and Leivestad 1980a).
Metals consistently exhibiting increased concentrations in acidic surface
waters, apparently as a result of increased solubility with decreasing pH
level, are aluminum, manganese, and zinc (Chapter E-4, Section 4.6.1.2). In
some regions, concentrations of cadmium, copper, lead, nickel, and other
metals are also elevated in acidic lakes. High concentrations of these
metals, however, probably result primarily from increased atmospheric loading
and deposition and occur principally in surface waters in close proximity to
pollutant sources (e.g., Sudbury, Ontario). As such, they are not specifi-
cally a result of acidic deposition but may still interact additively or
synergistically with toxic effects of low pH, aluminum, manganese, or zinc.
Unfortunately, with the exception of aluminum, data are not sufficient for a
thorough evaluation of possible adverse effects of metals on fish in acidic
waters. Spry et al. (1981) and Baker (1982) have reviewed the available
literature.
Total zinc concentrations measured in acidic surface waters in the Adirondack
region, in southern Norway and in southwestern Sweden ranged up to 0.056 mg
rl (Schofield 1976c, Henriksen and Wright 1978, Dickson 1980). Although
laboratory bioassays examining effects of zinc on fish are numerous (Taylor
et al. 1982), none of these studies considered low alkalinity water with pH
levels below 6.0, and results should not be automatically extrapolated to
conditions in acidified surface waters. For the most part, however, lethal
concentrations of zinc in bioassays are 10 times zinc concentrations found in
acidic waters (Spry et alI. 1981, Taylor et al. 1982). Sinley et al. (1974)
estimated that the maximum acceptable toxicant concentration (MATC) for
rainbow trout exposed to zinc in low alkalinity circumneutral water was
between 0.14 and 0.26 mg £-1. Benoit and Hoi combe (1978) determined that
the threshold level for significant adverse effects on the most sensitive
life history stage of fathead minnows was between 0.078 and 0.145 mg £-1.
Taylor et al. (1982) concluded from a review of the available literature that
concentrations of zinc that could be tolerated by aquatic organisms lie
between 0.026 and 0.076 mg £-1.
Manganese has been considered a relatively nontoxic element; thus
toxicological data are very limited. Total manganese concentrations measured
in acidic surface waters ranged up to 0.13 mg £-1 in the Adirondacks
(Schofield 1976b) and up to 0.35 mg £-1 in southwestern Sweden (Dickson
1975). Lewis (1976) determined that manganese concentrations up to 0.77 mg
5-123
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£-1 had no effect on survival of rainbow trout in soft waters with pH
levels of 6.9 to 7.6.
Relationships between pH and levels of cadmium, copper, lead, and nickel vary
markedly between regions. Excluding lakes within 50 krn of Sudbury, concen-
trations of cadmium, copper, lead, and nickel measured in acidic Ontario
surface waters ranged up to about 0.6, 9, 6, and 48 yg jr1, respec-
tively (Spry et al. 1981). In southwestern Sweden, concentrations of cadmium
in acidic waters were measured up to 0.3 yg «,-!; lead up to 5 yg
r* (Dickson 1980). Spry et al. (1981) reviewed available bioassay data
and noted no significant adverse effects on survival and reproduction at
concentrations up to 0.7 to 11 yg Cd
-------
(Muniz and Leivestad 1980a). Muniz and Leivestad (1980b) observed rapid loss
of sodium and chloride from the blood of brown trout exposed to aluminum
concentrations as low as 0.19 mg i-1 at pH 5.0. Schofield and Trojnar
(1980) noted moderate to severe gill damage at aluminum levels of 0.5 and 1.0
mg jr1 at pH 4.4 and higher. Aluminum was particularly toxic in over-
saturated solutions at pH levels 5.2 to 5.4 (Baker and Schofield 1982).
The pH level in acidic waters, therefore, affects fish survival both as a
direct toxicant and by controlling the concentration of inorganic aluminum.
5.6.5 Summary
5.6.5.1 Extent of Impact
Loss of fish populations associated with acidification of surface waters has
been documented for five areas—the Adirondack region of New York State, the
LaCloche Mountain region of Ontario, Nova Scotia, southern Norway, and
southern Sweden. The following summarizes major points from Section 5.6.2.1:
o The best evidence that loss of fish has occurred in response to
acidification is derived from observations of lakes in the LaCloche
Mountain region, Ontario (Section 5.6.2.1.2.1). Twenty-four percent
of 68 lakes surveyed had no fish present. Fifty-six percent of the
68 lakes are known or suspected to have had reductions in fish
species composition (Harvey 1975). Based on historic fisheries
information, 54 fish populations are known to have been lost, includ-
ing lake trout populations from 17 lakes, small mouth bass from 12
lakes, largemouth bass from four lakes, walleye from four lakes, and
yellow perch and rock bass from two lakes each (Harvey and Lee 1982).
The principal source of atmospheric acidic inputs to the LaCloche
area is sulfur dioxide emitted from the Sudbury smelters located
about 65 km to the northeast. Large acidic inputs have resulted in
relatively rapid acidification of many of the region's lakes. For
some lakes the development of increased lake acidity and the simulta-
neous decline of fish populations have been followed and recorded by
a single group of researchers (Beamish and Harvey 1972, Beamish et
al. 1975, Harvey and Lee 1982) from the mid-1960's to the present.
0 In Norway (Section 5.6.2.1.3.1), sharp drops in catch of Atlantic
salmon in southern rivers began in the early 1900's and are associ-
ated with current low pH levels and a recorded doubling of the
hydrogen ion concentration in one of these rivers from 1966 to 1976
(Jensen and Snekvik 1972, Leivestad et al. 1976). For almost 3000
lakes in Stfrlandet (southernmost Norway) data on the status of
brown trout have been recorded since about 1940 (Sevaldrud et al.
1980). Today, more than 50 percent of the original populations have
been lost, and approximately 60 percent of the remaining are in rapid
decline (Sevaldrud et al. 1980). Fish population declines have been
correlated with acidity, acidification and/or inputs of acidic
deposition (Wright and Snekvik 1978).
5-125
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0 Extensive surveys of fish population status and acidity of surface
waters in Sweden have not been completed (Section 5.6.2.1.3.2).
However, for 100 lakes in southern Sweden with historic records on
fish populations, loss of fish was correlated with present-day low pH
levels in lakes (Aimer et al. 1978). Forty-three percent of the
minnow populations, 32 percent of the roach, 19 percent of the char,
and 14 percent of the brown trout populations had disappeared.
0 In Nova Scotia, records of angling catch of Atlantic salmon in rivers
date back, in some cases, to the early 1900's (Section 5.6.2.1.2.3).
Of 10 rivers with current pH < 5.0 and historic catch records, 9 have
had significant declines in angling success over the time period 1936
to 1980. For 12 rivers with pH > 5.0, only one experienced a signif-
icant decrease in salmon catch. Decrease in salmon catch over time
is correlated with present-day pH values 5.0 and below. In addition,
6 former salmon rivers with current pH < 4.7 have no long-term catch
records, but surveys in 1980 indicated they no longer support salmon
runs. Acidification of rivers in the area between 1954 and 1974 has
been reported (Chapter E-4, Section 4.4.3.1.2.2). The high organic
content in many of the low pH waters (especially pH < 4.7) suggests,
however, that these rivers are naturally somewhat acidic, and perhaps
always had fairly low pH values and low fish production (Farmer et
al. 1981). The estimated lost or threatened Atlantic salmon produc-
tion potential represents 30 percent of the Nova Scotia resources but
2 percent of the total Canadian potential (Watt 1981).
0 Finally, fish populations in Adirondack lakes and streams have also
declined over the last 40 to 50 years (Section 5.6.2.1.1.1). The New
York State Department of Environmental Conservation reports that
about 180 lakes (2900 ha) out of a total of 2877 lakes (114,000 ha)
in the Adirondacks have lost their fish populations (especially brook
trout) (Pfeiffer and Festa 1980). The absence of fish in Adirondack
lakes and streams is clearly correlated with low pH levels (Schofield
1976c), although several factors may confound this relationship,
e.g., lake size, dystrophic conditions. Records of pH and other
information have not, however, been published to substantiate that
loss of fish in these 180 lakes resulted from acidification. For
very few individual lakes are historical data available that suggest
both lake acidification and simultaneous loss of fish. Acidification
probably contributed to the disappearance of fish for at least some
surface waters, but exactly how many lakes and streams (perhaps
substantially less than or more than 180) have been impacted cannot
be satisfactorily evaluated at this time.
0 In other regions of the world with low alkalinity waters and receiv-
ing acidic deposition (e.g., Muskoka-Haliburton area of Ontario and
Maine; Sections 5.6.2.1.1.2 and 5.6.2.1.2.2), acidification of
surface waters does not appear to have progressed to levels clearly
detrimental to fish (Schofield 1982). No damage to fish populations
has been reported.
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5.6.5.2 Mechanism of Effect—Three major mechanisms for the disappearance of
fish populations with acidification have been proposed: (1) decreased food
availability and/or quality; (2) fish kills during episodic acidification;
and (3) recruitment failure. Each probably plays some role, although
recruitment failure has been hypothesized as the most common cause of
population loss (Schofield 1976a, Harvey 1980, NRCC 1981, Overrein et al.
1980, Haines 1981b). The following summarizes major points from Sections
5.6.2.2 through 5.6.2.4, and 5.6.3.1:
o The influence of food chain effects on decreases in fish populations
in acidified waters has received little attention to date, but avail-
able information suggests it plays a relatively minor role (Beamish
1974b, Hendrey and Wright 1976, Mum'z and Leivestad 1980a, Rosseland
et al. 1980). With acidification, or in comparisons between acidic
and circumneutral lakes, fish growth is often unaffected or increased
with increasing acidity (Section 5.6.2.3). Some important prey orga-
nisms are sensitive to acidic conditions and disappear with acidifi-
cation yet fish seem capable of shifting to other suitable prey.
During the experimental acidification of Lake 223 (Section 5.6.3.1)
lake trout production remained unchanged in spite of the loss of
fathead minnows, a major prey item prior to acidification (Mills
1984). Few studies, however, have examined the potential effect of
reduced food quantity and/or quality on survival of early life
history stages of fish or on fish production at pH levels above those
that result in recruitment failures and reduced population size.
0 Fish kills have been observed during episodic acidification of
surface waters (Section 5.6.2.4) and in certain instances may play an
important role in the disappearance of fish from acidified surface
waters. For example, in the Tovdal River, Norway, in 1975 thousands
of dead adult trout were observed in association with the first major
snowmelt in spring (Leivestad et al. 1976). Dead and dying fish are,
however, seldom reported in acid-stressed waters relative to the
large number of lakes, streams, and rivers with fish populations
apparently impacted by acidification. In contrast, a substantial
portion of fish populations examined in acidified lakes lack young
fish (Section 5.6.2.2) and apparently have experienced recruitment
failure.
° Recruitment failure may result either from acid-induced mortality of
eggs and/or larvae or because of a reduction in numbers of eggs
spawned. The number of eggs spawned could be reduced as a result of
disruption of reproductive physiology and ovarian maturation or
inhibition of spawning behavior. Evidence exists that supports each
one of these proposed mechanisms (Sections 5.6.2.2, 5.6.4.1.2,
5.6.4.1.4, and 5.6.4.1.5). For example, Johnson and Webster (1977)
demonstrated experimentally that brook trout avoid spawning in waters
with pH below 5.0. Beamish and Harvey (1972) observed that recruit-
ment failure in several acidic lakes in the LaCloche Mountain region,
Ontario was associated with a failure of the female fish to spawn.
5-127
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Lockhart and Lutz (1977) hypothesized that a disruption in normal
calcium metabolism, induced by low pH, had adversely affected the
reproductive physiology of female fish in these lakes. Adverse
effects of low pH levels and elevated aluminum concentrations on
survival of fish eggs and larvae have been demonstrated in numerous
laboratory and field experiments (Sections 5.6.3.3 and 5.6.4.1.2).
In Norway, total mortality of naturally spawned trout eggs was
observed in an acidic stream a few weeks after spawning (Leivestad et
al. 1976).
0 It is likely that each one of these factors plays some role in re-
cruitment failure but the importance of each factor probably varies
substantially among aquatic systems, depending on the particular
circumstances.
o More research is necessary to define clearly the specific mechanism
for population decline in a given water. However, many studies in
the United States and Scandinavia (Schofield 1976a, Muniz and
Leivestad 1980a) emphasize increased mortality of eggs and larvae in
acidic waters as the primary cause of recruitment failures, and
recruitment failure as a common cause for the loss of fish popula-
tions with acidification of surface waters.
5.6.5.3 Relationship Between Water Acidity and Fish Population Response—To
assess tfieimpact of acidification on fish resources quantitatively, the
functional relationship between acidification and fish population response
must be understood. Unfortunately, loss of fish populations from acidified
surface waters is not a simple process and cannot be accurately summarized as
"X" pH (or aluminum concentration) yields "Y" response. The mechanism by
which fish are lost (Section 5.6.5.2) seems to vary between aquatic systems
and probably within a given system from year-to-year.
The water chemistry within a given aquatic system is likewise extremely
variable both spatially and temporally, and these variations are very
important to the survival or decline of fish populations. Lakes with
seemingly identical water quality may show marked differences in fish
response, perhaps reflecting, in part, the existence or lack of water quality
"refuge" areas for fish survival (Muniz and Leivestad 1980a). A circum-
neutral tributary or small segment of a lake may provide an area for success-
ful fish reproduction for a number of years following acidification of the
main body of a lake.
Fish species differ not only in their ability to tolerate acidic conditions
but also in their ability to exploit these chemical variations in their
environment (e.g., spawning time and location). Within a given fish species,
sensitivity to acidity varies with life history stage, age, condition,
previous exposures to acidity, associated water quality conditions (e.g.,
aluminum and calcium concentrations, temperature), and other parameters. In
addition, for reasons discussed at the beginning of Section 5.6.4, results
from laboratory experiments cannot be translated automatically into an
expected response in the field. Serious gaps exist in the understanding of
how to use laboratory results in a quantitative prediction of fish response
5-128
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in the field and in the analysis of the complexity of the natural environ-
ment, and the significance of this complexity in determining the impact of
acidification. It is therefore not surprising that development of an
accurate functional relationship between acidification and fish response is
impossible at this time.
First steps, however, in developing such a relationship are to: examine in a
semi-quantitative manner all of the available information connecting acidity
and fish (summarized in Table 5-14), produce an initial approximation of the
dose-response relationships (Figure 5-16), andtnen assesspatterns and
reasons for deviations from this initial approximation. In large part, the
analysis of deviations and variations must be done on a lake-by-lake,
population-by-population basis, and is the subject for further research.
Several points are, however, obvious.
Acidification adversely affects fish populations. Sensitivity of fish to
acidity is species-dependent and determined by aluminum and calcium concen-
trations, in addition to pH values. Loss of fish populations need not be
associated with large declines in annual average pH, but could result from
indirect effects on aluminum chemistry or episodic acidification.
5.7 OTHER RELATED BIOTA (R. Singer and K. L. Fischer)
5.7.1 Amphibians
Direct effects of acidity on vertebrates have been demonstrated only on fish
(Section 5.6) and amphibians. Amphibians are particularly sensitive because
many frogs, toads, and salamanders breed in vernal pools filled by acidic
snowmelt and spring rains. The salamanders Ambystoma maculaturn and A.
jeffersonianum breed in shallow woodland or meltwater ponds that have "pH
values l.b pH units less than nearby permanent ponds in New York State (Rough
and Wilson 1977). Spotted salamander (A. maculaturn) egg mortality increased
to > 60 percent in water less than pH 6.~Q, a substantial rise from the normal
mortality of < 1 percent at pH 7.0. In contrast, the Jefferson salamander,
t(. jeffersonianum, ^as most successful at pH 5.0 to 6.0 (Rough 1976). The
preference for neutrakwater by adult spotted salamanders may be a result of
the absence of their preferred prey, the stickleback (Eucalia), from acidic
water (Bishop 1941). Whe^i a stretch of stream was artificially acidified to
pH 4.0, "salamanders" weVe reported to leave the water (Hall and Likens
1980a). Elsewhere in its ^ange in central Ontario, the number of egg masses
of the spotted salamander positively correlated with pH (Clark and Euler
1980). Adults are not as sensitive to pH stress, but given a choice, adult
spotted salamanders U. maculaturn) preferred neutral substrates (Mushinsky
and Brodie 1975).
The mechanism by which acidity affects amphibians is not known. Huckabee et
al. (1975) suggest that the aluminum, manganese, and zinc mobilized by low pH
(Chapter E-4, Section 4.6) may be toxic agents for the shovel-nosed
salamander (Leurognathus marmoratus) larvae in the Great Smoky Mountains
National Park. Another mechanism may be their inability to control ion
fluxes across membranes against strong H+ gradients. This has been
5-129
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TABLE 5-14. SUMMARY OF FIELD OBSERVATIONS, FIELD EXPERIMENTS, AND LABORATORY EXPERIMENTS
RELATING HATER pH TO FISH RESPONSE
in
i
i— >
co
O
FIELD OBSERVATIONS
RecrutMent failure
Population loss
Fish ktll
FIELD EXPERIMENTS
Recrultaent failure
Population extinction
Adult cwrtallty
Ert>ryo MrUllty
LABORATORY EXPERIMENTS
Adult nortallty-
acute (2 day)
Adult mortal! ty-
chronlc (> 20 day)
Embryo Mortality
Fry aortal Ity
Reduced production
of viable eggs
Reduced growth
Avoidance
FIELD OBSERVATIONS
Recruitment failure
Population loss
Brook Lake Arctic
Trout Trout Char
5.0-5.5* 5.2-5.5" 5.2C
6.9e
5.7'
5.0* 5.0-5.5*
4. 5-4. 8 J 4.4e
5.9"
5.1-5.3*
4.8-5.2"
4.7-5.1*
4.4-4.6t
4.5-4.6*
3.8"
3.5°"
3.6«
4.4'' < 4.89g
•Ml
. ., J
.1"'
.4dd
< .6PP
<5.0kk
•4.4-4.9"
•4.5-48JJ
6.1*'
4.2«1
<5.4"
4.0-5. Oft
5.1"
6.5" 4.899
4.5«»
Lake Herring Lake SMllnoutll
UMteflsh Bass
4.5-4.70 5.0h 5.5-6.0"
5.0«
4.4« 4.4« 4.4*
6.0*
Brown Rainbow
Trout Trout
< 5.0C
4.7-5.1'
4.5-4.8J 5.5-6.0J
4.5-5.0'
5.1h
4.9-5.1"
4.6"
5.0k
4.0-5.0r
4.50
4.8k
5.1"
3.8» 4.01
3.8-4.8<:<:
4.0-4.2««
< 4.899 < 5.099
4.1"
4.0-4.5""
4.4"
4.899 4.899
< 5.0" 4.3-4. BUU
Largenouth Rock
Bass Bass
5.1« 4.7-5.2°
5.01 5.0*
4.8-5.0"
4.4' 4.3«
Atlantic White European Walleye Fathead Roach
Salmon Sucker Perch Minnow
4.7-5.0"1 5.2* 4.4-4.9c 5.5-6.0" 5.5C
4.7-5.2" 5.0-5.59 5.4e s.lh
5.0* <4.7h 6.5'
5.1k 5.1* 5.2« 4.7"
4.3e 5.5'
3.9-4.2°
5.0P
S.1-5.3Q 5.8-6.01
5.3-5.8Q
4.5-5.01 4.7-5.7U 5.4» 5.7U
5.0-5.5°
4.9k
3.9**
c 4.6hh
4.1'k *> 5.6JJ 5.09 5 9hh 5 011
5.5"" 5.2°° 5.611
5.5""
4.511
3.7-4.0rr '5.4-5.6JJ < 5 9"h
5.0k* 5.0-5.40°
6.6hh
5.3" 5.3«
Pumpklnseed Blueglll v.llow Comion aluntnose Lake
Sunflsh Perch Shiner Minnow Chub
5.01 4.5-4.7" 5.5-6.0* 4.5-4.7"
4.4« 5.7«
4.4-5.oyy
<-3« 4.4« 4.3' 5.7« 4.5-5.0*
Northern SI 1«ty Brown
Pike Sculpln Bullhead
5.0f 4.7-5.2"
5.0' 4.9e
4.4-4 .9' 5 21
4.7h
«-3« 4.7-5.0*
4.7e
5.3-5.81
4.2-5.2"
Creek Trout Burbot
Chub Perch
5.2-5.5" 5.5-6.0"
5.0* 5.4'
-------
REFERENCES
a Schofield 1976c aa
b Beamish 1976 bb
c Aimer et al. 1978 cc
d Watt et al. 1983 dd
e Harvey 1979 ee
f Hultberg 1977 ff
g Runn et al. 1977 gg
h Grahn et al. 1974 hh
i Beamish et al. 1975 ii
j Grande et al. 1978 jj
k Leivestad et al. 1976 kk
1 Harriman and Morrison 1982 11
m Overrein et al. 1980 mm
n Schofield and Trojnar 1980 nn
o Jensen and Snevik 1972 oo
p Farmer et al. 1981 pp
q Mills 1984 qq
r Harvey et al. 1982 rr
s Schofield 1965 ss
t Dunson and Martin 1973 tt
u Mil brink and Johansson 1975 uu
v Hulsman and Powles 1981 as vv
reported in U.S./Canada MOI 1982 ww
w Muniz and Leivestad 1980b xx
x D. W. Johnson 1975 yy
y Brown 1981
z Kwain 1975
Beamish 1972
Robinson et al. 1976
McDonald et al. 1980
Swarts et al. 1978
Lloyd and Jordan 1964
Swarts et al. 1978
Edwards and Hjeldnes 1977
Mount 1973
Menendez 1976
Baker and Schofield 1982
Johansson et al. 1977
Johansson and Milbank 1976
Carrick 1979
Peterson et al. 1980a
Trojnar 1977b
Trojnar 1977 a
Peterson et alI. 1980b
Daye and Garside 1976
Johansson and Kihlstrom 1975
Jacobsen 1977
Nelson 1982
Johnson and Webster 1977
Hoglund 1961
Ryan and Harvey 1977
Ryan and Harvey 1980
*Refers to laboratory experiments taking into account both low pH and inorganic
aluminum (at the expected concentration for that pH based on Driscoll 1980).
5-131
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SPECIES
YELLOW PERCH
NORTHERN PIKE
ROCK BASS
PUMPKINSEED SUNFISH
LAKE HERRING
LAKE WHITEFISH
BLUEGILL
LAKE CHUB
EUROPEAN PERCH
WHITE SUCKER
LARGEMOUTH BASS
BROOK TROUT
BROWN TROUT
SMALLMOUTH BASS
BROWN BULLHEAD
ATLANTIC SALMON
ROACH
LAKE TROUT
CREEK CHUB
RAINBOW TROUT
ARCTIC CHAR
SLIMY SCULPIN
TROUT-PERCH
BURBOT
WALLEYE
FATHEAD MINNOW
COMMON SHINER
BLUNTNOSE MINNOW
4.5 5.0 5.5
LEGEND RH
pH RANGE OF SUCCESSFUL REPRODUCTION
pH LEVELS AT WHICH POPULATIONS OCCUR
VARIATIONS IN OBSERVED LOWER pH LIMITS
6.0
15
6.5
Figure 5-16.
Initial estimates of relationship between acidity and
fish response, based on references in Table 5-14.
5-132
-------
indicated in fish (Section 5.6.4.1.5), invertebrates (Sections 5.3 and 5.5),
and frogs (Fromm 1981).
The species-specific tolerance of amphibians to low pH was confirmed by a
survey of newts in England (Cooke and Frazer 1976). Smooth newts (Triturus
vulgaris) were rarely encountered in water with pH < 6.0, but the palmate
newt (T_. helveticus) was routinely captured in bogs at pH 4.0 to 3.8. The
distributions of these species were correlated most strongly with potassium
and calcium concentrations, both of which co-varied with pH. The variable
sensitivity of newts to acid stress is demonstrated by the American red-
spotted newt, No to p h th alamu s viridescens, which one of us (RS) has observed
at 6 m in acidfc (pH 4.9) Woods Lake. TTiis same species has been reported at
13 m in neutral (pH 7.4) Lake George, also in the Adirondacks (George et al.
1977).
Many anurans are also sensitive to acidity. Calling densities (an estimate
of population size) of spring peepers (Hyla crucifer) were positively corre-
lated with the pH of water in which they occurred (Clark and Euler 1980).
Bullfrogs (Rana catesbeiana) (Clark and Euler 1980, Cecil and Just 1979,
Saber and Dunson 1978), wood frogs (j*. sylvatica) (Clark and Euler 1980), the
common frog (R,. temppraria) (Haagstrom 1977), and the leopard frog (R_.
pi pi ens) (Noble 1979) have all been reported to be sensitive to acidity below
pH 5.0. Evidence from counts of dead and moulded egg masses in the
Netherlands (Strijbosch 1979) supports the relationship between acidity and
mortality of frogs. The most serious effects occur in the immature stages
(Gosner and Black 1957). Cricket frog (Acris gryllus) and spring peeper
(Hyla crucifer) embryos exposed to pH 4.0 tor a few hours suffered 85 percent
mortal 1 ty. Rctole (1979) reported embryonic mortality in the leopard frog (R_.
pi pi ens) at pH < 4.7, and Schlichter (1981) observed sub-lethal reductions in
sperm mobility in this species below pH 6.5 and some embryonic mortality at
pH < 6.3. In spite of the sensitivity of £. pi pi ens to acidity in the
laboratory, one of us (RS) has seen adult leopard frogs in an acidic (pH 4.8)
Adirondack lake. The larvae may have developed in ponds that provided refuge
near the lake. Reports of only adult amphibians are of questionable value
because of the much greater sensitivity of the larval forms.
Toads, although terrestrial as adults, are also sensitive to acidity as
larvae and embryos. The common toad (Bufo bufo) was not reported below pH
4.2 in Sweden (Haagstrom 1977), and the natterjack toad (Bufo calamita) was
not found below pH 5.0 in England (Beebee and Griffin 1977)":
The contribution of salamanders to the energy flow of a forest aquatic
ecosystem is considerable. In one study (Burton and Likens 1975a), 20
percent of the energy available to birds and mammals passed through salaman-
ders, and these amphibians represented twice as much standing crop of biomass
as did birds and an amount equal to that of small mammals (Burton and Likens
1975b). Most (94 percent) of the salamanders were terrestrial, but all
salamanders are aquatic as larvae. Not only do amphibians provide energy for
birds and mammals, but they represent the top predators in many temporary
ponds (Orser and Shure 1972).
5-133
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5.7.2 Birds
Direct effects of acidity on birds are not expected, but indirect effects by
alterations in food resources and bioaccumulation of toxic metals are
possible.
5.7.2.1 Food Chain Alterations—Waterfowl that feed on fish are likely to
avoid lakes devoid of prey. Indeed, species richness of fish-eating birds
such as mergansers, loons, and gulls is positively correlated with pH (Aimer
et al. 1978, Nilsson and Nilsson 1978). The diet of the common loon (Gavia
immer) is approximately 80 percent fish, the remainder consisting of
crustaceans, molluscs, aquatic insects, and leeches (Barr 1973). The range
of the loon includes the sensitive areas of Canada's Precambrian Shield
(Godfrey 1966) and the Adirondack Mountains. Populations have declined in
the Adirondacks (Trivelpiece et al. 1979), but no causal relationship between
acidification and declining bird populations was implied (Mclntyre 1979). In
Quebec, the common merganser (Mergus merganser) and the kingfisher
(Megaceryle alcyon) were observed only on those lakes where the summer pH is
higher than 5.6 (DesGranges and Houde 1981). The distribution of the black
duck (Anas rubripes) has been restricted in some lakes in Maine because of
the lack of their preferred invertebrate prey (Reinecke 1979) but habitat
restriction unrelated to acidification is important in this area. Some
waterfowl may prefer acidic lakes if they can prey on the large predatory
insects which are often very common in these lakes (Section 5.3.2.5).
Goldeneye ducks (Bucephala clangula) were shown to favor acidic fishless
lakes that had large insect populations (Eriksson 1979) and to feed in larger
numbers around a lake after the fish were experimentally removed (Eriksson et
al. 1980b). As birds are opportunistic feeders, the alteration of a food
resource in a number of lakes may reduce the population but not cause a total
loss of the population. To a certain extent, birds may switch to other
resources and to other lakes in the region to sustain their feeding
requirements.
Birds such as swallows, flycatchers, and kingbirds that feed on the aerial
adult form of aquatic insects are forced to find alternative food sources if
the insect populations upon which they normally feed are depleted (Section
5.3.2.5). In early spring when many aquatic insects emerge, acid runoff to
lakes and ponds is at a peak. It is also in early spring that the birds
depend heavily on a supply of food to prepare for nesting and raising young.
This may be the explanation for the observation in southern Quebec, where the
tree swallow (Iridoproene bicolor) was more common during the breeding season
around moderately acid lakes studied; however, it was not seen on any of the
very acid lakes in northern Quebec (DesGranges and Houde 1981). Blancher
(1982) observed that weight gain of kingbirds was related to insect
emergence, not lake pH directly. Lake pH was not correlated with densities
of red-winged blackbirds (Agelaius phoniceus) and barn swallows (Hirundo
rustica).
5.7.2.2 Heavy Metal Accumulation—Alterations of food resources may not be
the only mechanism by which birds may be inhibited by acidity. The mobili-
zation of metals at low pH (Chapter E-4, Section 4.6.1.2) may result in
increased body burdens in the higher trophic levels. Studies by Nyholm and
5-134
-------
Myhrberg (1977) and Nyholm (1981) have suggested aluminum as the cause of the
impaired breeding of four species of passerines in Sweden. Aluminum is quite
insoluble in the alkaline conditions characteristic of vertebrate intestines,
but it might be actively transported across the intestinal barrier if calcium
or phosphorus is in short supply. Observed effects were manifested by
reductions in breeding success; formation of thin, porous eggs; small clutch
size; and lower egg weight near acidic lakes. The cause was suggested to be
the high aluminum content of the insects near the acidic lakes (Nyholm,
personal communication). Similar findings of decreased egg size and weight
were found for the eastern kingbird (Tyranus tyranus) in Ontario (Blancher
1982). A laboratory study proved that aluminum is toxic to bird embryos
(Gillani and Chatzinoff 1981) but results from aluminum injected into eggs
are not comparable to field responses to dietary aluminum. Another example of
increased metal levels in wildlife associated with acid lakes is the mercury
concentrations in the eggs of goldeneye ducks (Bucephala clangula) near
acidic Swedish lakes (Eriksson et al. 1980a,b). Across eastern North America
where extensive pesticide use has occurred, the mobilization of pesticides
and heavy metals by acidification may have even more serious effects, but
these considerations have not been researched. This whole area concerning
how acidification may affect metal and pesticide toxicity requires further
investigation.
5.7.3 Mammals
Mammals that feed on aquatic plants and animals, such as muskrats, minks,
otters, shrews, and raccoons, will be affected variously by acidification,
depending on the sensitivity of their food organisms to acidity and their
ability to choose alternate food sources and suitable habitats in acidified
areas. While many species are not directly affected, they are likely to
experience major changes in availability of food and habitat quality. An
increase in the concentration of heavy metals in the diet of certain species
of wildlife may occur (Newman 1979). Raccoons (Procyon lotor) from the
sensitive Muskoka area of Ontario contain mercury levels of 4.5 ppm in their
livers, a level five times greater than in raccoon livers from an area with
non-acidic waters (Wren et al. 1980). Metal contamination of roe deer
(Capreolus capreolus) resulted in reduced weight and antler size in an
industrialized region in Poland (Sawicka-Kapusta 1978, 1979; Jop 1979), but
this metal deposition is not related to the long-range deposition character-
istic of North America. In remote areas of Sweden, however, cadmium accumu-
lated in the body tissues of roe deer and moose (Alces alces) (Frank et al.
1981, Mattson et al. 1981).
The long-term effects of anthropogenic acidification on caribou (Rangifer
tarandus caribou) are of potential concern. The primary source of winter
browse for caribou (Thompson and McCourt 1981), the lichen Cladina stellaris,
is very sensitive to acidity (Chapter E-3, Section 3.2.2).Exposure of this
lichen to simulated acidic rain at pH 4.0 reduced photosynthetic rates by
about a quarter (Lechowicz 1982). Recovery time from drying was also
impaired. The caribou/lichen relationship is very sensitive, as the lichen
normally grows only 6 mm per year (Scotter 1963) and an adult caribou eats 5
kg of lichen per day (Hanson et al. 1975). Any impairment of lichen growth
rate may have a serious impact on the winter range of caribou, but it will
5-135
-------
take years for this effect to be noticeable as normal regeneration of lichen
communities requires in excess of 30 years.
Acidic deposition may affect mammals by causing changes in soil chemistry
that can sequester important minerals (Chapter E-2, Section 2.3.3.3). One
element that is likely to be made less available to herbivorous animals is
selenium. The solubility of selenium in soil pore water declines with pH
(Geering et al. 1968, C. M. Johnson 1975) and uptake by grasses is inhibited
by the presence of SOX in soils (Davies and Watkinson 1966, Gissel-Nielsen
1973), so concentrations of selenium in forage are reduced in areas of sensi-
tive soils receiving acidic deposition (Gissel-Nielsen 1975, Shaw 1981).
Furthermore, excess sulfur in the diet of animals can scavenge selenium from
tissues (Harr 1978). Dietary deficiency of selenium leads to degeneration of
the liver, kidney, and heart (Schwarz and Foltz 1957, Harr 1978). Selenium
deficiency leads to muscular dystrophy ("white muscle disease") in sheep,
cattle, swine, and horses (Muth et . al. 1958, Muth and Allaway 1963,
Hidiroglou et al. 1965, Harr 1978). Many soils in eastern North America are
naturally low in selenium and produce forage with concentrations below the
0.1 ppm level recognized as essential (Kubota et al. 1967, Levesque 1974,
Winter and Gupta 1979). Incidence of white muscle disease has been related
to the use of sulfur-containing fertilizers in areas naturally deficient in
selenium (Davies and Watkinson 1966, Allaway and Hodgson 1964, Allaway 1970).
Effects on the availability of other essential elements, such as molybdenum
(Chapter E-2, Section 2.3.3.3), may be equally important but have not yet
been considered. Speculation concerning mineral availability related to
acidic deposition must await resolution through future research.
5.7.4 Summary
Effects of acidification on vertebrate animals, not including fish (Section
5.6) are still largely speculative. The potential effects are diverse and
research is at an early stage. These data are summarized in Table 5-15.
Many of the effects are expected to take a number of years to appear;
therefore, long-term monitoring will be essential. The following tentative
conclusions can be drawn:
0 Direct effects are most severe on the embryos and larvae of
amphibians, including salamanders, newts, frogs, and toads. Sensi-
tivity to acidity varies widely within closely related taxa, but
total amphibian biomass may decline in areas exposed to acidic
rainfall and snowmelt.
0 Fish-eating birds (e.g., loons, mergansers) will be unable to rear
young in areas where fish populations are limited, resulting in
smaller population sizes for portions of the breeding range.
0 Some insectivorous bird populations may be limited by the reduced
availability of preferred prey (flycatchers, swallows, kingbirds)
around acidic lakes, but others (goldeneye ducks) seek out the
species of aquatic insects found in acidic lakes and may actually
prosper in impacted areas.
5-136
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TABLE 5-15. SUMMARY OF EFFECTS OF ACIDITY ON NON-FISH VERTEBRATES
en
i
CO
Taxa
AMPHIBIA
Ambystoma maculatum
A. jeffersonianum
Trituris vulgar is
T. helveticus
Notophthalamus
vlridescens
Hyla cruel fer
Rana catesbeiana
R. sylvatica
R. temporaria
R. pipiens
Acris gryllus
Bufo bufo
Common name
Yellow- spotted
salamander
Jefferson
salamander
Smooth newt
Palmate newt
Red- spotted newt
"Salamanders"
Spring peeper
Bullfrog
Wood frog
Common frog
Leopard frog
Cricket frog
Common toad
Observation Mechanism
Reproductive failure at Embryonic mortality
pH < 6.0
Egg number correlated ?
with pH
No effect of pH 5.0 ?
Not observed < pH 6.0 Cation concentration
Tolerant to pH 3.8 ?
Tolerant of pH 7.4-4.8 ?
Leave water at pH 4.0 Behavior change
Population density ?
correlated with pH
Mortality at pH 4.0 Embryonic mortality
Mortality below pH 5.0 ?
Mortality below 5.0 Embryonic mortality
Mortality below 5.0 ?
Mortality below 5.0 ?
Mortality below 4.7 Embryonic mortality
Reduction in sperm ?
mortality at pH < 6.5
Adults observed at pH ?
4.8
Mortality at pH 4.0 Embryonic mortality
Not observed < pH 4.2 ?
Evidence
Field obs.
Field obs.
Field obs.
Field correl .
Field obs.
Field obs.
Field pH manip.
Field obs.
Lab study
Field obs.
Lab study
Field obs.
Field obs.
Lab study
Lab study
Field obs.
Lab study
Field obs.
References
Mushlnsky and Brodie 1975,
Pough and Wilson 1977
Clark and Euler 1980
Pough 1976
Cooke and Frazer 1976
Cooke and Frazer 1976
George et al . 1977 ,
pers. obs. (RS)
Hall and Likens 1980a,b
Clark and Euler 1980
Gosner and Black 1957
Clark and Euler 1980
Saber and Dunson 1978
Clark and Euler 1980
Haagstrom 1977
Noble 1979
Schlicter 1981
pers. obs. (RS)
Gosner and Black 1957
Haagstrom 1977
-------
TABLE 5-15. CONTINUED
Taxa
Common name
Observation
Mechanism
Evidence
References
B. calamita
Natterjack toad Not observed < pH 5.0
Field obs.
Beebee and Griffin 1977
to
CO
BIRDS
Gavia inner
Common loon
Mergus merganser Common merganser
Megaceryle alcyon Kingfisher
Iridoprocne blcolor Tree swallow
foijs rubripes Black duck
Eucephala clangula Goldeneye duck
Habitat restriction in
sensitive areas
Avoidance of acid lakes
Avoidance of acid lakes
Avoidance of acid lakes
Avoidance of acid lakes
Preference for acidic
lakes
Elevated (Hg) in eggs
Land use changes,
fish losses?
Fish losses
Fish losses
Fish losses
Aquatic insect losses
Abundance of preda-
tory insect food
items
From Hg in insects
Field obs.
Field obs.
Field obs.
Field obs.
Field obs.
Field obs.
Lab analysis
Passerines
MAMMALS
Procyon lotor
Eastern kingbird
Songbirds (4 sp)
Raccoon
Capreolus capreolus Roe deer
Alces alces Moose
Decreased egg weight
near acidic lakes
Breeding failure, thin,
porous eggs
5 x normal (Hg)
Cd accumulation
Cd accumulation
Aluminum toxicity?
Aluminum in insect
prey
Bioaccumulation
Bioaccumulation
Bioaccumulation
Field obs.
Lab analysis
Field obs.
Lab analysis
Lab analysis
Lab analysis
Lab analysis
Rangifer sp.
Caribou
Loss of winter browse
over a long period
Sulfur sensitivity of Lab study
caribou lichen
Trivelpiece et al. 1979
Mclntyre 1979
DesGranges and Houde 1981
DesGranges and Houde 1981
DesGranges and Houde 1981
DesGranges and Houde 1981
Eriksson 1979
Eriksson et al. 1980b
Blancher 1982
Nyholm 1981, Nyholm and
Myhrberg 1977
Wren et al. 1980
Frank et al. 1981
Frank et al. 1981,
Mattson et al. 1981
Lechowicz 1982
-------
0 Mammals that feed on plants and animals in acidic lakes may accumu-
late higher than normal body burdens of heavy metals, but population
losses have not yet been demonstrated.
0 The large North American herds of caribou may be affected in the
long-term due to the sensitivity of the caribou lichen upon which
they depend for winter browse.
0 Other grazing animals, including some domestic cattle, may be
subject to mineral deficiencies, particularly selenium, if high
SOX deposition continues for extended periods. The seriousness of
this impact is difficult to quantify and is highly speculative at
this time.
0 Mechanisms of impact include disrupted ionic balances in amphibians,
metal toxicity in higher trophic levels of wildlife, alterations in
food chains, and nutrient deficiencies.
5.8 OBSERVED AND ANTICIPATED ECOSYSTEM EFFECTS (J. P. Baker, F. 0. Rahel,
and J. J. Magnuson)
Acidification may produce changes in either ecosystem structure or function.
Effects on structure involve changes in species composition caused by species
declines, extinctions, or replacements. Effects on ecosystem function refer
to changes in such processes as primary production, energy transfer between
trophic levels, detrital decomposition and rates of nutrient cycling. Most
studies have described the response of individual taxa to the acidification
process. Thus, most of our knowledge about the ecosystem-level effects of
acidification concerns changes in structure. Little is known about how these
structural changes influence ecosystem function. The object of this section
is to note the ecosystem changes which have been observed in acidic habitats
and to suggest potential ecosystem responses that need to be examined in
future studies.
5.8.1 Ecosystem Structure
Acidification produces changes in the basic structure of aquatic ecosystems
(Figure 5-17). Certain taxa (e.g., fish and Daphnia) disappear apparently as
a direct result of acid toxicity. Direct effects of acidity or aluminum are,
however, complicated by interactions among a complex web of consumers and
their food resources (Section 5.10.2.3). Important components of upper
trophic levels-fish populations decline or disappear. As a result, large-
bodied acid-tolerant invertebrates become top predators in the system
(5.3.2.5). Shifts in the importance of invertebrate predators may alter
zooplankton community structure which, in turn, may alter the phytoplankton
community structure. The reduction of grazers (snails, amphipods, etc.) may
allow periphyton to accumulate, while the inhibition of detritivores and
decomposers apparently causes detritus to accumulate. Within benthic and
planktonic communities the number of species generally decreases. The
overall result is a general decrease in ecosystem complexity.
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INVERTEBRATE
PREDATORS
SMALL GRAZERS AND
DETRITIVORES
MACROPHYTES AND
PERIPHYTON
NON-ACIDIFIED LAKEJ
INVERTEBRATE
PREDATORS
V
N
SMALL GRAZERS AND
DETRITIVORES
DETRITUS
ACIDIFIED LAKE
Figure 5-17.
Trophic interactions in a neutral pH, oligotrophic lake
compared to those in an acidified lake. Dotted lines
indicate trophic interactions which may be particularly
affected by acidification. Note the replacement of fish
by invertebrates as the top-level predators. Adapted from
Roberts et al. (1982).
5-140
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Woodwell (1970) considered simplification a system response common to all
types of environmental pollution and also to natural sources of environmental
stress. It is possible that simplification increases system instability
(e.g., Woodwell 1970, Van Voris et al. 1980), although the relationship
between system complexity and system stability is disputed (Allen and Starr
1982). Marmorek (1984) and Van et al. (1982) observed in field experiments
that acidification indirectly reduced the short-term stability and resilience
of the plankton community to nutrient additions.
The physical structure of the aquatic system may also be slightly altered
with acidification. The correlation between increasing acidity and increased
water clarity has been well established (Chapter E-4, Section 4.6.3.4). With
an increase in light penetration, some shift in the thermal budget and
patterns of thermal stratification may occur as has been demonstrated for
Lake 223 in the Experimental Lakes Area of Ontario (Schindler and Turner
1982).
5.8.2 Ecosystem Function
5.8.2.1 Nutri ent Cycling—11 has been suggested that nutrient cycling and
nutrient availability to primary producers are reduced in acidic aquatic
environments. The rate of nutrient cycling is thought to be slowed primarily
because of inhibition of bacterial decomposition and a sealing-off of mineral
sediments from the overlying water column with the accumulation of detritus
and periphyton on the lake bottom (Section 5.3.2.1). Grahn et al. (1974)
speculated that acidification stimulated lake oligotrophication as a result
of these changes, but definite confirmation of this hypothesis is lacking.
Nutrient availability could also be affected by chemical changes in the
water. Of particular importance may be decreased phosphorus availability
because of aluminum-phosphorus interactions (Chapter E-4, Section 4.6.3.5),
decreased levels of dissolved inorganic carbon due to the decrease in pH
(Section 5.5.4.2) and precipitation of organics (Chapter E-4, Section
4.6.3.3), and increased displacement of these materials into benthic
habitats. Although all of these postulated chemical changes are theoret-
ically plausible and potentially very significant, effects on nutrient
cycling in acidic waters have not yet been experimentally demonstrated.
5.8.2.2 Energy Cycling—Previous sections have discussed four types of
possible reactions to acidification that are relevant to energy cycling in
aquatic systems: 1) a potential decrease in primary productivity, 2)
decreased growth efficiencies, 3) decreased energy transfer between trophic
levels and 4) elimination of upper trophic levels. The evidence or lack of
evidence for these hypotheses is discussed below.
Biological productivity in aquatic ecosystems is supported by both
allochthonous organic carbon imported from sources external to the system
plus autochthonous production of organic carbon by primary producers within
the aquatic system. As a result of decreased nutrient availability, water
column primary productivity in acidic waters may be altered. Limited obser-
vations from field studies reviewed in Section 5.5.2.2 indicate, however,
that in most cases acidification has no consistent adverse effect on primary
5-141
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primary productivity. Adverse effects of decreased nutrient availability on
water column primary productivity may be counterbalanced by other changes
(especially increased light penetration) that stimulate primary production.
Although acidification does not consistently decrease primary productivity,
increased light penetration apparently does, to a certain extent, increase
the importance of benthic primary producers relative to planktonic primary
producers. The effects of acidification on total primary production (includ-
ing periphyton, macrophytes and phytoplankton) have not been studied.
Energy transfers within aquatic systems can be examined both within a given
trophic level and between trophic levels. Growth efficiency usually refers
to within stratum transfer, the fraction of a given quantity of energy (food
or light energy) consumed that is manifested as production (growth and
reproduction). Organisms that inhabit acidic waters may be inherently less
efficient or may be less efficient because of acid-induced stress, but exami-
nation of this phenomenon has been limited. Fish have been observed in
laboratory experiments to grow more slowly at lower pH levels (Section
5.6.4.1.3). Primary producers in some acidic waters (Sections 5.5.2.2 and
5.3.2.2.3) have lower instantaneous rates of production per unit biomass.
Possible reasons for this lower production are numerous, however, and have
not been clearly defined. No studies of growth efficiencies for zooplankton,
benthos, or other aquatic organisms have been completed. If growth effi-
ciencies are reduced in acidic environments, energy transfer through the food
chain would be reduced.
Energy transfers between trophic levels involve the percentage of available
food actually used by consumers, or relative productivities in successive
trophic levels. In Section 5.5.4, it is postulated that the transfer of
energy between phytoplankton and zooplankton may be inhibited by the inedible
nature of many of the phytoplankton species common in acidic lakes. In
stream systems, a reduction in populations of benthic invertebrate grazers
may decrease conversion of primary production into secondary production
(Section 5.3.2.2.3). Processing of detrital particles may also be affected.
Again, some evidence suggests energy cycling and energy transfer through the
food chain may be inhibited.
One of the best documented changes associated with acidification is the
decline and loss of fish populations that represent major components of upper
trophic levels in aquatic ecosystems. Loss of fish populations results in a
shortened aquatic food chain.
5.8.3 Summary
Structural changes in acidified aquatic ecosystems have been well documented
and include the loss of fish populations, reductions in the number and
diversity of benthic and planktonic invertebrates, and accumulations of
periphyton and detritus. How these structural changes affect ecosystem
processes such as primary production, energy transfers between trophic
levels, or nutrient cycling is largely unknown. Because acidification
potentially can have significant effects on these processes, the effects of
acidification on these key aspects of ecosystem function need to be address-
ed in future research.
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5.9 MITIGATIVE OPTIONS RELATIVE TO BIOLOGICAL POPULATIONS AT RISK
(C. T. Driscoll, C. A. Guthrie, arid G. C. Schafran)
The concept of surface water neutralization as a result of base and
phosphorus additions is discussed in Chapter E-4, Section 4.7. The biologi-
cal response to these additions and other rnitigative options for fish popu-
lations at risk from acidification of surface waters follows.
5.9.1 Biological Response to Neutralization
In lakes where neutralization has resulted in large, rapid pH changes (e.g.,
Ca(OH)2 addition, see Chapter E-4, Section 4.7.1), phytoplankton concentra-
tions have been observed to decline drastically. This phenomenon may be
either the result of stress associated with a drastic change in pH ("pH
shock") or removal of algal biomass with metals through flocculation and
precipitation processes (Scheider and Dillon 1976, Scheider et al. 1975).
Yan and Dillon (1981) noted that a small pH change, or a large pH change
initiated gradually, resulted in no change in biomass of lake phytoplankton.
After base addition, phytoplankton undergo a taxonomic shift. Certain
species will disappear while others appear. Species dominance has been
observed to shift and total number of species has been observed to increase.
Species dominance/composition are lake-specific, so response of the phyto-
plankton population cannot be generalized for all lakes. Subsequent to
liming, Scheider et al. (1975) observed a shift in dominance to the genera
Dinobryon and an unidentified chrysomonad. The appearance of diatoms
(Bacillariophycae--mostly Navicula and Nitzschia) and blue-green algae
(Cyanophyta-Oscillatoria) was also noted. Yan and Dillon (1981) observed an
increase in the contribution of dinoflagellates, while cryptomonads declined.
Among many species changes noted in a Swedish liming experiment were the
increase in small cryptomonads, diatoms, and chrysophyceans and the dis-
appearance of Merismopedia sp. (Hultberg and Andersson 1982).
After a population has been depleted by base addition, within a few months
phytoplankton biomass will increase and approach preneutralization levels.
The rapid recovery after base addition appears to be due to the short life
cycle of phytoplankton and decreased predation due to decreases in
zooplankton population.
Zooplankton populations are affected in much the same way as phytoplankton.
Additions of base that drastically increase lake pH immediately reduce
standing stocks of zooplankton. In three Canadian lakes where the most
frequently observed species were Cyclops vernal is, Chydorus sphaericus, and
Bosmina longirostris, the addition of base, which quickly increased pH more
than two units, caused immediate and drastic reductions in zooplankton
standing stock. Base additions that have resulted in smaller pH changes have
not affected the population negatively (Scheider et al. 1975, Dillon et al.
1979). Swedish lakes that have undergone a gradual increase in pH through
base application show a substantial increase in zooplankton biomass, shifts
in species composition, and increases in species diversity (Hultberg and
Andersson 1982).
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Recovery of zooplankton populations is much slower than that observed for
phytoplankton. For two full years following base addition, zooplankton
biomass was observed not to recover to pretreatment levels (Yan and Dillon
1981). This relatively slow recovery from base addition stress may be due to
slow life cycles and recolonization difficulties.
The literature is not consistent with respect to the response of benthic
fauna to base addition. In the first year following large pH increases due
to base addition, Scheider et al. (1975) observed numbers of benthic organ-
isms decrease substantially. Chironomids, which were observed to be dominant
prior to neutralization (Scheider et al. 1975, Yan and Dillon 1981), contrib-
uted significantly to this decline. This was attributed to an interruption
of a life cycle in response to the sudden pH change. However, this is not
consistent with Swedish observations. Hultberg and Andersson (1982) observed
that the groups Orthocladinae and Tanypodinae increased, while no change was
evident in trichopteran populations. With benthic fauna constituting an
important food source for fish, population perturbations resulting from
neutralization may affect fish positively or negatively.
In some regions, a feltlike structure of algal filaments, detritus, and
Sphagnum completely covers lake sediments and depletes normal populations of
submerged vegetation like Isoetes and Lobelia (Grahn et al. 1974, Hendrey and
Vertucci 1980). Hultberg and Andersson (1982) indicate that liming appears
to have a profound effect on Sphagnum. After base addition, Sphagnum was
rapidly eliminated from the littoral region where CaC03 was spread.
Populations were slowly depleted (1 to 2 years) in the remainder of the
treated lakes. The few plants that survived neutralization exhibited very
slow growth rates ( ~ 1 cm yr-1) as compared to acidic lake populations (8
to 10 cm yr-1) (Hultberg and Andersson 1982). In lakes that were allowed
to reacidify, Sphagnum was observed to recolonize the benthic region.
Neutralization to improve the water quality of acidified waters has both a
long- and short-term effect on fish. Immediately following base addition and
subsequent pH rise, aluminum hydrolysis generally occurs. This perturbation,
as previously described, may be detrimental to the existing fish population
(Baker and Schofield 1980). Mortality of fish may be lessened by incremental
addition of base, resulting in small pH changes. In some lakes this may not
be deemed necessary as the fish population may be negligible.
The long-term consequence of lake neutralization, provided reacidification is
not allowed to occur, is a much more hospitable environment for fish. An
immediate response (improvement) in reproduction and survival has been
observed in one-year-old fry (Hultberg and Andersson 1982). An increase in
recruitment and fish survival tends to increase the biomass of the younger
fish where previously the population had been dominated by older fish
(Dickson 1978). If neutral pH is maintained, fish reproduction and survival
will show marked improvement over acidified conditions and possibly restore
the population to pre-acidification levels. Restocking of native species,
lost because of acidification, may be necessary in some waters.
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5.9.2 Improving Fish Survival in Acidified Waters
Three major approaches for improving fish survival in acidified waters deal
directly with the fish. They are 1) screening existent fish strains to
determine which strains exhibit high acid tolerance, 2) selectively breeding
a given strain for improved tolerance to low pH, and 3) acclimating a group
of fish to increase their resistance to acidic water.
5.9.2.1 Genetic Screening—Several studies have shown differences in acid
tolerance between different strains within the same species (D. W. Johnson
1975, brook trout; Gjedrem 1976, brown trout; Robinson et al. 1976, Swarts et
al. 1978, Edwards and Gjedrem 1979, Rahel and Magnuson 1980, yellow perch;
and Schofield et al. 1981).
Edwards and Gjedrem (1979) determined that the method used for screening
different strains was important in determining the hierarchy of tolerance
among strains. They screened brown trout fingerlings (5.8 +_ 0.8 g) in water
synthetically acidified to pH values of 2.5, 3.0, and 4.0 and brown trout
eggs and fry in naturally acidic water (pH 4.7) and in water adjusted from pH
4.7 to 5.2 with sodium hydroxide. They found a high correlation of ranking
among strains tested at low pH values, indicating that the pH level used
within this range was unimportant. However, when they compared ranking
obtained from the fingerlings tested at very low pH values and those deter-
mined from the eggs and fry tested in the naturally acidic water, they found
a low rank correlation between strains. They concluded that the two
different procedures were apparently testing for different traits and thus
could not be used interchangeably.
The results of Edwards and Gjedrem (1979) indicate that a standardized
screening procedure is very important in determining the relative tolerance
of strains within species. Their results also indicate that the life cycle
stage screened is important in determining relative strain tolerance. Thus,
it is important to develop a screening procedure consistent with the goals of
the project. Edwards and Gjedrem (1979) concluded that a screening program
aimed at reestablishing viable populations in acidified waters must select
for strains with acid-resistant egg and larval stages because the major cause
of trout population losses is thought to be poor recruitment caused by egg
and fry mortality (Beamish and Harvey 1972, Jensen and Snekvik 1972,
Leivestad et al. 1976, Schofield 1977). However, if the goal of a screening
program is to find a strain to be used in maintaining stocked populations,
the screening procedure should target the life cycle stages that will be
stocked.
5.9.2.2 Selective Breeding—The logical extension of a genetic screening
program is to select for acid tolerance within a few superior strains and
improve their acid tolerance through selective breeding. Gjedrem (1976) and
Edwards and Gjedrem (1979) found high herilabilities (ratio of genetic
variance to total variance) for acid tolerance in eggs and alevins of brown
trout. They concluded that there was a good possibility of producing
acid-tolerant strains of brown trout through selective breeding.
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Selective breeding tests with brook trout have produced mixed results.
Swarts et al. (1978) performed a single selection with NYSV strain brook
trout (selecting to 80 to 90 percent loss of equilibrium at pH 3.4 to 3.5)
and found no increased tolerance in their offspring in field or laboratory
tests. Schofield et al. (1981) selected yearling (1977 year class) domestic
strain brook trout to 50 percent, using naturally acidified runoff water.
They then challenged the offspring (1979 year class) of the resistant and
non-resistant cohorts as fry in naturally acidified water. The offspring of
the resistant cohort were significantly more resistant (mean LT^o 195.5 hr)
than those of the non-resistant cohort (LT5Q 72.0 hr; P < 0.001). However,
when an identical test was performed on the 1980 year class offspring of the
1977 year class resistant and non-resistant cohorts, the offspring of the
resistant cohort exhibited performance inferior to that of the offspring of
the non-resistant cohort (LTso values 76.6 and 77.1 hr vs 84.7 hr,
respectively). Included in the 1980 year class tests were tests of hybrid
crosses between resistant and non-resistant cohorts and two wild strains from
Canada (Assinica and Temiscamie). In these tests the resistant X Assinica
and resistant X Temiscamie always performed better than the non-resistant X
Assinica and non-resistant X Temiscamie. From these results Schofield et al.
(1981) hypothesized that genetically inherent physiological acid tolerance
may be fixed within the selected cohorts.
In a preliminary field trial, Schofield et al. (1981) separated Assinica X
domestic yearlings into resistant and non-resistant cohorts in March of 1979,
stocked them in equal numbers in an acidified lake in May, and sampled them
in July. They observed a 3:1 return of resistant over non-resistant fish.
However, more extensive field trials performed in 1980 produced a resistant/
non-resistant ratio not significantly different from the expected 1:1 ratio
of the no difference case. Schofield et al. (1981) attributed the lack of an
unbalanced ratio to the relatively good water quality conditions in the
spring of 1980 caused by low snowfall during the winter of 1980. This study
appears to give some evidence of improved acid tolerance of brook trout
through selective breeding, but it is far from conclusive.
Hybrid vigor with regards to acid tolerance has been observed in several
studies. Robinson et al. (1976) found heterosis (hybrid vigor) in 66 percent
of the strain crosses tested. Edwards and Gjedrem (1979) observed mean
percent survival in hybrid crosses of brown trout to be twice that of the
parental strains. From this they suggest that the most efficient way to
produce acid-tolerant strains for restocking acidified waters would be to
identify the best strain crosses and then maintain just a few pure bred
strains in the hatchery. These strains could be improved by selective
breeding while hybrid fish for stocking could be routinely produced by
crossing a brood fish of the pure bred lines.
5.9.2.3 Acclimation—A conceivable method for improving the success of
stocked populations in acidified waters would be to acclimate the fish to the
acidic conditions before stocking. The question of whether fish can
acclimate to acidic conditions has been addressed by numerous authors, with
mixed results. Most of the studies in which fish were acclimated to sub-
lethal pH values and then tested for increased survival at lethal pH values
have produced negative results. Lloyd and Jordan (1964) acclimated
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rainbow trout to pH values of 6.55, 7.50, and 8.40 and found no difference in
survivorship when the fish were tested at pH values from 3.0 to 4.0.
Robinson et al. (1976) held brook trout at pH 3.75 for one week and then
tested them for survival at pH 2.5 and 3.0. They found that survival time
was 20 to 25 percent less in acclimated fish than in fish not previously
exposed to acidic water. Falk and Dunson (1977) exposed brook trout to
sublethal pH values of 5.0 and 5.8 for 2 or 24 hours prior to testing for
survival at pH 3.15 or 3.5. They found significant differences in survival
time between acclimated and non-acclimated fish in only three of nine tests.
Swarts et al. (1978) performed laboratory and field acclimation trials with
brook trout. In the laboratory they acclimated the fish to pH 4.25 for 10
days or pH 4.8 for 28 days and then tested them for improved survival at pH
3.25 or 3.6 respectively. They found no consistent differences between
acclimated and non-acclimated fish in their laboratory trials. In three
field trials in which fish were held in an acidified stream (pH 4.8 to 5.8)
and then tested in an acidic river (pH 4.2), the acclimated fish performed
better than non-acclimated fish in only one trial.
In a study with embryos and alevins of Atlantic salmon and rainbow trout
which had been incubated at pH values ranging from 4.5 to 6.8 for variable
time periods, Oaye (1980) could find no difference in tolerance between the
different groups and thus concluded no acclimation had occurred. In a simi-
lar study, performed by Trojnar (1977b), brook trout eggs were incubated at
pH 4.6, 5.0, 5.6, and 8.0 and then tested at swim-up for survival at pH
values from 4.0 to 7.86. He found that fish incubated at pH 5.6 and below
showed greatly increased survival at low pH as compared with fish incubated
at pH 8.0. He attributed the difference to acclimation.
Physiological evidence for acclimation in brown trout exposed to acidified
water was provided by McWilliams (1980b), who suggested that acclimation
might occur through a progressive decrease in the diffusional permeability of
the gills to sodium. However, actual resistance to lowered pH levels, in
terms of increased survivalship, was not determined in this study.
In all of the aforementioned studies, the acclimation procedure consisted of
holding the fish at a single sublethal pH for a fixed time period and then
transferring them to the test pH levels. Guthrie (1981) used a different
method. He hypothesized that previous acclimation attempts had failed for
three major reasons. First, if the acclimation pH was too high the fish
might not need to adjust physiologically to maintain homeostasis. The study
by Lloyd and Jordan (1964) might be an example of this. Second, if the
acclimation pH is too low then it might constitute a major stress in itself,
to which the fish are unable to adjust. The study by Robinson et al. (1976),
where the fish were acclimated to a pH of 3.75 before being tested at a lower
pH, is probably an example of this. Third, if the test pH is very low and
the adaptive response of the fish is overwhelmed, then no amount of previous
acclimation will improve survival. This probably occurred in the studies
where the test pH was below 4.0 (Lloyd and Jordm 1964, Robinson et al. 1976,
Swarts et al. 1978, Falk and Dunson 1977).
To avoid these problems, Guthrie (1981) developed a gradual acclimation
procedure in which the acidity and aluminum concentration were increased from
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control conditions to test conditions over a period of 4 to 5 days. He used
test pH values of 5.0, 4.5, and 4.0 with nominal aluminum concentrations of
0.2 and 0.4 mg Al £~1. In acclimation tests on brook trout sac fry and
swim-up fry, Guthrie (1981) found significantly improved survival at pH 5.0
and 4.5 at both aluminum levels, but no difference in survival between
acclimated and non-acclimated fish at pH 4.0. This lends credence to the
hypothesis that pH values below 4.0 are too low for testing for acclimation.
Guthrie also acclimated brook trout parr (55.7 _+ 6.8 mm) to naturally acidic
water (pH 4.9, 0.32 mg Al £-1) by gradually changing water from non-
acidified lake water (pH 6.5) to acidic brook water. After 6 days in the
acidic brook water, 80 percent of the acclimated fish remained alive while
only 40 percent of the non-acclimated fish (transferred into the acidic brook
water at the same time that the acclimation procedure was completed) were
still alive. In experiments with advanced fry (28 to 36 mm) and yearlings at
pH 5.0 with 0.4 mg Al £-1, dramatic improvements in the survival of the
acclimated fish were also observed. However, at pH 4.5 with the same
aluminum level, acclimation did not improve survivorship in these life
history stages.
The studies performed by Guthrie (1981) clearly demonstrate the ability of
brook trout to resist increased acidity and aluminum levels, within specific
limits of water quality and developmental sensitivities, as measured by
improved survival of fish in short-term gradual acclimation treatments.
These results indicate it may be possible, through acclimation prior to
stocking, to improve initial survival in hatchery-reared brook trout destined
for stocking in waters of low pH and high Al levels.
5.9.2.4 Limitations of Techniques to Improve Fish Survival—For the future
it appears that a combination of these three techniques could be a feasible
strategy for maintaining a sport fishery in waters where the extent of acidi-
fication is such that a natural fishery is no longer possible. This could be
accomplished by screening for the most acid-resistant strains of fish,
selectively breeding those strains and acclimating them to the acid water
before stocking.
This strategy would probably be successful in allowing the maintenance of a
sport fishery where none could exist otherwise; however, it would not be a
solution. It is doubtful that these techniques could ever be used to re-
establish a naturally-reproducing population where one had been lost due to
acidification. Also, because these techniques all require a great deal of
propagation work and clearly defined genetic strains, it would only be
possible to use game fish. The reestablishment of non-game fish in acidified
waters using these techniques would not be feasible.
When these techniques are used to reestablish sport fisheries in acidified
waters there is one foreseeable contraindication. Toxic metals such as
mercury may be mobilized as a result of acidification. This could result in
a hazardous situation if stocked fish accumulated these contaminants before
they were caught. Thus, it is important that fish stocked in acidified
waters be closely monitored for toxic metals contamination.
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5.9.3 Summary—Base addition to neutralize acidified waters will affect
aquatic organisms variously, and effects are likely to be lake-specific.
Phytoplankton undergo a taxonomic shift but their recovery approaches pre-
neutralization levels within a few months. Zooplankton are affected
similarly, though their recovery is much slower. The literature on benthic
fauna response to base addition shows no consistent response, but base
addition greatly reduces Sphagnum, a major component of algal mats covering
lake bottoms. Both long-term and short-term effects on fish populations can
be seen and, provided re-acidification is not allowed to occur, base addition
creates a much more hospitable environment for fish.
All three techniques for producing fish better able to survive in acidified
waters—genetic screening, selective breeding, and acclimation--show promise
as ameliorative strategies. However, all are still in the early stages of
development and require more laboratory and field testing before they will be
well enough defined to be useful as fish management tools.
5.10 CONCLUSIONS (J. J. Magnuson, F. J. Rahel, J. P. Baker, R. Singer,
and J. H. Peverly)
Although the literature regarding the response of aquatic biota to acidifica-
tion is sometimes conflicting, some effects have been well documented. These
are summarized below (Section 5.10.1). Emphasis is placed on those biologi-
cal changes that are supported by a combination of field observations, field
experiments, and laboratory experiments. Together, these species declines,
extinctions, and replacements represent major changes in the structure of
acidified aquatic ecosystems. The next section (5.10.2) focuses on the
mechanisms by which acidification affects aquatic ecosystems. Although
mechanisms by which acidification may affect processes such as primary
production, energy transfer between trophic levels, and nutrient cycling have
been hypothesized, few have been critically evaluated using field and labora-
tory experiments. The major conclusion is that many of these mechanisms are
speculative and need to be examined in future research. Section 5.10.3
describes potential mitigative options from a biological perspective. The
final section (5.10.4) presents an overview of biological changes expected if
current rates of acidic deposition continue in the northeastern United States
and southeastern Canada.
5.10.1 Effects of Acidification on Aquatic Organisms
The effects of acidification on aquatic organisms that are supported by
numerous observations and experimental studies are summarized in Table 5-16
and in the following statements.
Benthos
° The bottom community, which provides substrates for many organisms
and is the principal site of nutrient recycling, is severely altered
in clear waters low in pH, as compared to otherwise similar, but
neutral pH waters.
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TABLE 5-16. EFFECTS OF INCREASING ACIDITY ON AQUATIC ECOSYSTEMS. "NUMEROUS"
REFERS TO MANY OBSERVATIONS OR EXPERIMENTS, WHICH ARE DESCRIBED IN THE SECTIONS INDICATED
Taxa or Process
Type of Evidence
Field Observation Field Experiment Lab Experiment
Observed Effects
en
i
en
O
Benthos
Molluscs
(most species except
fingernail clams, family
sphaeriidae)
Crayfish
Amphipods
(Gammarus)
Mayfly larvae
(Ephemeroptera)
Water striders (Gerridae),
baclcswimmers (Notonectidac),
water boatmen (Corixidae),
beetles (Dytiscidae,
Gyrinidae), dragonflles
(Odonata)
Benthos community structure
Numerous
(Section 5.2 and
5.3)
Aimer et al. 1978 Mills 1984
K. Okland 1980c
Sutcliffe and
Carrick 1973
Numerous (Section
5.3)
Numerous (Section
5.3)
Numerous (Section
5.2 and 5.3)
Hall et al. 1980
Mai ley 1980
Costa 1967
Borgstrom and
Hendrey 1976
Bell and
Nebeker 1969,
Bell 1971
Hall et al. 1980
Benthic algae
(periphyton)
Numerous (Section
5.3)
Hall et al. 1980
Schindler 1980
Bell and
Nebeker 1969,
Bell 1971,
Mai ley 1980
Hendrey 1976
The calcareous shell of these animals
1s soluble under acidic conditions
making this group highly sensitive to
low pH. Few species present below pH
6.0 except for several species of
fingernail clams which may persist
down to pH 4.5 - 5.0.
In soft water lakes, calcium uptake
and exoskeleton formation Inhibited In
the pH range 5.0-5.8. Reproduction
impaired at pH 5.4.
Absent below pH 6.0, 1n the laboratory
avoids pH 6.2 and lower.
Most species decline or are absent in
the pH range 4.5 to 5.5.
Tolerant of acidity. Increase in
abundance in acidified lakes (below pH
5.0) after other Invertebrate groups
and fish have been eliminated.
With increasing acidity, species
richness declines. Entire groups of
aquatic organisms are absent or poorly
represented below pH 5.0 (e.g., mol-
luscs, amphipods, crayfish, mayflies).
Other taxa become dominant, particu-
larly after the loss of fishes (e.g.,
predacious beetles and true bugs).
Algal mass overgrow rooted plants
and cover bottom subtrates in
acidified lakes below pH 5.0
-------
TABLE 5-16. CONTINUED
Taxa or Process
Type of Evidence
Field Observation Field Experiment Lab Experiment
Observed Effects
Macrophytes
Eriocaulon sp.
Lobelia sp.
Grahn 1977, Best
and Peverly 1981,
Miller et al. 1982
Laake 1976 Rosette plant communities may
become overgrown by algal mats.
Tissue aluminum concentrations
Increase as pH decreases.
Photosynthesis of rosette species
decreases by 75% as pH declines
from 5.5 to 4.0.
Plankton
en
i
en
Zooplankton community
structure
Phytoplankton community
structure
Fishes
Fathead Minnow
(Pimephales promelas)
Numerous (Section
5.5)
Numerous (Section
5.2 and 5.5)
Davis and
Ozburn 1969
Parent and
Cheetham 1980
Numerous (Section
5.2 and 5.5.)
Van and Stokes
1978
Rahel and Magnuson
1983
Mills 1984
Mount 1973
Most species are acid-sensitive
and absent below pH 7.0 to 5.5
The number of species declines as
acidity increases. Taxa
characteristic of acid conditions
include certain genera of
rotifers (Keratella. Kellicottla.
Polyarthra); cladocerans
(BosminaTT and copepods
(Diaptonus).
The number of species declines as
acidity increases. Dinoflagellates
(Phylum Pyrrophyta) frequently
dominate acidified lakes (pH 4.0-5.0),
Dinoflagellates are a less palatable
food source for zooplankton compared
to the phytoplankton they frequently
replace.
One of the most acid-sensitive
fish species. Reproductive failure
occurs near pH 6.0. Generally
absent in waters below pH 6.5.
-------
TABLE 5-16. CONTINUED
Taxa or Process
Type of Evidence
Field Observation Field Experiment Lab Experiment
Observed Effects
en
ro
Darters
(Etheostoma exile. £.
nigrum, Percina caprodes)
and Minnows (several
Motropis spp. Pimephales
notatus)
Smallmouth Bass
(Micropterus dolomleui)
Lake Trout
(Salve! inus namayeusch)
White Sucker
(Catostomus commersoni)
Rainbow Trout
(Salmo gairdneri)
Atlantic Salmon
(Salmo salar)
Brown Trout
(Salmo trutta)
Brook Trout
(Salvelinus fontinalis)
Sunfishes
(Amblopl i tes rupestris.
Microptcrus sTfmoidesT
Lepomis "spp.l
Yellow Perch
(Perca flavescens)
Decomposition
Harvey 1980
Rahel and Magnuson
1983
Beamish 1976,
Harvey 1980, Rahel
and Magnuson 1933
Beamish 1976, Mills 1904
Beamish et al. 1975
Harvey 1980, Rahel
and Magnuson 1983
Numerous (Section
5.6)
Numerous (Section
5.6)
Numerous (Section
5.6)
Numerous (Section
5.6)
Harvey 1980.
Rahel and Magnuson
1983
Svardson 1976,
Keller et al. 1980,
Harvey 1980, Rahel
and Magnuson 1983
Hendrey 1976,
Leivestad et al.
1976
Mills 19B4
Hall et al. 1980
Smith 1957
Scheider et al.
1976, Gahnstrom
et al. 1980, Hall
et al. 1980
Rahel and Very acid-sensitive. Generally
Magnuson 1983 absent below pH 6.0 in both
naturally acidic and anthro-
pogenically acidified waters.
Reproduction ceases and populations
become extinct below pH 5.2-5.5
Beamish 1972 Experiences reproductive failure near
Trojnar 1977a pH 5.0. Generally absent below pH 5.0
In both naturally acidic and
anthropogenlcally acidifed waters.
Numerous Adversely affected by pHs below
(Section 5.6) 5.0-5.5
Numerous Adversely affected by pHs below
(Section 5.6) 5.0.
Numerous Lower pH limit between 4.5 to
(Section 5.6) 5.0.
Numerous Lower pH limit between 4.2 to
(Section 5.6) 5.0.
Lower pH limit near 4.5.
Rahel 1983 Lower pH limit 4.2 to 4.5. May
become very abundant after other
species have become extinct.
Leivestad et Bacterial decomposition is signifi-
al. 1976 significantly reduced in the pH
range 4.0 to 5.0. In many
cases, fungi replace bacteria as
the primary decomposers
-------
Bacterial metabolic rates are decreased between pH 6.0 and 4.0, and
shredding invertebrate populations are reduced in numbers, bringing
about an increased accumulation of undecomposed organic materials.
Most substrates are covered with an encrusting mat of algae and
detritus in acidic lakes and streams below pH 5.0.
Many predatory insects (beetles, true bugs, dragonflies) increase in
numbers below pH 6.0 in lakes and streams. Their effect on the
plankton and on benthic detritivores is not known.
Several preferred food sources for game fish (e.g., Gammarus snails,
many mayflies and stoneflies) do not survive below pH 5.0, but
fisheries impacts due to food shortages have not been observed.
Macrophytes
Dominant macrophyte species are the same in both acidified {pH less
than 5.6) and non-acidified (pH 5.6 to 7.5), oligotrophic North
American lakes.
-Shifts to Sphagnum-dominated macrophyte communities have been
documented in six Swedish lakes acidified for at least 15 years.
However, this does not seem to be a general property of acidified
lakes as there is currently no trend toward dominance of macrophyte
communites by Sphagnum spp. in 50 oligotrophic, softwater lakes
surveyed in North America.
Standing crops of macrophytes vary widely (5 to 500 g dry wt m~2)
in softwater, oligotrophic lakes, and acidification produces no
consistent changes in standing crop. In Lobelia dortmanna, a common
plant in softwater, oligotrophic lakes, oxygen production was
reduced 75 percent at pH 4.0 vs pri 4.3 to 5.5 in one flow-through
laboratory experiment.
In the two published studies of metal concentrations in macrophytes
from acidic lakes, tissue concentrations of iron, lead, copper and
especially aluminum are higher, while cadmium, zinc and manganese
are lower compared to tissue concentrations in plants from non-
acidic lakes.
Plankton
Changes in species composition, standing crop, and productivity of
the plankton community with acidification are complex and probably
result from not only lower pH levels and higher metal concentra-
tions, but also decreased fish predation, increased water clarity,
and perhaps decreased nutrient availability.
The structure of the plankton community in acidic lakes (pH 4.0 to
6.0) is markedly different from that in non-acidic lakes within the
5-153
-------
same region. With increasing acidity, the total number of species
decreases (by 30 to 70 percent) and biomass is dominated by fewer
species.
Comparisons between acidic and non-acidic lakes within the same
region and experimental acidification of a lake indicate no
consistent change in water column primary productivity with
increased acidity.
Data on zooplankton productivity are not available. In three
studies, the biomass and/or numbers of zooplankton were lower in
more acidic lakes (pH 4.0 to 5.0).
Fish
The clearest evidence for impacts of acidification on aquatic biota
is adverse effects on fish.
Loss of fish populations associated with acidification of surface
waters has been documented in Nova Scotia, southern Norway, and the
LaCloche Mountain range of Ontario. Available data for these
regions include historic records of declining fish populations
coupled with historic records of increasing water acidity.
Additional evidence for loss of fish populations is available from
the Adirondack region of New York State and southern Sweden.
In the United States, only in the Adirondack region have adverse
effects of acidification on fish populations been observed. The
presence of fish in Adirondack lakes and streams is correlated with
pH level. Particularly below pH 5.0, the occurrence of fish is
reduced. Loss of fish populations has been documented for about 180
Adirondack lakes (out of a total of approximately 2877), although
historic records are not available at this time to relate each loss
specifically to acidification or acid deposition.
Fish kills have been observed during episodic acidification of
surface waters in Norway and Ontario. In addition, in hatcheries
receiving water directly from lakes or rivers, unusually heavy
mortalities of adult and young fish have occurred in the Adirondack
region, Nova Scotia, and Norway. These mortalities are typically
associated with rapid decreases in pH (generally to pH levels below
4.5 to 5.0) during snowmelt.
Many fish populations in acidic waters (pH 4.5 to 5.0) lack young
fish, implying that failure to reproduce is a common, although not
the only, cause for extinction of fish populations with
acidification. In Sweden, neutralization through lake liming
resulted in the recurrence of young fish.
Field observations of growth of adult fish in acidic (pH 4.0 to 5.0)
versus non-acidic waters, or through time with acidification,
5-154
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typically indicate increased growth or no change with increased
acidity. In some cases, increased growth may be a result of reduced
competition for food as fish populations decline.
Experiments in the laboratory and the field have established a
direct cause and effect between acidification and adverse effects on
fish. In the field, acid additions to Lake 223 in the Experimental
Lakes Area of Ontario produced pH declines from pH 6.5 to 5.9 in
1976 to pH 5.1 in 1981 and resulted in reproductive failures and/or
extinction of several fish populations. In laboratory bioassays, pH
and aluminum levels typical of acidified surface waters were toxic
to fish.
Other Related Biota
Effects of acidification on amphibians, birds, and mammals are still
largely speculative. Research is at an early stage. Decreased pH
levels have been demonstrated in the laboratory to decrease
amphibian reproductive success, but the significance and extent of
breeding habitats acidified or sensitive to acidification have not
yet been evaluated.
Ecosystem Effects
Changes in ecosystem structure have been well documented in acidi-
fied aquatic habitats and include species declines, local extinc-
tions and reduced species richness in many taxonomic groups. In
some cases, acid-tolerant taxa which formerly were rare, may become
abundant.
The effects of acidification on ecosystem processes such as primary
production, energy transfer between trophic levels, and nutrient
cycling have not been well studied and should be addressed in future
research efforts.
5.10.2 Processes and Mechanisms by Which Acidification Alters Aquatic
b. cosy steins
5.10.2.1 Direct Effects of Hydrogen Ions—Effects of low pH on aquatic
organisms are the best studied aspect of the acidification process. Numerous
laboratory bioassays have documented both the toxicity of hydrogen ions to
aquatic organisms and differences in sensitivity to acid stress among taxo-
nomic groups. These studies provide insight into physiological mechanisms of
toxicity and offer guidelines for predicting effects of various pH levels on
aquatic biota. Mechanisms by which various taxa are affected by low pH have
been discussed elsewhere (Section 5.3 through 5.6; Fromm 1980) and include
disruptions in ion transport, acid-base balance, osmoregulation, and enzyme
function. Low pH stress seldom exists alone in acidified waters and thus its
effect on aquatic organisms will be influenced by other stresses (Sections
5-155
-------
5.10.2.2, 5.10.2.4, 5.10.2.7) and biological interactions (Section
5.10.2.3).
5.10.2.2 Elevated Metal Concentrations—The acidification process has
resulted in elevated concentrations of aluminum and other metals in many
waters (Chapter E-4, Section 4-6). Aluminum leached from the soil in
response to acidic deposition has been implicated in fish kills in field
observations, field experiments, and laboratory studies (Section 5.6.4.2).
The interaction of acidity and aluminum is especially important as fish may
be killed by aluminum at a pH value not considered harmful by itself. The
toxicity of aluminum is greatest in the pH range 4.5 to 5.5.
In laboratory experiments, aluminum precipitates phosphorus from water, with
the greatest effect occurring in the pH range 5.0 to 6.0 (Aimer et al . 1978).
Phosphorus is the nutrient that typically limits plant growth in oligotrophic
lakes. While increased aluminum due to acidification would be expected to
reduce phosphorus concentrations and thereby reduce productivity, this
process has not been confirmed by in-lake studies.
Aluminum concentrations are higher in macrophytes from acidified lakes than
in macrophytes from non-acidified lakes. The biological significance of
these higher aluminum concentrations is not known.
High mercury concentrations in fish are correlated with low pH levels for
lakes in Sweden, Ontario, and the Adirondack Mountains of New York (Section
5.6.2.5). In laboratory experiments, biological uptake of most metals is
enhanced at low pH, but whether lake acidification will significantly enhance
bioaccumulation of mercury has not been definitively demonstrated.
Furthermore, there is considerable variation in fish mercury concentrations
between lakes and not all acidified lakes contain fish with elevated mercury
concentrations. Other factors, in addition to pH, which may contribute to
between-lake variability of fish mercury concentrations include dissolved
organic carbon, conductivity, bioproductivity, and watershed geology.
Other metals which consistently exhibit increased concentrations in acidic
surface waters are manganese and zinc (Chapter E-4, Section 4.6.1). Currently
available toxicity data indicate that concentrations of these metals in
acidic surface waters (unless local sources of metal emissions exist) are
below toxic levels. However, a lack of sufficient bioassay data collected in
soft, acidic waters and the potential for additive or synergistic effects
with other toxic components make this statment tentative.
5-10.2.3 Altered Trophic-Level Interactions—The loss of fish from acidified
lakes has beendocumentedTnScandinavia, Canada, and the United States
(Section 5.6.2.1). As the top predators in aquatic habitats, fish are known
to exert control over trophic structure, trophic dynamics, and nutrient
cycling in lakes (Brooks and Dodson 1965, Shapiro et al. 1975, Kitchell et
al. 1979, Clepper 1979, Zaret 1980). For example, zooplanktivorous fish, by
influencing the species composition and size distribution of zooplankton, can
alter the rate of primary production in lakes (Shapiro et al. 1975).
5-156
-------
Changes in aquatic ecosystems following the loss of fish populations are
evident in non-acidified lakes where fish have been intentionally removed
(Stenson et al. 1978, Eriksson et al. 1980a, Henrikson et al. 1980a,b).
Large invertebrate predators (e.g., corixids, dytiscid beetles, Chaoburus)
normally kept at low abundance by fish predation become abundant.
Zooplankton community composition changes and dinoflagellates become dominant
among the phytoplankton. Many of these same changes have been observed in
lakes which have lost their fish populations as a result of acidification.
Thus, biological and limnological changes in a complex aquatic ecosystem
undergoing acidification may be difficult to ascribe directly to the toxicity
of increased acidity or metal concentration. Understanding the role of
trophic-level interactions in producing biological changes during
acidification will require holistic, manipulative studies of consumer
regulation of ecosystem dynamics.
5.10.2.4 Altered Water Clarity--Mater clarity typically increases with
increased acidity (Section 5.5.4.2 and Chapter E-4, Section 4.6.3.4). This
may be due to a reduction in algal biomass in the water column, the precipi-
tation of dissolved organics by aluminum, or changes in the light-absorption
capacity of aquatic humic materials. Increased light penetration would allow
macrophyte and phytoplankton growth at greater depths and would warm the
water to a greater depth.
5.10.2.5 Altered Decomposition of Organic Matter—Decomposition of organic
material releases nutrients for reuse by plants. Reductions in decomposition
rates have been reported in some acidified lakes as a result of decreased
bacterial metabolic rates and declines in populations of shredding inverte-
brates. It has been suggested that decreases in nutrient recycling as a
result of decreased decomposition would lead to decreased productivity at all
trophic levels, but this hypothesis has not been adequately tested nor have
consistent decreases in productivity been observed.
5.10.2.6 Presence of Algal Mats—Algal mats which cover the lake bottom down
to the limit of light penetration are characteristic of acidified lakes.
While these mats would be expected to interfere with water column-sediment
interactions important in the recycling of nutrients, this hypothesis has not
been experimentally tested. The degree to which the physical alteration of
the bottom substrate affects benthic invertebrates and fish is unknown.
5.10.2.7 Altered Nutrient Availability—Increased aluminum concentrations
could decrease the concentration of phosphorus via precipitation of aluminum-
phosphorus complexes. Reducing phosphorus availability should decrease
biological production but this result needs to be quantitatively evaluated.
Nitrogen added via acidic deposition is used as a nutrient, but overall
biological effects on production would be negligible since phosphorus is the
limiting nutrient in most oligotrophic waters.
5.10.2.8 Interaction of Stresses—Predicting the response of a particular
lake or stream to acidification is difficult because acidification results in
many limnological changes besides increased acidity. These changes interact
with biotic responses in complex and often counterbalancing ways. This is
illustrated by the response of the phytoplankton to acidification.
5-157
-------
Phytoplankton biomass and productivity have shown increases, decreases, or no
change with respect to decreasing pH (Section 5.8). Certain types of algae
(dinoflagellates) are frequently dominant in acidic lakes, yet exceptions are
not uncommon. Algal species that are rare one year may dominate a lake the
following year (Van and Stokes 1978). Variation in the response of plankton
communities to acidification may result from the interaction of many factors.
Acidification eliminates sensitive algal species, may decrease phosphorus and
inorganic carbon concentrations, and may depress nutrient cycling. These
changes would tend to decrease phytoplankton biomass and productivity. Yet
acidification may increase water clarity, allowing light to penetrate into
the thermocline and hypolimnion, where nutrient levels are generally higher.
This would tend to increase productivity. Zooplankton are similarly affected
by numerous factors besides pH, including changes in their food supply and
the loss of fish predators.
The response of fish to acidification is likewise complicated. Aluminum and
hydrogen ions interact to cause fish mortalities. Yet this interaction may
be most important during short time periods (e.g., spring snowmelt) and may
not be detected during stream or lake surveys done at other times of the
year. Laboratory experiments predict decreased fish growth in acidified
waters (Section 5.6.4.1.3), yet increased fish growth has been observed in
the field. The reason may be that the increased metabolic demands at low pH
are outweighed by the greater abundance of forage organisms available to a
continually dwindling fish population. Reproductive failures, not decreased
growth, the loss of food items, or adult mortality, appear responsible for
most fish extinctions.
Contradictory responses should not be interpreted as evidence that acidifi-
cation has no effect, but rather as an indication that poorly understood
interactions among stresses may be involved. The infrequency of manipula-
tive, whole-system experiments has contributed to this lack of resolution.
5.10.3 Biological Mitigation
Techniques for mitigating the effects of acidification on aquatic organisms
include base additions to neutralize the acidity (Section 5.9.1 and Chapter
E-4, Section 4.7.1) and development of acid-tolerant fish strains (Section
5.9.2). Immediately after base addition dramatic reductions in
phytoplankton, zooplankton, and benthic fauna have been reserved. However,
the long-term consequence of lake neutralization, provided that
reacidification is not allowed to occur, is repopulation by aquatic organisms
and an environment that is more hospitable for fish.
Fish survival in acidic waters may be enhanced by genetic screening, selec-
tive breeding, and acclimation. These techniques appear to be a feasible
strategy for maintaining a sport fishery in waters acidified to the point
where a natural fishery is no longer possible. It is doubtful, however, that
they could be used to reestablish naturally-reproducing fish populations and
they do not address the problem of restoring other components of the biota to
preacidified conditions. Because of the potential for increased metal
concentrations in fish from acidified waters (Section 5.6.2.5), fish stocked
in such waters should be monitored for toxic metal accumulation.
5-158
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5.10.4 Summary
Biological effects due to acidification occur for a few species near pH 6.0
(Table 5-16). Because the biological response to acidification is a graded
one, continuing pH declines below pH 6.0 will result in escalating biological
changes, with many species adversely affected in the range pH 5.0 to 5.5.
Long-term declines in pH, commonly to pH 4.5 to 5.0, have been observed for a
number of lakes and streams in areas receiving acidic deposition (Chapter
E-4, Sections 4.4.3.1.2 and 4.4.3.2.2). For the same waters, historical data
and paleolimnological analyses indicate that pH levels were often mid 5's or
higher prior to acidification. In addition, episodic depressions down to pH
4.4 to 4.9 often occur in low alkalinity waters during periods of snowmelt
and heavy rainfall and can affect systems with a pH as high as 7.0 (Table
4-4, Chapter E-4). These pH levels, along with other changes associated with
the acidification process (e.g., increased aluminum clarity, accumulation of
detritus and algal mats), will have significant harmful effects on aquatic
organisms. In waters where pH values average below 5.0, most fish species,
virtually all molluscs, and many groups of benthic invertebrates will be
eliminated Increased aluminum concentrations may eliminate fish species
otherwise tolerant of low pH. The plankton community will be simplified and
dominated by a few acid-tolerant taxa. Benthic algal mats will often cover
the lake bottom, and water clarity may increase. These represent the best
documented effects of acidification. Effects on ecosystem processes remain
largely unconfirmed and are an important area for future research efforts.
5-159
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS
E-6. INDIRECT EFFECTS ON HEALTH
6.1 INTRODUCTION (T. W. Clarkson)
Indirect effects on health that may be causally related to acidic deposition
have not been demonstrated in human populations. This lack of documented
effects may mean that no such effects exist in individuals or populations.
On the other hand, interest in the phenomenon of acidic deposition is recent
and few investigations, if any, have been made into the possibility of
indirect health effects. In principle, acidic deposition may influence human
exposure to toxic chemicals via two main pathways: the accumulation of
chemicals in food chains leading to man and the contamination of drinking
water. The format of this chapter is organized according to these exposure
pathways, i.e., Food Chain Dynamics (Section 6.2) and Ground, Surface and
Cistern Waters (Section 6.3).
The substances requiring special attention are methyl mercury, due to its
accumulation in aquatic food chains, and lead, due to the potential for
contaminating drinking water. Aluminum is a special case as its presence at
elevated concentrations in water used in dialysis therapy may cause brain
damage. Other elements and chemicals will only be briefly mentioned as
information is limited. These include arsenic, asbestos, cadmium, copper,
and nickel. Furthermore, reference will be made to other metals and elements
that may interact with mercury, lead, and aluminum to modify human exposure
and toxicity.
6.2 FOOD CHAIN DYNAMICS (T. W. Clarkson)
6.2.1 Introduction
Human exposure could result from bioaccumulation processes. Aquatic
organisms, particularly predatory fish at the top of the food chain, may
concentrate certain toxic elements, leading to substantial human exposure as
in the case of mercury. Accumulation may occur in wildlife that is in
contact with the contaminated water or consumes aquatic organisms. Water
used for irrigation could lead to contamination of edible vegetation.
Concentrations of toxic elements in meat, eggs, and diary products could be
produced by contamination of livestock. This could occur from drinking water
or from contamination of livestock food.
Each of these potential bioaccumulation pathways to humans should be con-
sidered in light of possible health hazards. Data, however, are very limited
with regard to measurement of the toxic elements and to the kinetics of
transfer and uptake in bioaccumulation processes. This discussion will,
6-1
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therefore, be limited to only a few toxic elements and the major pathways of
exposure.
6.2.2 Availability and Bloaccumulation of Toxic Metals
Mercury and its compounds have been extensively studied in terms of avail-
ability and bioaccumulation. The impetus for this work came from a discovery
made in the late 1960"s (see below) that inorganic mercury may be methylated
in the aquatic environment to the highly neurotoxic species, methyl mercury,
and thereby accumulate in aquatic food chains leading to man. Mercury is the
most dramatic example of a change in speciation produced in the environment
that ultimately leads to increased levels in human food. Alkylation of
certain other toxic metals may also occur in the environment (Wood 1974).
Organic forms of arsenic are known to accumulate in shellfish but organic
arsenic is much less toxic to man and animals than the inorganic species.
Cadmium accumulates in plants and certain marine Crustacea, although the role
of aquatic acidification in these accumulation processes is not well
documented. In short, this section will deal primarily with our knowledge
concerning the bioaccumulation of methyl mercury in aquatic food chains and
the possible role of acidification. Other metals and elements will be
discussed briefly as a group.
6.2.2.1 Speciation (Mercury)—The different chemical and physical forms of
mercury each have their own distinctive biological activity (for a detailed
review, see Carty and Malone 1979). Each differs from the others in the
extent of bioaccumulation in food chains and in toxicity to human life. The
speciation of mercury in natural bodies of water is, therefore, an important
consideration in assessing potential hazard to man.
Mercury exists in a variety of physical and chemical forms. The inorganic
forms have three oxidation states: Hg° or "metallic" mercury is in the zero
oxidation state. It is a liquid metal ("quicksilver") and possesses a high
vapor pressure. The vapor is a monatomic gas, is highly diffusible, and
possesses a low solubility in water. It is commonly referred to as "mercury
vapor" despite the fact that certain other forms of mercury (e.g., dimethyl
mercury) also readily vaporize. If Hg° is produced in aquatic bodies of
water, it will readily diffuse into the atmosphere.
Mercury vapor in the presence of water and oxygen is readily oxidized to the
first oxidation state Hg22+, called mercurous mercury and to the second
oxidation state, Hg2+, known as mercuric mercury. Indeed, the inter-
conversion of these three oxidations states via the disproportionate
reaction
Hg22+ t Hg2+ + Hg°
is an important reaction in the environmental transport of mercury (Wood
1974). The direction of the reaction is affected not only by the relative
concentrations of the three species of mercury but by the ambient redox
potential and by certain microorganisms capable of reducing Hg2+ to Hg°
(Wood 1974).
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Most mercurous salts of mercury possess a low solubility in water.
Furthermore, the mercurous action disproportionates to Hg° and Hg2+ in the
presence of protein and other substances containing ligands having a high
affinity for Hg2+. Thus, inorganic mercury in the environment tends either
to be present as Hg° (usually as the vapor) or Hg2+.
The mercuric cations are capable of forming a wide variety of chelates and
complexes with electron donating groups (ligands). For example, four com-
plexes are formed with chloride anions--HgCl+, HgCl^, HgCl3~, and
HgCl42~. The mercuric cation possesses such high affinities for many
organic ligands expected to be present in sediments, water, and aquatic biota
that it is unlikely that the free cations, Hg2+, will ever be detected in
measurable quantities. Its highest affinity is for sulfur anions $2-, S-H,
and the sulfhydryl anion in proteins and amino acids, R-S~, where the
affinity constants are usually in the range of 10 to 20. It is not
surprising, therefore, that the naturally-occurring ore of mercury, cinnabar,
is the sulfide complex HgS. The reaction of Hg2+ with sulfide ions is
important in the geochemical cycles of mercury (see below). Mercuric sulfide
is highly insoluble in water, (solubility product 10-53 M) t So reaction of
mercury with sulfides in water and sediments leads to immobilization of the
metal. However, in the presence of well-oxygenated water (Jensen and
Jernelov 1972) and also in the presence of aerobes, HgS can be oxidized to
the much more soluble sulfite and sulfate salts, thus leading to remobili-
zation of mercury (see below).
Mercuric mercury can form a wider variety of organometallic compounds in
which the mercuric atom is linked covalently with at least one carbon atom.
These organometallic compounds are usually referred to as "organic mercury."
Phenyl mercury has long been used as a fungicide in the paint industry and as
a slimicide in the paper pulp industry. The latter use led to contamination
of many bodies of freshwater in Europe and North America, and its use has now
been banned. Phenyl mercury may be broken down rapidly to inorganic mercury
(Hg2+) by microorganisms present in the aquatic environment and by enzymes
in mammalian tissues. It has a low toxicity to man.
Methyl mercury possesses unique environmental and toxicological properties
that make it the most dangerous mercury compound to human health and one of
the most hazardous chemicals found in the natural environment. Methyl
mercury is known to be produced by methylation of inorganic (Hg2+) mercury
by methanogenic bacteria present in sediments in natural bodies of water (for
review, see Wood 1974). It is readily accumulated in fish and attains the
highest concentration in species of predatory fish. Like Hg2+, it has a
high affinity for organic ligands, prticularly the sulfhydryl anion in
proteins. It appears to have a low toxicity to fish and other aquatic
species but is highly toxic to the human central nervous system (see Section
6.2.4.2).
Dimethyl mercury (CH3)2Hg is also produced by methanogenic bacteria.
Like mercury vapor, it possesses a low solubility in water and has a high
vapor pressure. Thus, dimethyl mercury tends to escape from the aquatic
system into the atmosphere, where it may be broken down by sunlight to Hg°
and methyl free radicals.
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6.2.2.2 Concentrations and Speciations In Water (Mercury)--The early
findings of Stock and Cucuel(1934)that rainwater contains mercury between
50 to 500 ng Hg £-1 is generally supported by more recent findings.
Brune (1969) reported average values of approximately 300 ng Hg s,-l in
Sweden, and Eriksson (1967) also in Sweden found most samples of rainwater in
the range of 0 to 200 ng £-1.
Values for snow depend greatly on the collection conditions and how long the
snow has laid on the ground. Straby (1968) found values of 80 ng g-1 in
fresh snow, but values as high as 400 to 500 ng Hg g-1 were found in snow
samples that had partly melted and evaporated over the winter. Analysis of
the samples deposited in Greenland prior to the 1900s yielded values of 60 ng
g-1 (Weiss et al. 1971).
Bodies of freshwater for which there is no known source of contamination
generally yield values less than 200 ng £-1. Most values fall in the
range of 10 to 40 ng £-1 and drinking water usually has values less than
30 ng £-1 (WHO 1976).
Few reports exist on the speciation of mercury in water, probably because of
analytical difficulties. A recent review by McLean et al. (1980) found that
methyl mercury accounted for a small fraction of the total of the order
of 1 percent. However, a more recent report by Kudo et al. (1982) found that
methyl mercury accounted for about 30 percent of total mercury in samples
taken from Canadian and Japanese rivers. Mercuric mercury (Hg2+) accounted
for about 50 percent.
Two important conclusions may be drawn from these data. First, that precipi-
tation is an important source of mercury to freshwater (see next section),
and second, that mercury in drinking water offers no health threat.
Concentrations on the order of a few hundred nanograms per liter would result
in a negligible intake of mercury on the assumed intake of two liters per day
(U.S. EPA 1980a). This intake, less than 2 yg day-1, is well below the
advised maximum safe intake of 30 ug Hg day-i (WHO 1972b); thus,
additional mobilization of mercury into water by acidic deposition should not
pose a health threat in terms of contaminated drinking water.
6.2.2.3 Flow of Mercury in the Environment—This topic has been the subject
of a number of reviews (WHO 1976, MAS 1978, U.S. EPA 1980a) and will be
briefly summarized here. The subject is one of intensive research, parti-
cularly by the Coal-Health-Environment Project (KHM 1981) in Sweden. This
topic's development is hampered by the need for more sensitive and more
specific methods for measuring the various physical and chemical species of
mercury believed to be present at extremely low concentrations in the
atmosphere and in bodies of natural water.
6.2.2.3.1 Global cycles. The global cycles of mercury have recently been
reviewed by Nriagu (1979) and by the National Academy of Sciences (1978).
The global cycle of mercury involves degassing of the element from the
Earth's crust and evaporation from natural bodies of water, atmospheric
transport believed to be mainly in the form of mercury vapor, and deposition
of mercury back onto land and water. Mercury ultimately finds its way to
6-4
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sediments in water, particularly to oceanic sediments where the carry-over is
very slow. The ocean and oceanic sediments are believed to be the ultimate
destination of mercury in the global cycle.
Andren and Nriagu (1979) have indicated that mercury's residence time in the
atmosphere may vary from approximately 6 to 90 days. Residence times of
mercury in soils are on the order of 1000 years, in oceans on the order of
2000 years, and in sediments on the order of millions of years.
Estimates of the quantities of mercury entering the atmosphere from degassing
of the surface of the planet vary widely, but a commonly quoted figure is
30,000 tons yr-1 (Table 6-1). Estimates of the proportion of the mercury
in the atmosphere due to anthropogenic sources vary greatly; figures from 10
percent to 80 percent of atmospheric mercury have been credited to man.
Estimates of the yearly amount of mercury finding its way to the ocean
indicate that atmospheric deposition accounts for the major amount, approxi-
mately 11,000 tons yr-1, with land runoff accounting for about 5,000 tons
yr-1.
The measurement of mercury in extremely low environmental concentrations is
frequently close to the limit of detection of many current methods. With
this caveat, it would appear that the vastly predominant reservoir for
mercury is the ocean water, containing on the order of 40 million tons (Table
6-2). In contrast, the atmosphere and freshwater contain much less. As one
might expect, therefore, the impact of man-made release of mercury is much
greater on these smaller reservoirs, especially those to which man-made
release is direct. Thus, the impact on levels of atmospheric mercury and
mercury in freshwaters is appreciable, whereas it is estimated that oceanic
concentrations have not appreciably changed in recent history. For example,
it is estimated that the mercury content of lakes and rivers may be increased
by a factor of 2 to 4 due to man-made release (Nriagu 1979).
6.2.2.3.2 Biogeochemical cycles of mercury. This overall global cycle of
mercury results from extremely complex physical, chemical, and biochemical
processes occurring in the main reservoirs and interfaces between these
reservoirs. Most of these processes are poorly understood; nevertheless,
certain very important fundamental discoveries have been made in recent years
and are summarized below.
The most important single discovery in understanding the chemical and
biogeochemical cycles of mercury in the environment was made by Swedish
investigators in the 1960s (for a review see MAS 1978, Nriagu 1979). An
intensive investigation into the source of the methyl mercury compound in
freshwater fish revealed that microbial activity in aquatic sediments can
result in the methylation of inorganic mercury (Jensen and Jernelov 1967).
The most probable mechanism involves the non-enzymatic methylation of
mercuric mercury ions by methyl-carboning compounds (Vitamin BI?) that are
produced as a result of bacterial synthesis. However, other pathways, both
enzymatic and non-enzymatic, may play a role (Beijer and Jernelov 1979).
The methylation of ionic mercury in the environment appears to occur under a
variety of conditions: in both aerobic and anaerobic waters; in the presence
6-5
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TABLE 6-1. SOURCES OF MERCURY IN THE ENVIRONMENT 1971
(WHO 1976, NRIAGU 1979)
Source Amount
Metric tons yr~l
Natural
degassing of earth's crust ~ 30,000
Anthropogenic
worldwide mining 10,000
combustion of coal 3,000
combustion of oil 400-1500
smelting of metal sulfide ores 1,500
steel cement phosphates 500
6-6
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TABLE 6-2. THE AMOUNT OF MERCURY IN SOME GLOBAL RESERVOIRS (NAS 1978)
Reservoir Mercury Content
(metric tons)
Atmosphere 850
Fresh water 2,000
Freshwater biota 400
Ocean water 41,000,000
Oceanic biota 200,000
6-7
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of various types of mlcroblal populations, both anaerobes and aerobes; and in
different types of freshwater bodies such as both eutrophic and oligotrophic
lakes.
The methylation of mercury can result in a formation of either monomethyl or
dimethyl mercury compounds (Figure 6-1). The monomethyl mercury compound is
avidly accumulated by fish and shellfish, whereas the dimethyl compound,
having a low solubility and high volatility, tends to vaporize from the water
phase to the atmosphere where it may be subjected to photolytic decomposition
(Figure 6-1).
However, these reactions are not understood in detail and there does not
appear to be general agreement in the literature as to those conditions that
favor the formation of the monomethyl or the dimethyl form; neither is there
complete agreement as to the extent that the dimethyl species actually
vaporizes from the water phase into the atmosphere.
Methyl mercury compounds are subject to decomposition in the water phase
probably by the action of a variety of microorganisms. These demethylation
microbes appear to be widespread in the environment, occurring in water
sediments and soils and in the gastrointestinal tract of mammals, including
humans. This biogeochemical cycle involving bacterial methylation and
demethylation is part of a more general cycle of mercury that describes
global transport of mercury. Professor Brosset and colleagues (KHM 1981)
have described a large-scale cycle that has the following aspects.
1) Mercury is introduced to the atmosphere from the ground and water
surfaces. It occurs primarily in the form of mercury vapor (Hg°).
2) The total concentration of mercury diminishes while the proportion
of water-soluble mercury increases as a function of height over the
ground. The origin of the soluble mercury is not yet completely
understood.
3) Water soluble mercury is deposited in wet and dry forms in the
water phase of terrestrial and aquatic systems and probably in
other phases if the mercury compounds are soluble in those phases.
4) The deposited forms of water-soluble mercury, once in the water or
terrestrial phase, partly undergo reduction to Hg°, and are partly
absorbed temporarily or permanently on sediments.
5) The rates of deposition into and removal from the water phases
determine the steady-state levels of each mercury species in
water.
6) The concentration of each mercury species in the water phase
determines the concentration on the sediment in contact with the
water phase.
7) The reduction product Hg° returns (i.e., is re-emitted) to the
atmosphere.
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AIR
FISH
SHELLFISH
WATER
CH3Hg
Hgl
j)2Hg CH3S-HgCH3
SEDIMENT
BACTERIA
Figure 6-1. The mercury cycle, demonstrating chemical transformation
by chemical and biological processes and the accumulation
of monomethyl mercury by fish. Adapted from NAS (1978).
6-9
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Neither the detailed chemical mechanisms nor the kinetics of these processes
are understood at this time; for example, the extent to which mercury may be
deposited and re-emitted from water or land surfaces to the atmosphere is
still not understood in quantitative terms. Nevertheless, the general
picture that emerges is one in which long distance transport of mercury in
the vapor phase is possible, its deposition in water and re-emission probably
occurs extensively, and the chemical conversion of mercury from the elemental
to the ionic and to the organic forms is much more extensive than was
originally believed. Therefore, methyl mercury may occur not only as a
result of microbial action in aquatic sediments as indicated in Figure 6-1
but may have a more general source, including the atmosphere.
6.2.3 Accumulation in Fish (T. W. Clarkson and J. P. Baker)
Once methyl mercury enters the water phase as a soluble compound, it is
rapidly accumulated by most aquatic biota and attains highest concentrations
in the tissues of large carnivorous fish. Indeed, it is generally believed
that the major amount of methyl mercury compounds in bodies of water are
contained in the biomass of the system. The bioconcentration factors, that
is, the ratio of the concentration of methyl mercury in fish tissue to
concentrations in fresh water can be extremely large, usually on the order of
10,000 to 100,000 (U.S. EPA 1980a).
In principle, fish can accumulate methyl mercury both directly from water and
from the food supply. Hultberg and Hasselrot (1981) have reviewed available
data and suggested that pike obtain virtually all their methyl mercury from
their food supply. Methyl mercury concentrations correlate well between
different trophic levels of fish and other aquatic organisms, implying the
importance of the food chain. In a survey of several lakes, levels of methyl
mercury in pike were closely correlated (r = 0.92) with methyl mercury
concentration in plankton (Hultberg and Hasselrot 1981). Thus, factors that
affect bioaccumulation of methyl mercury at this early stage of the food
chain should also affect methyl mercury levels at the highest level (e.g., in
predatory fish).
The concentration of methyl mercury in fish tissue is of special interest in
terms of human exposure. Bioaccumulation of methyl mercury in fish is the
main if not the sole source of human exposure, barring episodes of accidental
discharge or misuse of man-made methyl mercury compounds. Thus, factors that
affect concentrations of methyl mercury in edible fish tissue are of con-
siderable importance in assessing potential human health risks from this form
of mercury.
6.2.3.1 Factors Affecting Mercury Concentrations in Fish—In general, for
any body of water one might expect to see an eventual steady-state distrib-
ution of methyl mercury--a balance of synthetic and degradation reactions.
Concentrations of methyl mercury in sediment, water, and biomass at steady
state are influenced by a wide variety of experimental conditions, perhaps
only a few of which have so far been identified. No detailed review will be
given in this chapter, but the reader is referred to other references that
give a more specific treatment of this topic (kriagu 1979).
6-10
-------
Theoretical considerations, experimental data, and observations in field
studies have indicated or suggested that methyl mercury concentrations in
fish are affected by: (1) the species of fish, (2) the age of the fish, (3)
concentrations of mercury in surface sediments and/or in water, (4) the
biomass or biomass production index, (5) salinity, (6) concentrations of
dissolved organics, (7) the microbial activity associated with sediments, (8)
the degrees of oxygenation of water and redox potential, and (9) the pH
and/or alkalinity of water (Hultberg and Hasselrot 1981, Jensen and Jernelov
1972, Fagerstrom and Jernelov 1971, Jernelov 1980). This list is not
exhaustive and, indeed, recent evidence suggests that other as yet unknown
factors are involved (for discussion see Hultberg and Hasselrot 1981). In
view of the current interest in the relationship between the use of fossil
fuels, particularly coal, and possible acidification of large bodies of fresh
water, the influence of aquatic pH on levels of methyl mercury in fish will
be given special attention here.
An indirect result of acidification of surface waters may be increased
accumulation of mercury (and perhaps other metals) in fish. Evidence for
this relationship derives from correlations between metal concentrations in
fish and lake and stream pH levels, and from evalutions of metal chemistry
and availability in oligotrophic, acidic waters.
Elevated levels of mercury in fish from acidic waters have been measured in
Sweden, Norway, Ontario, and the Adirondack region of New York (Hultberg and
Hasselrot 1981, Overrein et al. 1980, Suns et al. 1980, Jernelov 1980,
Schofield 1978). In each case, although fish mercury content was statis-
tically correlated with pH level, the data points still exhibit significant
scatter. At any particular pH level, for a given age and species of fish,
the range observed between lakes in values of mg Hg kg-1 flesh was
considerable, even to the extent that not all lakes with low pH exhibited
elevated mercury concentrations in fish and some lakes without low pH had
fish with high mercury content. Obviously, other factors in addition to pH
control the accumulation of mercury in fish as noted above. Waters of low
productivity (oligotrophic lakes) and low alkalinity tend to be more sensi-
tive to mercury contamination and mercury accumulation in fish. Because
these conditions are also strongly associated with low pH levels, the effect
of pH on mercury bioaccumulation may be somewhat confounded. The correlation
between pH and fish mercury content may in part be a result of the observa-
tion that low pH waters tend to be oligotrophic softwaters with low
alkalinities. On the other hand, the association between low alkalinity and
elevated mercury content may directly reflect that pH influences mercury
accumulation and low pH waters have low alkalinities. Results from these
correlations must be interpreted carefully.
The most extensive studies on factors controlling mercury levels in fish have
been carried out in Sweden. In the 1960's pike and other edible fish were
found to have unacceptably high levels of mercury (greater than 1 yg Hg
g"1, FDA 1979). For some lakes, local industrial "mercury emitters" with
direct outlets to the lakes were identified as the cause. Many lakes,
however, had inexplicably high mercury levels in fish. This led to extensive
studies in Sweden on the dynamics of mercury chemistry and uptake by fish and
the role of acidity in these processes.
6-11
-------
Data collected by Jernelov et al. (1975), Grahn et al. (1976), Landner and
Larsson (1972), and Hultberg and Jernelov (1976), as reported by Jernelov
(1980), all Indicated an overall strong correlation between mercury levels in
fish and pH values of lakes. Jernelov (1980) concluded that in Swedish lakes
in general, extremely few lakes with pH values below 5.0 have pike (weighing
1 kg) with mercury concentrations of less than 1 mg kg-1. At a pH value of
6.0. the normal level for the same pike would be approximately 0.6 mg
kg-1.
Hultberg and Hasselrot (1981) reviewed ten years of Swedish work on factors
affecting mercury in fish. In a study involving over 152 Swedish lakes
mercury level in pike muscle was inversely correlated with water pH (Figure
6-2). Water samples collected during the fall overturn were analyzed for pH,
humic material (water color at an adjusted pH), and specific conductivity
(salt content). Multiple linear regression analysis (Table 6-3) suggested
that a one unit decrease in pH would elevate mercury in the muscle tissue of
pike (weighing 1 kg) by 0.14 ppm. The influence of pH on fish mercury
content was generally greater than that associated with humic content or
conductivity.
Hakanson (1980), also using the Swedish data base, developed (based on a
combination of statistics and deductive reasoning) a quantitative model
expressing mercury content in a 1-kg pike as a function of pH, the mercury
content in the top one cm of lake sediments, and a bioproduction index. The
model was validated using an independent data set from 107 Swedish lakes.
The correlation coefficient between observed and predicted mercury content
was 0.79.
Hakanson's formula was as follows:
4.8 x log (1 + H950)
F(Hg) = 200
(pH-2) x log(BPI)
where
F(Hg) = the concentration of methyl mercury in a 1 kg pike in
yg g-1 wet weight,
Hgso = the weighted mean mercury content of surface sediments,
0 to 1 cm, in ng Hg g-1 ds (ds = dry substance),
pH = the mean pH of the water system, i.e., the mean of
at least five measurements of which two should be
obtained at different seasons, and
BPI = Bioproduction Index - for details, see Hakanson (1980).
6-12
-------
to
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METHYL MERCURY CONCENTRATION IN PIKE MUSCLE
(yg Hg g-1 wet wt.)
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-------
TABLE 6-3. THE RESULTS OF A STATISTICAL ANALYSIS INDICATING THE
CONTRIBUTION OF pH, HUMIC CONTENT AND SPECIFIC CONDUCTIVITY
TO METHYL MERCURY CONCENTRATIONS IN THE MUSCLE TISSUE OF 1 KG PIKE
(ADAPTED FROM HULTBERG AND HUSSELROT 1981)
Decrease In water pH Increase in mercury concentration
mg Hg kg-1
one pH unit 0.14
two pH units 0.28
three pH units 0.42
Color increase
10 mg Pt jr1 0.015
50 mg Pt £-1 0.075
100 mg Pt r1 0.150
Increase in specific conductivity
5 mS nrj 0.075
10 mS m-J 0.150
20 mS m-1 0.300
6-14
-------
Calculations based on the Hakanson formula yield results similar to those
from Hultberg and Hasselrot (1981). For example, if it is assumed that a 1
kg pike at pH 6.0 contains 0.75 ppm Hg (e.g., Figure 6-2), then a pH change
from 6.0 to 5.0 would increase fish mercury concentration by approximately
0.13 ppm. Overlap in the data bases used by both Hakanson and Hultberg-
Hasselrot may have occurred, however, accounting in part for this close
agreement.
If the Hakanson formula is valid, then
appropriateness of linear regression
concentration (e.g., in Figure 6-2 and
Table 6-3). The Hakanson formula has
hyperbole:
a question might be raised on the
analyses relating pH to mercury
the multilinear analysis used for
the general form of a rectangular
F(Hg) = —
pH - 3
where Hgso and BPI are constant.
Regression analysis of the data in Figure 6-3 according to a hyperbolic
equation yielded a value of the correlation coefficient (r = 0.81) appreci-
ably higher than that obtained by linear regression analysis (r = 0.3). Thus,
for the Swedish study, change in pH accounted for about 80 percent of the
total variance in methyl mercury concentrations in 1 kg pike. The hyperbolic
aspects will become more pronounced at lower pH values and will be discussed
later with regard to apparent scatter of points around linear regression
lines.
Additional, and as yet unknown, factors seem to be operative in determining
mercury concentrations in fish. For example, Hultberg and Hasselrot (1981)
noted that lakes in more northern regions of Sweden tend to have higher
concentrations of mercury in pike. Possible explanations include 1) the
impact of snow on water quality during the spring melt, 2) loss of sensitive
prey species (in this case roach, Rutilus rutilus) adversely affected by acid
episodes during spring melt and a shift to predation on higher trophic levels
(in this case perch, Perca gluvicotilis) that contain greater amounts of
mercury, 3) the importance of snow itself as a source of mercury including
methyl mercury (Brouzes et al. (1977), and 4) lower water temperature and
salinity generally associated with northern latitudes.
In Norway, concentrations of mercury in muscle of trout, perch, char, and
pike were studied by Muniz, Rosseland, and Paus (Overrein et al. 1980).
Again, fish populations in acidic waters generally had higher levels of
mercury than did reference populations from areas without acidified lakes.
Studies in Canada (Suns et al. 1980) have also found a statistically signi-
ficant (r = 0.65, p < 0.05) inverse correlation between water acidity and
mercury levels in fish, for yearling perch in 14 pre-cambrian lakes in
Ontario (Figure 6-3). For lakes with similar pH, mercury levels were higher
in fish from lakes with a higher drainage area/lake volume ratio.
6-15
-------
200
180
160
~ 140
i
en
en
~ 120
z
o
g
!= 100
o
o
80
60
40-
20-
LEGEND
• 1980
O 1981
r *
P <
0.63
0.05
4.5
5.0
5.5
6.0
6.5
7.0
7.5
PH
1.
2.
3.
4.
5.
6.
7.
8.
9.
DUCK LAKE
LITTLE CLEAR LAKE
HARP LAKE
BIGWIND LAKE
NELSON LAKE
CHUB LAKE
CROSSON LAKE
DICKIE LAKE
LEONARD LAKE
10.
11.
12.
13.
14.
15.
16.
17.
18.
HENEY LAKE
CRANBERRY LAKE
HEALEY LAKE
CLEAR LAKE
FAWN LAKE
BRANDY LAKE
McKAY LAKE
LEECH LAKE
MOOT LAKE
Figure 6-3.
Mercury concentrations in yearling yellow perch and
epilimnetic pH in lakes in the Muskoka-Haliburton area
of Ontario (Suns et al. 1980, U.S./Canada 1983).
6-16
-------
Suns et al. (1980) failed to see a relationship between mercury In fish and
water alkalinity, whereas Scheider et al. (1979) reported that for walleye
(Stizostedion vitreum) of equal length, those caught in Ontario lakes with
alkaline water (£ 15 mg CaCOa a*1) had significantly higher mercury
levels than walleye caught in lakes with high alkalinity (> 15 mg CaCOs
£-!). Comparisons based on fish length may, however, be somewhat mis-
leading. If fish from waters with lower alkalinity grow slower (possibly as
a result of lower primary productivity or lower temperatures), than the
higher mercury content at a given length may actually only reflect the older
age of the fish.
Statistical evaluations of mercury in fish and water acidity have not been
published for freshwater fish caught in the United States. A graph of mer-
cury levels in brook trout muscle as a function of fish length for Adirondack
lakes indicated that fish from acid drainage lakes (pH < 5.0) in general had
higher mercury levels (for a given length) than fish from limed, seepage, or
bog lakes (Schofield 1978). However, high mercury level in fish were also
found in some lakes without low pH, indicating that the unusual mercury
bioaccumulation may be, in part or in total, independent of pH level.
In summary, field studies in Sweden, Norway, and Canada have identified
several factors that correlate (positively or negatively) with mercury levels
in fish. This includes fish species and age (length and weight are frequent-
ly used instead of age), mercury levels in surface sediments, the biomass or
bioproductivity of the lake, the salinity (specific conductivity), and pH.
Other factors may also be operative, such as morphometric parameters (drain-
age area/lake volume ratios) and geographic (northern latitude). However, in
virtually all such studies published to date, elevated mercury levels in fish
muscle (most notably pike and perch) have been statistically associated with
higher levels of acidity.
However, a number of factors influencing mercury levels in fish may also
change in parallel with acidity. Thus, a true cause-and-effect relationship
between acidity and elevated mercury in fish has not been established by the
available data. Absolute proof may be unattainable in field studies, given
"the large number of variables and the probability that, in any given field
study, not all of these will be controlled or even measured.
To resolve whether correlations observed between lake pH level and mercury
content in fish actually reflect a cause-and-effect relationship and whether
acidification will enhance bioaccumulation of mercury, the effects of pH and
acidity on mercury chemistry, mobilization, and uptake must be understood.
Field and laboratory research on mercury cycles have resulted in several
proposed mechanisms (Jernelov 1980, Wood 1980, Haines 1981):
1) Acidic precipitation may scavenge mercury from the atmosphere more
effectively than non-acidic precipitation.
2) The rate of methylation of inorganic mercury by microorganisms is
pH-dependent, the maximum occurring at pH 6.0; methylation is
higher from pH 5.0 to 7.0 than above 7.0. Thus, at lower pH more
methyl mercury would be present and, because methyl mercury is the
6-17
-------
form most rapidly taken up by fish, bioaccumulation presumably
would be enhanced.
3) Low pH levels favor the formation of monomethyl mercury rather than
dimethyl mercury. Dimethyl mercury is unstable and volatile and
thus more quickly lost from the aquatic system (Figure 6-1).
4) Under aerobic conditions, inorganic mercury is more soluble at
reduced pH and thus more available for methylation reactions.
Retention of mercury in the water column is enhanced with increased
acidity (Jackson et al. 1980), thus increasing the exposure of fish
to mercury.
5) Since the biomass of fish is often lower in acidic lakes, the
available mercury is concentrated in a smaller biomass, resulting
in higher body burdens per fish. Also, if growth rate is reduced,
fish in an acidic lake would be older than fish of an equivalent
size in a non-acidic lake and would have been accumulating mercury
longer.
Laboratory experiments will be useful, if not essential, in order to unravel
mechanisms associating pH change with mercury accumulation in fish.
Laboratory experiments have shown that, for a given amount of total mercury
in an aquatic ecosystem, higher levels of mercury were found in fish at low
pH values than at high pH values (for review, see Jernelov 1980).
Miller and Akagi (1979) presented experimental evidence that low pH levels
mobilize methyl mercury absorbed on sediments. Natural water from the Ottawa
River was incubated with various types of sediment materials for periods of
approximately three weeks. Irrespective of the type of sediment, a reduction
in water pH shifted, by a factor of 2 for each unit change in pH, the
distribution of methyl mercury from the sediment to the water phase (Figure
6-4). Miller and Akagi (1979) asserted that the effect of pH on the equili-
brium of methyl mercury between water and sediment, may be the principal
factor responsible for higher levels of mercury in fish in low pH aquatic
environments.
That acidification of surface waters will significantly enhance bioaccumu-
lation of mercury has not been definitively demonstrated. The chemistry and
environmental sampling of mercury are extremely complex. More research is
needed to identify all factors that affect mercury accumulation in fish and
the relative importance of each. The significance of a one unit pH decrease
(or a decline in alkalinity by 100 yeq £-!) relative to the effects of
the large number of other factors that influence bioaccumulation has not been
quantified. This need is especially urgent in the United States, where few
data are available at this time.
Other metals in addition to mercury occur at elevated concentrations in
acidic waters and potentially may accumulate in fish and other biota. Data
on these accumulations are, however, very limited. Dickson (1980) reported
that concentrations of cadmium in pike increased with increased acidity.
Harvey et al. (1982) determined manganese concentrations in the vertebrae of
6-18
-------
150
H& 100 -
1*
el
o:
50-
-
,1
LEGEND
03 SAND
E3 SAND CHIP SEDIMENT
D WOOD CHIPS
i
73 MONOMETHYL MERCURY
gH; x;.;::';: I::::;:::;!;:;:;:]
. :-:f" "| •••:•:•.-.: •.-.•.•••••***J 1 -s
6
pH
Figure 6-4. The partition coefficient of methyl mercury between water
and three different types of sediments. The units of the
ordinate have been multiplied by a factor of 10,000. The
data are taken from Miller and Akagi (1979).
6-19
-------
white suckers from six lakes in sourthern Ontario. Fish from the most acidic
lake, George Lake (pH 4.65), had particularly high manganese content. The
remaining five lakes had pH levels from 5.02 to 6.59, and fish manganese
level appeared relatively independent of pH. George Lake also had aqueous
manganese concentrations that were 50 percent greater than in any of the
other lakes. The Ontario Ministry of Environment (U.S./Canada 1983) analyzed
yearling yellow perch for body burdens of lead, cadmium, aluminum, and
manganese in 14 Ontario lakes (Figure 6-5). Lead (p < 0.01) and cadmium (p <
0.05) were significantly correlated with lake pH level. No data are avail-
able to evaluate the environmental significance of these accumulations. No
correlations between lake acidity and body levels of aluminum or manganese
were evident. Aluminum has, however, been observed to accumulate on gills of
fish during fish kills in Plastic Lake, Ontario, and in two lakes in Sweden.
Grahn (1980) measured 40 to 47 P g Al g-1 wet weight of tissue on gills
from dead ciscoe from lakes Ransjon and Amten, Sweden, but only 6 yg Al
g-i for fish from reference lakes without fish kills. Aluminum concen-
trations on fish gills from dead and moribund pumpkinseed and sunfish from
Plastic Lake ranged from 83 to 250 mg g-1 dry weight (Harvey et al. 1982).
6.2.3.2 Historical and Geographic Trends in Mercury Levels in Freshwater
Fish—Presently it is difficult to assess quantitatively the contribution of
acidic deposition to elevations of mercury concentrations in freshwater fish.
The problem in part is a lack of data showing temporal and regional changes
in mercury as related to water pH and in part due to the operation of other
processes affecting mercury levels in fish.
Bloomfield et al. (1980) have reviewed the results of an extensive mercury
screening involving some 3500 freshwater fish collected in New York State
from 1960 to 1972. Less than 10 percent of the fish had mercury levels in
excess of the current federal guideline of 1.0 ppm. A sizeable portion of
the high mercury fish came from Onondaga Lake—known to be polluted by a
local industrial source of mercury. Predatory species of fish such as
walleye, pike, and smallmouth bass had levels sometimes exceeding 1 ppm in
certain Adirondack Lakes remote from known sources of mercury. Bloomfield et
al. (1980) quote unpublished work indicating that concentrations in small-
mouth bass were still high in 1975, and Armstrong and Sloan (1980) reported
elevated mercury levels in predatory fish species collected in certain
Adirondak Lakes (Cranberry, Great Sacandaga, Raquette) in 1978. In contrast,
fish from rivers and lakes previously contaminated with mercury now show
declining fish levels (Armstrong and Sloan 1980). For example, following
cessation of mercury discharge, levels of mercury in smallmouth bass in Lake
Onondaga declined by 55 percent over the period 1972 to 1978. The Ontario
Ministry of Environment (1977) has reported substantial declines in mercury
in fish caught in Lake St. Clair following curtailment of industrial
discharge of mercury.
Based on very limited data in the United States, a general picture emerges of
declining mercury levels in freshwater fish caught in areas where direct
discharge of mercury has been curtailed but of continued high levels of
mercury in certain lakes remote from industrial activity. Reasons for these
high mercury levels are being investigated (Section 6.2.2.3). Wet deposition
6-20
-------
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400
300
200
100
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,
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i i i i i ii
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.5 5.5 6.5 7.5 8.
pH pH
Figure 6-5. Metal concentrations in yearling yellow perch and
epilimnetic pH in 1981 in lakes in the Muskoka-Haliburton
area of Ontario (U.S./Canada 1983).
6-21
-------
of mercury from the atmosphere has been shown to occur in several Adirondack
Lakes. These lakes, in general, are characterized by low pH and low
alkalinity. The role of long distance transport of mercury and lake
acidification merits careful investigation.
6.2.4 Dynamics and Toxicity in Humans (Mercury)
6.2.4.1 Dynamics in Man (Mercury)—The U.S. EPA (1980a) has reviewed in-
formation "^h~lJptal
-------
Methyl mercury readily crosses the placenta! barrier and enters the fetus.
It distributes to all tissues in the fetus, including the fetal brain, which
is the principal target for prenatal toxicity of methyl mercury. Levels of
methyl mercury in cord blood are usually higher than the maternal blood
concentrations.
Methyl mercury is secreted in milk. Thus body burdens of methyl mercury
acquired by the infant before birth may be maintained by breast feeding if
the nursing mother continues to be exposed to methyl mercury.
The rate of elimination of methyl mercury from the human fetus and suckling
infant is not known. Experiments on animals indicate that elimination in
suckling animals is much slower than in adults. The adult rate of excretion
appears to commence at the end of the suckling period.
In brief, methyl mercury accumulates in the human body over a period of about
one year. Blood and hair analyses may be used as indicators of human ab-
sorption of mercury. In assessing hazard to human health, chronic exposure
over weeks or months is important.
6.2.4.2 Toxicity in Man--Methy1 mercury damages primarily the human central
nervous system. When ingested in sufficient amounts, methyl mercury destroys
neuronal cells in certain areas of the brain, the cerebellum and the visual
cortex, resulting in permanent loss of function. Symptoms of damage include
loss of sensation, constriction of the visual fields, and impairment of
hearing. Coordination functions of the brain are also damaged, leading to
ataxia and dysarthria. Severest damage causes mental incapacitation, coma,
and death. The mildest and earliest effect in adults is usually a complaint
of paresthesia, an unusual sensation in the extremities and around the mouth.
In the Japanese population poisoned by methyl mercury from contaminated fish,
paresthesia was usually permanent. In the Iraqi population, paresthesia was
frequently reported to be transient. This population had consumed homemade
bread from wheat contaminated with a methyl mercury fungicide.
The effects on the fetal brain differ qualitatively from those seen in
adults. Methyl mercury interferes with the normal growing processes of the
brain. It inhibits migration of neuronal cells to their final destination,
thus affecting the brain's architecture. This damage manifests itself as
diminished head size (microcephaly) and gross neurological manifestations
such as cerebral palsy. The mildest effects are delayed achievement of
developmental milestones in children and the presence of abnormal reflexes
and mild seizures.
Brain concentrations associated with the onset of human methyl mercury
poisoning are in the range of 1 to 5 yg Hg g-1 wet tissue. Blood
concentrations for the onset of the mildest effects have been established to
be between 200 and 500 ng Hg ml-1 whole blood. Corresponding hair concen-
trations would be 50 to 125 yg Hg g-1 hair (Table 6-4). The chronic
daily intake of methyl mercury that would lead to a maximum blood level of
200 ng ml-1 has been established to be 300 ug Hg. However, in the mother
during pregnancy, the blood level associated with the earliest damage to the
fetus has not yet been determined.
6-23
-------
TABLE 6-4. THE CONCENTRATIONS OF TOTAL MERCURY IN INDICATOR MEDIA AND
METHYL MERCURY ASSOCIATED WITH THE EARLIEST EFFECTS IN THE
MOST SENSITIVE GROUP IN THE ADULT POPULATION3
(ADAPTED FROM WHO 1976)
Concentrations in indicator media
Blood Hair Equivalent long-term daily intake
(ng ml'1) (yg g~M (ug kg'1 body weight)
200 to 500 50 to 125 3 to 7
risk of the earliest effects can be expected to be between 3 to 8
percent, i.e., between 3 to 8 percent of a population having blood levels
in the range 200 to 500 mg ml'l, or hair levels between 50 to 125 yg
g"1 would be expected to be affected (for further details, see text).
6-24
-------
The conclusions reported in Table 6-4 were based on observations of affected
populations in outbreaks of poisoning in Niigata, Japan and in Iraq (the
1971-72 outbreak). In effect, the numbers in Table 6-4 refer to the lowest
effect levels observed in an outbreak of poisoning from methyl mercury
contaminated fish in Niigata, Japan (Swedish Expert Group 1971) and lowest
effect levels estimated from an affected population in the Iraqi outbreak of
1971-72 (Bakir et al. 1973). With such low observed effect levels in humans,
it is usual to apply a safety factor of ten (WHO 1972a) to arrive at an
acceptable "safe" body burden or "allowable daily intake."
A direct estimation of absolute risks associated with a given long-term daily
intake of methyl mercury was reported by Nordberg and Strangert (1976, 1978).
In their approach they combined the data from dose-response relationship
published in the Iraqi outbreak (Bakir et al. 1973) with the range of biolog-
ical half-times, also obtained in the Iraqi outbreak (Al-Shahristani and
Shihab 1974), to calculate the relationship depicted in Figure 6-6.
Their calculations indicated that an intake of 50 yg dayl in an adult
gives a risk of about 0.3 percent of the symptom of paresthesia, whereas an
intake of 300 yg day'1 would give a risk of about 8 percent of symptoms
of paresthesia. As pointed out by Nordberg and Strangert (1976), the back-
ground frequency of these non-specific symptoms such as paresthesia plays a
key role in determining the accuracy of the estimates of response of low
frequencies. They estimated from the same Iraqi data the background frequen-
cy of paresthesia of 6.3 percent. However, there is considerable uncertainty
in determining the precise value of the background frequency, and this
uncertainty becomes the dominant cause of error at low rates of response.
Since the studies on the Iraqi outbreak, a major epidemiological study has
been carried out in Northwestern Quebec on Cree Indians exposed to methyl
mercury in freshwater fish (Methyl Mercury Study Group 1980). The authors
claim to find an association in men over age 30 and women over age 40 of a
set of neurological abnormalities and the estimated exposure to methyl
mercury. However, it should be pointed out that this association has been
seen by only four of seven observers who reviewed video taped recordings of
the neurological screening tests. The severity of these neurological
abnormalities was assessed by neurologists as mild or questionable. It was
not possible to estimate any threshold body burden or hair levels because
this population had been exposed possibly for most of their lives; peak
values in previous years are unknown. However, observations on this
population over several years indicate that maximum blood concentrations are
below 600 ppb and most are below 200 ppb (Wheatley 1979). A WHO expert group
(1980), on examining the reports from these studies, raised the possibility
that this might be the first example of an endemic disease due to exposure to
methyl mercury in freshwater fish. However, another epidemiological and
clinical study of the same population of Cree Indians failed to find any
effects associated with methyl mercury (Kaufman, personal communication to
EPA).
The safety factor of ten applied to the lowest effect levels in Table 6-4 was
intended to take into account inter alia the greater sensitivity of the
6-25
-------
100
I/O
o
CL
l/>
UJ
CC
o
UJ
Q.
X
50
0.1 0.4 1.0 2
DAILY INTAKE (mg)
Figure 6-6. The calculated relationship between frequency of
paresthesia in adults and long-term average daily intake of
methyl mercury. The calculations were performed by
Nordberg and Strangert (1978). The broken line is the
estimated background frequency of paresthesia in the
population. Data are taken from publications on the
Iraqi outbreak of methyl mercury poisoning (Bakir
et al. 1973; Al-Shahristani and Shihab 1974).
6-26
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fetus. Since the WHO evaluation of 1976, data have been published relating
methyl mercury levels in the mother during pregnancy to effects such as
psychomotor retardation in the offspring (Marsh et al. 1980). These data
were the basis of a recent risk estimate (Berlin 1982) relating concentra-
tions of mercury in maternal hair to risk of mental retardation in prenatally
exposed infants (Figure 6-7). Berlin calculated a background frequency of
mental retardation in the Iraqi children of approximately 4 percent as com-
pared to a background frequency in Sweden of 2 percent. He also noted that,
in the case of adults, the error in determining background frequency is
probably the major source of error when researchers look at low rates of
responses. Berlin calculated that there was a risk of doubling the back-
ground frequency of mental retardation at methyl mercury levels in the mother
on the order of 20 ppm in hair and a risk of a 50 percent increase in
background frequency at hair concentrations of about 10 ppm.
The McGill Group (Methyl Mercury Study Group 1980) in their study of Cree
Indians exposed to methyl mercury in fish, found an association "... between
findings on examination of tone and reflexes in Cree boys and the concentra-
tion of methyl mercury in the mothers' hair during pregnancy. This associ-
ation was shown at levels of methyl mercury exposure which are very low in
relationship to those previously reported to be associated with effects of
methyl mercury in utero.... These findings were isolated and the variation
from normal was mild." The highest range of maternal hair concentration was
13 to 23.9 yg g-1.
These hair levels overlap the range estimated by Berlin associated with the
earliest detectable effects in Iraq. However, the association noted in the
McGill study may have been due to chance as their observations on tone and
reflexes were part of a number of observations, the rest of which did not
correlate with mercury levels.
These observations on human infant-mother pairs agree with animal data
indicating the greater sensitivity of prenatal life to methyl mercury (for
review, see Clarkson 1983). However, the risk estimations described in
Figure 6-6 should only be regarded as approximate, as they are based on small
numbers. We greatly need to obtain more precise estimates of human health
risks associated with prenatal exposure to methyl mercury.
6.2.4.3 Human Exposure from Fish and Potential for Health Risks—Dietary
intake accounts for the greatest fraction of total mercury intake by man
(Table 6-5). Methyl mercury intake is exclusively from the diet and almost
entirely from fish and fish products. The evidence comes from dietary
studies showing close correlation of blood levels with fish consumption
(Swedish Expert Group 1971) and from large-scale analyses of food items in
several countries, indicating that significant concentrations of methyl
mercury are found only in fish and fish products (U.S. EPA 1980a).
Based on data from the National Marine Fisheries, Cordle et al. (1978) have
reported a ranking of species of fish according to annual consumption in the
United States (Table 6-6). The table clearly demonstrates that oceanic fish,
especially tuna, account for the major amount consumed. However, when
consumption is expressed according to the consumer use, a different picture
6-27
-------
100 F
a.
o
Q.
o
UJ
O.
Figure 6-7.
30
20
10
10 30
MERCURY IN HAIR (ppm)
50
A dose-response relationship between the frequency of mental
retardation in a population of children prenatally exposed
to methyl mercury and the maximum hair concentrations of the
mothers during pregnancy. The maximum hair concentrations
in the mothers during pregnancy was used as a measure of the
prenatal dose. The curves are drawn according to logit
analysis, assuming the presence of a background frequency.
Figure 6-7A gives the complete dose-response curve. Figure
6-7B gives the low frequency end of the dose-response rela-
tionship, indicating the presence of a background frequency,
i.e., the vertical intercept at zero mercury concentration
in the mothers' hair. The analysis was carried out by Berlin
(1982) on data from the Iraqi outbreak (Marsh et al. 1980).
6-28
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TABLE 6-5. ESTIMATES OF AVERAGE INTAKES OF
MERCURY BY THE "70 kg MAN" IN THE UNITED STATES POPULATION
(ADAPTED FROM U.S. EPA 1980a)
Media Mercury intake yg day-1 70 kg-1 Predominate
(average) form
Air 0.3 Hg°
Water 0.1 Hg2+
Food 3.0 CHsHg+
6-29
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TABLE 6-6. ESTIMATED FISH AND SHELLFISH CONSUMPTION IN THE UNITED
STATES RANKED ACCORDING TO ANNUAL CONSUMPTION FOR THE PERIOD
SEPTEMBER 1973 TO AUGUST 1974 (ADAPTED FROM U.S. EPA
1980a AND CORDLE ET AL. 1978)
Amount
Rank IC* lb yr-1
Total
Tuna (mainly
Canned)
Unclassified
(mainly
breaded,
Including fish
sticks)
Shrimp
Ocean Perchd
Flounder
Clams
Crabs/lobsters
Salmon
Oysters/scallops
Troutf
Codd
Bassf
Catfish*
Had dock d
Pollockd
Herring/smelt
Sardines
Pikef
Halihutd
Snapper
Whiting
All other
classified
1
2
3
4
5
6
7
8
9
9
11
12
12
12
15
16
17
18
18
20
2957
634
542
301
149
144
113
110
101
88
88
78
73
73
73
60
54
35
32
32
25
152
Percent of
total by
weight
100.0
21.4
18.4
10.2
5.0
4.9
3.8
3.7
3.4
3.0
3.0
2.7
2.5
2.5
2.5
2.0
1.8
1.2
1.1
1.1
0.9
5.1
Number of
actual users
(millions)
197.0
130.0
68.0
45.0
19.0
31.0
18.0
13.0
19.0
14.0
9.0
12.0
7.6
7.5
11.0
11.0
10.0
2.5
5.0
4.3
3.2
Mean Amount
per user,
(g dayl)
18.7
6.1
10.0
8.3
9.7
8.6
7.6
10.6
6.7
7.8
12.3
8.1
12.0
12.1
8.6
6.8
6.7
17.4
8.0
9.3
9.7
Average cone.
of mercury
vg Hg g-1*
0.14b
0.27
0.35
c
0.05
0.13
0.10
0.05
0.07-0
0.08
0.03
0.42
0.14
c
0.15
0.11
0.14
0.02
0.61
0.19-0
.14S
.53
0.45-369
c
c
aU.S. Chamber of Commerce (1978).
^Average values for skipjack, yellow fin, and white tuna, respectively.
cData not available.
dflainly Imports.
eKing crab - all others, respectively.
fFresh Water.
9Red Snapper - other.
6-30
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emerges. On this basis, freshwater fish dominate the rankings, with northern
pike consumed at 17.4 g day1, followed by freshwater trout at 12.3 g
day1, bass (freshwater) and catfish at 12.1 g day1. The highest user
consumptions of seafood are crabs and lobster at 10.6 g day1, with tuna
down to 6.1 g day1.
The highest average mercury concentrations are also found in freshwater
fish--pike at 0.61 yg Hg g"1 and trout at 0.42 yg Hg g"1. Thus a
pike consumer would have a daily average intake of methyl mercury of 10.4
yg exclusively from pike, and a trout consumer would have had an average
intake of 5.2 yg Hg. These average values are well below the recommended
maximum safe intake of 30 yg day1.
The National Marine Fisheries developed an extensive data bank on fish
consumption by individuals according to fish species (U.S. Department of
Commerce 1978). These data were based on a Diary Panel Survey of approxi-
mately 25,000 individuals chosen to be representative of the U.S. population.
These data, along with additional information on mercury concentration in
edible tissues of various fish species, allowed a calculation of the number
of individuals who would be expected to exceed the maximum safe daily intake
of 30 yg. It was calculated that 47 individuals would exceed this limit by
a small margin from consumption of fish and that 23 of these were consumers
mainly of freshwater fish. According to calculations by Nordberg and
Strangert (Figure 6-6) the risk at this level of intake will be small--on the
order of 0.3 percent.
The risk of prenatal poisoning cannot be estimated with any precision, given
the small number of cases used in Figure 6-7. The daily intake of about 30
yg Hg roughly corresponds to a hair concentration of 6 to 10 ppm. The
dose-response data in Figure 6-7 would indicate that the background frequency
of mental retardation would be increased by less than 50 percent.
Estimates of increased rates risks due to acid precipitation would depend
upon a number of assumptions, including whether increases in freshwater
acidity would elevate levels of methyl mercury in freshwater fish and by how
much, the effect of acidity on the supply of freshwater fish, as well as
actions taken by local, state and federal agencies to limit fishing and sales
of fish if methyl mercury levels increase. Nevertheless, information on
methyl mercury is now reaching the point where rough estimates can be made of
health risks in this country for consumption of methyl mercury from fresh-
water fish, and information may be forthcoming on the impact of acidity on
methyl mercury levels in fish. At least the direction of future research is
now more clear—to obtain more quantitative information on human dose-
response relationship and to further test hypotheses on cause-effect
relationship between pH and methyl mercury levels in freshwater fish.
6.3 GROUND, SURFACE AND CISTERN WATERS AS AFFECTED BY ACIDIC DEPOSITION
(W. E. Sharpe and T. W. Clarkson)
For reasons given in Section 6.1, this section will deal only with those
metals whose concentrations and/or speciation in drinking water may be
affected by acidic deposition. As discussed in the previous section, mercury
6-31
-------
concentrations, including any potential changes due to pH, should not offer
any conceivable threat to human health. Lead is the one metal of greatest
concern and will be given special attention in this section. Other metals
such as aluminum, cadmium, and copper, will be discussed briefly.
6.3.1 Water Supplies
An understanding of the modes of hydrologic interactions between acid
deposition and various types of water supplies is essential to assessing the
potential indirect health effects to users of drinking water obtained from
such systems. In addition, the physical facilities used to store, treat, and
distribute water are of primary importance, as are the chemical methods used
to treat water prior to use. Principal water sources in continental North
America are usually either surface or groundwater, with other sources such as
direct use of precipitation of much lesser importance. Health risk is
directly related to the source of drinking water.
Health risk in drinking water supplies is also closely related to the manage-
ment of the drinking water supply. Risks are generally greater the smaller
the water supply, with small privately owned water systems serving a single
dwelling at greatest risk. These systems typically do not routinely monitor
water quality nor do they provide even rudimentary water treatment. Data on
the impacts of atmospheric deposition on drinking water quality are extremely
scarce; however, by using available information on the impacts to surface
water aquatic ecosystems, we may assess impacts.
6.3.1.1 Direct Use of Precipitation (Cisterns)—The direct use of precipita-
tion by collection in artificial catchments is one of the oldest forms of
water supply, having been used widely by ancient civilizations. The Romans
used lead-lined water conveyances and lead-lined cisterns for the storage of
water. Lead also was used in cooking utensils and wine storage containers.
It has been reported that plumbism (chronic lead poisoning) was a major
reason for the fall of the Roman Empire (Gil fill an 1965, Nriagu 1983).
Direct use of precipitation has been practiced in North America from very
early times and is still common where there are no other water supply
alternatives. Island communities in the equatorial regions of the world
still rely heavily on rainwater cisterns to supply their freshwater needs,
and this method of water supply is being, seriously considered as appropriate
technology for the developing counties or the world.
Roof catchments consist of an impervious surface, usually a house or auxili-
ary building roof, connected by means of conventional roof gutters and
downspouts to a below ground concrete or cinder block cistern. Water is
pumped from cistern storage to points of use within the house. Because in
most systems, precipitation is used directly with no treatment, the quality
of precipitation and the amourit of dry deposition on the catchment between
precipitation events are of paramount importance to the quality of drinking
water at the user's tap. The major impacts are two-fold. First, direct
deposition of atmospheric pollutants such as lead and cadmium may occur and,
6-32
-------
second, the acid components of atmospheric deposition may cause increased
corrosion of metallic plumbing system components.
In a study of 40 roof-catchment cistern systems in western Pennsylvania,
Young and Sharpe (1984) report that lead in atmospheric deposition accumu-
lates in the sediments that collect at the bottoms of cisterns and that this
particulate lead could appear in the drinking water of cistern users when
conditions allowing the suspension of this material in cistern water are
present. They did not report on the frequency of such conditions, but they
did point out that in the systems they studied there were no safeguards to
prevent the ingestion of lead-contaminated cistern sediments. However,
cistern systems with gross particulate filters for incoming catchment runoff
had much lower lead concentrations in sediments.
Young and Sharpe (1984) also report accumulations of cadmium in cistern
sediments, although such accumulations were less frequent than lead. The
cadmium concentrations in atmospheric deposition in the Young and Sharpe
study were generally very low, indicating that some other source such as
corrosion of galvanized gutters and downspouts might have been present.
Young and Sharpe (1984) found that precipitation was highly corrosive as
measured by the Langelier Saturation Index (LSI) and that cistern water,
although still corrosive in all but a few systems, was less corrosive than
bulk precipitation. The decreased corrosion potential of cistern water was
attributed to dissolution of the calcium carbonate building materials in the
cistern, a fact confirmed by the much higher LSI's of cisterns with imperme-
able vinyl liners.
Young and Sharpe measured the concentrations of copper and lead in tapwater
that had stood in the plumbing system overnight. In nine of the 40 systems
studied (22 percent) average lead concentrations exceeded drinking water
limits (U.S. EPA 1979b), copper exceeded drinking water standards (U.S. EPA
1979b) in 11 of the 40 systems. All of the systems (100 percent) having all
copper plumbing showed an increase in copper concentration in standing
tapwater as compared to cistern water, indicating that corrosion was taking
place.
Francis (1983) estimated that there are 133,000 individual water systems of
the roof-catchment cistern type in the United States. Of these, 12,000 are
located in the Northeast, 92,000 are located in the South, and 29,000 are
located in the North Central regions of the United States. No cistern
systems were reported in the West. These systems typically serve one single
family residence. Determination of the population at risk is difficult, but
these data indicate that it is likely to be substantial.
Cistern systems can be modified to minimize the risk (Young and Sharpe 1982).
However, these modifications are likely to be expensive with minimum estimat-
ed costs of $500 to $1000 per household for water treatment equipment and the
necessary changes to plumbing systems (Sharpe 1980).
Young and Sharpe (1984) conclude that "The presence of lead and copper in the
tapwater of cistern water supplies in western Pennsylvania was sufficient
6-33
-------
to constitute a hazard to users of such systems. Users involved in the study
were advised to discontinue use of cistern water for drinking purposes until
such time as proper safeguards were employed to reduce the hazards implicit
from this study."
6.3.1.2 Surface Vlater Supplies—Very little work has been done on the speci-
fic effects of atmosphericdeposition on surface water supplies, although
quite a bit can be inferred from the surface water quality work done to
determine impacts on aquatic biota. In most regions where atmospheric
deposition is of concern the same types of surface water are used for both
water supply and fish propagation; consequently, the water quality changes
reported for one are applicable to the other. The chief area of concern is
for surface water supplies providing drinking water for humans.
Two main drinking water impacts exist. The quality of the source water may
be impaired and/or increases in the corrosivity of the water could lead to
the same types of tapwater quality problems evident with cistern water
supplies. As reported elsewhere in this document (Chapter E-5), aluminum
concentrations may be increased in surface waters. In a 1981 study of the
surface water quality of a stream (Card Machine Run) feeding a small water
supply reservoir, DeWalle et al. (1982) reported that total aluminum concen-
trations in the stream directly above the water supply intake increased from
0.05 mg £~1 to 0.70 mg JT* in response to a February rain and snow-
melt event on the watershed. These data are illustrated in Figure 6-8. High
concentrations of aluminum have been reported elsewhere by Cronan and
Schofield (1979) and Herrmann and Baron (1980). The health significance of
aluminum concentrations of this magnitude are addressed elsewhere in this
chapter. Other metals not as readily leached from acidified soils are not
likely to increase as dramatically as aluminum.
Increasing corrosivity is probably the most significant potential impact of
atmospheric deposition on surface water supplies. The corrosivity of the
dilute water often used for surface water supplies in the northeastern United
States is mostly controlled by H+ concentration. As the H+ concentration
increases so does the corrosivity of the water (Figure 6-9).
Corrosivity in surface water supplies has been widely reported, and its
impacts are well documented. Where lead water distribution pipes are in use,
clinical lead poisoning of children has been reported as a consequence of
corrosive drinking water conveyance. A notable example of such a problem is
Boston, Massachusetts. Less well known is the case of Mahanoy City, PA
(Kuntz 1983). A case of copper toxicity from a corroded water fountain has
also been reported by Semple et al. (1960). Where pipes are of other metals
such as copper, iron, or galvanized steel the respective corrosion products
of copper, lead, iron, zinc, and cadmium can be problems.
Because these corrosion problems can lead to elevated concentrations of toxic
metals in drinking water, the U.S. EPA (1979a) has recommended that all
drinking water supplies be noncorrosive and that a minimum pH of 6.5 be
maintained. Numerous studies of surface water chemistry have shown dramatic
increases in the H+ concentration of surface waters in response to acidi-
fication by atmospheric deposition (Jeffries et al. 1979, Galloway et
6-34
-------
CM
I
CO
1.5
1.0
CARD MACHINE RUN
LEGEND
ALUMINUM
DISCHARGE
to
FEBRUARY
Figure 6-8. Aluminum concentration and discharge for Card Machine Run. Adapted from DeWalle et al
(1982).
-------
(Tl
OJ
cn
LEGEND
pH
RYZNAR INDEX
22 23 24
FEBRUARY
29
X
UJ
o
Qi
Figure 6-9. pH and Ryznar Index for Card Machine Run.
-------
al. 1980, Herrmann and Baron 1980, Corbett and Lynch 1982, DeWalle et al.
1982). In dilute surface waters such increases are almost certain to produce
corresponding increases in the corrosivity of that water. If the pH and
computed Ryznar Stability Index (RI) for the data of DeWalle et al. (1982)
are plotted for a rain and snowmelt event on Card Machine Run in February
1981 (Figure 6-9) a strong relationship between the two is identified.
Linear regression techniques were used to quantify the relationship between
pH and RI for this runoff event, and a correlation coefficient of r = -1.00
was obtained. Good correlation coefficients for pH and RI were also obtained
for three other streams in this area (Wildcat, McGinnis, and Linn Runs).
This indicates that large changes in the pH of dilute surface waters, weakly
buffered by CaC03, are almost certain to produce correspondingly large
increases in the corrosivity of such waters.
If RI values are plotted with streamflow (discharge) for the same event on
Card Machine Run (Figure 6-10), it is obvious that as streamflow increases as
a result of acid snowmelt and rainfall runoff, the corrosivity as indicated
by the Ryznar Index also increases dramatically. Regression analysis again
yields a very good correlation (r = 0.80) for these two variables.
Although the data presented are limited, there would appear to be strong
indications that the corrosivity of raw water entering surface water supplies
located in headwater areas of the Laurel Hill is increased substantially as a
result of acid snowmelt and rainfall runoff. If this model for the relation-
ship of pH and RI holds true for all dilute surface waters, then increased
corrosivity is likely anywhere that the pH of such waters changes dramatical-
ly subsequent to acid runoff events. Where surface water storage facilities
are small, necessitating the direct use of raw water during stormflow
periods, and where corrosion control is not practiced in the water system,
populations served are at increased risk of being exposed to higher
concentrations of corrosion products such as Cu, Pb, Cd, and Zn.
6.3.1.3 Groundwater Supplies—Acidification of groundwater as a consequence
of atmospheric deposition has been reported in Sweden by Hultberg and Wenblad
(1980). Such changes have not as yet been well documented in North America.
Fuhs (1981) reports that atmospheric deposition in sensitive regions of New
York State has decreased the pH and increased the Al concentration of shallow
groundwater and indicates that pH of groundwater is significantly correlated
with depth, with deeper groundwater sources having higher pH. Fuhs also
reports on the concentrations of Pb and Cu in private individual water
supplies obtaining water from shallow circulation springs and shallow wells.
Fuhs indicates that the Al concentrations measured in these types of water
sources would make such water unsuitable for hemodialysis units. Although
Fuhs demonstrates that standing tapwater derived from shallow groundwater
systems in atmospheric deposition sensitive areas of New York contains high
concentrations of Cu and Pb, he does not make a clear case linking these
results to the acidity of atmospheric deposition. As Fuhs correctly states,
shallow groundwater in these areas would be corrosive even without acid
deposition; consequently, the degree to which atmospheric deposition makes
these waters more corrosive and the concomitant increases in tapwater metals
concentrations must be determined. Neither has yet been demonstrated
conclusively.
6-37
-------
8£-9
c
ro
en
i
N
3
O)
-5
3
O-
ro
x
O-
Q.
O
3-
o>
-s
o
-s
o
EU
Q.
O>
O
3"
-j.
3
05
70
C
RYZNAR INDEX
DISCHARGE (i
ha-1)
-------
Unpublished data collected by Sharpe and DeWalle indicate a probable link
between acid recharge water and the decreasing pH and alkalinity of a deep
circulation spring on Pennsylvania's Laurel Hill. The data were collected
during an acid snowmelt and rainfall runoff event in March of 1982 and are
depicted in Figure 6-11. Unfortunately, flow data for the spring are not
available; consequently, flow data for Wildcat Run, a stream whose watershed
makes up a significant part of the spring's recharge area, are used for
comparison. Wildcat Run, at the point of flow measurement, is only several
hundred feet from the spring discharge and groundwater is an important
component of its total flow. Thus, the run's temporal response to acid
runoff recharge is likely to be quite similar to that of the spring. The pH
and alkalinity of the spring water appear to drop in concert with the
increased streamflow in Wildcat Run, with the most dramatic change occurring
in alkalinity.
As discussed in an earlier section of this chapter there is a strong corre-
lation between pH change and corrosivity for dilute waters; therefore, it
could be reasonably assumed that the corrosivity of the water in this spring
increased during the acid recharge event.
The lack of data is greatest with respect to groundwater impacts from atmos-
pheric deposition. Much additional work is indicated, but preliminary
information seems to indicate that adverse impacts to drinking water quality
are possible in water supplies using shallow groundwater in areas edaphically
and geologically sensitive to atmospheric deposition.
6.3.2 Lead
6.3.2.1 Concentrations in Noncontaminated Maters—The U.S. national interim
primary drinking water standard For lead is 50 yg £-1. The United
States Environmental Protection Agency (U.S. EPA 1979a) summarized data in
two surveys on lead in drinking water. The median lead concentration in
municipal drinking water supplies is about 10 yg £-1. In certain
areas, such as Metropolitan Boston, it may contain lead in excess of the 50
vig £-1 standard. This is believed to be due to very soft water (low
pH) and the presence of lead piping in the domestic water distribution system
(The Nutrition Foundation Expert Advisory Committee 1982). Lead piping is no
longer used for new potable water systems in the United States (U.S. EPA
1979a).
A recent national survey of Canadian drinking water supplies involving 71
municipalities representing 55 percent of the population, indicated a median
level of lead equal to or less than 1 yg £-1 and values ranged from < 1
yg £'! to 7 yg £-•*.
Most natural ground waters have concentrations ranging from 1 to 10 yg
£ .
6.3.2.2 Factors Affecting Lead Concentrations in Water, Including Effects
of pH—In areas where the home water supply is stored in lead-lined tanks and
where it is conveyed to the household taps by lead pipes, the concentration
6-39
-------
8.5
8.0
7.5
7.0
6.5
21
18
15
CO
o
o
ro
O>
6.0 -
7
I
11
LEGEND
ALKALINITY (Spring)
pH (Spring)
DISCHARGE (Wildcat Run)
12
13
14
15
MARCH
16
17
18
0.091
0.084
0.077
0.070
0.063 -
i
a
0.056 ^
i
0.049 I
^_-
Lul
0.042 <3
0.035
0.028
0.021
0.014
0.007
o
19
Figure 6-11. Alkalinity and pH for unnamed spring and discharge for Wildcat Run.
-------
mav reach several hundred micrograms per liter and even exceed 1000 ug
£-1 (Beattie et al. 1972). The concentration of lead in water conveyed
through lead pipes is affected by several factors. The longer the water is
held in the pipes, the higher the lead concentrations (Wong and Berrang
1976). The so-called "first flush" sample generally has lead concentrations
about three times higher than free-running tapwater (Nutrition Foundation
Expert Advisory Committee 1982). The lower the pH of the water and the lower
the concentration of dissolved salts, the greater the solubility of lead in
water.
Leaching of lead from plastic pipes has also been reported (Heusgem and
DeGraeve 1973). The source of lead was probably lead stearate, which is used
as a stabilizer in the manufacture of polyvinyl plastics.
6.3.2.3 Speciation of Lead in Natural Water—Lead does not present the wide
range of chemical and physical forms that mercury does. Metallic lead and
its inorganic compounds possess a negligible vapor pressure at room
temperatures, so volatile forms of lead are not important in the geochemical
cycle. The organometallic forms of lead, such as the tetra-alkyl leads,
although synthesized for use as antiknock compounds in gasoline, do not occur
naturally as in the case of methyl mercury compounds. The inorganic salts of
lead are numerous. The solubility of these compounds differs greatly.
The soluble salts will dissociate in water to liberate the reactive lead
cation Pb2+, which will form complexes and chelates with a variety of
organic ligands present in water and sediments. Sibley and Morgan (1977)
have described different forms of lead in freshwater: complexed ions, lead
absorbed to precipitate, solid precipitate, and free lead ions. Lead present
as the complexed ion is by far the most predominant species.
No studies have reported on the effect of acidic deposition on the speciation
of Pb in natural bodies of water. Lead has been reported to bind to a wide
range of organic fractions in river water (Ramanoorthy and Kusher 1975). As
pointed out in Chapter E-4 of this document, decreasing water pH will reduce
the fraction of heavy metals bound to organic components and increase the
concentration of free inorganic metal species. This should increase lead
levels in aquatic biota, possibly affecting human dietary intake.
6.3.2.4 Dynamics and Toxicity of Lead in Humans--Excellent reviews of this
topic have been published frirecent years TW"HO 1977, U.S. EPA 1980b,
Nutritional Foundation Expert Advisory Committee 1982).
6.3.2.4.1 Dynamics of lead in humans. The uptake, distribution, and
excretion of lead have recently been reviewed in detail (U.S. EPA 1980b).
Approximately 8 percent of dietary lead is absorbed in the gastrointestinal
tract in adults. Children absorb about 50 percent of the ingested lead.
Lead in water and other beverages may be absorbed with greater efficiency
than lead presented in food.
Lead is distributed to all tissues in the body and to all compartments within
cells. Most of the lead in blood is associated with the red blood cells.
The skeleton is the main site of lead storage, with about 95 percent of the
6-41
-------
total lead in the body in the skeleton of adults. Lead readily crosses the
placenta. It also crosses the blood-brain barrier but more readily in
children than in adults.
Lead is excreted in urine and feces, with the human urinary route probably
being more important. The half-time of lead retention in soft tissues is
about six weeks following exposure of a few months. The half-time may be
longer following years of occupational exposures to lead. Lead is accumu-
lated in the skeleton throughout most of the human life-span, and the
half-time in skeletal tissue is very long.
Lead concentration in whole blood is the most commonly used indicator for
assessing the burden of lead in soft tissues. The relative contributions of
airborne lead, lead in food, and other sources of lead are often assessed in
terms of their contributions to the blood-lead concentration.
A positive correlation exists between the concentration of lead in domestic
water supply and the concentration of lead in blood. The United States
Environmental Protection Agency, based on a study by Moore et al. (1977), has
estimated blood concentrations associated with levels of lead in free-running
tap water (Table 6-7).
If the relationship is valid, the impact of lead concentrations in running
tapwater is greatest in the lower range of lead in water. According to Table
6-7, the median lead level in U.S. drinking water (10 yg r1) would
contribute approximately 3.4 yg dl"1. Assuming the median blood level in
the absence of the water contribution to be 11 yg dl"1, the U.S. water
supply contributing about 30 percent additional blood lead and lead present
in tapwater at the current interim primary drinking water standard would
contribute about 10 yg dl"1 to blood lead concentration, i.e., about
equal to the lead contribution from all other sources. However, blood levels
in the United States are affected by a number of factors such as age, sex,
and urban versus non-urban locations. Urinary excretion of lead may be used
on a group basis to indicate the soft tissue burden. Lead in hair, unlike
the case of methyl mercury, is not a useful indicator because it represents
external contamination of the hair sample.
6.3.2.4.2 Toxic effects of lead on humans. Lead damages a variety of human
organs and tissues. Damage to the human hemopoietic system is usually the
first observable effect of lead (Figure 6-12). The inhibition of enzymes
involved in synthesizing hemoglobin results in the accumulation of precursor
substances: 6-aminolevulinic acid (6-ALA) in plasma and urine, and free
erythrocyte protoporphyrin (FEP) on the red blood cells. Measuring FEP has
become a routine method for checking the earliest effects of lead.
During recent years, measurement of FEP has come into wide use as the most
practical screening tool in both epidemiologic studies and in monitoring
populations at high risk for lead toxicity. Figure 6-13 shows the curvi-
linear relationship between FEP and lead concentration in blood. The
curvilinear shape is typical of the relationship between blood lead and other
intermediate metabolites of porphyrin synthesis, such as
-------
TABLE 6-7. THE ESTIMATED RELATIONSHIP BETWEEN LEAD
CONCENTRATIONS IN RUNNING TAP WATER AND HUMAN
BLOOD LEAD LEVELS (MOORE ET AL. 1977 IN U.S. EPA 19805)
Lead in running Total lead Lead in blood
tap water in blood (PbB) due to water
(yg rl) (yg dl-1) (yg dl-1)
0
1
5
10
25
50
100
lia
14.4
16.7
18.4
21.0
23.6
26.8
0
3.4
5.8
7.4b
10.0
12.6
15.8
aThe blood level of 11 yg dl-1 is strongly associated with air emis-
sions of lead, primarily resulting from the use of leaded gasoline.
Since 1977, such emissions have decreased by more than 50% on an annual
basis in the United States.
t>The NAS (1980) interpretation of the EPA's estimated relationship of
excess lead attributes only 5 yg dl-1 of blood lead due to 10 yg
fc-1 in drinking water. A concentration of 50 yg £-1 in drinking
water would add an additional 3.4 yg of lead per dl of blood lead.
6-43
-------
ENZYMIC STEPS
INHIBITED
BY LEAD
NORMAL PATHWAYS
METABOLITES AND
ABNORMAL PRODUCTS ACCUMULATED
IN HUMAN LEAD POISONING
PROPHYRIN FORMATION
IRON UTILIZATION
1
3
4
5
fifh .. .
•jpu
Pb
MITOCHON
'
CYTOPLASM
S
o
a:
o
o
\—
z:
rKREBS CYCLE 1 Fe TRA
SUCCINYL CoA + GLYCINE RETICL
| ALAS
1
-AnlNULtVULlNll' «^1U ^HLH>
I ALAD
1URO I SYN
UROIICOSYN
UROPORPHURINOGEN III
i UROGENASE
COPROPORPHYRINOGEN III
COPROGENASE
i HEMESYNTHETASE
1 Fe++
NSFERRIN
) INTO
LOCYTES
L
Pb
T Pb
iirnr
*• i
Pb
r
HEMOGLOBIN
Serum Fe
may be increased
ALA in urine (ALAU)
and serum increased
= urine
urine
COPRO in rbc urine (CPU)
Zn Protoporphyrin
(ZnP) in RBC
Ferritin, Fe micelles
in rbc
Damaged Mitochondria and
immature rbc fragments
(basophilic stippled cells)
Globin
Figure 6-12. The initial and final steps associated with disturbances
in the biosynthesis of hemoglobin due to lead are mediated
by intramitochondrial enzymes and the intermediate steps by
cytoplasmic enzymes. The enzymes most sensitive to lead
(steps 2 and 7) are the SH-dependent enzymes, 6-amino-
levulinate dehydrase (ALAD) and heme synthetase. Accumulation
of the substrates of these enzymes (ALA and FEP) is charac-
teristic of human lead poisoning as is increased urinary
coproporhyrin excretion. Although zinc protoporphyrin (ZnP)
accumulates in erythrocytes in lead poisoning (and iron
deficiency), it is usually measured as "free" erythrocyte
protoporphyrin (FEP). Lead reduces the bioavailability of
iron for heme formation. A compensatory increase in the
activity of the first enzyme in the pathway, 6-amino-
levulinic acid synthetase (ALAS), occurs in response to
reduced heme formation. Other compensatory responses
include erythroid hyperplasia , reticulocytosis and micro-
cytosis. Non-random shortening of erythrocyte life span
has been demonstrated in lead workers. Amicrocytic,
hypochromic anemia results including some morphological
features noted above. Adapted from Chisolm (1978).
6-44
-------
FEP = 0.043 x {blood lead) + 0.45(blood lead) - 2.14
r ' 0.79
n = 1056
10
u
o
o
o
1200
1080
960
840
720
600
480
360
240
120
0
. • ! -,sXV
':••••:
0 15 30 45 60 75 90 105 120 135 150
BLOOD LEAD (yg 100 ml'1)
Figure 6-13. Free erythrocyte protoporphyrin (FEP) vs blood level
Shoshone County, Idaho, August 1974. Adapted from
Landrigan et al. (1976).
6-45
-------
blood level of lead at which FEP or other metabolites attain abnormal values.
At first, levels of FEP increase slowly with blood lead, but as lead rises
about 40 to 50 yg dl~l the rate of increment of FEP rises rapidly. Roels
et al. (1978) defined abnormal blood FEP levels as those in excess of the
upper 95 percent confidence limit of the controls and published a dose-
response relationship relating blood lead levels to the frequency of individ-
ual having FEP values equal to or in excess of their defined abnormal value
(Figure 6-10). Children and adult females tend to show a greater response
than adult males. This analysis indicates that most of a population having
blood leads in the range of 30 to 40 yg Pb dl'1 will have abnormally high
FEP values.
Higher doses of lead cause anemia and damage to both the peripheral and
central human nervous system (Table 6-8). The central nervous system in
children appears to be more sensitive than the mature central nervous system.
A growing body of knowledge suggests that lower blood levels of lead exposure
than those previously recognized are associated with altered neuropsychologi-
cal function and intelligence deficits. For example, reduced general intel-
ligence quotients, reduced auditory and speech processing, and attention
deficits have been reported in children with higher dentine lead than those
with lower dentine lead (Needleman et al . 1979).
Piomelli (1980) has reported that heme synthesis is impaired in children with
blood levels less than 30 yg Pb 100 ml'1, consistent with findings of
Roels et al. (1978) reported in F-gure 6-14. Several other metabolic changes
associated with low level lead exposure of children have been identified.
Plasma levels of the vitamin D metabolite, 1,25-dihydroxy vitamin D, which is
active in stimulating the gastrointestinal absorption of calcium and phos-
phorous, decreased as the blood level increased (Rosen et al. 1980). Plasma
levels of the vitamin D metabolite exhibited a strong negative correlation
with blood lead concentrations in the range of 12 to 120 yg 100 ml"*,
with no difference in slope of the regression line from blood lead levels
over or under 30 yg 100 ml'1 (Mahaffey et al. 1982b).
Lead produces both acute and chronic effects on kidney function (Nutritional
Foundation Expert Advisory Committee 1982). The acute effects manifested as
dysfunction of the proximal tubular cell, such as amino aciduria, glycoseria,
and hyperphosphoturia, usually do not occur until blood levels exceed 70 yg
dl~l. Chronic lead nephropathy is not usually recognized in humans until
it has reached an irreversible stage. The disease is characterized by the
slow development of contracted kidneys with pronounced arteriosclerotic
changes, fibrosis, glomerular atrophy and hyaline degeneration of blood
levels. These changes portend progressive disease sometimes resulting in
acute renal failure. The duration of excessive exposure to lead is believed
to play an important role in the development of the disease. Although
information on blood levels is inadequate, it is unlikely that levels in the
general child and adult populations, even in the upper 2 to 5 percentile of
the "normal" U.S. range are sufficient to produce chronic renal effects.
Studies in the 19th and early 20th centuries indicated that occupational
exposures to lead (presumably higher than current exposures) caused increased
6-46
-------
TABLE 6-8. NO. DETECTED EFFECT LEVELS IN RELATION TO PbB
(ADAPTED FROM WHO 1977)
No-detected effect
level (yg 100 ml"1)
Effect
Population
< 10
20-25
20-30
25-35
30-40
40
40
40
40-50
50
50-60
60-70
60-70
> 80
Erythrocyte ALAD inhibition
FEP
FEP
FEP
Erythrocyte ATPase inhibition
ALA excretion in urine
CP excretion in urine
Anemia3
Peripheral neuropathy
Anemia9
Minimal brain dysfunction
Minimal brain dysfunction
Encephalopathy
Encephalopathy
Adults, children
Children
Adults, female
Adults, male
General
Adults, children
Adults
Children
Adults
Adults
Children
Adults
Children
Adults
aThe term anemia here is used to denote earliest statistically
demonstrable decrease in blood hemoglobin. In adult workers a decrease
in blood hemoglobin within the normal range has been reported during
the first 100 days of employment. Other studies of workers indicate
that frank anemia is not statistically demonstrable until PbB > 100
yg, as cited elsewhere in the WHO report. An increased frequency of
early anemia has been reported at PbB > 40 yg of groups of children
in whom concurrent iron deficiency anemia was not ruled out but is
highly likely.
6-47
-------
to
c
CD
cr>
i
PREVALENCE OF ABNORMAL VALUES (%.POPULATION)
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frequency of abortions and stillbirths (Oliver 1911). Indeed, following the
publication of Oliver's findings, women have largely been excluded from
occupational exposures to lead until very recently.
Lancranjan et al. (1975) have reported reduction in sperm counts and abnormal
sperm morphology in occupationally exposed men. The functional significance
on fertility is not known.
Prenatal exposure to lead may be associated with mental retardation in
children (Moore 1980). The human data are consistent with experimental
findings on animals that show modestly elevated blood levels, ~ 40 yg
dl~l, during prenatal and early postnatal life may be associated with
subtle and long lasting adverse consequences to the offspring.
Lead has been shown to be a carcinogen in animal tests, but epidemiological
studies have failed to reveal an association between lead exposure and human
cancer. Measurement of precursor metabolites of heme synthesis such as FEP
or 6-ALA provide the earliest warning of the effects of lead.
6.3.2.4.3 Intake of lead in water and potential for human health effects.
Mahaffey (1977) estimated that the daily intake of drinking water ranged from
300 ml for children to as much as 2000 ml for adults. An expert group of the
National Academy of Sciences (NAS 1980) stated a value of 1630 ml day1 for
water intake of adults (not including amounts used to prepare foods and
beverages) and a range of 100 ml to 3000 ml for children.
A study in Canada by Armstrong and McCullough quoted by the Nutrition
Foundation Expert Advisory Committee (1982) indicated that the total daily
intake including water used as a food ingredient was 760 ml averaged for 0 to
6 years, and 1140 ml for the 6- to 18-year-old group. The highest average
daily intake was 1570 ml for the 55 and older age group. However, up to 3000
ml total water per day was consumed by some children in the 0- to 6-year-old
age group and up to 4300 ml total water was consumed by certain individuals
in each of the remaining age groups.
Using the NAS reported range of 100 to 3000 ml for children and a U.S. median
level of 10 ug *,-!, the range of intake for children would be 1 to 30
yg Pb and for adults 16 ug, assuming a water intake of 1600 ml day"1
(Table 6-9). If average lead concentrations attained the interim drinking
water standard of 50 yg £-1, these intake values would be five times
greater.
The review of the human toxicity of lead in Section 6.3.2.4.2 identified
children as the most susceptible group in the general population. Blood lead
levels in children in the United States cover a broad range of values
(Mahaffey et al. 1982a). A criterion of 30 yg Pb 100 ml-1 whole blood
has been used in estimating the prevalence of elevated blood lead (Center for
Disease Control 1978). If this concentration of blood lead is accompanied by
an erythrocyte protoporphyrin concentration of 50 to 250 ug 100 ml-1 Of
whole blood, the child is thought to have undue lead absorption. Community
6-49
-------
TABLE 6-9. DAILY INTAKE OF LEAD FROM DRINKING WATER
Age Group Daily Water Intake9 Daily Lead Intakeb
ml yg Pb
Children 100 - 3000 1 - 30
Adults 1630 16
aNAS (1980).
bAssumes U.S. median concentration of lead in drinking water to be 10
yg Pb £-1.
6-50
-------
based lead poisoning prevention programs report that approximately 75 percent
of children with blood lead levels of >_ 30 yg 100 ml-1 also have
erythrocyte protoporphyrin values of > 50 yg 100 ml-1 (Mahaffey et al.
1982a). The review of human toxiciTy data in Section 6.3.2.4.2 also
indicates that blood lead levels in children >^ 30 yg 100 ml-1 indicates a
risk of biochemical, if not neuropsychological, dysfunctions.
A survey of blood lead levels in children in the years 1976 to 1980 in the
United States indicated that substantial numbers of children have blood lead
levels > 30 yg dl~l (Table 6-10). The prevalence of elevated blood lead
values ~Ts highest in children of low income families (approximately 11
percent of children in families having an income less than $6000) and in
children living in large cities (7.2 percent of children living in cities of
population more than one million). However, elevated blood lead is widely
distributed in the general population, including children in families earning
more than $15,000 annual income (1.2 percent) and in children living in rural
areas (2.1 percent).
Section 6.3.1 reviewed available data to indicate that reduced pH increases
the corrosivity of water and can mobilize metals such as lead, resulting in
increased concentrations in drinking water. Lead piping in home plumbing is
rare and no longer used in this country except in certain parts of New
England. However, lead can be mobilized from other types of piping where it
is used as a solder (copper piping) or in stabilizers (certain types of
plastic pipes). Homes using roof-catchment cisterns for collecting drinking
water seem especially vulnerable to corrosive rain water. Young and Sharpe
(see Section 6.3.1.1) noted that 22 percent of such systems yielded lead
concentrations in tapwater (having stood overnight) in excess of the drinking
water standard of 50 yg Pb £~1.
From the point of view of human health risks, any increases of lead concen-
trations in drinking water should be viewed as an additional burden of lead.
This is especially important with children where substantial numbers already
have elevated blood levels. Drinking water at the median concentration of 10
yg Pb jr1 already makes an appreciable contribution to blood lead
levels (approximately 30 percent added on to other sources of lead; see
Section 6.3.2.4.1). Thus the drinking water standard of 50 yg Pb £-1
will not provide sufficient protection to those children already having high
blood lead from other sources of exposure.
Unfortunately quantitative data are lacking on the contribution of acidic
deposition to lead in drinking water. Roof-catchment cistern systems
believed to be widely used in rural areas of Ohio and Western Pennsylvania
appear to be a probable target for the effect of acidic deposition. Thus, it
is of great importance to ascertain the extent of usage of these systems in
those areas of the U.S. subject to acidic deposition and to check the extent
to which changed corrosivity of this water affects lead levels in tapwater.
6.3.3 Aluminum
Inorganic aluminum is toxic to fish and may be the main cause of fish kills
due to acidification of natural bodies of water. Acidic deposition dissolves
6-51
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TABLE 6-10. BLOOD LEAD LEVELS IN CHILDREN 6 MONTHS THROUGH 5 YEARS
BY ANNUAL FAMILY INCOME AND DEGREE OF URBANIZATION OF PLACE OF
RESIDENCE IN THE UNITED STATES FROM 1976 TO 1980*
Demographic variable
Estimated
population
(thousands)
No. of
persons
examined
B1oodb lead
ug 100 ml'1
Prevalence of
blood lead
levels > 30
yg 100 ml"-1
% persons
exami ned
Annual Family Income0
< $6000 2465
$6000 - 14,999 7534
> 15,000 6428
448
1083
774
20 +_ 0.6
16 4- 0.5
14 + 0.4
10.9 +_ 1.4
4.2 + 0.7
1.2 + 0.4
Degree of Urbanization
urban > 106 persons 4344
urban < 106 persons 6891
Rural 5627
544
944
884
18 + 0.5
16 T 0.7
14 + 0.6
7.2 + 0.7
3.5 + 0.6
2.1 + 0.9
aAdapted from of Mahaffey et al. (1982a).
bMean +_ S.E.M.
CA11 values shown for this variable reflect the exclusion (from
analysis and tests of significance) of children in households that
declined to reported their income.
6-52
-------
aluminum in clay materials in soils and sediments, thereby increasing
concentrations of the A13+ ions and inorganic salts of aluminum (for
details, see Chapter E-2). Fish mortality appears to be due to damage to the
gills of the fish. The toxic properties of aluminum are self-limiting with
regard to bioaccumulation; when the aluminum levels in water reach toxic
levels, the ensuing mortality of fish stops further accumulation in aquatic
food chains. The behavior of aluminum is thus in sharp contrast to methyl
mercury, which is of lower toxicity to fish and is avidly accumulated.
Aluminum in drinking water, unlike lead, is not directly toxic to humans.
However, a special circumstance may lead to human toxicity—that is the use
of aluminum containing water in hemodialysis procedures. This is believed to
lead to direct entry of aluminum into the blood stream and eventually damage
to the central nervous system.
6.3.3.1 Concentrations in Uncontaminated Water—Burrows (1977) has reviewed
the literature on concentrations of aluminum in natural bodies of water. He
draws attention to two factors that are important in assessing published
values. First, many publications do not clearly distinguish between dis-
solved and suspended aluminum in water. He notes that many investigators now
use a 0.45 ym millipore filter to distinguish between dissolved and parti-
culate aluminum. The second factor is that procedures for trace analysis of
aluminum have only recently become available and most of the literature data
have been collected without using these techniques. Burrows states that, as
a general rule, all aluminum values reported before 1940 should be regarded
with skepticism. Unfortunately, very few analyses have been reported for the
most recent times (from 1970). The Maumee River Basin (Ohio) was reported to
have a mean value of 0.01 mg £-1 for the period 1971-73. A phosphate
limestone lake in Florida had a mean value of 0.05 mg £~1 at a water pH
7.0 to 9.6. Tributaries to Lake Michigan had mean values of 0.353 in 19/2
but pH was not specified. The above values have been taken from Burrows
(1977).
6.3.3.2 Factors Affecting Aluminum Concentrations in Water—Burrows (1977)
notes a number of factors that influence aluminum concentrations in bodies of
natural water:
1) Acidic waters consistently contain much more soluble aluminum
than neutral or alkaline waters. Schofield and Trojnar (1980)
report that in a brook in the Adirondack Wilderness region of
New York State, aluminum concentrations rose from about 0.2 mg
n~l at pH 5.5-6.5 to 0.8-1.0 mg rl as the pH fell to less
than 5.0 during the spring snowmelt.
1Editor's note: Several reviewers felt references to hemodialysis (this
page and page 55) are irrelevant in that water used in such units should be
deionized. However the literature indicates that effects due to aluminum in
dialysate, traced to the aluminum concentration in water, have occurred and
may be an important factor in long-term dialysis treatment.
6-53
-------
2) Highly saline waters contain higher aluminum concentrations
than freshwaters.
3) Hot waters (e.g., hot water springs) tend to have higher levels
of aluminum than cold water.
4) Moving waters tend to give higher aluminum analysis than
quiescent waters. This effect is probably due to mobilization
of suspended material.
6.3.3.3 Speciation of Aluminum in Water—The species of aluminum in bodies
of natural water have been discussed in Chapter E-4. Most of the dissolved
aluminum is present as complexes with organic ligands. The inorganic
fractions consist of Al^+ and aluminum fluoride, hydroxide, and sulfate
complexes. The fluoride complex is probably the predominant inorganic
species, according to thermodynamic calculations (Driscoll et al. 1980).
The inorganic monomeric species are more toxic to fish than are the organic
complexes of aluminum. Of the inorganic species, the fluoride complex is
probably the least toxic because addition of fluoride ion reduces the
toxicity of aluminum. Lowering the pH in natural bodies of water increases
the labile (inorganic) monomeric aluminum and thereby increases toxicity to
fish. Driscoll et al. (1980) found that seasonal variations in organically-
chelated aluminum were not affected by seasonal variations in pH in lakes in
the Adirondack region of New York State. The organic aluminum correlated
with total carbon measurements in water.
6.3.3.4 Dynamics and Toxicity in Humans--This topic has been the subject of
a number of reviews (Norseth 1979).
6.3.3.4.1 Dynamics of aluminum in humans. Data on absorption, distribution,
and excretionof"aluminumcompoundsfh man have been reviewed recently
(Norseth 1979). Aluminum is absorbed in the gastrointestinal tract. The
fraction of dietary intake absorbed into the blood stream is believed to be
small, but precise figures are not available. When aluminum was given as the
hydroxide salt to uremic patients, approximately 15 percent of the dose was
absorbed, with considerable differences between individuals (Clarkson et al.
1972). Unfortunately, information is not available on the absorption of
other forms of aluminum or in people with normal kidney function. Aluminum
is distributed to all tissues in the body and has been reported in fetal
tissues. When aluminum in food was given to rats, increased levels were
reported in blood, brain, liver, and testes (Ondreicka et al. 1966).
Little information on the relative importance of urine versus fecal pathways
of excretion is available. Renal clearance of aluminum may be as high as 10
percent of the glomerular filtration rate (creatinine clearance) as indicated
in patients with compromised renal function. These data would suggest a high
urinary rate of excretion in normal subjects and a correspondingly short
biological half-time (on the order of days or hours). Animal experiments
indicate that biliary excretion of aluminum contributes to fecal excretion of
the metal.
6-54
-------
Aluminum is found in both cow and human milk. Normal levels of aluminum in
human blood and other biological fluids exhibited a very wide range of values
relative to the different laboratories making the analyses. Apparently
considerable problems remain, particularly those related to chance contamina-
tion by the ubiquitous metal, in determining reliable values for the low
levels in human plasma.
6.3.3.4.2 Toxic effects of aluminum in humans. Toxic effects in terms of
fibrosis of lung tissue have been reported in workers inhaling aluminum or
its compounds. The situation with regard to toxic effects in humans due to
oral intake of aluminum is equivocal. An early claim (Crapper et al. 1973)
that Alzheimer's Disease—a chronic degenerative disease of the central
nervous system leading to presenile dementia--was associated with accumu-
lation of aluminum in the brain has not been substantiated by later studies
(Markesbery et al. 1981). However, a chronic neurological disease "Dialysis
Dementia," that develops in a number of patients receiving dialysis therapy
may be associated with elevated aluminum intake (Alfrey et al. 1976,
McDermott et al. 1978). Intake of aluminum may be directly from the water
used in the dialysis fluid or from the aluminum hydroxide compounds given
orally to remove phosphate from the uremic patients. Aluminum has been shown
to be harmful to the central nervous system in animals when directly adminis-
tered in brain tissue (Kopeloff et al. 1942) and to damage neuroblastoma
cells in culture (Miller and Levine 1974).
6.3.3.5 Human Health Risks from Aluminum in Water—Acute or chronic disease
in man has not been related to normal dietary intake of aluminum from food or
drinking water. However, a potential risk may exist under the special
circumstances of patients with compromised kidney function who undergo
regular therapeutic dialysis. Driscoll et al. (1980) have reported levels of
aluminum as high as 800 yg Al £"1 in natural bodies of freshwater in
the Adirondack Region of New York State under the influence of acidic
deposition. A concentration of 50 yg ji"1 of aluminum in dialysis water
is claimed to be dangerous (Registration Committee, European Dialysis and
Transplant Association 1980).
Of the various species of aluminum known to exist in bodies of natural water,
only data on aluminum hydroxide are available. This is absorbed across the
human gastrointestinal tract. In areas of the country where drinking water
is fluoridated or where elevated fluoride concentrations occur naturally, it
is likely that aluminum flouride complexes will be present in tapwater in
substantial amounts. Unfortunately, we know nothing of the gastrointestinal
absorption or about its potential toxicity in humans.
6.4 OTHER METALS
A number of other metals such as cadmium, copper, manganese, and zinc have
been mentioned with regard to the possibility of indirect health effects. In
general, evidence to justify a detailed report for each metal is lacking.
However, it should be noted that this chapter has not considered at least one
potential pathway of human intake of environmental chemicals, i.e., the food
supply other than fish and fish products. Cadmium is known to be accumulated
6-55
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by plants, including cereals, and the possible effects of acidic deposition
have not been considered chiefly because of a lack of studies.
6.5 CONCLUSIONS
Chapter E-6 examines the indirect effects on health and in doing so mainly
discusses lead and mercury availabilities as affected by acidic deposition.
The following statements summarize the content of this chapter.
o No adverse human health effects have been documented as being a con-
sequence of metal mobilization by acidic deposition. On the other
hand, interest in the phenomena of acidic deposition is recent and
few investigations, if any, have been made into the possibility of
indirect effects on human health (Section 6.2.1).
° The substances requiring special attention are methyl mercury, due
to its accumulation in aquatic food chains, and lead due to the
potential for contaminating drinking water (Section 6.2.1).
" In virtually all studies published to date, elevated methyl mercury
levels in fish muscle (most notably pike and perch) have been
statistically associated with higher levels of acidity in water.
However, a number of factors influencing mercury levels in fish may
also change parallel to acidity (Section 6.2.3).
o More research is needed to identify all the factors that affect
mercury accumulations in fish and the relative importance of each.
This need is especially urgent in the United States where few data
are available at this time (Section 6.2.3).
0 The contamination of freshwater fish by direct discharge of mercury
has been curtailed in recent years. The role of long-distance
transport of mercury merits careful investigation as an explanation
for high mercury levels in lakes remote from mercury-related
industries (Section 6.2.2).
0 Potential impacts of acidic deposition on methyl mercury concentra-
tions in freshwater are of interest for several reasons (Section
6.2).
a) Fish and fish products are the major if not only sources of
methyl mercury for humans.
b) Consumers of freshwater fish have a greater possibility of
exceeding a allowable daily intakes of methyl mercury than do
consumers of other forms of fish.
c) Pike and trout, freshwater fish among the most likely species to
be affected by acidic deposition, have the highest user consumption
figures and the highest average methyl mercury levels.
6-56
-------
o Prenatal life is a more sensitive stage of the life cycle for methyl
mercury effects. More information is needed on fish consumption
patterns of women of child-bearing age in order to quantitatively
assess the potential impact on human health of elevated methyl
mercury levels in freshwater fish (Section 6.2.4).
Data on the impacts of acidic deposition on drinking water quality are
scarce. However, by using available information, tentative assessments of
impacts on ground and surface water systems were made.
° The lack of data is greatest with respect to groundwater.
Preliminary information seems to indicate that adverse impacts to
drinking water quality are possible in water supplies using shallow
groundwater in areas edaphically and geologically sensitive to
acidic deposition (Section 6.3.1.3).
o Increasing corrosivity is probably the most significant potential
impact of acidic deposition on surface water supplies. Populations
are at increased potential risk of being exposed to higher concen-
trations of corrosive toxicants, such as lead and possibly cadmium,
where surface water storage facilities are small, necessitating the
direct use of raw water during stormflow periods and where corrosive
control is not practiced in the water system (Section 6.3.1.2).
° People receiving drinking water from roof-catchment cistern systems
should be considered at potential risk of increased intake of lead
in areas of acidic deposition and especially if cisterns that are
used have no particulate filters (Section 6.3.2).
o From the point of view of human health risks, any increases of lead
concentrations in drinking water should be viewed as an additional
burden of lead. This is especially important where substantial
numbers of children already have elevated blood lead levels (Section
6.3.2.4).
o Acute or chronic diseases in humans have not been related to normal
dietary intake of aluminum from food or drinking water. However, a
potential threat exists for patients undergoing hemodialysis if
aluminum concentrations in the water used in this treatment exceed
50 yg of aluminum per liter (Section 6.3.3).
Generally, the indirect effects on human health attributable to acidic
deposition require further study. Data are very limited with regard to
measurement of the toxic elements and their speciation and to the kinetics of
transfer and uptake by accumulation processes. Studying less toxic essential
metals may be important in that elevated concentrations of some or all of
them might affect the food chain dynamics or the toxicity of lead or mercury.
6-57
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Wood, J. M. 1980. The role of pH and oxidation-reduction potentials in the
mobilization of heavy metals. In Polluted Rain, pp. 223-232. T. Y.
Toribara, M. W. Miller and P. E. Morrow, eds. Plenum Press, New York.
World Health Organization. 1972a. Evaluation of certain food additives and
the contaminants mercury, lead and cadmium. World Health Organization
Technical Report Series No. 505, Geneva.
World Health Organization. 1972b. The Hazards of Persistent Substances in
Drinking Water. World Health Organization, Copenhagen.
World Health Organization. 1976. Environmental Health Criteria 1. Mercury.
World health Organization, Geneva.
World Health Organization. 1977. Environmental Health Criteria 3. Lead.
World Health Organization, Geneva.
World Health Organization. 1980. Health Risk Evaluation for methylmercury
- An Interim Review. EHE/EHC/80. 22. World Health Organization, Geneva.
Young, E. S. and W. E. Sharpe. 1982. Rainwater cisterns design construction
and management. Cooperative Extension Service, The Pennsylvania State
University, University Park, PA. 15 pp.
Young, E. S. and W. E. Sharpe. 1984. Atmospheric deposition and
roof-catchment cistern water quality. J. Environ. Qua!. 13:38-43.
6-66
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS
E-7. EFFECTS ON MATERIALS
7.1 DIRECT EFFECTS ON MATERIALS (J. E. Yocom and N. S. Baer)
7.1.1 Introduction
In the popular press many articles ascribe damaging effects to acidic depo-
sition (LaBastille 1981). Damage to non-living materials and structures is
commonly listed as one of the important effects of this phenomenon. Further-
more, damage to irreplaceable historic buildings and monuments, works of art,
and other cultural properties is emphasized as one of the most important
aspects of such damage. If one narrowly considers the "acid rain syndrome"
as precipitation that has been rendered more acidic as a result of long-range
transport of acid rain precursors, this mechanism alone probably accounts for
only a small fraction of the total damage to materials attributable to the
effects of air pollutants.
In general, the distinction between the effects on materials of near or in-
termediate sources from distant sources is difficult if not impossible to
make.1 If the discussion is broadened to "acidic deposition," which
includes all of the mechanisms by which acidic pollutants (gases and solid
and liquid particulate matter) may contact and damage surfaces, one is able
to point to a considerable body of experimental evidence for damage to
materials by acidic deposition. For most cases, in urban areas where most
materials are located, the atmospheric load from local sources tends to
dominate over the smaller amounts of pollutants arriving from remote upwind
sources (U.S./Canada 1982). This broad definition is used for this chapter.
This chapter deals with the effects on materials of anthropogenic acidic air
pollutants. Later in this chapter several typical broad mechanisms for
acidic deposition are discussed. They include adsorption and absorption of
acidic primary pollutant gases such as S02 and N0£ on moist surfaces and
their conversion to strong acids, and the processes in which precipitation is
acidified by condensation around acidic particles or washout of acidic pri-
mary gases. While this chapter's scope is extremely broad in concept, the
literature describing research on any one specific contact-and-effect sce-
nario may be limited or even non-existent.
A significant body of literature describes the effects of primary air pol-
lutants on materials as determined by both laboratory and field experiments.
This literature has been summarized in detail by the U.S. Environmental
ln Chapter A-9 of this document the following definitions for the scales
of pollutant transport are given: short range (< 100 km), intermediate
range (100 to 500 km), and long range (> 500 km).
7-1
-------
Protection Agency in its criteria documents supporting the establishment of
air quality standards, for example, the document on sulfur oxides and par-
ticulate matter (U.S. EPA 1982). Several other reviews have also been
published (Yocom and Grappone 1976, Yocom and Upham 1977, Yocom and Stankunas
1980). A recent review in draft form by Haagenrud et al. (1982) deals
primarily with effects of sulfur compounds. The draft U.S./Canada (MOI)
Transboundary Report contains a review of the literature on the effects of
acidic deposition on materials (U.S./Canada 1982).
Among the documented effects of air pollution on materials are many that may
be broadly described as associated with acidic deposition. Table 7-1 sum-
marizes the potential damaging effects of air pollutants and other environ-
mental conditions on several classes of materials. One should note that
sulfur oxides, other acidic gases, and particulate matter figure prominently
among the important, potentially damaging pollutants, and note that moisture
(as atmospheric humidity and surface wetness) is an extremely important
factor.
Damage to materials from acidic deposition takes a variety of forms including
the corrosion of metals, erosion and discoloration of paints, decay of build-
ing stone, and the weakening and fading of textiles. All of these effects
occur to a significant degree as a result of natural environmental condi-
tions, even in unpolluted atmospheres. Moisture, atmospheric oxygen, carbon
dioxide, sunlight, temperature fluctuations, and the action of microorganisms
all contribute to the deterioration of materials. Quantifying the specific
contributions of anthropogenic air pollutants to such damage is a formidable
task. Furthermore, distinguishing the relative amount of damage caused by
specific pollutant transformation and contact processes (for example, acid
precipitation) becomes even more elusive.
7.1.1.1 Long Range vs Local Air Pollution—Acidic pollutants whether they
are preseTvtasprimarypollutantgases(e.g., SOg and NOX), as fully
oxidized acids or salts (e.g., sulfates and nitrates), or in the form of
acidified precipitation may have arrived at a material surface from local
pollutant sources or may have been transported many miles from distant
sources. Table 7-2 summarizes the characteristics of long-range and local
air pollutants and their effects. As the table shows, several mechanisms may
be described as acidic deposition. The separation of long-range and local
characteristics is somewhat artificial because phenomena associated with
long-range transport may be generated by local sources under the appropriate
conditions. For example, acidic deposition may be produced close to sources
of primary pollutants under the proper meteorological conditions. The
distinction between different acidic deposition scenarios is especially
important when the cost of damage related to such deposition is considered
and when control strategies to ameliorate such damaging effects are being
developed. The transport, deposition, damage, and cost scenarios of greatest
economic importance must be defined before the effectiveness of any control
strategy can be estimated.
7.1.1.2 Inaccurate
literature
property.
contains
In most
Claims of
frequent
cases no
"Acid Rain
references
attempt i s
" Damage
to "acid
made to
to Material s--The popular
rain" damage to cultural
distinguish between local
7-2
-------
TABLE 7-1. AIR POLLUTION DAMAGE TO MATERIALS
I
GO
Materials
Metal s
Building
Stone
Ceramics
and Glass
Paints and
Organic
Coatings
Paper
Photo-
graphic
Materials
Textiles
Textile
Dyes
Leather
Rubber
Type of
Impact
Corrosion,
tarnishing
Surface erosion,
soiling, black
crust formation
Surface erosion,
surface crust
formation
Surface erosion
discoloration,
soiling
Embrittleraent,
discoloration
Microbleraishes
Reduced tensile
strength,
soiling
Fading, color
change
Weakening,
powdered surface
Cracking
Principal air
pollutants
Sulfur oxides
and other acid
gases
Sulfur oxides
and other acid
gases
Acid gases,
especially
fluoride-
containing
Sulfur oxides,
hydrogen
sulfide, ozone
Sulfur oxides
Sulfur oxides
Sulfur and
nitrogen
oxides
Nitrogen
oxides and
ozone
Sulfur oxides
Ozone
Other
environmental
factors
Moisture, air,
salt, particulate
matter
Mechanical ero-
sion, particulate
matter, moisture,
temperature
fluctuations,
salt, vibration,
C02, micro-
organisms
Moisture
Moisture,
sunlight,
particulate
matter, mechan-
ical erosion,
microorganisms
Moisture, phys-
ical wear,
acidic materi-
als introduced
in manufacture
Particulate
matter,
moisture
Particulate
matter,
moisture,
light, physical
wear, washing
Light,
temperature
Physical wear,
residual acids
introduced in
manufacture
Sunlight,
physical wear
Methods of measurement
Weight loss after removal of
corrosion products, reduced
physical strength, change in
surface characteristics
Weight loss of sample, surface
reflectivity, measurement of
dimensional changes, chemical
analysis
Loss in surface reflectivity
and light transmission, change
in thickness, chemical
analysis
Weight loss of exposed painted
panels, surface reflectivity,
thickness loss
Decreased folding endurance,
pH change, molecular weight
measurement, tensile strength
Visual and microscopic
examination
Reduced tensile strength,
chemical analysis (e.g.,
molecular weight) surface
reflectivity
Reflectance and color value
measurements
Loss in tensile strength,
chemical analysis
Loss in elasticity and
strength, measurement of crack
Mitigation measures
Surface plating or coating,
replacement with corrosion-
resistant material, removal to
controlled environment.
Cleaning, impregnation with
resins, removal to controlled
environment.
Protective coatings,
replacement with more
resistant material , removal to
controlled atmosphere.
Repainting, replacement with
more resistant material
Synthetic coatings, storing
in controlled environment,
deacidification, encapsula-
tion, impregnation with
organic polymers.
Removal to controlled
atmosphere
Replacement, use of substi-
tute materials, impregnation
with polymers
Replacements, use of
substitute materials, removal
to controlled environment.
Removal to controlled
environment, consolidated with
polymers, or replacement
Add antioxidants to
formulation, replace with more
frequency and depth
resistant materials
-------
TABLE 7-2. CHARACTERISTICS OF LONG-RANGE AND LOCAL AIR POLLUTION
Pollutant
or Effect
Long-range
Local
Pollutant Concen-
tration Patterns
Sulfur Oxides
Nitrogen Oxides
Particulate Matter
(includes
aerosols)
Ozone and Other
Oxidants
Dry Acidic
Deposition
Acidic
Precipitation
Acidic Fog
(includes liquid
aerosols)
Low concentrations and uniform
distribution.
SOg tends to be oxidized to
particulate sulfates.
Significant conversion to
particulate nitrates.
Only the smallest primary
particle sizes persist. Large
component of material converted
from gases and vapors to
particulate form such as
sulfates.
Ozone and other oxidants are
produced from hydrocarbons and
NOX over moderate to long-range
transport in presence of
sunlight.
Dry deposition of acidic
particles (for example, sul fates)
is possible.
Acidic rain mechanisms appear to
be dominated by processes
involving condensation on acidic
particles and oxidation of
dissolved S02 in cloud
droplets.
Acidic fog may be formed by drop
condensation around small acidic
particles or other acidic
condensation nuclei.
High to moderate
concentrations and strong
gradients in time and space.
Exist primarily as SOg;
however, under light winds and
stable atmospheric conditions
conversion to particulate sulfate
can occur.
Exist primarily as NO and NOg,
but under low wind speed, stable
conditions and sunlight, conversion
to organic or inorganic nitrates in
particulate form is possible.
Exists in wide range of sizes which
may be bimodal. Particles are
capable of producing surface
soiling and participates in the
formation of corrosion layers
(e.g., black crust on stone).
The formation of ozone and other
oxidants is likely only under low
winds and sunlight if precursors
are present.
Dry deposition of acidic particles
is possible, especially under
stable conditions, often enhanced
by moist surfaces.
Acidic rain formation may be
predominantly through rain washout
of acidic particles and pollutant
gases.
Same as for long-range.
7-4
-------
pollution sources and long-range transport. In some cases the damage is
caused by factors entirely independent of acidic deposition.
Perhaps the most egregious example is the damage to the granite Egyptian
obelisk, "Cleopatra's Needle," located since 1881 in Central Park in New York
City. In one account, it was stated that, "The city's atmosphere has done
more damage than 3 1/2 millenia in the desert, and in another dozen years the
hieroglyphs will probably disappear" (New York Times 1978a). A careful study
of the monument's complex history makes it clear that the damage can be at-
tributed to advanced salt decay, high humidity of the New York climate, and
unfortunate attempts at preservation (New York Times 1978b, Winkler 1980).
7.1.1.3 Complex Mechanisms of Exposure and Deposition—The work done to date
to measure damage to materials from acidic deposition has not considered to
any significant degree the specific mechanisms of exposure, deposition, and
subsequent damage. As will be discussed, most of the studies that have used
laboratory chamber exposure or field exposure in the ambient atmosphere are
unable to isolate specific deposition mechanisms from the many interrelated
chemical and physical processes involved. The following list presents a
series of simplified mechanisms that the authors believe occur in one form or
another. These mechanisms are based upon the presence of acidic gases such
as S02 and N02, their transformation products, and moisture in some form.
1. Dry Gas, Dry Surface: An acid gas is adsorbed on a relatively dry
material surface (for example, building stone) and exposure to
moisture forms acids that attack the material.
2. Dry Gas, Wet Surface: An acid gas is absorbed in moisture (con-
densed dew or collected precipitation) already on surfaces and
results in acid attack.
3. Large, Dry Particle, Dry Surface: Large particles containing acid
components fall on the material's surface and lead to damage di-
rectly. An example would be acid-containing soot from an oil-
fired boiler.
4. Small Particle, Dry or Wet Surface: A small particle containing
acidic compounds such as sulfuric or nitric acid salts capable of
reacting with moisture to form acids settles on or impacts on a
dry or wet surface and subsequently leads to acid attack.
5. Acid Precipitation: Rain or snow containing acidic components
falls on the material surface and leads to damage directly.
The above group of simplified mechanisms is not intended to be exhaustive or
completely rigorous. They are illustrative of the wide spectrum of processes
that operate to produce acidic deposition and each of the listed mechanisms
may have one or more variations. For example, in mechanism 1 (Dry Gas, Dry
Surface) it is likely, in the case of $03 contact, that some surface oxi-
dation may take place within a relatively dry adsorbed layer or that S02
7-5
-------
may react directly with a reactive surface to produce a sulfite salt. Never-
theless, as will become apparent later in this chapter, acidic deposition and
subsequent damage accelerates in the presence of moisture.
The end result of each of these mechanisms is acidic deposition capable of
damaging materials. Yet certain of these mechanisms are undoubtedly more
important than others in causing economically significant damage. In large
population centers where levels of primary, gaseous pollutants and total
material inventories are high, mechanisms 1, 2, and 3 may be more important
than 4 and 5. In rural areas, where the inventory of exposed materials is
likely to be different from urban areas and the pollutant mix may include a
higher portion of secondary, particulate pollutants, mechanisms 4 and 5 may
dominate.
These factors and others such as the distinction between wet and dry deposi-
tion mechanisms are important because of the link between pollutant levels
and meteorological factors. For example, if a local source has an elevated
emission point, the kind of surface inversion associated with radiational
cooling and dew formation may also act to keep the pollution from reaching
ground level. Thus, mechanism 2 may not be especially important, even though
all the critical components (active pollutant, susceptible material, wet
surface) are all present on an annual average basis. Conversely, materials
on elevated terrain may be subject to pollutant plume impact only rarely, but
when they are affected, the conditions (such as wetness) may be such that the
maximum degree of damage occurs.
Note in Table 7-2 that mechanisms 4 and 5 (small particle, dry or wet sur-
face; and acid precipitation) may occur both locally and after long-range
transport. Stable atmospheric conditions and low wind speeds may provide the
time necessary for atmospheric transformations to create effects on a local
scale that would otherwise be associated with long-range transport.
7.1.1.4 Deposition Velocities—Chemical reaction between exposed surfaces
and air pollutants leads to removal of the pollutant from the atmosphere.
Deposition rates are quantified using the expression:
Flux = Vg C , [7-1]
which relates the flux of a pollutant gas to a surface to the atmospheric
concentration C above the surface. The deposition velocity, Vg, depends on
the specific gas/surface combination. Other factors influencing Vg are
humidity, surface roughness, air velocity, and turbulence. The determination
of Vg is usually made by measuring the change in concentration above the
surface or measuring the rate of deposition at the surface. Judeikis (1979)
has compiled deposition velocities for various materials in contact with
sulfur dioxide and ozone. Table 7-3 presents the deposition velocities for
sulfur dioxide. (More extensive discussion of deposition processes can be
found in Chapter A-7.)
7.1.1.5 Laboratory vs Field Studies—The effects of acidic deposition on
materials have been studied under both laboratory and field conditions. In
laboratory studies, the conditions of exposure can be controlled, and the
7-6
-------
TABLE 7-3. MEASURED DEPOSITION VELOCITIES FOR S02 ON VARIOUS SURFACES
(COMPILED BY JUDEIKIS 1979)
Surface9 Va (m min~1)b
Cement (5)
Limestone (6)
Copper
Leather (18)
Steel
Fabric (2)
Wood (7)
Aluminum (2)
Gloss Paint
Asphalt
Carpeting (3)
Wallpaper (17)
Solid Floor Materials (25)
0.6
> 0.021
> 0.001
~~ > 0.1
> U.001
~ 0.010
0.016
0.001
0.001
0.024
0.005
0.002
0.0003
-1.6
- 0.63
- 0.26
- 0.2
- 0.13
- 0.033
- 0.031
- 0.029
- 0.025
- 0.014
- 0.010
- 0.003
aNumber in parentheses indicates the number of different
materials examined if greater than one.
bAs defined by Equation 7-1 (x 1.667 = cm s"1.).
7-7
-------
specific effects of a single pollutant or environmental parameter can be
isolated. However, to produce measurable material damage in a reasonable
time period, the material is often exposed continuously to severe environ-
mental conditions (e.g., extremely high pollutant concentrations and/or high
humidity) completely unrepresentative of field conditions. Furthermore, the
exposure conditions are programmed through predetermined cycles that may only
remotely resemble the complex interactions of temperature, humidity, surface
wetness, sunlight, pollutant concentration, and other environmental factors
occurring in the ambient atmosphere. In this context, laboratory experiments
have thus far been unable to represent a true picture of the effects of pol-
lutants under conditions of long-range transport, where such transformation
would have had ample opportunity to take place.
Field studies normally consist of exposing samples of materials to ambient
atmospheres representing various combinations of pollutant concentrations and
other environmental factors. By comparing damage level (e.g., loss of sur-
face material) with pollutant concentration and other environmental factors
(e.g., humidity, "time-of-wetness", or pH of rainwater), statistical models
may be developed for the damage. The principal difficulties with this ap-
proach are:
° Materials exposed may not represent materials in actual use.
o In normal use materials are found in combination. Field studies
may not include interactions of other materials in contact with
test materials.
o Damage is a complex function of many environmental conditions, and
the effect of one condition is difficult to isolate.
° Measured variables may be interrelated (e.g., pH of rain may be
dependent upon S02 level).
Material damage is usually measured by noting quantitative changes in some
physical or chemical feature of the material (e.g., weight or thickness of a
sample; surface color, reflectivity or appearance degradation; chemical anal-
ysis and identification of corrosion products). Measurement methods will be
discussed in the appropriate subsections of Section 7.1.2.
7.1.2 Damage to Materials by Acidic Deposition
A wide range of sensitive materials can be damaged by acidic deposition.
However, this chapter will deal only with those choices judged to be econom-
ically and culturally important. These material classes are:
0 metals
0 masonry
0 paint and other coatings
7-8
-------
0 cultural property (historically and culturally valuable structures
and objects)
° other materials (paper, photographic materials, textiles, and
leather)
7.1.2.1 Metals--The atmospheric corrosion of metals is generally an elec-
trochemicalprocess governed by diffusion of moisture, oxygen, and acidic
pollutants (e.g., SOg) to the surface. The EPA Criteria Document for
Sulfur Oxides and Particulate Matter (U.S. EPA 1982) provides a review of the
primary mechanisms governing the corrosion of metals in the presence of SO?
and moisture. This review is based on the research of many workers, and it
deals primarily with the effects of S02 and moisture on metals and other
materials. However, most of the scenarios discussed fall within the general
definition of acidic deposition.
Moisture is always required for metal corrosion, each metal tending to have a
critical humidity above which corrosion tends to accelerate. Depending on
the specific metal, these critical humidities are in the range of 60 to 80
percent RH. The relative length of time a metal surface is wet ("time-of-
wetness") is the single most important variable affecting the acceleration of
corrosion by acidic deposition. Some workers (U.S. EPA 1982) have found that
hygroscopic corrosion products (e.g., iron sulfate) cause metal surfaces to
remain wet at lower RH than if these products were not present.
The position of metals in the electromotive series determines their relative
reactivity. However, the solubility of the particular metal salt and the
stability of the metal oxide coatings that tend to form in the atmosphere
determine metals' abilities to corrode as a result of acidic deposition. For
example, aluminum is high in the series, but aluminum oxide coatings that
form in the atmosphere resist corrosion even in the presence of significant
amounts of acidic deposition. However, even aluminum may be pitted in atmos-
pheres containing sea salt or large, acidic particles.
Thermodynamic considerations governing electrochemical corrosion are conven-
iently examined with the help of Pourbaix potential-pH diagrams. Plotting
electrical potential against solution pH can indicate regions of stability
for various chemical species. In simplified form, when reactions that form
soluble species occur, one has "corrosion"; when the free metal is stable the
region is designated "immune" to corrosion; and when a chemically stable
oxide or salt film forms on the surface, leaving the metal resistant to sub-
sequent attack, the region is one of "passivation" or mitigation of corro-
sion. Pourbaix (1966) has developed diagrams that show areas of stability,
corrosion, and passivity for various combinations of electrode potential and
pH, several of which are presented as Figure 7-1.
When using these diagrams to determine the effect of lowered pH on corrosion,
one must determine the potential attained by the metal in the natural envi-
ronment. Moreover, reduced pH tends to increase the solubility of corrosion
products. While the corrosion products in unpolluted atmospheres may be
relatively insoluble, in polluted atmospheres quite different corrosion
products that may be considerably more soluble may form. This potentially
7-9
-------
GOLD
SILVER
COPPER
LEAD
IRON
TIN
ZINC
CHROMIUM
ALUMINUM
LEGEND:
STABLE (IMMUNE)
CORROSION
PASSIVATION
Figure 7-1. Pourbaix diagrams for various metals. The ordinate is in
volts (electron potential standard hydrogen electrode) and
the abscissa is in units of pH. The upper thin diagonal line
is the 02 evolution line while the lower line is that for
\\2 evolution. Adapted from Pourbaix (1966).
7-10
-------
synergistic problem is sometimes overlooked in traditional writings on cor-
rosion. The Pourbai* jiam can give much insight into this process.
However, caution m „ exorcised in interpreting these diagrams because
kinetic factors witi, non-eq .brium behavior may govern corrosion.
Corrosion of metals may be measured by weight changes resulting from the
accumulated corrosion products before and after a predetermined exposure
period. However, during long exposures, corrosion products tend to spall or
wear off. Thus, corrosion products formed during the exposure period are
usually removed chemically to determine damage by weight of metal lost.
Another method applicable to metals is measurement of changes in sample
thickness, which in some cases may be obtained from the electrical resis-
tance. Mechanical tests involving bending are frequently used to test f^or
stress corrosion.
Physical methods such as scanning electron microscopy, x-ray diffraction, and
x-ray fluorescence can be used to characterize the physical and chemical na-
ture of corrosion products.
7.1.2.1.1 Ferrous metals. Corrosion of iron and steel in polluted atmos-
pheres has received a great deal of attention over the years. Steel, unless
it is an alloy designed for unprotected exposure, is usually coated by paint
or plating (e.g., zinc) when used in outdoor exposures. Nevertheless, data
on iron and steel corrosion provide valuable information on the relative
importance of acidic deposition components and the mechanisms causing damage.
The Pourbaix diagram for the iron system is presented as Figure 7-2. It
illustrates the relationships among normal corrosion products and the
equilibrium pH and potential conditions for their stability.
Some of the earliest work on the nature of iron corrosion in atmospheres con-
taining acid gases and moisture was that of Vernon (1935). He showed that in
the presence of S02 and moisture, iron corrosion proceeds from randomly
distributed centers he associated with the deposition of particulate matter.
Metal rusting is an oxidation process that is accelerated by the presence of
acidic pollutants. Barton (1976) has proposed the following set of reactions
involving the oxidation of 302 to sulfate on iron surfaces:
S02 + 02 + 2e- -»• 5042- [7-2]
4 HS03- + 3 02 + 4e~ -> 4 SCty2' + 2 H20. [7-3]
The electrons are provided by the oxidation of the metal (M):
M -»• Mn+ + ne- [7-4]
Barton (1976) noted that rusting of iron occurs first at isolated sites and
then spreads across the entire surface. This phenomenon is not well un-
derstood but may relate to a variety of factors including differential
deposition rates of S02 or acidic particulate matter, the influence of rust
deposits on subsequent corrosion, and variations in "time-of-wetness" in
7-11
-------
14
O OT
n.
UJ CO
o >
o
LU
0.0
-1.0 -
Figure 7-2. Pourbaix diagram for the system Fe, Fe2+, Fe3+,
Fe304, and Fe203. The thin diagonal lines indicate
regions of water stability. Compare with Figure 7-1 for
designated regions of "corrosion," "immunity, —'
"passivation." The reactions considered are:
and
1) Fe = Fe2+ + 2e-
2) 3Fe + 4H20 = Fe304
3) 3Fe2+ + 4H20 = Fe3
4) 2Fe2+ + 3H?0 =
5) Fe2+ =
6)
8H+ + 8e-
+ 8H+ + 2e-
+ 6H+ + 2e-
b. -^
+ e
u; dre- -r 3H20 = Fe203 + 6H+
7) 2Fe304 + H20 = 3Fe203
2e~
7-12
-------
relation to electrolyte concentrations at various points on the surface.
Rice et al. (1982) believe that moisture forms in "clusters" on metal sur-
faces even in indoor environments and at the site of these clusters, cor-
rosion is initiated. While rust deposits increase the absorption of SOg, a
thin layer of iron oxide on steel will provide some degree of protection from
subsequent atmospheric corrosion. In fact, special steel alloys whose iron
oxide layers provide considerable protection against further corrosion have
been developed for bold, unprotected exposures. The corrosion products on
several nonferrous metals (zinc, copper, and especially aluminum) tend to
suppress the absorption of S02-
According to Nriagu (1978), once corrosion has been initiated, the progress
of the reaction is controlled largely by sulfate ions produced from the
oxidation of absorbed or adsorbed SOg. However, the actual mechanism of
S02 oxidation on the surface is poorly understood. The work of Johnson et
al. (1977) appears to show that sulfur or sulfates are only a minor con-
stituent of the corrosion products of steel. Mild steel samples were exposed
to two urban areas near Manchester, England. One area was heavily polluted,
and the other was lightly polluted. Scanning electron microscopy, energy
dispersive x-ray analysis, and x-ray diffraction analysis of corrosion pro-
ducts showed them to be predominantly Y-Fe203-H20, «-Fe203 • H20 anda-FeOOH.
Some minor amounts of sulfur were found in a few of the samples. While not
discussed in the article, the possibility exists that any sulfates formed
were soluble and washed away. The relative amount of corrosion produced was
strongly dependent on whether the sample was initially wet at the beginning
of the exposure.
An iron oxide corrosion layer tends to reduce the rate of further corrosion
of iron and steel. Nriagu (1978) and Sydberger (1976) showed that steel
samples exposed initially to low concentrations of sulfur oxides were more
resistant to further corrosive attack than samples exposed continuously to
high concentrations. This suggests that the composition of the initial layer
is critical in determining the nature and extent of subsequent corrosion.
7.1.2.1.1.1 Laboratory studies. Exposing iron and steel samples to
S02 and moisture under controlled laboratory conditions has two principal
advantages:
1. The pollutant concentrations and other influencing factors can be
independently controlled in a factorial experiment and permit the
quantification of each factor's impact.
2. Exposure conditions can be made more severe than in nature to
accelerate the corrosion effect, thereby reducing the duration of
the experiment.
While many of the early experiments showed clearly that corrosion rates
correlate with both S02 and humidity, exposure consisted of S02 concen-
trations many times higher than those found in the ambient atmosphere, or in
what are referred to as "reflux" conditions, where water and excess S02
were continuously flushing the surface of the samples.
7-13
-------
The set of laboratory experiments most clearly approximating field conditions
was conducted by Haynie et al. (1976). Various materials were exposed to
controlled pollutant concentrations and moisture conditions at levels en-
compassing those found in ambient urban atmospheres. Sunlight and the for-
mation of dew were also simulated. Steel corrosion was determined in terms
of weight loss of the steel panels by chemically removing the corrosion
products, and the results showed a strong, statistically significant rela-
tionship between steel corrosion and S02 concentration, together with high
humidity.
7.1.2.1.1.2 Field studies. A inherent problem with field studies is
that iron and steel corrosion occurs even in unpolluted atmospheres, and the
impact of specific acidic deposition scenarios is difficult to isolate
completely. Therefore, the effects of acidic deposition can only be inferred
by statistical treatment of the data.
Upham (1967) exposed mild steel samples in a number of sites in and around
St. Louis and Chicago. He showed that corrosion correlated well with sulfur
oxide levels and increased with length of exposure. Starting in 1963, Haynie
and Upham carried out a five-year progam in which three different types of
steel were exposed in eight major metropolitan areas in the United States.
Multiple regression analyses showed significant correlations between average
S02 concentrations and corrosion for all three types of steel. No attempt
was made to relate damage to the joint occurrence of S02 and moisture
(relative humidity or time-of-wetness).
In 1964, Haynie and Upham (1971) exposed steel samples for 1 and 2 years at
57 stations of the National Air Sampling Network. Pollutants of interest
were S02, total suspended particulate matter, and the sulfate and nitrate
content of the particulate matter. An empirical function was developed
relating sulfate in particulate matter and humidity to corrosion. However,
the authors believed that S02 rather than sulfate was the causative agent
in producing corrosion, and the relationship was transformed into one based
on S02 from a linear regression between sulfate and S02- The corrosion
or damage function is:
cor = 325 /t etO-00275 S02-U63.2/RH)] [7-5]
where
cor = depth of corrosion, vm,
t = time, years,
S02 = S02 concentration, pg m~3, ''
RH = average annual relative humidity, percent.
Figure 7-3, based on the above damage functions, shows the relationship
between pseudocorrosion rate (cor v^-), relative humidity, and S02
concentrations. This graph shows that the corrosion rate is much more sen-
sitive to humidity than to S02, especially at levels of S02 normally
experienced in urban areas.
7-14
-------
cc
•z.
o
1/1
o
cc
C£.
O
o
o
AVERAGE S02 CONCENTRATION, jug m
-3
Figure 7-3. Steel corrosion behavior as a function of average sulfur
dioxide concentration and average relative humidity. Adapted
from Haynie and Upham (1974).
7-15
-------
For example, referring to Figure 7-3, if one were comparing relative cor-
rosion at 55 percent RH in two areas with average S02 levels of 100 and 150
yg nrj, a very significant difference in relative air pollutant levels,
the difference in relative corrosion would be approximately three pseudo-
corrosion units. On the other hand, if one were comparing relative corrosion
at a constant S02 level of 100 yg nr3 between two areas with a moderate
difference in average relative humidity (55 and 65 percent), the difference
in relative corrosion rate would be approximately 15 pseudocorrosion units.
This damage function shows that the sensitivity of corrosion to humidity is
far greater than that to S02, especially at levels of S02 normally ex-
perienced in urban areas.
A number of other damage functions relating steel corrosion to S02 and
humidity (or time-or-wetness) have been developed by several other workers
and have been summarized by U.S. EPA (1982) and Haagenrud et al. (1982). It
should be noted that nearly all metal corrosion damage functions have been
developed by regression analysis and most do not include terms for
precipitation.
A recent study of material damage in the St. Louis area in 1974-75 by
Mansfeld (1980) included the use of special atmospheric corrosion monitors
which measured the length of time that a corrosion panel was wet enough for
electrochemical corrosion to take place (time-of-wetness). Mansfeld1s sample
exposure array included weathering steel, galvanized steel, house paint, and
Georgia marble. Concentrations of S02 measured in this study were an order
of magnitude lower than those measured in Upham's earlier study (Upham 1967).
Mansfeld was unable to show any significant correlation between corrositivity
and pollutant levels.
Some of the experiments of Vernon (1935) showed that moist air polluted with
S02 and particles of charcoal produced corrosion much more rapidly than air
containing S02 and moisture alone. He reasoned that the effect of the
particles was primarily physical in that they increase the S02 concen-
tration. Sanyal and Singhania (1956) stated that particulate matter had a
"profound" effect on corrosion rates. They believed that the influence of
particulate matter on corrosion was related to its electrolytic, hygroscopic
and/or acidic properties, and its ability to absorb corrosive pollutant
gases. While these laboratory studies appear to show a strong influence of
particulate matter with corrosion, field studies have not confirmed this
effect.
Haynie (1983) has attempted to address the effects of small particles on
materials. Lacking a significant body of experimental data, he has ap-
proached the question theoretically, using data on deposition velocities. He
considered four species of small particles: carbon, sulfuric acid, ammonium
sulfates, and ammonium nitrate. He concludes that data from one study
(Harker et al. 1980) confirmed the chemical models for damage, and based on
calculated pollutant fluxes, S02-induced damage will tend to dominate over
H2S04 effects in most urban areas.
7-16
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Measurement of the effects of pollutants associated with long-range transport
(e.g., acid precipitation) as compared with locally-generated pollutants
(e.g., primary pollutant gases) is just getting under way in the United
States. The Scandanavians have been addressing this question for some years.
In summarizing several years' work in Norway, Haagenrud (1978) states that
monthly corrosion rates for carbon steel are strongly influenced by long-
range transport of both acid precipitation and S02- However, episodes of
precipitation of < pH 4.0 occur so seldom that these episodes do not strongly
influence long-term corrosion rates. Similarly, episodes of high S02 con-
centration also affected monthly corrosion rates, but had little effect on
long-term values because they occurred so seldom.
7.1.2.1.2 Nonferrous metals. The corrosion rates of commercially important
nonferrous metals in polluted atmospheres are generally less than those for
steel but cover a wide range. Figure 7-4, from the work of Sydberger and
Vannenberg (1972), shows adsorption of S02 with time at 90 percent relative
humidity for iron and three nonferrous metals. Copper and aluminum have
relatively low adsorption capacities for S02> confirming the lower sen-
sitivity of these metals to attack by S02 in the presence of moisture.
These tests were carried out by exposing polished metal surfaces to the test
conditions over very short exposure periods. While the results appear to
confirm the relative sensitivity of these metals to acidic deposition and
attack, the exposure conditions bear little relationship to real life con-
ditions. Rice et al. (1982) point out that a pure metal surface rarely
presents itself to the atmosphere for more than a few microseconds. Water is
rapidly absorbed in the surface films and may exist as moisture clusters as
pointed out in Section 7.1.2.1. Furthermore, corrosion products and salts
from surface contamination (e.g., chlorides) greatly influence corrosion
rates, principally through lowering of the critical humidity--the point where
corrosion rates begin to accelerate.
Only limited evidence links NOX with damage to nonferrous metals, though a
number of corrosion problems with telephone equipment have been traced to
NOX and high nitrate concentrations in airborne dust. In a laboratory
study of nickel-brass wire springs, stress corrosion cracking was observed
when surface concentrations of nitrate reached 2.1 mg cnr2 and RH was about
50 percent. To avoid the nickel-brass corrosion problem, zinc has been
eliminated from the alloy, and the cooling systems for existing equipment
have been modified to keep the RH below 50 percent in N0x-impacted areas
(Harrison 1975). Such damage to components in communications switch gear is
a serious problem because a simple malfunction can put a large system out of
service.
7.1.2.1.2.1 Aluminum. Aluminum is quite resistant to S02-related
acidic deposition. However, the presence of particulate matter may produce a
pitted or mottled surface in the presence of S02 and moisture. In view of
the reductions in emissions of S02 and particulate matter, especially
larger particles or agglomerates that could act as centers for corrosion
initiation, S02-related acidic deposition and surface corrosion of aluminum
do not appear to be a significant problem (Fink et al. 1971).
7-17
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OJ
en
OJ
o
(S)
CD
ce.
o
in
Q
I T
I T
ZINC
COPPER
ALUMINUM
10
EXPOSURE TIME, hr
Figure 7-4. Adsorption of sulfur dioxide on polished metal surfaces is
shown at 90 percent relative humidity. Adapted from Sydberger
and Vannenberg (1972).
7-18
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7.1.2.1.2.2 Copper. Copper and copper alloys in most atmospheres de-
velop thin, stable surface films, which inhibit further corrosion. Initial
atmospheric corrosion is a brown tarnish of mostly copper oxides and sulfides
that can thicken to a black film. Then in a few years, the familiar green
patina forms. Analysis of this film indicates it to be either basic copper
sulfate or, in marine atmospheres, basic copper chloride. However, in
coastal urban areas, the sulfate may still predominate (e.g., the Statue of
Liberty) because of the continuous availability of SOe over many years.
Nevertheless, both the sulfate- and chloride-based patinas are generally
resistant to further attack (Yocom and Upham 1977).
7.1.2.1.2.3 Zinc. Zinc is used primarily for galvanizing steel to make
it resistant to corrosion in the atmosphere and as an alloying metal with
copper to produce brass. Zinc as a coating on steel is anodic with respect
to steel, such that when zinc and steel are in contact with electrolyte, the
current flow protects the steel from corrosion at the expense of some oxi-
dation of zinc.
Because of its economic importance, the behavior of zinc in the presence of
acidic deposition has been studied intensively by a number of workers.
Guttman (1968) carried out long-term measurements of atmospheric corrosion of
zinc from which he developed a damage function for zinc corrosion in relation
to S02 concentrations and time-of-wetness. Time-of-wetness was measured by
means of a dew detector. S02 was measured by lead peroxide sulfation can-
dles and conductiometric S02 measurements. Guttman1s damage function is
Y = 0.005461(A)°-8152x (B + 0.02889)f [7.5]
where
Y = corrosion loss, mg for a 3 x 5 inch panel,
A = time of wetness, hr,
B = atmospheric S02 content during the periods that the panels
were wet, ppm.
Haynie and Upham (1970) carried out an extensive zinc corrosion study in
eight cities wherein zinc panels were exposed, while concurrently collecting
data on S02, temperature, and humidity. They developed the following em-
pirical damage function relating zinc corrosion to S02 levels and relative
humidity:
y = 0.001028 (RH - 48.8) S02, [7-7]
where
y = corrosion rate, m yr*1,
RH = average annual relative humidity,
S02 = average S02 concentration, yg m-3.
Note that in Equation 7-6 moisture is in terms of time-of-wetness while in
Equation 7-7 annual average relative humidity is used. Time-of-wetness is a
7-19
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far more relevant indication of surface moisture than average relative hu-
midity when corrosion and other forms of moisture-enhanced material damage
are being considered. For example, if Equation 7-7 is applied in an area
that has annual average relative humidity significantly less than 48.8 per-
cent, no corrosion is implied. Yet in such areas, surfaces become wet with
dew or seasons of high humidity occur and corrosion proceeds even when annual
average relative humidity is below the critical value obtained by regression
analysis. Similarly, the equation indicates no damage in the absence of
S02, ignoring damage due to moisture, etc., in the absence of S02-
The damage coefficients for these two functions plus functions developed from
other studies were compiled by U.S. EPA (1982). These coefficients are com-
pared in Table 7-4. Additional zinc damage functions have been reviewed by
Haagenrud (1982).
7.1.2.2 Masonry—The term "masonry" is applied to a large number of building
and decorative materials exhibiting a broad range of surface reactions to
physical and chemical stresses imposed by the environment. The importance of
acidic deposition to this class of materials may be related to the effect
produced directly on a single material (e.g., limestone or marble) or direct
or indirect damage to composite masonry systems. An example of direct damage
to composite systems involves the rusting of steel reinforcing bars in con-
crete, which expand and crack the concrete. Indirect damage includes damage
to brick-mortar systems in which the relatively reactive mortar is damaged
directly by acidic materials and rainfall; then the salts released by these
reactions diffuse into the brick, causing stress and subsequent spall ing.
Samples of building materials such as stone, mortar, and concrete can be
weighed before and after exposure to determine erosion rates. Caution must
be exercised in interpreting such data because conversion to new phases may
involve weight gain without obvious change in physical appearance. Discol-
oration of such samples from exposure to dark particulate matter can be
measured photometrically. A series of photographs of buildings taken over
sufficient time periods may provide a qualitative assessment in the form of
soiling and/or loss of surface detail. Dimensional changes and analysis of
concrete sections may also provide useful indications of damage.
7.1.2.2.1 Stone. The accelerated decay of stone buildings and monuments in
highly industrialized areas has been documented by comparing current con-
dition with historic photographs and plaster casts. Photographs taken in
1908 and 1969 of a sandstone sculpture carved in 1702 in Westphalia, West
Germany, demonstrate a dramatic loss of material during the past 60 years and
virtual obliteration of the object (Winkler 1982). Similarly, comparison of
a plaster cast made in 1802 with a photograph taken in 1938 demonstrates
substantial deterioration of a sculpture on the west frieze of the Parthenon
(Plenderleith and Werner 1971). A detailed account of the restorations of
the Acropolis and measurements of the thickness of gypsum layers formed on
its exposed marble surfaces is presented by Skoulikidis (1982). The de-
teriorating conditions of the Caryatids of the Erechthion led to their
replacement with replicas and their removal to the controlled environment of
the Acropolis Museum (Yocom 1979).
7-20
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TABLE 7-4. EXPERIMENTAL REGRESSION COEFFICIENTS WITH ESTIMATED
STANDARD DEVIATIONS FOR SMALL ZINC AND GALVANIZED STEEL
SPECIMENS OBTAINED FROM SIX EXPOSURE SITES
Study
Time of wetness
coefficient S02 coefficient^
(urn yr'l) (ym yr'Vyg m~I)
Number
of
data
sets
Field Studies
CAMP (Haynie and Upham
1970)
ISP (Cavender et al.
1971)
Guttman 1968
Guttman and Sereda 1968
St. Louis (Mansfeld 1980)
1.15 j^O.60
1.05 j^O.96
1.79
2.47 +_ 0.86
2.36 + 0.13
0.081^0.005 37
0.073^0.007 173
0.024 < 400
0.037^0.008 136
0.022 + 0.004 153
Chamber Study
Haynie et al. 1976
1.53 + 0.39
0.018 + 0.002
96
al ppm S02 = 2620 pg nr3
7-21
-------
Stones composed almost entirely of calcium carbonate (limestone, marble,
travertine, etc.) or stones whose cementing material is calcium carbonate are
particularly vulnerable to damage from acidic deposition. The attack of
sulfur dioxide on such carbonate stones has been studied for over a century.
Yet, no quantitative relationship has been developed between ambient SO?
levels and resulting materials damage.
The general decay mechanism includes aerodynamic factors controlling delivery
of SOe to the stone surface, oxidation of S02 to sulfate and the sub-
sequent reaction with the carbonate surface, mechanical stress by which
reaction products destroy the stone structure, and removal of the stone and
its alteration products by rainfall and other weathering phenomena
(Livingston and Baer 1983).
Although the primary air pollutants causing damage to stone are sulfur com-
pounds, a comprehensive decay mechanism must include the roles of nitrogen
compounds, carbon dioxide, and water. For the carbonate stone/sulfur com-
pound system three general modes of attack pertain:
Gaseous S02
S02 + CaC03 •*• CaSOa + C02 (Step 1) [7-8]
CaSOs + 1/2 02 + CaS04 (Step 2) [7-9]
Wet Deposition
H2S04 + CaC03 •*• CaS04 + H20 + C02 [7-10]
Dry Deposition is exemplified by the reaction between sul fates in parti cul ate
matter and calcium carbonate either in the form of sul f uric acid as in wet
deposition, or in the form of ammonium sul fates (Stevens et al 1980).
(NH4)2S04 + CaCOa + CaS04 + (NH4)2C03 [7-11]
NH4HS04 + CaCOs •> CaS04 + NH4HC03 [7-12]
The anhydrous CaS04 is hydra ted to form gypsum, which is highly susceptible
to surface erosion.
Humidity plays a key role in all aspects of the interactions of SOX with
carbonaceous stone. In autoradiographic experiments using sulfur-35,
Spedding (1969b) showed surface saturation of oolitic limestone samples by
S02 at 81 percent RH occurring in less than ten minutes. However, at the
same concentrations but at 11 percent RH only a few distinct sites showed
reaction after 20 minutes exposure, with approximately 25 percent of the
total S02 uptake measured for the high humidity case. Tombach (1982) has
summarized the many factors contributing to stone decay as shown in Table
7-5.
Few quantitative studies of air pollution damage to stone have been reported,
although the increased rate of erosion for marble tombstones in the urban
7-22
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TABLE 7-5. CLASSIFICATION OF MECHANISMS CONTRIBUTING TO STONE DECAYC
(ADAPTED FROM TOMBACH 1982)
Mechani sm
Temper-
Rainfall Fog Humidity ature
Solar
Insolation Wind
Gaseous
Pollutants Aerosol
I
r\>
CO
External Abrasion
Erosion by wind-borne particles
Erosion by rainfall
Erosion by surface ice
Volume Change of Stone
Differential expansion of mineral grains
Differential bulk expansion due to uneven heating
Differential bulk expansion due to uneven moisture
content
Differential expansion of differing materials at
joints
Volume Change of Material in Capillaries and Interstices
Freezing of water
Expansion of water when heated by sun
Trapping of water under pressure when surface freezes
Swelling of water-imbibing minerals by osmotic pressure
Hydration of efflorescences, internal impurities, and
stone constituents
Crystallization of salts
Oxidation of materials into more voluminous forms
Dissolution of Stone or Change of Chemical Form
Dissolution in rainwater
Dissolution by acids formed on stone by atmospheric
gases or particles and water
Reaction of stone with S02 to form water-soluble
material
Reaction of stone with acidic clay aerosol particles
Biological Activity
Chemical attack by chelating, nitrifying, sulfur-
reducing or sulfur-oxidizing bacteria
Erosion by symbiotic assemblages and higher plants
that penetrate stone or produce damaging excretions
aSo!1d circles denote principal atmospheric factors; open circles denote secondary factors.
-------
environment of Edinburgh was observed as early as 1880 (Geike 1880). A
study of tombstones in U.S. National Cemeteries (Baer and Berman 1983) has
developed methodology for measuring damage to marble headstones exposed to
the environment for 1 to 100 years. The study's data base2 consists of
measurements of some 3,900 stones in 21 cemeteries distributed throughout the
United States. The factors affecting damage rates include grain size, total
precipitation, and local air quality.
In the United States, measured rates of marble deterioration have generally
been small, on the order of 2.0 mm per 100 years (Winkler 1982). This is
substantially less than values reported for stones exposed in urban areas in
Europe although direct comparison is difficult because the stones exposed in
Europe are generally more reactive.
Comparing the condition of similar samples of sandstone exposed in different
areas of Germany for about 100 years, Luckat associated the large differences
in observed deterioration with trends in local air quality (Luckat 1981,
Schreiber 1982). These results presented in Table 7-6 describe stones openly
exposed to the environment. For similarly reactive test stone specimens
protected from the direct action of rain and placed at 20 locations in West
Germany, the following functions correlating reaction with SOe immission
(uptake) rate were obtained:
Baumberg sandstone U = 0.54 D; r2 = 0.92 [7-13]
Krensheim shell limestone U = 0.22 D; r2 = 0.72 [7-14]
When similar test samples were exposed to the rain the following damage
functions were obtained:
Baumberg sandstone L = 0.03 D + 0.5; r2 = 0.36 [7-15]
Krensheim shell limestone L = 0.018 D + 0.6; r2 = 0.80 [7-16]
where:
U = S02 immission rate of the stone in {mg nr2 day-1) by weight
gain of standard stone,
D = by weight gain S02 immission rate, IRMA measured value (mg
nr2 day'1),
L = loss in weight, and
r = correlation coefficient.
2Note: the data base has continued to grow and so is larger than that of
the study cited.
7-24
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TABLE 7-6. DETERIORATION OF SCHLAITDORF SANDSTONE EXPOSED FOR
100 YEARS IN WEST GERMANY (AFTER SCHREIBER 1982)
Relative S02
immission Rate,3
Monument Location mg nr2 day1 Deterioration
Neuschwanstein
Castle
Ulm Cathedral
Cologne Cathedral
Fussen
Ulm
Cologne
6
48
111
Practically
Moderate
Very severe
none
aRelative immission or uptake rate of S02, annual average (August
1973 - July 1974) measured by IRMA method. (See Baer et al. 1983 for
details of the technique.)
7-25
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Hie high contribution of the non-SOx factors for stones exposed to rainfall
suggests that damage functions for stone must specifically address such var-
iables as other pollutants and rainfall as well as the initial physical and
chemical properties of various stones.
A series of measurements made at St. Paul's Cathedral, London on the Portland
stone (biosparite limestone) balustrade, demonstrate a high rate of weath-
ering (Sharpe et al. 1982). Using lead plugs filled in openings in the stone
in 1718 as base level references, a mean rate of lowering of 0.078 mm yr-1
was obtained for the period 1718-1980. The balustrades represent conditions
of exposed rain flow. Similarly, by use of a micro-erosion meter (dial mi-
crometer gauge mounted on reference studs) a current erosion rate of 0.139 mm
yr"1 was obtained for six sites on the cathedral. These sites represent
drip erosion zones. Though the two sets of data are not strictly comparable,
both represent substantially higher rates of loss than observed for marble in
the United States.
7.1.2.2.2 Concrete. World production of concrete amounts to some 3 billion
cubic meters per year. Cement, concrete, and steel reinforced concrete
structures are all subject to complex actions and many important structures,
e.g., bridge decks, highways, military installations, and naval shore struc-
tures suffer from severe durability problems (NMAB 1980). Similarly, concern
has been exposed over leaching of possibly toxic components of cement cul-
verts transporting acidified water (see Section 7.2).
The alkaline nature of cement has led to general neglect of the effects of
acid deposition and acidified water runoff on concrete/cement durability
although it is recognized that any reaction reducing matrix alkalinity will
be harmful. The role of chloride ion as a major contributor to corrosion of
reinforced concrete is well established (Volkwein and Springenschmid 1981,
Browne 1981). The alkalis in the hardened cement passivate the reinforcing
steel while penetrating chlorides depassivate the iron. Other factors in
corrosion of the steel include the development of electrolytic corrosion
cells and the penetration of atmospheric Og through the concrete to the
steel. The reaction of S02 and $042- with cement involves the for-
mation of cacium sulfate and calcium sulfate aluminum hydrate (ettringite).
The highly alkaline nature of cement/concrete leaves such surfaces vulnerable
to acidic deposition. The principal mode of attack on concrete is loss of
alkalinity by reaction with COe- Spedding (1969b), reporting on the con-
tamination/decontamination of laboratory surfaces accidentally exposed to
sulfur-35/sulfur dioxide, observed that good decontamination was obtained by
simple water washing. This suggests that the reaction products of the depo-
sition of S02 on concrete are water soluble. The high volume of water flow
through rain collecting and distribution culverts in drinking water systems
also raises questions about the possible release of toxic materials leached
from the concrete matrix.
Similar concerns have been expressed over the erosive effects of acidified
streams on concrete bridge piers. The literature reveals only limited
research on the effects of acidified water runoff on concrete durabil-
ity. Cements used in dams and culverts require a special formulation for
7-26
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sulfate resistance when exposed to concentrations in excess of 200 ppm in
water (Nriagu 1978). Specialized concretes in which sulfur replaces cement
as the binding agent have been developed by the Bureau of Mines for resis-
tance to acid and salt attack and damage from freeze-thaw cycling (Sulphur
Institute 1979).
7.1.2.2.3 Ceramics and Glass. Although enamels and glasses are quite resis-
tant to chemical attack by air pollutants, in certain circumstances damage
has been observed. In a three-year exposure study on porcelain enamels
placed in seven U.S. cities, some change in surface condition of the enamel
was observed although the base metal was protected (Moore and Potter 1962).
Glass weathering is the process of removing alkali cations (e.g., Na+ and
K+) from glass by reaction with water or sulfur dioxide. The reaction with
water involves the exchange of sodium ions by hydrogen ions with the rate of
reaction limited by the diffusion of sodium ions to the surface. The reac-
tion with sulfur dioxide in the range 20 to 100 C in gas saturated with S02
involves the same process at approximately the same rate as with water alone
(Douglas and Isard 1949).
Fluorides, especially HF, are capable of attacking on a wide variety of cera-
mic materials and glasses. Restrictive legislation on fluoride emissions
has, for the most part, eliminated fluoride-induced damage.
Perhaps the most serious glass damage problem is that associated with the
decay of medieval stained glass windows. The unique composition of these
glasses combined with their open exposure to the atmosphere makes them
particularly susceptible to deterioration. This problem is discussed in
detail in Section 7.1.2.5.3.
Properly fired brick is highly resistant to attack by air pollutants while
poorly fired brick is highly susceptible to chemical attack. Acidic sol-
utions accelerate such damage, increasing the rate of reaction 10-fold over
water alone. Residual sulfates from decay of mortars can combine with other
salts to produce failure in brick (Robinson 1982).
7.1.2.3 Paint—Paint damage from acidic deposition is strongly related to
the paint formulation. Such factors as the ratio of pigment and extenders to
film-forming ingredients determine the hardness, flexibility, and perme-
ability of the surface. It has been shown that the presence of extremely
high concentrations of S02» a reducing gas, can interfere with the drying
process, which is an oxidation-polymerization reaction (Holbrow 1962).
However, it is doubtful that S02 concentrations at present in any area of
the United States would be high enough to cause this potential problem.
The most realistic mechanism for damage to paint by acidic deposition is
reaction between acidic materials and pigments (e.g., ZnO) and extenders such
as CaC03. The long-term effect is the loss of paint surface through ero-
sion, so measurement is most conveniently done by measuring weight loss of
painted panels. Surface darkening by deposits of particulate matter or
7-27
-------
reactions between pigments and air pollutants are usually measured photo-
metrically.
Paint consists of pigment and vehicle. Pigments, such as titanium dioxide
and zinc oxide, provide color, hiding power, and durability. Sometimes fil-
lers such as calcium carbonate or inorganic silicates are also added. The
vehicle provides the film-forming properties of the paint and contains resin
binders, solvents, and additives. Together, the pigment (along with fillers)
and vehicle protect the underlying surface and enhance the appearance of the
exposed surface. Air pollution may limit both of these functions by damaging
the protective coating, thus exposing the underlying surface to attack and/or
spoiling the appearance of the surface. The most important potential effects
of S02 on paints are interference with the drying process and acceleration
of the normal erosion process.
The primary effect of particulate matter on paint is soiling. Soluble salts
such as iron sulfate contained in deposited particles can also produce stain-
ing. Chemically active large particles such as acid smut (or soot) from
oil-fired boilers, mortar dust near building demolition sites, or iron par-
ticles from grinding operations can severely damage automotive paint (Yocom
and Upham 1977). The effects range from discoloration of the paint film to
corrosion of the underlying metal in the vicinity of individual particles.
Large particles becoming imbedded in a freshly painted surface can act as
wicks to transfer moisture and corrosive pollutants such as S02 to the
underlying material's surface.
Hoi brow (1962) has reported a number of experiments to determine effects of
sulfur dioxide on newly applied paints. Drying times for various oil-based
paints exposed to extremely high concentrations of S02 (1 to 2 ppm) were
increased 50 to 100 percent. Thus far no experiments have been carried out
on the effect of S02 on drying time of water-based latex paints.
Campbell et al. (1974) carried out an extensive study of paint erosion for a
variety of paint types and exposure conditions (including S02 and 03).
Both chamber and field experiments were conducted. The researchers evaluated
four important types of paint:
1. Acrylic latex and oil-based house paints,
2. Urea-alkyd coil coating for sheet metal in coil form,
3. Nitrocellulose-acrylic automotive refinishing paint, and
4. Alkyd industrial maintenance coating.
Table 7-7 presents the principal findings of this work.
Generally, exposures to high concentrations (1 ppm of both S02 and ozone)
produced statistically significant erosion rate increases compared to clean
air (zero pollution) conditions. Oil-based house paint experienced the
largest erosion rate increases. The greater susceptibility of oil-based
7-28
-------
TABLE 7-7.
PAINT EROSION RATES AND PROBABILITY DATA (T-TEST) FOR CONTROLLED ENVIRONMENTAL
LABORATORY EXPOSURES (ADAPTED FROM CAMPBELL ET AL. 1974)
Type of Paint
House paint
011
latex
Coll coating
Automotive refinlsh
Industrial maintenance
Mean Erosion Rate {nm hr"1 with 95
confidence limits) for unshaded
Clean air S02
control (1.0 ppm)
5. 11 +_ 1.8 35.8^4.83*
0.89+^0.38 2.82^0.253
3.01^0.58 8.66^1.193
0.46 +_ 0.02 0.79^0.66
4.72^1.30 5.69 +_ 1.78
percent
panel s
°3
(1.0 ppm)
11.35 +_ 2.673
2.16 +_ 1.50b
3.78 ^0.64"
1.30 +_ 0.33a
7.14 +_ 3.56
PAINT EROSION RATES AND PROBABILITY DATA (T-TEST)
FOR FIELD EXPOSURES (ADAPTED FROM CAMPBELL ET AL. 1974)
Mean Erosion Rate (nm hr"1 with 95 percent confidence
limits) for panels facing 'south
Type of Paint
House paint:
oil
latex
Coil coating
Automotive refinish
Industrial maintenance
Rural
(clean air)
109
46
53
23
91
i 191
± 13
+_ 20
+ 28
± 41
Suburban
376
76
254
58
208
+_ 124a
i 183
+ 48a
+_ isb
+_ 361b
Urban
(SO? dominant),
- 60 yg fir3
361
97
241
41
168
+_ 124b
±8b
+_ 203
+ 10
^99
Urban
(oxidant dominant),
- 40 vg m-3
533
165
223
43
198
+_ 157a
+ 142
+ 433
+ 10
+_613
Significantly different from control at p = 0.01.
bSignificantly different from control at p = 0.05.
Note: 1 ppm S02 = 2620 wg m-3
7-29
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house paint to S02 was attributed to the use of extenders such as
or metal silicates. Latex and coil coatings experienced moderate increases,
and the industrial maintenance coating and automotive refinish experienced
the smallest increases. In general, exposures to S02 produced higher
erosion rates than ozone. Unshaded panels eroded more than shaded panels.
Exposures to 0.1 ppm pollutants did not produce significant erosion rate
increases over clean air exposures. It should be noted that even these lower
concentrations are high when compared with average concentrations found in
the ambient air of urban areas.
In the field portion of this same study, painted panels were exposed at four
locations with different environments:
1. Rural - clean air (Leeds, North Dakota),
2. Suburban (Valparaiso, Indiana),
3. Urban - sulfur dioxide-dominant (Chicago, Illinois), and
4. Urban - oxidant-dominant (Los Angeles, California).
In most cases, southern exposures produced somewhat larger erosion rates,
which agreed with the unshaded versus shaded results of the laboratory study.
Oil-based house paint again experienced by far the largest erosion rate in-
creases, followed in order by the urea-alkyd coil coating, latex house paint,
industrial maintenance paint, and automotive refinish. Generally, the field
exposures showed that the relative paint erosion rate was about the same for
the sulfur dioxide-dominant as for the oxidant-dominant location, which ap-
peared to contradict the chamber studies. However, the authors believed that
differences in the pollutant mix at the two locations and especially the
presence of nitrogen oxides at the oxidant-dominant site could have enhanced
the erosion rate at this location, bringing it up to the level of damage at
the sulfur dioxide-dominated location (Campbell et al. 1974).
It is noteworthy that the oil-based house paint and urea-alkyd coil coating
experienced the largest erosion rate increases in both the field and labo-
ratory sulfur dioxide exposures. These coatings were the only ones that
contained a calcium carbonate extendei—a substance sensitive to attack by
acidic materials.
Spence et al. (1975) summarized the results of paint exposure to several
gaseous pollutants from the full-scale chamber studies reported by Haynie et
al. (1976) and discussed earlier in relation to metal exposures. Four
classes of painted surfaces were evaluated: oil-based house paint, vinyl-
acrylic latex house paint, vinyl coil coating, and acrylic coil coating. A
strong correlation was found between paint erosion for the oil-based house
paint and S02 and humidity. The vinyl and acrylic coil coating were
unaffected, but blistering was noted on the latex house paint. It was not
certain if the blistering was the result primarily of SOg or moisture.
A multiple regression relationship was developed for the joint influence of
S02 and relative humidity on the oil-based house paint:
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E = 14.3 + 0.0151 S02 + 0.388 RH [7-17]
where
E = erosion rate of ym yr"1,
S02 = concentration of S02 in ^9 "i~3»
RH = means annual relative humidity in percent.
This relationship indicates that paint erosion is significantly more sensi-
tive to changes in humidity than to S02 concentration. However, one must
be careful in using models based on accelerated chamber tests for actual
exposures because Equation 7-17 would predict that in an atmosphere with no
S02 present, with an average relative humidity of 50 percent, the paint
erosion rate would be about 34 ym yr"1. Assuming a typical paint thick-
ness of 50 ym, the paint film would be completely eroded away within 1.5
years.
The present understanding of damage to paint from air pollution is based
primarily upon two sets of chamber studies and one set of field exposures.
Because the field studies were carried out in the early 1970s, further labo-
ratory and field studies are needed to determine the importance of paint
damage from present levels of sulfur oxides. Furthermore, these studies
should include present formulations (especially water-based paints) that may
have a different response to air pollutants from those used earlier. At the
present time, the effects of air pollutants on paint films are not well
enough understood to provide meaningful dose-response relationships including
all relevant causes of damage (e.g., moisture, insolation, oxidation). In
addition, one should note that paint formulations change frequently so that
the compositions of paints currently in use may bear little resemblance to
the formulations examined in earlier studies.
7.1.2.4 Other Materials—In addition to coatings, a wide range of organic
materials are found to be susceptible to attack by atmospheric pollutants.
These materials, including paper, photographic materials, textiles and
leather, were not considered in the EPA1 s criteria documents, so they are
considered here, although the indoor locations in which they are normally
found dictate gaseous transport mechanisms for deposition.
Most organic materials exposed to the atmosphere are quite resistant to the
effects of acidic deposition. Deterioration of such materials is determined
primarily by the effects of atmospheric oxygen, ultraviolet (UV) light, and
atmospheric oxidants such as ozone.
The degradation of paper and textiles is dominated by three factors: light,
humidity, and acidity. Paper and other cellulosic materials (e.g., cotton,
linen, and rayon) are highly susceptible to acid hydrolysis at the glucosidic
linkage in the cellulose chain. Among proteinaceous textile materials silk
is most susceptible to damage by light. In bright light silks may lose 60
percent of their strength in as little as 8 weeks of exposure (Leene et al.
1975).
7-31
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7.1.2.4.1 Paper. The embrittlement of paper is accelerated by exposure to
acidic deposition. Excess acidity can be observed by combination surface
electrode pH measurements. Resulting damage may be determined by measuring
folding resistance.
The role of sulfur dioxide in the deterioration of paper has been accepted
since the 1930's. Early experiments (Langwell 1952, 1953) relied on un-
real istically high S02 concentrations of 5,000 ppm interacting with damp
paper. Hudson and Milner (1961) used sulfur-35 as a radioactive tracer to
demonstrate that measureable amounts of S02 were rapidly deposited in
paper. Working with concentrations of 10 ppm, Grant (1963) showed that S02
deposition increased with increasing aluminum sulfate/resin sizing of the
paper.
A comparative study of identical copies of twenty-five 17th and 18th cen-
tury books in two British libraries, one in an unpolluted atmosphere in
Chatsworth, the other in the badly polluted urban atmosphere of Manchester,
revealed a significant increase in paper acidity in the Manchester library
(Hudson 1967). This acidity was greatest at the page edges and decreased
greatly toward the center of the page, which might be considered the initial
sheet acidity.
Wallpapers form an important part of the indoor surface area available for
SOg sorption. Spedding and Rowlands (1970) measured the sorption charac-
teristics of PVC and conventional wallpapers on exposure to maximum initial
S02 concentrations of 150 yg nr3. Sorption depended largely on surface
finish and design pattern, with greater sorption by conventional wallpapers.
The researchers suggested that S02 sorption accelerated the deterioration
of wallpaper.
7.1.2.4.2 Photographic Materials. Under normal conditions of temperature
and relative humidity, paper, acetate film, and other photographic materials
are oxidized at a very slow rate. One of the most serious factors in the
preservation of photographic materials is the presence of large quantities of
oxidizing gases: hydrogen sulfide, sulfur dioxide, and to a lesser extent
NOX, peroxides, formaldehyde, and ozone (Eastman Kodak 1979).
The effect of these pollutants is usually yellowing and fading of the silver
image. The paper base may also be degraded and stained. Acidic gases will
degrade gelatin, paper, and the film base of negatives (Eastman Kodak 1979).
Agfa produces a colloidal silver test strip which is 8 to 10 times more sen-
sitive to gaseous pollutants than ordinary photographic materials. In a
survey of major libraries and archives using this technique many examples of
significant air quality problems were observed (Weyde 1972).
7.1.2.4.3 Textiles and Textile Dyes. Certain textile materials are weakened
by acidic deposition. Such damage is best determined by measuring loss in
tensile strength. Cotton is also weakened by biological processes (e.g.,
mildew), and methods have been developed to differentiate between acidic
deposition and these biological mechanisms by determining the relative mo-
lecular weight of the exposed material. Damage from acidic chemical attack
7-32
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causes depolymerization and reduction in average molecular weight, while
biological attack causes essentially no reduction in average molecular
weight.
Textile dyes are affected by N02. Changes in color values from such damage
are measured by specially designed colorimeters or spectrophotometers capable
of detecting small changes in color within narrow ranges of the visible
spectrum.
Sulfur oxides are capable of causing deterioration to natural and synthetic
fibers. Cotton, like paper, a cellulosic fiber, is weakened by sulfur di-
oxide. Under circumstances where sulfuric acid comes in contact with a
cellulosic surface, the product of reaction is water soluble with little
tensile strength (Petrie 1948). In field tests in St. Louis, cotton duck
exposed to varying SOX levels showed a direct relationship between loss in
tensile strength and increasing SOX concentration (Brysson et al. 1967).
Zeronian (1970) exposed cotton and rayon fabrics under accelerated aging
conditions of light and water spray with and without 0.1 ppm S02. Loss in
strength was 13 percent in the absence of S0£ and 22 percent in the pre-
sence of S0£. In a study of nylon fabrics exposed to 0.2 ppm S02 under
similar conditions, he found that nylon fabrics lost 40 percent of their
strength under the S02 free conditions and 80 percent of their strength in
the presence of S02 (Zeronian et al. 1971).
The degradation of nylon 66 by exposure to light and air is increased by the
addition of 0.2 ppm of S02 to the air. Chemical properties and yarn ten-
sile properties both reflect this damage (Zeronian et al. 1973). Results
demonstrated that the mode of degradation is not changed although S02 ac~
celerates the rate of reaction.
Among proteinaceous textiles, silk is most vulnerable to the effects of
light, acidity, and sulfur dioxide, demonstrating much greater loss in
strength than wool (Leene et al. 1975).
Damage to textiles has been attributed to NOX (Harrison 1975). Such damage
has been caused both by loss of fiber strength and by fading of textile dyes.
Significant reduction in breaking strength and increase in cellulose fluidity
were observed for combed cotton yarns exposed in Berkeley, California, to
unfiltered air compared to those exposed to carbon-filtered air (Morris et
al. 1964). Both sets of samples were unshaded and exposed at a 45° angle
facing south. Though the authors did not isolate the effects of individual
pollutants,, they implied that compounds associated with photochemical smog,
especially NOX, were the probable cause of increased damage.
In an EPA chamber study of the effects of individual pollutants on 20 dyed
fabrics, it was demonstrated that N02 at 0.1 to 1.0 ppm produced appre-
ciable dye fading, and S02 at 0.1 to 1.0 ppm caused visible fading on wool
fabrics" (Beloin 1973). It was also concluded that higher temperatures and
relative humidities increase dye fading and that the rate of fading as a
function of exposure time appears to be nonlinear.
7-33
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7.1.2.4.4 Leather. Michael Farady is credited (Parker 1955) with having es-
tablished in 1843 a link between the rotting of leather armchairs in the
London Atheneum Club and sulfur dioxide emitted by its gas illumination.
Plenderleith (1946), Innes (1948), and Smith (1964) describe the sequence of
chemical deterioration for leather and consider possible mitigative actions.
It has been observed that leather initially free of sulfuric acid will
accumulate up to one percent acid by weight per year if exposed to an atmos-
phere containing S02. The mechanism is thought to involve metal ion-
catalyzed conversion to sulfuric acid of the S02 absorbed by the collagen
of the leather. Using sulfur-35 labelled S02, Spedding et al. (1971)
showed that it is sorbed evenly over the leather surface, with the limiting
factor in uptake being gas-phase diffusion to the surface. Weakening of
leather caused by acidic deposition can be quantified by means of tensile
strenght tests.
7.1.2.5 Cultural Property—It has been estimated that the United States has
over 6,000 museums, historical societies, and related institutions; more than
10,000 entries on the National Register of Historic Places, and in excess of
26,000 libraries and archives of substantial size (NCAC 1976). Light, oxi-
dation, fluctuations in humidity, and chemical pollutants threaten this
precious cultural heritage.
Damage to cultural property cannot be quantified in simple dose-response
terms. Just as an electrical conponent may require replacement due to cor-
rosion of a fraction of its mass, or stress-corrosion fracture may lead to
failure of a mechanical system, damage to the texture of sculpture or the
surface of a fresco exposed to the environment diminishes their aesthetic
importance far in excess of the amount of material damage. Still more
critical is the circumstance that, for most cultural property, replacement is
impossible. What is lost is lost.
7.1.2.5.1 Architectural Monuments. Historic and artistic structures re-
present the single most visible aspect of our history and culture. For
the United States, legislation providing a mandate for preservation began
with the Antiquities Act of 1906, followed most recently by the Historic
Preservation Act Amendments of 1980. In Canada, the Archaeological Sites
Protection Act and the Historic Sites and Monuments Act were adopted in 1953.
Architectural monuments are universally threatened by the effects of pol-
lution and urbanization as well as by weathering cycles and other natural
phenomena (NAS 1979). Although damage to these monuments is frequently
attributed to acid precipitation, no clear evidence providing a cause and
effect relationship between acid precipitation and damage to a specific
monument exists. In general, it appears that while acidic deposition can
effect significant damage to cultural property, the sources are predominantly
of local origin.
7.1.2.5.2 Museums, Libraries and Archives. As discussed above, the sorption
of SOX and NOX by organic materials in the indoor environment is well
established. In some cases, as in paper and leather embrittlement, dye
fading, and "red-ox" blemishes on microfilm, a direct relationship between
7-34
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pollutant sorption and damage has been established. This has led major
museums, libraries, and archives to install scrubbers for the removal of acid
gases.
Among the systems in use are activated charcoal and Purafil (activated alu-
mina impregnated with KMntty) dry scrubbers and alkaline wash wet scrubbers.
Such systems have been introduced as part of new construction or retrofitted
at the National Gallery (London), the Library of Congress (Washington, D.C.),
the Newbury Library (Chicago), and the National Gallery (Washington, D.C.).
Many other collections of cultural artifacts are preparing for the eventual
retrofitting of their air handling systems to use scrubbers for removing air
pollutants.
The universal nature of concern for the effects of polluted air on cultural
property is reflected in a Japanese study of ambient and indoor SOx and
NOX concentrations for buildings where important screen and panel paintings
are housed (Kadokura and Emoto 1974). Six sites in Kyoto were investigated.
Average concentrations for SOX and NOX were found to be about one-third
of those in Tokyo. Seasonal concentrations for SOX peaked in winter and
were highest for a site near a dyeing factory whose liquid wastes emitted
S02- The NOX concentrations were found to be more evenly distributed
throughout the city. Tight buildings showed higher NOX levels indoors than
were found for ambient conditions. Although they did not cite specific ex-
amples of damage, the authors called for protective measures to prevent air
pollution damage to paintings.
7.1.2.5.3 Medieval Stained Glass. Some evidence exists that medieval
stained glass exposed to the atmosphere has deteriorated more rapidly since
World War II than in previous centuries. This accelerated deterioration has
been attributed to the effects of air pollution (Frenzel 1971, Froedel-Kraft
1971, Korn 1971) because gypsum and syngenite (CaSO^KgSO^^O) are found in
the weathering crust. However, such crusts are found even in locations with
low S02 concentrations, suggesting that background S02 levels are suf-
ficient to produce the sulfates observed. An alternative mechanism of decay
suggests that storage of the windows under damp conditions during the war
permitted the formation of a fissured hydrated layer that led to enhanced
corrosion after reinstallation of the windows. The sulfates found in the
weathering crusts are thought to be by-products of the deterioration process
(Newton 1973).
A broad range of preservation techniques has been employed, including la-
mination, coating with inorganic and organic materials, and "isothermal
glazing." In the latter process, the ancient glass is moved just inside the
building and modern glass is placed in the grooves in the stone.
7.1.2.5.4 Conservation and Mitigation Costs. Some indication of the prob-
lem's magnitude is given by cost for mitigative actions taken for cultural
property in West Germany (Table 7-8). Similar cost estimates exist for
national preservation programs in the United Kingdom, Greece, France, Italy,
and the United States. For example, the Italian Parliment designated
$200,000,000 in 1980 for a 5-year program to restore and maintain the ancient
7-35
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TABLE 7-8. ESTIMATED COSTS ASSOCIATED WITH AIR POLLUTION DAMAGE TO
CULTURAL PROPERTY IN WEST GERMANY (AFTER SCHREIBER 1982)
CO
01
Location
Federal
Republic
of Germany
Objects
All municipal bronze
monuments and sculptures
All metal sculptures in
Measures
Desirable
cleaning
Desirable
Period
Annual
Annual
Costs DM
4,000,000
1,000,000
Munster
Cologne
Cologne
Freiburg
Ulm
museums and open air
All medieval stained
glass
Artifacts in museums
Castle facade
Cathedral stained glass
windows
Cathedral facade
Cathedral stained glass
windows
Cathedral stained glass
windows
cleaning,
conservation
Desirable
conservation
Air condition-
ing with air
improvement
Cleaning,
restoration,
conservation
Conservation
Cleaning,
restoration,
conservation
Restoration,
conservation
Desirable
restoration
10 year cost
estimate
During
construction
1965-1973
1978
Annually
1977-1997
1978
Total cost
200,000,000-
300,000,000
15% of construc-
tion costs
1,000,000
448,000
3,000,000-
60,000,000
(estimated)
150,000
3,000,000
-------
monuments in Rome (Hofmann 1981) and it is estimated by a British
Parliamentary Committee that restoration of the fabric of the Houses of
Parliament will cost up to £5,000,000 (International Herald Tribune 1980).
7.1.2.6 Economic Implications—The possibility of determining the economic
costs of air pollution's damaging effects has long attracted environmental
policy makers. If reliable cost estimates could be developed for such
effects in relation to the pollutant levels that produced them, it then might
be possible to compare the costs for achieving various levels of air quality
control through emission control with the cost savings from reduced damage—a
significant step toward developing cost-benefit relationships for air pol-
lution control. The many attempts to estimate costs associated with air
pollution-induced material damage have recently been summarized by Yocom and
Stankunas (1980). Without exception, all of the generalized estimates of
material damage costs related to all types of air pollution existing at the
time of this review are of questionable value. The reasons for this include
the following:
° As was pointed out earlier, it is usually not possible to isolate the
specific portion of damage and therefore the associated costs created
by a given air pollution effect.
° Improper assumptions and inaccurate estimates of the quantities of
materials in place and exposed to pollutants.
° Unrealistic or improper scenarios of use, repair, and replacement of
materials susceptible to air pollution damage, together with improper
or inaccurate assignment of costs to the scenarios.
0 Incomplete knowledge of substitution scenarios where more expensive
material systems may replace more susceptible materials.
0 Inadequate knowledge of the exposure conditions of susceptible mater-
ials, for example, coexistence of pollutants with other environmental
effects such as moisture and temperature, and the physical aspects of
exposure such as orientation and degree of sheltering.
A recent study by Stankunas et al. (1981) has addressed many of the above
difficulties. In this study the quantities of potentially susceptible ma-
terials were determined within 357 randomly selected 100 x 100 foot square
areas covering the Boston metropolitan area. Teams of observers using survey
techniques determined the areas of various types of exposed painted surface,
bare metal of several types, brick, stone, concrete, and several other types
of surfaces. Of the 357 areas selected, 183 were found to contain manmade
structures. The total areas of each material found at the survey sites were
extrapolated to the entire Boston metropolitan area. Then, using air quality
records for SOg in the Boston area, together with humidity data and pub-
lished air pollution damage functions for given materials, the researchers
computed the total damage to a given material for the entire area. In the
case of painted surfaces, assumptions were made on the average thickness of
7-37
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typical paint films. Costs were assigned to the increase in painting fre-
quency, based on the SC^-related increase in paint erosion, to arrive at a
total S02-related damage cost to paint in the Boston metropolitan area.
The excess painting costs for the Boston metropolitan area attributable to
S02 damage for the year 1978 were estimated to be $31.3 million. This is
equivalent to a per capita cost between $11 and $12. Costs for damage to
zinc-coated materials were two orders of magnitude lower.
Haynie (1982) estimated costs for damage to zinc-coated transmission towers
and to galvanized roofing, siding, and guttering. Different approaches were
used for transmission towers than for the other materials. Costs for trans-
mission tower damage were based on a single group of towers serving the
Colbert Steam Plant in the TVA system, Measurements were made by TVA of the
thickness of the zinc coating at several points on 19 towers likely to be
affected by S0£ from the plant in question. Using S02/moisture damage
functions for zinc corrosion and an estimate of how height above ground would
affect S02 deposition velocity (based primarily on changes in wind speed
with height), estimates were made of change in zinc thickness with time for
the group of towers. Then, using several scenarios of painting, repair, and
replacement, researchers estimated annual costs for mitigating the effects of
the damage, based on local S02 and humidity levels. Since TVA owned the
towers, such costs could be internalized and were estimated at 0.0028 mills/
Kwh + 0.0011 to be added to customers' electric bills. These estimates were
based" on an S02 concentration of 17 yg nr3. If S02 levels were al-
lowed to reach the ambient air quality standard of 80 yg m"3, the annual
extra maintenance cost would rise to an estimated 0.0132 mills/Kwh _+ 0.0052.
Cost estimates for damage to galvanized roofing, siding, and guttering re-
quired estimating the relative quantity of these materials in place. One of
the complicating factors in making this determination was the trend in recent
years of replacing bare galvanized materials exposed to the outdoor
atmosphere with coil-coated galvanized steel or bare aluminum. Various
models were used to convert data on shipments of the materials in question
and anticipated use of alternate materials to a realistic picture of the
amount of bare galvanized materials in these categories in 1980. Damage
functions for the effects of $02 and moisture on zinc, together with
estimates for the thickness of zinc coatings and various maintenance sce-
narios and their costs were used to estimate per capita costs. These costs
were computed to be in the range of $0.60 to $1.50 with the best estimate
being $1.05 at an annual average S02 concentration of 30 yg m~3. At
the primary standard of 80 yg nr3, the best estimate of per capita costs
would be $1.80.
Such approaches as these should be refined and extended so that realistic
estimates may be made of the total costs of damage from acidic deposition.
7.1.2.7 Mitigative Measures—Assuming that some degree of damage to ma-
terials results from acidic deposition, a wide range of mitigative actions
may be taken in response to damage. Table 7-1 listed several of these in
relation to various material categories. The particular mitigative measure
and whether it will be implemented will depend on many factors, including:
7-38
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° Physical and chemical nature of the material,
o Age and state of repair of the materials system,
o Availability and cost of substitute materials,
° Feasibility of isolating the object or surface of concern from the am-
bient environment,
° The importance of aesthetics in the appearance of the materials,
o The impact of damage on structural integrity, and
° The attitudes of those responsible for the objects made of the mater-
ials in question regarding the relative importance of the damage.
As stated earlier, material damage from acidic deposition is generally in-
distinguishable from damage caused by the natural environment. However,
chemical analysis of corrosion or damage products can often distinguish
various damage mechanisms. In general, superimposing acidic deposition on
these natural phenomena only tends to shorten the time before some mitigative
measure must be considered. It does not change the mitigative actions them-
selves. Thus mitigative measures taken to protect, replace, repair, and
maintain materials exposed to the ambient environment will generally not
change whether any acidic deposition has an effect. Only the frequency of
implementing these measures will change.
7.2 POTENTIAL SECONDARY EFFECTS OF ACIDIC DEPOSITION ON POTABLE WATER PIPING
SYSTEMS (G. J. Kirmeyer)
7.2.1 Introduction
The potential effects of acidic deposition on materials in potable water
piping systems represents a special concern because of the potential for
indirect effects on human health. Chapter E-6 has discussed the effects on
health from contaminants in water supplies, contaminants that may occur in
greater concentrations under acidic conditions. This section, dealing with
potable water piping systems, discusses the potential effects of acidic
deposition on piping materials. The effects of acidification may lead to
increased concentration of metals in the water and may increase the cost to
maintain piping systems in serviceable condition.
7.2.2 Problems Caused by Corrosion
The problems caused by corrosion can be grouped into three categories:
health, aesthetic, and economic.
7.2.2.1 Health--Corrosion of materials in plumbing and distribution systems
increases the concentrations of metal compounds in the water. Lead, cadmium,
and other heavy metals are present in various amounts in pipe material, and
there is concern for the possible health hazards created by corrosion and
7-39
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subsequent leaching and ingestion of these materials (see Chapter E-6,
Section 6.3). The health-related compounds are regulated by the U.S. EPA
through the Safe Drinking Water Act, PL 93-523.
7.2.2.2 Aesthetics—Contaminants, such as copper, iron, and zinc are also
leached from plumbing and distribution systems. These contaminants, when
present in concentrations above the limits suggested in the National
Secondary Drinking Water Regulations, can render the water aesthetically
undesirable for consumption because of taste, color, or staining character-
istics. Corrosion of piping can cause red water, blue stains on fixtures,
stains on laundry, and can impart a metallic taste to the water.
Acidification of water can increase these problems.
7.2.2.3 Economics--Deterioration of plumbing and distribution systems be-
cause of corrosion frequently results in extensive and costly replacement.
Corrosion of copper pipe is usually characterized by a uniform etching or
thinning of the pipe wall. Failure occurs when corrosion has damaged the
structural integrity of the pipe so much that leakage becomes a problem.
Corrosion of galvanized steel is normally characterized by pits that develop
in the pipe surface. These pits may eventually penetrate the pipe wall and
cause leakage. As the pipe deteriorates, tubercles build up over the devel-
oping pits. These tubercles increase the roughness of pipe surfaces as well
as tend to form a blockage of the pipe. Tuberculation of the interior sur-
faces of metal pipes will cause the loss of carrying capacity of the pipe. To
overcome the resistance to flow, higher pressures have to be maintained at
the pumping stations, which in turn requires additional energy. Acidifi-
cation of waters can render them more corrosive; thus, these waters will
require more intensive measures for corrosion control. This in turn will
increase the economic burden of processing the water.
7.2.3 Principles of Corrosion
The word corrosion is derived from the Latin word "rodere", meaning "to
gnaw." Corrosion may be thought of as the gnawing away or attack of a ma-
terial, usually by some chemical or electrochemical means. Internal piping
corrosion occurs in several widely differing forms, which are usually clas-
sified according to the appearance of the corroded metal, and can be either
uniform or localized. Uniform corrosion occurs when the material corrodes or
thins at approximately the same rate over the entire surface. Localized
corrosion occurs when a material surface is attacked unevenly so that some
areas are severely affected while adjacent areas are not. Types of localized
corrosion include galvanic, crevice, pitting, and erosion.
Electrochemical corrosion can be viewed in terms of oxidation and reduction
reactions. For corrosion to occur, all the components of an electrochemical
are required (see Figure 7-5; see also Section 7.1.2.1). At the anode, the
oxidation of a metal occurs as follows:
Metal -> Metaln+ + ne~.
7-40
-------
WATER
02 + 4H+ + 4e~ •> 2H20
acid solutions
02 + 2H20 + 4e~ + 40H-
basic solutions
HOC1 + H+ + 2e- -> Cl- + H20
reduction of chlorine
2H+ + 2e- -> H2
hydrogen evolution
Metal + Metaln+ + ne~
general equation
Fe + 2H20 + Fe(OH)2 + 2H+ + 2e~
low concentration of carbonate
Fe + HC03~ ->- FeC03 + H+ + 2e~
high concentration of carbonate
Figure 7-5. Some anode and cathode reactions for iron pipe contacting water.
-------
7.2.4 Factors Affecting Internal Piping Corrosion
The factors affecting corrosion are many and varied and each water is dif-
ferent. Generally, the factors favoring corrosion include a low pH, low
buffering capacity, and a high concentration of oxidizing substances, such as
dissolved oxygen and free chlorine. Some of the different factors which
control the rate and degree to which this corrosion reaction occurs are pre-
sented in Table 7-9.
A review of Table 7-9 indicates that acidic deposition can potentially affect
corrosivity by several methods. In soft, acidic, poorly buffered waters such
as those waters in the Northeast, Southeast and Pacific Northwest, acidic
deposition could affect the factors that cause piping corrosion. Such waters
are prone to corrosivity in their natural state and acidic deposition, in
sufficient quantity, could further reduce pH and the water's alkalinity or
buffering capacity, thereby aggravating the problem. In addition, the
sulfate ion present in the acidic deposition environment is considered an
aggressive anion, and increased sulfate could increase corrosion rates,
especially a destructive form of corrosion called pitting. In poorly buf-
fered waters, acidic deposition could increase the molar ratio of strong
mineral acids to alkalinity and shift the carbonate balance toward carbon
dioxide; both changes could increase corrosivity.
Based on water quality parameters, several corrosion indices have been de-
veloped. These include the Langelier Index (LI), the Aggressiveness Index
(AI), the Ryznar Index (RI; or Stability Index), the Larson's Ratio, the
Buffer Intensity, the Momentary Excess (ME), the Driving Force Index (DPI),
the Casil Index, the Riddick Index, and the Calcite Saturation Index. The
effects of acidification and sulfate addition on various corrosion indices
are presented in Table 7-10. In nearly every instance, acidic deposition
increases corrosivity as indicated by the movement of the index towards a
more corrosive environment.
One of the most comprehensive studies conducted in the United States, Acid
Precipitation and Drinking Water Quality in the Eastern United States (Taylor
et al. 1984), was recently completed by the New England Water Works
Association under cooperative agreement No. CR 807808010 with the U.S.
Environmental Protection Agency. It evaluated the quality of drinking water
in the New England, Appalachian, and coastal States and the potential effects
of acid precipitation on water supplies.
Hundreds of raw water supplies were sampled in that study (Taylor et al.
1984). Data indicated that pH was seldom below 5.0 for raw waters, but
almost half of the raw surface water alkalinities were below 5 mg £~1 as
CaC03, and well over half of the raw surface waters had calcium concen-
trations below 5 mg £~1. One fifth or more of the finished waters from
both ground and surface sources fell outside the pH range of the Federal
Secondary Drinking Water Standards.
Taylor et al. summarized the potential raw water corrosiveness using indices
presented in Table 7-11. The water quality and the high frequency of indices
in the "Highly aggressive" range indicate the corrosive nature of many New
7-42
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TABLE 7-9. FACTORS AFFECTING CORROSIYITY OF DRINKING WATER
Factor
pH
Dissolved oxygen
Low buffering capacity
Free chlorine residual
High halogen and sulfate:
alkalinity ratio
Carbon dioxide
Total dissolved solids
Calcium
Silica
Tannins
Flow rate
Effect on Corrosivity
Low pH's generally accelerate corrosion. Acidifi-
cation would lower pH's and tend to increase cor-
rosivity.
Dissolved oxygen in water induces active corro-
sion, particularly of ferrous materials.
Low alkalinity waters have little capacity to
resist change in pH. Acidification lowers alka-
linity and buffer capacity, generally increasing
corrosivity.
The presence of free chlorine residual promotes
corrosion of ferrous metals and copper.
A molar ratio of strong mineral acids much above
0.5 results in conditions favorable to pitting
corrosion. Acidification and addition of sulfates
from acidic deposition would increase the molar
ratio, tending to increase corrosion - especially
in soft, poorly buffered waters.
Carbon dioxide is particularly corrosive to copper
piping. Acidification reduces pH and increases
carbon dioxide.
Higher concentrations of dissolved salts increase
conductivity and may increase corrosiveness. Sul-
fates would increse the salt content.
Calcium generally reduces corrosion by forming
protective films with dissolved carbonates.
Silica forms protective films over metal surfaces.
Tannins form protective organic films over metals.
Turbulence at high flow rates allows oxygen or
carbon dioxide to reach the surface more easily,
removes protective films, and causes higher cor-
rosion rates.
7-43
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TABLE 7-9 CONTINUED
Metal Ions Certain ions, such as copper, can aggravate corro-
sion of downstream materials. For example, copper
ions can increase corrosion of galvanized pipe.
Temperature High temperature increases corrosion reaction
rates. High temperature also lowers the solu-
bility of calcium carbonate and calcium sulfate
and thus may cause scale formation in hot water
heaters and pipes.
7-44
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TABLE 7-10. EFFECT OF ACIDIC DEPOSITION ON VARIOUS CORROSION INDICES
Index
Casll Index (CI)
Larson's Ratio
(LR)
Formula
CI = Ca + Mg + HS103
- Anjons
LR = (C1-) + (S042-)
(Alk)
Water Quality
Parameter
Calcium, magnesium,
silica, anions
such as CI", F~,
$04 (Expressed 1n
mllllequlvalents
per liter)
Chloride, sulfate,
and alkalinity
expressed molar
Theoretical Effect
of Acidic Deposition
on Index
Adds corrosive
anions, lowers
Casll Index
Increase SO^-
and decrease
alkalinity,
Increase LR
Probable
Effect on
Corroslvity
Increase
Increase
I
-1^
en
R1dd1ck Corrosion
Index
Corrosion Index =
75 [COz + 1/2 (Hardness - Alk)
TTk
Sat DO
Alkalinity (mg je"1
CaCOa). CO?
(mg £-1), hardness
(mg d-1 CaCOa),
Cl~ (mg s."1 as
CI"), nitrate 1on
(mg t-1 as N),
dissolved oxygen (DO In
mg *~1), and oxygen
saturation (sat DO 1n
mg ft'1).
Reduce alkalinity
and increase
CO?, Increase
Rlddlck corrosion
index
Increase
Calclte Saturation
Index (CSI)
CSI = log K - log (Ca2+) - log
HC03 - pH
log K = 2.582 - 0.024
T°C; Ca2+ and ,
HC03 1n mol £-1;
HC03 = total alk.
as CaC03 for pH <
9.3 and HCO? is
less than H*.
Reduce HC03
and pH, Increase
CSI
Increase
-------
TABLE 7-10. CONTINUED
Index
Langeller
Saturation Index
(LSI)
Aggressiveness
Index (AI)
— i
-pa
°^ Ryznar Index
(or Stability
Index: SI)
Buffer Intensity
(BI)
Driving Force
Inrlpx fnFTl
Formula
LSI = pHa - pHs
where
pHs = -log - log HC03~
- log [K2'/KS']
AI = pH + log (Ca) (Alk)
SI = 2pHs - pHa
BI = Shape of Alkalinity
Tltratlon Curve at actual
pH of the water
DFI = Ca+(mg JT1) -C03'2(mg "M
Water Quality
Parameter
pH, Alkalinity,
Calcium, Temperature
Ionic strength (I)
pH, Ca, alkalinity
PH
pH, alkalinity in
(meq)
Calcium, carbonate
Theoretical Effect
of Acidic Deposition
on Index
Reduce pH and
alkalinity, thus
lowers LSI
Reduce pH and
alkalinity, thus
lowers AI
Reduce pH, thus
Increase SI
Reduce pH and
alkalinity. Depends
on Initial pH and
alkalinity of water.
Reduce C032~,
lowpr<; DPI
Probabl e
Effect on
Corroslvity
Increase
Increase
Increase
Increase or
decrease
Increase
-------
TABLE 7-11. PERCENT OF RAW WATER SUPPLIES INVESTIGATED INDICATING CORROSIVENESS.
ADAPTED FROM TAYLOR ET AL. (1984).
Index
Calcite Saturation
Index
Langelier Index
Aggressiveness
Value Category
> 3 Susceptible or
highly susceptible
to change
< -2 Highly aggressive
< 10.0 Highly aggressive
Round 1
Ground &
Surface Water
Percent
63
85
85
Round 2
Ground &
Surface Water
Percent
79
97
91
Round 2
Groundwater
Percent
72
91
88
Index
Ryznar Stability
Index
> 8 Highly aggressive
97
96
96
-------
England water supplies. These soft, low pH, poorly buffered waters are the
ones that would be most affected by a sufficient quantity of acidic depo-
sition.
7.2.5 Corrosion of Materials Used in Plumbing and Vlater Distribution Systems
Each type of material used in plumbing or water distribution systems will
react differently to the various water qualities. Engineering professionals
have generally selected distribution system piping based on its structural
strength and resistance to external and internal corrosion. Many pipes are
lined with cement or coal tar to separate the metal from the water, thus
affording protection against corrosive waters. Some piping has not been
lined and is more susceptible to internal corrosion. Piping for home and
building plumbing systems is generally metal having a small diameter and is
unlined. This piping was installed to meet building codes, normally with
little thought given to its ability to resist internal corrosion. In a
corrosive environment, home and building plumbing systems are subject to
deterioration. Acidic deposition in sufficient quantity could affect the
parameters that cause increased corrosion in piping systems.
7.2.5.1 Corrosion of Iron Pipe--Corrosion of iron pipe is characterized by
pitting and the formation of iron oxide tubercles. In general, life spans
for unlined iron pipe in water with low pH and alkalinity are quite short.
Failure can usually be attributed to plugging and pinhole leaks. The pitting
and tuberculation process is initiated when, for any reason, the rate of iron
dissolution is momentarily increased at particular points on the pipe sur-
face. The initiation process usually occurs in a surface scratch, near
surface irreguarities, in standing water, and near iron oxide deposits. Once
initiated, pitting occurs by an autocatalytic process.
Within the pit, rapid dissolution of iron occurs and oxygen is depleted.
Because iron dissolution continues, an excess positive charge develops. This
positive charge is balanced by the migration of chloride and other ions into
the pit to maintain electroneutrality. Thus, the pit contains a high
concentration of Fed2, and as a result of hydroysis,
FeCl2 + 2H20 •* Fe(OH)2 (s) + 2H+ + 2C1-,
a high concentration of hydrogen ions exists. Both hydrogen and chloride
stimulate the dissolution of iron, and the entire process accelerates with
time.
At the interface between the pit and the adjacent surface, iron hydroxide
tubercles form because of the interaction between the hydroxide produced by
the cathodic reaction and the dissolved iron:
Fe2+ + 20H- + Fe (OH)2(s).
Dissolved oxygen can further oxidize the iron (II) hydroxide to other oxides.
The different colored layers in a tubercle are evidence of this oxidation by
dissolved oxygen. Favorable conditions can result in formation of FeC03
and subsequent iron compounds, which form a protective layer on iron pipe and
7-48
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inhibit corrosion. A discussion of the mechanism of corrosion inhibition by
formation of FeC03 has been presented previously (Sontheimer et al. 1981).
Acidic deposition in sufficient quantity could influence corrosion of iron
pipe by reducing the capacity of the water to neutralize local areas of low
pH. Sulfate ions in sufficient quantity could increase the process of cor-
rosion by increasing the aggressive anion concentration.
7.2.5.2 Corrosion of Galvanized Pipe—Deterioration of galvanized pipe
occurs in two stages.Initially, only the galvanized or zinc layer corrodes
until iron is exposed. Corrosion of the galvanized layer depends on pH,
carbonate concentration, and flow. Once the galvanized layer is penetrated
and iron is exposed, the galvanized pipe begins to perform as a iron pipe.
Initially, the zinc sacrificially corrodes, offering protection to the iron.
Eventually, the iron base metal of the pipe begins to pit and iron oxide
tubercles are formed.
7.2.5.3 Corrosion of Copper Pipe—The corrosion of copper pipe is generally
uniform. In the presence of dissolved oxygen, a thin film of cuprous oxide
is formed over most of the metal's surface. This film promotes a constant
corrosion rate that is normally only a fraction of the corrosion rate of iron
or galvanized pipe. However, in a softwater environment, thinning of copper
pipe can proceed quickly. Copper corrosion is highly dependent on pH and low
pH waters can cause rapid deterioration of piping—resulting in leaks. Thin-
walled copper pipe allowed by building codes will require replacement sooner
because initial wall thickness is less.
Pitting of copper pipe can also occur. Pitting is usually caused by a break-
down of the passivation film. The film can be disturbed by high-velocity
water flow or dissolved by either carbonic or organic acids that are found in
some freshwaters. Chlorides also tend to promote pitting by increasing the
porosity of the passivation film. Chlorine increases the oxidation of copper
and prevents the establishment or continuation of the protective film of cu-
prous oxide. Acidic deposition in sufficient quantity could affect corrosion
of copper pipe by lowering pH and increasing the carbon dioxide content of
the water. The lower pH values could cause more rapid thinning of this pipe
and the increased carbon dioxide and carbonic acids could aggravate pitting.
Near a soldered joint in copper piping, a galvanic couple is formed between
the copper pipe and the solder. The copper acts as the cathode and the
solder acts as the sacrificial anode. In the case of 50-50 lead-tin solder,
the principal anodic reaction is the dissolution of lead and the subsequent
leaching of lead into the drinking water. The problem with leaching of lead
is eliminated with the use of 95-5 tin-antimony solder, which corrodes to
form a passivation film that inhibits metal leaching from the solder.
7.2.5.4 Corrosion of Lead Pipe—Lead pipe has been in service for many years
in several older water distribution systems. Although lead pipe is durable
for use with potable water, the toxicity of trace amounts of dissolved lead
should preclude its use to distribute any potable water, especially those of
soft, acidic nature.
7-49
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The corrosion of lead pipe depends very much on pH and alkalinity. At very
low alkalinities, lead is soluble throughout the pH range of drinking water.
In water containing carbonate alkalinity, an insoluble film of basic lead
carbonate forms in the intermediate pH region. For example, at a total al-
kalinity of 20 mg £~1 and a pH of 9.5, the concentration of lead in water
circulating through lead pipe was less than 50 vg £-1 in an U.S. EPA
experiment (Schock and Gardels 1983). The film of basic lead carbonate per-
forms two functions: (1) by adhering to the metal surface, the film forms a
physical barrier between the metal and the water, and (2) the basic carbonate
or carbonate solid phase limits the lead solubility and, therefore, reduces
the amount of lead that can be leached into the drinking water. However,
even in systems containing high pH values, corrosion can occur at very low
and high alkalinities. Corrosion of lead pipe and subsequent leaching of
lead into the water has been a concern of water utilities and health offi-
cials for many years. Acidification of low pH, poorly buffered waters could
increase the potential for leaching of lead by further reducing pH and alka-
linity.
7.2.5.5 Corrosion of Non-Metalic Piping—Very little work has been done on
the effect of water quality on non-metallic water piping. For the purposes
of ths discussion, non-metallic pipe is divided into two categories: (1)
cement pipe (A/C pipe, mortar-lined pipe, etc.), and (2) plastic pipe (poly-
vi nyl chl oride, chlorinated polyvinylchloride, polyethylene, polybutylene,
etc.). Obvious properties would indicate that most plastic pipes are not
affected by the water quality variations normally experienced from one
utility to the next.
Cement pipes, on the other hand, do show deterioration under certain con-
ditions, particularly acidic water of low mineral content. Softwaters attack
concrete pipe by removing calcium oxide (CaO) from the cement matrix.
However, the mechanism is only poorly understood.
Several of the indices presented in Table 7-10 rely on the solubility of cal-
cium carbonate to explain corrosion potential. Low pHs in poorly buffered
water would tend to dissolve calcium carbonates and the cement matrix, thus
causing pipe deterioration. Acidic deposition in sufficient quantity could
lower pH and alkalinity and increase deterioration of cement piping.
7.2.6 Metal leaching
In addition to pipe deterioration, corrosion causes changes in the quality of
the water distributed to the customer. To detect these changes and determine
the extent of deterioration, a water sampling program is used. Changes in
water quality can occur both in the purveyor distribution system and in the
customer home plumbing system.
7.2.6.1 Standing vs Running Samples—Time is an important parameter in
leaching of metals because more metals will be leached as water stands idle
overnight in a plumbing system than will be in water that is flowing quickly
through the piping during heavy demands. Thus, the metal leaching survey
must account for different residence times in the water piping system. The
U.S. EPA has described a procedure for collection of samples to represent
7-50
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different contact times and locations in the plumbing system. The first
sample is collected immediately upon opening the faucet to provide water that
has been standing overnight in the home plumbing. The second sample is de-
signed to represent the service line and is collected as soon as the water
temperature changes from warm to cool. The cool water has been in the ground
outside the foundation. The third sample is designed to represent water from
the utility distribution system, and is collected from the tap after 3 to 5
minutes of flushing, depending on the length of service piping. This third
sample would have had minimum residence time in the plumbing and represents
ambient distribution water quality.
Water quality changes can also occur in the utility distribution and trans-
mi sion system. These changes can be detected by sampling at the water source
and then at various points in the distribution system. The selection of
sampling sites should account for system variables such as different sources
of supply, different pressure zones, chlorination stations in the system,
differences in the transmission/distribution pipe material, and type and age
of plumbing system.
7.2.6.2 Metals Surveys—When planning a leaching survey, researchers should
select water quality parameters based on the types of materials in the dis-
tribution or home plumbing system. For example, in a mortar-lined ductile
iron pipe, parameters of interest are iron, calcium, pH, alkalinity, con-
ductivity, color, and dissolved oxygen. Table 7-12 contains a listing of
pipe materials and water quality parameters of interest. This list is not
inclusive of all components needing analysis under potential corrosion
conditions, but indicates the rationale to be followed in selecting analysis
parameters. Most older plumbing systems are a mix of plumbing types, in-
cluding galvanized steel, black steel, copper, and possibly lead. In this
case, the minimum set of parameters should include lead, cadmium, zinc, iron,
and copper.
Some very extensive tap metal surveys have been conducted on water supplies
in the eastern United States in areas that receive acidic deposition. Taylor
et al. (1984) collected early morning samples from 43 locations in Maine,
Connecticut, Massachusetts, New Hampshire, New York, Rhode Island, Vermont,
Pennsylvania, New Jersey, and North Carolina. Taylor concluded that water
standing overnight in the plumbing system had the highest concentration of
metals about two thirds of the time, when compared to samples from the ser-
vice line and free flowing water from the main. Fourteen percent of the
samples from plumbing systems exceeded the primary MCL (maximum contaminant
level) of 50 yg &-i for lead and only 2 percent of the samples from
service lines exceeded that level. No distribution system samples exceeded
the lead MCL. The same survey also indicated that household water exceeded
the secondary MCL of 1 mg £-1 for copper in 42 percent of the samples and
21 percent of the time in service line samples.
An extensive tap survey in Boston, MA was conducted on the Quabbin Reservoir
in 1976-77 before pH adjustment was initiated. Of the total 443 tap samples
taken the average lead content of 0.079 mg £-1 and 195 samples exceeded
the 0.05 mg £-1 primary MCL (Taylor and Symonds 1984).
7-51
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TABLE 7-12. WATER QUALITY PARAMETERS OF INTEREST FOR DIFFERENT PIPING MATERIALS'
Pipe Material
Primary Parameters
Secondary Parameters
ro
Steel, ductile and cast iron (unlined)
Ductile and cast iron (mortar lined)
Copper
Lead
Galvanized steel
Copper alloys
Asbestos-cementb
Concrete cylinder
Iron
Iron, calcium
Copper, lead from
solder joints
Lead
Zinc, cadmium, lead,
iron
Copper, zinc, lead
Calcium, asbestos
fibers
Calcium
pH, conductivity, color, DO, manganese
pH, alkalinity, conductivity, color, DO
pH, alkalinity
pH, alkalinity
pH, conductivity, color, DO
pH, alkalinity
pH, alkalinity
pH, alkalinity, conductivity
Extracted from American Water Works Association (1983).
bSamples should be collected directly from distribution/transmission mains,
-------
7.2.7 Corrosion Control Strategies
A corrosion control strategy should include an evaluation of existing water
qualty data and collection of additional information if needed. Samples of
pipe may need to be collected and evaluated as part of the program. Corro-
sion control treatment plans should include blending of water sources to
reduce the waters' corrosiveness, and bench and pilot studies to evaluate the
effectiveness of corrosion inhibitors. The treatment plan should include
definitive water quality goals. A material selection program for corrosion
resistant materials should also be considered. In most water sources, pro-
tective coatings, such as cement mortar lining, provide good protection.
Plastic piping for home and service piping should be considered. Building
and plumbing code changes can be enacted to preclude materials that rapidly
deteriorate and to encourage use of corrosion resistant materials.
To ensure that the program goals are being met, corrosion monitoring is
needed. Monitoring can include water qualty analysis, pipe tap and coupon
evaluation, customer surveys and complaint records, and continuing cost
effectiveness evaluations.
7.2.8 Economics
The significance of corrosion costs to the overall economy of the United
States has been reported by the National Bureau of Standards (NBS) (Bennett
1979). Corrosion costs in the United States were estimated at $70 billion
annually in 1975. The direct-cost portion of this amount is approximately 25
percent, which in proportion to the gross national product is 4.2 percent.
The proportion of costs that can be avoided by corrosion control measures is
approximatly 15 percent of the direct cost portion. The same NBS report
indicated that the annual costs in the water supply field total approximately
$700 million and that 20 percent of the water supply corrosion costs were
thought to be avoidable by control measures. These costs are only for dis-
tribution systems. Often, however, a far greater portion of corrosion costs
are incurred through damage to interior piping and plumbing systems. Re-
placing a water heater could cost the homeowner $200 to $300. The cost of
replacing accessible plumbing in a home would be several hundred dollars. If
all of the plumbing in a home has to be replaced, costs could easily reach
$2000 to $3000.
Corrosion damage can be quantified in monetary terms, and benefits of cor-
rosion reductions can also be calculated. A detailed study in the Seattle
area conservatively estimated costs of corrosion at an annual cost of $7
million primarily for residential premise piping sytems and $410,000 for
transmission and distribution systems (Kennedy Engineers 1978). The approach
to such calculations is based on knowing the service life of pipe under
existing conditions and then projecting service life as the water quality
changes. By performing a present-worth analysis and comparing monetary
benefits of longer service life versus the costs of water treatment, one can
calculate a cost-benefit ratio. A brief hypothetical example of calculating
a cost-benefit ratio follows. Although the example is hypothetical, it is
based on pipe service life data and replacement costs gathered in the Seattle
study (Kennedy Engineers 1978, Ryder 1980).
7-53
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Assuming that under existing water quality conditions, galvanized pipe in a
single-family dwelling has a 30-year service life and the house is 20 years
old, then the pipe has 10 years of useful life until it becomes so plugged
with rust that neither flow nor pressure can be maintained at adequate
levels. Interim repairs, leak damage, and total replacement are assumed to
occur as shown in Table 7-13. If a treatment method is instituted to reduce
corrosion by 40 percent, the repairs, pipe damage, leaks, and replacement
will be delayed to some time in the future. Tables 7-14 and 7-15 list
present-worth values of the corrosion damage without and with treatment,
respectively. The cost aspect of corrosion is a powerful tool in presenting
persuasive arguments for reducing corrosion damage.
7.3 CONCLUSIONS
From the review of the available literature on the effects of acidic deposi-
tion (as defined in this chapter) on materials the following conclusions are
drawn:
° Several scenarios and mechanisms exist for damage to materials from
acidic deposition as a result of both long-range transport and local
source emissions (Section 7.1.1).
o Without question acidic deposition causes significant incremental dam-
age to materials beyond that caused by natural environmental phenomena
(Section 7.1.1).
° Because very few research efforts have attempted to isolate the
effects of specific acidic deposition scenarios, it is presently
impossible to determine quantitatively if any one scenario is more
important than another in causing material damage. However, based on
the juxtaposition of primary acidic pollutant (e.g., S0£) sources
and large quantities of susceptible material surfaces in urban areas,
damage to materials from primary pollutants directly or in oxidized
form together with surface moisture (e.g., dew) is believed to be due
more to acidic deposition from local sources than to acidified rain
produced from long-range transport of pollutants and their reaction
products (Section 7.1.1)
o Reliable cost estimates for material damage from acidic deposition are
at present fragmentary because they deal with only selected material
systems or limited geographical areas. Available estimates of total
material damage costs on a nationwide basis are unreliable. There is
a need for improved inventories of materials in place in various parts
of the country (Sections 7.1.1 and 7.1.2).
o Damage to cultural property from acidic deposition is a complex
problem because of the high value placed upon such objects, their
often irreplaceable nature, and the wide range of material types
represented. Highest priority should be placed on identifying and
quantifying actual and potential damage to such artifacts and develop-
ing methods to prevent damage (Section 7.1.2.5).
7-54
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TABLE 7-13. SCHEDULE OF REPAIRS
Cost
Repairs Year dollars
Replace service line 1984 300
Interim repair 1989 75
Leak repair 1991 100
Leak damage 1991 150
Replace accessible plumbing 1994 1000
TABLE 7-14. PRESENT WORTH VALUES WITHOUT CORROSION TREATMENT
Item
Replace service line
Interim repair
Leak repair
Leak damage
Replacement
Total
Year
1984
1989
1991
1991
1994
Cost
dollars
300
75
100
150
1000
PWFa
1.00
0.62
0.51
0.51
0.38
Present Worth
dollars
300
46
51
78
385
860
aPresent worth factor for 10 percent (rounded off)
TABLE 7-15. PRESENT WORTH VALUES WITH CORROSION TREATMENT
Item
Replace service line
Interim repair
Leak repair
Damage repair
Replacement
Total
Year
1984
1991
1996
1996
2001
Cost
dollars
300
75
100
150
1000
PWFa
1.00
0.45
0.33
0.33
0.20
Present Worth
dollars
300
34
33
49
200
616
aPresent worth factor for 10 percent (round off)
7-55
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0 Further research directed at isolating damage caused by specific
acidic deposition processes and identifying those processes that are
most important and/or amenable to control is needed (Sections 7.1.1
and 7.1.2).
° Studies that accurately assess damage costs associated with acidic
deposition are needed (Section 7.1.2.6).
0 Further research is needed in developing mitigative measures such as
reliable surface protection systems when damage has already been
observed and when protection cannot wait for improvement in air
quality (Section 7.1.2.7).
From the review of potential secondary effects of acidic deposition on pot-
able water piping systems the following conclusions are drawn:
o Three categories of problems are caused by water piping corrosion:
health, economic, and aesthetic. Health concerns are primarily
associated with leaching of lead into the potable water by corrosion
of lead pipe and solder containing lead. Economic concerns are as-
sociated with pipe blockage, leaks, and pipe deterioration causing
premature replacement. Corrosion-related aesthetic deterioration
causes colored water (i.e., red water), unappealing taste, and stain-
ing of fixtures and clothes (Section 7.2.2).
o Several factors that influence water piping corrosion include pH,
temperature, dissolved oxygen, alkalinity and buffer intensity,
aggressive anions, chlorine residual, total dissolved solids, natural
protective scales, velocity, metal ions, and external electric cir-
cuit. Acidic deposition in sufficient quantity could increase a
water's corrosivity if it caused decreases in pH and akalinity, and
increases in the SO^- level (Section 7.2.4).
° Several corrosion indices were evaluated as to the theoretical effect
of acidic deposition on each index. The evaluation showed that water
affected by a sufficient quantity of acidic deposition would tend
towards increased corrosivity, with respect to every index except one
(Section 7.2.4).
o Soft, low pH, poorly buffered waters prevalent in the Northeast,
Southeast, and Pacific Northwest are more prone to corrosivity than
more highly buffered, mineralized waters. Acidic deposition in
sufficient quantity in these types of waters, would tend to aggravate
a corrosion problem that is already present (Sections 7.2.4 and
7.2.5).
o Metal leaching surveys in several locations including the Northeast
and the Pacific Northwest have demonstrated that corrosive water can
leach metals, including lead, from plumbing systems in quantities that
exceed the primary MCL's of 50 yg &-1. This occurs primarily in
7-56
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water standing in the plumbing system and service line and is caused
by corrosion of lead and galvanized piping, and lead solder used to
join copper piping (Section 7.2.6).
Corrosion control strategies must address two elements: corrosive
water and susceptible piping materials. Treatment of the water with
corrosion inhibitors should be considered along with use of corrosion
resistant materials. Water quality and corrosion monitoring should
continue to ensure that the corrosion control plan is meeting its
goals in a cost effective manner (Sections 7.2.7 and~7.2.8).
7-57
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7-65
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TECHNICAL REPORT DATA
(Please read Instructions on the reverse before completing)
1. REPORT NO.
600/8-83-016BF
2.
3. RECIPIENT'S ACCESSION NO.
4. TITLE AND SUBTITLE
The Acidic Deposition Phenomenon and Its Effects:
Critical Assessment Review Papers
Volume II - Effects Sciences
5. REPORT DATE
6. PERFORMING ORGANIZATION CODE
7. AUTHOR(S)
Editors - Rick A. Linthurst and Aubrey P. Altshuller
8. PERFORMING ORGANIZATION REPORT NO.
I. PERFORMING ORGANIZATION NAME AND ADDRESS
NCSU Acid Precipitation Program
North Carolina State University
1509 Varsity Drive
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15. SUPPLEMENTARY NOTES
This project is part of a cooperative agreement between EPA and North Carolina
State University
ID. AOoTHACT
This document is a review and assessment of the current scientific
knowledge of the acidic deposition phenonemon and its effects. The areas
discussed include both atmospheric (Volume I) and effects (Volume II) sciences.
Specific topics covered are: natural and anthropogenic emissions sources;
transport and transformation processes; atmospheric concentrations and
distributions of chemical substances; precipitation scavenging and dry deposition
processes; deposition monitoring and modeling; and effects of deposition on
soils, vegetation, aquatic chemistry, aquatic biology, materials and human
health, indirectly through ingested food or water. Each of the above topics is
reviewed in detail using the available literature, with emphasis on U.S. data,
and where possible, conclusions are drawn based on the available data.
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