United States Off ice of EPA-600/8-83-016B
Environmental Protection Research and Development May 1983
Agency Washington, DC 20460
Research and Development
&EPA The Acidic Deposition
Phenomenon and
Its Effects
Critical Assessment
Review Papers
Volume II Effects Sciences
Public Review Draft
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS:
CRITICAL ASSESSMENT REVIEW PAPERS
Aubrey P. Altshuller, Editor
Atmospheric Sciences
Co-editors
John S. Nader
Lawrence" E. Niemeyer
Rick A. Linthurst, Editor
Effects Sciences
Co-editors
William W. McFee
Dale W. Johnson
James N. Galloway
John J. Magnuson
Joan P. Baker
Project Staff
Rick A. Linthurst-Director
Betsy A. Hood-Coordinator
Gary B. Blank-Manuscript Editor
Clara B. Edwards-Production Staff
C. Willis Williams-Crapses
Mike Conley-Graphics
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 U.S. EPA/NCSU Acid
Precipitation Program, a cooperative agreement between the U.S.
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 Program and was funded by
U.S. EPA.'
NOTICE
This document is a public review draft. It has not been formally
released by EPA and should not at this stage be construed to represent
Agency policy. It is being circulated for comment on its technical
accuracy.
U.S. Environr;
Region V, LiL.
230 South D,--.
Chicago, Iliinc-i:
,rscy
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Authors
chapter
Lawrence E.
Chapter A.2 - Natural and Anthropogenic M..1... Sources
aP E1TOr Robinson - Washington State U.
Jim B. Homolya - TRW
Chapter A-3 - Transport Processes
Moor V. Gillani - Washington U.
jac* D. Shannon - Argonne National Lab
David. E. Patterson - Kashington U.
Chapter A-4 - Transformation Processes .
David F. Mfller - U. of Nevada
Dean />. Hegg - U. of Washington
Peter V. Hobbs - U. of Washington,
Noor V. Gillani - Washington U.
Michael R. Whitbeck - U. of Nevada
Chapter A-5 - Atmospheric Concentrations and Distributions of Chemical
Substances
A. Paul Altshuller - Consultant
Chapter A-6 - Precipitation Scavenging Processes
Jeremy M. Hales - Battelle, Pacific Northwest Lab
Chapter A-7 - Dry Deposition Processes
Bruce B. Hicks - National Oceanographic and Atmospheric
Administration
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Chapter A-8 - Deposition Monitoring
A. . u
Chapter A-, - Long.Range Transpon
Chandrakant M. Bhumralkar - National rv
E. ™ - », ""
Chapter E-l - Introduction
Rick A. Linthurst - North Carolina State U.
Chapter E-2 - Effects on Soil Systems
William W. McFee - Purdue U.
Fred Adams - Auburn U.
Christopher S. Cronan - U. of Maine
Mary K. Firestone - U. of California, Berkeley
Charles D. Foy - U.S. Department of Agriculture
Robert D. Harter - U. of New Hampshire
Dale W. Johnson - Oak Ridge National Lab
Chapter E-3 - Effects on Vegetation
Dale W. Johnson - Oak Ridge National Lab
Boris I. Chevone - Virginia Polytechnic Institute
Patricia M. Irving - Argonne National Lab
Samuel B. McLaughlin - Oak Ridge National Lab
Dudley J. Raynal - Syracuse U.
David S. Shriner - Oak Ridge National Lab
Lorene L. Sigal - Oak Ridge National Lab
John M. Skelly - Pennsylvania State U.
William H. Smith - Yale U.
Jerome B. Weber - North Carolina State U.
ii
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Chapter E-4 - Effects on Aquatic Chemistry
James N.Galloway - U. of Virginia
Dennis S. Anderson - U. of Maine
M. Robbins Church - U.S. EPA
Christopher S. Cronan - U. of Maine
Ronald B. Davis - U. of Maine
Peter J. Dillon - Ontario Ministry of Environment
Charles T. Driscoll - Syracuse U.
Steve A. Norton - U. of Maine
Gary C. Schafran - Syracuse U.
Chapter E-5 - Effects on Aquatic Biology
John J. Magnuson - U. of Wisconsin
Joan P. Baker - North Carolina State U.
Peter G. Daye - Daye Atlantic Salmon Corp.
Charles T. Driscoll - Syracuse U.
Kathleen Fischer - Environment Canada
Charles A. Guthrie - N.Y. State Dept. of Environ. Conservation
John H. Peverly - NY State College Agric. & Life Sciences
Frank J. Rahel - U. of Wisconsin
Gary C. Schafran - Syracuse U.
Robert Singer - Colgate U.
Chapter E-6 - Indirect Effects on Health
Thomas W. Clarkson - U. of Rochester
Joan P. Baker - North Carolina State U.
William E. Sharpe - Pennsylvania State U.
Chapter E-7 - Effects on Materials
John Yocom - TRC Environ. Consultants, Inc.
Norbert S. Baer - New York U.
iii
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PREFACE
The Acidic Deposition Phenomenon and Its Effects: Critical
Assessment Review Papers public review draft, is a technical review
document in two volumes, prepared and released for a 90-day period of
public technical comment. The Environmental Protection Agency will
develop an interpretive summary, The Acidic Deposition Phenomenon and
Its Effects: Critical Assessment~Pocument, based upon the content of
the Review Papers and the public comments.
The Acidic Deposition Phenomenon and Its Effects: Critical
Assessment Review Papers was requested by the Clean Air Scientific
Advisory Committee (CASAC) of EPA's Science Advisory Board and will be
reviewed by that committee. The CASAC is comprised of independent
scientists who are quite knowledgeable in matters pertaining to
atmospheric pollution and its effects. These scientists will evaluate
the scientific adequacy of the Critical Assessment Document. As part of
this evaluation, the CASAC considers the comments and criticisms of the
general public and scientific community as they pertain to scientific
issues and questions. (Although the science of an issue may obviously
have implications for policy decisions, matters of policy per se are not
in the province of the document.) This review process is essential to
developing a scientifically unimpeachable assessment.
The document's original charge was to prepare 'a comprehensive
document which lays out the state of our knowledge with regard to
precursor emissions, pollutant transformation to acidic compounds,
pollutant transport, pollutant deposition and the effects (both measured
and potential) of acidic deposition.1 It was the decision of the
editors to provide the following guidelines to the authors writing the
Critical Assessment Review Papers to meet this overall objective of the
document:
1. Contributions are 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.
iv
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6. Potential vs known processes/effects is 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.
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
presented, is advised to consider and understand these directives before
reading.
Again, the document has been designed to address our present status
of knowledge relative to 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 con-
clusions are those of the authors after their review of the literature.
The success of the Critical Assessment Review Papers has depended
on the coordinated efforts of many individuals. The document involved
the participation of over 54 scientists contributing material on their
special areas of expertise under the broad headings of either
atmospheric processes or effects. 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 Award in Pollution Control, and recently
retired director of EPA's Environmental Sciences Research Laboratory;
Dr. Linthurst is an ecologist and serves as Program Coordinator for the
Acid Precipitation Program at North Carolina State University. Dr.
Bennett is the Director of the Acid Deposition Assessment Staff in EPA's
Office of Research and Development and provides liaison between the
section editors/contributors and CASAC scientific reviewers.
The United States and Canada in 1980 signed a Memorandum of Intent
to seek agreement on transboundary air pollution issues. A number of
working groups are compiling technical information to support the
negotiations called for by the Memorandum. Although the Critical
Assessment Document and the U.S.-Canada working group reports come from
different origins, and are intended for different purposes, there is
likely to be some overlap in their areas of coverage.
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The written materials to follow are contributions from one to eight
authors per chapter, integrated by the editors. Approximately 75
scientists, with expertise in the fields being addressed, have
participated in reviewing earlier drafts of the chapters. In addition,
200 individuals participated in a public workshop held for the review
of these materials in November of 1982. Numerous changes resulted from
these reviews, and this document reflects those comments. This is the
final public review draft and comments are welcome. However, several
guidelines and forms should be used to submit formal comments. Please
consult the last section of the volume for details.
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, recommendations, and deadlines, leading to this fourth version
of the document. These dedicated persons 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 producing 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. Their concern for the truth, as we know it now,
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.
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS:
CRITICAL ASSESSMENT REVIEW PAPERS
Table of Contents
Volume I
Atmospheric Sciences
Note: Comment forms and guidelines to be used by reviewers can be found at the ends of
Volumes I and II
Page
GLOSSARY (not available)
ACRONYM LIST xxiii
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 1-1
1.2.5 Atmospheric Concentrations and Distributions of Chemical 1-2
Substances 1-2
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 Biogenic Emissions of Sulfur Compounds 2-5
2.2.1.4 Geophysical Sources of Natural Sulfur Compounds 2-16
2.2.1.4.1 Volcanism 2-16
2.2.1.4.2 Marine sources of aerosol particles and
gases 2-20
2.2.1.5 Scavenging Processes and Sinks 2-22
2.2.1.6 Summary of Natural Sources of Sulfur Compounds 2-23
2.2.2 Nitrogen Compounds 2-24
2.2.2.1 Introduction '....I!'. 2-24
2.2.2.2 Estimates of Natural Global Sources and Sinks 2-25
2.2.2.3 Biogenic Sources of NOX Compounds 2-29
2.2.2.4 Tropospheric and Stratospheric Reactions 2-31
2.2.2.5 Formation of NOX by Lightning 2-32
2.2.2.6 Biogenic NQX Emissions Estimate for the United States ... 2-33
2.2.2.7 Biogenic Sources of Ammonia 2-34
2.2.2.8 Oceanic Source for Ammonia 2-38
2.2.2.9 Biogenic Ammonia Emissions Estimates for the United
States 2-39
2.2.2.10 Meteorological and Area Variations for NOX and Ammonia
Emi ssions 2-40
2.2.2.11 Scavenging Processes for NOX and Ammonia 2-40
2.2.2.12 Organic Nitrogen Compounds 2-40
2.2.2.13 Summary of Natural NOX and Ammonia Emissions 2-41
2.2.3 Chlorine Compounds 2-41
2.2.3.1 Introduction \\ 2-41
2.2.3.2 Oceanic Sources 2-42
2.2.3.3 Vol cam' sm 2-46
2.2.3.4 Combustion 2-46
VII
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Table of Contents (continued)
Page
2.2.3.5 Total Natural Chlorine Sources 2-47
2.2.3.6 Seasonal Distributions 2-47
2.2.3.7 Environmental Impacts of Natural Chlorides 2-47
2.2.4 Natural Sources of Aerosol Particles 2-49
2.2.5 Precipitation pH in Background Conditions 2-50
2.2.6 Summary 2-54
2.3 Anthropogenic Emissions , 2-55
2.3.1 Origins of Anthropogenically Emitted Compounds and
Related Issues 2-55
2.3.2 Historical Trends and Current Emissions of Sulfur Compounds 2-58
2.3.2.1 Sulfur Oxides 2-58
2.3.2.2 Primary Sulfate Emissions 2-66
2.3.3 Historical Trends and Current Emissions of Nitrogen Oxides 2-72
2.3.4 Historical Trends and Current Emissions of Hydrochloric Acid (HC1) 2-75
2.3.5 Historical Trends and Current Emissions of Heavy Metals Emitted
from Fuel Combustion 2-79
2.3.6 Historical Emissions Trends in Canada 2-87
2.3.7 Future Trends in Emissions 2-96
2.3.7.1 United States 2-96
2.3.7.2 Canada 2-96
2.3.8 Emissions Inventories 2-98
2.3.9 The Potential for Neutralization of Atmospheric
Acidity by Suspended Fly Ash 2-100
2.4 Conclusions 2-105
2.5 References 2-109
A-3 TRANSPORT PROCESSES
3.1 Introduction 3-1
3.1.1 The Concept of Atmospheric Residence Times 3-1
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 3-11
3.3.1 The Planetary Boundary Laye>" 3-11
3.3.2 Structure of the Transport Layer 3-13
3.3.3 Dynamics of the Transport Layer 3-15
3.3.4 Effects of Mesoscale Complex Systems on Transport Layer Structure
and Dynamics . , 3-28
3.3.4.1 Effect of Mesoscale Convective Precipitation Systems
(MCPS) 3-28
3.3.4.2 Complex Terrain Effects 3-32
3.3.4.2.1 Shoreline environment effects 3-32
3.3.4.2.2 Urban effects 3-35
3.3.4.2.3 Hilly terrain effects 3-36
3.4 Mesoscale Plume Transport and Dilution 3-39
3.4.1 Elevated Point-Source Emissions 3-39
3.4.2 Broad Areal Emissions Near Ground 3-62
3.5 Continental and Hemispheric Transport 3-68
3.6 Conclusions 3-91
3.7 References 3-94
A-4 TRANSFORMATION PROCESSES
4.1 Introduction 4-1
4.2 Homogeneous Gas-Phase Reactions 4-3
4.2.1 Fundamental Reactions 4-3
vi i i
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Table of Contents (continued)
Page
4.2.1.1 Reduced Sulfur Compounds 4-3
4.2.1.2 Sulfur Dioxide 4-4
4.2.1.3 Nitrogen Compounds 4-10
4.2.1.4 Halogens 4-16
4.2.1.5 Organic Acids 4-16
4.2.2 Laboratory Simulations of Sulfur Dioxide and Nitrogen Dioxide
Oxidation 4-18
4.2.3 Field Studies of Gas-Phase Reactions 4-21
4.2.3.1 Urban Plumes 4-21
4.2.3.2 Power Plant Plumes 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 Acid 4-32
4.3.3 Production of HC1 in 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 SO?'H20 to H2S04 ... 4-48
4.3.5.4 Heterogeneous Production of H2S04 in Solution 4-53
4.3.5.5 The Relative Importance of the Various H2S04
Production Mechanisms 4-54
4.3.6 Neutralization Reactions 4-62
4.3.6.1 Neutralizati-on by NHs 4-62
4.3.6.2 Neutralization by Particle-Acid Reactions 4-63
4.3.7 Summary 4-64
4.4 Transformation Models 4-64
4.4.1 Introduction 4-64
4.4.2 Approaches to Transformation Modeling 4-67
4.4.2.1 The Fundamental Approach 4-67
4.4.2.2 The Empirical Approach 4-70
4.4.3 The Question of Linearity 4-70
4.4.4 Some Specific Models and Their Applications 4-75
4.4.4.1 Detailed Chemical Simulations 4-75
4.4.4.2 Parameterized Models 4-77
4.4.5 Summary 4-81
4.5 Conclusions _. 4-83
4.6 References ." 4-87
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
5.2.3 Sulfate 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 5-15
5.2.3.4 Nonurban Composition Measurements 5-19
5.2.3.5 Concentration and Composition Measurements at Remote
Locations 5-22
5.2.4 Particle Size Characteristics of Particulate Sulfur Compounds .... 5-23
ix
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Table of Contents (continued)
Page
5.2.4.1 Urban Measurements 5-23
5.2.4.2 Monurban Size Measurements 5-25
5.2.4.3 Measurements at Remote Locations 5-26
5.3 Nitrogen Compounds 5-27
5.3.1 Introduction 5-27
5.3.2. Nitrogen Oxides 5-27
5.3.2.1 Historical Distribution Patterns and Current
Concentrations of Nitrogen Oxides 5-27
5.3.2.2 Measurements Techniques-Nitrogen Oxides 5-28
5.3.2.3 Urban Concentration Measurements 5-28
5.3.2.4 Nonurban Concentration Measurements 5-29
5.3.2.5 Measurements of Concentrations at Remote Locations 5-33
5.3.3 Nitric Acid 5-35
5.3.3.1 Urban Concentration Measurements 5-35
5.3.3.2 Nonurban Concentration Measurements 5-38
5.3.3.3 Concentration Measurements at Remote Locations 5-43
5.3.4 Peroxyacetyl Nitrates 5-44
5.3.4.1 Urban Concentration Measurements 5-44
5.3.4.2 Nonurban Concentration Measurements 5-46
5.3.5 Ammonia 5-48
5.3.5.1 Urban Concentration Measurements 5-50
5.3.5.2 Nonurban Concentration Measurements 5-50
5.3.6 Particulate 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 5-55
5.3.7 Particle Size Characteristics of Particulate 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 5-65
5.5.3 Concentration Measurements in Rainwater 5-65
5.6 Chlorine Compounds 5-66
5.6.1 Introduction 5-66
5.6.2 Hydrogen Chloride 5-66
5.6.3 Particulate Chloride 5-67
5.6.4 Particle Size tharacteristies of Particulate Chlorine Compounds .. 5-67
5.7 Metallic Elements 5-68
5.7.1 Concentration Measurements and Particle Sizes in Urban Areas 5-69
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
Composi tion 5-74
5.8.1 Fine Particle Concentration and Light Scattering Coefficients .... 5-74
5.8.2 Light Extinction or Light Scattering Budgets at Urban Locations .. 5-75
5.8.3 Light Extinction or Light Scattering Budgets at Nonurban
Locations 5-77
5.8.4 Trends in Visibility as Related to Sulfate Concentrations 5-79
5.9 Conclusions 5-79
5.10 References 5-85
A-6 PRECIPITATION SCAVENGING PROCESSES
6.1 Introduction 6-1
6.2 Steps in the Scavenging Sequence 6-3
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Table of Contents (continued)
Page
6.2.1 Introduction 6-3
6.2.2 Intermixing of Pollutant and Condensed Water (Step 1-2) 6-7
6.2.3 Attachment of Pollutant to Condensed Water Elements (Step 2-3) ... 6-8
6.2.4 Aqueous-Phase Reactions (Step 3-4) 6-15
6.2.5 Deposition of Pollutant with Precipitation (Step 4-5) 6-15
6.2.6 Combined Processes and the Problem of Scavenging Calculations .... 6-18
6.3 Storm Systems and Storm Climatology 6-18
6.3.1 Introduction 6-18
6.3.2 Frontal Storm Systems 6-19
6.3.2.1 Warm-Front Storms 6-20
6.3.2.2 Cold-Front Storms 6-25
6.3.2.3 Occluded-Front Storms 6-25
6.3.3 Convectlve Storm Systems 6-28
6.3.4 Additional Storm Types: Nonideal Frontal Storms, Orographlc
Storms and Lake-Effect Storms 6-28
6.3.5 Storm and Precipitation Climatology 6-30
6.3.5.1 Precipitation Climatology 6-32
6.3.5.2 Storm Tracks 6-32
6.3.5.3 Storm Duration Statistics 6-35
6.4 Summary of Precipitation-Scavenging Field Investigation 6-35
6.5 Predictive and Interpretive Models of Scavenging 6-51
6.5.1 Introduction 6-51
6.5.2 Elements of a Scavenging Model 6-54
6.5.2.1 Material Balances 6-54
6.5.2.2 Energy Balances 6-55
6.5.2.3 Momentum Balances 6-56
6.5.3 Definitions of Scavenging Parameters 6-56
6.5.4 Formulation of Scavenging Models: Simple Examples
of Microscopic and Macroscopic Approaches 6-62
6.5.5 Systematic Selection of Scavenging Models:
A F1 ow Chart Approach 6-65
6.6 Practical Aspects of Scavenging Models: Uncertainty Levels and Sources
of Error 6-68
6.7 Conclusions 6-72
6.8 References 6-75
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 Phoretic Effects and Stefan Flow 7-12
7.2.5 Surface Adhesion ; 7-15
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 7-19
7.2.9 Fog and Dewfall 7-20
7.2.10 Resuspension and Surface Emission 7-21
7.2.11 The resistance Analog 7-22
7.3 Methods for Studying Dry Deposition 7-28
7.3.1 Direct Measurement 7-28
7.3.2 Wind Tunnel and Chamber Studies 7-31
7.3.3 Micrometeorological Measurement Methods 7-33
7.4 Field Investigations of Dry Deposition 7-39
7.4.1 Gaseous Pollutants 7-39
7.4.2 Particulate Pollutants 7-46
7.4.3 Routine Handling in Networks 7-51
XI
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Table of Contents (continued)
Page
7.5 Micrometeorological Models of the Dry Deposition Process 7-52
7.5.1 Gases 7-52
7.5.2 Particles 7-55
7.6 Summary 7-56
7.7 Conclusions 7-60
7.8 References 7-63
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 .... 8-4
8.2.4 Wet Deposition Network Data Bases 8-7
8.3 Monitoring Capabilities for Dry Deposition 8-11
8.3.1 Introduction 8-11
8.3.2 Methods for Monitoring Dry Deposition 8-17
8.3.2.1 Direct Collection Procedures 8-18
8.3.2.2 Alternative Methods 8-20
8.3.3 Evaluations of Dry Deposition Rates 8-21
8.4 Wet Deposition Network Data With Applications to Selected Problems 8-28
8.4.1 Spatial Patterns 8-28
8.4.2 Remote Site pH Data 8-50
8.4.3 Precipitation Chemistry Variations Over Time 8-59
8.4.3.1 Nitrate Varfation Since 1950's 8-59
8.4.3.2 pH Variation Since 1950's 8-61
8.4.3.3 Calciiro Variation Since the I960' s 8-65
8.4.4 Seasonal Variations 8-67
8.4.5 Very Short Time Scale Variations 8-68
8.4.6 Air Parcel Trajectory Analysis 8-68
8.5 Glaciochemical Investigations as a Tool in the Historical Delineation of
the Acid Precipitation Problems 8-70
8.5.1 Glaciochemical Data 8-70
8.5.1.1 Sulfate - Polar Glaciers 8-71
8.5.1.2 Nitrate - Polar Glaciers 8-72
8.5.1.3 pH and Acidity - Polar Glaciers 8-72
8.5.1.4 Sulfate - Alpine Glaciers 8-73
8.5.1.5 Nitrate - Alpine Glaciers 8-73
8.5.1.6 pH and Acidity - Alpine Glaciers 8-73
8.5.2 Trace Metals - General Statement 8-74
8.5.2.1 Trace Metals - Polar Glaciers 8-74
8.5.2.2 Trace Metals - Alpine Glaciers 8-76
8.5.3 Discussion and Future Work 8-76
8.6 Conclusions 8-79
8.7 References 8-83
A-9 LONG-RANGE TRANSPORT AND ACIDIC DEPOSITION MODELS
9.1 Introduction 9-1
9.1.1 General Principles for Formulating Pollution Transport and
Diffusion 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 9-5
9.1.2.3 Reaction Mechanisms 9-5
9.1.2.4 Removal Mechanisms 9-5
xn
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Table of Contents (continued)
Page
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 9-8
9.2 Types of LRT Models 9-8
9.2.1 Eulerian Grid Models 9-8
9.2.2 Lagrangian Models 9-11
9.2.2.1 Lagrangian Trajectory Models 9-11
9.2.2.2 Statistical Trajectory Models 9-13
9.2.3 Hybrid Models 9-13
9.3 Modules Associated with Chemical (Transformation) Processes 9-14
9.3.1 Overview 9-14
9.3.2 Chemical Transformation Modeling 9-14
9.3.2.1 Simplified Modules 9-15
9.3.2.2 Multireaction Modules 9-15
9.3.3 Modules for NOX Transformation 9-16
9.4 Modules Associated with Wet and Dry Deposition 9-20
9.4.1 Overview 9-20
9.4.2 Modules for Wet Deposition 9-21
9.4.2.1 Formulation and Mechanism 9-21
9.4.2.2 Modules Used in Existing Models 9-22
9.4.2.3 Wet Deposition Modules for Snow 9-24
9.4.2.4 Wet Deposition Modules for NOX 9-24
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 : 9-26
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 9-27
9.5.1 Overview 9-27
9.5.2 Model Application 9-27
9.5.2.1 Selection Criteria 9-27
9.5.2.2 Regional Concentration and Deposition Patterns 9-28
9.5.2.3 Use of Matrix Methods to Quantify Source-Receptor
Relationships 9-29
9.5.3 Data Requirements 9-34
9.5.3.1 General 9-34
9.5.3.2 Specific Characteristics of Data Used in Model
Simulations 9-37
9.5.3.2.1 Emissions 9-37
9.5.3.2.2 Meteorological Data 9-38
9.5.4 Model Performance and Uncertainties 9-38
9.5.4.1 General , 9-38
9.5.4.2 Data Bases Available for Evaluating Models 9-40
9.5.4.3 Performance Measures 9-40
9.5.4.4 Representivity of Measurements 9-41
9.5.4.5 Uncertainties 9-41
9.5.4.6 Selected Results 9-42
9.6 Conclusions 9-47
9.7 References 9-49
xm
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS:
CRITICAL ASSESSMENT REVIEW PAPERS
Table of Contents
Volume II
Effects Sciences
Note: Comment forms and guidelines to be used by reviewers can be found at the ends of
Volumes I and II
Page
E-l INTRODUCTION
1.1 Objectives 1-1
1.2 Approach 1-1
1.3 Chapter Organization and General Content 1-2
1.3.1 Effects on Soil Samples 1-3
1.3.2 Effects on Vegetation 1-3
1.3.3 Effects on Aquatic Chemistry 1-4
1.3.4 Effects on Aquatic Biology 1-4
1.3.5 Indirect Effects on Health 1-5
1.3.6 Effects on Materials 1-5
1.4 Acidic Deposition 1-5
1.5 Linkage to Atmospheric Sciences 1-6
1.6 Sensitivity 1-6
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 ... 2-1
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 Acid Soils 2-5
2.2.1 Development of Acid 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 in Soils 2-8
2.2.2.2 Exchangeable Bases and Base Saturation 2-8
2.2.3 Exchangeable and Solution Aluminum in 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 Acidic Deposition on Soil Chemistry and Plant Nutrition 2-19
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 2-25
2.3.3.1 Effects on Cation Nutrient Status 2-29
2.3.3.2 Effects on S and N Status 2-31
2.3.3.3 Acidification Effects on Plant Nutrition 2-34
XIV
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Table of Contents (continued)
Page
2.3.3.3.1 Nutrient deficiencies 2-34
2.3.3.3.2 Metal ion toxicities 2-34
2.3.3.3.2.1 Aluminum toxicity 2-35
2.3.3.3.2.2 Manganese toxicity 2-36
2.3.4 Reversibility of Effects on Soil Chemistry 2-36
2.3.5 Predicting Which Soils will be Affected Most 2-37
2.3.5.1 Soils Under Cultivation 2-37
2.3.5.2 Uncultivated, Unamended Soils 2-37
2.3.5.2.1 Basic cation-pH changes in forested soils .... 2-40
2.3.5.2.2 Changes in aluminum or other metal concen-
tration in soil solution fn forested soils ... 2-41
2.4 Effects of Acidic Deposition on Soil Biology 2-41
2.4.1 Soil Biology Components and Functional Significance 2-41
2.4.1.1 Soil Animals 2-41
2.4.1.2 Algae 2-42
2.4.1.3 Fungi 2-42
2.4.1.4 Bacteria 2-42
2.4.2 Direct Effects of Acidic Deposition on Soil Biology 2-43
2.4.2.1 Soil Animals 2-43
2.4.2.2 Terrestrial Algae 2-44
2.4.2.3 Fungi 2-44
2.4.2.4 Bacteria 2-45
2.4.2.5 General Biological Processes 2-45
2.4.3 Metals--Mobilization Effects on Soil Biology 2-47
2.4.4 Effects of Changes in Microbial Activity on Aquatic Systems 2-48
2.4.5 Soil Biology Summary 2-48
2.5 Effects of Acidic Deposition on Organic Matter Decomposition 2-49
2.6 Effects of Soils on the Chemistry of Aquatic Ecosystems 2-50
2.7 Conclusions 2-56
2.8 References 2-59
E-3 EFFECTS ON VEGETATION
3.1 Introduction 3-1
3.1.1 Overview 3-1
3.1.2 Background 3-2
3.2 Plant Response to Acidic Deposition 3-5
3.2.1 Leaf Response to Acidic Deposition 3-5
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 3-10
3.2.3 Summary 3-17
3.3 Interactive Effects of Acidic Deposition with Other Environmental
Factors on Plants 3-18
3.3.1 Interactions with Other Pollutants 3-18
3.3.2 Interactions with Phytophagus Insects 3-21
3.3.3 Interactions with Pathogens ....; 3-21
3.3.4 Influence on Vegetative Hosts That Would Alter Relationships
with Insect or Microbial Associate 3-24
3.3.5 Effects of Acidic Deposition on Pesticides 3-25
3.3.6 Summary 3-26
3.4 Biomass Production 3-27
3.4.1 Forests 3-27
3.4.1.1 Possible Mechanisims of Response 3-28
3.4.1.2 Phenological Effects 3-30
3.4.1.2.1 Seed germination and seedling establishment .. 3-31
3.4.1.2.2 Mature and reproductive stages 3-33
3.4.1.3 Growth of Seedlings and Trees in Irrigation
Experiments 3-33
XV
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Table of Contents (continued)
Page
3.4.1.4 Studies of long-Term Growth of Trees 3-34
3.4.1.5 Dieback and Decline in High Elevation Forests 3-37
3.4.1.6 Summary 3-41
3.4.2 Crops 3-42
3.4.2.1 Review and Analysis of Experimental Design 3-42
3.4.2.1.1 Dose-response determination 3-43
3.4.2.1.2 Sensitivity classification 3-44
3.4.2.1.3 Mechanisms 3-45
3.4.2.1.4 Characteristics of precipitation simulant
exposures 3-45
3.4.2.1.5 Yield criteria 3-46
3.4.2.2 Experimental Results 3-46
3.4.2.2.1 Field studies 3-47
3.4.2.2.2 Controlled environment studies 3-51
3.4.2.3 Discussion 3-59
3.4.2.4 Summary 3-62
3.5 Conclusions 3-62
3.6 References 3-65
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 4-1
4.2.1 Receiving Systems 4-1
4.2.2 pH, Conductivity, and Alkalinity 4-4
4.2.2.1 pH 4-4
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 4-6
4.3.1 Atmospheric Inputs 4-6
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 Di stribution of Deposition 4-8
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 4-9
4.3.2 Characteristics of Receiving Systems Relative to Being Able to
Assimilate Acidic Deposition 4-10
4.3.2.1 Canopy 4-10
4.3.2.2 Soil 4-12
4.3.2.3 Bedrock 4-14
4.3.2.4 Hydrology 4-15
4.3.2.4.1 Flow paths 4-15
4.3.2.4.2 Residence times 4-17
4.3.2.5 Wetlands 4-17
4.3.2.6 Aquatic 4-18
4.3.2.6.1 Alkalinity 4-18
4.3.2.6.2 International production/consumption
of ANC 4-22
4.3.2.6.3 Aquatic sediments 4-24
4.3.3 Location of Sensitive Systems 4-25
4.3.4 Summary--Sensi tivi ty 4-30
4.4 Magnitude of Chemical Effects of Acidic Deposition on
Aquatic Ecosystems 4-31
XVI
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Table of Contents (continued)
Page
4.4.1 Relative Importance of HN03 vs H2S04 4-31
4.4.2 Short-Term Acidification ..'. 4-37
4.4.3 Long-Term Acidification 4-38
4.4.3.1 Analysis of Trends based on Historic Measurements of
Surface Water Quality 4-44
4.4.3.1.1 Methologlcal problems with the evaluation
of historical trends 4-44
4.4.3.1.1.1 pH 4-44
4.4.3.1.1.1.1 pH-early metho-
dology 4-44
4.4.3.1.1.1.2 pH-current metho-
dology 4-46
4.4.3.1.1.1.3 pH-comparability
of early and cur-
rent mesurement
methods 4-47
4.4.3.1.1.1.4 pH-general
problems 4-47
4.4.3.1.1.2 Conductivity 4-48
4.4.3.1.1.2.1 Conductivity
methodology 4-48
4.4.3.1.1.2.2 Comparability of
early and current
measurement
methods 4-48
4.4.3.1.1.2.3 General problems.. 4-48
4.4.3.1.1.3 Alkalinity 4-49
4.4.3.1.1.3.1 Early methodology. 4-49
4.4.3.1.1.3.2 Current
methodology 4-49
4.4.3.1.1.3.3 Comparability of
early and current
measurement
methods 4-50
4.4.3.1.1.4 Summary of measurement
techniques 4-51
4.4.3.1.2 Analysis of trends 4-51
4.4.3.1.2.1 Introduction 4-51
4.4.3.1.2.2 Canadian studies 4-53
4.4.3.1.2.3 United States studies 4-61
4.4.3.1.3 Summary—trends in historic data 4-74
4.4.3.2 Assessment of Trends Based on Paleol imnological
Technique 4-77
4.4.3.2.1 Calibration and accuracy of paleol imnological
reconstruction of pH history 4-78
4.4.3.2.2 Lake acidification determined by
paleol imnological reconstruction 4-78
4.4.3.3 Alternate Explanations -for Acidification-Land Use
Changes 4-79
4.4.3.3.1 Variations in the groundwater table 4-79
4.4.3.3.2 Accelerated mechanical weathering or
land scarification 4-79
4.4.3.3.3 Decomposition of organic matter 4-80
4.4.3.3.4 Long-term changes in vegetation 4-80
4.4.3.3.5 Chemical amendments 4-80
4.4.3.3.6 Summary--Effects of land use changes
or acidification 4-80
4.4.4 Summary—Magnitude of Chemical Effects of Acidic Deposition 4-81
4.5 Predictive Modeling of the Effects of Acidic Deposition
on Surface Waters 4-82
xvn
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Table of Contents (continued)
Page
4.5.1 Aimer/Dick son Relationship 4-83
4.5.2 Henriksen's Predictor Nomograph 4-88
4.5.3 Thompson Cation Denudation Rate Model (CDR) 4-91
4.5.4 Summary of Predictive Modeling 4-94
4.6 Indirect Chemical Changes Associated with Acidification
of Surface Waters 4-94
4.6.1 Metals » 4.94
4.6.1.1 Increased Loading of Metals From Atmospheric
Deposition 4.95
4.6.1.2 Mobilization of Metals by Acidic Deposition 4-97
4.6.1.3 Secondary Effects of Metal Mobilization 4-98
4.6.1.4 Effects of Acidification on Aqueous Metal Speciation 4-98
4.6.1.5 Indirect Effects on Metals 1n Surface Waters 4-98
4.6.2 Aluminum Chemistry 1n Dilute Acidic Waters 4-99
4.6.2.1 Occurrence, Distribution, and Sources of Aluminum 4-99
4.6.2.2 Aluminum Speciation 4-102
4.6.2.3 Aluminum as a pH Buffer 4-102
4.6.2.4 Temporal and Spatial Variations 1n Aqueous
Aqueous Levels of Aluminum 4-104
4.6.2.5 The Role of Aluminum In Altering Element Cycling
Within Acidic Waters 4-106
4.6.3 Organics 4-108
4.6.3.1 Atmospheric Loading of Strong Acids and Associated
Organic Micropollutants 4-108
4.6.3.2 Organic Buffering Systems 4-109
4.6.3.3 Organo-Metallc Interactions 4-109
4.6.3.4 Photochemistry 4-110
4.6.3.5 Carbon-Phosphorus-Alumlnupi Interactions 4-110
4.6.3.6 Effects of Acidification on Organic Decomposition
1n Aquatic Systems 4-110
4.7 Mitigative Strategies for Improvement of Surface Water Quality 4-111
4.7.1 Base Additions 4-111
4.7.1.1 Types of Basic Materials 4-111
4.7.1.2 Direct Water Addition of Base 4-115
4.7.1.2.1 Computing base dose requirements 4-115
4.7.1.2.2 Methods of base application 4-119
4.7.1.3 Watershed Addition of Base 4-123
4.7.1.3.1 The concept of watershed
application of base 4-123
4.7.1.3.2 Experience In watershed liming 4-124
4.7.1.4 Water Quality Response to Base Treatment 4-126
4.7.1.5 Cost Analysis, Conclusions and Assessment of Base
Addition 4-128
4.7.1.5.1 Cost analysis 4-128
4.7.1.5.2 Summary—base additions 4-130
4.7.2 Surface Water Fertilization 4-130
4.7.2.1 The Fertilization Concept 4-130
4.7.2.2 Phosphorous Cycling in Acidified Water 4-132
4.7.2.3 Fertilization Experience and Water
Quality Response to Fertilization 4-133
4.7.2.4 Summary-Surface Water Fertilization 4-134
4.8 Conclusions 4-134
4.9 References 4-137
E-5 EFFECTS ON AQUATIC BIOLOGY
5.1 Introduction 5-1
5.2 Biota of Naturally Acidic Waters 5-3
xvm
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Table of Contents (continued)
Page
5.2.1 Types of Naturally Acidic Waters 5-3
5.2.2 Biota of Inorganic Acldotrophlc Waters 5-4
5.2.3 Biota In Acidic Brownwater Habitats 5-6
5.2.4 Biota In Ultra-Ollgotrophic Waters 5-8
5.2.5 Summary 5-9
5.3 Benthic Organisms 5-15
5.3.1 Importance of the Benthic Community 5-15
5.3.2 Effects of Acidification on
Components of the Benthos 5-16
5.3.2.1 Mlcroblal Community 5-17
5.3.2.2 Perlphyton 5-18
5.3.2.2.1 Field surveys 5-18
5.3.2.2.2 Temporal trends 5-19
5.3.2.2.3 Experimental studies 5-21
5.3.2.3 Microlnvertebrates 5-22
5.3.2.4 Crustacea 5-23
5.3.2.5 Insecta 5-25
5.3.2.5.1 Sensitivity of different groups 5-25
5.3.2.5.2 Sensitivity of Insects from different
mlcrohabitats 5-30
5.3.2.5.3 Acid sensitivity of Insects based on food
sources 5-31
5.3.2.5.4 Mechanisms of effects and trophic
Interactions 5-31
5.3.2,6 Mollusca 5-32
5.3.2.7 Annelida 5-33
5.3.2.8 Summary of Effects of Acidification on Benthos 5-34
5.4 Macrophytes and Wetland Plants 5-39
5.4.1 Introduction ; 5-39
5.4.2 Effects on Acidification on Aquatic Macrophytes 5-43
5.4.3 Summary 5-45
5.5 Plankton 5-45
5.5.1 Introduction 5-45
5.5.2 Effects of Acidification on Phytoplankton 5-47
5.5.2.1 Changes in Species Composition 5-47
5.5.2.2 Changes in Phytopl ankton Bloroass and Productivity 5-54
5.5.3 Effects of Acidification on Zooplankton 5-57
5.5.4 Explanations and Significance 5-70
5.5.4.1 Changes in Species Composition 5-70
5.5.4.2 Changes in Productivity 5-72
5.5.5 Summary 5-75
5.6 Fishes 5-76
5.6.1 Introduction 5-76
5.6.2 Field Observations 5-77
5.6.2.1 Loss of Populations 5-78
5.6.2.1.1 United States 5-78
5.6.2.1.1.1 Adirondack Region of
New York State 5-78
5.6.2.1.1.2 Other regions of the eastern
United States 5-81
5.6.2.1.2 Canada 5-82
5.6.2.1.2.1 LaCloche Mountain Region of
Ontario 5-82
5.6.2.1.2.2 Other areas of Ontario 5-86
5.6.2.1.2.3 Nova Scotia 5-86
5.6.2.1.3 Scandinavia and Great Britain 5-91
5.6.2.1.3.1 Norway 5-91
5.6.2.1.3.2 Sweden 5-95
5.6.2.1.3.3 Scotland 5-95
XIX
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Table of Contents (continued)
Page
5.6.2.2 Popul ati on Structure 5-97
5.6.2.3 Growth 5-100
5.6.2.4 Episodic Fish K111 s 5-103
5.6.2.5 Accumulation of totals in Fish 5-105
5.6.3 Field Experiments 5-105
5.6.3.1 Experimental Acidification of Lake 223 Ontario 5-105
5.6.3.2 Experimental Acidification of Norris
Brook, New Hampshi re 5-108
5.6.3.3 exposure of Fish to Acidic Surface Waters 5-108
5.6.4 Laboratory Experiments 5-112
5.6.4.1 Effects of Low pH 5-113
5.6.4.1.1 Survival 5-113
5.6.4.1.2 Reproduction 5-116
5.6.4.1.3 Growth 5-123
5.6.4.1.4 Behavior 5-124
5.6.4.1.5 Physiological responses 5-124
5.6.4.2 Effects of Aluminum and Other Metals in Acidic Waters ... 5-127
5.6.5 Summary 5-129
5.6.5.1 Extent of Impact 5-129
5.6.5.2 Mechanism of Effect ; 5-131
5.6.5.3 Relationship Between Water Acidity and Fish
Population Response 5-133
5.7 Other Related Biota 5-137
5.7.1 Amphibians 5-137
5.7.2 Birds 5-138
5.7.2.1 Food Chain Alterations 5-138
5.7.2.2 Heavy Metal Accumulation 5-139
5.7.3 Mammals 5-140
5.7.4 Summary ; 5-141
5.8 Observed and Anticipated Ecosystem Effects 5-144
5.8.1 Ecosystem Structure 5-144
5.8.2 Ecosystem Function 5-146
5.8.2.1 Nutrient Cycling 5-146
5.8.2.2 Energy Cycling 5-146
5.8.3 Summary 5-147
5.9 Mitigative Options Relative to Biological Populations at Risk 5-148
5.9.1 Biological Response to Neutralization 5-148
5.9.2 Improving Fish Survival in Acidified Waters 5-150
5.9.2.1 Genetic Screening 5-150
5.9.2.2 Selective Breeding 5-151
5.9.2.3 Acclimation 5-152
5.9.2.4 Limitations of Techniquest to Improve Fish Survival 5-153
5.9.2.5 Summary 5-154
5.10 Conclusions 5-154
5.10.1 Effects of Acidification on Aquatic Organisms 5-155
5.10.2 Processes and Mechanisms by Which Acidification
Alters Aquatjc Ecosystems 5-161
5.10.2.1 Direct Effects of Hydrogen Ions 5-161
5.10.2.2 Elevated Metal Concentrations 5-161
5.10.2.3 Altered Trophic-Level Interactions 5-162
5.10.2.4 Altered Water Clarity 5-162
5.10.2.5 Altered Decomposition of Organic Matter 5-162
5.10.2.6 Presence of Algal Mats 5-163
5.10.2.7 Altered Nutrient Availability 5-163
5.10.2.8 Interaction of Stresses 5-163
5.10.3 Biological Mitigation 5-164
5.10.4 Summary 5-164
5.11 References 5-165
XX
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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 Bioaccumulation of Toxic Metals 6-2
6.2.2.1 Speciation (Mercury) 6-2
6.2.2.2 Concentrations and Speciations in Water (Mercury) 6-5
6.2.2.3 Flow of Mercury In the Environment 6-5
6.2.2.3.1 Global cycles 6-6
6.2.2.3.2 Biogeoc hem leal cycles of Mercury 6-6
6.2.3 Accumulation in Fish 6-10
6.2.3.1 Factors Affecting Mercury Concentrations in Fish 6-11
6.2.3.2 Historical and Geographic Trends in Mercury Levels in
Freshwater Fish 6-22
6.2.4 Dynamics and Toxicity in Humans (Mercury) 6-24
6.2.4.1 Dynamics in Man (Mercury) 6-24
6.2.4.2 Toxicity in Man 6-25
6.2.4.3 Human Exposure from Fish and Potential for Health
Risks 6-31
6.3 Ground Surface and Cistern Waters as Affected by Acidic Deposition 6-34
6.3.1 Water Supplies 6-34
6.3.1.1 Direct Use of Precipitation (Cisterns) 6-35
6.3.1.2 Surface Water Supplies 6-36
6.3.1.3 Groundwater Supplies 6-40
6.3.2 Lead 6-43
6.3.2.1 Concentrations in Noncontaminated Waters 6-43
6.3.2.2 Factors Affecting Lead Concentrations
in Water, Including Effects of pH 6-43
6.3.2.3 Speciation of Lead in Natural Water 6-45
6.3.2.4 Dynamics and Toxicity of Lead in Humans 6-45
6.3.2.4.1 Dynamics of lead 1n humans 6-45
6.3.2.4.2 Toxic effects of lead on humans 6-46
6.3.2.4.3 Intake of lead in water and potential for
human health effects 6-53
6.3.3 Aluminum 6-57
6.3.3.1 Concentrations in Uncontaminated Waters 6-57
6.3.3.2 Factors Affecting Aluminum Concentrations in Water 6-58
6.3.3.3 Speciation of Aluminum in Water 6-58
6.3.3.4 Dynamics and Toxicity in humans 6-58
6.3.3.4.1 Dynamics of aluminum in humans 6-59
6.3.3.4.2 Toxic effects of aluminum in man 6-59
6.3.3.5 Human Health Risks from Aluminum in Water 6-59
6.4 Other Metals 6-60
6.5 Conclusions 6-60
6.6 References 6-63
E-7 EFFECTS ON MATERIALS
7.1 Introduction 7-1
7.1.1 Long Range and Local Effects 7-2
7.1.2 Inaccurate Claims of Acid Rain Damage to Materials 7-5
7.1.3 Complex Mechanisms of Exposure and Deposition 7-5
7.1.4 Laboratory vs Field Studies 7-7
7.1.5 Measurement of Materials Damage 7-7
7.1.5.1 Metals 7-7
7.1.5.2 Coatings 7.3
7.1.5.3 Masonry 7-8
7.1.5.4 Paper and Leather 7-8
7.1.5.5 Textiles and Textile Dyes 7-8
XXI
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Table of Contents (continued)
Page
7.2 Mechanisms of Damage to Materials 7-8
7.2.1 Metals 7-9
7.2.2 Stone 7-10
7.2.3 Glass 7-12
7.2.4 Concrete 7-12
7.2.5 Organic Materials 7-12
7.2.6 Deposition Velocities 7-13
7.3 Damage to Materials by Acidic Deposition 7-13
7.3.1 Metals k. 7-13
7.3.1.1 Ferrous Metals 7-15
7.3.1.1.1 Laboratory Studies 7-18
7.3.1.1.2 Field Studies 7-19
7.3.1.2 Nonferrous Metals 7-23
7.3.1.2.1 Aluminum 7-23
7.3.1.2.2 Copper 7-25
7.3.1.2.3 Zinc 7-25
7.3.2 Masonry 7-26
7.3.2.1 Stone 7-26
7.3.2.2 Ceramics and Glass 7-30
7.3.2.3 Concrete 7-30
7.3.3 Paint 7-31
7.3.4 Other Materials 7-35
7.3.4.1 Paper 7-35
7.3.4.2 Photographic Materials 7-35
7.3.4.3 Textiles and Textile Dyes 7-36
7.3.4.4 Leather 7-36
7.3.5 Cultural Property 7-37
7.3.5.1 Architectural Monuments 7-37
7.3.5.2 Museuns, Librarties and Archives 7-37
7.3.5.3 Medieval Stained Glass 7-38
7.3.5.4 Conservation and Mitigation Costs 7-38
7.4 Economic Implications 7-40
7.5 Mitigative Measures 7-42
7.6 Conclusions 7-43
7.7 References 7-44
XXI1
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Acronym and Abbreviation List
ADI (acceptable daily intake) E-6
AL (Aeronomy Laboratory, NOAA)
6-ALA (s-aminolevulinic acid) E-6
ANC (acid neutralizing capacity) E-4
ARL (Air Resources Lab, NOAA)
ARS (Agricultural Research Service, DOA)
BCF (biconcentration factor) E-6
BLM (Bureau of Land Management, DOI)
BLMS (boundary layer models) A-9
BM (Bureau of Mines, DOI)
BNC (base neutralizing capacity) E-4
BNC aq (aqueous base neutralizing capacity) E-4
BOD (biologic oxygen demand)
BS (base saturation) E-4
BSC (base saturation capacity) E-4
BUREC (Bureau of Reclamation, DOI)
BWCA (Boundary Water Canoe Area)
CANSAP (Canadian Sampling Network for Acid Precipitation)
CB (base cation level) E-4
CDR (cation denudation rate) E-4
CEC (cation exchange capacity) E-2
CEQ (Council on Environmental Quality)
CSI (calcite saturation index) E-4
CSRS (Cooperative State Research Service, DOA)
xxm
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DOA (Department of Agriculture)
DOC (dissolved organic carbon) E-4
DOD (Department of Defense)
DOE (Department of Energy)
DOI (Department of Interior)
DOS (Department of State)
ELA (experimental lakes area) E-4
ENAMAP (Eastern North America Model of Air Pollutants)
EPA (Environmental Protection Agency)
EPRI (Electric Power Research Institute)
ERDA (Energy Research and Development Agency (defunct)
ESRL (Environmental Sciences Research Laboratory, EPA)
FA (fulvic acid) E-4
FDA (flourescein diacetate) E-2
FEP (free erythrocyte protoporphyrin) E-6
FGD (Flue Gas Desulfurization)
FS (Forest Service, DOA)
FWS (Fish and Wildlife Service, DOI)
GTN (Global Trends Network)
HHS (Department of Health and Human Services)
ILWAS (Integrated Lake Watershed Acidification Study) E-4
LAI (leaf area index) A-7
LIMB (Limestone Injection/Multistage Burner)
LRTAP (Long-Range Transboundary Air Pollution)
LSI (Langelier Saturation Index) E-6
MAP3S (Multi-State Atmospheric Power Production
Pollution Study)
xxiv
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MCPS (Mesoscale convective precipitation systems) A-3
MOI (Memorandum of Intent, U.S.-Canada)
NADP (National Atmospheric Deposition Program)
NASA (National Aeronautics and Space Administration)
NATO (North Atlantic Treaty Organization)
NBS (National Bureau of Standards, DOC)
NCAR (National Center for Atmospheric Research)
NECRMP (Northeast Corridor Regional Modeling Program) A-2
NOAA (National Oceanic and Atmosperic Administration, DOC)
NPS (National Park Service, DOI)
NSF (National Science Foundation)
NSPS (New Source Performance Standards)
NTN (National Trends Network)
NWS (National Weather Service, NOAA)
OECD (Organization for Economic Cooperation and
Development)
OMB (Office of Management and Budget)
ORNL (Oak Ridge National Laboratory)
OSM (Office of Surface Mining, DOI)
PAN (peroxyacetyl nitrate) E-3, A-5
PBCF (practical biconcentration factor) E-6
PBL (planetary boundary layer) A-4
PGF (pressure gradient force) A-3
PHS (Public Health Service)
RSN (Research Support Network)
SAC ($04 adsorption capacity) E-4
SEAREX
xxv
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SAES (State Agricultural Experiment Station, DOA)
SCS (Soil Conservation Service, DOA)
SURE (Sulfate Regional Experiment, EPRI)
TFE (total fixed endpoint alkalinity) E-4
TIP (total inflection point alkalinity) E-4
TVA (Tennessee Valley Authority)
USGS (United States Geological Survey, DOI)
VOC (Volatile Organic Compounds)
WMO (World Meteorological Organization)
xxvi
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS
E-l. INTRODUCTION
(R. A. Llnthurst)
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 understanding 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 diagramatically presents a conceptual flow of wet
and dry deposition 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 ground water system. At any point along this
pathway, the chemistry of precipitation can be significantly altered.
As a result, the complexities of quantifying effects in relation to a
chemical dose becomes increasingly 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.
1-1
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INPUTS
GASEOUS
OUTPUT
ROOT
TURNOVER
LEACHING
(biological export)
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. Research
18(3):449-461)
1-2
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A common example of this concept is lake and watershed interactions.
Small lakes surrounded by large watersheds are more greatly influenced
by those waters which pass through the terrestrial landscape prior to
entering the lake; since 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 addition to these direct alterations of the system
components, indirect effects can also occur. Soils, for example, if
chemically altered, ultimately affect vegetation responses; soils being
the medium in which plants grow. 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 only be fully understood with
this perspective in mind. However, for convenience of presentation,
each major ecosystem component has been somewhat artificially separated
from the others and subsequently 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 deposition 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 watershed perspective, continuing to rebuild
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; in
the interest of establishing a general sense of what will be found in
more detail in the chapters to follow.
1-3
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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 vegetation. Therefore, in this chapter, emphasis is placed on
the natural processes that contribute to acidification, nutrient status,
and metal movement in soils. The effects of acidic and acidifying
substances on these natural processes is then superimposed as an
additive factor, and their contribution 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 sections of the chapter are
chemically oriented and some basic soil chemistry is also included.
Nutrient cycling aspects of acidic deposition influences on soils
is the primary emphasis of the chapter. Both the chemical and
biological components 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.
The 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 variables. 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 vegetation collectively includes the primary producers in the
food web, its importance 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 section discusses a
diverse range of acidic deposition/pi ant 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 deposition, plant, and environmental
condition interactions, leading to quantification of plant response, are
presented. Special attention is given to the concept of cumulative
effects on forests over time and the lack of data in this field of
acidic deposition effects at the present time. 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 any
other component 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
1-4
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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 makes it
difficult to isolate single causes of response, whether they be
beneficial or detrimental. At the present level of understanding of
plant response as influenced by general stress factors, the direct and
indirect effects of acidic deposition that can be definitely stated are
extremely 1 imi ted.
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
ecosystan 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 discussions of methodology and historical trends are
relevant to drawing conclusions 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 components 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, at this time, our 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 rather, define the link
between biological response and acidification, whatever the cause.
The chapter discusses the biota found in naturally acidic systems,
recognizing that such systems have and will always exist. Such
1-5
409-262 0-83-2
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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 oligotrophic 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 response, field
and laboratory evidence for changes in aquatic biota resources, and
biological mitigation options are also presented and evaluated.
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 deposition 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 is available on the potential and known effects of
acidic deposition on human health. Food chain dynamics are discussed in
a bioaccumulation 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 highlighted. 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. 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. Deposition, both wet and dry, contains both
1-6
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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 the inputs of sulfur and nitrogen can be acidic
upon delivery, or can become acidifying as they cycle through the
system, these substances are the critical elements for discussion.
Because the data bases were not sufficient to conclusively define input
limits for 'protection' of biological resources, there was no need to
deal with a separation of wet and dry forms of deposition. When
simulated treatments are involved, differentiation of deposition forms
is noted as necessary; e.g., in the crop productivity discussion.
Although an effort to separate the components 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 conclusion 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 'cause', 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 'sensitivity' lends itself
to varied interpretations. Sensitivity, as used in the effects
chapters, refers to the relative potential 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, to acidic inputs than
will one that is generally classified 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 sensitivity is clear. For
example, reference to a sensitive soil is not meaningful. 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 'sensitive1 to change. Likewise,
particular tree species, aquatic organisms, processes, and/or materials
can be sensitive to change due to acidic deposition. However,
1-7
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developing sensitivity classifications 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 discussions 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 influencing variables has been considered.
What is presented is the authors'/editors' perspective of key issues.
This does not infer that other issues are 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. W. McFee, F. Adams, C. S. Cronan, M. K. Firestone,
C. D. Foy, R. D. Harter, and D. W. Johnson)!
2.1 INTRODUCTION
Soil plays a key role in ecosystems. It is one of their most
stable components 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-piant relationships, and to soil-water
relationships.
2.1.1 Importance of Soils to Aquatic Systems
Aquatic systems receive diverse outputs from terrestrial
ecosystems. Influences 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
understanding of the processes is rather incomplete.
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
lAll of these authors have contributed to this chapter. Because of
subsequent integration of the material, these authors are not
identified by section.
2-1
-------
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 (Ultisols 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; physical changes, such as accelerated weathering
2-2
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rates or changes in aggregation; or biological changes, such as
reductions in nitrification or other processes.
2.1.3.1 Soil Physical Properties--Soil physical properties are never
independent 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 composition. 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 Soil Microbiology—Biological processes in soils may be
influenced by acid precipitation 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 precipitation, the impact could be
significant. Studies of relationships of soil acidity to biochemical
activity are plentiful. However, most have doubtful applications to the
acid rain 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 rain (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 microbe in the soil contribute to resistance to acid
rain 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 terrestrial ecosystem is shown in Figure 2-1. In a forested ecosystem
(to a lesser degree on cropland, also), a major portion of the
precipitation is intercepted by foliage. The chemical properties of the
resultant throughfall and stemflow can be substantially altered from the
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-------
•^v
*r
INTERCEPTION
DIRECT DEPOSITION
SURFACE FLOW
=>—==—-—'
Minimum to
Moderate soil
interaction
CHANNELIZED FLOW
Minimum soil interaction
GROUNDWATER FLOU
DIFFUSION FLOW
Maximum soil interaction
IMPERVIOUS ZONE —
- _ -
/ __ i k ,
Figure 2-1. Flow paths of precipitation through a terrestrial system.
2-4
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incipient precipitation (see Section 3.2.1.2). While this alternation
may be of no importance in 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 laterally 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 surficial materials to a greater extent than does
surface flow in cultivated areas. The amount of interaction will be
proportional to path length and flow rate.
In uncultivated areas, many large channels are 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 dominant 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 is presented here as background for understanding the sections
that follow. Those already familiar with these concepts may wish to
proceed to 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 determine 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 rain-
fall 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 mobile anions.
The H+ ions can be donated from a variety of sources.
2-5
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2.2.1.1 Biological Sources of H+ Ions— Al though H+ ions may be
generated by chemical weathering of minerals through hydrolytic
reactions, the significant sources of H+ production in soils are all
based on biological reactions.
Oxidation of sulfur and sul fides can be important natural sources
of acidity. Much of the sulfur in soils is present in a highly reduced
state. This includes 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 = $042- + 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 sul fide minerals, e.g., pyrite, are exposed to atmospheric
oxygen, oxidation 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 sul fide 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 in only 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 = NOa- + 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.
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 atmospheric air. Soil air commonly contains up to 1
percent CO?. in comparison to the 0.033 percent of a normal atmosphere
(Patrick 1977). 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.
2-6
-------
(H+) = [1.50 x 10-10 x % c02]l/2.
If atmospheric C02 is 0.033 percent, then H+ activity of rainwater
is 2.2 X 10-6M (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, NfyNOs, (NHa^CO, NH4H?P04,
(NH/^HPOa, and (Nfy^SO^ Because nitrogen is often used at
rates of TOO to 200 kg ha-1, fertilizers alone may generate H+ 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 NH4+ 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 dissociate, 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 ^ 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 downward through the soil
profile, it carries basic cations equivalent to its anionic content.
Meanwhile, the adsorbed H+ remains in place with the soil particles,
causing the soil to become more acid.
2-7
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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 Soi Is—To have a CEC,
soil particles must have a net negative charge.Soil clay particles may
have a negative charge due to isomorohous substitution of AP+ for
Si4+ in tetrahedral layers and of Mg*+ 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 h^O. At low pH, the coordinating
ligand tends to be H20, 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 contents. In the highly weathered Ultisol soils common to the
Southeast, surface-soil clays are usually kaolinite and hydroxy-Al
intergrade vermiculite. 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~*. In soils with a more temperate climate in the
eastern half of the United States, soil organic matter is usually higher
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 Saturati'on--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 of the CEC that is satisfied by basic cations is
defined as "base saturation." For a particular soil and CEC method, a
2-8
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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 extraction, (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 NH4OAc extraction at pH 7.0, method (2) above. For soils with
simiTar characteristics, pH can be used as a reasonable estimate of base
saturation. For example, the "soil pH" - "base saturation" relationship
of 111tisols 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
exchangeable cations can be expressed in terms of saturation of the CEC.
This concept 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 Soils
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.
As H+ concentration increases in soil solution, the stability of clay
minerals decreases, resulting in the release of A13+ 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
2-9
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8.0
7.0
6.0
2 5.0
to
ID
O
UJ
Z3
cr
4.0
3.0
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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case, A13+ activity in soil solution is a function only of pH because
of the reaction
A1(OH)3 + 3H+ = Al3+ + 3^0.
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 20 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
aluminosilicates, 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 A^SigOjjfOH)^ 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+ + i0g Si (OH) - 3 log
H+, varies from 5.6 for amorphous halloysite to 3.25 for crystalline
kaolinite. If Si(OH)4 in soil solution is 0.2 mM (a reasonable value
for acid soils), then Al3+ 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.
The relative solubilities of Al oxides and aluminosilicates in
soils show that soil-solution Al3+ activity, at the same pH, varies
according to the solubility of the Al -control ling 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 nonagricul tural 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
such insoluble precipitates as variscite, A1(OH)2H2P04. Dilute,
acid solutions of Al react with sul fate to form insoluble compounds but
these compounds will be the controlling 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 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 sul fate are present.
2-11
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Inorganic aluminum is present in acid soil solutions primarily as
monomeric ions, the most common ones being Al3+, A10H2+,
A1(OH)2+, Al(OH)3o, A1S04+ and A1H9POA2+. 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 limiting plant growth in acid soils (Adams and Pearson 1967).
Because of the pH-dependent solubility of Al, phytotoxic levels of
solution Al can be expected 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 (Mn2+) 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.
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 accompanied by
the rapid decomposition of organic matter. Consequently, over the
short-term, toxic levels of Mn are more likely under poorly aerated
conditions. 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 symtoms 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
accelerates 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,
2-12
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phosphates, and 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. Nitrogen 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 decomposition 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 atmosphere 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 weathering are the important
neutralizing forces. These forces do not normally reverse the natural
acidification trends, but modify the rate and distribution of
acidification within the soil profile.
The effectiveness with which soil acidity can be neutralized by
liming depends 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.
2-13
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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 Al3+ and H+ to the soil solution from exchange sites;
the other is lime dissolution and the hydrolysis of COs2'. When
exchangeable Al3+ js 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 CaCOa + 3 1^0 = 3 Ca-soil + 2 A1(OH)3 + 3 C02-
With thorough mixing of small lime particles with an acid soil, the
neutralization 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 C02 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
resul ts.
2-14
-------
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 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 CaCl?
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
horizontally, 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 ^S042~).
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 of adsorb 5042-. Sulfate adsorption
results in the displacement of OH° or OH2 from iron or aluminum
hydroxide surfaces (Rajan 1978). This results in an increased negative
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 liquids.
The question of reversibility of sulfate adsorption is crucial to
the long-term effects of acidic deposition on soil leaching. If sulfate
is irreversibly adsorbed, sulfate adsorption can be viewed as increasing
2-15
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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 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 H20) under field conditions but not in
laboratory conditions (Johnson and Henderson 1979), a phenomenon likely
related to slow reactions. Microbial immobilization 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.
2-16
-------
oo
CQ
C£.
O
1/1
O
-------
In the case 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 equilibrium sulfate concentration 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.
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 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 (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 C02» organic matter, nitrification, and
sulfur oxidation. The buffer capacity of soils partially neutralizes
H+ input by reactions with carbonates (> 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 Ultisol 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 H+ 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 alumino-silicate 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.
2-18
-------
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 depositional materials could be expected to have acidifying
effects: H2S04, HN03, H2SOs, S02, S, NHs, NH4S04, 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 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 demonstrated 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 leaching processes due to carbonic acid
formation, organic acid formation, vegetative cation uptake, and a
variety of management practices (Reuss 1977, Johnson et al. 1977,
Andersson et al. 1980, Soil ins et alI. 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 deposition 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,
2-19
-------
TABLE 2-1. RELATIVE ACIDIFYING AND NEUTRALIZING POWER OF
MATERIALS ADDED TO SOILS
Source
Potential acid or base effect
Agricultural liming operation
5000 kg CaCOa ha'1
Neutralizing or basic effect
100 keq ha'1 10 eq m"z
Nitrogen fertilization with
reduced form of N, such as
urea or Nfy
70 kg N ha'1
Acidifying effect9
10 keq ha'1
1 eq
Atmospheric deposition Acidifying effect
1 year (100 cm) pH 4.0 rain 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 m"2
Internal acid production in
soils due to carbonic and
organic acids in one year
from Table 2.5
Acidifying effect
.23-22.7 keq ha-1
.023-2.27 eq nr2
N fertilization usually has somewhat less actual acidifying effect.
This is the maximum assuming complete nitrification of the N
fertilizer.
2-20
-------
but in other cases it would be quite small and of little consequence in
natural ecosystems. Unfortunately, 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, Parrel! et al. 1980,
Wainwright 1980, Stuanes 1980, Bjor and Teigen 1980) either used
accelerated application rates far exceding natural precipitation or
applied concentrated acid. Both 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, Soil ins 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.
2-21
-------
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-1, 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.
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 cc"1, 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 keq ha-1 yr-1 generated by N-fixation-metabolism, organic
matter decay and C02 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.
2-22
-------
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'1 YR'1)
Soil
CEC
meq 100 g~l
Assumption
1 2 3
years
Midwestern Alfisol
Southeastern Ultisol
Ouartzipsamnent
15
9
3
75
45
15
110
67
22
220 oo
125 °°
45-90
with low organic matter
Assumption 1.
Assumption 2.
Assumption 3.
All incoming H+ exchanges for (replaces) basic cations
on the soil exchange complex. There are no accompanying
basic cations and no weathering or other input of basic
cations. This is the "worst case" situation and cannot
exist in nature.
The incoming H+ is accompanied by 0.3-0.5 keq ha~*
yr-1 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'1 yr-1 (for example, 20 kg Ca ha"1 of 15 kg Ca
plus 3 kg Mg ha~l yr"1 within range calculated for
Hubbard Brook (Likens et al. 1977) and the cycling of
basic cations back to soil surface by plants.
2-23
-------
The evidence for acidification of soils by the present rate of
acidic deposition 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 cations unless the associated sulfate or nitrate anions in the
soil are mobile. Evidence indicates that sulfate is not always mobile
(Section 2.2.8) particularly 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 predictive models
seriously limit the models' applicability to natural systems. 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~i yr~l and 20 kg N ha~l yr~l deposited in precipitation in
2-24
-------
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 agronomic 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~l
yr-1. In northern Alabama, Kentucky, Tennessee, and Virginia the
levels were much higher (10 to 30 kg ha~l yr-1) (Jordan et al.
1959). These can be compared with the recent NADP data for wet
deposition of S (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-1 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 are much higher. Therefore, it is generally
accepted that atmospheric N 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 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. Solubility (availability) and weathering
reactions are discussed in Section 2.2.
Acid rain has created a major concern because of the potential for
accelerated cation leaching from forest soils and eventual losses of
productivity (Engstrom et al. 1971). This concern was the driving force
for numerous empirical studies of acid rain 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 rain's effects on forest
2-25
-------
TABLE 2-3. AMOUNTS OF SULFUR FOUND IN PRECIPITATION IN VARIOUS STATES
State
Southern States
Al abama
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 Major source
kg
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
General
General
Steam Plant
Steam Plant
General
Urban
General
General
General
General
Industry
General
General
General
General
Industry
General
Industry
General
Average
S ha-1 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
2-26
-------
TABLE 2-3. CONTINUED
State
Northern States
Indiana
Michigan
Nebraska
New York
Wisconsin
Adapted from Terman
Location Sit
in state
Gary
Others 1
Various 5
Various 7
Ithaca
Industrial Site
Urban
Rural 1
(1978). See origir
es
1
0
1
1
9
3
al
Years
1946-47
1946-47
1959-60
1953-54
1931-49
1969-71
1969-71
1969-71
for data
Major source
Industry
General
Industry
General
Urban &
Industry
Industry
Urban
General
sources.
Average
kg S ha-1 yr-1
142.2
30.0
11.3
7.2
54.9
168.0
42.0
16.0
2-27
-------
TABLE 2-4. SULFUR CONTENT OF CROPS
Crops
Yield Total S Content
tons ha-1 kg ha"1
Grain and oil crops
Barley (Hordeum vulgare L.)
Corn (Zea Mays L.)
Grain sorghum (Sorghum bicolor L. Moench)
uats (Avena satlva 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 dactylon 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.)
Burley
Flue-cured
Fruit, sugar, and vegetable crops
Beets
Sugar (Beta saccharifera)
Table (Beta vulgaris L.)
Cabbage (Brassica oleracea)
Irish potatoes (Solanum tuverosum L.)
uranges (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.
Terman (1978).
Adapted from
2-28
-------
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 rain can cause increases as well as decreases
in forest productivity (Abrahamsen 1980b, Cowling and Dochinger 1980).
The net effect of acid rain 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 $042-
and N03" are already in the ecosystem and that H+ is generated
naturally by the plant community (111rich 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
S042~ deposition in an S-deficient ecosystem? In short, the problem
of assessing the effects of acid rain 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
depressed soil pH. It is important in hydrology because cations leached
from soils may be deposited in 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 rain can
accelerate cation leaching rates, but the magnitudes of these increases
must be evaluated within the context of natural, internal 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, Soil ins et al.
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.
2-29
409-262 0-83-3
-------
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
Thompson,
Washington
Soiling,
W. Germany
Jadrass,
Sweden
Findley,
Washington
H.J. Andrews,
Oregon
Precipitation
Species Age H+
(yr) input3
Pseudotsuga 42
menziesii
Fagus sylvatica 59
Pinus sylvestris
Abies amabilis, 175
Tsuga mertensiana
Pseudotsuga 450
menziesii
240<*
(4.8)
9009
1909
90h
(5.6)
289
Cation
Cation removal by
leaching by harvesting0
internal acid
production^5 Bole WTH
(eq ha"1 yr"1)
420<* 3806
(5.9)
19509 2209
2269
1410" 2726
(4.5)
227009 606
6606
3706
4606
1066
aWeighted average [H+] times precipitation amount; weighted average [H+] as
pH appears in parenthesis where available.
bCalculated from net increase in weighted average HC03~ 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.
CNutrient content divided by age; WTH = whole tree harvest, removal of all
aboveground biomass.
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 H?C03
production values are included.
2-30
-------
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 HgCOs production reported by Andersson et al. (1980)
exceed atmospheric H+ inputs as measured by open-bucket collectors.
In this case, the comparison is misleading, however, since dry
deposition to the forest canopy at Soiling is known to be exceedingly
high (Ulrich et al. 1980), and, consequently, 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 inconsistent 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
considerably from sulfate in that it is very poorly adsorbed to most
soils (Johnson and Cole 1977). However, biological processes in
N-l invited ecosystems quickly immobilize nitrate, and since N limitations
are common in forested regions of the world, nitrate is rarely mobile
(Abrahamsen 1980b). 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 1951, 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 conservative 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
2-31
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(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 S04Z"
adsorption capacity. In these situations continual input of moderate
amounts of H2$04 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
deposition 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~l are sufficient to satisfy S requirements
in most forest ecosystems (Humphreys et al. 1975, Evans et al. 1981,
Johnson et al. 1982). Inputs of S042- in acid rain affected regions
frequently exceed this value (often by a factor of 2 to 4), implying
that Sis currently being deposited in excess of forest requirements
(Table 2.3).
Several studies have shown that excess S cycles within vegetation
and accumulates in soils as $042- 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 $042-
component of acidic deposition. Similarly, further emphasis ought to be
given to the effects of N inputs, since they appear to be increasing
(Abrahamsen 1980b) and N is commonly the limiting nutrient in forest
ecosystems.
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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 N03~ immobilization on the leaching potential
of HN03 is the same as that of S042~ immobilization on the
leaching potential of H2$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
(111rich et al. 1980).
If atmospheric N inputs increase to the point where N deficiencies
are alleviated 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 (111 rich et al. 1980). Thus, nitrogen
"saturation" of forest ecosystems could result in significant increases
in cation leaching and, under extreme circumstances, 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 widespread, imminent
nitrogen saturation of forests since N deficiencies are still quite
common and most ecosystems are still accumulating N (Abrahamsen 1980,
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 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 availability and
nitrate leaching were stimulated by H2S04 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.
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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 Section 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 rain 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, Foy
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; Foy 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, Graham 1978, Pritchett and Smith 1972).
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
sufficient 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 frequently 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) .
Aluminum and Mn toxicities are the most prominent growth-limiting
factors in many, if not most, acidic soils (Foy 1973, 1974b, 1981;
2-34
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Tanaka and Hayakawa 1974). 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 Al uminum toxi city. 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 £-1) beneath the beech stand
and ~ tenfold (from 1-2 mg £-1 to 15-18 mg £-1) 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 process at Soiling is nitrification in mineral soil
layers during warm, dry years. Nitrification during these periods
(thought to be caused by decomposition of previously accumulated N-rich
root residues) causes a pulse of acid production. According to Ulrich
et al. (1980), systems that have been acidified 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~l;
comparatively, H+ input values of this magnitude are not uncommon in
forests of the United States (Evans et al. 1981). However, at Soiling
2-35
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H+ flux in throughfall is 2 to 5 times greater than in open
precipitation due to dry deposition in the forest canopy.
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 H2$04 ranging from current
background levels (approximately 0.8 keq ha-1 yr-1) up to
approximately 30 times that amount (26 keq ha-1 yr-1). 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 (Abrahamsen
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 review 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 capacity, root
phosphatase activity, internal concentrations of Si, NH4+ - NOs"
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
2-36
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gradually by weathering, by recycling by deep rooted species, and by
dust inputs if the acidic inputs are reduced. Since 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.
Since microbial activity in soils responds quickly to changing
environments, important soil processes it moderates can be expected to
return 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 practices 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 Section 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 or semi-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 deposition.
Various schemes for assessing site sensitivity to acidic deposition
effects have been proposed. Those directed toward aquatic effects have
emphasized bedrock geology (Hendry et al. 1980, Norton 1980), while
those concerned with terrestrial effects have emphasized cation exchange
capacity and base saturation (McFee 1980, Klopatek 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 unfortunately, the data base for the latter
is limited. In considering soil sensitivity to adverse effects of
acidic deposition, it is helpful to separate the effects into two
2-37
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Figure 2-4. Generalized soil map of the United States (SCS USDA, 1975) showing
regions dominated by suborders or groups of suborders. The most common
suborder is named. Many other suborders exist within the boundaries of
each area.
Al f i sol s Mol 1 i sol s
Al Aqualfs Ml Aquolls
A2 Boralfs M2 Borolls
A3 Udalfs M3 Udolls
A4 Ustalfs M4 Ustolls
A5 Xeralfs M5 Xerolls
Aridisols Spodosols
Dl Argids SI Aquods
02 Orthids S2 Orthods
^ Entisols Ultisols
i El Aquents Ul Aquults
oo E2 Orthents U2 Humults
E3 Psamments U3 Udults
Histosols Vertisols
HI Hemists VI Uderts
H2 Hemists and Saprists V2 Usterts
H3 Fibrists, Hemists, and Saprists
Inceptisols
II Andepts
12 Aquepts
13 Ochrepts
14 Umbrepts
-------
U. S DEPARTMENT OFAGRICUITURE
GENERAL SOIL MAP OF THE UNITED STATES
SOIL CONSERVATION SERVICE
ro
I
OJ
E3c Vok
0 100 7uO 300 '00 500 600 M.k
.H3a
Obi' .uv
-------
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-1, those subject to frequent flooding, or
those with free carbonates in the upper 25 cm of the sol urn to be
insensitive. Non-calcareous, non-alluvial 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, 1980b) 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 saturation, and the presence of carbonates and
ferromagnesium minerals. Considering 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
sensitivity criteria proposed by McFee (1980), Wiklander (1980b), and
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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 meet criteria 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 01 sen 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 criteria 2 and do
not currently receive significant acidic deposition.
2.3.5.2.2 Changes in aluminum or other metal 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 concentration in soil
solution or leachate from the soil. This is most likely to occur where
total quantity of the controlling Al compounds exposed to chemical
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 characteristics. Many members of this group, such as
earthworms, mites, ants, and termites are involved in mixing the organic
2-41
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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 microorganisms 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 Algae--Chlorophyta (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 pioneering species, frequently
supplying the organic material required for subsequent colonization by
other life forms. Some blue-green algae (bacteria) can convert
atmospheric N2 to organic compounds. In many environments, such as
flooded paddy fields, this ability to fix nitrogen provides a critical
input 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 constituents, 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 effectively draw nutrients. In deficient soils, the
fungal partner can substantially 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 investigation.
2.4.1.4 Bacteria--The procaryotic microflora of soils are also
extremely important in the decomposition of plant litter and the
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synthesis and breakdown 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 NH4+ to N03~ (nitrification) is dominantly catalyzed by
autotrophic soil bacteria. Nitrogen is lost from the soil through
anaerobic bacterial reduction of N(h- to the gaseous species N2
and N20 (denitrification). Most nitrogen enters ecosystems through
bacterial reduction of atmospheric N£ to NHA+ (No-fixation).
Fixation by bacteria living symbiotically with plants can contribute
significant amounts of nitrogen to both agricultural and forest systems.
Nitrogen nutrition of many leguminous plants is enhanced through
N2~-fixation by bacteria of the genus Rhizobium. Fixation by
actinomycetes, such as Frankia. in association 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 availability 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 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 deposition. 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.
2.4.2.1 Soil AnimalIs—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,
high populations of other soil animals, such as springtails (Collembola)
2-43
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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~l H2S04 for 6 years,
showed that the numbers of Collembola increased, Enchytraeidae
decreased, but mites (Acarina) were generally unaffected by both
application rates. In a two-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 Telatively 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 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 No-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, however, have used traditional plate count methods, which do not
2-44
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necessarily reflect viable fungal biomass. Baath et al. (1980) found
that FDA (fluorescein 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 pollution, 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 researchers suggested that cation carriers of
mycorrhizal roots may be more susceptible 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 (Butlin 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 H2$04, although the magnitude of these effects was not reported.
In soils transferred to a site receiving pH 3.0 rain and dry deposition,
Wainwright (1980) found that over a one-year period bacterial numbers
did not change significantly, 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 H2S04 inputs for six 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 C02 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
2-45
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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 decomposition are summarized in Section 2.5.
Several reports now indicate that acid inputs can slightly
accelerate mineralization 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 more retarded than mineralization (a
hypothesis for which no substantiating data existed). Conversely,
Francis et al. (1980) found lower NH4+ production 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, however, are compatible
with the fact that "natural" soil acidity does not have a uniform effect
on N-mineralization (Alexander 1980b).
Since nitrification is generally believed to be catalyzed by
relatively autotrophic nitrifiers (known to be acid-sensitive on labora-
tory 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 N0o~ 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 ^$04, 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 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. Since 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 (Remade 1977,
Johnsrud 1978).
2-46
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Little published data concern effects of acidic deposition on soil
denitrification. While slight soil acidification may not alter the
overall rate of this process, it should be expected to increase NgO
production relative to N£ (Firestone et al. 1980).
A substantial amount of work on the sensitivity of N2-fi*ation 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 No-
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 mobilization of Al and Mn from mineral forms indigenous to the
soil. Can mobilization 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 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 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 N2-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.
2-47
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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 demonstrate 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 electronegativity 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, Calluna 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 availability 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
deposition on microbial activity in terrestrial ecosystems is limited,
extrapolations 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.
2.4.5 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
2-48
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composition of simulated rain, as well as differences in important soil
characteristics, makes comparing data difficult. 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 shortcomings of
long-duration experiments involving infrequent sampling should also be
recognized. Acid rain 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
soilsare not significant. In less acid soils, pH 3.0 simulated
rain significant changes in litter animals.
o 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 No-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.
o 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
2-49
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forest soils will result in decreased decomposition rates for organic
matter. This hypothesis 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; Cronan 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; and
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
following 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 response 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 biologically-mediated
mineralization of organic matter in forest soils will be
only slightly inhibited by acidic deposition (< 1 to 2
percentdecrease 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., N. M.
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.
2-50
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TABLE 2-6. REVIEW OF STUDIES CONCERNED WITH THE IMPACT OF ACIDIC DEPOSITION ON ORGANIC DECOMPOSITION
Author
Soil Type
Duration
of
Experiment
Treatments
Results
1. Abrahamsen et al. 1980 Lodgepole pine needles 75-90 days
Norway spruce needles
3-9 mos.
ro
i
en
Raw coniferous humus • unspecified
2. Abrahamsen and Dollard 1978 General Review
3. Abrahamsen et al. 1976
Lodgepole pine needles 90 days
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
nun mo.
-1
Raw humus in litterbags ex-
posed to pH 5.3, 4.3, and
3.5 treatments.
Needles moistened with di-
lute H2$04 solutions.
Cellulose/Wood
Unspecified
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.~l. At
200 mm mo."1, 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.
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
3.2 acid rain treatments and
were incubated with C^
glucose.
pH 4.1 treatment had no
effect on glucose
mi neralization
pH 3.2 treatnent decreased
glucose mineralization rate
by 30-66%.
5. Alexander 1980b
Spodosols from the 14-61
central Adirondacks days
Soils were exposed to 100 cm of
pH 3.5 and 5.6 artificial rain
for 14 consecutive or 35
intermittent days.
ro
tn
ro
In 14 day consecutive rain,
the rate of total organic
carbon (TOC) leaching was
initially greater at pH 3.5
than 5.6. This later re-
versed.
In 35 day intermittent treat-
ment: pH 3.5 leached more
TOC than pH 5.6.
6. Baath et al. 1980.
Coniferous iron Podzol
12 mo.
7. Cronan 1980a.
Coniferous forest floor 4 mo.
Litterbags were placed in
field plots exposed to
H2S04 treatments P 50
and 150 kg ha'1.
Forest floor microcosms were
exposed to pH 5.7, 4.0 and
3.5 artificial rains.
C02 evolution response variecf
with soil pH -- inhibition in
more acid soils, but
stimulation by pH 3.5 rain in
less acid soil.
No significant difference com-
pared to controls for Scots
pine needle litter.
Root litter exposed to 150 kg
ha~l had 21* decrease in
decomposition rate.
Increased rainfall acidity
caused increased leaching of
Ca, Mq, K, and NH4 +.
Compared to the pH 4
treatment, the pH 3.5 rain
caused 50-150% more K, Ca, and
Mg leaching.
-------
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
IX)
i
en
CO
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 leachinq and 65%
more Mq leachinq at pH 4.0.
Coniferous forest floors
showed 40% more Ca and 25%
more Mg 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
cellulose 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'l.
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
Author
Duration
Soil Type of Treatments
Experiment
Results
ro
i
en
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 Porizol
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 respirtion.
Litterbaqs showed significant
increase in weight loss (15%)
with increased acidity.
Found decreased C02
respiration with increased
H2S04.
-------
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 Rosenqvist1s hypothesis by applying NaOH to one of
the mini-catchment watersheds 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 deposition.
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+. 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 Al3+ concentrations, regardless of
whether the anion is introduced as a salt or an acid. This mechanism
has been known to soil scientists for decades as the "salt effect,"
wherein soil pH is usually more acid in CaCl2 solutions than in H20
(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 practices 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 (Drabltfs 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
2-55
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Brook, NH ecosystem in which H+ in acid rain is initially neutralized
by dissolution of reactive 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.
Kilham (1982) reports a case in which deposition appears to have
caused an increase in lake alkalinity. Alkalinity in Weber Lake,
Michigan, has increased twofold 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
hydro!ogical 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, metal mobility, and
rnicrobial activity. The following conclusions, relative to these
general categories, can be drawn from Chapter E-2:
2-56
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Soils amended in agricultural practice will not be harmed by
acidic deposition (Section 2.3.5).
Soil acidification is a natural process in humid regions.
It is obvious that acidic deposition contributes to this
process; however, at current levels, it is a minor
contribution (Section 2.3.5).
Most soils that were easily acidified are already acid;
therefore, soils likely to become perceptibly more acid due
to deposition are limited. They are the soils that have low
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
HN03 inputs generally will be beneficial. Exceptions may
occur on sites with adequate or excessive 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).
The most likely damage to forest productivity, and the one for
which some evidence exists, would result from Al toxicity.
This may occur 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
2-57
-------
dissolution of Al- and Mn-containing clay minerals, any H+
input will increase solution of Al and Mn concentration
(Section 2.3.3).
The increased mobility of Al in uncultivated, acid soils is
probably the most significant effect of acidic deposition on
soils as they influence terrestrial plant growth and aquatic
systems (Section 2.3.3).
Based upon shorter term studies, we can expect that increased
H+ loading will generally cause increased loss of cation and
organic components from forest litter. Over the longer term,
it appears that the biologically-mediated mineralization of
organic matter in forest soils will be only slightly inhibited
by acidic deposition (< 1 to 2 percent decrease in
decomposition rate). In general, experimental data suggest
that decomposition processes are relatively unaffected by
simulated precipitation pH's above 3.0. Thus, unless average
precipitation inputs were to drop to pH 3.0 or below,
significant impacts of acidic deposition or 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 availability. Evidence for effects of acidic
deposition Rhizpbium or actinomycete is reduced by artificial
acid inputs, but no evidence exists that current rates common
in the United States will cause such a decrease. Slight
decreases and increases in N mineralization rates result from
short-term acid inputs, but long-term responses are not
documented. Possible effects of acidic deposition as soil
microbial activity in natural systems have not been ruled out,
but important effects under field conditions have not been
clearly demonstrated (Section 2.4).
2-58
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS
E-3. EFFECTS ON VEGETATION
3.1 INTRODUCTION
3.1.1 Overview
This chapter examines diverse plant-pollutant relationships to
assess potential 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.
Since the close relationship between soils and plants bears
examination in terms of how soil acidification affects productivity. It
is important to recall from the previous chapter the following points:
o soils amended in agricultural practice will not likely be
negatively 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
o soils with low buffering capacity, relatively high pH, low
sulfate adsorption capacity, no carbonates, and no basic inputs
are susceptible to increased acidification rates from
atmospheric inputs of acidic and acidifying substances.
Keeping these points in mind, Chapter E-3 will deal more with the
direct effects of acidic deposition on plant response, and to
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, however, this document does not attempt to make economic
assessments of recognized 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
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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 manuscript, "Air and Rain: The Beginnings of a Chemical
Climatology," included 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 Nedbj6rs 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 and a number of research programs to
investigate the effects were initiated in the mid-1970's.
Anthropogenic and natural air contaminants are usually inventoried
on a separate basis (e.g., chemical speciation) when information is
sought as to sources, dispersion, or induced effects (see Chapters A-2
and A-5). Categorically, the National Ambient Air Quality Standards
(NAAQS) for criteria pollutants (ozone, sulfur oxides, hydrocarbons,
nitrogen dioxides, carbon monoxide, 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, 1982). As distances from
pollutant sources increase, chances for combinations to occur also
increase, or, as in the case of large metropolitan/industrial areas,
pollutant combinations are the rule rather than the exception.
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 isopaths of acidic precipitation must also be
3-2
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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.
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.
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 dispersion), 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 pollutants
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 pollutants (see Chapter A-2). Comparable conditions
also appear to exist in coastal California, where severe air stagnation
has led to very high levels of phdtochemical 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 to
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 dispersion have
combined to produce the highest oxidant concentrations in the United
States and widespread mortality and decline of forests in the nearby San
Bernadino Mountains (Miller et al. 1977).
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1-BOREAL FOREST ECOSYSTEM
J-UAICE STATES FOREST ECOSYSTEM
3-EASTERN DECIDUOUS FOREST ECOSYSTEM
4-SOUTH EASTERN PINE FOREST ECOSYSTEM
6-TROPICAL FOREST ECOSYfEM
6-WESTERN MONTANE FOREST ECOSYSTEM
7-SUBALPINE FOREST ECOSYSTEM
8-PACIFIC COAST FOREST ECOSYSTEM
9-CALIFORNIA 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. Isopleths are shown in
relation to major forest types of the United States
(adapted from Miller and McBride 1975} and in relation to
mean annual hydrogen ion (kg ha'1 yr-1) deposition in
precipitation (adapted from Henderson et al. 1981).
3-4
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3.2 PLANT RESPONSE TO ACIDIC DEPOSITION
3.2.1 Leaf Response to Addle 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
are 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 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 proportion 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 deposited 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 experi-
mental evidence with simulated rain, a wide range of plant species 1s
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 (Haines et al. 1980). These results suggest that
generalizations about sensitivity to Injury may be difficult, 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) influences
the foliage wettability; 2) makes the foliage more vulnerable to injury
(e.g., through differential permeability of the cuticle); or 3) retains
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 1on, 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,
3-5
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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 surface (e.g., roughness,
pubescence, waxiness) or the chemical composition 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 penetrate 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 micropores 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 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 epidermal 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
penetration 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 physical 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
3-6
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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 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 Populus sp. exposed to
simulated acidic precipitation experienced abnormal cell proliferation
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,
3-7
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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 precipitation
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 precipitation 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 photosynthesis were significantly increased at pH 2.0.
Ferenbaugh concluded that because growth of the plants was significantly
reduced, photophosphorylation was uncoupled by the treatments. Irving
(1979) reported increased photosynthetic rates in some soybean
treatments, attributing them to increased 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 nutrients 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 absorption. Throughfall can be defined as
that portion of the gross, or incident, precipitation that reaches the
3-8
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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 variety of forest species (Tamm 1951, Madgwick and Ovington 1959,
Nihlgard 1970, Patterson 1975, Lindberg and Harriss 1981). This
enrichment has three potential sources: 1) reactions on the leaf
surface in which cations on exchange sites of the cuticle are exchanged
with hydrogen from rainfall; 2) movement of cations directly from the
translocation stream within the leaf 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 atmospheric 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 ($042- or NOa") 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 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 1n the forest canopy
or 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
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
3-9
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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 (washoff) 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 enrichment ranged from 13
to 100 percent, or from 0.3 to 14.4 kg ha~l yr-1. Parker et al.
concluded that for temperate hardwood forests in industrialized regions,
40 to 60 percent of annual net throughfall (throughfall enrichment) for
sul fate 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). Phaseplus yulgaris 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
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. Si gal)
The objective of this section is to review the literature on the
effects of acidic deposition on lichens and mosses and also to review
the literature that describes the effects of realistic, low levels of
gaseous sulfur dioxide (S02) on lower plants (Grennfelt et al. 1980).
Several researchers (Skye 1968, Turk and Wirth 1975) have concluded that
S02 toxicity and pH effects are not independent factors.
3-10
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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,
southeastern
Alaska
Tropical rain forest,
Costa Rica
Douglas fir,
Reference
Cronan 1978
Feller 1977
Haughbotn 1973
Haughbotn 1973
Haughbotn 1973
Heinrichs and Mayer
1977
Heinrichs and Mayer
1977
Johnson 1975
Johnson 1975
Johnson 1975
S deposition
Incident
24.4
11. Qa
32. 3b
17.7
10.0
24. ld
24.1
0
12.5
4.0
kg ha"1 yr"1
Throughfall
46.4
40.0
111.2
69.1
21.1
47.6
80.0
16.4
23.3
5.2
Precipitation
amount
(cm)
203d
245C
77
77
77
106
106
270
390
165
Washington
-------
TABLE 3-1. CONTINUED
Forest system
Subalpine silver fir,
Washington
Hardwoods,
Amazonian Venezuela
Hardwoods,
Amazonian Venezuela
Hard beech,
New Zealand
CO
jL Beech,
^ Southern Sweden
Spruce,
Southern Sweden
Oak,
Southern France
Loblolly pine,
North Carolina
Chestnut oak,
Tennessee
Mixed oak, Tennessee
Mixed oak, Tennessee
Reference
Johnson 1975
Jordan et al. 1980
Jordan et al . 1980
Miller 1963
Nihlgard 1970
Nihlgard 1970
Rapp 1973
Wells et al. 1975
Lindberg et al . 1979
Kelly 1979
Kelly 1979
S deposition
Incident
16. 8f
44.5
46.6
8.4
7.9<1
7.9"
16.4
7.93
13.2b»e
8.73
11.3a»b
kg ha"1 yr'1
Through fall
5.3
16.7
19.6
10.4
18.5
54.2
22.6
9.9
32.0
15.0
14.0
Precipitation
amount
(cm)
300
391
412
135
95
95
NA
NA
143
154
75
-------
TABLE 3-1. CONTINUED
aScaled up from a subannual estimate.
In vicinity of factory or power plant.
°Mean of extreme estimates.
Includes stem flow.
eSeveral years data.
fLittle throughfall.
CO
I
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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 bioindicators 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 nirate (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'corroborated 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 solutions reduce photosynthesis (Ing!is
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 S02 concentration of 130 yg nr3 (Ferguson et al. 1978).
It has been suggested that sulfate at "feasible" atmospheric
concentrations has no effects upon photosynthesis in mosses; however,
the fall in pH that accompanies the oxidation of atmospheric SOg to
504 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 S02 injury .is greatest when lichens are in a moistened or saturated
3-14
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condition (Rao 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 buffering
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 Galvin 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 plant when
the annual S02 average exceeded 50 yg nr3. Will-Wolf (1980) found
that Parmelia caperata and P. bolliana showed morphological alterations
in areas where maximum S02 Tevels were 389 yg m-3, 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 m-3 for 96 days, and plasmolysis of algal cells in both
U^ hirta and Parmelia chlorochroa after 31 days of S02 at the same
concentration!Le Blanc and Rao (1975) concluded that long-range
3-15
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average concentrations for S02 between 16 to 79 yg m-3 (0.006 to
0.03 ppm) cause chronic injury to epiphytes.
In the Ohio River Valley, maximum annual averages of $03 ranged
from about 50 to 80 yg m-3 in 1977 and 1978. Maximum 1-hr averages
ranged from 300 to 500 yg m-3 (Mueller et al. 1980). At the same
sites (Rockport and Duncan Falls), mean rainfall pH for August 1978 to
March 1980 were 4.12 and 4.36, with ranges of 3.17 to 5.64 and 3.24 to
6.03, 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 photosyn-
thesis was reduced by 40 percent in the lichen Cladina stellaris by
field fumigations with fluctuating S02 concentrations of less than 655
vg m-3 (0.25 ppm; Moser et al. 1980). Laboratory exposures of the
same lichen species wetted by artifical precipitation having a pH = 4.0
and a sulfate concentration = 10.00 mg £-1 reduced photosynthesis by
27 percent (Lechowicz 1981). 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 (1981). 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 L_^
pregana 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 1980, Rhoades 1981), in 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, Plummer 1980) is
3-16
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inconclusive because separating pH effects from potential ambient SO?
(or other gaseous pollutant) toxicity is impossible. 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
lichens 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 thallus and substrate are Important for the survival and
regeneration of lichens 1n polluted areas because pH and buffer capacity
control the distribution and proportions of toxic compounds in solution
and the rates of breakdown of 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 precipitation. Conclusive linkage to decreased photosynthetic
rates 1s currently 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 extreme 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 of H+ to soils, but they do not prevent cation
losses from the ecosystem. If the anion associated with acidic
3-17
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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+ concentration 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
deposition 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
precipitation and gaseous air pollutants on terrestrial vegetatation
consists of only three separate studies. Shriner (1978b) examined the
intreaction 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 reduction 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
combination 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.
3-18
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ACID DEPOSITION
INSECT
FOLIAGE
FEEDER
MICROBE
FOLIAR
PATHOGEN
MICROBE
STEM PATHOGEN
INSECT
BARK BEETLE
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.
3-19
t09-262 0-83-5
-------
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 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 exposed 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 additional 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).
After more than a decade of research in Scandinavia, the observed
changes in forest soil chemical properties that can be attributed to
acidic precipitation 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 combination 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, 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 mecham'sm(s) involved, increased dissolution of
3-20
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gaseous sulfur dioxide will occur 1n 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.
3.3.2 Interactions with Phytophagus 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 regulating 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, mutation rates, dispersal, fecundity, mate finding, host finding,
and mortality. Indirect effects may occur through changes in host age
structure, distribution, 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
(Exotelela 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 (S1erp1nsk1 1967, Novaskova 1969, Lebrun 1976)
environments. Further specific information 1s available on the general
influence of polluted atmospheres on population characteristics of
forest insects (Tempiin 1962; Schnaider and Sierpinski 1967; Sierplnski
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,
3-21
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under appropriate environmental conditions, to impact by a heterogeneous
group of microblal pathogens including vlroids, viruses, mycoplasmas,
bacteria, fungi, and nematodes. As with insect interactions, microbes
and the diseases they cause play important roles in succession, species
composition, density, composition, and productivity. In the short term,
the effects of microbial pathogens may conflict with forest management
objectives and assume a considerable economic or managerial as well as
ecologlc significance (Smith 1970).
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 summarized 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 1n a variety of nonwoody and woody hosts.
Treshow (1975) has provided a detailed review concerning the influence
of sulfur dioxide, ozone, fluoride, and partlculates 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
m1crob1ally-1nduced 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
3-22
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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 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 Cronartiurn fusiforme (fungus)
on willow oak, (2) a 66 percent inhibition of Meloidogyne ha^Ta
(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 phaseolicola (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 a§e 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 Gremmeniena abietina has increased in
importance in the northeastern United StatelTThe disease, termed
Scleroderris canker, was first reported on red pine in New York in 1959.
Currently, G. 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
3-23
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America, Paul D. Manlon, SUNY, Syracuse, Initiated an acidic rain
Scleroderrls 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
1s, acidic rain may reduce the Importance of the canker disease (Raynal
et al. 1980, Bragg 1982, Manlon and Bragg 1982).
Armillarlella me!1ea is an extremely important forest tree root
pathogen throughout the temperate zone. The fungus 1s 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 Indicated that soil pH is related to disease development
caused by A. mellea. If acidic deposition influences soil pH (see
Chapter E-7) or tree vigor, it may indirectly impact tree susceptibility
to A_. mellea Infection. In the northeast, spruce decline 1n high
elevation forests has been a recent concern. A_. 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.
roellea are buffered relative to plant-surface habitats, so for acidTc
deposition to Influence these pathogens an alteration of soil pH or
chemistry or host susceptibility would have to occur.
Fusiform rust caused by £. fuslforme 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 basidlospores. 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 tree leaf
microfloras. Lacy et al. (1981) observed that populations of Erwinia
herbicol a and Pseudomonas syringae were reduced on soybean 1 eaves 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 Hofjts That Would Alter Relationships with
Insect or Microbial Associate (W. H. Smith)
As Section 3.2 discussed, exposure to acidic deposition may lead to
acidification of plant surfaces, leaf cuticle erosion, and foliar
lesions. Foliar lesions could release plant volatHes 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, amino acids,
3-24
-------
vitamins). Leaching of toxic radioelements 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 differential 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 Individuals. Trees
exhibiting vigorous growth, on the other hand, may be predisposed 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 systems. The majority of these materials are organic chemicals
that selectively control unwanted and injurious Insects, pathogens, or
weeds. 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 fate of pesticides in the environment depend upon the
following:
(1) method of application of the pesticide;
(2) chemical properties of the pesticide;
(3) edaphic properties of the system;
(4) dissipation routes of the pesticide; and
(5) climatic conditions.
No studies on effects of acidic deposition on pesticides were found
in the literature; however, pH changes have been reported to affect
factors 1 through 5 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 phenoxyacetic acid (Crafts 1961b) and dinitrophenol (Crafts and
Reiber 1945) type compounds. Acidic conditions promote formation of the
un-ion1zed 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-till age crop
production systems, or aquatic systems where the foliage has had ample
time to accumulate acidic deposition.
Chapter E-4 (Section 4.4) reports that acidic deposition causes
significant lowering of the pH of inland waters in certain instances.
3-25
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This would have a substantial effect on the direct biological activity
and longevity of herbicides used 1n aquatic weed and algae control. One
would expect a significant Increase 1n the herblddal activity of the
phenoxyacetlc add compounds. Aquatic herbicides such as slmazlne would
perform less satisfactorily under acidic conditions since many
Investigators (Armstrong et al. 1967, Jordan et al. 1972) have reported
that chloro-s-tr1az1nes decompose at a much faster rate under acidic
conditions. ~T"h1s would make 1t necessary to Increase the rates of
chloro-j;-tr1az1ne herbicides and to make more frequent applications for
satisfactory aquatic weed control 1n waters where the pH levels were
below normal levels.
Organic pesticides are categorized Into five major types depending
on Ionizing characteristics. Examples of the five types are:
(1) catlonlc (dlquat, paraquat);
(2) basic (atrazlne, slmazlne, prometryn);
(3) acidic (2,4-D, fenac, plcloram)
(4) phosphates and arsenates (glyphosate, DSMA); and
(5) nonlonlc (alachlor, carbaryl, methomyl).
Changes In pH levels In waters or soil solutions affect the
Ionizing properties of basic and acidic properties to the greatest
extent. At lowered pH levels acidic and basic pesticides tend to be
more readily adsorbed by soil parti oil ate matter, hence less
biologically 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
mlcroblal populations. Captan, dlcamba, amltrole, vernolate,
chloramben, crotoxyphos (Hamaker 1972), metrlbuzln (Ladlle 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, dlazlnon and dlazoxon (Hamaker 1972)
were degraded more readily at lower pH levels.
Pesticides degraded chemically are directly affected by soil pH
levels. Malathlon and parathlon (Edwards 1972) persisted much longer In
acidic soils than 1n neutral soils, while atrazlne (Best and Weber 1974)
and slmazlne were degraded much more rapidly under acidic conditions
than under neutral conditions.
3.3.6 Summary (W. H. Smith)
A review of the evidence on the Interaction of acidic deposition
with other pollutants, Insect and mlcroblal pests, does not allow
3-26
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generalized statements concerning stimulation or restriction of blotlc
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 pollutants,
although the potential for such effects 1s very real.
Future research must combine field and controlled environment
studies. Mechanisms for addle deposition Impact on predisposition/
protection of forest trees to/from disease caused by mlcroblal
pathogens, and Infestation 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 mlcroblal pathogens.
No known studies demonstrate that acidic deposition on plant
surfaces directly affects the biological 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
inactivation, 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, and A. H. Johnson)
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 receiving 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
succession and the sensitivity of forest trees to natural environmental
stresses add further to the challenge of quantifying effects of
anthropogenic pollutants on forest productivity.
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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.
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 beyond 1975. The
most recently published estimates of historical changes in isopleths of
precipitation acidity (Likens and Butler 1981) suggest that initial
intensification of acidity of northeastern precipitation 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 documentation 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 hypothesized 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
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WET
DEPOSITION
ACID DEPOSITION
-ACID RAIN (WET)
-POLUTANT GASES (DRY)
i
1DIRECT EFFECTSI
GROWTH
VIGOR
REPRODUCTION
I
THROUGHFALL
I
I
{INDIRECT EFFECTSI
NUTRIENT AVAILABILITY
TOXIC EFFECTS
31
MICRQBIAL PROCESSES
NUTRIFICATION
DENITRIFICATION
IMMOBILIZATION
RELEASE
MYCORRHIZAE
I FOREST |
PRODUCTIVITY
Figure 3-3. Key components and processes to be considered in evaluating
effects of acidic deposition on forested ecosystems.
3-29
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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 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 substances in a manner which 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 assessment of the influences of acidic deposition on plants
must include consideration 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 deposition 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, including marketable crops, 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, development, 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 possible 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 performance, plant dieback or reduction in
growth or yield must be evaluated in terms of physiological stress, soil
toxicity and nutrient deficiency problems, plant disease, and direct and
indirect effects of acidic precipitation, if chronic effects of
deposition of acidic substances are to be fully characterized.
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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 contrasting response of seed germination of three tree
species to different substrate acidity levels.
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, Abougendia and Redman 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, Monitor 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 sensitive 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 conditions
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 1980) 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
3-31
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g
GERHINAT]
•j*
100
90
80
70
60
50
40
30
20
10
0
• * *
(a) ,fi|ft
:
*
.* SUGAR MAPLE
B
a*
I A pH 5.6
| o pH 4.0
I • pH 3.0
••' 1 1 1 1 1
10 20 30 40
DAYS SINCE START OF GERMINATION
50
3
M
70
60
50
40
30
20
10
WHITE PINE
J I
pH 5.6
pH 4.0
,..., pH 3.0
J_
I
2 4 6 8 10 12 14 16
DAYS SINCE START OF GERMINATION
5 10 15 20
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 acidiy: (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).
3-32
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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 affecting growth is not fully
known (Abrahamsen 1980). When nutrients in soil are not limiting for
plant growth, detrimental effects of deposition of acidifying substances
are not likely to occur. 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 acidic deposition, however,
foliar injury may simultaneously occur in some species (Raynal et al.
1980, 1982b). Thus, competitive promotive and inhibitory effects of
acidic deposition may concomitantly 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 variable results, ranging from no apparent
effects, stimulation of growth, and inhibition of growth (Wood and
Bormann 1974, Matiziris 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, effects on 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 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.
Controlled-environment studies of effects of simulated acidic
precipitation on agricultural crops indicate differential sensitivity of
species and contrasting effects on different plant parts (Evans and
Lewin 1981, Lee et al. 1981; see also Section 3.4.2). Research by Evans
and Lewin (1981) indicates that yield of pinto beans may be reduced by
simulated acid rain because of a decrease in the number of seeds per
pod. In contrast, under similar exposure conditions, soybean yield
showed an increase due to a larger dry mass per seed.
3.4.1.3 Growth of Seedlings and Trees in Irrigation Experiments--
Abrahamsen (1982) has reviewed field experiments in Sweden and Norway
designed to determine the effects of artificial acidification on growth
3-33
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of forest trees and tree seedlings. In Swedish experiments (Tamrn and
Wlklander 1980), young (18-yr-old) Scots pine were Irrigated below the
canopy with dilute sulfurlc add (annual application, 50 to 150 kg
ha"1 HeSOd In one application per year) 1n 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 highst acidity) whereas
growth responded positively (approximately + 30 percent increase at
highest acidity level) on the unfertilized plots. Increased nitrogen
uptake was considered a probable cause of positive responses. Results
of these studies were complicated by changes 1n nutrient availability in
the soil and associated with the effects of high acidity on soil fungi,
bacteria, and competing understory vegetation (Tamm and Wlklander 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 ^$04 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 experiments 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 1n the 5th year.
Norway spruce showed reduced diameter growth at all acid treatment
levels on the year after 6 years of prior treatment. Height growth of
silver birch was stimulated by rainfall acidity. Lime application had
little or no effect on observed responses. Effects of add irrigation
on foliar nutrient levels were also generally small (Tveite 1980b).
In evaluating the results of the Scandinavian Irrigation
experiments Abrahamsen (1980) concluded that the data give "no
substantial evidence of effects on tree growth at acidity levels
presently found In precipitation." However, he cautions that add
effects produced particularly at highest acidity levels may be partly
attributable to soil effects that were artifacts 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 1n 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
3-34
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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 in Norway
spruce and Scots pine a decrease in growth (generally less than 2.3
percent per year) 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 1n 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 pronounced 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 distribution 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 between susceptible and non-susceptible regions were
estimated to be in the range of 0.3 to 0.6 percent per year.
Since this original study, a second study has been initiated
covering an additional 9 years, 1965-1974, 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. For sites of
relatively poor quality, growth of both pine and spruce in the 1970's
had increased substantially since Its minimum in the mld-60's but was
still substantially less than that attained up to 1940. The overall
trend was still downward over the Interval 1910-1974. 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
mathematical model that was used has not compensated for or
caught those effects in the ground that are the results of more
3-35
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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 forest!and and its fertility."
The final report on the Norwegian SNSF project makes the point
that:
"decreases in forest 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 regional acidification at levels
currently 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
terrestrial 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 S02 in Sweden increased 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 interal 1955 to 1964, but then remained constant
through 1974. By contrast, nitrate levels increased by 33 percent (15
to 20 meq 2,"1) from 1955 to 1964 and by 1974 had reached 35 meq
jf1, 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 most
recent changes in growth patterns.
In the United States, Cogbill (1976) examined growth of beech,
birch, and maple in the White Mountains of New Hampshire and red spruce
3-36
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1n the Smoky Mountains of Tennessee. From analysis of tree-ring
chronologies, he concluded 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) 1n 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 Ulrlch et al. (1980)
have reported dieback and decline of Norway spruce and beech and 1n
inland areas of the eastern United States, contributions of gaseous
pollutants, primarily 03 and $03, can be considered to have changed
over the same time spans as has add precipitation and thus should be
considered 1n 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 deposition are those at high elevations in the northeast.
Forests characterized by varying proportions of spruce, fir, and white
birch occur at the high elevations 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
3-37
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receive extraordinarily high rates of H+, sulfur, and heavy metal
deposition (Lovett et al. 1982, Fried!and et al. 1983). In addition,
the vegetation is subjected to very acid cloud moisture for a
considerable portion of the year (Johnson et al. 1983). 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 quantitative 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 (Betula papyrifera
var. cordifolia) in the canopy, mountain ash (Pyrus americana) and
mountain maple (Acer spicatum) as important understory trees, and an
herb layer dominated by ferns (Dry op ten's sp.) and pxalis montana
(Siccama 1974). The lowermost elevation to which 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 vegetation is
believed to be related to the incidence of cloud moisture, with the
boreal vegetation occupying the often cloud-capped upper slopes, and
hardwoods 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. 1983).
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 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 frequently and
for a total of 30 to 80 days per year.
In the mountainous areas of New England, precipitation increases
with altitude. Lovett (1981, in Cronan 1983) estimates precipitation
rates of 240 cm yr-1 in the balsam fir forests of New Hampshire.
Low-elevation precipitation in New England ranges from about 100 to 150
cm yr-1. Siccama (1974) determined that growing season throughfall
increased by 2.9 cm per 100 m~2 in the Green Mountains of Vermont due
3-38
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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 (1983) estimated H+ input to the canopy at 77 to 100 meq
m-2 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+ m-2 during this period. Based on Cronan1s 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 estimate of H+ deposition has been
made by Lovett et al. (1982), who used field collection of cloud
moisture samples and modeling of cloud droplet interception to estimate
H+ deposition in the subalpine zone of the White Mountains to be ~
340 meq m-2 yr-1.
As a result of the substantial input and the inferred low
neutralization capacity of the canopy (Cronan 1983), the potential for
accelerated leaching of bases is high, but to date, no quantitative data
from high elevation forests indicate if 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. (1983) 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 observed in New York and Mew Hampshire (Johnson et al.
1983). 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
3-39
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spruce Indicate the presence of secondary fungal pathogens, with
Arm-Mi aria me! 1 ea, Fomes pini, and Cytospora kunzi'1 most prominent
(Hadfleld 1968). Hadfleld (1968) speculated 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 Manlon (1981), the spruce decline has
the characteristics of a complex blotlc-ablotlc disease related to
environmental stress. Currently, there are no data which Implicate
acidic deposition 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 dleback of spruce (Plcea ables) and fir (Abies
alba) Is under study 1n Germany. From long-term, Intensive, ecosystem-
Tevel studies, UlHch (Ulrlch et al. 1980; Ulrlch 1981a,b, 1982)
suggested that acidic deposition has contributed to changes 1n H+
generation and consumption which have caused soil acidification,
mobilization of Al, mortality of fine roots, and ultimately, dleback and
decline 1n spruce, fir, and beech (Fagus sylvatica). That contention 1s
based on careful documentation of changes In soil solution chemistry, a
nearly parallel decrease In fine root blomass and Increase 1n soil
solution Al concentrations during the growing season, and nutrient
solution studies which Indicated that the ratio of uncomplexed Al
(I.e., Al3+) to Ca found In the soil solution was sufficient to cause
abnormal root growth and development. While those findings suggest the
possibility of Al toxldty, 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 toxlclty or Ca deficiency In those cases. More recently,
however, 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 reponsible (see Section 3.2.1.2).
Rehfuess points out that the parallel change in soil solution Al and
fine root blomass noted by Ulrlch was not synchronized 1n that marked
decreases in fine root blomass preceded the increase in soil solution
Al. Rehfuess cites several studies (Goettsche 1972, Deans 1979, Persson
1980) in support of his contention that late simmer declines 1n fine
root blomass 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, since 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). Data on spruce root
chemistry from Camels Hump, VT, indicate that CarAl ratios increase with
increasing elevation. As mortality increases with elevation, it 1s
3-40
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not likely that imbalances of Al and Ca in root tissue are the major
cause of spruce decline (Lord 1982, Johnson et al. 1983).
Whether the red spruce decline is related to acidic deposition has
been the focus of considerable speculation. The deline is widespread,
easily discerned, 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 intensely 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, where
the trees are probably predisposed to stress by the site conditions
where 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 those possibilities warrant research in the future.
3.4.1.6 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
U.S. and in some areas of Europe with high atmospheric deposition
levels. Recent increases in mortality of red spruce in the northeastern
U.S. 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 of immediate alarm yet are difficult
to interpret because of potential artifacts of experimental protocols.
Detecting responses of mature forest trees is made difficult by the
complexities 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
3-41
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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 are ever conclusively established or refuted.
3.4.2 Crops (P. M. Irving)
A considerable number of studies on the vegetative effects of
acidic precipitation have been published in the last 5 years. However,
because of 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
approach 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. Since 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 characteristics; (3) determination of
mechanisms of action. Both field and controlled-environment
methodologies with simulated rain have been used in these approaches.
3-42
-------
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 from add rain exposure Include
dose-response studies to mathematically relate yield to pollutant dose.
The term 'dose-response' suggests a unlvarlate relationship; however, a
number of potentially Important variables comprise 'acid rain dose1 (see
next section). Complex, factorial designs and multlvarlate 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) control!ed-env1ronment 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 1f 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 1n 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 differences
caused by environmental variability or a large number of treatment 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.
A lack of comparable unpolluted (control) plots 1s 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 1n
the Kona district of Hawaii during a volcanic eruption period. The
plants growing outside the ralnshelter 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 non-reactive 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 1t occurs In conjunction with wet deposition 1s one
facing all field researchers.
Controlled-envlronment studies are useful Indicators of potential
effects and may suggest subtle changes not measureable 1n an
3-43
-------
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 (1981) 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.
Pollutants rarely occur alone, and since 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 (Heggestadt and Heck 1971; see also Section
3.4.1.2). Gaseous pollutants (i.e., ozone, sulfur dioxide) have been
found to affect certain crop cultivars 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 plant characteristics that can be used to indicate a plant's
3-44
-------
relative sensitivity or resistance to acidic deposition. For example,
leaf wettabillty, which 1s related to surface morphology, has been
suggested as a parameter that may Indicate sensitivity to acidic
precipitation (Evans et al. 1977a).
It has been suggested that crop classes can be grouped according to
their sensitivity to add precipitation. Based on a study of 28
different crops, Lee et al. (1981) reported that Inhibition of
marketable yield was observed only 1n the dicotyledons that were
studied, and within this group root crops, leaf crops, cole crops, tuber
crops, legumes and fruit crops were ranked 1n 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 having the objective of determining
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 photosynthesis, 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
J^cog assimilation technique In the field, reported that S02
exposures reduced both photosynthesis and yield of soybeans but that
acid rain treatments had apparently stimulated the photosynthetlc rates
with no effect on soybean yield. Usually physiological determinations
alone are 1n adequate to estimate the economic damage of pollutants to
crops.
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 affecting plant response than the pH of
the rain (Irving and Sowlnski 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
3-45
-------
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 individual 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] +
[$042- x cm] + [N03~ x cm]. Response to a given dose of gaseous
pollutant is frequently 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 SCtyZ-,
N03-, and H+), rate or intensity (i.e., cm hr-1), duration, and
frequency. Physiological systems can be quite resilient due to
activation 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 (1981) 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 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 pre-
cipitation 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
3-46
-------
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 1n
units were made only for comparison purposes. None of the experimental
results described below have been changed from those of the orglnal
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 1n Table 3-2 were
performed in the field, using accepted agricultural practices to the
extent experimental design would permit. Because hydrogen, sulfate, and
nitrate ions are those components of precipitation that are believed to
most likely affect the growth and yield of crops, they were used in
describing the precipitation dose. In all experiments, simulated rain
was applied at regular intervals during the Hfe 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 cultlvars (9 species) studied, only one exhibited
a consistently negative yield effect at all acidity levels used (garden
beet), three were negatively affected by at least one of the acidity
levels used 1n the study ('So. Giant Curled' mustard green, 'Pioneer
3992' field corn, and 'Amsoy1 soybean), and six had higher yields from
at least one acidity level ('Champion' 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 Curled' mustard
green, 'Improved Thick Leaf spinach, and 'Vernal' alfalfa). Some
experiments demonstrated both positive and negative response to acid
rain, depending on the H+ concentratipn. There 1s little evidence for
a linear response function, however, since 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 decrease 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:nitrate ions 1n the precipitation simulant also
affected the response of some plants (i.e., alfalfa, fescue, mustard
green; Table 3-2), independent of pH.
A comparison of studies on five different cultlvars of soybeans by
four different investigators appears to indicate that the 'Amsoy'
cultivar may be more susceptible to acidic deposition than 'Beeson',
3-47
-------
TABLE 3-2. FIELD RESEARCH ON CROP GROWTH AND YIELD AS AFFECTED BY ACID PRECIPITATION
GO
CO
Total deposition
kg ha"1
(simulant ••• ambient)
H+ S042- N03-
Simulant concentration
mg l"1
H+ S042- N03- S(
Rate
cm hr'
J4^~:N03~
Events
1 f
hr
events-1
Droplet
size
PH
Effect*
Alfalfa, 'Vernal', Hedlcago saliva L. (Lee and Neely 1980)
0.017
0.171
0.833
2.611
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»2o
2.26
2.26
2.26
0.30
2.26
7.89
19.64
60.85
0.30
0.'OQ25 0.53 0.753
0.10 4.83 0.753
0.316 14.67 0.753
1.00 46.19 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
'Perfected Detroit V-9041, Beta vulgarls
'Pioneer 3992' Zea
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
4T7F
17.33
43.54
135.47
0.39
O.OOZ 1.26 3TD4
0.010 5.47 3.04
0.079 37.07 3.04
1.995 106.6 3.04
0.087
mays L. (Lee and Neely 1980)
JJ.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
0.7
6.4
19.5
61.3
2.5
0.7
1.1
1.5
1.5
2.5
L. (Evans
0.4
1.8
12.2
35.1
0.7
1.1
1.5
1.5
2.5
'Alta1, 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
12.99
5.57
5.57
5.37
0.0025 0.53 0.753
0.10 4.83 0.753
0.316 14.67 0.753
1.00 46.19 0.753
0.016 1.07 0.434
0.0025 0.30 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
0.7
6.4
19.5
61.3
2.5
0.7
1.1
1.5
1.5
0.67
0.67
0.67
0.67
0.67
0.67
0.67
0.67
et al
35.0
35.0
35.0
35.0
0.67
0.67
0.67
0.67
(Lee
0.67
0.67
0.67
0.67
0.67
0.67
0.67
0.67
26
26
26
26
26
26
26
26
. 1982)
19
19
19
19
58
58
58
58
and Neely
26
26
26
26
26
26
26
26
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
0.001
0.001
0.001
0.001
1.5
1.5
1.5
1.5
1980)
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1200
1200
1200
1200
1200
1200
1200
1200
353
353
353
353
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
5.6
4.0
3.5
3.0
4.8
5.6
4.0
3.5
3.0
4.8
5.7
4.0
3.1
2.7
4.1
5.6
4.0
3.5
3.0
4.8
5.6
4.0
3.5
3.0
4.8
5.6
4.0
3.5
3.0
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
101 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
Control
9% lower yield; no effect on growth compared
pH 5.6
No effect on growth or yield compared to pH 5
No effect on growth or yield compared to pH 5
Ambient
Control
24% greater yield than pH 5.6
19% greater yield than 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
to
.6
.6
, Phaseolus vulgarls L. (Shrlner and Johnston 1981)
0.001 O.OZ 0.12
0.2
0.631 50.0 0.12 417
2.30 0.95
2.4
3.0
3.0
27
27
0.17
0.17
900
900
6.0
3.2
Control
No effect on growth or yield compared to pH 6
Ambient
.0
aEffects are reported when statistical significance 1s < 0.05 level.
-------
TABLE 3-2. CONTINUED
co
Total deposition Simulant
concentration
kg ha'1 mg ir1
(simulant + ambient)
H+ S042- N03- H+ S042-
Hustard
0.033
0.189
0.535
1.629
0.029
0.033
0.189
0.535
1.629
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' , Brassica
2.78 1.95 0.0025
9.66 1.95 0.10
25.40 1.95 0.316
75.83 1.95 1.00
1.93 0.78 0.016
2.78 1.95 0.0025
7.05 5,45 0.10
20.20 12.68 0.316
56.33 38.04 1.00
1.93 0.78 0.016
'Champion', Raphanus satlvls
0.0025
0.06
0.32
1.585
0.17
0.0025
O.OS
0.32
1.58
0.16
0.53
4.83
14.67
46.19
1.07
0.53
3.20
11.42
34.00
1.07
N03-
Rate
cm hr'l
S042~:N03~
Events Droplet
1 hr
events-1
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
L. (Trolano et
0.72 0.31
2.9
11.7
55.6
0.72
2.90
11.70
55.60
'Cherry Belle', Raphanus satlvus L.
0.0025
0.10
0.316
1.00
0.026
0.0025
0.10
0.316
1.00
0.026
0.53
4.83
14.67
46.17
0.96
0.53
3.20
11.42
34.00
0.96
'Cherry Belle', Raphanus satlvus L.
0.002
0.010
0.079
1.995
0.087
1.26
5.47
37.07
106.6
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 0.7
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
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
size
Mm
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
pH
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
Effect*
Control
No effect on growth or yield compared to pH 5.
No effect on growth or yield compared to pH 5.
No effect on growth or yield compared to pH 5.
Ambient
6
6
6
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 5% 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
Ambient
•12% lower root wt (yield), 71 higher shoot wt
than ambient
31 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
25% greater yield than pH 5.6
No effect on growth or yield compared to pH 5
Ambient
Control
No effect on growth or yield compared to pH 5
No effect on growth or yield compared to pH 5
No effect on growth or yield compared to pH 5
Ambient
No effect on growth or yield compared to pH
4.06 (ambient)
No effect on growth or yield compared to pH 5
No effect on growth or yield compared to pH 5
No effect on growth or yield compared to pH 5
Ambient
.6
.6
.6
.6
.6
.7
.7
.7
aEffects are reported when statistical significance Is £ 0.05 level.
-------
TABLE 3-2. CONTINUED
tn
O
Total deposition Simulant concentration
kg ha-1
(simulant + ambient)
H+ SO*2' N03- H*
Soybean. 'Amsoy',
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
,b 'Bee son
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 max (L.)
0.10
0.794
1.995
10.0
0.79
' , Glyclne max (L
-------
'Davis1, 'Williams1 or 'Wells'; however, the experimental conditions
such as soil type and characteristics of the rain simulant 1n the
'Amsoy' study were markedly different than the other studies and may be
responsible for the observed effect. Figure 3-5 Indicates the location
and results of the four soybean field studies 1n relation to the
principal production regions and soil types. The one cultlvar that
responded negatively to add rain treatments ('Amsoy') was grown 1n an
area with a sandy soil, unlike the other studies. The simulated rain
used in this 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. 1977) that were not present in the rain simulants used by
other Investigators. The 'Beeson' and 'Williams' cultivars, which were
studied in a location near the 'Amsoy1 study, but with more structured
soil, reponded 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 the response because of differences 1n major soil types, cultivars
grown, climatic conditions, and ozone concentrations. Perhaps of equal
Importance is the presence of toxic heavy metals which become more
soluble with Increases 1n acidity of simulated rain.
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 differences in total deposition of H+,
$042-, N03~, or $042- to NOa- ratios.
Experimental results from some of these studies also demonstrate
that the response 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 and
whether the average concentration or total deposition of H+, $042-,
N03~ is important in determining yield response.
3.4.2.2.2 Controlled environment studies. As with the field studies,
experimental 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 m-2 (Table
3-3) by 10 to obtain kg ha-1 (Table 3-2). A comparison of effects on
the same species grown 1n a controlled environment as opposed to in the
field indicates a similar response 1n most species (alfalfa, spinach,
mustard green, soybean) although radishes exhibited a negative effect in
a controlled environment and a positive effect In the field. In
3-51
409-262 0-83-6
-------
SOYBEANS
Crop yield - kg ha~! (harvested)
1978 Census of Agriculture
0 - 1500
1500 - 2000
> 2000
oo
en
ro
ANL-Argonne National Laboratory
Soil: silt loam 'Martinton1
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. 1981d
BTI-Boyce Thompson Institute
Soil: sandy loam
Cultivar: 'Beeson1, 'Williams'
Acidity Effect: Positive
Troiano et al. 1983
NCS-North Carolina State University
Soil: sandy clay loam 'Appling'
Cultivar: 'Davis'
Acidity Effect: None
Heagle et al. 1983
NCS
Figure 3-5.
Location of four soybean field studies, indicating
production regions and soil types.
-------
TABLE 3-3. CONTROLLED ENVIRONMENT STUDIES ON CROP GROWTH AND YIELD AS AFFECTED BY ACID PRECIPITATION
01
oo
Total deposition
Simulant concentration
g m"z ( i events)
H* S042- N03- H+
Alfalfa,
0.001
0.056
0.177
0.56
Barley,
0.001
0.045
0.142
0.45
'Vernal', Mod 1c ago
0.300 0.417
2.744 0.417
17.35 0.417
54.92 0.417
'Steptoe', Hordeum
0.238 0.335
2.205 0.335
13.945 0.335
44.13 0.335
satlva
0.0025
0.10
0.316
1.00
vulgare
0.0025
0.10
0.316
1.00
mg r1
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
S042-:N03-
Rate
cm hr-1
Events Droplet Fertilizer
1
hr ,
events"1
size
vm
N-P-K
PH
Effect*
1981, Lee et al. 1981)
0.7
6.6
41.8
132.5
0.67
0.67
0.67
0.67
56
56
56
56
1.5
1.5
1.5
1.5
1200
1200
1200
1200
67-252-252b
67-252-252
67-252-252
67-252-252
5.6
4.0
3.5
3.0
Control
No effect market yield, increased
shoot wt
31% greater market yield, Increased
shoot/root wt
No effect growth or market yield
1981, Lee et al. 1981)
0.7
6.6
41.8
132.5
Beet, 'Detroit Dark Red'. Beta vulgarls L. (Cohen et al. 1981.
0.001
0.026
0.082
0.26
0.140 0.193
1.274 0.193
8.057 0.193
25.50 0.193
Bibb lettuce, 'Limestone',
0.0002
0.009
0.028
0.09
0.048 0.067
0.441 0.067
2.789 0.067
8.826 0.067
Bluegrass, 'Newport', Ppa p
0.002
0.072
0.227
0.720
Broccol 1
0.0006
0.022
0.070
0.22
0.382 0.472
3.528 0.472
22.31 0.472
70.61 0.472
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
iratensls L. (Cohen et al
0.0025
0.10
0.316
1.00
, 'Italian Green Sprouting
0.117 0.164
1.078 0.164
6.319 0.164
21.58 0.164
0.0025
0.10
0.316
1.00
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
0.7
6.6
41.8
132.5
et al . 1981
0.7
6.6
41.8
132.5
. 1981. Lee
0.7
6.6
41.8
132.5
', Brassica 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
0.67
0.67
45
45
45
45
1.5
1.5
1.5
1.5
1200
1200
1200
1200
112-224-224b
112-224-224
112-224-224
112-224-224
5.6
4.0
3.5
3.0
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
, Lee et al.
0.67
0.67
0.67
0.67
et al. 1981)
0.67
0.67
0.67
0.67
Botrytls 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
, Lee
1200
1200
1200
1200
112-224-2240
112-224-224
112-224-224
112-224-224
112-224-2240
112-224-224
112-224-224
112-224-224
224-448-443b
224-448-448
224-448-448
224-448-448
et al. 1981)
168-224-2240
168-224-224
168-224-224
168-224-224
5.6
4.0
3.5
3.0
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
Control
No effect market yield or growth
No effect market yield or growth
25% lower market yield
'Effects are reported when statistical
OFerttltzer as kg ha"1 of N-P»05-K20.
fertilizer as percentage of N-PjOs-l^f
significance Is < 0.05 level.
-------
TABLE 3-3. CONTINUED
CO
-fi
Total deposition Simulant concentration
g m-z ( r events)
H* 304?- NOa- H*
mg l-1
S042- HOa- $04*':NO
Bush bean, 'Blue Lake 274', Phaseolus vulgarls L.
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 0.0025
0.008 0.001 0.10
0.001 0.001 0.0025
0.007 0.001 0.10
0.041 0.001 0.631
0.60
5.33
0.60
5.33
30.70
'Golden Acre', Brassfca oleracea
0.270 0.379 0.0025
2.499 0.379 0.10
15.80 0.379 0.316
50.02 0.379 1.00
0.53
4.90
30.99
98.07
0.83
0.70
0.83
0.70
0.75
L. var.
0.74
0.74
0.74
0.74
Rate
cm hr-1
3"
1
Events
hr ,
events"1
Droplet Fertilizer
size N-P-K
ym
PH
Effect*
(Johnston et al. 1982)
0.7
7.6
.7
7.6
40.9
Capltata L.
0.7
6.6
41.8
132.5
'Danvers Half Long', Daucus carota L. var. Sativa DC
0.230 0.327 0.0025
2.156 0.327 0.10
13.636 0.327 0.316
43.15 0.327 1.00
0.53
4.90
30.99
98.07
-0.74
0.74
0.74
0.74
Cauliflower, 'Early Snowball', Brasslca oleracea
0.0006
0.023
0.073
0.23
0.122 0.171 0.0025
1.127 0.171 0.10
7.128 0.171 0.316
22.56 0.171 1.00
Corn, 'Golden Midget', Zea mays L.
0.0005
0.020
0.063
0.20
Fescue,
0.001
0.059
0.186
0.59
0.11 0.149 0.0025
0.980 0.149 0.10
6.198 0.149 0.316
19.61 0.149 1.00
'Alta', Festuca elatlor L.
0.31 0.439 0.0025
2.891 0.439 0.10
18.20 0.439 0.316
57.86 0.439 1.00
0.53
4.90
30.99
98.07
(Cohen
0.53
4.90
30.99
98.07
0.74
0.74
0.74
0.74
et al.
0.74
0.74
0.74
0.74
0.7
6.6
41.8
132.5
L. var. Botr
0.7
6.6
41.8
132.5
1981, Lee et
0.7
6.6
41.8
132.5
.64
.64
.64
.64
.64
(Cohen et al
0.67
0.67
0.67
0.67
(Cohen et al.
0.67
0.67
0.67
0.67
18
18
16
16
16
. 1981,
51
51
51
51
1981,
44
44
44
44
ytls L. (Cohen et al
0.67
0.67
0.67
0.67
al. 1981)
0.67
0.67
0.67
0.67
var. arundlnacea Schreb. (Cohen et al.
0.53
4.90
30.99
98.07
0.74
0.74
0.74
0.74
0.7
6.6
41.8
132.5
0.67
0.67
0.67
0.67
23
23
23
23
20
20
20
20
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
900
900
900
900
900
. 1981)
1200
1200
1200
1200
1981)
1200
1200
1200
1200
. 1981, Lee et al.
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1981, Lee et al.
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
0-20-OC
0-20-0
0-20-OC
0-20-0
0-20-0
224-224-224b
224-224-224
224-224-224
224-224-224
224-224-224b
224-224-224
224-224-224
224-224-224
1981)
224-224-224D
224-224-224
224-224-224
224-224-224
168-336-336b
168-336-336
168-336-336
168-336-336
168-336-336l>
168-336-336
168-336-336
168-336-336
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
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 yield; older leaves age
-------
TABLE 3-3. CONTINUED
Total deposition
g m~z ( i events)
H* S042- N03-
Simulant concentration
H*
Green pea, "Marvel '. PI sum satlvum
0.001
0.028
0.088
0.28
0.150 0.208
1.372 0.208
8.677 0.208
27.46 0.208
Green pepper, 'California
0.001
0.038
0.128
0.380
0.20 0.283
1.86 0.283
11.78 0.283
37.27 0.283
Kidney bean. 'Red Kidney'.
0.0004
Y* 0.029
Oi
CJi 0.095
0.105
0.107
0.134
0.200
0.229
Lettuce.
0.00096
0.03024
0.03024
0.03024
Mustard
0.0004
0.014
0.044
0.14
0.007 0.043
2.274 0.043
7.564 0.043
8.32 0.043
9.831 0.043
10.59 0.043
15.88 0.043
18.14 0.043
0.0025
0.10
0.316
1.00
Wonder' ,
0.0025
0.10
0.316
1.00
mg fl
S042- N03-
S042-:N03-
Rate
cm hi-1
Events
1
Droplet
hr , size
events*1 u»
Fertilizer
N-P-K
PH
Effect*
L. (Cohen et al, 1981, Lee et al. 1981)
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
Capsicum annuum L.
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
0.7
6.6
41.8
132.5
(Cohen et
0.7
6.6
41.8
132.5
0.67
0.67
0.67
0.67
al. 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
l.S
. 1981)
1.5
1.5
1.5
1.5
1200
1200
1200
1200
1200
1200
1200
1200
67-224-2240
67-224-224
67-224-224
67-224-224
224-448-448*
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
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
Phaseolus vulgaMs 1. (Shrlner 1978a)
0.001
*
*
*
*
*
*
0.631
'Oakland'. Lactuca satlva
0.02304 0.02976
0.13824 1.7856
0.78384 0.95232
1.38336 0.11904
green, 'So. Giant
0.074 0.104
0.687 0.104
4.339 0.104
13.73 0.104
0.002
0.63
0.63
0.63
Curled.'
0.0025
0.10
0.316
1.00
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
50 0.12
0.2/416
0.2/416
0.2/416
0.2/416
0.2/416
0.2/416
0.2/416
416.67
3.0
3.0
3.0
3.0
3.0
3.0
3.0
3.0
24
24
24
24
24
24
24
24
0.17
0.17
0.17
0.17
0.17
0.17
0.17
0.17
900
900
900
900
900
900
900
900
6.0/3.
3.2/6.
6.0/6.
3.2/3.
6.0/3.
3.2/6.
6.0
2/6 .Od
0/6.0
0/3.2
2/6.0
2/3.2
,0/3.2
3.2
Control
751 Increased pod number; greater
shoot and root wt
SOT lower pod number; greater
shoot wt
501 lower pod number; greater
shoot wt
No effect pod numher; lower shoot/
root wt
751 greater pod number; greater
root wt
No effect pod number; lower shoot/
root wt
50% greater pod number; lower
shoot wt; greater root wt
L. (Jacobson et al. 1980)
0.48 0.62
2.88 37.20
16.33 19.84
28.82 2.48
Brasslca japonlca
0.53 0.74
4.90 0.74
30.99 0.74
98.07 0.74
0.77
0.08
0.82
11.6
O.RO
0.80
0.80
0.80
Hort. (Cohen et al .
0.7
6.6
41.8
132.5
0.67
0.67
0.67
0.67
3
3
3
3
1981, Lee
14
14
14
14
2.0
2.0
2.0
2.0
et al.
1.5
1.5
1.5
1.5
900
900
900
900
1981)
1200
1200
1200
1200
Half-strength 5.7
Hoaglands
112-224-2240
112-224-224
112-224-224
112-224-224
3.2
3.2
3.2
5.6
4.0
3.5
3.0
Control
No effect growth or yield
7% Increase root wt; 24% Increase
apical leaf wt
10% Increase root wt; 29% Increase
apical leaf wt
Control
14% lower market yield
No significant effect
31% lower market yield
*0.001/0.631.
'Effects are reported when statistical significance Is £ 0.05 level.
dpH sequence Is: 10 events prior to'Halo blight Infection/3 events during Infection period/11 events post Infection.
-------
TABLE 3-3. CONTINUED
CO
9l
Total deposition
g «rz ( i events)
H* »42- H03-
Simulant concentration
mgt-
1
N03-
»,'-,
Rate
cm hr-
Events
1 » hr i
events"1
Droplet Fertilizer pH
size N-P-K
Effect*
Oats. 'Cay use', Avena satlva L. (Cohen et al. 1981, Lee et al. 1981)
0.001
0.048
0.152
0.48
Onion, '
0.002
0.065
0.205
0.65
0.254 0.357
2.354 0.357
14.87 0.357
47.07 0.357
Sweet Spanish'.
0.34 0.484
3.185 0.484
20.14 0.484
63.75 0.484
Orcnardgrass, 'Potomac'
0.001
0.035
0.111
0.35
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
Allllum cepa L. (Cohen et al.
0.0025 0.53
0.10 4.90
0.316 30.99
1.00 98.07
, Dactyl Is glomerata
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
Pinto bean, 'Univ. Idaho 111'. Phaseolus vulgar Is L.
0.003
1.355
2.149
3.405
5.396
Potato,
m
0.164
0.52
Radish.
0.0003
0.012
0.033
0.12
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', Solanum tuberosu* L.
litt 8-J887-
16.11 0.387
51.00 0.387
'Cherry Belle',
0.064 0.089
0.588 0.089
3.719 0.089
11.77 0.089
Red clover, 'Kenland' ,
0.001
0.056
0.177
0.56
0.300 0.417
2.744 0.417
17.35 0.417
54.92 0.417
8:?825 l-M
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
Trl folium pratense L
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
Sift
0.74
0.74
0.7
6.6
41.8
132.5
1981,
0.7
6.6
41.8
132.5
et al
0.7
6.6
41.8
132.5
(Evans
6.2
46.9
74.8
119.0
189.1
0.67
0.67
0.67
0.67
Lee et al.
0.67
0.67
0.67
0.67
. 1981, Lee
0.67
0.67
0.67
0.67
and Lewln
0.72
0.72
0.72
0.72
0.72
al. 1981, Lee et
u
41.8
132.5
8:l77
0.67
0.67
48
48
48
48
1981)
65
65
65
65
et al. 1981)
35
35
35
35
1981)
45
45
45
45
45
al. 1981)
if
52
52
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
0.33
0.33
0.33
0.33
0.33
hi
1.5
1.5
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
1200
ass
353
353
353
353
ifoo
1200
1200
112-224-2240
112-224-224
112-224-224
112-224-224
336-336-3360
336-336-336
336-336-336
336-336-336
112-224-2240
112-224-224
112-224-224
112-224-224
manure and
1 Imestone
added
247-224-2240
224-224-224
224-224-224
224-224-224
5.6
4.0
3.5
3.0
5.6
4.0
3.5
3.0
5.6
4.0
3.5
3.0
5.7
3.1
2.9
2.7
2.5
5.6
4.0
3.5
3.0
Control
No effect market yield or growth
No effect market yield; Increased
root growth
No effect market yield or growth
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
281 lower seed yield
29% lower seed yield
39% lower seed yield
Control
No effect yield; Increased shoot
growth
lit greater market yield; Increased
shoot growth
8% lower market yield
(Cohen et al. 1981, Lee et al. 1981)
0.74
0.74
0.74
0.74
0.7
6.6
41.8
132.5
. (Cohen et al.
0.74
0.74
0.74
0.74
0.7
6.6
41.8
132.5
0.67
0.67
0.67
0.67
12
12
12
12
1.5
1.5
1.5
1.5
1200
1200
1200
1200
112-224-224»
112-224-224
112-224-224
112-224-224
5.6
4.0
3.5
3.0
Control
No effect growth or market yield
Lower market yield
Lower market yield; decreased shoot
growth
1981, Lee et al. 1981)
0.67
0.67
0.67
0.67
56
56
56
56
1.5
1.5
1.5
1.5
1200
1200
1200
1200
67-336-3360
67-336-336
67-336-336
67-336-336
5.6
4.0
3.5
3.0
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 Is £0.05 level.
'"Fertilizer as kg ha-1 of N-P
-------
TABLE 3-3. CONTINUED
co
Total
g m-2
H*
deposition Simulant concentration
( i events) mg i-l
5042- N03- H+ S042-
Ryegrass, "L1nn", Loll urn 'perenne L.
0.001
0.055
0.183
0.58
Spinach,
0.0004
0.014
0.044
0.14
Soybean,
0.004
1.549
6.169
Soybean ,
0.002
0.002
0.002
0.105
0.105
0.105
0.700
0.700
0.700
0.31 0.432 0.0025
2.842 0.432 0.10
17.97 0.432 0.316
56.88 0.432 1.00
' Improved
0.074
0.687
4.339
13.73
'Amsoy 71
9.755
73.20
295.12
'Wells',
0.669
0.980
1.113
3.738
2.485
5.600
20.55
25.52
29.40
c
N03- S042":N03~
(Cohen et al. 1981, Lee et al
0.53
4.90
30.99
98.07
0.74'
0.74
0.74
0.74
Thick Leaf, Sp1nac1a oleracea L.
0.104 0.0025
0.104 0.10
0.104 0.316
0.104 1.00
', Glydne max (L
1.561 0.002
1.561 0.794
1.561 3.162
Glydne max (L.)
0.721 0.0025
0.490 0.0025
0.371 0.0025
3.731 0.15
5.530 0.15
3.619 0.15
18.55 1.0
12.82 1.0
9.800 1.0
0.53
4.90
30.99
98.07
.) Merr.
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, 'Qulnalt', Fragarla chlloensls
0.002
0.080
0.253
0.800
0.42
3.920
24.79
78.46
0.595 0.0025
0.595 0.10
0.595 0.316
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
Rate
:m hr-1
. 1981)
0.67
0.67
0.67
0.67
Events Droplet Fertilizer
1
58
58
58
58
hr .
events"1
1.5
1.5
1.5
1.5
size
1200
1200
1200
1200
N-P-K
112-224-224b
112-224-224
112-224-224
112-224-224
pH Effect*
•
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; decrease!
root growth
(Cohen et al. 1981, Lee et al. 1981)
0.7
6.6
41.8
132.5
al. 19810)
6.2
46.9
189.1
0.67
0.67
0.67
0.67
0.72
0.72
0.72
14
14
14
14
78
78
78
1.5
1.5
1.5
1.5
0.17
0.17
0.17
1200
1200
1200
1200
353
353
353
112-224-224°
112-224-224
112-224-224
112-224-224
manure and
1 Imestone
added
5.6 Control
4.0 No effect growth or yield
3.5 No effect'growth or yield
3.0 No effect growth or yield
5.7 Control
3.1 11% greater seed yield; decreased
shoot growth
2.5 11% lower seed yield; decreased
shoot growth
Sowlnskl 1980)
1.0
2.0
3.0
1.0
2.0
1.5
1.0
2.0
3.0
21.2
21.2
21.2
21.2
21.2
21.2
21.2
21.2
21.2
Duchesne var. ananassa (Cohen
0.74
0.74
0.74
0.74
.7
6.6
41.8
132.5
0.67
0.67
0.67
0.67
10
10
10
10
10
10
10
10
10
et al.
80
80
80
80
0.33
0.33
0.33
0.33
0.33
0.33
0.33
0.33
0.33
1981, Lee et
1.5
1.5
1.5
1.5
2300
2300
2300
2300
2300
2300
2300
2300
2300
al.
1200
1200
1200
1200
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
1981)
224-336-3360
224-336-336
224-336-336
224-336-336
5.6 1:1 S04:N03; control
5.6 2:1 S04:N03; control
5.6 3:1 S04:N03; control
3.8 No effect growth or yield compared
to 1:1 control
3.8 No effect growth or yield compared
to 2:1 control
3.8 Lower root nodule wt compared to 3:1
control; no effect yield
3.0 No effect growth or yield compared
to control
3.0 No effect growth or yield compared
to control
3.0 25% greater yield than 1:1 control,
19% greater than 2:1 control
5.6 Control
4.0 51% greater market yield; Increased
shoot growth
3.5 72. greater market yield; Increased
shoot/root growth
3.0 72% greater market yield; Increased
shoot/root growth
'Effects are reported when statistical significance Is £0.05 level.
fertilizer as kg ha"1 of N-P
-------
TABLE 3-3. CONTINUED
Total deposition Simulant concentration
g m-2 ( £ events) mg
H+ S042- M03- H* S042-
l-l
N03-
Swiss chard, 'Lucullus', Beta vulgar Is var. clcla
0.001
0.032
0.101
0.32
Timothy
0.001
0.033
0.104
0.33
V Tobacco
en
CD o.OOl
0.024
0.076
0.24
Tom to,
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
0.238 0.0025 0.53
0.238 0.10 4.90
0.238 0.316 30.99
0.238 1.00 98.07
0.74
0.74
0.74
0.74
, Phleum pratense L. (Cohen et al
0.256 0.0025 0.53
0.256 0.10 4.90
0.256 0.316 30.99
0.256 1.00 98.07
21' . Nlcotlana tabacun L
0.179 0.0025 0.53
0.179 0.10 4.90
0.179 0.316 30.99
0.179 1.00 98.07
Lycoperslcon esculentum
0.379 0.0025 0.53
0.379 0.10 4.90
0.379 0.316 30.99
0.379 1.00 98.07
0.74
0.74
0.74
0.74
. (Cohen
0.74
0.74
0.74
0.74
S042":MO
Rate
cm hr-1
3~
L. (Cohen et al. 1981, Lee
0.7
6.6
4V. 8
132.5
. 1981, Lee
0.7
6.6
41.8
132.5
et al. 1981
0.7
6.6
41.8
132.5
Mill. (Cohen et al.
0.74
0.74
0.74
0.74
'F1el
-------
general, total deposition of H+, $042-, and N0s~ applied was
greater 1n 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 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,
'Limestone' 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 Discuss1on--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 1n experimental design or exposure conditions. For example,
in all field studies except those of 'Champion* radish and 'Beeson* and
'Williams1 soybeans, the ratio of sulfate to nitrate in the rain
simulant differed among treatments and was usually much higher than the
sulfaternitrate 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., an
range of pH 3.7 to 6.8 for New York) while the 50*2- to N03~
ration appears to be Independent of pH (Figure 3-6). Because
preliminary evidence Indicates that plants are affected by the
sulfate:nitrate ratio 1n rain (Irving and Sowlnski 1980, Lee et al.
1980), the differences reported among treatments in these Investigations
may be the result of this ratio rather than the hydrogen 1on 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.
(1981) 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.
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 response. The 9 percent reduction in the yield of
field corn exposed to pH 4.0 rain (0.594 kg ha-1 depositon of H+) 1s
an alarming result; however, treatments with greater acidity levels
3-59
-------
13
12 -
11 -
10
9
8
co 7
o '
%r 6
to
5
4
3
2
1
N
N
0
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
N P
N
I I 1
SYMBOL IS LETTER OF STATE
i I
3;5
4.0
4.5
5.0
5.5
6.0
6.5
7.0
pH
Figure 3-6. Ratio of
to N0o~ versus pH of precipitation in New
York (N), Pennsylvania (P), and Ohio (0) during the
growing season. Data are from the NADP network, 1979.
3-60
-------
produced no effect on the corn yield. The experiment was repeated the
following year with similar (although not statistically significant)
results and for a third year with no negative effects observed (J. J.
Lee, pers. comm.). The reduction in the yield of one ("Amsoy") of the
five cultivars of soybeans that have been studied suggests that genetic
or soil factors or the presence of heavy metals 1n the rain simulant may
control plant response to acidic rain. If these possibilities are
substantiated, ramifications of the negative effects of acid rain
observed in the above two studies could be considerable since soybeans
and field corn are two of the most economically important agricultural
crops 1n 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 production.
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 documented 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 importance 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 nutrients may stimulate plant growth and
yield (Garcia and Hanway 1976). The repeated exposure of plants to
rain, especially during the critical reproductive stage, suggests that
nutritional benefits from rain may be significant, 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 responds 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
3-61
-------
agricultural soil scientists generally accept that the Influence of
acidic deposition on the need for additional fertilizer and Hrne
application 1s probably mlnlscule.
Another consideration that may be Important 1n controlling the
Impact of ambient add precipitation, 1s that crop cultlvar
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 pollutants.
3.4.2.4 Summary--
1) Because of limitations In research design, differences In
methodologies and Inconsistent results, 1t 1s difficult to
compare research results directly or arrive at an overall
conclusion regarding crop response to acidic deposition without
a thorough description and comparison of experimental methods.
2) Complex factorial research designs and multlvarlate analyses
may be necessary to describe adequately the relationship
between acid rain dose and plant response rather than the
simple unlvarlate 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
controlled environment experiments Indicates that the majority
of crop species exhibited no effect on growth or yield as a
result of exposure to simulated acidic rain (acidity treatments
had pH values of 4.2 or less). Growth and yield of a few crops
1n 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 1s the result of
the Interaction between 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.
Available experimental results appear to Indicate that the
effects of acidic precipitation on crops are minimal and that
when a response occurs 1t 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 divers
plant-pollutant relationships. Documented experiments concern widely
varying situations, from controlled-envlronment studies to field
3-62
-------
studies, and from intensively managed agricultural systems to natural
plant communities. Control1ed-env1ronment 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:
0 Leaf structure may play two roles 1n the sensitivity of foliar
tissues to acidic deposition: 1) leaf morphology may
selectively enhance or minimize surface retention of incident
precipitation, 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).
0 Leaching mechanisms are major factors in nutrient cycling in
terrestrial ecosystems and are critical to the redistribution
of nutrients within these cycles. If the rate of leaching
exceeds the rate of mineral nutrient uptake, plant growth and
yield reductions are likely (Section 3.2.1.).
0 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).
0 A review of the evidence on the interaction of forest trees,
Insect and microbial pests, and acidic deposition does not
allow generalized 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 combine field and controlled
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
depend 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).
3-63
-------
At present we have no direct evidence that acidic deposition
currently 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 environment studies Indicate that the deposition of
acidic and acidifying substances may have stimulatory,
detrimental, or no apparent effects on pi ant growth and
development. Response depends upon species sensitivity, plant
life cycle stage, and the nature of exposure to acidity. Some
simulation studies have Indicated 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 negatively affected by acidic deposition, while
others exhibited positive responses (Section 3.4.2).
A crop's net response to acidic deposition results from a
combination 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 outweight the
positive effects, however, many crops and agricultural systems
have not been properly or adequately studied (Section 3.4.2).
3-64
-------
3.6 REFERENCES
Abougendia, Z. M. and R. Redman. 1979. Germination and early seedling
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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, acidification of streams and lakes, with
subsequent biological damage, has become a well reviewed effect of
acidic deposition (NAS 1981, U.S./Canada 1981, NRCC 1981). However,
confusion, ignorance, and debate still cloud our knowledge of past,
current, and future trends in the acidification of aquatic systems, the
key processes that control the acidification, and the degree of
permanency of biological effects.
This chapter critically reviews how aquatic chemistry responds to
acidic deposition. Initially, basic definitions and concepts regarding
acidic deposition, terrestrial and aquatic systems, measurements of
sensitivity using alkalinity, and' the different time scales of
acidification are presented. These definitions are followed by a
detailed listing of the important characteristics of terrestrial and
aquatic system subcomponents that ameliorate or enhance the effect of
acidic deposition. The theoretical and practical sensitivity of aquatic
systems to acidic deposition is discussed with locations of sensitive
and affected sensitive systems documented. Aquatic systems considered
include lakes, streams, and estuaries. The degree of acidification of
these systems is examined in the light of existing models, and the
models are critically reviewed. The role of S and N in the
acidification process is addressed. The status of our knowledge on
acidification of aquatic systems is presented in the context of asking
what will happen if depositions of S and N compounds from the atmosphere
increase or decrease. As a final section, the interaction of aquatic
acidification with the metal and organic biogeochemical cycles is
addressed and an assessment of knowledge is presented.
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
4-1
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(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 1n
another component. The hydro!ogle pathway controls which components are
affected by (or linked to) other components. For example, water (preci-
pitation) 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 'upstream1 from the component of Interest.
For example, the effect of atmospheric deposition on aquatic systems
depends on what is in the atmospheric deposition and its effect on all
components of the terrestrial and wetland systems that it contacts prior
to discharge into the aquatic system (Figure 4-1).
Therefore, when discussing effects on components of terrestrial,
wetland, or aquatic systems we are incorrect to investigate only the
component of interest. Rather, all components either directly or
indirectly linked to the specified component should also be included in
the study. This point will become especially pertinent in Section
4.3.2, when it is shown that decreases in sulfur deposition from the
atmosphere may not result in decreases in lake acidification until the
terrestrial system above the lake recovers. Instances where the
terrestrial 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
systems.
The composition of aquatic systems is controlled by not only
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,
4-2
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TERRESTRIAL
ECOSYSTEM
AQUATIC
ECOSYSTEM
METEOROLOGIC
V
GEOLOGIC
BIOLOGIC
Figure 4-1. Diagrammatic model of the functional linkages between
terrestrial and aquatic ecosystems. Vectors may be
meteorologic, geologic, or biologic components moving
nutrients or energy along the pathways shown. Adapted
from Likens and Bormann (1974).
4-3
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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 the terrestrial systems are found in this
section (following) and in Chapters E-2 and E-3 on soils and vegeta-
tion, respectively. The next chapter, Chapter E-5, discuss the effects
of acidification of aquatic systems on biota.
4.2.2 pH, Conductivity, and Alkalinity (M. R. Church)
Three important analytical measures for evaluating acidification of
ground or surface waters are pH, conductivity, and alkalinity.
Definitions of these three quantities are briefly given here. A later
section (4.3.3.2.1) examines problems concerning the comparability of
historical and more recent pH, conductivity, and alkalinity data.
4.2.2.1 j)H--In 1909 the Danish chemist S. P. L. Stfrensen 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 - icrp [4-i]
where CH was the hydrogen ion concentration and P was the hydrogen ion
exponent, which Sdrenson 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 Stfrensen is not a measure of either
hydrogen ion activity or concentration (Feldman 1956, Bates 1973).
Fortunately, this fact, 1n and of Itself, does not adversely affect
measurements of surface water acidification. A practical (Feldman 1956)
or operational (Bates 1973) pH scale has been defined:
(Ex - EC) F
P"(x) = PH(S) + — C4-2]
RT In 10
where pH(s) 1s the assigned pH of a standard solution, Es the 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).
4.2.2.2 Conductivity—Conductivity (or specific conductance) measures a
solution's ability 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
4-4
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usually reported as ymho cnrl or the equivalent ySiemens cm'1.
Distilled water may have a conductivity as low as 0.5 ymho cm~l, and
some naturally occurring surface waters in the United States may have
conductivities as high as 1500 ymhos cm'1 (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 acidification (Ontario Ministry of the Environment
1979). {In some cases, however, organic compounds may contribute to
buffering but only very little to conductivity.) 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 water under some
circumstances can be used to infer acidification of that water body
(Nilssen 1980). Hydrogen ions have extremely high mobilities in
solution and contribute greatly to conductivity. As 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 cnrl, us-jng a value
of 0.313 ymho cm'1 per yeq JT1 free H+; see Wright and Snekvik 1978).
In the absence of other data (e.g., pH, alkalinity, acidity) such a
change with time could possibly be used to infer acidification,
provided, of course, no other reasonable explanation was apparent.
4.2.2.3 Alkalinity—Al kalinity measures the ability of an aqueous
solution to neutralize acid. For this reason it is also known as acid
neutralizing capacity, or ANC. In most natural freshwaters, buffering
is predominantly due to species of the carbonate system (Stumm and
Morgan 1981). 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
so that an appropriate expression for this quantity is
ANC = (HC03~) + 2 (C032') + (A10H2+) + 2 (A1(OH)2+)
+ 4 (A1(OH)4~) + (RCOO~) + (OH") - (H+) [4-3]
where (RCOO-) represents dissociated organic acids (Bisogni and
Driscoll 1979). In some waters organic acids may dominate both the pH
and buffering of natural waters. North American areas with waters of
this type include parts of the south and southeast, the upper midwest,
4-5
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locations in the northeast (see General Introduction for map defining
regions) and the Atlantic maritime provinces of Nova Scotia and
Newfoundland—all regions noted for their highly colored brownwater
lakes and streams. Naturally acidic brownwater lakes and streams are
discussed further in Section 5.2.1. For discussion of buffering due to
organic systems see Bisogni and Driscoll (1979), Wilson (1979), and
Section 4.6.3.2.
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, Loewenthai and Marais
1978, Skougstad et al. 1979, Stumm 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. Loss of alkalinity can be either chronic or acute,
identified as long-term acidification and short-term acidification,
respectively. Short-term acidification refers to the development of
strong acidity (i.e., alkalinity < 0) 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). Long-term acidification, on the
other hand, refers to the gradual loss of alkalinity over periods of
years or decades. As a result of chronic exposures, biological effects
may occur at alkalinity < 100 peg i~l (Chapter E-5, Section
5.10.4), and waters with alkalinity < 200 ueq £-1 are generally
considered sensitive as defined in Section 4.3.2.6.1.
4.3 SENSITIVITY OF AQUATIC SYSTEMS TO ACIDIC DEPOSITION (J. N.
Galloway and P. J. Dillon)
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 then
examine what is known about this response.
4.3.1 Atmospheric Inputs
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
4-6
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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" measurement. These data are inadequate for
assessing impact on aquatic and terrestrial ecosystems because they
underestimate total deposition, 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) "calibrated" lakes and watersheds have been used to
infer dry or total deposition of acidic substances. In other cases,
"bulk" deposition measurements (made with a continuously open collector)
have been used. Although these collect an undefined portion of the dry
deposition, this information is more useful for chemical budget
calculations than "wet only" 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
contributions of SCty2- and N03~ and NH4 (see Section 4.4.1 and
Chapter A-8). In some systems, N03~ is significantly utilized (chemically
or biologically), resulting in the internal production of ANC (Harvey et al.
1981, Dillon et al. 1982). In some cases, $042- is stored in terrestrial
watersheds by the process of sulfate adsorption (N. M. Johnson 1979), a process
that may also generate ANC if the S042~ is reduced or if strong acid is
simultaneously stored. $042- may also be reduced in lakes, resulting in
production of ANC (Cook et al. 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 $64 on an annual basis. In other systems
S042- apparently acts as a conservative substance within the limits
of error in the measurement of the dry deposition fluxes (Likens et al.
1977, Galloway et al. 1983c).
Once wet or dry deposited, $03 and SCty- have the same
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.
Virtually all of the ammonium ion, NH4+, deposited from the
atmosphere on terrestrial and aquatic systems is used chemically or
biologically in those systems (Likens et al. 1977, NAS 1981) in many
cases resulting in a decrease in ANC. NH4+ deposition is
"significant" (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+ deposition (Dillon et al. 1979). Therefore,
measuring only free acid (H+) is inadequate for assessing the impact
of acidic deposition on systems.
4-7
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The input rate of basic cations is also important because without it
the net loss of base cations from the watershed cannot be calculated.
4.3.1.2 Loading vs Concentration—Because the ANC of some components of
the 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 the concentration as well as the total loadings
of the acids. The assimilation capacities of components that have a
continually renewed ANC, (e.g., a lake's 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
passing through components as well as the concentration of acid.
In general, current measurements of acidic deposition include both
the concentration of important substances and the total loading rate 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—Deposition of acidic substances is
well measured in most areas where the geological terrain has low
capability to neutralize acids and where the wet deposition is known to
be relatively high (> 20 meq strong acid m~2 yr~l; 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 more poorly understood. For example, the ability of the
vegetation in a terrestrial system, particularly the forest canopy, to
enhance deposition of acidic substances relative to other components of
the system (e.g., bedrock or soil surfaces, water surfaces) 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 on 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 and subsequent release of a major portion of the annual
deposition during snowmelt to groundwaters and surface waters occur
(Jeffries et al. 1979, Galloway et al. 1980b). In some cases (e.g.,
Central Ontario) during snowmelt the ground is frozen. As a result,
4-8
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the release of Ions occurs at a time when the terrestrial system cannot
assimilate the ions as efficiently 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 more poorly studied and need
further quantification.
4.3.1.5 Importance of Atmospheric Inputs to Aquatic Systems--
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 Pin 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 et al. 1974). With human influences in the
watershed (urbanization, 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 oligotrophic lakes. Mean concentrations
averaged about 6 mg C £-1 in precipitation and accounted for 28
percent of the total allochthonous inputs of organic carbon for a small
oligotrophic lake in New Hampshire (Jordan and Likens 1974).
4.3.1.5.2 Sulfur. Two sources provide sulfur for surface waters:
rock weathering and atmospheric deposition. In the absence of
significant sulfur sources in bedrock, atmospheric deposition is the
primary source (Cleaves et al. 1970). This is especially true in areas
receiving acidic deposition, where atmospheric sulfate becomes the
dominant anion in low alkalinity waters (Gjessing et al. 1976, Oden
4-9
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1976a, Henriksen 1979, Wright et al. 1980, Galloway et al. 1983c). This
dependence is illustrated by plotting the mean and range of excess
S042- (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 $042- (Figure 4-2). The wet
deposition of excess S042~ at each location is shown in the same
figure, with estimated total S042- deposition shown at four
locations. There is a clear positive relationship between excess
S042~ deposition and SC)42~ in the runoff. The influence of the
increased S042- deposition on aquatic chemistry is large, for on an
equivalent basis, the increase in the SO*2- 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.
1983c). 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 cations from the terrestrial system. Both effects can
potentially alter biological communities in the respective ecosystem
(see Sections 4.4 and 4.6).
4.3.2 Characteristics of Receiving Systems Relative to Being Able to
Assimilate Acidic Deposition
The anthropogenic acids transported via the atmosphere may be
deposited directly into 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 acid deposition received. This section
discusses the factors that determine the quantitative capability of the
subsystems to assimilate acidic deposition.
4.3.2.1 Canopy--Throughfal1 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 signficant assimilation capacity. However, the relative
importance of each process is generally unknown. Although there are
conflicting reports, some generalizations may be made.
Stemflow 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
4-10
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-3-
LABRADOR
ISLAND OF
NEWFOUNDLAND
HALIFAX
NEW BRUNSWICK
LAFLAMME
MAURICIE
ADIRONDACKS
N. OF OTTAWA
ALGONQUIN
HALIBURTON
SUDBURY
ALGOMA
THUNDER BAY
QUETICO
ELA
,_ _
ui baiu
Figure 4-2. Mean and range of basin specific yield of excess sulfur
(|—<•>—|) (U.S./Canada 1982) compared with atmospheric excess
sulfur deposition (|—•—|) in precipitation for 1980
(Barrie and Sirois 1982). Also shown are the ranges of wet
plus dry deposition of sulfur (| — |) calculated from the
1980 measurements of SOX in the atmosphere at 4 Canadian
Acid Precipitation Network Stations (Barrie 1982).
4-11
-------
Miller 1980) also may contribute to the acidity of the throughfall. The
throughfall pH of in coniferous forests has been reported to be
decreased relative to pH of precipitation (Horntvedt and Joranger 1976),
although cation content is increased.
However, the amount of throughfall or stemflow is 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.
In general, surficial materials containing carbonate minerals have
abundant exchangeable bases and can assimilate acidic deposition to an
almost unlimited extent. Regions of North America with soils formed in
situ on limestone, 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 assimilation 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, silicate-mineral dissolution reactions and, in some cases, Fe
and Al oxide dissolution. The result of these reactions is an increase
in the concentrations of major cations (particularly Ca2+, Mg^+, 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
4-12
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TABLE 4-1. TYPICAL CATION EXCHANGE CAPACITIES OF SOIL COMPONENTS
(FROM MCFEE ET AL. 1976)
SOIL COMPONENTS
(meq 100 g'1)
Organic matter (humus) 200
Silicate clays
vermiculite 150
montmorillorite 100
kaolinite 10
illite 30
Hydrous oxide clays 4
Silts and sands negligible
aVariation is commonly 40% of these mean values.
4-13
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Mg2+, Na+, and K+). The organic layer of the soil has a high CEC
(McFee et al. 1976). Fresh organic litter has a substantial BS
component, so it has the potential to assimilate acidic deposition if
the BS is high, but potential is low if the BS is low.
The permeability of the soil layers is also important because it
determines the contact time of the percolating water with the soil
particles. 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 increasing 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 surficlal 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. Further evaluation of soils as they affect aquatic systems is
found in Chapter E-2.
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.
Generally, soils in unglaciated regions have much greater sulfate
adsorption capacities (see Chapter E-2, Section 2.2.8) and hence will be
protected from extreme acidification of aquatic systems (relative to the
northeastern United States) until the sulfate adsorption capacity is
saturated. Estimates of the length of that period are decades to
centuries (Galloway et al. 1983a).
4.3.2.3 Bedrock—The ability of bedrock to neutralize acidic deposition
is controlled by:
1) chemical composition of the bedrock,
2) effective reaction surface area, and
3) retention time or contact time of water with the bedrock.
4-14
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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. Surficial geology, including glacial deposits, soils, and
unconsolidated material, has a greater influence on a system's ability
to assimilate acidic depositon. Bedrock influence on surface water
chemistry is mainly indirect through derived unconsolidated material.
4.3.2.4 Hydrology—
4.3.2.4.1 Flow paths. The extent of reaction of the strong acid
components of deposition 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.
The contact time of runoff waters in the organic and inorganic
layers of the soil profile depends on many factors: topography (e.g.,
basin slope, soil thickness), meteorology (e.g., precipitation rate,
season), etc. These factors affect the degree of soil saturation, which
in turn determines contact time.
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
4-15
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TABLE 4-2. 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
4-16
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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 central Ontario, the
upper 1.0 to 1.5 m of the soil is usually frozen each winter (Jeffries
1981, pers. comm.) so spring runoff may flow 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.
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 proportion 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 on an annual
basis (Galloway et al. 1980a, Wright and Johannessen 1980, Jefferies et
al. 1981). Therefore, 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 sensitive) 1f 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.
4-17
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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 because the high
organic content and relatively shallow water depth 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, have been shown to
produce mineral acidity (Clymo 1963) by means of cation exchange
reactions.
4.3.2.6 Aquatic—The ability of an aquatic system to assimilate acidic
deposition must be considered with respect to time frame, i.e., the
ability of the system to prevent long-term acidification vs short-term
acidification. This, in turn, depends on the system's ability to
assimilate acidic deposition at all times so that no fluctuations in pH
or alkalinity result in mineral acidity at any time, even during major
hydrologic events such as snowmelt or stormflow.
4.3.2.6.1 Alkalinity. A threshold alkalinity, below which an aquatic
system would have the potential for becoming acidic to a point where
biological effects might occur, can be estimated. The following
material provides support for a quantitative estimate of such a
threshold. It should be clearly recognized that the presentation
emphasizes the long-term (years/decades) concept of lake and stream
acidification. Also, the computed threshold ignores any acid
assimilation of acidic deposition by the watershed, the importance of
which is noted in the preceding sections. As a result, the estimate to
follow represents an upper threshold limit for systems receiving present
levels of acidic deposition. Using this threshold to estimate the
number of systems sensitive to the long-term inputs of acidic deposition
would likely result in an overestimate. However, this same threshold
would underestimate the number of aquatic systems that are susceptible
to short-term (days/weeks) acidification.
The instantaneous ability, i.e., excluding watershed influences,
of lake or stream water to assimilate acidic deposition is quantita-
tively measured as the ANC or alkalinity of the water (Stumm and Morgan
1970). HCOs" is an adequate measure of ANC in most lakes because
other contributing species (e.g., ammonia--NH3, borate--B(OH)4~)
are of minor importance.
4-18
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Alkalinity frequently has been used to assess the sensitivity of
lakes to acidic deposition, and subjective criteria have been
established to "classify" lakes; e.g., lakes in Ontario were classified
as having extreme sensitivity if 0 to 40 yeq alkalinity £-! was
measured, moderate sensitivity if 40 to 200 yeq alkalinity £-1 was
measured, etc. (Anon 1981). The boundary between "sensitive" and
"insensitive" that is commonly used is 200 yeq£-l of alkalinity
before the onset of acidification (Hendrey et al. 1980b). Altshuller
and MacBean (1980) classified lakes as "susceptible" if alkalinity was
measured as < 200 yeq s,'1. Calcite saturation index (CSI)~~a
measure of the degree of saturation of water with respect to CaC03
(calcite) that integrates alkalinity, pH, and Ca concentration—has
also been used (Kramer 1979, unpub. manuscript; Harvey et al. 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.
To understand why an alkalinity of < 200 yeq A'1 was selected
as indicative of sensitivity, it 1s first necessary to explore the
relationship between pH and alkalinity in oligotrophic systems. A
typical relationship between alkalinity and pH for oligotrophic systems
1s obtained by plotting pH versus alkalinity for 928 streams and lakes
in New York State (Figure 4-3, individual points not shown). This
graphical relationship between pH and alkalinity will be used in two
ways. First, it will illustrate dependence of pH changes on the
magnitude of the initial alkalinity. Secondly, it will be used to
relate alkalinity changes to pH changes and subsequent biological
effects. This linkage is needed because the literature on the chemical
effects of acidic deposition on aquatic ecosystems uses alkalinity as
the critical variable and not pH. However, those researchers
investigating biological changes due to acidic deposition relate the
biological changes to changes in pH and not alkalinity. Therefore, to
be able to relate changes in alkalinity of aquatic systems due to acidic
deposition to changes in biological systems due to changes in pH, we use
Figure 4-3.
The logic behind alkalinities < 200 yeq £-1 as an upper
long-term sensitivity limit is as follows:
o On a regional basis, the maximum increase of $04, due to
acidic deposition, in aquatic systems is ~ 100 yeq £-1
(Harvey et al. 1981, NRCC 1981). Therefore, the maximum
alkalinity decrease that could have occurred over time is -
100 yeq £-1 (see explanation, Section 4.4.3).
o Biological effects due to acidification become apparent as
alkalinities decline to about 65 to 35 yeq £-1 and pH's
between 6.5 and 6.0 (Figure 4-3; Chapter E-5, Table 5-16).
o Given the above two points, if systems with alkalinities
< 200 yeq £-1 are acidified to the maximum amount
(Aalkalinity = 100 yeq £-!), then resulting
4-19
-------
-100
ALKALINITY (peq i'1
Figure 4-3. The change in pH for a given change alkalinity at two
alkalinity levels and an example of pH-alkalinity 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,
50, 84, 98) computed from the data at each 0.1 pH point.
Adapted from Hendrey (1982).
4-20
-------
alkalinities will be < 100 yeq A-lf which is near the
threshold for biological effects on a long-term basis. In
addition, there is a short-term consideration. If an aquatic
system that was originally at 200 yeq x,'1 is acidified to
100 yeq £-1, there may be no biological effects on a
long-term basis, but there could be some on the short-term
basis. During spring snowmelt, alkalinity reductions of > 100
yeq £~1, lasting several weeks, have been reported
(Galloway et al. 1980b, Galloway and Dillon 1982).
This is graphically illustrated in Figure 4-3, where 200 yeq £-1
is represented by Xi. If acid precipitation is added to the system,
the alkalinity could decrease to - 100 yeq fc-1 (X2). This assumes
that acidification is immediate and that the only chemical effect of the
added acid is the loss of HC03~. Thus, it represents an upper limit
(see Section 4.4.3.1.1.1). Note that this acidification has only
resulted in a small pH change (~ pH 7.4 to 7.1). However, future
inputs of acidic deposition on a long-term basis (years/decades) or any
short-term acidification (days/weeks) will probably result in large pH
decreases, with subsequent biological effects, because the system is now
poorly buffered.
Systems of alkalinities ~ 200 yeq £-1 that are exposed to
high current levels of acidic deposition (annual average pH _< 5.0) are
only minimally sensitive. [Natural pH of precipitation in eastern North
America is most certainly greater than pH 5.0 due to the absence of
large sources of sulfur and nitrogen oxides (Galloway and Whelpdale
1980).] However, systems with < 100 yeq Jr1 of alkalinity that
are exposed to the same acidic deposition are very sensitive. For
example, point X2 represents a system that started with 100 yeq
£ . Point \3 represents the same system after it is exposed to
current levels of acidic deposition. The subsequent large decrease in
pH and alkalinity will influence a large variety of biological
responses.
In summary, the boundary between sensitive and nonsensitive aquatic
systems is chosen to be 200 yeq £-1 after consideration of (1)
current levels of acidic deposition (annual average pH £ 5.0), (2) the
relationship between pH and alkalinity in oligotrophic systems, and (3)
the pH and alkalinity values at which changes in them 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
underestimates those systems sensitive to short-term acidification.
Although parameters such as alkalinity and calcium saturation index
(CSI) probably provide reasonable estimates of the instantaneous ability
of the lake water to neutralize acidic deposition, they have two
drawbacks: the interpretation of degree of sensitivity is subjective,
and the methodology relies on only a static measure. The renewal rate
of ANC is undoubtedly more important and ultimately governs a lake's
4-21
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sensitivity. Renewal rate of ANC in a lake depends on the rate of ANC
supply from the lake's watershed which, in turn, depends on many factors
(see preceding sections); the importance of direct atmospheric
deposition on the lake surface (since this will have alkalinity < 0)
relative to the total water budget; and the internal renewal rate of
alkalinity. As long as the total input of alkalinity, including
external and internal sources, remains positive, the lake will not
become acidic (i.e., will not have mineral acidity) because alkalinity
is a conservative parameter. However, short-term acidification (e.g.,
at snowmelt) may occur in lakes that will never experience long-term
acidification.
4.3.2.6.2 Internal production/con sumption 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 (Brewer and Goldman 1976, Goldman and
Brewer 1980). The generation of alkalinity depends upon the use of
as a nitrogen source by algae:
106 C02 + 16 N03~ + HP042' + 122 H20 + 18 H+ ->
+ 138 02. [4-4]
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 most north-temperate lakes are such that N03" "se
surpasses Nfy use (Harvey et al. 1981; Dillon, unpub. studies).
Any NH4 use results in a decrease in alkalinity in the lakewater
(D. W. Schindler, unpub. studies). However, a net gain of ANC can
result from NOs- uptake during primary production since the
inorganic NOo" is converted to organic nitrogen and stored
permanently in the lake's sediments.
The uptake of N03~ (corrected for uptake of NH4+) is
often in the range of 10 to 20 ueq £-1 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
yeq £-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~l, 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 generation if N03" is used as the N-source, on a net
basis, and the organic N is lost permanently to the sediments.
Fertilization with NHA+, on the other hand, may result in lake
acidification (D. W. Schindler 1981, pers. comrn.). Nutrient status is
therefore very important in determining the sensitivity of a lake to
acidic deposition.
4-22
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In most lakes (I.e., those with vernal circulation), the alkalinity
of the whole lake 1s uniform at the Initiation of stratification. If no
Internal production or use of alkalinity occurred In the hypollmnlon, It
would serve as a "reserve" pool for the lake, because the external
Inputs of strong acids (atmospheric deposition, runoff) enter the
epil Imnion. However, Internal processes within the hypollmnlon may also
deplete or produce alkalinity (Schlndler et al. 1980, Cook 1981, Harvey
et al. 1981, Kelly et al. 1981, unpub. manuscript).
Acidification of lakes by acidic deposition results 1n Increased
transparency (Dillon et al. 1978, Schlndler et al. 1980, Harvey et al.
1981, Schindler and Turner 1982, Van 1983). Therefore, hypolimnetlc
primary production (by phytopl ankton or perlphyton), and associated
production of ANC, may be elevated relative to non-acidic lakes of
equivalent nutrient and morphometrlc status.
Under oxic conditions, respiration of organic matter (produced
principally in the epilimnion and metal imnion) results in a decrease in
alkalinity or depletion of ANC:
C106H263°110N16P1 + 138 °2 * 106 C02 + 122 H2°
[4-5]
+ 16 HN03 + H3P04.
This reaction may occur in the hypol Imnetic 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 anoxlc conditions, several microbial processes that occur in
the hypol imnion (or in the surficial sediments) and that require organic
material produce alkalinity:
o
$04 reduction
53 so42' + 106 H+ +
[4-6]
106 C02 + 16 NH3 + H3P04 + 106 H20 + 53 H2S
NQ3~ reduction
5 C6H12°6 + 24 N03~ + 24 H+ -»• 30 C02 + 12 N2 + 42 H20 [4-7]
Mn4+ reduction
C106H263°110N16P1 + 236 Mn02 + 472 H+ ->
94. [4-8]
236 Mn2+ + 106 C02 + 8N2 + H3P04 + 366 H20
4-23
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Fe3+ reduction
c106H263°ll()Nl6pl + «4 FeOOH + 848 H+ ->
[4-9]
424 Fe2+ + 106 COg + 16 NH3 + ^04 + 742 H20.
However, the alkalinity produced by some of these processes may
be temporary. Fe2+ and Mn2+ (and NH4+) production Is probably
largely temporary, with the reverse reaction occurring as soon as oxlc
conditions again prevail at overturn. NOa" reduction occurs In
hypollmnla or In lake sediments, but the M2 evolution makes the
reaction Irreversible; therefore, this represents a source of
permanent alkalinity. SCty2- reduction results 1n permanent
alkalinity 1f the $2- formed Is Irreversibly lost to the sediments.
Any S2~ (HS-, H2$) left 1n the water column at fall circulation is
re-oxidized to S042', with concurrent loss of alkalinity.
Therefore, the critical factor with respect to the ability of a
lake's hypolimnlon to assimilate acidic deposition is its oxygen regime.
At the Experimental Lakes Area, Schlndler et al . (1980), Kelly et al .
(1981), and Cook (1981) studied fertilized and unfertilized lakes that
had anoxic hypolimnia and consequent summer alkalinity production.
Increased S04Z" input resulted in increased alkalinity generation.
In Muskoka and Haliburton counties (Dillon et al., unpub. results) and
in the Sudbury area (Van and Miller 1982), most study lakes did not have
large anoxic zones in their hypolimnia and $04 loss was not
observed. Fertilized lakes (Yan and Lafrance 1982) were an exception,
however.
4.3.2.6.3 Aquatic sediments. The potential for lake sediments to
assimilate acidic deposition is not quantitatively understood. The same
mlcrobial processes that occur In hypolimnia occur in lake sediments,
but the contribution 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 neutralization of three study lakes
(with CaC03 plus Ca(OH)2), the pH of the upper sediments increased
to the same levels as the deep sediments. Sediment 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~2 of BNC.
Over the subsequent five years, one of the three neutralized lakes was
reacidlfied. 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
result in acidification of lakewater by Ca2+ exchange for H+. He
suggested, however, that the reverse process will occur with increased
4-24
-------
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.
Unfortunately, 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 does reflect a combination
of many relevant factors. Aquatic systems with alkalinity < 200 yeq
£-1 have been defined in Section 4.3.2.6.1 as sensitive to
acidification by acidic deposition. These systems can be located by
direct analyses of alkalinity over large areas or by use of geological
and soil maps to identify areas that will have aquatic systems with low
alkalinities. The advantage 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 the sensitivity on a regional basis and
that in the absence of measurements no mechanism to estimate the
alkalinity exists. The advantage of the second method is that broad
regional determinations can be made, but the major disadvantage is that
no fine detail is available. Therefore, the proper way to address this
issue is to use the regional data on bedrock and soil characteristics to
determine general areas of sensitivity and then to follow up with
alkalinity surveys in the 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
approximation of sensitive areas in North America (Figure 4-4). Their
identification was improved by the addition of soils and surficial
geology information to determine sensitive systems for eastern Canada
(NRCC 1981; Figure 4-5). For the soils of the eastern United States,
McFee (1980) used CEC and percent BS to indicate the areas where soils
would be expected to be sensitive to acidic deposition (Figure 4-6).
Significant portions of the soils within the areas designated
"sensitive" or "slightly sensitive" would provide low buffering capacity
and therefore would have little ability to neutralize acidic inputs as
they passed through the soil.
As a check on the use of soil characteristics and bedrock geology
as predictors of low alkalinity waters, Hendrey et al. (1980b) using
Norton's (1980) methods, compared surface water alkalinities waters with
sensitivity predicted on the basis of geology on a county by county
(U.S.) basis; they found clear correlations. Haines et al. (1983b)
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 bedrock 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 Figure 4-7.
4-25
-------
Figure 4-4. Regions in North America containing lakes that are potentially
sensitive, based on bedrock geology, to acidification by
acid precipitation. Adapted from Galloway and Cowling (1978).
4-26
-------
HIGH SENSITIVITY
I Granite, granite gneiss,
orthoquartzite, syenite
INTERMEDIATE-HIGH SENSITIVITY
Volcanic rocks, shales, greywacke
jsandstones, ultramafic rocks, gabbro,
mudstone, and metamorphic equivalents
INTERMEDIATE-LOW SENSITIVITY
Calcareous clastic rocks, carbonate rocks
interbedded or interspersed with non-calcareous
sedimentary, igneoussand metaporphic rocks
Limestone, dolomite and metamorphic
equivalents
Figure 4-5. 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).
-------
LEGEND
NONSENSITIVE SOILS
SLIGHTLY SENSITIVE SOILS
SENSITIVE SOILS
Figure 4-6. Regions with significant soil areas that are potentially
sensitive to acidic deposition. Adapted from McFee (1980),
See also discussion of soil sensitivity in Chapter E-2,
Section 2.3.5.
4-28
-------
LEGEND
-------
The map is a useful presentation of regions where waters of low
alkalinity might be found. In essence, the map was created using a
predictive technique. Specifically, existing data on surface water
alkalinity were compiled and then correlated with geology, soils,
climatic, physiographic, and human factors. These correlations were
used to predict mean annual alkalinities for regions without data.
There are problems with this predictive technique, that while not
terminal, need to be realized. First, if the compiled data are not
themselves representative 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, and 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 alkalinities of most surface
waters in a given region. In areas where substantial heterogeneities in
soil, geology, elevation, etc. occur there may be large variations from
the mean. Unfortunately, sensitive areas generally occur in regions
with large variations in elevation and soil thicknesses.
Some of these problems can be resolved by field testing the
predictions of the method. However, these field tests will be site
specific. A good test in one area would not necessarily mean that all
regions would be as well behaved.
We know large regions in North America contain aquatic systems with
low alkalinity that are presumably sensitive to acidic deposition.
These regions are found throughout much of eastern Canada; New England;
the Allegheny, Smoky, and Rocky Mountains; and the Northwestern and
North Central United States (Galloway and Cowling 1978, NAS 1981, NRCC
1981) and parts of the south and east coasts of the United States
(Omernik and Powers 1982). However, a large amount of more detailed
survey work is 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 composition of the deposition, the total rate of the loading (wet
plus dry deposition), the temporal distribution, and the characteristics
of the receiving system.
Atmospheric deposition is a major source of chemicals to aquatic
systems. The chemicals supplied by atmospheric deposition in important
quantities include P, Ca, Mg, S, N, Pb, Zn, H,. Cl, Na, and Cd. Of
these, the concentrations of Pb, Zn, Cu, Cd, S, NOX, and H, in
atmospheric deposition in eastern North America are apparently directly
controlled by anthropogenic activities (Chapters A-2 and A-3; see also
Sections 4.3.1.5 and 4.6.1.1). The effects of S on aquatic systems are
independent of type of deposition (wet or dry) or chemical speciation
(e.g., S02, $04). It is the total deposition of the total element
that controls the effect in aquatic systems.
4-30
-------
The effects of acidic deposition on aquatic systems depend upon the
characteristics of the receiving systems. Three such characteristics
determine the ability of the receiving systems to assimilate acidic
deposition. The characteristics are size, composition, and hydrological
residence time. The smaller the receiving system, the less likely it
will be able to assimilate the acidic deposition. The greater the
watershed to surface water ratio, in general, the greater the ability to
assimilate acids. The ability of systems to assimilate the acidic
deposition also depends upon the composition and characteristics of the
soil in exchange. Small systems with calcareous rock, for example, are
much better able to assimilate the acidic deposition than small systems
with granite bedrock and low CEC, percent BS, and sulfate adsorption
capacity. The hydrologic residence time is also important. For,
generally, the longer the acidic deposition stays in contact with the
system the more it is assimilated. The longer the hydrologic time in
the terrestrial system, the less the effect of the acidic deposition on
the aquatic system. The aquatic systems that tend to be the most
sensitive to acidic deposition are located in the areas 'downstream of
terrestrial systems that are small, have slowly weathering soil and
bedrock, have short hydrologic residence times, and are unable to
assimilate the acidic deposition that falls into them.
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 alka-
linities < 200 yeq £-1 (see Section 4.3.2.6.1). Such systems are
located through much of eastern Canada, New England, the Allegheny,
Smoky and Rocky Mountains, the Northwestern and North Central United
States and parts of the south and east coasts of the United States.
4.4 MAGNITUDE OF CHEMICAL EFFECTS OF ACIDIC DEPOSITION ON AQUATIC
ECOSYTEMS
The previous sections have laid a foundation of important
definitions, concepts 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 acidification.
4.4.1 Relative Importance of HNO^ vs H?S04 (J. N. Galloway)
is the more important in acidification of aquatic systems
for two reasons. First, in most areas impacted by acidic deposition,
atmospheric H2S04 loading exceeds the HN03 loading. The second
reason that increases the importance of ^$04 relative to HN03 has
to do with the terrestrial system associated with the aquatic system.
Specifically, for aquatic systems dependencies are controlled by the
ability of the terrestrial system to retain $042- and N03~. In
the case of $042-, because its concentration varies little
seasonally or year-to-year in a given volume of surface water (e.g.,
stream or lake epilimnion), the spatial variability is the most
important. This variability is controlled by the $042- adsorption
4-31
-------
capacity (SAC) of the soils in the terrestrial system. In the
northeastern United States and eastern Canada, the SAC of the soils is
low; thus the $042- concentrations in the surface waters are higher
than those in the mid-Atlantic region of the United States, which has
soils with a higher SAC (Chapter E-2, Section 2.3.3.2; sulfate
adsorption in soils).
In the case of N03~, regional variability is overwhelmed by
seasonal variability. During most of the year, the hydrologic residence
time in the soil is sufficient to allow for rapid uptake of NOs"
(Likens et al. 1977). Of the HQ-$~ released from the terrestrial
system to the aquatic system, 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 uptake within the terrestrial system. In addition to
constraints of the hydrologic residence times on N03~ transport
through soil systems, a temperature dependency also exists. During warm
periods (e.g., summer), when biological acitivity is the highest,
N03~ is efficiently retained by the terrestrial systems. However,
during colder periods (e.g., winter) there is often a maximum of
concentrations (Likens et al. 1977).
Therefore, the two periods that allow a larger flux of N from the
soil system to the lakes with subsequent increased N03" concentrations
are winter base flow and spring snowmelt. This is illustrated by a
37-month record of the N03- concentration in the outlet of Woods Lake,
a small oligotrophic lake in the Adirondack Mountains, NY, (Figure 4-8)
and by two of the inflows to Harp Lake in Southern Ontario (Figure 4-9).
The NOj values are highest in the winter and during spring snowmelt
(usually in March and April). Therefore, during most of the year,
because H2$04 deposition is greater than HN03 deposition in most
areas and because there is greater retention of HN03 by the
terrestrial system, H2S04 is more important than HN03 in causing
acidification of aquatic systems.
Due to the increased importance of HN03 (relative to H2S04)
in acidification of the spring snowmelt and the effects of the large pH
and alkalinity decreases (Table 4-3) on sensitive life forms and life
stages, it is necessary to explore further the relative roles of HN03
and H2S04 during spring acidification. Galloway et al . (1980b)
studied the role of N03 as a source of acidity for Woods and Panther
Lakes, acidified during the 1979 snowmelt. They found that in the two
lakes alkalinity decreased during snowmelt because of dilution of base
cations (Co) and an increase in HN03 in the lake epilimnion (Figure
4-10). Although $04 concentrations changed only slightly 1n Woods and
Panther Lakes during snowmelt, S042~ still contributed to the
acidification in an Indirect manner, namely, by causing long-term
alkalinity reductions (as opposed to episodic). Thus, the episodic
reduction of alkalinity due to NOg is added to the~long-term
reduction in alkalinity due to $04 (see Sections 4.4.2 and 4.4.3) .
Galloway et al. U980b) concluded that the primary cause of the
increased N03 concentration was release from the snowpack. An
4-32
-------
u>
co
Figure 4-8. The concentration of N03 in the outlet of Woods Lake, Adirondack Mountains, NY. Adapted
from Galloway and Dillon (1982).
-------
o>
2000
1600
1200
• 800
ro
o
400
Ou
HARP INFLOW 3A
2000
1600
1200
800
400
HARP INFLOW 5
^••••^••fc^™****"""*
1976 1977 1978 1979 1980
Figure 4-9. 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 (1982).
4-34
-------
TABLE 4-3. MAGNITUDE OF pH AND ALKALINITY DECREASES OF LAKES AND
STREAMS DURING SPRING SNOWMELT (GALLOWAY ET AL. 1980a;
1980b; JEFFRIES ET AL. 1979)
Adirondack Mountains, New York, USA
Panther Lake, 1979
Woods Lake, 1979
£ Sagamore Lake, 1979
Southern Ontario, Canada
Harp Lake #4, 1978
Dickie Lake #10, 1978
Paint Lake #1, 1978
Prior
PH
6.6
4.8
6.1
6.6
4.8
5.5
to Melt
Alk
yeq A-l
162
-38.5
28.8
108
-16
61
During Melt
pH Alk
yeq A'1
4.8
4.5
4.9
5.4
4.5
5.0
-18
-42.2
-16.7
8
-32
8
PH
1.8
0.3
1.2
1.2
0.3
0.5
A
Alk
yeq A-1
180
4.0
45.5
100
16
53
-------
.£»
CO
I
o?
cr
0)
240
200
160
-80
m
RTOT
MIDWINTER
THAW
SPRING
"JAW
LEGEND
PANTHER LAKE
WOODS LAKE
JL
J L
M A M J J A
1978
0 N
F M A M
1979
Figure 4-10. Temporal trends in alkalinity at outlets of Woods and Panther Lakes. Adapted from
Galloway et al. (1980a).
-------
analysis of two additional snowmelt periods (1978, 1980) supports this
conclusion (Galloway et al. 1983b).
The chemical changes that accompany the decreases in pH and
alkalinity, however, are not consistent from study area to study area.
For example, Jeffries and Snyder (1981) found that $042- levels
increase in several streams in the Muskoka-Haliburton area of Ontario at
peak flow during snowmelt. On the other hand, Johannessen et al. (1980)
reported decreasing $042- during snowmelt in streams in Norway.
Three of six streams studied by Jeffries and Snyder (1981) exhibited
declining N03~ concentrations associated with peak H+ concentrations,
a finding opposite to that of Galloway et al. (1980b) in the
Adirondacks.
In sunmary, during most of the terrestrial biological productivity,
$04^" is the most important anion causing acidification. However, in
winter and in areas with snowpacks, in the spring N03 can become more
important both 1n an absolute sense and relative to $042-. The
effects of H2S04 and HN03, on acidification of aquatic ecosystems
are:
0 H2S04 causes long-term (decades) alkalinity reductions on a
regional basis.
0 HN03 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)
Acidification of lakes and streams during major hydrologic events
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.
1979, Jeffries et al. 1979, Jeffries and Synder 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 have resulted in decreases
in alkalinity and pH 1n at least one case (Scheider et al. 1979).
Episodic events have resulted in decreases in pH of greater than or
equal to one pH unit in several reported cases (Table 4-3). 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
vieq A-l. During the 1979 spring snowmelt, the pH and alkalinity
decreases in Panther Lake epilimnion were 1.8 pH units and 180 yeq
A"1, respectively (Galloway et al. 1980a,b). Streams in Ontario,
Canada and New York, USA with lower pre-melt pH's and alkalinitles had
correspondingly smaller decreases (Table 4-3).
Based on the studies cited above and on other available data sets
(Leivestad and Muniz 1976, Schofield 1980) it is reasonable to expect
4-37
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that the pH levels during spring snowmelt may approach pH 4.3 to 4.9.
This is the same pH range observed for chronic, long-term acidification
(Section 4.4.3; Pfeiffer and Festa 1980, Haines et al. 1983b). The
difference is that in episodic acidification, aquatic systems with pH's
as high as 7.0 can be acidified to pH < 5.0. While 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.
To predict the importance of episodic events in aquatic ecosystems,
one must be able to evaluate the probability of chemical (pH, alkalinity,
aluminum, etc.) change of specific magnitude in a lake or stream for a
specified duration. To construct a model to make these predictions, one
must know the following:
a) the amount of the chemical of interest and of water stored
(mass/area in the
watershed,
the hydrological pathway that the snowmelt follows (Section
4.3.2.4), and the biogeochemical interactions that occur en
route (e.g., ion exchange, biological uptake),
c) the volume of the lake that interacts with the runoff
(determined partly by temperature patterns),
d) the chemical composition and amount of the snowmelt relative to
the composition and amount of the baseflow.
e) the relative importance of natural vs anthropogenic induced
changes in runoff composition during snowmelt.
Factors a, d, and e have been measured in a few studies (e.g.,
Johannessen et al. 1980, Galloway et al. 1982a). Factors b and c have
rarely been investigated. Before a predictive model is available,
long-term (i.e., multi-year) measurements of all of these factors are
required in a variety of different geographical locations.
4.4.3 Long-Term Acidification (J. N. Galloway)
Affected sensitive aquatic systems must have three characteristics.
First, they must have waters with alkalinities < 200 yeq A""1
(Figures 4-4 to 4-7). Second, they must receive acidic deposition (pH £
4.5; Figure 4-11). Third, they must be shown to have been acidified by
acidic deposition. Overlaying Figures 4-4 and 4-7 (sensitivity to acidic
deposition) on Figure 4-11 (occurrence of acidic deposition) suggests
regions that potentially may have been impacted. Documentation that
aquatic systems have been acidified (lost alkalinity) by acidic
deposition has been provided by three techniques: (1) analysis of
temporal trends in alkalinity and pH, (2) paleolimnological analysis, and
(3) investigation of the importance and source of $04 in aquatic
systems.
Studies that have used the first technique, historical pH/alkalinity
data, to identify waters acidified by acidic deposition are reviewed in
4-38
-------
Figure 4-11. pH from weighted average hydrogen concentration
for 1980 for wet deposition samples (reproduced
from Barrie et al. 1982)
4-39
-------
Section 4.4.3.1. In some of the studies, problems exist with the
comparison of old data to recent data. The errors associated with the
comparison may preclude an absolute statement that each of the aquatic
systems has been acidified. However, the fact that many studies point to
decreased alkalinity is strong circumstantial evidence for acifidication
by acidic deposition.
Supporting this circumstantial evidence is the analysis of diatoms
in lake sediment cores. While 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 circumnavigating 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 ( zc^ = z a.,- where c-j is the normality of the ith
cation and ai 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 of the drainage basins. Marine aerosol
content corrections should also be performed.
The basis for the technique is that the concentration of S042~
in clear water lakes and streams has increased due to atmospheric
deposition. With the increase in S042~ has to come an increase in a
positive ion, H+, Ca2+, Mg2+, etc. If H+ increases, the aquatic
system is acidified (i.e., alkalinity decreases). If the concentration
of Ca or another non-protolytic cation increases, only, then no loss of
alkalinity occurs. For example, Figure 4-12 shows the two extremes with
chemical changes that can occur to an aquatic system when the $04
concentration increased by a factor of five. In one extreme, the
increase in the $04 anion is balanced by an increase in the
non-protolytic base cations (Alternative I, Figure 4-12). In the second
extreme, the increase in $04 is balanced by an increase in H+, which
causes a reduction of alkalinity (Alternative II, Figure 4-12). Since
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 and short hydrologic
path lengths), Alternative 2 appears to be a closer approximation to the
process that has occurred. As support of this Henriksen (1982), in an
analysis of long-term time series for the concentrations of Ca and Mg
over gradients of acidic deposition, concludes the increases in $04 in
lakes are balanced by increases in H+ (^60 percent) and increases in
base cations (£40 percent). Therefore, in aquatic systems with a
predominant atmospheric source of S042~ and with alkalinities less
than 200 eq -1, the increases in S042- will cause decreases
in alkalinity, i.e., acidification, although the magnitude/significance
of the decrease is dependent on watershed characteristics.
The beginning of this section stated that affected sensitive systems
had three character!sties—they were sensitive, receiving acidic
deposition, and had been shown to be acidified. Using the information on
temporal trends in Section 4.4.3.1.2 and studies on the role of
4-40
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PRE ACIDIC
DEPOSITION PERIOD
BASE CATIONS
S04
HC03-
ALTERNATIVE 1
ALTERNATIVE 2
ACIDIC
DEPOSITION PERIOD
BASE CATIONS
$04
HC03-
BASE CATIONS
H+
S04
Figure 4-12. Two extremes for the response of aquatic systems to a
5-fold increase in $04. The length of the boxes relates
to yeq £~1.
4-41
-------
Figure 4-13. The response of aquatic systems to atmospheric deposition
of acidic and acidifying substances from local or regional
(long-range) sources. The numbers refer to the following
references, which conclude that acidic deposition has
caused acidification of aquatic systems.
1. Gordon and Gorham (1960), Beamish et al. (1975), Dillon
et al. (1978)
2. Bobee et al. (1982)
3. Watt et al. (1979, 1983, Wiltshire and Machell (1981)
4. Davis et al. (1978), Norton et al. (1981b)
5. Schofleld (1976c), Pfeiffer and Festa (1980),
Galloway and Dillon (1982), Galloway et al. (1983c)
6. Shaffer and Galloway (1983)
7. Gordon and Gorham (1963)
8. A. H. Johnson (1979)
9. Burns et al. (1981), Johnson et al. (1981)
The letters refer to the following references where
possible acidification of aquatic systems has been studied
but not found.
A. McCarley (1983)
B. Schindler and Ruszczynski (1983)
C. Logan et al. (1982)
D. Zeman (1973), Feller and Kimmins (1979)
E. Melak et al. (1983)
4-42
-------
IQ
CD
I
t— •
co
-P.
CO
-------
atmospheric sulfur in aquatic systems, Figure 4-13 indicates (by numbers)
the areas that have been shown to be acidified. All numbers fall in
sensitive areas receiving acidic deposition. In addition to the studies
showing where acidification has occurred, the letters on Figure 4-13
represent studies where the atmospheric sulfur has been shown not to have
acidified the aquatic system because of its low concentration.
The maximum degree of acidification of freshwaters by acidic
deposition depends on the total increase in acid anions (primarily
S042-, Section 4.4.1). For each peq r-1 increase in S042', the
maximum loss of alkalinity 1s 1 yeq jr1 (Section 4.4.1). Studies
of $04 in aquatic systems across depositional gradients (Figure 4-2),
and sulfur budget studies for watersheds and lakes (Galloway et al.
1983c, Dillon 1981) and determination of excess S042' in aquatic
systems (Likens et al. 1977, Harvey et al. 1981, NRCC 1981) indicate
that the maximum increase in $64, and therefore maximum loss of
alkalinity in aquatic systems as a result of acidic deposition, is 100
peq JT • The actual loss will certainly be less and will depend on
the avail a- bility of base cations in terrestrial systems receiving
acidic deposi- tion. This maximum alkalinity decrease is merely a
boundary condition that can be compared to measured or estimated degrees
of acidification.
4.4.3.1 Analysis of Trends based on Historic Measurements of Surface
Water QuaTity (M. R. Church) —
4.4.3.1.1 Methological problems with the evaluation of historical
trends. In assessing the effects of acidic precipitation 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, conductivity, 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 are presented here.
4.4.3.1.1.1 £H
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
colorimetrically with acid-base indicators. Materials for visual
colorimetry are inexpensive and readily portable and, thus, highly
amenable to use in rugged, remote field locations, often the site of
"acidification" 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 (1928) for even more
exhaustive accounts, descriptions, and discussions of colorimetric
indicator use.
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
4-44
-------
HIn (Color A) £ In" (color B) + H+ . [4-10]
Indicators are used to measure the pH of an unknown aqueous solution as
follows. 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 different 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 transformation 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. (1983a) 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 unknown 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
,.„.,*„,,+ !.,£_+log JS- [4-11]
HIn
where pa^ is defined by Equation 4-2, pKuin is the thermodynamic
dissociation constant of the acid form or the indicator, is the
fraction of the indicator in the form In, and Yin andYHIp are
the activity coefficients of the dissociated and undissociated forms of
the indicator, respectively. Color matching (by eye or instrument)
indicates only that the term log a/l-a is the same for the unknown
and the standard solutions. However, if the activity coefficient ratio
(the last term in Equation 4-11) of the indicator is not the same in both
the standard and the unknown solutions, the pH of the solutions will not
be the same when the colors are identical. This is called the "salt
error" 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 actually 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 so it is close enough to the pH of the
unknown solution that no pH change occurs when the two are mixed. This
4-45
-------
can be accomplished by a trial and error technique using portions of the
sample to be determined plus a variety of indicators.
Bates (1973) indicates that when the above cautions are observed and
correction or adjustment is made for salt effects, an accuracy and a
precision of 0.05 to 0.1 pH unit 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, probably
seldom approach such accuracy or precision (Boyd 1977, 1980; Haines et
al. 1983a). Indeed, many details (including exact methpdology) of
historical pH measurements made with indicators are often "lost in
antiquity," lending further uncertainty to their reliability.
Fortunately, many of the investigators of acidification trends in surface
water pH values appreciate these 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 electrometrically (potent?"ometrically) both in the laboratory
and, with the advent of more reliable portable pH meters, in remote field
locations as well.
The "practical" or "operational" pH was defined in Equation 4-2 (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. [4-12]
The reference electrode is usually either a calomel or si Tver-siTver
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 articles, and laboratory manuals (e.g., Feldman
1956, Golterman 1969, Bates 1973, Durst 1975, American Public Health
Association 1976, Westcott 1978, Skougstad et al. 1979).
An important potential source of error in electrometric pH measure-
ments of surface waters is the residual liquid-junction potential.
Li quid-junction potentials arise at the point of contact of the reference
electrode and the solution being tested. Liquid-junction 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 measure-
ments 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
4-46
-------
(e.g., KC1) to the dilute test solution (this may Introduce Impurities,
thus altering the pH), 2} dilute the standard solution (which alters Its
pH--a correction must be applied), or 3) use dilute strong add standards
(these are not normally reliable pH standards--they must be frequently
calibrated by tHratlon) (Bates 1973, Galloway et al. 1979).
Another potential source of error 1n electrometrlc 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 samples (Galloway et al. 1979). To eliminate such
error, measurements should be made only on quiescent solutions.
Under rigorous conditions 1n a properly equipped laboratory, routine
electrometrlc 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—Inasmuch as both colorlmetrlc and electrometrlc measurements
(using secondary cells) are based on operational or practical pH
(designated by the primary pH cell and scale), the methods are directly
comparable. Attention has been, and should continue to be, placed on the
limits of reliability of the measurement methods as discussed above.
4.4.3.1.1.1.4 pH-general problems—Independent of the methodology
employed, 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 1s the variation 1n the pH of
surface waters over relatively short time Intervals. The most dramatic
and Important "short-term" changes 1n 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 add held 1n Ice and
snow (Wright 1977, Overreln et al. 1980, Galloway et al. 1980b, Hendrey
et al. 1980a). 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 we compare past and present pH
measurements in an effort to assess acidification.
Less Important but potentially meaningful effects are pH changes
associated with the uptake and release of 003 and/or HC03~ by
aquatic plants. Most lakes studied In conjunction with acidification
problems are usually ol1gotroph1c, and these changes are probably small.
Yet another factor to consider (especially In streams) Is the occurrence
of local sources of groundwater high 1n C02« One method sometimes used
to account for variable COg concentrations 1s to report the pH value
after a sample has been thoroughly agitated to equalize Its C02 partial
pressure with that in the laboratory. It must be noted, however, that
the CO? concentration in a laboratory can vary considerably from day to
day and 1s nearly always well above that commonly considered to be the
global mean (Church 1980). A nunber of methods may be employed to
4-47
-------
overcome this problem and to insure comparability 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 equilbrate all samples by bubbling with bottled air of a
known and standardized C02 content.
4.4.3.1.1.2 Conductivity
4.4.3.1.1.2.1 Conductivity 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. Conductivity is routinely reported as ymho cm~l at 25.0 C.
Detailed instructions 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 conductiv-
ity as low as those often reported in studies of acidification of surface
waters (American Public Health Association 1976, Skougstad et al. 1979).
Inasmuch as this figure pertains to measurements made under laboratory
conditions it is to be expected that measurements made with portable
battery-powered conductivity meters in the field would be less precise.
4.4.3.1.1.2.2 Comparability of early and current measurement
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.
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 1n conductivity
certainly may be associated with both increases and decreases in pH and
alkalinity. For this reason observed Increases in conductivity should
not be used by themselves to infer that acidification has occurred.
4.4.3.1.1.2.3 General problems--Conductivity can be expected to
vary seasonally (e.g., it may be much higher during snowmelt than at
4-48
-------
other times). Therefore, comparison of historical and recent
measurements to assess acidification should take Into account time of
year when the measurements were made. Temporary changes 1n 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 acidification as the decrease 1n 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,
National Research Council of Canada 1981).
4.4.3.1.1.3.1 Early methodology—Historically, acidimetrlc
titratlons have usually been performed to an endpoint of pH 4.5
determined electrometrlcally or to an end point determined by a
colorlmetric 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.
First, the assumption that the endpoint of pH 4.5 Is close to the
equivalence point for tltratlon of the predominating inorganic carbon
species Is true only for samples with relatively high total Inorganic
carbon content (I.e., ~ 2.5 mM total Inorganic carbon; Golterman 1969)
and ANC. For samples with relatively low total inorganic carbon and ANC
(like those usually considered 1n studies of surface water acidification)
the tltratlon endpoint should be at a higher pH (e.g., near pH 5.0 for
total Inorganic carbon of ~ 0.2 mM) to approximate the equivalence
point (Golterman 1969, American Public Health Association 1976, Dillon et
al. 1978, Ontario Ministry of the Environment 1979, Zimmerman and Harvey
1979, Jeffries 1980, National Research Council of Canada 1981). Methods
currently 1n routine use account for this fact.
Second, unless detailed notes have been kept of titratlons to some
endpoint determined with a colorlmetric 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 1s 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 Indicator Is added so they can tell by comparison to
the sample being titrated precisely when the tltratlon has reached the pH
that they have a priori selected as the endpoint. Unfortunately, many
early tltratlon 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 tltratlon.
4.4.3.1.1.3.2 Current methodology--Determining ANC of surface water
samples Is now commonly done by acldlmetric tltratlon to the (HC03~
- IF) equivalence point (inflection point) of the tltratlon curve.
4-49
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This point can be readily determined by using differential electrometric
titration methods or Gran's (1952) procedure (see Stumm and Morgan 1981).
ANC determined in this fashion is termed total inflection point
alkalinity or TIP (Dillon et al. 1978, Ontario Ministry of the
Environment 1979, Jeffries 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+) equiva-
lence 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 Comparability of early and current measurement
methods--As stated by the Ontario Ministry of the Environment (1979)
"Almost all past water quality surveys conducted on Precambrian
Shield waters have employed a TFE method [note added: i.e., to
~ pH 4.5] and hence, unrealistically high estimates of acid
buffering capabilities will be drawn from the data. This fact
is true no matter whether an extremely sensitive potentiometrie
TFE procedure was used or an insensitive field titration
(indicator, eye dropper titrant addition, etc.) method was
employed. The most appropriate use of the TFE alkalinity data
obtained from past surveys is (a) to define water systems which
are not acid susceptible, (b) to.suggest where further sampling
is warranted, and (c) in the case of high quality potentiome-
tric data, to infer relative levels of susceptibility for lakes
in the 0-20 mg a~l (CaC03) (0-400 yeq rl)
alkalinity range. The data base on "absolute" or correct water
alkalinity values is very small; there is a great need for
improving this situation as quickly as possible."
Discussions of the differences between TFE and TIP have been pre-
sented by Zimmerman and Harvey (1979), Jeffries (1980), and the National
Research Council of Canada (1981). As pointed out in some detail by the
National Research Council of Canada (1981), a rigorous correction may be
made from TFE to TIP in the situation where (1) TFE has been precisely
determined to a known endpoint pH and (2) the total inorganic carbon
concentration of the sample is known (or can be closely estimated).
Unfortunately, in most cases TFE has historically been determined using
colorimetric titration procedures, and total inorganic carbon concentra-
tions are not known. Conversion of such values to accurate TIP alka-
linities has proven to be nearly impossible (Jeffries 1980). Obviously,
such difficulties exist in comparing past and present data regardless of
whether the data come from Canada, Scandinavia, or the United States.
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,
4-50
-------
rainstorms, uptake of \\CQ^~ by aquatic plants) will not distort
evaluations of long-term trends.
4.4.3.1.1.4 Summary of measurement techniques. Each of the three
types of measurements (i.e., pH, conductivity, alkalinity) discussed here
has something to recommend its continued use in the study of surface
water 'acidification1. 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 in comparison to current pH data historical
measures of pH are somewhat unreliable, 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 'acidification1.
It is clear that in relation to historical data no overall best
analytical measurement of surface water chemistry exists for evaluating
acidification of lakes and streams. The authors recommend that pH,
alkalinity, and conductivity should continue to be routinely measured in
surface water acidification studies, taking into careful consideration
the detailed sampling and analytical techniques outlined in the articles
and manuals referenced above.
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 4-10 and 4-14 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 higher 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 "differences" 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.
4-51
-------
4-52
-------
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 information presented in the original
technical report or journal article is reviewed here. In some cases the
information presented by the original authors does not demonstrate
"beyond a shadow of a doubt" that their sampling was completely unbiased.
But this does not mean then that their sampling was necessarily biased,
and it is not the duty or intent of this reviewer to focus unduly on such
omissions or to speculate irresponsibly on their importance. Major
critical discussions are presented here only on important points of
reasonable debate for which sufficient information was presented by the
authors.
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 1n lake pH in North America attributable to acidic
precipitation. They studied chemistry changes and loss of fish
populations 1n lakes of the La Cloche Mountains, an area that has
quartzlte geology and that receives acidic precipitation. The acidity of
the precipitation is directly attributable to smelters at Sudbury,
Ontario, 65 km to the northeast. During the period of their study
(1969-1971) 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 oligotro-
phic lake 1n 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 (pre 1968) data were
available from studies performed 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 1n 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 degassing). In studies prior to 1968 the Ontario
Department of Lands and Forests measured pH with a HelHge comparator
(Beamish and Harvey 1972). 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
4-53
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that for lakes In and to the east of the La Cloche Mountains pH had
decreased with time (Table 4-4). For 11 lakes sampled prior to 1961 H+
concentration had increased 10- to 100-fold by 1971. The average change
in the mean annual 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 experienced lesser decreases in pH
(Table 4-5). 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).
South-Central Ontario (Beamish et al. 1975).
Beamish et al. (1975) reported on the relationship between various
fish populations and water chemistry in George Lake, Ontario, for the
period 1967-1973. In that report they cited evidence for a trend of pH
decrease in the lake.
Over the period 1968 to 1973 they measured pH electronic trie ally 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. Discarding the 1961 data point, 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 precipitation 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 x,"1).
This was a decrease from a TIP alkalinity of 33 (yeq &"1) 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 (yeq r*1) for the epilimnion
and 130 (yeq &-1) during a non-stratified period for Walker Lake in
1976. These were decreases from TFE values of approximately 180 (yeq
A'1) during 1974 (from unpublished data of the Ontario Ministry of
the Environment) and approximately 400 (yeq £-1) during 1971
(several samples on a single date; Michalski 1971).
The authors did not find any noticeable differences between the TFE
alkalinities of Harp Lake (137 to 152 yeq r-1) or Jerry Lake (137
4-54
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TABLE 4-4. EARLIEST AND 1971 pH MEASUREMENTS ON LAKES IN AND TO THE
EAST OF THE LA CLOCHE MOUNTAINS (BEAMISH AND HARVEY 1972)
Lake
Broker3
Carlyle
David
Free! and
George
Grey3
Johnnie
Kakakise
Killarney
L * F 24
Lumsden
Lumsden II
Lumsden III
Mahzenazing3
Nellie
Norway
O.S.A
Spoon3
Sun fish3
Threenarrows
Township
Attlee
Carlyle
Stalin and Goschen
Klllarney
Killarney
Sale
Goschen and Carlyle
Killarney
Killarney
Carlyle
Klllarney
Klllarney
Klllarney
Carlyle and Hunboldt
Roosevelt
Klllarney
Klllarney
Kilpatrlck and
Humboldt
Hunboldt
Klllarney, Roosevelt,
and Stalin
Date
Sept/61b
Aug/71
May/68b
Aug/71
^ * •
Aug/61b
Aug/71
June/69
SepV71
•»••— r ~ L.
Sept/61b
SepV71
Sept/59b
SepV71
Aug/61b
Aug/71
June/68b
Aug/71
Aug/69
Sept/71
w
Sept/67b
Aug/71
Sept/61b
Aug/71
June/69
OcV71
June 69
Oct/71
Sept/61b
Aug/71
Sept/69
Aug/71
Sept/69b
Aug/71
/61b
SepV71.
Sept/61b
Aug/71
Sept/61b
Apr/71
T * •
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
4-55
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TABLE 4-4. CONTINUED
Lake
Tyson9
Unnamed Lake3
(46001I30"NS1°24'W)
Mean of 22 lakes
Township
Sale and Humboldt
Killarney
Date
Aug/55b
June/69
Oct/71
pH
7.4
5.7
5.2
Avg
annual
change in
pH units
-0.16
-0.25
-0.16
aLocated east of the La Cloche Mountains.
bpH determined by the Ontario Department of Lands and Forests.
4-56
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TABLE 4-5. EARLIEST AND 1971 pH MEASUREMENTS ON LAKES NORTH OF THE
LA CLOCHE MOUNTAINS (BEAMISH AND HARVEY 1972)
Lake
Anderson
Annie
Bear
Brazil
Deerhound
El i zabeth
Frank
Fox
Griffin
Hannah
Han wood
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
Gosch
Merrit
Foster, Truman,
Curtin and Roosevelt
Roosevelt
Curtin
Roosevelt
Roosevelt
Truman
Louise and Dieppe
Eden, Waters, and
Broder
Merritt and Foster
Roosevel t
Foster
Date
Aug/60a
Oct/71
/61a
Aug/71
Aug/68a
Aug/71
Aug/67a
Aug/71
Sept/68a
Aug/71
Sept/68a
SepV71
/60a
Oct/71
July/60a
SepV71
Aug/60a
Oct/71
Aug/68a
Aug/71
Aug/67a
OcV71
Aug/68a
Oct/71
Aug/67a *
OcV71
Aug/68a
OcV71
Aug/68a
Aug/71
July/68
May/70
Nov/69a
SepV71
Sept/68a
OcV71
Aug/68a
OcV71
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
4-57
-------
TABLE 4-5. CONTINUED
Lake
Simon
Spring
Stratton
Walker
WMtefish
White Oak
Mean of 26 lakes
Township
Graham
Merritt
Fo ster
Truman and Roosevelt
Whitefish Indian
Reserve
Tilton and Halifax
Date
Aug/60a
Sept/71
Aug/66a
Oct/71
Sept/68a
Aug/71
Aug/68a
Aug/71
Aug/60a
Oct/71
Nov/69a
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.
4-58
-------
to 168 yeq £-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 possible short-term variations in alkalinity.
Halifax, Nova Scotia (Watt et al. 1979).
Gorham (1957) 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 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" (Wattetal. 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 equilibra-
tion) (Gorham 1957). Watt et al. (1979) plotted 1977 pH values vs 1955
pH values (Figure 4-15) 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 significantly (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 1n 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 atmospheric transport) for an explanation of the
acidic condition of lakes in the Halifax area" (Watt et al. 1979).
Nova Scotia and Newfoundland (Thompson et al. 1980).
Thompson et al. (1980) reported temporal trends in the pH of Nova
Scotia and Newfoundland rivers. In their report they discussed data
4-59
409-262 0-83-9
-------
7.0
6.0
£ 5.0
Q.
4.0
3.0
3.0
4.0
5.0
pH 1955
6.0
7.0
Figure 4-15.
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).
4-60
-------
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 laboratory, at room temperature" (Thompson et al.
1980). Thompson et al. (1980) compared the discharges on days of sam-
pling 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 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, near mid-day,
during the simmer. Wallace-Pieman colorimetry was used to measure pH
"until the 1960's"; then pH was measured with portable meters. "The two
methods were found to agree within 0.1 pH units (sic)" (Davis et al.
1978).
The authors noted initially that the mean pH of 196 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 euthrophi-
cation beginning in the 1950's. The authors realized that these pre-
liminary results might have been affected by regional edaphic differences
4-61
-------
in lake types and also by differences 1n 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 v.alues 1937-42), the final (1974) pH
would be 5.79, an increase on acidity of 12.6 times. Using a t-test, the
authors also found that the 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 1.15 x 10-7 M yr-l ^+
concentration 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+
concentration) slope linearly in inverse proportion to the time interval
between each reading. Tliese 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 1n 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 yeq &'1 from 236 to 109 yeq A"1) 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).
4-62
-------
The authors concluded from their study that between the years
1937-74 H+ concentration in Maine lakes increased about 10"6 M 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.
The pH values of the recent samples were measured in the field with
(1) a portable pH meter with combination electrode, and (2) a Hell ige
color comparitor. 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
electrometrlc pH (Figures 4-16 and 4-17). They concluded that 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 (New Hampshire) (Hendrey et al. 1980b, Burns et al. 1981).
During 1936-39 the New 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 1n 1979). Mean H+ concentration was 0.22 (yeq
r1) in 1936-39 and 0.87 (yeq JT1) in the 1979 samples. A
t-test showed this increase to be significant at the p < 0.02. "However,
4-63
-------
8.5
8.0
1 7.5
S 7.0
i—i
Q£
O
S 6.5
o
O
6.0
5.5
I I
4.5 5.0 5.5 6.0 6.5 7.0 7.5
NEW COLORIMETRIC (pH)
Figure 4-16. Old lake water pH (colonmetric) vs recent lake water pH
(colorimetric). Adapted from Norton et al. (1981a).
4-64
-------
7.5r-
7.0
5 6.5
o
6.0
_
O
o
5.5
5.0 -
4.5
I I I I
I
4.5 5.0 5.5 6.0 6.5 7.0 7.5
NEW ELECTRODE (pH)
Figure 4-17. Recent lake surface water electrode pH vs recent colorimetric
pH.
4-65
-------
when the errors associated with comparing the colorimetric data to the
electrometric data are considered, the difference 1n 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).
Schofield (1976a) 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 lakest pH data exist from the period 1929-37.
Frequency distribution plots (Figure 4-18) of lake pH for the two data
sets illustrate the apparent pH decrease with time (Schofield 1976a).
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) (Schofield 1976b). Schofield (1976a) did not present any
information on sampling or analytical methodology for pH for the data
sets, stating only that they were "comparable data."
New York (Pfeiffer and Festa 1980).
In the simmer 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 Takes in 1979 to the
values of the period 1930-34.
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 (1981) 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."
4-66
-------
20
10 •-
1930's
r-n
8
20
10
1975
8
NO FISH PRESENT
FISH PRESENT
Figure 4-18.
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 (1976a).
4-67
-------
To minimize any potential bias In the comparison of pH measurements
over time, Pfelffer and Festa (1980) used only colorlmetrlc measurements
In their data analysis. They presented their results graphically (Figure
4-19). 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" (Pfelffer and Festa
1980). The authors did not explain the apparent differences between the
pH distribution of the lakes on which they reported and the pH
distribution of the lakes described by Schofield (1976a). Like Schofield
(1976a), however, they attributed the observed deterioration of water
quality to the acidic precipitation in the region.
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) 1n the New Jersey P1ne Barrens. Precipitation
In the area had a mean pH of 4.4 1n 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 1n 1978-79. The
US6S 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 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 4-6 and Figure
4-20) 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 1n the amount of precipitation were responsible
for the long-term trends. He concluded "it appears that the decrease 1n
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 post illation of an atmospheric source for the
increased H+."
Pennsylvania (Arnold et al. 1980).
In an effort to assess temporal changes 1n pH and alkalinity of
Pennsylvania surface waters, Arnold etal. (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
4-68
-------
5.5
6.0 6.5 7.0 7.5
DETERMINED COLORIMETRICALLY (pH)
8.0
Figure 4-19. Cumulative comparison of historic and recent pH values for a
set of 138 Adirondack lakes. Adapted from Pfeiffer and
Festa (1980).
4-69
-------
TABLE 4-6. REGRESSIONS OF STREAM pH ON TIME: N IS THE NUMBER OF SAMPLES,
r IS THE CORRELATION COEFFICIENT, AND P IS THE LEVEL OF SIGNIFICANCE;
an AND ai ARE COEFFICIENTS IN THE REGRESSION pH = an + aix,
WHERE x IS THE NUMBER OF MONTHS AFTER JUNE 1963 (A. H. JOHNSON 1979)
Data source
A yeq H+
per liter
(1963-
1978)
USGS data, 1963-78
USGS data + UP data3
USGS data, anion 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
^Includes 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.
4-70
-------
6
5
4
^6
5
4
I
OYSTER CREEK
•
A
• .
I I
MCDONALDS BRANCH
1960
1970
1980
Figure 4-20.
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.
4-71
-------
(generally within one mile with no major tributaries or Influences
between)." Of these 314 cases, 107 (34 percent) showed decreases 1n 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
recent1 was 6.94 (range 4.9 to 8.3). The mean change 1n pH was a de-
crease 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 (yeq a"1)
(range 100 to 4000 yeq A'1), and the mean of the "most recent" was
532 (yeq jr1) (range 40 to 3720 yeq A"1). The mean net
change was a decrease of 302 (yeq &'1) and the range was (-2100 to
+360 yeq A"1). The average time span between the "earliest" and
"most recent" 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, 1t seems clear that there Is a definite overall
trend toward Increasing acidity 1n 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 dfscuss 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. 1980). Yet another
consideration 1n studies such as this has been noted by Schofleld (1981);
"It 1s obvious that detection of significant, long-term pH changes 1n
acidifying systems, still In a bicarbonate buffered state, cannot be made
reliably because normal metabolism Induced changes 1n C02 levels would
likely obscure any pH change resulting from decreased alkalinity. Thus
Interpretations of long-term pH changes 1n 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 variety 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 evidence 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 measured the pH and alkalinity of a number of North Carolina
headwater mountain streams. Burns et al. (1981) resampled 38 of these
streams In 1979, attempting to discern any changes 1n stream chemistry
that might have occurred In association with the acidic precipitation
that falls 1n the area (weighted annual pH 4.7 to 5.2 1n 1955-56 and
< 4.5 In 1979). The data discussed by Burns et al. (1981) were also
presented and discussed by Hendrey et al. (19805).
4-72
-------
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 previous ones."
Significant differences existed in the analytical methods used for
the early and recent data sets. For the 1961-64 samples, pH was measured
with a Hellige colorimetric kit and alkalinity was determined by
titration to a colorimetric (methyl-orange) endpoint. For the 1979
samples, pH was determined electrometrically and alkalinity by Gran's
plots. The authors compared pH measurements by Hellige kit to those with
their pH meter and found that they "agreed to within + 0.15 of a pH unit"
(Burns et al. 1981). To correct for the possible ove7titration of
alkalinity (past the true equivalence point to some arbitrary endpoint)
and thus the overestimation of alkalinity for the 1961-64 samples, the
authors subtracted 32 (yeq jrl) from each of the historical values.
Unfortunately, this may not have been a valid procedure. Crude
calculations using the mean pH and alkalinity of the 38 early samples
indicate that the equivalence point for the alkalinity titrations is near
pH 5.0. A correction of 32 (peq jr1) assumes that the actual
titration endpoint was at pH 4.5. No records exist to indicate that this
was the case (Burns, personal communication). All that is known is that
a methyl-orange technique was used. The pKa of methyl-orange is 3.5 and
its transition range (yellow-orange-pink-red) is roughly pH 4.5 to 3.1
(Bell 1967, Golterman 1969). Not only is it impossible to know what
color (and thus pH) the original analyst used as an endpoint, it may be
that a color transition is not even observable at the pH 4.5 point
assumed by the authors. If, for example, the titration endpoint for the
early samples was actually pH 4.0, a correction of 100 (peq A'1)
would be required. This level of uncertainty in the correction between
techniques is quite large compared to the mean differences in alkalinity
found by the authors--!46 (yeq £-1) in 1961-64 compared to 80
(yeq £"i) by Gran's titration in 1979. This would seem to cast
doubt upon the authors' findings 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). The authors did not find a
significant temporal trend in pH (mean 6.77 in 1961-64 and mean 6.51 in
1979).
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
4-73
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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.
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 Mesozolc
granite and are subject to acidic deposition resulting from NOX and
SOg pollution generated in the Bay area (McColl 1930, 1981). Measure-
ments 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 4-21 and 4-22.
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 significantly on a seasonal basis, and
(3) yearly precipitation did explain a small percentage of the variance
in mean annual (H+) 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 similarly considered and dismissed as unimportant the
possible effects of atmospheric increases 1n C02.
McColl (1981) concluded from his analyses "It Is clear that the
(H+) of waters in both reservoirs has Increased since at least 1954, 1f
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 depositions, especially those resulting from
emissions of nitrous oxides by automobiles."
4.4.3.1.3 Summary—trends in historic data. Numerous studies have
examined temporal changes in surface water chemistry 1n areas that have
"sensitive" terrain and that receive precipitation more acidic than pH
4.7. A consistent (and sometimes major) drawback of these studies 1s a
lack of clear documentation of the "historic" data used. Often It is
unproven that these crucial data are unbiased, either by sampling or by
the analytical procedures used. Many authors recognize this problem;
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 necessary by the deficiencies of the only data set
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200
160
120
80
40
LEGEND •
- o HETCH HETCHY
• PARDEE
6.7
6.8
6.9
7.0
7.1
7.2
7.3
7.4
7.6
7.8
1955 1960 1965 1970 1975 1980
YEAR
Figure 4-21. Increasing acidity at Pardee and Hetch Hetchy, shown by
hydrogen ion activity vs year, for the period 1954-79.
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en
ro
o
o
(O
o
20
18
16
14
12
10
i i
1945 1950 1955 1960 1965 1970 1975 1980
YEAR
Figure 4-22. Decreasing alkalinity at Pardee, shown by alkalinity as
vs year, for the period 1944-79.
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available." However, In every case reviewed here, the scientists who
performed these studies concluded that pH and/or alkalinity decreased 1n
at least some of the waters studied.
Surely In regions where acidic substances are deposited there exist
lakes and rivers (1n otherwise undisturbed watersheds) that have not
experienced recent decreases 1n pH or alkalinity. Such an occurrence,
however, Is not a valid argument against the phenomenon of surface water
acidification. The body of evidence In toto from the studies reviewed
above Indicates that in some regions of acidic deposition, otherwise
undisturbed lakes and streams are being acidified. Particularly
noteworthy by Its absence Is any body of data Indicating consistent
decreases In alkalinity or pH of surface waters In otherwise unaffected
regions (see Section 4.4.3.3) jiot receiving acidic deposition.
This reviewer Is unaware of any natural process that would cause
decreases In pH and/or alkalinity at the rates Indicated by the studies
(of apparently otherwise unaffected regions) reviewed here. Until
appropriate evidence 1s presented In support of some such natural process
or until some other reasonable explanation of the data presented above Is
put forth, the only logical conclusion Is that acidic deposition 1n these
regions Is causing acidification of some surface waters. Furthermore, 1t
Is only reasonable to assune that In regions of similar sensitivity that
receive similar levels of acidic deposition other surface waters are
being acidified.
4.4.3.2 Assessment of Trends Based on Paleollmnologlcal Technique (R. B.
Davis and u. 5. Anderson)—
To assess the Impact of acidic precipitation and associated
pollutants on lake ecosystems, scientists have begun to analyze the
record contained In the lake sediment (Norton and Hess 1980, Davis et al.
1980). The sediment contains a diversity of physical, chemical, and
biological evidence which starts thousands of years ago deep In the
sediments and proceeds upward toward the sediment surface to cover the
period of the Industrial revolution and recent technological activities.
By applying paleollmnologlcal techniques Including the dating of the
sediment (B1rks and Blrks 1980), 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
1n that each 1s more or less restricted to a different pH range. By
careful study of these pH relationships for present-day diatom
assemblages, 1t Is possible to calibrate the sedimentary diatom record so
that the past pH of lake waters can be Inferred. Thus, a dated record of
lake acidification can be constructed by studying sediments cores.
The paleollmnologlcal approach 1s useful for assessing the Impact of
acidic precipitation, because for the vast majority of acidification-
susceptible lakes no record of past, direct pH measurement exists. Where
such direct data exist, they (1) postdate 1920, (2) are usually only for
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one year or a short series of consecutive years, (3) are ordinarily only
for mid-summer when pH's are highest, and (4) are usually made by
colorimetric pH indicators (pre 1965) which themselves may alter the
pH of poorly buffered waters. Paleolimnological pH reconstruction
provides a nearly continuous record based on a single technique, from the
present to the past (including pre-1920). It solves the difficult
problem of the 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 nunber of years' deposits.
4.4.3.2.1 Calibration and accuracy of paleolimnplogical reconstruction
of pH history^Davis et al. (1983) have been calibrating the sedimentary
en atom record of pH by deriving "transfer functions" (Webb and Clark
1977) from the study of subfossil diatoms in surface-sediments (uppermost
0.5 or 1.0 cm) from the deepest part of 31 lakes in northern New England
and 36 lakes in Norway. Davis et al. have developed regression equations
relating these subfossil diatom assemblages to pH of the surface waters
in the lakes. The regression coefficients are used as transfer functions
to infer down-core pH. These regressions have standard errors (se)
ranging from + 0.23 to + 0.54 pH units. The errors for the New England
data are greater than th"ose for Norway, partly because the pH readings in
New England are limited to mid-simmer. Regressions on Nygaard's (1956)
alpha index, based on Hustedt (1937-39) pH preference categories, provide
less precise pH inferences, especially for lakes pH < 6.2. This probably
is a result of the semi-quantitative nature of HustecTt's categories, the
uncertainty in assigning individual taxa to categories and possibly
additional uncertainties. Several factors responsible for variance in
the surf ace-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 Cl) 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 surf ace-sediment data would suggest.
4.4.3.2.2 Lake acidification determined by paleolimnological
recons true 11 on". Quantifying this pal eol imnological approach and applying
it to lakes affected by acidic precipitation are quite recent techniques.
The methods are time-consun ing, and pH reconstructions have been
completed for only a small number of lakes. In southern Norway, recon-
structions for seven acidic (pH < 5.7) lakes indicate that acidification
started between 1850 and 1930 (different dates at different lakes) and
that the total decrease in pH by 1980 was 0.06 to 0.83 units (depending
on lake; average decrease 0.40 units) (Davis et al. 1983). Before this
acidification, these lakes were "naturally" all quite acidic (pH 5.0 to
6.0) and were highly susceptible to further acidification. 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 Harvatten 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.
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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 0.20 to 0.35 units (depending on lake; average decrease 0.26
units). However, in at least three cases the pH decrease may have
resulted in part from a recovery from an earlier, mild eutrophication
(and elevated pH) associated with limbering or other disturbance. Del
Prete and Schofield (1981) report a pH decrease of 0.6 units for another
Adirondack lake, but this is based on only one sample (0 to 1 cm vs all
deeper samples), and the deeper sediment is not dated. In two other
Adirondack lakes, the diatom record indicated no pH change.
To clarify these relationships, Davis et al. are continuing
paleolimnological studies of pH change in acidification-susceptible lakes
in the northeastern United States. An impediment to this research in the
United States is the scarcity of recently well-monitored (for pH)
acidification-susceptible lakes. This hampers efforts to develop more
precise transfer functions for inferring past pH.
4.4.3.3 Alternate Explanations for Acidification-Land Use Changes (S. A.
Norton)—
Land use changes may directly affect the pH (and related chemistry)
of surface waters in a number of ways, including variations in the
groundveter table; accelerated mechanical weathering or land scarifica-
tion; 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 conditions. 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)a or FeO(OH) or Fe20a + 2H+ + S042"
02 + H20 -> Mn02 + H2X
Organic matter + Decay •*• NOs" + 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
disturbances of the land surface. The exposure of relatively unweathered
material 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 witn cation
concentrations. This results in eutrophi cation trends in downstream
4-79
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waters (Pierce et al. 1972). Readjustment of the system may take
decades, with concurrent long-term changes in surface water chemistry,
including pH.
4.4.3.3.3 Decomposition of organic matter. Long-term trends in the net
production of biomass result in near steady-state chemistry for aquatic
ecosystems but, depending on the direction of the trend (aggrading or
degrading), the water quality parameters have different values (Nilsson
et al. 1982). The net loss of organic matter generally results in
accelerated production of nitric acid, C02, and increases in
cations—all other conditions being kept the same. However, a change in
stored biomass is generally accompanied by other changes such as changes
in canopy interception of aerosols, changes in evapotranspiration, or
changes in surface water temperatures, so the individual effects are
difficult to sort out.
4.4.3.3.4 Long-term changes in vegetation. Long-term changes in
vegetation bring about various physical and chemical changes in the soils
and watershed which result in long-term changes in surface water
chemistry. For example, Harriman and Morrison (1980) have demonstrated
that spruce reforestation in Scotland resulted in acidification of
streams and increased export of cations. It is not clear whether this is
due to indigenous tree-related processes as compared to the pre-existing
peaty soil vegetation or is due to changes in aerosol capture of acidic
components or to changes in hydrology.
Certain vegetation types (e.g., conifers) produce abundant humic
material which can produce acidity. Thus, the appearance of these
vegetation types in a succession could yield long-term declines in pH as
well as dissolved organic material concentrations. The appearance of
Sphagnum sp. because of changes in the moisture regime could also result
in acidification of surface waters due to the highly effective cation
exchange capacity of Sphagnum with associated release of H'1". Malmer
(1974) reviewed the Swedish literature relating to reversion of farmland
to forests and finds that the chemical changes (increased organic
content, lower pH, lower exchangeable metals) are the same as those that
have also been attributed to acidic precipitation.
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 phosphate). This potential acidification is generally
recognized and the affected soils are amended with a base, CaCOa, with
subsequent elevation of pH. In regions where agriculture is on the wane
and reforestation is underway, 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—effects of land use changes or acidification.
Drabltfs 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. Thus, although
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changing land use may locally alter the pH regime of lakes and streams,
It appears clear that regional lake acidification and episodic pH
depression occur 1n response to Increased atmospheric deposition of
strong acid, primarily ^$04.
4.4.4 Summary—Magnitude of Chemical Effects of Acidic Deposition
(J. N. Galloway)
The aquatic systems that are most likely to be influenced by
atmospheric deposition are those with alkalinity of less than 200 peq
&"1. Large areas of Canada and the United States contain such
systems. For example, approximately 80 percent of New England, by virtue
of the geology, has surface waters with less than 200 ueq £-1.
Eastern Canada provinces range from 90 percent (Quebec) to 20 percent
(New Brunswick).
Of the aquatic systems that are potentially susceptible to
acidification (Figures 4-4 to 4-7), only ones located in eastern North
America and small regions of western North America are receiving acidic
deposition (pH^B.O; Figure 4-11; see also Chapter A-8, Section 8.4).
Acidification of aquatic systems receiving acidic deposition has been
noted in several instances (Figure 4-13).
Acidification of aquatic systems and acidic deposition is supported
by the following lines of evidence:
o Due to acidic deposition, $04 concentrations have increased in
aquatic systems in most of eastern North America. The increase
in $04 has to have been matched by an increase in Cp or
H+. Since aquatic systems with original low alkalinities are
characterized by watersheds with low CB/H+ ratios in the
soil, a large portion of the increase in $04 will have to be
matched by an increase in H+, i.e., decreased alkalinity.
o Although there are problems with comparing old and" new data,
overall, the analysis of temporal records shows decreases in
alkalinity and pH in aquatic systems of eastern North America
receiving acidic deposition.
0 The limited application of paleolimnologic indicators shows
acidification of aquatic systems.
o Acidified aquatic systems are only found in areas receiving
acidic deposition (pH £ 5.0). In areas not receiving acidic
deposition (pH >_ 5.0), acidification of sensitive aquatic
systems is not found.
0 No other possibilities exist to explain the regional scale of
acidification that has occurred. For example, changing land use
is at times advanced as one explanation. However, in areas with
identical changes in land use, it is only those areas receiving
acidic deposition that are acidified.
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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 the northeastern United States and southeatern
Canada this is about 100 yeq «,-!. The actual decrease in alkalinity
depends on how much of the increased S042" is balanced by increases in
base cations. One estimate (Henriksen 1982) is that for a 100 yeq fc-1
increase in S042- and N03~ there will be an approximately 60 yeq £-1
decrease in alkalinity. The pH change associated with an alkalinity
decrease of 60 yeq Jr* can range 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.
The time scales of long-term acidification are on the order of years
to decades in areas with low $04 adsorption capacities (Northeast and
North Central United States) and decades to centuries in areas with high
$04 adsorption capacities (Southeast United States). If acidity of
deposition decreases, it is reasonable to believe the time scales of
recovery will be of similar magnitude (Galloway et al. 1983a) .
In addition to long-term acidification (years and decades) by
H2S04» short-term acidification (days to weeks) occurs as a result of
the combined action of ^$04 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 jr1 and reduction of pH from 7.0 to
4.9 have been reported due to the action of both $042- and
4.5 PREDICTIVE MODELING OF THE EFFECTS OF ACIDIC DEPOSITION ON SURFACE
WATERS (M. R. Church)
The predictive modeling of the effects of acidic precipitation on
the chemistry of natural waters is an extremely complicated task requir-
ing 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
possess no time component; they cannot predict the time required to reach
a new output level .
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Mechanistic models, of course, have a different set of pros and
cons. The principal attraction of mechanistic modeling 1s that 1f
accurate mathematical representations of all (or the most Important) of
the phys1cal/chem1cal/b1olog1cal 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, 1s that the
knowledge may not exist to allow formulation of accurate representations
of all (or even the most Important) phys1cal/chem1cal/b1olog1cal
processes of Interest. Second, mechanistic models (especially of
lake-watershed eco-systems) 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 predlctlvely, 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, 1s somewhat
Involved and time consuming. The last drawback to the mechanistic
approach 1s 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, 1982; Chrlstophersen and Wright 1981; Thompson 1982;
Chen et al. 1982; Chrlstophersen 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 1s
beyond the scope of this chapter. Instead, a fairly complete yet brief
review 1s presented of those three empirical models that are, so far, the
best known and most referenced of existing approaches.
4.5.1 Almer/Dlckson Relationship
Aimer et al. (1978) plotted lake pH vs lake sulfur loading (gm S
m"z yr"1 "concentration of 'excess1 sulfur multiplied by yearly
runoff) for Swedish lakes. They found "tltratlon curve"-type patterns
for data from sets of lakes occurring In areas of similar bedrock. They
plotted two curves (Figure 4-23): One for waters "with extremely
sensitive surroundings" and one for waters with "slightly less sensitive
surroundings" (Aimer et al. 1978). The authors did not define any
4-83
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7.Or
Figure 4-23.
6.0
5.0
4.0
0 1 2
EXCESS S IN LAKE WATER (g nT2 yr"1)
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 "excess" sulfur multiplied by the yearly
runoff. Adapted from Aimer et al. (1978).
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objective method for classifying lakes with regard to their surroundings
and responses to sulfur loadings (e.g., "extremely sensitive" or
"slightly less sensitive"). 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 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 4-23 has led to
several mislabelings of reproductions of the original figure (e.g., Glass
1980, Loucks et al. 1981, U.S./Canada 1981). If Figures 4-24 and 4-25
(adapted from Aimer et al. 1978) can be compared" (note 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
4-23 from a representation of "excess S in lake water" to some more
primary or causative factor (e.g., area! rate of total excess sulfur
deposition, areal 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
concentrations 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 4-23, 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 wouTd "return" along the curves of
Figure 4-23 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.
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1.5
Figure 4-24.
Atmospheric load of "excess" sulfur from precipitation and
dry deposition, 1971-72 (g S nr2 yr~l). Dry deposition
calculated from a deposition velocity of 0.8 cm s"l.
Adapted from Aimer et al. (1978).
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Figure 4-25.
"Excess" sulfur in lake water per year (g S m~2 yr"1).
(Concentration of "excess sulfur multiplied by the yearly
runoff.) Adapted from Aimer et al. (1978).
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As a minimum condition, before the Aimer/Dick son 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.
4.5.2 Henriksen's Predictor Nomograph
The contributions of Henriksen (1979, 1980, 1982) 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
4-26) 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 (1982) 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, 1982)
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 bound to organics; thus ionic concentrations of excess Ca
plus excess Mg+2 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
4-88
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300
o- 200
-------
for use of the nomograph. He has subsequently studied this question in
more detail, using data from lakes in North American and Scandinavia
(Henriksen 1982).
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 magnesium in certain waters. Unfortunately, he found no
clear cut answer to the question. In some cases, there was evidence of
increase in base cation concentrations (up to 0.63 for Lake
Rishagerodvatten, Sweden). In other cases} there was none. In an effort
to overcome these difficulties and conflicting data, Henriksen (1982)
used his best judgment to designate a maximum value of "base cation
increase factor" of 0.4 yeq (Ca* + Mg*)/yeq SO**. That is, for
every yeq JT1 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 (1982) 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
purposes. 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) and Haines et al. (1983b) in
attempting to apply the nomograph to lakes in Quebec and New England,
respectively, should be noted in this regard. It may very well be that
4-90
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the predictor nomograph will have to be modified to accommodate local
relationships for whatever region for which application is attempted.
4.5.3 Thompson Cation Denudation Rate Model (CDR)
As seen in the previous discussions of the.Almer/Dickson
relationship and the Henriksen 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 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 constant bicarbonate concentration.
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 at 1 m yr"1, then an
equivalent representation applicable to lakes or streams is generated
(Thompson and Hutton 1981, Thompson 1982) (see Figure 4-27).
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 assumption relates
to both the constancy and quantification of PQQ2 in any set °f waters
to which the model may be applied. Significant variations in C02
partial pressures in surface waters are well known.
Yet another point worth considering is the fact that Thompson (1982)
tested this approach in some highly colored lakes and rivers of Nova
Scotia (Figure 4-28). 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." A way to resolve this
question is through Gran titrations for weak and strong acids.
Apparently, such a study has not been conducted. The CDR model has yet to
be thoroughly verified with any other data sets.
4-91
409-262 0-83-10
-------
10
-2.5
PC02
RUNOFF = 1 m yr
-1
Intercepts are at
£.5 and 10
Figure 4-27.
ACID LOAD (meq nr2 yr'1) or EXCESS S042' (yeq jf
A plot of the model that relates pH and sum of cations to
excess $04 - in concentration units, or pH and CDR to rate
of excess $04 - loading in rate units.
4-92
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WALLACE.
200
CM
I
^ METEGHA
g LA HAV
PIPERS HOLE
ST.
TUSKEI.
N.E. POND, MEDWAYt
MERSEY- •
ROSEWAY. -
RUNOFF = 1 m
PpC02 = 2'5
0
EXCESS S042" (meq m'2
Figure 4-28. CDR plot for rivers with mean runoff near J m yr-1, 1973
excess $04 - loads, and mean or median river pH.
4-93
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4.5.4 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/watersheds respond to acidic deposition, and thus the ways in which
these responses 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; 5) the presence or
absence of hysteresis in those 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 three models discussed briefly here have 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+, S042-, N03", H+, and basic
cations. The previous sections have discussed the chemical effects
acidic precipitation has in aquatic systems by direct 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 1n
acidified surface waters have been demonstrated to be toxic to aquatic
biota (Chapter E-5, Section 5.6) 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 precipitation or, more broadly, atmospheric
deposition on metal mobility in aquatic ecosystems may be divided into
four areas:
4-94
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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 1n aqueous spedatlon of metals and consequent
biological effects.
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+, S042-, N03% and H+ In acidic deposition. Although
this excessive of metals Is apparently related to Industrial activities,
historic measurements of metals 1n 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 (which uses 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) Ombrotrophlc peat bogs, those having no source of water other
than precipitation, receive all their nutrients and
non-essential metals 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 1n 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 the precise calculation of absolute
increases of atmospheric deposition of these metals.
c) 'Continuously' 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
4-95
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deposition rates. Nonetheless, it 1s clear that the deposition
of Pb and Zn (fossil fuel-related elements) has greatly
accelerated over the last 100 to 150 years 1n 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 1n modern-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 than 1n older sediments and established that the
ubiquitous (1n northern New England) and essentially
synchronous (ca. 1860-80) Increases correlate with the initial
rapid Increase 1n the consuuptlon of fossil fuel in this
country. Because these lakes are relatively undisturbed, these
changes are interpreted to be caused by increases 1n the rate
of atmospheric deposition of these metals, starting prior to
1860.
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 (Lazrus et.
al. 1970). Groet (1976) demonstrated spatial variation in the
northeastern United States of concentrations of heavy metals 1n
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 period 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 1n the Utter layer), suggesting increased
deposition of Pb (Reiners et al. 1975, Hanson et al. 1982).
Although Pb can be removed mechanically by erosion andvertlcal
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 1n recent litter. Many workers
have demonstrated spatial and temporal trends for other
elements (e.g., Zn) which parallel those for Pb, but
quantitatively assessing increased deposition rates cannot be
done because of the nonconservative nature of the other
elements.
4-96
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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
precipitation 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
(1976a) 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 vegetation 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 indicate continuously increasing values for Pb for
all lakes for about 100 years. The values for Zn increase continuously
to the present for lakes with a pH > 6.0 and decrease in younger
sediments for lakes with pH < 5.5, suggesting recent acidification of
those lakes with decreasing Zn (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
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 precipitation, causing Al
leaching where before Al was accumulating.
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 precipitation (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 concentrations.
Accelerated cation release (from aquatic sediments) has also been
demonstrated during experimental acidification of surface waters. In the
field, Hall and Likens (1980) observed increased release of Al, Ca, Mg,
- 4-97
-------
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. (1980) 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 Mobil izati 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 is 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 immobil-
ization of both Pb2+ and S042~ as the insoluble salt, PbSO/j. Similarly
Nriagu (1973) has suggested that excess Pb2+ may immobilize P042~. 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 S042-. Increased A13+ in acidified soil waters could also
immobilze phosphate. Alternatively, desorption from or solution of FeOOH
from "B" soil horizons in well drained soils could liberate adsorbed
phosphate. These potentially important mechanisms have not been thoroughly
investigated in the context of acidic precipitation. Very probably P04
availability will be strongly affected by increased concentrations of
and Al3+ in soil and and 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 few HC03~, C032~, 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.
4-98
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4.6.2 Aluminum Chemistry in Dilute Acidic Waters (C. T. Driscoll)
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 (Blcornfield 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 (1970) report a
median aluminum value of 10 yg Al jT1 for terrestrial waters, while
Bowen (1966) gives an average concentration of 240 g Al -1 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
Robertson (1967) have shown that crystals having a diameter near 0.1 m
were relatively stable chemically. Filtration of samples through 0.4
m 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 soluble 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 4-7).
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. 1983). 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. 1983), basic
aluminum sulfate (Eriksson 1981) and exchange on soil organic matter
4-99
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TABLE 4-7. ALUMINUM CONCENTRATION IN DILUTE ACIDIC WATERS
o
o
Location
Lakes
Sweden
Norway
Scotland
Bel gi urn
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
Ontario various locations, 1980
Sudbury, Ontario
Adlrondacks, 1977-1978
Adirondacks, 1977
pH
4.0
4.2
4.4
3.5
3.9
4.1
4.3
4.0
4.4
Range
- 7.4
- 7.8
- 6.4
- 8.5
- 7.2
- 6.5
- 7.0
- 7.6
- 6.5
Al
ng
10
0
25
300
4
6
150
92
Range
Al a-1
- 670
- 740
- 310
- 8000
- 850
- 856
- 1150
- 1170
100 - 1000
Reference
Dick son 1978a
Wright et al. 1977
Wright et al. 1977
Vangenechten and
Vanderborght 1980
Driscoll 1980
Kramer 1981
Scheider et al .
1975
Driscoll 1980
Schofield and
Trojnar 1980
-------
TABLE 4-7. CONTINUED
I
I—1
o
Location Description
Streams (cont.)
USA Hubbard Brook stream order
average
pH
1
2
3
2
3
3
4
4
5
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 A-l
710 Johnson et al.
320 1981
210
200
150
190
180
160
150
230
Groundwaters
Sweden
USA
West Coast, 1977-1978
Hubbard Brook seepwater, 1979
3.8 - 5.7
4.6 - 6.5
100 - 2600
0 - 700
Hultberg and
Johansson 1981
Mulder 1980
-------
(Bloom et al. 1979). Johnson et al. (1981) and Driscoll et al. (1983)
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) $04, 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. Unfortu-
nately, 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 controlling 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. 1983).
4.6.2.2 Aluminum Speciation--Dissolved monomeric aluminum occurs as aquo
aluminum, as well as hydroxide, fluoride, sulfate, and organic complexes
(Robertson 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 et al. 1980, Driscoll
1980, Driscoll et al. 1983) 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. (1983) reported that organic complexes were the predominant form of
monomeric aluminum in Adirondack waters, on the average accounting 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--Pilute water systems are character-
istically low in dissolved inorganic carbon (DIG) due to limited contact
with soil. Because dilute waters are inherently low in DIG, they are
limited with respect 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 BNC of Swedish lakes. Waters were strongly
4-102
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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 reactions, and therefore decreased the buffering of
aluminum. Natural organic matter also participated in proton
donor/acceptor reactions; the extent to which total organic carbon (TOO
would dissociate/associate protons was 7.5 peq mg organic carbon-1.
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.
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 yeq £~* could
be attributed to dissolved aluminum and silicon, while 20 to 50 yeq
JT1 could be attributed to natural organic acids. Solution
titrations were characterized as having a major proton dissociation
constant (Ka) of 10~6 to 5 x 10'7, in addition to some less well
defined iom'zation 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 (1983) quantatively 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
4-29 for a hypothetical system that has some properties in common with
4-103
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Adirondack waters (Driscoll and Bisogni 1983). Buffering capacity is
defined as the quantity of strong acid or base (mols £~M which would
be required to change the pH of a liter of solution by one unit.
Conditions specified for the construction of Figure 4-29 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.
It is noteworthy 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 but 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. 1983).
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.
Johnson et al. (1981) observed a shift in aluminum speciation with
increasing stream order. Aquo aluminum and aluminum hydroxide complexes
decreased substantially with increasing stream order. Alumino-fluoride
complexes remained constant throughout the experimental reach.
Alumino-organic forms increased in concentration with decreasing
elevation.
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CQ
Q.
ALUMINUM
CARBONATE
ORGANIC SOLUTES
WATER
PH
Figure 4-29.
Buffer capacity diagram for dilute Adirondack water systems
(Driscoll and Bisogni 1983). Equilibrium with aluminum
trihydroxide (pKso = 8.49), organic solutes (CTorg =
2 x 10~5, pKorg =4.4) and atmospheric carbon dioxide
(Pco2 = 10~3-5 atm) were assumed.
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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 conditions, the water table was lowered and pools of
reduced sulfur within the soil become oxidized by molecular oxygen
entering the zone. Very low values (< 4.0) and very high aluminum levels
(> 40 mg Al £-1) 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 conclusively attribute groundwater 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 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
Waters—In acidic water systems conditions or supersaturation with
respect to aluminum trihydroxide have been reported (Driscoll et al.
1983). During conditions of supersaturation, aluminum will hydrolyze,
forming partial!ate aluminum oxyhydroxide. The acid-soluble aluminum
fraction mentioned earlier would include the mi croc rystal line 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.
Mi croc rystalline particles have a very large specific surface area and
may adsorb or co-precipitate organic and inorganic solutes. The cycling
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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. 1983, Hall et al. 1982) 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
AlpOs- 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), 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 yg P
£'*), 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 1n 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. (1983) observed that DOC was
removed from the water column of an acidic lake after CaC03 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 super
saturation 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- A1203/
water interface. He observed that natural organic matter adsorbs by
complex formation between the surface hydroxyls of alumina and 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 anticiapted for alumina. Organic coatings
may be important with respect to subsequent adsorption of trace metals
and anions.
Hall et al. (1982) observed a decrease in DOC levels of a third
order stream in New Hampshire after aluminum chloride (A1C13) addition.
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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 1983). DOC also serves as an
aluminum complex!ng 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 Orgam'cs (C. S. Cronan)
4.6.3.1 Atmospheric Loading of Strong Acids and Associated Organic
Microponutants--This first subsection deals with the association (but
not 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, phthalic acid esters, fatty acid ethyl
esters, and many other chemicals of industrial origin. Concentrations
ranged from one to several hundred ng £~1, with 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 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.
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These investigations by Alfhelm 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 add
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 surf Ida! 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 adds. Likewise, natural himic materials may
themselves have sufficiently low pka constants that they contribute to
the free acidity of surface waters. Organic weak anlons may be
particularly significant in providing ANC below pH 5.0, with the greatest
buffer intensity for the organlcs exhibited in the range of pH 4.5
(Figure 4-29) (Driscoll 1980).
The organic species responsible for contributing to the buffer capacity
of these soft water lakes Include a range of hydrophlUc and hydrophoblc,
low and high molecular we.ight compounds. These organic solutes may range
from simple carboxyllc acids like malic acid to complex poly phenolic
compounds like the model fulvic acid described by Schnitzer (1980). On
the average, these organic adds 1n natural waters might be expected to
contribute 5 to 10 yeq of anionlc 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). 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-Metalic Interactions—Acidification of surface waters may
affect metal-organic associations and 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 4.1, whereas it 1s 2.6 at pH 3.0;
likewise, the Zn2+-FA stability constant at pH 5.0 1s 3.7, but is 2.4
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 organlcally-complexed metals and
correspondingly higher 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 adds. This process could affect the adsorption
of humlc materials on mineral surfaces and could also affect Interactions
between hum1c/fulv1c monomers. The net result of this could be to
4-109
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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 significantly 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-AIuminum 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. There-
fore, 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 tefid to preserve the availability
of phosphorus in aluminum-rich waters.
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
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availability (Chapter E-5, Section 5-8). Traaen (1980) found that
organic matter decomposition 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 populations. 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 Additions
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 (CaC03)] 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 al. (1982) and Fraser and Britt (1982)
compiled a detailed review of base addition technology and effects which
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). There are many considerations in selecting a base
material to be used in neutralization. Scheider et al. (1975) have
summarized these considerations.
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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
predictable 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 is because atmospheric carbon dioxide will dissolve into solution,
neutralize the hydroxide, and eventually form a bicarbonate solution.
($2 introduction
Ca2+ + 20H" = Ca2+ + 20H" + 2C02 = Ca2+ + 2HC03-
Acidified waters generally have a low aqueous buffering capacity. As a
result, large 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 after strong base addition. This may result in
pH shock to organisms. These problems are accentuated within certain
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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 CaCOa 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 acidified surface water
applications, 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 neutralization 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 characteristic 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
CaCOa 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 CaCOa treatment than with strong base addition.
Sodium carbonate (NagCOa, 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 dissolution 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 (Ca(OH)2,
(see Section 4.7.1.2.2).
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Olivine (Mg, Fe)2 Si04, is a natural silicate mineral that has
been used in neutralization efforts (Hultberg and Andersson 1981).
Olivine is a continous reaction series in which magnesium and ferrous
iron can freely substitute for each other. Upon dissolution of Peg
$1*04, iron will oxidize and precipitate as Fe(OH)3 and thereby
contribute to the acidification 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
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 OePinto
1978). Basic fly ash has been shown to be effective in the neutraliza-
tion of acidified 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
mi dwe stern 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 contamination. 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
neutralization 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
(SiOo) oxides. While basic slags may contain similar calcium (CaO)
levels, dissolution rates and therefore neutralization characteristics
can vary considerably. The dissolution 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 determination 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
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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.
4.7.1.2 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 requirements. 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. For cost effective
use, a rational method for base dose determinations should be used;
however, to date none have 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] + 2[Al(OH)+2]
+ [A1(OH)2+] + 3[A1-F] + 3[A1-S04] + [RCOOH] + [H+]
- [AKOHU-] - [OH-]
where Al-F is the sum of all aqueous aluminum-fluoride complexes
(mols £-1),
Al-SO^ is the sum of all aluminum - sulfate complexes
(mols r1), and
RCOOH is the dissolved organic carbon base neutralizing capacity
(mols £-1).
Driscoll et al. (1983) 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 (1983) developed an empirical formulation
relating aquatic humus (dissolved "organic carbon, DOC) to the mols of
proton-dissociable organic acid/base:
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CT = 2.62 x 10-6 (DOC) + 7.63 x 10-6
where DOC is the dissolved organic carbon concentration (mg C £-1)
and
Cj 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:
CT x
[RCOOH] =
KJ. PUT I .
d L n J
Other metals (Cu, Mn, Zn, 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 concentrations are high. High concentrations would
most likely be found in acidified 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 jr1 (Scheider et al. 1975, Van and Dillon 1981).
If equilibrium with atmospheric carbon dioxide is assumed, an upper
limit of the aqueous BNC may be estimated. Driscoll and Bisogni (1983)
have made such an analysis to neutralize a "typical" Adirondack lake
(Table 4-8) to pH 6.5 (Table 4-9). It is apparent that a substantial
portion of the aqueous BNC is associated with the hydrolysis of
aluminum, and this should not be overlooked 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 acidification 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 neutralization, 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)2) of 477, 196 and 477 yeq
£-1 were applied to Middle, Lohi, and Hannah Lakes, respectively, in
the Sudbury region of Ontario, Canada (Dillon and Scheider 1983). Of
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TABLE 4-8. COMPONENTS OF BASE NEUTRALIZING CAPACITY IN
TYPICAL ADIRONDACK LAKE WATER
(ADAPTED FROM DRISCOLL AND BISOGNI 1983)
Parameter Value
pH 4.95
Inorganic Monomerlc Aluminum 0.2 mg Al &
Aluminum Fluoride forms 0.105 mg Al
Aluminum Sulfate forms 0.005 mg Al
Free Aluminum 0.04 mg Al J
A1(OH)2+ 0.03 mg Al J
AKOH)2+ 0.02 mg Al
TOC 5.0 mg C H'
TABLE 4-9. AMOUNT OF BASE REQUIRED TO NEUTRALIZE
BASE NEUTRALIZING CAPACITY OF
TYPICAL ADIRONDACK LAKE WATER TO pH 6.5
(ADAPTED FROM DRISCOLL AND BISCOGNI 1983)
Acid component Base required (eq £~
Hydrogen Ion 1.1 x (lO"5)
Carbonate 1.3 x (10~5)
Aluminum 2.0 x (10~5)
Organic Carbon 0.4 x (10~5)
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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 sediments from Sudbury lakes by titrating sediments with
Ca(OH)2 to a pH of 8.0 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
southeastern 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:
SDmax + ANCa
K + ANCa
where: SD is the sediment demand of base (meq m~2),
SDmax is the maximum demand of base (meq m~2) »
ANCa is the increase in water column ANC after base addition
(yeq r1), and
K is the sediment demand constant (yeq jr1).
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 simle 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)?, Na2COa)
a dissolution efficiency of 100 percent may be a reasonable assumption.
However, the dissolution efficiency of sparingly soluble bases (e.g.,
CaCOs, 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 CaCOa coated with
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organic detritus and metals within the sediments of a limed lake.
Conversely, Dillon and Schelder (1983) 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 neutraliza-
tion 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) has reported dose requirements of
7340 yeq CaCOa ha of lake surface area"1 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 acidified waters by
adding chemicals 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 1982, Sverdrup and Bjerle 1982) have developed a model to
describe CaCOs dissolution after application in acidic lakes. The
major parameters influencing CaC03 dissolution are particle settling
depth and solution characteristics. Sverdrup (1982) 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
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liming. Calcium carbonate resting on (or within) lake sediments has
very slow dissolution rates. This may be attributed to burial, limited
turbulence, or coating of CaC03 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 (1982) calculations suggest that CaCOs 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 CaC03 placement. If
CaC03 is placed in organic sediments, particles may become buried or
coated with metal and/or organic matter. If applied to the littoral
region, CaCOa should be dispersed so only a thin layer accumulates on
sediments. This will ensure that a large surface area of base directly
contacts the water and increases dissolution efficiency. Driscoll et
al. (1982) observed that when CaCOa 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 CaC03 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 determines 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 transferred to a boat. This 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 103 kg was applied in an average
working day.
Large amounts of powdered CaCQ$ 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 application. Transporting bagged lime by helicoper into
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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 103 kg
capacity) suspended from a helicopter. Upon flyi.ng 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) has 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.
Sodium carbnoate (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 4-10) and is therefore
economically competitive with more conventional basic materials (e.g.,
Ca(OH)2, CaCOs). Lindmark' s 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 acidified lake (Lindmark
1981). The sodium carbonate was mixed on land and pumped to a moveable
raft, and a land-based compressor that 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
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TABLE 4-10. CHEMICAL COST COMPARISON OF NEUTRALIZATION AGENTSa
Chemical
CaC03
Ca(OH)2
Ca(OH)2
CaO
Na2C03
(Mg,Fe)2Si04b
H3P04
Form Equivalent
supplied weight
(g eq-1)
bags (325 50
mesh)
bulk 37
bags 37
bulk 28
bulk 53
bulk (100 86
mesh)
agricultural 5.75C
grade (70%
solution)
Costa
Mass Equivalence
basis basis
(dollars x (dollars
10-3 j
-------
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 1981). 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 BMC (H+, Al) 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
CaC03) is a common agricultural practice. However, forest ecosystems
are considerably different than 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 (Ca, Mg, Na, K) and/or acidic cations
(Al, H). At near neutral pH values, the exchangeable cation pool is
largely comprised of basic cations. As soil pH decreases, the
exchangeable acidity (Al, H) is thought to increase. Another reaction
of interest is biological uptake of cations. Forest biomass requires
cationic nutrients (e.g., Ca, Mg, K) for growth. An aggrading forest
will assimilate basic cations and tend to deplete soil pools.
Cation exchange and biological uptake reactions are significant
considerations with regard to watershed liming. Forest soils are
generally nutrient poor and elevated in exchangeable acidity. Upon
addition of base [Ca(OH)2, CaCOs] 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.
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409-262 0-83-11
-------
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
neutralization 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 biogeochemistry. When an acidified lake
is limed, it will still experience the introdution of BNC (Al, 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).
Agricultural lime (powdered CaC03, 0 to 0.5 mm) has been transported
to the watershed in large trucks and applied as a slurry with a sprayer
truck. In this manner one man is able to apply 40 x 103 kg 9f CaC03
per day (Hinckley and Wisniewski 1981). The CaCOa dose required to
achieve adequate neutralization of water systems is generally two orders
of magnitude greater than that of direct water addition (Bengtsson et
al. 1980). This is 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 CaC03 ha~* yr-1. 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 1981). 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
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.
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Four approaches have been attempted to achieve stream
neutralization. The simplest approach is CaC03 addition to the
streambed (Hultberg and Andersson 1981). This has been attempted with
both coarse (5 to 15 mm) and fine (0 to 0.5 mm) CaCOa- Coarse CaCOs
will tend to stay in the stream bed, but neutralization is generally
inadequate because of the rather low surface area of the stone. Fine
CaC03 will more readily dissolve (due to a greater surface area) but
is more influenced by stream turbulence. Powdered CaC03 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 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 CaCOa 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 conditoins 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
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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 by direct 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.
0 Transparency increases immediately following base addition
especially in colored waters (Van and Dillon 1981, Hultberg and
Andersson 1981). This appears to be due to the removal of dis-
solved organic matter by co-precipitation with metals (Yan and
Dillon 1981). The long term consequence, however, has been the
reduced transparency in neutralized lakes. Decreases in
epilimnion thickness and decreased hypolimnetic temperatures
have been associated with these transparency changes (Yan and
Dillon 1981). Upon reacidification, transparency has
increased.
o 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 C02 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.
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When a slightly soluble base (e.g., CaC03) 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.
o 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 ^003 (C02[aq] + ^003) 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., CafOH^)
is added, the increase in inorganic carbon is due entirely to
atmospheric CO?, whereas when a carbonate base (e.g.,
03003) 1S added, inorganic carbon increases are due to the
base itself as well as atmospheric 003.
° 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 &"1 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, Yan 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. 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
4-127
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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 (CaCC^). 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 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. This 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 acidified 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 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 neutraliza-
tion efforts are listed in Table 4-10, 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, Na2COa) are undoubtedly the most
efficient means to add base, while slightly soluble bases (CaCOs,
MgFeSi04) and phosphorus (Section 4.7.2.2) are potentially less
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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-1 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 day1 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 day"1 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 per surface hectare. 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 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
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(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 acidified 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 prior to base addition of acidified lakes.
Base addition has become a popular strategy to mitigate water
quality problems associated with acidification. However, before base
addition is implemented as a regional, acidified lake management
alternative it should be more thoroughly evaluated.
4.7.2 Surface Vlater 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 effectively
evaluate 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 N0a~ + H2P04~ + 122 H20 + 17 H+ (+ Trace Elements, Energy)
photosynthesis + * respiration
C106 H262 °110 N16 p (alQal protoplasm) + 138 02
This 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 essentially 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, assimila-
tion 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., Ca£+ t
Mg2+, Fe). The uptake of cations by algal protoplasm would
diminish the quantity of ANC generated through photo-
synthesis.
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 vg of net phosphorus fixed by
algal uptake.
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The second reason for nutrient addition is to replenish the biomass
of acidified lakes. Hendrey et al. (1976) have suggested that
phytopiankton biomass is reduced by lake acidification. Dillon et al.
(1979) suggest that phytoplankton biomass is better correlated with
total phosphorus levels than 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
reasonably well understood in circumneutral lakes (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 phytoplankton 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
precipitation or adsorption reactions. Aluminum can directly
precipitate with orthophosphate to form A1P04 (varacite). 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 significant vertical diffusion of
phosphorus occurs. If fixed nitrate associated 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
4-132
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generation of ANC to the water column resulting). Schindler et al.
(1973) have indicated that nitrogen sedimentation and removal are less
efficient than phosphorous sedimentation and removal.
4.7.2.3 Fertilization Experience and Water Quality Response to
Fertilization—As mentioned previously there has been limited experience
with fertilization of acidic lakes. Most of the work has been accom-
plished by Canadian 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 ^04 is applied because of its
low cost, ease of handling, and solubility (Table 4-10). 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 addition. However, Dillon and Scheider (1983) observed
decreases in inorganic nitrogen (largely nitrate) and increases in total
organic nitrogen folowing nominal orthophosphate additions of 10 to 15
yg p £-1 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 yeq
£~1 greater than the control lake. In addition, the ANC generated
from nitrogen transformations declined dramatically after phosphorus
additions were terminated.
Observed changes in aquatic biota have been more significant.
Small additions of total phosphorus resulted in significant increases in
phytoplankton biomass of neutralized Canadian lakes (Dillon et al.
1979). No single observation 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 dominance 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
4-133
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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 in
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 (1981)
investigated nutrient addition as a means of supplementing liming
efforts in Sweden. They reported few results except for 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
critically assess 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 4-10) 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 (NAS 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 acidification
trends, of critical processes that control acidification, and of the
degree of permanency of biological effects remains incomplete and
subject to debate.
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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.
0 Each of several components of aquatic or terrestrial systems may
assimilate some or all acidic deposition falling in a watershed.
These components are vegetative canopy, soils, bedrock,
hydrology, wetlands, or an aquatic system itself (Section
4.2.1).
0 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).
0 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 meteorology (Section 4.3.2.4).
0 Alkalinity or acid neutralizing capacity (ANC) determines a
lake's instantaneous ability 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 very
important in determining the sensitivity of a lake to
acidification (Section 4.3.2.6).
0 Aquatic systems sensitive to acidification by acidic deposition
are commonly waters with pH and alkalinities towards the lower
end of the spectrum. The .boundary between sensitive and
insensitive that is used is 200 yeq £-1 of alkalinity
4-135
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(Section 4.3.2.6.1). This concentration is chosen because in
North America acidic precipitation has resulted in about 100
peq r-1 of potential acidification of surface water
(Harvey et al. 1981, Church and Galloway 1983), and because
biological effects due to acidification begin when aquatic
systems reach alkalinities of ~ 100 yeq £-1.
Regions in North America contain aquatic systems sensitive to
acidification. These regions are found throughout much of
eastern Canada; New England; the Allegheny, Smokey, and Rocky
Mountains; and the Northwest and North Central United States
(Galloway and Cowling 1978, 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).
Studies uniformly point to acidification of some surface waters
in eastern Canada and the northeastern United States (Section
4.4.3.1.2).
Although changing land use may locally alter the pH regime of
lakes and streams, it appears that regional lake acidification
and episodic pH depression occur in response to increased
atmospheric deposition of strong acid, primarily H2S04
(Section 4.4.3.3).
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).
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lakes with special reference to acid precipitation. Limnol. Oceanogr.
23:487-498.
Wright, R. F. and M. Johannessen. 1980. Input-output budgets of major
ions at gauged catchments in Norway, pp. 250-251. Jji Ecological Impact
of Acid Precipitation. Proceedings of an International Conference,
Sandefjord, Norway, March 11-14, 1980. Drabljos, D. and A. Tollan,
eds. SNSF-project, Oslo-As, Norway.
Wright, R. F. and E. Snekvik. 1978. Acid precipitation chemistry and
fish populations in 700 lakes in southernmost Norway. Verh. Internat.
Verein. Limnol. 20:765-775.
Wright, R. F., T. Dale, A. Henriksen, G. R. Hendrey, E. T. Gjessing, M.
Johannessen, C. Lysholm and E. Storen. 1977. Regional Surveys of Small
Norwegian Lakes. SNSF project IR33/77. Oslo, Norway.
Wright, R. F., N. Conroy, W. T. Dickson, R. Harriman, A. Henriksen, and
C. L. Schofield. 1980. Acidified lake districts of the world: a
comparison of water chemistry of lakes in Southern Norway, Southern
Sweden, Southwestern Scotland, the Adirondack Mountains of New York, and
South Eastern Ontario, pp. 377-379. ^Ecological Impact of Acid
Precipitation. Proceedings of an International Conference, Sandefjord,
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Yan, N. D. and P. J. Dillon. 1981. Studies of Lakes and Watersheds
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Van, N. D. and C. Lafrance. 1982. Experimental fertilization of lakes
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Ont. Min. of Environ.
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Van, N. D. and W. A. Scheider, and P. J. Dillon. 1977. Chemical and
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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 affecting 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 animals 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 distributions 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 acidification 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
5-1
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acidification is suspected to have occurred or 2) contemporary
comparisons of water bodies thought to exhibit different degrees of
acidification. Problems exist with this type of correlation approach.
For example, before and after studies may be difficult to interpret if
methodologies have changed in the interim, 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 intentionally 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 precipitation, 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).
5-2
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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 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 (J. J. Magnuson and 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 chapter'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. Methods for distinguishing between naturally acidic and
culturally acidified waters are discussed in Chapter E-4, Section 4.4.3.
5.2.1 Types of Naturally Acidic Waters
Naturally occurring acidic waters fall 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 (FeS?). 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
conditions 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
5-3
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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
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 brownwaters may mean that even physiologically
tolerant species may be excluded due to food unavailability (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 (HeCOa). These lakes tend to be small
and have low concentrations of dissolved ions (Chapter E-4, Section
4.3.2). 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 precipitation
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 Cynanidi urn
caldarium was found at pH 0.05, while the bacterium 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)
5-4
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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). Al though
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, Chironpmus
riparius, and the rotifer, Brachionus urceolaris, were abundant at pH
2.8, but no copepods or cladocerans were present.
Among the few insects reported from acidic thermal waters is the
ephydrid fly Ephydra thermpphila (Brock 1978). This fly breeds in
streams at pH 2.0 and is the basis of a 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 which may be tolerated by Morth
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 3.01Diatoms (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, Cnydorus, Macrocyclops) and a rotifer (Braehionus). The
teleost Tribolodon 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 outstanding examples of extreme pH tolerance, in general, these
waters have very low species diversity and monocultures of tolerant
species are common.
5-5
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5.2.3 Biota jn 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, characterized 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, Ceratpphyllum, Ispetes, Juncus, Limnobiuiri,
hluphar, Potamogeton and Utricularia (Jewell and Brown 1924, Griffiths
1973, Stoneburner and Smock 1980). Many 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 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 nannoplankton) in these waters (Bricker and Gannon 1976).
Although species from most phy topi ank ton phyla have been reported,
certain genera of desmids (Xanthidium, Euastrum, Hyalotheca) and diatoms
(As ten' one! la, Eunotia, Ac tin ell a, Anomoeoneis, Pinnularia, Melosira)
are especially characteristic (Woelkerling and Gough 1976, Marshall
1979, Patrick et al . 1979, Stoneburner and Smock 1980). As with the
phy topi ank ton, the zooplankton in acidic dystrophic lakes are frequently
dominated by small -bodied forms, particularly rotifers (Brachionus,
K era tell 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,
Chydorus, Daphm'a, Diaphanpsoma, Eubosmina, Leptodora, and Pleurpxus
(Marshall 1979, Von Ende 1979, Stoneburner and Smock 1980). In lakes
where fish are absent or where darkly stained water and low hypolimnetic
oxygen offer some protection from fish predation, dipteran larvae of the
genus Chaoborus are an important part of the zooplankton community (Von
Ende
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).
5-6
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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 damsel flies (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 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 enough 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
5-7
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sucetta). eastern mudminnow (Umbra pygmaea), pirate perch (Aphredoderus
sayanus), and the yellow perdu
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
(Mi cropterus do!omieui), and walleye (Stizostedion vitreum).
5.2.4 Biota in Ultra-Oligotrophic 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 phytopiankton, for example, crysophytes 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 (Ratalas
1971, Torke 1979). Factors like 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, Chaoborus) along with fingernail clams (Pisidium),
the amphipod Pontoporeia, and" the mysid My sis 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 concentrations, even though
the pH is circumneutral. Crayfish, for example, were absent from
softwater Wisconsin lakes having calcium concentrations below 2 mg
£-! regardless of lake pH (Capelli 1975).
5-8
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Aquatic macrophytes typical of oligotrophic waters have been
summarized by Hutchinson (1967) and Seddon (1972). Among the
representative genera are Bidens, Elatine, Eriocaulpn, Ispetes, Juncus,
Lobelia, and Sparganium. Most of these have a distinct physical 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 smallmouth bass, Micropterus dolpmieui) and rock bass (Ambloplites
rupestris) (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
bigger influence on the species composition than pH per se. For
example, many small-bodied fish species (e.g., minnows and darters) are
commonly absent from oligotrophic 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 only successfully coexist 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
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TABLE 5-1. LOWER pH LIMITS FOR DIFFERENT GROUPS OF ORGANISMS IN
NATURALLY ACIDIC WATERS
Group
Bacteria
Approx.
Lower
PH
limit
0.8
2-3
Examples of Species
Occurring at Lower pH Limit
Thiobacillus thiooxidans,
Suifolobus acidocaldarius
Bacillus, Streptomyces
Reference
Brock 1978
Brock 1978
Plants
Fungi
Eucaryotic
algae
Blue-green
algae
Vascular
plants
Mosses
Animals
Protozoa
Rotifers
Cl adocera
Cope pods
Insects
Amphipods
Clams
Snails
Fish
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
Aconti urn velatum
Cyanidium caldarium
Euglena mutabl1 i s,
Chlamydpmonas acfdophila.
Chi orel la
Eleocharis, Carex,
Ericacean plants,
Phragmites
Sphagnum
Amoebae, Heliozoans
Brachionus, Lecane, Bdelloid
Col1otheca, Ptygura
SimocephaTus, Chydorus
Mac rocy clops
Cyclops
Ephydra thermophila
Chironomus riparius
Mayflies
Gammarus
Pi si di urn
Most other species
Amnicola
Most other species
Tribolodon hakonensis
Umbra limi
Sunfishes (Centrarchi dae)
Brock 1978
Brock 1978
Brock 1978
Mastigocladus, Synechococcus 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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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 rankings of fish species,
based on distributions among naturally acidic Wisconsin lakes (Figure
5-1), were correlated with acid tolerance rankings from culturally
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
misinterpretation 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 concentrations, 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, acidic waters (below pH 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 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 laboratory 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, indicating 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 especially true of metal toxicants, which are common in waters
affected by acidic mine drainage or acidic precipitation (Parsons 1977,
5-11
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COMMON NAME
CENTRAL MUDMINNOW
YELLOW PERCH
BLACK BULLHEAD
BLUEGILL
LARGEMOUTH BASS
WHITE SUCKER
YELLOW BULLHEAD
PUMRKINSEED
GOLDEN SHINER
NORTHERN REDBELLY
BROOK STICKLEBACK
NORTHERN PIKE
WALLEYE
ROCK BASS
MOTTLED SCULPIN
SMALLMOUTH BASS
MUSKEILUNGE
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 > 70)
FAMILY 7.
P
Ce
Ce
Ca
Ce
Cy
s*
E
P
Ce
Co
Ce
E
Ce
Ga
iy
P
P
Ca
Cy
Cy
Pe
V
Cy
Cy
pH RANGE NUMBER OF
0 6.0 5.0 4.0 LAKES
.ii.
, 1 . 1 L
1 1 1 ._ 1 . 1 ' -
50
114
50
84
80
83
35
78
63
13
10
65
51
56
29
44
40
59
25
13
16
24
37
23
31
22
10
45
15
15
7
23 ' 17 17 15
Figure 5-1. The distribution of 31 fish species in relation to pH for
138 northern Wisconsin lakes. Family names are abbreviated
as follows: Catostomidae (Ca), Centrarchidae (Ce), Cyprinidae
(Cy), Esocidae (E), Gadidae (Ga), Gasterosteidae (G),
Ictaluridae (I), Percidae (P), Percopsidae (Pe), Salmonidae
(S), Umbridae (U). Adapted from Rahel and Magnuson (1983).
5-12
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V
_*
<0
XJ
o>
-------
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 spring fish kills in lakes affected by acidic precipitation (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
especially 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 (Percindae)
below pH 6.0
o several species of sunfish (Centrarchidae) below pH 4.5
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:
0 too low if other stresses (e.g., aluminum) are present in
culturally acidified lakes, or
o 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.
5-14
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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,
macroinvertebrates, Crustacea, Insecta, Mollusca, and Annelida (Table
5-3). 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 precipitation 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 partially 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
microbial 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), dissolved 02, and other constituents (Hutchinson
1957). Losses or alterations 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. (1981b)
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. 1981b). 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.
5-15
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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 0kland
1980); however this prediction has not been verified. Rosseland et al.
(1980) reported that trout in acidified waters shifted their diet from
acid-sensitive invertebrates like 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 sensitive 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 investigated, but Eriksson et al. (1980b) 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 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.
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 is
one of the last biological effects of acidification, and 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 (pH
< 6.5) 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.
5-16
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Using historical collections and known water quality requirements
of organisms 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 fluctuations in past years.
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 cautioned (1977) that many other factors besides 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
scrapping insects in an acidic (pH 4.3 to 5.9) 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 Verticci 1980) due to the senescent, colorless state of the common
blue-green algal (Phormidium spp.) component of the mat.
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
5-17
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changes in the planktonic bacterial populations from seven lakes of pH <
5.0 as compared to seven lakes of pH > 5.0. Traaen argued that acidic
inputs should affect the plankton populations prior to affecting benthic
algae. His results showed that the distribution of bacterial
populations was more strongly influenced by organic inputs and temporal
and spatial (depth) patchiness than by pH. Gahnstrom et al. (1980)
reported that inhibition of oxygen uptake by sediments increased in
acidic lakes as compared to reference lakes only in the littoral
sediments. They argued that the inhibition 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 necessarily 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 Periphyton--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 which are a principal source of food for fish. Algal
seasonal growth and decomposition store and periodically release
nutrients and other ions.
5.3.2.2.1 Field surveys. Acidic lakes develop periphytic communities
dominated 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 chemically isolates the lake bottom from the
overlying water. The mat varies 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):
1) Cyanophycean mats, dominated by the blue-green algae
pscillatoria sp., Lyngbya sp., and Pseudoanabaena sp. in
Sweden at pH 4.3 to 4./ Uazarek 1980) and Phormidium sp.
in New York at pH 4.8 to 5.1 (Hendrey and Vertucci 1980).
These mats are dark blue-green with occasional flecks of
orange-colored carotene-rich material. They are thick,
5-18
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felt-like, and encrusting. 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 like Mougeotia
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 Mougeotla sp., Oedogonium, and
BulbochaeteT 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),
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
planktonic and benthic diatoms to sediments produces a record of the
past populations 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 stratigraphic record. This technique is subject to
variances caused by macroinvertebrate mixing, local changes in pH
sensitivity of species, and the numerous other factors besides pH that
determine the distribution of species (Norton et al. 1981).
Nonetheless, inferred pH generates a value that reflects the real water
5-19
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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
related 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. 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 which are currently acidic (pH < 5.0) have
decreased 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 and 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 alternatively, 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 indicate acidic (pH < 5.0) conditions are
primarily benthic, whereas those from circum-neutral (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
5-20
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an increase in dominance by acid-tolerant species in the most
acid-impacted lakes. A trend towards species tolerant of low nutrient
waters was also reported. label!aria fenestrata and Cyclotella
stelligera increased in numbers most directly with increasing acidity,
although some 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 eutrophication. 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 oligotrophication. The eutrophic ponds
remained well-buffered and unchanged. Thus, the effects on community
composition brought on by cultural eutrophication 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
acidification 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, 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 artificially acidified 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 dominance 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 (14C) showed no trend with pH. The dominance of the
periphyton by Chlorophyta in the acidified samples was due almost
entirely to the growth of Mougeotia sp., which by June represented 96
percent of the biomass and cell numbers at pH 4.0. This taxon was
responsible 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 superimpose the effects of acidification on
aquatic ecosystems. The only other report of seasonal changes in
periphyton (Lazarek 1982) dealt with algae living attached to Lobelia
dortmanna and verified the succession from diatoms to green algae
(Mougeotia spp.) during the growing season.
5-21
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Higher standing crops but lower rates of C-fixation per unit
chlorophyll occurred in periphyton growing in artificial stream channels
at reduced pH (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 grazing by the reduced
macroinvertebrate population, and 3) inhibition of microbial
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 £ 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 Mi'croinvertebrates--The responses of several minor groups of
invertebrates 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 (COa?-) 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
5-22
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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.
Aquatic mites (Acarina) are not generally collected in surveys of
benthic fauna, but Raddum (1976) noted that mites occurred in great
abundance in the shallow water of an acid-impacted lak-e. 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
like crayfish (Decapoda), sow bugs (Isopoda), and scuds (Amphipoda), but
also smaller forms such as benthic copepods, mysids, cladocerans, and
other branchiopods (e.g., Lepldurus). 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 amphi pod 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 stressedTby
either acidic inputs or organic enrichment. A. aquaticus was widely
distributed in acid-stressed lakes at pH 5.0 TK. 0kland 1980b) but £.
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. G_. lacustris achieved 96 hr Tl-50 at pH 7.26 in Montana, but
5-23
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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, G. fossarum, from Germany, showed
no mortality at pH 6.0, and had a 96 Tfr TL-5Q of ~ 4.7. At pH 5.0,
30 percent of the laboratory population survived for 10 days (Matthias
1982). K. 0kland (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 cannibalistically, 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-dwelling crustacean
characteristic 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 j^. 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 Daphnia middendorffiana,
Diaptomus arcticus, Lepidurus arcticus and Branchinecta paUidpsa have
provided additional evidence (Havas and Hutchinson 1982) of the
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 North West 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, Mi, 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~l 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.
5-24
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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 softwater of ~ 22 ymhos cm-1 conductivity
and Ca2+ of 2.8 mg JT*, 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 £-1, and conductivity of 96
ymhos cm-1. Cambarus bartoni was found in three acidic lakes (pH
4.6 to 4.9, - 3 mg £-1 Ca^"1") and Orconectes propinquis was
collected in one of three acidic lakes (Collins et ai. 1981). I have
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 viri1is has difficulty recalcifying its
exoskeleton after molting at pH < 5.5. 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 parasitic protozoan Thelohonia 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 predation by fish is eliminated in some
water 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
5-25
409-262 0-83-13
-------
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 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 ignita, and Heptagem'a
lateral is were absent only from the acidic 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 indicated 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
distributional pattern of the mayfly, Ephemerella funeral is, was studied
in the Hubbard Brook, NH, watershed during a 2-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 Libel!ula pulchella
tolerated pH 1.0 for several 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
5-26
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TABLE 5-2. RESULTS OF LABORATORY STUDIES ON pH TOLERANCE OF SELECTED
INSECT NYMPHS. TLso 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
96 hr
PH
Long-term 50% successful
TL50 emergence References3
Ephemeroptera
Baetls sp. 4.5
Clnygniula par 6.11
EphemereTIa doddsl 4.10
Ephemere]fa grandls 3.6 5.8(48)
EphemerelTa subvarfa 4.65 5.38(30)
Heptagenia~sp. 6.17
Hexagenla Umbata 5.66 5.5(33)b
LeptppMebia sp. 5.20
RhlthrogenaTrobusta 4.60
StenonemaTubrurn 3.32
Odonata
Boyerla vlnosa 3.25
Ophtogomphus rtplnsulensis 3.50
Plecoptera
Acroneun'a lycorias
Acroneuri'a paclfi'ca
Arcynopteryx parallel a
Isogenus aestivails
Isogenus frontalts
Isoperla fulva
Nemoura cinerea
PterorTarcella badla
Pteronarcys caHfornica
Pteronarcy? dorsata
Taem'opteryx maura
5.
3.
3.32
3.8
4.37
.08
.68
4.5
2.6
3.92
4.44
4.25
3.25
Trichoptera
Hydropsyche betteni 3.15
Hydropsycffe sp. 3.28
Arctopsycfie" grandls 3.4
Ltmnephtlus ornatu? 2.82
BrachycenTrus americanus 1.50
Brachycentrus occldentails
Cheumatopsyche sp.
4.42(30)
4.30(30)
3.85(30)
5.8(90)
4.50(30)
4.52(90)
4.95(90)
5.00(30)
3.71(30)
3.38(30)
2.45(30)
4.3(90)
4.52(90)
5.9
5.2
5.2
5.0
6.6
5.8
4.0
4.7
4.0
M
G
G
G
B, BN
G
G
G
G
B
B, BN
B, BN
B,BN
G
G
G
B, BN
G
M
G
G
B, BN
B, BN
B, BN
G
G
G
B, BN
G
G
5-27
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TABLE 5-2. CONTINUED
Organl sms
Diptera
Atherix variegata
Holorusia sp.
Simulium vlttatum
96 hr
TL50
2.8
3.63
pH
Long-term 50% successful
TLso emergence
4.2(68)
References3
G
G
G
aReferences: B = Bell 1971, BN = Bell and Nebeker 1969, G = Gaufin
1973, M = Matthias 1982.
^Seventy of 90 survived.
5-28
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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 Taeniopteryx, 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 near pH 4.5 in field
surveys (Sutcliffe and Carrick 1973, Leivestad et al. 1976, Raddum 1976)
but not at pH 4.0 (Raddum 1979; Hall and Likens 1980a,b). Raddum (1979)
observed that the running water caddisflies Rhyacophila nubila,
Hydropsyche sp., Polycentropus flavomaculus, and Plecforenemia conspersa
a 11 survived pH 4.0 in the laboratory, but only P. conspersa did well in
situ at pH 4.8. Raddum explained the loss of Rhyacophila and
Hydropsyche in the field by alterations in their food supply. P_.
navomacuTatus became cannibalistic 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,
Limnephilus pal lens, 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, Sialis (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.. LlmnochTronomous sp., Sergentia coracina,
5-29
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Stlchtochironomus sp. 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 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
(Henri ptera) like 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 metabolically very active and receive
most of their 03 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 microhabitats.
Important generalizations are better made by analyzing the data after
grouping the taxa by functional guilds and microhabitats rather than by
phylogenetic associations (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
decreases in populations of animals living in sediments (infauna).
Observations of epifauna by scuba divers concurred with the general
observation that acidic lakes have depauperitic mollusc and insect
populations.
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, Hexagenia. Metabolic and physical actions of
Hexagenia nymphs increased the Eh, NHa, inorganic S, SOa, and
decreased the pH as compared to control microcosms lacking nymphs or
5-30
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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
invertebrate 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 corroborated
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 other factors besides H"1" concentration stress organisms in
acidified waters. (Overrein et al. 1980). Malley's (1980) work (see
Section 5.4.2.4) 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, Glaenpcorisa propinqua propinqua a
predator on zooplankton and other small invertebrates, 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 yg &-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 modifications 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
5-31
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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 Mollusca--Molluscs provide food for vertebrates (fish, ducks,
muskrats, etc.). Clams are filter feeders and are important bio-
indicators 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 molTuscan 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. flkland 1969a,b, 1976, 1979a,b, 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. Sphaeriidae live in
sediments (infaunal) and no surveys of the more epifauna! 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 CaCOa 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, indicating that pH may stress snails by reducing
the CaCOs availability (J. 0kland 1979b). These 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).
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).
5-32
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These later authors emphasized 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 ir 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-acidifica-
tion water quality was suggested.
5.3.2.7 Anne!ida--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 fy concentrations, the benthic community is
typically dominated by Tubifex spp. and Limnodrilus hoffmeisteri
(Brinkhurst 1965, Howmiller 1977). Considering their tolerance of other
stresses 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
oligotrophic 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 (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 concentrations of pollutants (Hart and Fuller
1974, Chapman et al. 1980). An alternative 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,
Glossiphpnia heteroclita) were absent from even mildly acidic lakes.
Raddum (1980) 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
5-33
-------
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 extensively noted in field
surveys and have developed in artificially acidified chambers and
stream channels below pH 5.0.
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 observations and laboratory bioassays.
Insects are often limited by mechanisms not related to direct
toxicity. Some dragonflies and many predacious 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 predaceous insects receive their air supply
from the surface.
5-34
-------
TABLE 5-3. SUMMARY OF EFFECTS OF pH 4.5 WATER ON BENTHOS
Taxon
Common name
Microhabitats
Sensitivity to acid
(pH 4.5) stress
References
Bacteria
Periphyton
Algae
Blue-greens (Cyanophyta)
Greens (Chlorophyta)
Diatoms (Bacillariophyceae)
Dinoflagellates (Pyrrophyta)
oo
en
Crustacea
Decapoda
Crayf 1 sh
Isopoda
Amphipoda
Aquatic sowbug
Scud
Al 1 substrates.
On plants (epiphytic)
On rocks (epipl ithic)
On mud (epipelic)
On "encrusting mat".
Burrowers deposit
feeders and grazers
lakes and streams.
Deposit feeder in
lakes and streams,
under rocks and in
littoral vegetation.
Detritivore in lakes
and slow-flowinq
areas of streams.
Found among pi ant
stems.
Growth rate or Og
uptake inhibited.
Increasing standing
crop in lakes and
streams.
Development of dis-
tinct types of peri-
phyton communities.
Sensitivity variable
and highly species
specific. Effects
vary depending on
other cation concen-
trations.
Asellus aquaticus
tolerant to - pH
5.0.
Not generally found
below pH 6.0. Sen-
sitive differences
between species have
been described.
Bick and Drews
1973, Baath et al.
1979, Gahnstrom et
al. 1980
Hendrey 1976, Hall
et al. 1980, Van
and Stokes 1978,
Hendrey and
Vertucci 1980,
Muller 1980
Singer et al. In
press, Stokes 1981
Mai ley 1980,
Collins et al.
1981, Shaw 1960
K. Okland 1978a,
1980b.
K. 0kland
1980a,b,c;
Sutcliffe and
Car rick 1973;
-------
TABLE 5-3. CONTINUED
Taxon
Common name
Microhabitats
Sensitivity to acid
(pH 4.5) stress
References
Eubranchiopoda Tadpole shrimp
Small, often tenpo-
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
Include burrowing and Sensitivity varies
en
i
co
CD
Odonata Dragonflies (Anisoptera)
Damselflies (Zygoptera)
Plecoptera StonefHes
Trichoptera Caddlsflies
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.
between groups but
generally not
tolerant
Tolerant to pro-
longed severe acid
stress.
Most genera are
sensitive but some
are tolerant (see
text).
Some genera are very
tolerant, hut others
are sensitive (see
text)
Sutcliffe and
Carrick 1973,
Nilssen 1980,
Pratt and Hall
1981, Leivestad
et al. 1976,
Borgstrom et al.
1976, Raddum 1976
Stickney 1922,
Borgstrom et al.
1976, Bell and
Nebeker 1969, Bell
1971
Sutcliffe and
Carrick 1973;
Leivestad et al.
1976; Hall and
Likens 1980a,b;
Raddum 1979
Sutcliffe and
Carrick 1973;
Leivestad et al.
1976; Hall and
Likens 1980a,b;
Raddum 1976, 1979
-------
TABLE 5-3. CONTINUED
Taxon
Common name
Microhabitats
Sensitivity to acid
(pH 4.5) stress
References
Diptera
01
i
GO
Hemiptera
Coleoptera
Mollusca
Pel ecypoda
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 (Gyrinidae)
Clams
Major detritivores in
lakes and enriched
streams, living in
mud or on substrates
in tubes.
Predator on sub-
strates and in 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 clans which are
rarely found as low
as pH 5.0.
Raddum and Saether
1981, Uutala 1981
Wi ederholm and
Eriksson 1977,
Leivestad et al.
1976
Leivestad et al.
1976
Raddum 1976,
Raddum et al.
1979, Nilssen 1980
Raddum 1976,
Raddum et al.
1979, Nilssen
1980
J. (Bkland 1976,
1979b, 1980; K.
0kland 1971,
1979b,c, 1980b;
(Jkland and
flkland 1978,
1980; 0kland
and Kuiper 1980;
Singer 1981b
-------
en
i
to
CO
TABLE 5-3. CONTINUED
Taxon
Common name
Mlcrohabitats
Sensitivity to acid
(pH 4.5) stress
References
Annelida
01igochaeta
Hirudinea
Aquatic earthworms
Leeches
Detritivores in lakes Standing crops low
and streams with soft in acidic waters
substrates.
Predators, detriti-
vores.
Anecdotal observa-
tions report no
leeches below pH
5.5.
Raddum 1976, 1980
Nilssen 1980,
Raddum 1980
-------
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 may-
flies). In those groups that have been studied in the laboratory
(crayfish, backswimmers, molluscs), high calcium concentrations (>
2 mg £-1 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 hard waters 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, Mi tell a 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 in turn are a major food source for the
larger aquatic animals, such as 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 (.02 mM C02 at pH
5.0) and major nutrient minerals such as P, K, N, and Ca are in limited
supply (Hutchinson 1975). However, these aquatic macrophytes have
5-39
-------
TABLE 5-4. PLANTS COMMONLY OBSERVED IN SOFTWATER (LOW ALKALINITY)
OLIGOTROPHIC LAKES
Species
Common Name Type
Response to
Acidification
Pontederia
cordata L.
Pickerel
weed
Emergent
Unknown
Juncus sp.
Spargam'um spp.
Brasenia
schreberi Gmel
Nuphar
advena Ait.
Nymphaea
odorata Ait.
Isoetes spp.
Lobelia
dortmanna L.
Eriocaulon
septangulare With
Myriophylliim
Tk
tenel1umBigel
Potamogeton spp.
Rush
Burreed
Water
shield
Yellow
lily
White
lily
Quillwort
Pipewort
Pond weeds
Emergent
Emergent
Floating leaves,
rooted
Floating leaves,
rooted
Floating leaves,
rooted
Submerged,
rooted
(iosetids)
Submerged,
rooted
(iosetids)
Submerged,
rooted
(iosetids)
Submerged,
rooted
(iosetids)
Submerged,
rooted
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)
5-40
-------
TABLE 5-4. (CONTINUED)
Species
Common Name Type
Response to
Acidification
Eleocharis spp.
Utricu'iaria spp.
Sphagnum sp.
Nitella spp.
Spike rush Submerged,
rooted
Bladderwort Submerged,
unrooted
Moss
Drepanocladus spp. Moss
Fontinalis spp. Moss
Submerged,
attached
Submerged,
attached
Submerged,
attached
Stonewort Submerged,
attached
Unknown
Unknown
Stimulated
growth (Grahn
1977)
Unknown
Unknown
Unknown
5-41
-------
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 C02
and store it in an internal gas chamber systan 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
accomplished 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 (Wiurn-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 nutients 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
CO^ 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, concentration differences between sediment
and water column levels of metals 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 £-M in the
Adirondacks and Ontario were elevated above Hutchinson's (1975) mean
value of 0.36 g Kg-1 (Best and Peverly 1981, Miller et al. 1982).
5-42
-------
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 metabolically, but also physically adsorb then 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
snow melt. 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 photosynthesis, nutrient depletion, increased metal
concentrations, and decreased rates of nutrient recycling (Grahn et al.
1974, Andersson et al. 1978b, 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, Lobelia 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
5-43
-------
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 species by exchanging 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 5.8.2.1) 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 nonacidic 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 nonacidified 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. 1982). 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. 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).
5-44
-------
If metal concentrations Increase in tissues, but do not inhibit
growth, there is a potential for increased cycling of metals. However,
Sphagnum spp. growth may be a positive factor, removing metals from the
water by adsorption (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). 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
communities by Sphagnum sp. in 50 oligotrophic, softwater lakes
surveyed in North America. In fact, dominant species are the
same in both acidified (pH less than 5.6) and non-acidified (pH
5.6 to 7.5) lakes.
0 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.
o 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 nonacidified 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 ofincreased 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
cadimium and zinc are lower in the tissues of plants from
acidified lakes.
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
5-45
-------
distribution by water movements (Wetzel 1975). The plankton community
consists of animals (zoo plank ton), plants (phy to plank ton), and microbes.
Effects of acidification on zooplankton and phy to plank ton 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 communites 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 taxonomic 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 interactions 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 difficult. In some cases,
results appear contradictory. With an increased understanding of causal
mechanisms, many of these apparent contradictions should be resolved.
5-46
-------
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.
uberrimum (both dinoflagellatesT^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 Peridinium limbatum and Peridinium inconspicuum. In Carlyle
Lake (pH 4.8 to 5.1) near Sudbury, acidification experiments within
limnocorrals resulted in the proliferation of Peridinium limbatum (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 (National Research Council Canada 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 lakes sampled, these
algae accounted for only 21 and 23 percent of the standing crop. On the
5-47
-------
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 mi sslng
1n acid water
General
comments
Swedish West
Coast (Aimer et
al. 1974 and
1978)
Numbers of species per 100 ml
sample:
pH 6-8: 30 to 80 species
pH < 5: about 12
pH < 4: 3
en
CO
In most acid waters:
dlnoflaqellates (Pyrrophyta)
Peridinium inconspicuum
Gymnodinium cf. uberrimum
In a few lakes with pH about 4:
green algae (Chlorophyta) -
Ankistrodesmus convolutus
pocystis submarina
Oocystf? lacustrTs
Other common species:
chrysophyceans (Chrysophyta)
Dinohryon crenulatum
Dinohryon sertularia
qreen algae (Chlorophyta)
Chlamydomonas sp.
The classes Chlorophyceae
(Chlorophyta) and
Chrysophyceae (Chrysophyta)
had greatly reduced numbers
of species
Absence of diatoms (class
Bacillarlophyceae, Phylum
Chrysophyta) and bluegreen
alqae (Cyanophyta) at pH <5:
Chroococcus limneticus
Heri snipped ia~tenu1ssima
Species common 1n oligotrophlc
lakes, but absent at pH <6:
bluegreen algae (Cyanophyta) -
(lOmphpsphaeria lacustrls
green algae (Chlorophyta) -
Scenedesmus serratus
chrysophyceans (Chrysophyta) -
Dinobryon divergens
Dinobryon bavarlcum
DinobryoF borgel
Dinobryori sucecicum
Kephyripn spirale
Stichogloea doederlelnll
diatoms (Chrysophyta) -
Rhizosolenia lonqiseta
CyclotellaTodanica
cryptophytes (Cryptophyta) -
Rhodomonas mi nuta
dinoflagellates (Pyrrophyta) -
Ceratium hirundinella
One stop survey of
115 lakes in August
1972 and 60 lakes 1n
Greatest change in
species composition
occurred in the pH
Interval 5 to 6
2.
Swedish West
Coast (Hultberg
and Andersson
1982)
Following Hming, number of
species generally increased
Dominant species in acid,
ollqotrophic lakes:
dinoflagellates {Pyrrophyta) -
Perldinlum Inconspicuum
Gymnodinium sp.
Following Umlnq, the Importance
of genus Perldinlum declined
Prevalent (30 to 40% of the
blomass) 1n humlc lake:
bluegreen alqae (Cyanophyta) -
Merismopedia sp.
green alqae (Chlorophyta) -
Oocystls sp.
Diatoms Insignificant In all
acid lakes
Following liming, the Importance
of species of green algae
(Chlorophyta) and
chrysophyceans (Chrysophyta),
and, 1n some cases, diatoms
(Chrysophyta) Increased
pre- and post-1imlnq
studies; long-term
monitoring of four
lakes
-------
TABLE 5-5. CONTINUED
Location
(reference)
Reductions 1n
species diversity
dominant species
1n add water
Species missing
1n acid water
General
comments
Southern Norway Number of species Identified
(Hendrey and per lake:
Wriqht 1976, pH > 4.5: 10 to 25 species
Leivestad et pH 4 to 4.5: < in
al. 1976)
Decrease 1n importance of
species of green algae (Class
Chlorophyceae, Phylum
Chlorophyta)
No consistent trend relating pH
to numbers of species of
diatoms (Chrysophyta) or
bluegreen algae (Cyanophyta)
One-stop survey of 55
lakes in October 1974
4.
Southern Norway
(Raddum et al.
1980)
The number of algal species
collected at any one time
was generally lower 1n 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
blomass consisted of
dlnoflagellates (Pyrrophyta)
especially-
Peridinium limbatum
Peridinium InconspTcuum
However, this was not the case
1n a naturally acidic
dystrophlc 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 phytoplankton
6.
Sudbury Region
of Ontario
(Van 1979)
Number of taxa observed in
acidic lakes was less than
in non-acidic lakes
Biomass 1n acid lakes dominated
by a dinofl agellate
(Pyrrophyta) Peridinium
Inconspicuum
Proportion of the biomass
contributed by dlnoflagellates
was correlated with hydrogen
1on activity, hut not with
phosphorus concentration
Most common genera in acid
lakes:
dinoflaqcllates (Pyrrophyta) -
Peridinium
cryptophytes (Cryptophyta) -
Cryptomonas
chrysophyceans (Chrysophyta) -
Ilinohryon
green algae (Chlorophyta) -
Chlamydomonas Oocystis
Non-acidic oligotrophic lakes
typically dominated by
chrysophyceans (Chrysophyta)
and diatoms (Chrysophyta),
but In acidic lakes sampled
a dinoqlagellate (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 1n community
structure apparently
occurs over a pH range
of 4.7 to 5.6
-------
TABLE 5-5. CONTINUED
Location Reductions in
(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 dinofl agel lates
(Pyrrophyta) and cryptophytes
(Cryptophyta) to the
chrysophyceans (Chrysophyta)
more typically observed in
circumneutral waters
neneral
comments
Three of the acidic lakes
sampled by Yan (1979)
were limed 1973-1975;
pH levels were raised
from <4.7 to above 6
Sudbury Region
of Ontario
(Conroy et al.
1976)
(Jl
CD
In the two acidic lakes, a few
qenera usually dominated the
bicmass, resulting \n a low
diversity index. In the
non-acidic lakes, the
blomass was more evenly
distributed throughout a
large number of qenera
In acidic Wavy Lake, the
dominant genus was a
chrysophycean (Chrysophyta)
Dinohryon. 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
90% of the standing crop.
In the 2 non-acidic lakes,
these 2 groups accounted for
only 217. 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
Millerd 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
LaCloche
Mountain Region
of Ontario
(Kwlatkowski
and Roff 1976)
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 1n their numbers of
species with Increasing
add conditions. Comparing
the highest pH lake sampled
(pH about 6.7) with the
most add 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 in. Numbers of
species of diatoms
The important species 1n each lake shifted according to pH level.
In the more neutral lakes, the green algae (Chlorophyta)
comprised between 40 and 50% of the total algal flora, with
bluegreen algae (Cyanophyta) accounting for only 30%. In acidic
lakes, however, bluegreen algae constituted about 60%, and green
algae only about 25% of the algal flora.
Species common 1n acidic waters:
Bluegreen algae (Cyanophyta) -
Aphanocapsa sp.
Chroococcus Prescottii
fiscillatorla sp.
Rhabdodgrma 1Ineare
Green algae (Chlorophyta) -
Carteria sp.
Chiamydomonas sp.
Chlorella eTTlpsoidea
Closterium sp.
Aphanocapsa sp.
Chroococcus dispersus
Chroococcus 1imneticus
Osci1latoria sp.
Ankistrodesmus sp.
Carteria sp.
Chiamydomonas sp.
Oocystis sp.
Scenedesmus sp.
6 la,kes 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 similarlHy
of community. Thus,
community structure
in these lakes
reflected the pH
gradient
-------
TABLE 5-5. CONTINUED
Location
(reference)
Reductions 1n
species diversity
Dominant species
1n acid water
Species missing
1n acid water
General
comments
9. Cont.
(Chrysophyta) collected 1n
samples were also greatly
reduced 1n the two most
acidic lakes relative to the
other lakes sampled
Cryptomonads (Cryptophyta) -
(considered by the authors as 1n the Phylum Pyrrophyta)
Cryptomonas erosa
Cryptomonas ovata
Dinoflaqellates (Pyrrophyta)
species of the genera Peridinium
and Glenodinlum, although present
1n 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.
Ul
LdCloche
Mountain Region
of Ontario
(Van and Stokes
1978)
Phytoplankton community
dominated by Peridinium
limbatum (a dinoglagellate,
Phylum Pyrrophyta), and
Cryptomonas ovata
(a cryptomonad, Phylurn
Cryptophyta, but considered by
the authors in the Phylum
Pyrrophyta)
These 2 groups formed between
50-90% of the biomass in all
collections
Intensively sampled one
acid lake, Carlyle
Lake (pH about 5.0),
also studied by
Kwiatkowksi 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 circumneutral lake
Many species of the Class
Chrysophyceae (Chrysophyta),
the class Myzophyceae
(Cyanophyta; bluegreen algae),
and diatoms (Class Bacillario-
phyceae, Phylum Chrysophyta)
developed in the circumneutral
lake that were absent or
occurred in only small numbers
in the 2 acidic lakes
Three lakes - one
circumneutral and two
acidic lakes,
acidified as a result
of contamination by
acid leachate from
processing of local
urani um 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
Homlnant species
1n acid water
Species missing
In acid water
General
comments
12. Adirondack
Region of New
York State
(Hendrey 1980,
Hendrey et al.
1980b)
13. Adirondack
Region of Hew
York State
(Charles 1982)
Total number of species
identified 1n each lake
decreased with increasing
acidity:
clrcumneutral lake - 64
intermediate - 38
acidic - 27
Species of the Class
Chrysophyceae (Chrysophyta)
dominated the bioraass of the
most acidic lake, although
dinoflaqellates (Pyrrophyta),
especially Peridinium
inconspicuum. comprised a
significant fraction of the
biomass In the ice-free
season
cn
en
ro
Numbers of species of green
algae (Chlorophyta) and blue-
green alqae (Cyanophyta)
decreased most markedly
Dinobryon spp. (Chrysophyta) are
the typical dominant phyto-
plankters during the summer 1n
Adirondack lakes
All lakes with pH > about 5.8
had euplanktonic diatoms
(class Badllarlophyceae,
Phylum Chrysophyta) present in
their surface sediments.
Lakes with a lower pH had
none.
Intensive sampling of
three lakes - one
acidic (pH about 4.9),
one intermediate (pH
about 5.5), and one
circumneutral (pH
about 7.0)
Survey of sediment
diatom assemblages and
lake water
characteristics for
39 lakes
14. Florida
(Crisnan et al.
1980)
Mean number of taxa in acidic
lakes was 10.8 vs. 16.5 for
non-acidic lakes
In most acidic lakes (pH 4.5 to
5.0), green algae (Chloro-
phyta) accounted for 60% 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
Ankistrodesmus
Staurastrum
and several species of small
coccoid green algae
Rlnoflagellates (Pyrrophyta) -
Peridinium
In lakes of pH 6.5 to 7.0,
bluegreen algae (Cyanophyta)
made up 63% 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 1n
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
add lake
Study of a very acid
lake (pH 3.2 to 4.1)
affected by strip
mining)
-------
TABLE 5-5. CONTINUED
Location
(reference)
Reductions 1n
species diversity
Dominant species
In acid water
Species ml sslng
In acid water
General
comments
16. England
(Hargreaves et
al. 1975)
Number of alga) specfes
present per water was
negatively correlated with
total acidity
Study of 15 waters with
pH levels of 3.0 or
less; most affected by
strip mining
activities
cn
i
en
CO
17. Smoking Hills
Region, North-
west Terr.,
Canada
(Hutchinson et
al. 1978)
In these very acidic ponds,
phytoplankton populations
were depleted
Even at these extremely low pH
levels, some species of algae
still commonly occurred:
Euglenoids (Euglerrophyta) -
Euglena mutablis
Diatom (Chrysophyta) -
Nitzshia sp.
Rinoflagellate (Pyrrophyta) -
Gymnodinium ordinatura
Ponds affected by
spontaneous burning
of bituminous shale
deposits. pH values
as low as 1.8
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, Waimangu
Cauldron
-------
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.
Hultherg and Andersson (1982) noted that blue-greens (again Merismopedia
sp.) were prevalent in humic acid lakes in Sweden. Stokes (1980) noted
that 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 conditions.
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.
5.5.2.2 Changes in Phytoplankton Biomass and Productivity--Available
data on acidification and primary productivity in acidic lakes yield no
clear correlation 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
5-54
-------
TABLE 5-6. THE RELATIONSHIP BETWEEN LAKE ACIDITY AND PHYTOPLANKTON BIOMASS AND/OR
PRODUCTIVITY—OBSERVED RESPONSE TO LOW pH
CJl
CJ1
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
(KwiaUowski 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 £ concentrations were
siqnficantly lower in acidic lakes (1.88
nig m"3) than in non-acidic lakes (7.53 mg
ra-3) (Crlsnan et al. 1980).
Significant Increase
In 58 lakes along the west coast of Sweden,
the larqest biomass of phytoplankton occurred
in the most acidic lakes (pH 4.5), and the
lowest biomass at Intermediate pH levels
(pH 5.1 to 5.6) (Aimer et al. 1978).
In acidification experiments within Hmno-
corrals 1n 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
(Van 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
1n 1976 to pH 5.4 1n 1980. Over that time
period, chlorophyll and algal bionass Increased
significantly, associated with hypoHmnetlc
algal blooms of Chlorella, and apparently in
response to increased water clarity (Schlndler
and Turner 1982).
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, no,r did
phytoplankton biomass (Van 1979). Experimental increases
in phosphorus levels 1n acidic lakes (with or without
neutralization) have, however, induced significant
increases in phytoplankton biomass (Dillon et al. 197R,
Hendrey et al. 1980b).
Within eight plastic enclosures in Eunice lake, an
oligotrophic lake with pH 6.5-in British Columbia, add
addition (minimum pH 5.5) resulted 1n no significant change
in chlorophyll content. Additions of acid plus nutrients
(minimum pH 5.0) increased algal biomass (Marmorek 1983).
In three Swedish lakes sampled from 13 to 15 May 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 1n all three
lakes (Aimer et al. 1978).
In 13 lakes in southern Norway, chlorophyll ^ content was
not significantly correlated with lake pH (Raddum et al.
1980).
-------
status and not a cause-and-effect relationship between pH and
phytoplankton response.
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 examination 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 m~3 hr"l) is well within the range normally observed in
non-acidic lakes in the region (0.3 to 6.9 mg m-3 hr"1) (National
Research Council Canada 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
National Research Council Canada 1981) and Norway (Raddum et al. 1980)
found no correlation between lake pH and algal biomass; in Sweden (Aimer
et alI. 1978), the largest biomass occurred in the most acidic lakes.
Acidification experiments within limnocorrols yielded no change
(Marmorek 1983) or an increase (Van 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 phytoplankton 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, Loefer 1973, Moss 1973,
Cassin 1974). For each species, growth responses to pH form an inverted
U-curve, with an optimum 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.
5-56
-------
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).
The presence of an alga at a low pH level does not necessarily
imply a preference for acidic conditions or that photosynthesis and
growth are optimal (Hendrey et al. 1980b). The proliferation of
Perl 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).
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 (Mailey 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 of 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
5-57
409-262 0-83-14
-------
TABLE 5-7. SUMMARY OF OBSERVATIONS RELATING SPECIES COMPOSITION, SPECIES DIVERSITY, AND
BIOMASS OF THE ZOOPLANKTON COMMUNITY TO ACIDITY
en
en
00
Changes in species composition and abundance of:
Location
(Reference)
1. Southern
Sweden
(Aimer et
al. 1974
and 1978)
General
Observations
Number of species
lower add 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
Bosmina coreqoni , and
niaptomus qracilis
often dominate
Rotifers
Polyarthra remata,
Polyarthra vulqarls,
Keratella cochlearis, and
Kellicottia lonqispina
common at most pH levels,
4.4 to 7.9
Polyarthra remata dominant
in several lakes with pH
< 5.5
Conochllus unlcornls
present in many lakes but
less prevalent in acid
waters
Many of the other rotifers
Cladocerans
Bosmina coregonl common
and occurred at all pH
levels
All Daphnia species
were sensitive to low
pH levels. Only a few
individuals found at
pH < 6
niaphanosoma
brachyurum, Holopedium
qlbberum, and Leptodora
kindti common but
mainly at pH > 4.9
Bythotrephes longimanus
Copepods
Dlaptomus gracllis and
Cyclops spp. common at all
pH levels
Heteroeope appendlculata
occurred mostly at pH>5.5
Others Comments
One-stop survey
of 84 lakes In
August 1971
Samples collected
with 75 11 mesh
net
appear to have preferences
above 5.5
found more frequently
in lakes with pH < 5.4.
At higher pH levels,
fish predation may keep
the population at low
levels
Common 1n non-act die
lakes:
Dlaphanosoma
2. Southern
Sweden
(Hultberg
and
Anderson
1982)
In acidic lakes,
zooplankton community
dominated by a few
species
Dominants 1n acid lakes:
Polyarthra spp.
Keratella cochlearis
Kelllcottia longisplna
Common after liming:
Polyarthra spp.
Keratella cochlearis
Asplanchna prlodonta
Conochilus mincornis
Holopedium
Daphnia cri stata
Bosmina
Dominant 1n acid lakes:
Bosmina coreqoni
Common after liming:
Bosmina coreqoni
niaphanosoma sp.
Daphnia cristata
Limnosida froutosa
Holopedium qibberum
Ceriodaphnia
quaaranqula
Dominants in add lakes:
Eudlaptomus gracllis
Cyclops spp.
Common after liming:
Eudiaptomus qracilis
Pre- and post-
liming studies.
Effects of lining
on zooplankton
are difficult to
evaluate due to
simultaneous
rotenone treat-
ments
-------
TABLE 5-7. CONTINUED
Changes 1n species composition and abundance of:
Location
(reference)
3. Southern
Norway
(Hendrey
and Wright
1976)
General
observations
Total number of
species collected
decreased with
decreasinq pH
Rotifers Cladocerans
Daphnia galeata absent
at pH < 6.9
Eubosmina lonqispina
common at al 1 pH
levels, 4.1 to 7.7
Hoi oped 1 urn gibberum
occurred frequently at
Copepods Others
Eudiaptomus qracilis
common over wide ranqe of
pH, 4.1 to 6.6 most
frequently dominant at low
pH levels; rarely dominant
at pH>5.5
Heterocope sal lens
occurred pH 4.1 to 6.6
Comments
One-stop survey
of 57 lakes
durlno fall 1974
Samples collected
with single
vertlcle haul of
a 75 v mesh
net
01
01
UD
pH levels 4.2 to 7.2
Daphnia longispina
appeared in samples pH
4.6 to 6.8
Ancanthodiaptomus
denticornis and
Mixodiaptomus lacinlatus
did not occur at pH < 5
Cyclops scutifer appeared
at pH 4.6 to 7.7
4. Norway
(Raddum et
al. 1980;
Raddum,
1978;
Hobaek and
Raddum
1980)
Number of species
lower in add lakes.
In southern Norway,
clear-water lakes with
pH < 5 held on the
average 7.1 species;
equally acid humic
lakes, 11.7 species;
less acid (pH > 5.5)
clearwater lakes,
16.1 species on
average
All major groups
contributed to the
lowered number of
species, but
cladocerans
apparently most
affected
Species occurring with
equal frequency in acid
and non-acid clearwater
lakes:
Kellicottia lonqispina
Keratella serrulata
Species more frequent in
non-acidic lakes:
Conochllus spp.
Asplanchna sp.
Keratel 1 a
cochlearis
Keratella
hlemalis
Species more frequent in
acidic lakes:
Polyarthra spp.
Species occurinq with
equal frequency in acid
(pH < 5) and non-acid
(pH > 5.5) clearwater
lakes:
Bosmina (Eubosmina)
lonqispina
Species more frequent
1n non-acid lakes:
Holopedium gibberum
Diaphanosoma
TTrachyurum
Ceriodaphnia
nuadranqula
Daphnia lonqispina
Daphnia galeata
Bythotrephes
lonqimanus
Polyphemus pediculus
Leptodora kindti
Species occurrlnq with Chaeborus
equal frequency in add flavicans
and non-acid lakes: absent 1n
Eudiaptomus qracilis clearwater
Heterocope saliens acid lakes
Species more frequent in
non-acid lakes:
Cyclops scutifer
Cyclops abyssorum
Mesocyclops leuckartl
Survey of 27
lakes; sampled
(3 vertical net
hauls, with 90
u mesh net , per
visit) from June
to September 1977
- 1979
-------
TABLE 5-7. CONTINUED
Location
(reference)
General
observations
Changes
Rotifers
in species composition and abundance of:
Cladocerans Copepods Others Comments
en
i
4. cont. Some species tolerate
acid conditions 1n
the presence of
humus, but are ahsent
from add clearwater
lakes
The species number of
filter-feeders
reduced in clear-
water acid lakes.
Changes for
rapturial species
not as obvious
5. Southern Lower abundance of
Norway Daphnia longlsplna and
(Nilssen Daphnia lonqlremus at
1980) pH<5.5
Bosmina longlsplna a
dominant species at all
pH levels, 4.5 to 7.0
Leptodora klndtl absent
at pH levels 5.0 to 5.5
In sediment cores from
lake with pH 4.2 to
5.0, Daphnia Ignqispina
present below 3 to 4
cm, hut absent in
surface sediments and
In the lake
Eudiaptomus qracilis
a dominant species at
all pH levels, 4.5 to 7.0
Cyclops scutifer
common at all pH levels,
4.5 to 7.0
Heterocope sal lens
increased 1n abundance
1n more acid lakes
(pH < 5.0)
Chaoborus
flavicans
absent at pH
levels <
4.5 to 5.0
In sediment
cores from
lake with pH
4.2 to 5.0,
Chaoborus
flavicans
remnants
common below
3 to 4 cm,
but absent 1n
surface
sediments and
1n lakes
Samples collected
with 90 v mesh
net June -
October 1973 and
1974 at fourth
nlqht intervals
-------
TABLE 5-7. CONTINUED
en
en
Changes In species composition and abundance of:
Location General
(reference) ohservations
6. LaCloche Signifacant reduction
Mountain in number of species
Region of and numbers of
Ontario individuals at lower
(Roff and pH levels (pH 4.4 to
Kwiatkow- 4.8)
ski, 1977)
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
turocephala dominated. As
the pH increased,
Keratella cochlearis.
Kellicottia bostoniem's,
and Kellicottia longispina
increased in occurrence
Polyarthra euryptera and
Polyarthra dolichoptera
rare at pH 4.7 to 5.0;
absent pH<4.4
In highly acidic waters
(pH about 4.4), Keratella
taurocephala dominated.
As the pH increased,
Keratella cochlearis,
Kellicottia bostoniensis,
and Kellicottia longispina
Increased in occurrence.
Polyarthra euryptera and
Polyarthra dolichoptera
rare at pH 4.7 to 5.0;
absent pH < 4.4
Cladocerans
Standing crop of
Cladocerans reduced at
pH levels below 5;
maximum at pH 5 to 6
Leptodora klndti found
only at pH>5.0
Daphnia galeata
mendotae'Daphnia
retrocurva, and
Diaphanosoma
leuchtenbergianum found
in all lakes but rare
at pH 4.4 to 4.8
Bosmina longirostris,
Eubosmia tubicen, and
Holopedium gibberum
common in all lakes
Copepods Others
Standing crops of
cyclopoid copepods but not
calanoid copepods reduced
at pH levels 4.4 to 4.8
Diaptomus minutus 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
thomasi and Hesocyclops
edax found in all lakes,
pRT.4 to 6.8
Comments
Six lakes with
pll levels 4.0 to
7.1 samples at
weekly intervals
June and August
197? and Hay and
July 1973
Vertical haul
with 60 |i mesh
net; and
Schindler-Patalas
trap at various
depths.
-------
TABLE 5-7. CONTINUED
en
i
01
Changes in species composition and abundance of:
Location
(reference)
7. LaCloche
Mountain
Region of
Ontario
(Sprules
1975a,
1975t>)
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.
64% of all species
identified occurred
never or rarely at
pH < 5.0.
Rotifers Cladocerans
Tolerant species
distributed Independent
of pH:
Bosmlna
Diaphanosoma
leuchtenhergianum
Holopedium gibberum
Never occur pH < 5.0:
leptodora klndtl
Daphnia galcata
mendotae
Daphnia retrocurva
Daphnia ambiqua
Daphnia lonqiremis
Copepods Others
Tolerant species
distributed independent of
pH:
Mesocyclops edax
Cyclops bicuspidatus
thomasi
Diaptomus mlnutus
Never occur pH < 5.0:
Tropocyclops prasinus
mexicanus
Epischura lacustris
Diaptomus oregonesis
Comments
One-time sampling
of 47 lakes f'-om
July to early
September 1972 -
1973 .
Vertical hauls
with either
75 v or 110 u
mesh net
pH ranged from
3.8 to 7.0
In some lakes, only
Diaptomus minutus
remains. Above pH
5.0, pH had little
effect on tolerant
species and 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
Occur primarily in
lakes with pH < 6.0:
Polypemus pediculus
Daphnia~catawba
Daphnia pulicaria
Oiaptomus rcinutus dominant
in most lakes pH < 5.0; In
some cases the only
species present
-------
TABLE 5-7. CONTINUED
GO
Changes
Location General Rotifers
.(reference) observations
8. Sudbury Numbers of species
Region of reduced 1n 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 biomass
resul ted from both a
decrease In 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
In species composition and abundance of:
Cl adocerans
Major species 1n non-
acid lakes:
Bosmina longirostris
Holopedium qibberum
Diaphanosoma
leuchtenbergi anum
Daphnia galeata
mendotae
In acid waters, Bosmina
longirostris accounted
for an average of 79%
of the total crustacean
biomass vs 3% in
non-acid lakes
In acidic Clearwater
Lake zooplankton
community characterized
by the importance of
Bosmina longirostris,
and the absence of
Daphnia sp. and the
other common
cl adocerans Holopedium
gibberum and
Diaphanosoma
leuchtenbergianum
Copepods Others
Major species 1n non-acid
lakes:
Cyclops bicuspidatus
thomasi
Tropocyclops praslnus
mexicanus
Diaptomus minutus
Copepods contributed an
average of 65% of the
total biomass and 85% of
the total individuals 1n
non-acid lakes
Diaptomuro minutus formed
between 44 and 73% of all
crustacean zooplankton,
and dominant in all
non-acid lakes
In acidic Clearwater Lake,
zooplankton community
characterized by the
absence of Tropocyclops
prasinus mexicanus and
Mesocyclops edax and by
the scarcity of Cyclops
bicuspidatus thomasi and
Diaptomus minutus
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-Halib
urton Reglor of
Ontario
Acidic lakes also
have high levels
of copper and
nickel which nay
adversely effect
zoo-plankton
Samples collected
summers 1973-1977
as vertical hauls
with 80 u mesh
tow net and at 2-
to 3-n Intervals
with a plastic
trap
plankton community over
and above effects of
low pH
-------
TABLE 5-7. CONTINUED
en
i
OY
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-add lake, Rotifers generally form
collections Included only about IX of total
7 species on the zooplankton biomass In
average vs 3.7 from non-acidic oligotrophic
the add lake lakes in the Sudbury area
Standing crop
generally greater 1n
very acid (pH 4.7 to
5.2) than in slighly
acid or alkaline
ponds
About 14 species
Cladocerans
Cladocerans unimportant
1n non-acid lake,
forming <5% of the
average biomass
In the add lake,
Bosmina longirostris
comprised 403 of the
crustacean zooplankton
biomass
Species occurring in
all ponds independent
of pH:
Bosmina longirostris
Ceriodaphnia
quadrangula
Diaphanosoma
leuchtenbergianum
Copepods
Diaptomus minutus major
contributor to total
zooplankton biomass 1n the
non-acid lake. Cyclops
scutlfer, Mesocyclops
edax, and Tropocyclops
prasinus mexicanus also
important
In the acid lake,
Diaptomus- minutus
comprised 321 of the
crustacean zooplankton
biomass. Chydorus
sphaericus and Cyclops
vernal is also common
Species occuring in all
ponds independent of pH:
Diaptomus reighardl
Cyclops vernal is
tycTops bicuspidatus
thomasi
Mesocyclops edax
Others
2 individuals
of Chaoborus
flavicans
collected in
non-add
lake.
In the acidic
lake,
Chaoborus
flavicans,
Chaoborus
albatus, and
Chaoborus
americanus
occurred
Commons
Pre- and post-
fertll Ization
study of one
acidic (pH about
4.6) and one non-
acidic (pH about
6.0) lake only
pre-fertll Izotion
data included
here
Samples collected
in plexiglass
trap at 1-, 4-,
and 7-m; 76 u
mesh net
32 ponds sampled
to 10 times ove-r a
3-year period.
Samples collected
with a Clarke-
Dumpus or trans-
parent zooplankton
trap
present 1n non-add
waters were absent
from the very acid
lakes
Species occurring only
in less acid and
alkaline ponds:
Leptodora kindta
Daphnia ambiqua~
Daphnla retrocurva
Ceriod
Tac"u
jipsmina coregonl
coreqom
Holope'jrum gibberum
Species occuring only in
less acid and alkaline
ponds:
Epischura lacustrls
Diaptomus minutus
Diaptomus oregonensls
Tropocycfops prasinus
mexicanus
The acidity in
these waters Is
attributed mainly
to large amounts
of organic (humlc)
acids
-------
TABLE 5-7. CONTINUED
Location
(reference)
Changes In species composition and abundance of:
General
observations
Rotifers
Cladocerans
Copepods
Others
Comments
10. cont.
tn
Bosmlna Ipnglrostrls
was the most conslst-
ently abundant crusta-
cean In all ponds. Its
greatest numbers were
usually found In the
very add ponds.
11. Smoking
Hills area
in North-
west Terr.,
Canada
(Hutchlnson
et ai. 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 Branchlonus
urceolarls the dominant
form
number of zooplankton
species and
zooplankton biomass
sharply related to pH
(p < 0.01). For each
unit decrease on 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:
Bosmuna longlrcotrls,
5.2-5.1
Bosmina corrgoni ,
4.5-7.2
Daphnia catawba,
S.Z-6./
Daphnia ambigua,
4.5-6.6
Holopedium gibberum,
4.5-7.2
Diaphanosomean
leuchtenberglanum,
4.7-6.6
Polyphemus pedlcutus,
'4.7-7.2
Leptodora kindtH
6.3-6.4
Diaptomus minutus domi-
nated at pH < 5,
Identified pH range for
dlstlbutors of species:
Diaptomus minutus,
4.5-7.2
Cyclops scutlfer.
5.4-7.2
Mesocyclops edax,
4.5-6.7
Tropocyclops praslmus,
5.3-7.2
Epischura lacustris,
5.3-7.2
Chaoborus sp. Two stop survey
occurred (July and August
throughout 1979) of 20 lakes
the pH range (10 in NY; 13 'n
4.5 to .6.7 NH)
Samples collected
with Van Oorn
sampler at 3-5
depths per lake
All headwater
lakes; pH ranged
from 4.5 to 7.2
-------
TABLE 5-7. CONTINUED
tn
i
CTl
Changes In species composition and abundance of:
Location General Rotifers Cladocerans
(reference) observations
13. Great Identified pH range for
Britain distribution of
(Lowndes species:
1952) Diaphanosoma
brachyurum,
4.3-9.iJ
Daphnia pulex,
5.8-9.2
Daphnia longispina,
HP? .2
Ceriodaphnia
retlculata, 6.2-9.2
Ceriodaphnia
quadrangula,
3.2-9.2
Bosmina longirostrls,
6.9-9.2
Polyphemus pedlculus,
1.6-9.2
Bythotrephes
longimanus, 6.7-7.2
Leptodora kindtl ,
6.7-8.4
14. Great Number of species Fond 1n waters with
Britain lower In low pH pH < 5.0:
(Fryer waters In pH range 3 Diaphanosoma
1980) to 7. brachyurum
Ceriodaphnia
quadranguTa
Bosmina coregoni
Polyphemus pedTculus
Copepods
Identified pH range for
distribution of species:
Diaptomus gracills,
"'4.7-9.2
Cyclops abyssorum,
6.2-7.3
Cyclops vernalis,
4.4-9.2
Cyclops bicuspidatus,
4.1-9.?
Found in waters with
pH < 5.0:
Cyclops abyssorum
Tropocyclops prasinus
Others Comments
One-time sampling
of 70 w.itcr
bodies
Acidity attrib-
utable primarily
to high levels of
organic (humic)
acids
-------
NUMBER OF SPECIES PER COLLECTION
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lakes were Keratella cochlearls, Keratella serrulata, Kellicottia
longispina, Polyarthra remata, and Polya'rthra vulgaris. Species reduced
in abundance with acidification indtided Asplanchna priodonta,
Conochilus unicprm's, Conochilus mincorm's, and KeFatella hTemalis
(Aimer et al. 1974, 1978; Raddum 1978; Hultberg and Andersson 1982). In
Ontario, species of Keratella and Kell icottia were also important in
acidic lakes (Keratella taurocephala, Keratella cochlearls, Kellicottia
bostoniensis, Kellicottia longisp'Tinia) (Roff and KwiatkowskT 19/7).
Experimental acidification of Lake 223, to pH 5.4, resulted in increased
numbers of Polyarthra yulgaris, Polyarthra remata, Keratella
taurocephala, and Kell icottia longi'spina (Malley et al. 1982).
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, Van and
Strus 1980, Nilssen 1980). In addition, in laboratory experiments with
Daphnia magna and Daphnia pal ex, reductions in survival and
reproduction, and physiological imbalances 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 coregoni 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, Van and Strus 1980, Van et al. 1982). Other cladocerans common in
temperate, oligotrophic lakes (e.g., Diaphanosoma brachyurum,
Diaphanosoma leuchtenbergianum, Leptodora kindti, Holopedium gibberum,
Polyphemus pediculus, Ceriodaphnia quadrTngula. and Bythotrephes
Tongimariu?) often are less abundant in waters with pH 1 eve!s 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 the numbers of Holopedium gibberum (Malley 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
5-7).Tn addition, frequently reported as common in acidic waters are
Heterocope saliens in Scandinavia, and Cyclops vernal is, Cyclops
bicuspidatus thomasi, and Mesocyclops edax in North America"! S~pecies
noted as being more frequent in non-acidic lakes include Epischura
lacustris, Diaptomus oregonensis, Tropocyclops prasinus mexicanus,
Heterocope appendiculata, Ancanthodiaptomus denticorni's, and
MIxodiapTblmus laclniatus'. Similarly, experimental acidification of Lake
223 to pH 5.4 resulted in no consistent change in populations of
Diaptomus minutus, Cyclops bicuspitatus thomasi, and Mesocyclops edax,
but a dec!ine in numbers of Tropocyclops~prasinus mexicanus, and
5-68
-------
extinction of Epischura lacustris below pH 5.8 and Diaptomus sicilis
below pH 6.1 (Mailey et al. 1982).
Experimental acidification of Lake 223, Ontario also resulted in
the extinction of the opposum shrimp, Mysis relicta, an important
planktonic predator, below about pH 5.6 (Mailey et al. 1982).
Of the insects, midge larvae Chaoborus spp. are important
zooplankters in many lakes. Little is known about effects of acidifi-
cation on Chaoborus. although it appears to persist in some acidic
environments down to pH 4.2 to 4.5 (Scheider et al. 1975, Van et al.
1982, Confer et al. 1983, Marmorek 1983). On the other hand, Hobaek and
Raddum (1980) observed that Chaoborus flavicans was absent in
clear-water acid lakes (pH < 5.0). fTilssen (1980) reported the
extinction of Chaoborus in an acidic lake (pH 4.2 to 5.0), where
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 m-2 per unit drop in pH. Roff
and Kwiatkowksi (1977) concluded that standing crops of rotifers,
cladocerans, and cyclopoid copepods (but not calanoid copepods) were
reduced at pH levels below 5.6. The mean size of crustacean
zooplankters 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
1imno-corrals within Lake Eunice, British Columbia, resulted in no
change, or even a slight increase, in zooplankton standing crops (Malley
et al. 1982, Marmorek 1983). 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
5-69
-------
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 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,
concentrations 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, and, Epischura lacustris and Mysis
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.
5-70
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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), Daphnia ambigua (1.7 mm),
HolopedTum gibberum (1.2 mm), Diaphanosoma brachyurum (1.1 mm), and
Ceriodapffnia quad^angula (0.9 mm) (Nilssen and Pejler 1973, Makarewicz
and Likens 1979, Lynch 1980). 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 Ipngirostris) 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 dominates in acidic Cheat Lake (impacted by acid mine
drainage) because of its high reproductive potential and the intolerance
of its major predator, Mesocyclpps edax, to acidic conditions.
Populations of a number of crustacean pianktonic predators including
Epischura lacustris, Leptodora kindti, and Mysis relicta do seem to be
reduced in acidic lakes, (Nilssen 1980, Schindler and Turner 1982; Table
5-7). Data on abundance of Chaoborus are scarce and somewhat
contradictory (Section 5.5.3TIThe 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 predaceous
zooplankton in acidic lakes are probably direct consequences of acidi-
fication. Changes in these populations occur while their food supplies
are still abundant (National Research Council Canada 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
pnytoplankton community (Section 5.5.2), an important food source for
zooplankton. Some algae are more edible than others (Porter 1977). A
5-71
-------
high proportion of the phytoplankton in many acidic lakes are
dinoflagellates, a relatively large phytoplankter 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 Clear-water Lake, was
14 ym. Yet, the maximum size of a particle likely to be ingested by
Bosmina longirostris, the dominant zooplankter, was 10 to 14 ym, 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 zooplankton in this acidic lake was 5 to 18 times
lower than estimated rates for non-acidic oligothrophic 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 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 systans
phytoplankton biomass accumulates (Aimer et al. 1978, Van and Stokes
5-72
-------
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 phytopl ankton. 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 decomposition and the accumulation of coarse
detritus, benthic algae, and macrophytes (especially Sphagnum) on the
bottom of acidic lakes decreased 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
however to confirm this hypothesis.
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 C02 rather than as bicarbonates or carbonates. The
National Research Council Canada (1981) calculated that for a typical
soft-water 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 dissolved inorganic carbon available in
5-73
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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 nay 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
production (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 phytoplankon 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 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 suggestss 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
5-74
-------
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 4.4.1). 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 significant role in driving and controlling the
metabolism of lakes (Section 5.5.1.2). 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
acidification 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
° Acidification results in a marked shift in the structure of
the plankton community. For both phytoplankton and
zqolankton, the total number of species represented decreases
with increasing acidity. For zooplankton. the greatest
change in species composition occurs in trie pH range 5.0 to
5.3; for phytoplankton, in the pH interval 5.0 to 6.0.
0 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.
0 Generalizations about changes in community structure for
phytoplankton populations with acidification are more
difficult to make. In many acidic waters (but certainly not
all), dinoflagellates (Phylum Pyrrophyta) predominate.
Dinoflagellate species Peri dim'urn inconspicuum and Perl dim'urn
11mbaturn in particular are reported as dominants, often
5-75
-------
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 chrysophyceans.
0 Despite the altered structure of the plankton community,
productivity may remain unaffected. Relative to levels of
primary (phytoplankton) productivity in circumneutral lakes,
primary productivity in acidic lakes in some cases islower,
in others equal. A cause-and-effect relationship between
primary productivity 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.
0 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 whose changes may affect the plankton
community. Because of the complexities 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.
5.6 FISHES (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;
5-76
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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
physiological responses of fish to low pH, elevated aluminum
concentrations, and other water quality conditions associated
with acidification.
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 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
5-77
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and absence of various fish species. Unfortunately, only In a few
Instances are historic records available that provide concurrent
documentation 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 TCatostomus commersoni), creek chub
(Semotnus atromaculatus), lake chub (Couesiou? plurebeus), and canmar
shiner (Notropis comutus) (Greeley and Bishop 1932). Tn~ addition, a
variety of other species (e.g., smallmouth bass, Micropterus dolqmieui;
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
populations 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 610 m (Schofield 1976b); 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.
5-78
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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 (1976b, 1981, 1982) noted fewer fish species in lakes with pH
levels below 5.0 (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 populations 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 from the 1930's (Chapter E-4, Figure 4-18) (Schofield 1976b).
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 5.0 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.
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
5-79
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a:
40
30
20
50
40
30
20
10
0
NORWAY
WRIGHT et al.
1975
LA CLOCHE MOUNTAINS, ONTARIO
HARVEY 1975
ADIRONDACK MOUNTAINS, NEW YORK
SCHOFIELD 1976
LEGEND
0 NO FISH PRESENT
D FISH PRESENT
4.0 4.5 5.0 5.5 6.0 6.5 7.0 7.5
pH
Figure 5-4. Distribution of fish in relation to lake pH.
5-80
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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. Nongame 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 campgrounds 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.
5.6.2.1.1.2 Other regions of the eastern United States. Schofield
(1982) summarized available data relating water acidity and fish
5-81
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population status for areas in the eastern United States with waters
potentially acidified 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) discussed 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 (Micropterus salmoides) and
bluegull sunfish (Lepoim' s 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). 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 LaClpche 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 £-1, 8 to 12 yg Ni jr1, 24 to 36
yg Zn £-1, and 1 to 4 yg 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.
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; 63 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
5-82
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1972, Beamish 1974a,b; Beamish et al. 1975, Harvey 1975). Major sport
fishes common in these lakes include lake trout, small mouth bass, and
walleye (Stizostedion vitreum). Other fish occurinq very frequently are
yellow perch, pumpkinseed sunfish (Lepomis gibbpsus), rock bass
(Ambloplites rupestris), brown bullhead, lake herring (Coregorus
artedii), 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, smallmouth 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 smallmouth
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
5-83
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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, rockbass,
brown bullhead, and white sucker fail to spawn
Last capture of lake white fish
Lake trout rare
5-84
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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 muskellunge rare
White sucker, brown bullhead, rock bass, lake
herring, and yellow perch present
5-85
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yielded only four 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) concluded 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 alkalinity 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 Novia 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 sustained 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 distributed 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)
5-86
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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 MINNO
NUMBER OF LAKES
CONTAINING SPECIES
I—
1 __
h"
1 —
« ' ' 1 >
40
37
29
25
20
23
6
6
19
7
g
7
10
20
11
6
q
4.0 4.5 5.0
) 11 | 19 I
5.5
pH
19 I 13 | 8 I 10
NUMBER OF LAKES
6.0 6.5 7.0
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).
5-87
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TABLE 5-9. MAJOR RIVERS IN NOVA SCOTIA ON THE ATLANTIC COAST,
pH LEVELS AND STATUS OF ATLANTIC SALMON STOCKS
Mean3
PH
River 1980-81
Musquodoboit6
St. Mary's
LeHave*
Ecum Secum
Petit
Ship Harbour
Gold
Salmon (Digby)
East Ship Harbour
West Ship Harbour
Moser
Quoddy
Kirby
Medway6
Salmon (Port Dufferin)
Gaspereau
Mersey6
Middle
Li scomb
Ingram
Tangier
East
Tusket
Issacs Harbour
Nine Mile
Salmon (Lawrencetown)
Clyde
Barrington
Jordan
Sable
Broad
Roseway6
Larry 'sf
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
4.8
4.7
4.6
Recorded
Presence (+)
Rangeb or Absence (-) Regression of
pH of Salmonc Angling Catch
1979-80 <1960 1980-82 on Year3
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 +
+ NS
D
NS
NS
NS
D
D
D
NS
NS
NS
+ NS
NS
NS
D
+
NS
+
— _
+
_
_
_
_ _
.. _
_
_
_
_
-
5-1
-------
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; <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.
5-89
409-262 0-83-15
-------
200
3
CO
Of.
o
100
80
60
40
en
«3
O
Z
UJ
o
a:
o
o
20
10
8
—Q—MEAN FOR 12 RIVERS WITH pH >5.0 (1980)
O—MEAN FOR 10 RIVERS WITH pH<5.6 (1980)
1935 1940 1945 1950 1955 1960
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).
-------
probably act uniformly over the entire area (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 has increased in recent
years. Acidic conditions 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 salmon in New Brunswick,
Prince Edward Island, and other areas of Novia 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 Britian
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
5-91
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i/o
o
1900
1980
Figure 5-7. Yearly yield for Atlantic Salmon fisheries in seven rivers
from the southermost part of Norway (bottom curve) compared
with 68 rivers from the rest of the country (top curve).
(Leivestad et al. 1976).
5-92
-------
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 and
fingerlings 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-sized 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 gairdTieri). and brook trout. "ATxMTt 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+, Al3+,
and HCOs" indicated that pH and Ca2* were the two most important chemical
variables (r = 0.53).
The original data base on fish populations in Stfrlandet 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 provided 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 in 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 Sjfrlandet, the population status of brown trout has been
5-93
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LOST POPULATIONS
TROUT PRESENT
Figure 5-8. Status of brown trout populations from the affected areas
in the four southermost 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).
5-94
-------
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 Stfrlandet 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). 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 acidification, e.g., changes in land
use, specifically as they relate to historic changes in fish populations
(Drabljte and Sevaldrud 1980, Drablsfs 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 h.as 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 rut11 us), 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's-40's and the simultaneous elimination 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 definitely
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
5-95
-------
3000
2500
2000
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failed to yield any trout In most streams draining forested catchments
(mean pH 4.34), while moorland streams (mean pH 5.40) Invarlbly 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 (National Research
Council Canada 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: year 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 variations, 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 agre 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). 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 1n 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 1n 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).
5-97
-------
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 lake yielding young-of-the-year yellow perch was characterized by
a pH of 4.4, for rock bass by a pH of 4.8.
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 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 1976a, 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.
5-98
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PERCH POPULATION LAKE ST. SKARSJ0N 1976
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Figure 5-10. Liming of Lake St. Skarsjon, Sweden, in 1975 reestablised
reproduction of perch population (Muniz and Leivestad
1980a).
5-99
-------
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, in some cases loss of older
fish has been observed in acidic waters. Three lakes in the Tovdal
River, Norway, were test fished 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. Test fishing 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 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 documented 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
5-100
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CATCH YEAR |l979ll978|1977il976|l975
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LEGEND
ULE FISH(I-II)
JIT SPANNERS (III-IV)
SPAWNERS (VII- )
1967
Figure 5-11. Age distribution of brown trout in Lake Tveitvatn,
Tovdal, Norway (Rosseland et al. 1980)
5-101
-------
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 3 years, but
thereafter grew more slowly. In addition, yellow perch collected from
George Lake 1n 1973 and 1974 (pH 4.6) at age 1 to 4 years were
significantly larger than perch of the same age collected 1n 1966 (pH
5.8); this trend was reversed for age groups 5 years and older. Up to
age 4, 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 1n acidic lakes and streams.
Observations of increased or unchanged growth 1n 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 competition 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.
5-102
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Detailed studies of effects of food availability on fish, at all
life history stages, in acidic waters are not, however, available.
Therefore, the conclusion 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
acidifying 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 associ-
ated 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 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 obser-
vations 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. pH measurements (colormetric) 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.
5-103
-------
Leivestad and Mum'z (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 snowmelting. Leivestad and Muniz
(1976) proposed that-increased acidity interferred with osmoregulation
via perhaps 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 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 £-!). Mortalities occurred over a 5-day period 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
5-104
-------
concentrations in the lakes, followed by a long period of dry, sunny
weather preceded fish kills in both lakes. pH levels in the lake
epilimnion 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 precipitates out of solution. Dlckson
(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, SchofieW 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 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
Morris 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 1n 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 Precambrlan Shield of western Ontario.
Prior to acidification, surface waters had an average alkalinity of
about'80 yeq sr1 and pH of 6.5 to 6.9. Five species of fish were
present: lake trout, white sucker, fathead minnow (P1mephales promelas),
5-105
-------
pearl dace (Semotolus margarita) and slimy sculpin (Cottus cognatus).
Beginning in 1976, additions of 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, Malley et
al. 1982, Schindler and Turner 1982, Mills 1982, 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
relicta), a henthic/planktonic crustacean (Section 5.5.3), and the
collapse of the fathead minnow population. Although both these species
were important prey for lake trout in the'lake, 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.
The population of bottom-dwelling 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 disappeared 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,
5-106
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TABLE 5-10. BIOLOGICAL CHANGES IN LAKE 223 IN RESPONSE TO
EXPERIMENTAL ACIDIFICATION (MILLS 1982, 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 dlpteran emergence
5.8-6.0 Disappearance of the opossum shrimp (Mysis relicta)
Reproductive Impairment of the fathead minnow (Pimephales
promelas)
Possible Increased embryonic mortality of lake trout
(Salyelinus namaycush)
Inhibition of calcification of exoskeleton of crayfish
(Orconectes vlrlUs)
Disappearance of the copepod Dlaptomus sicills
5.3-5.8 Increased hypollmnetlc primary production
Development of Mougeotea algal mats along shoreline
Increased infestation of crayfish with a parasite Thelohania
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 commersonl)
Increased abundance of lake trout
Disappearance of copepod Epischura lacystris
First appearance of the cladoceran DaphnlaTcatawba x
schoedleri
Below 5.3 Recruitment failure of lake trout
Recruitment failure of white sucker
5-107
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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
severe 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--
Norn's Brook, a third order stream in the Hubbard Brook Experimental
Forest, 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 srl. 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 (H^SO^, HN03, HC ) into
nonacidic background water. These artificially acidic waters may not
precisely mimic acidified surface waters, and, as a result, fish
responses recorded in laboratory bioassays may not always accurately
represent what would occur in the field. In this section, in situ
exposures of fish to 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 four locations in a control, nonacidic lake (Harvey et al.
5-108
-------
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 for
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 simultaneous 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
finger!ing 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.
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
5-109
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TABLE 5-11. SUMMARY OF FIELD EXPERIMENTS WITH FISH EGGS
EXPOSED TO ACIDIC SURFACE WATERS
Species
Brook trout9
Brown troutd
Brown troutf
Atlantic
salmon*1
Location
Hatchery with
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 Reference
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
g
d
f
d
Atlantic
sal monk
Atlantic
salmon3
River and a
near-neutral
tributary,
Norway
In situ at
several rivers
in Stfrlandet
Norway
In situ in
streams,
Scotland
5.0
5.5
4.2
4.4
4.9
5.8
0
0
54
30
Critical pH
for hatching
Comparison of
forested vs non-
forested catchments
5-110
-------
TABLE 5-11. CONTINUED
Species
Roache
Walleyec
Location
pH % Survival Comments
Reference
Perche
In situ in
Lakes
Stensjon,
Trehorningen,
and Malaren,
Sweden
4.7
5.7
7.5
28
50
89
e
As above
In situ in
series of
small streams
in LaCloche
Mt. area,
Ontari o
4.7
5.7
7.5
4.6
6.7
0
14
100
Hatching success
significantly
reduced at pH
References
aHarriman and Morrison 1982
bHendrey and Wright 1976; Muniz and Leivestad 1980a
cHulsman and Powles 1981
Leivestad et al. 1976
eMilbrink and Johansson 1975
fMuniz and Leivestad 1980a
gSchofield 1965
5-111
-------
of fish from the same streams in an electrofishing survey. Finally,
Milbrink and Johansson (1975) incubated perch (Perca fluviatilis) 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
disappeared from Lake Trehorningen. Roach are still prevalent in both
Lake Stensjon and Malaren.
5.6.4 Laboratory Experiments (J. P. Baker and P. G. Daye)
One of the best ways to prove cause and effect is to conduct
experiments in a carefully controlled environment, i.e., the laboratory.
Experimental conditions and fish response can be clearly quantified and
dose-response relationships 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 laboratory 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 qualitative 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 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
5-112
-------
(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 4 months (finger!ings) 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., temperature,
flow-through or statij: 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 2 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, and Tathead minnow
indicated no substantial reductions in survival at the lowest pH levels
tested, 5.0, 4.8 and 4.6, respectively.
5-113
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TABLE 5-12. MEDIAN SURVIVAL TIME (HR) FOR FISH EXPOSED TO pH LEVELS
en
i
Age/ 2.0- 2.6-
Species size 2.5 2.8
Brook trout * 10-60 g < 1
fngl*
2 9 1 2
90 g 1 4
60-130 g < 1
50 g
50-90 g
Rainbow trout * 1 g
* 130 g 1-4
* 200-300 g 2 2
2-5 g
* 2-5 g
* 5-15 cm
5-15 cm
Brown trout * 1-5 g
* 6 g 1-2
* 60-80 g 3 4
3
Arctic char * 100-170 g 3
White sucker 7 mo
Roach 7-13 cm
*Experiments using low alkalinity water.
tfngl = fingerling, age 0+, weight usually
References -
a. Daye and Garside 1975
b. 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
3.0-
3.1
2-3
3-6
9
1
4
5
1
< 1
1
2
3-7
9
4
1
< 1
< 50 g.
PH level
3.2- 3.4- 3.6- 3.8- 4.0- 4.2- 4.4-
3.3 3.5 3.7 3.9 4.1 4.3 4.5
7
3-6 6-18 10-38 14-51 20-270
12-14 45
18 61-66 334
5-9
25 66-70
10-32
8 37
23
8 18
2 3 6 17 83 117 133
1 2 6 27 133
2 3 8 22 70
3 7 18 55
120
25 40 2-4
2 5 10 30-200 350 1000
1 3
h. Edwards and Hjeldnes 1977
1. 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
Reference
a
b
c
c
d
e
f
g
g
h
i
i
j
j
k
1
h
h
m
j
-------
TABLE 5-13. PERCENT SURVIVAL OF FISH FOLLOWING CHRONIC EXPOSURE TO LOW pH LEVELS
Species
Brook trout
*
*
Rainbow trout*
Brown trout*
Arctic char*
Fathead Minnow
Hagflsh*
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
86
80
36
5.2-
5.6
100
96
98
100
75
79
5.9- 6.5- 7.0-
6.2 6.8 7.5
100
100
75 100
97
95
100
85 75 85
100 93
Reference
a
b
c
d
e
f
f
f
9
h
•Experiments using low alkalinity water.
References
'Dively et al. 1977
bDaye and Garslde 1975
C8aker 1981
dSwartz et al. 1978
eMenendez 1976
fEdwards and Hjeldnes 1977
SMount 1973
"Craig and Baksl 1977
-------
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 deposition 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 responsible 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 fish!ess lakes
was correlated with both pH level and calcium level, with the greatest
number of fishless lakes having both low pH and low calcium (Wright and
Snekvik 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
development stages and reproductive 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 n. nevadensls)' (Lee and Gerking 1981) (Figure
5-12). pH level had some eTfect 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 production appeared particularly
sensitive to low pH, with reductions in eggs produced per female at pH
levels between 6.0 and 7.0. Lee and Gerking (1981) concluded that
5-116
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Q O
UJ t-
o. o
oo
M-
OO O
C5
O &«
UJ
UJ TJ
i S-
rj 80
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CJ
60
tt 40
UJ
CO
2 20
0' ' ' ' '
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-------
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 yeq £"1
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 acidification 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 hatching, 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
relationship 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).
5-118
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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
a ATLANTIC SALMON
Figure 5-13. Effect of low pH on survival of fish through hatching.
References:
a Baker and Schofield 1982 g
b Swarts et al. 1978 h
c Trojnar 1977a i
d Trojnar 1977b j
e Johansson et al. 1977 k
f Mount 1973
Carrick 1979
Runn et al. (in 1975) 1977
Johansson and Mil brink 1976
Peterson et al. 1980a
Peterson et al. 1980b
5-119
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100
80
60
40
20 -
3.5
1
I /
/ /
I /
I /
LEGEND
/ • BROOK TROUT
/ o WHITE SUCKER
• ATLANTIC SALMON
°BROWN TROUT
xPIKE
pH
PH
Figure 5-14.
Effect of low pH on survival of fish as sac fry. Solid
line, sac fry survival through swin-up following
development of eggs and hatching of larvae in low pH water
(expressed as percent normal hatch); Dashed line, sac fry
survival without previous exposure to low pH.
PART (A)
References
a Baker and Schofield 1982
b Swarts et al. 1978
c Johansson et al. 1977
d Trojnar 19775
PART (B)
a Daye and Garside 1975
b Johansson and Kihlstrom 1975
c Johansson et al. 1977
5-120
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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 Milbrink
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). Zotin (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 intermediate periods of
development occurs but is limited (Kalman 1959, Zotin 1965).
Given the evidence that timing of exposure substantially affects
the sensitivity 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 laboratory 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).
5-121
t09-262 0-83-16
-------
100
oo
°- 40 -
20
LEGEND
• BROOK TROUT
O WHITE SUCKER
Figure 5-15.
pH
Effect of pH on survival of fry exposed for 14 days
after swim-up and initiation of feeding.
jjBaker and Schofield 1982
bTrojnar 1977a; previous exposure during development at
pH 8.'0 (o); previous exposure at pH 4.6 to 5.6 (•).
5-122
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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) and Swarts et al. (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 sensitive 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
increased 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 7.0 to 7.2 (Nelson 1982).
The decrease in growth at low pH represents a sublethal response to
elevated hydrogen ion concentrations and suggests that fish are
5-123
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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. Mum'z and Leivestad (1980a) reported observations of trout
concentrated into "refuge areas" 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 the ability of
fish to detect and use these areas are 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.Fish were able to detect and avoid waters with pH
< 5.4, a value slightly above the lethal level of pH 5.0. Hoglund
T1961) 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
5-124
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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.
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; McWilllams 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 solutions
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 freshwater 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 4.3.2.6)
and previous exposure of fish to high acidity are crucial factors in
determining the response of fish exposed to sudden acid episodes.
5-125
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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 flux of
hydrogen ions across gill membranes into the blood. McDonald et al.
(1980) noted that in moderately high alkalinity 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
r1), 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 metabolic reactions. The oxygen carrying capacity of fish
blood drops sharply below a blood pH level of 7.0 (Green and Root 1933,
Prossner and Brown 1961). 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 (pOg). 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, Frornm 1980).
Finally, Nelson (1982) reported that ossification, amount of
calcium deposited 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 surrounding water and from food intake. A
decrease in skeletal ossification at low pH level could be partially
responsible for increased incidence of skeletal 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
5-126
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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
deecressing pH levels in acidified surface waters (Chapter E-4, Section
4.6). 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 regulation 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 southern 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 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 4.3.4.1). 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 specifically 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.
5-127
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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 jr* (Schofield 1976b, 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 al. 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 £-!. 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 «,-!. 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 «,-!.
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 jr1 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 r^ 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 km of
Sudbury, concentrations of cadmium, copper, lead, and nickel measured in
acidic Ontario surface waters ranged up to about 0.6, 9, 6, and 48 yg
Ni £-1, respectively (Spry et al. 1981). In southwestern Sweden,
concentrations of cadmium in acidic waters were measured up to 0.3 yg
A-l; lead up to 5 yg t-1 (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 £-1, 9.5 to 77 yg Cu A-l, 13 to 253 yg Pb £-1, and 380 yg Ni £-1.
In general, all of these reported toxic concentrations and/or
maximum acceptable concentrations for zinc, manganese, cadmium, copper,
lead, and nickel are above the highest levels of these metals measured
in acidic surface waters of Scandinavia and eastern North America
(unless a local source of metal emissions exists). However, the lack of
sufficient bioassay data collected in low alkalinity, acidic waters
makes this statement tentative. In addition, sublethal and additive or
synergistic effects with other toxic components in acidified surface
waters cannot be ruled out.
Aluminum, on the other hand, has been found to be toxic to fish at
concentrations as low as 0.1 to 0.2 mg jr1 (Schofield and Trojnar
1980, Muniz and Leivestad 19805, Baker and Schofield 1982), a level
within the range of concentrations measured in acidic surface waters.
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Total aluminum levels measured ranged up to 1.4 mg a~l in the
Adirondack region, New York (Schofield 1976b), 0.76 mg i'1 in
southwestern Sweden (Dickson 1975, Wenblad and Johansson 1980), 0.6 ing
JT1 in southern Norway (Wright et al. 1980), and 0.8 mg jr1 in
the Pine Barrens of New Jersey (Budd et al. 1982). In addition,
analysis of survival of brook trout stocked into 53 Adirondack lakes as
a function of 12 water quality parameters indicated aluminum to be a
primary chemical factor controlling trout survival (Schofield and
Trojnar 1980; Section 5.6.2.1.2.1).
Baker (1981, 1982), Baker and Schofield (1982), and Driscoll et al.
(1980) examined the effect of aluminum complexation on aluminum toxicity
to fish in laboratory experiments. Complexation of aluminum with
organic chelates appeared to eliminate aluminum toxicity to fry, and
survival of brook trout and white sucker fry in acidic Adirondack waters
correlated most accurately with inorganic aluminum concentrations and
pH. The toxicity of a given inorganic aluminum concentration varied at
different pH levels and with different life history stages. At low pH
levels (4.2 to 4.8), the presence of aluminum was beneficial to egg
survival. In contrast, in experiments with sac fry and fry, aluminum
concentrations of 0.1 mg &~1 (for white suckers) or 0.2 mg &"1
(for brook trout) and greater resulted in measurable reductions in
survival and growth at all pH levels (Schofield and Trojnar 1980, Baker
and Schofield 1982, Muniz and Leivestad 1980a).
The toxic action of aluminum seems to be a combined effect of
impaired ion exchange and respiratory distress caused by mucous clogging
of the gills (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 £~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 £-1 at pH 4.4 and higher.
Aluminum was particularly toxic in solutions over-saturated with
aluminum 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 the 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).
5-129
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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, including lake trout populations from 17
lakes, smallmouth 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
simultaneous 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-19601 s to the present.
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 associated 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 has been recorded
since about 1940 (Sevaldrud et al. 1980). Today, more than 50
percent of the original populations have been lost, and
approximately 60 correlated with acidity, acidification, and/or
inputs of acidic deposition (Wright and Snekvik 1978).
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.
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 significant 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
5-130
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(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 production
potential represents 30 percent of the Novia Scotia resources
but only 2 percent of the total Canadian potential (Watt 1981).
° 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
1976b), 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.
o In other regions of the world with low alkalinity waters and
receiving acidic deposition (e.g., Muskoka-Haliburton area of
Ontario and Maine) (Sections 5.6.2.1.1.1 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.
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:
0 The influence of food chain effects on decreases in fish
populations in acidified waters has received little attention to
date, but available information suggests it plays a relatively
minor role (Beamish 1974b, Hendrey and Wright 1976, Muniz and
Leivestad 1980a, Rosseland et al. 1980). With acidification, or
in comparisons between acidic and circumneutral lakes, fish
5-131
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growth is often unaffected or increased with increasing acidity
(Section 5.6.2.3). Some important prey organisms are sensitive
to acidic conditions and disappear with acidification 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
1982). 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.
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 snow melt 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 5.6.4.1.51. 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 recruitment failure in
several acidic lakes in the LaCloche Mountain region, Ontario
was associated with a failure of the female fish to spawn.
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).
It is likely that each one of these factors plays some role in
recruitment failure but the importance of each factor probably
varies substantially among aquatic systems, depending on the
particular circumstances.
5-132
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0 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 populations with acidification of surface
waters.
5.6.5.3 Relationship Between Water Acidity and Fish Population
Response—To assess the impact 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.
The1 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 circumneutral tributary or small segment of a lake may
provide an area for successful 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 in the field and
in the analysis of the complexity of the natural environment 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),
and then assess patterns and reasons for deviations from this initial
5-133
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TABLE 5-14. SUMMARY OF FIELD OBSERVATIONS, FIELD EXPERIMENTS, AND LABORATORY EXPERIMENTS
RELATING WATER pH TO FISH RESPONSE
tn
*^
+^
FIELD OBSERVATIONS
Recruitment failure
Population loss
Fish kill
FIELD EXPERIMENTS
Recruitment Failure
Population extinction
Adult mortality
Embryo mortality
LABORATORY EXPERIMENTS
Adult mortality-
acute (2 day)
Adult mortality-
chronic (> 20 day)
Embryo mortality
Fry mortality
Reduced production
of viable eggs
Reduced growth
Avoidance
FIELD OBSERVATIONS
Recruitment failure
Population lots
Brook Lake Arctic
Trout Trout Char
5.0-5.5* S.2-5.50 5.2*
6.9«
5.7*
5.0« 5.0-5.5'
4.5-4.8J 4.4*
5.9"
5.1-5.34
4.8-5. jn
4.7-5.1*
4.4-4. 6l
4.S-4.6S
3.8«
3.500
4i4ff < 4.899
4.511
•4.1JJ
6.1"
4. add
<4.6PP
•4.4-4.9"
•4.5-4.8JJ
,;2dd
<4.4kk
4.0-5.0PP
5.1"
6.5" 4.899
4.5»»
Lake Herring Lake Smallmouth
Uhtteflsh Bass
4.5-4. 7b 5.0* 5.5-6.QO
5.0C
4.4C 4.4« 4.4<=
Brown Rainbow Atlantic Unite European Walleye Fathead
Trout Trout Salmon Sucker Perch Minnow
< 5 .IK 4.7-5.011 5.2* 4.4-4.9C 5. 5-6. Ob
4.7-5. if 4.7-5.2° 5.0-5.59 5.4«
5.0e <4.7h 6.5'
4.5-4.8J 5.5-6.0J S.I* 5.1«
4.5-5.0' 4.3* 5.2<
5. 1" 5.5<
4.9-5.1*
4.6J 3.9-4.2°
5.01 5.0"
5.1-S.jt 5.8-6.01
5.3-5.14
4.0-5.01"
4.5° 4.5-5.0' 4.7-5.7" S.4»
4.8* ' 5.0-5.5°
5.1" 4.9*
3.8y 4.0* 3.9** < 4.6"
3.8-4. 8«
4. 0-4 .Zee
< 4.899 < 5.099 5.09 S.91*
5.6"
4.1" 4.1k* +> s.&JJ
4.0-4.5"" S.snn s.joo
S.S"11*
4i544
4.4** 3.7-4.0T «5.4-S.63J < S.*"!"
5.0** 5.0-5.4«°
6.6hh
4.899 4.899
< 5.0" 4.3-4.8U"
S.ww
Largenouth Rock Pumpklnseed Blueglll Yellow Common Bluntnose
Bass Bass Sunflsh Perch Shiner Minnow
5. 1C 4.7-5.20 5.0<1 4.5-4.7° 5.5-6.01
5.011 S.O* 4.4' 5.7C
4.4C 4.8-5.01" 4.3<: 4.4« 4.4-5.0* 5.7°
6.0<1 4.3' 4.3C
Roach Northern SI Imy Brown
P1ke Sculpln Bullhead
5.5« 5.0e 4.7-5.20
5.1' 5.01 4.9«
4.4-4.9C 5.2'
4.7h
4.7' 4.3« 4J-5.0*
4.7e
5.3-5.84
5.7°
5.6"
4.2-5,2"
ww .
•
Lake Creek Trout Gadldae
Chub Chun Perch Burbot
4.5-4.7" 5.2-5.50 5.5-6.0>>
4.5-5.0* 5.0* 5.»d
-------
REFERENCES
a Schofield 1976c
b Beamish 1976
c Aimer et al. 1978
d Watt et al. 1983
e Harvey 1979
f Hultberg 1977
g Runn et al. 1977
h Grahn et al. 1974
i Beamish et al. 1975
j Grande et al. 1978
k Leivestad et al. 1976
1 Harriman and Morrison 1982
m Overrein et al. 1980
n Schofield and Trojnar 1980
o Jensen and Snevik 1972
p Farmer et al. 1981
q Mills 1982
r Harvey et al. 1982
s Schofield 1965
t Dunson and Martin 1973
u Milbrink and Johansson 1975
v Hul sman and Powles 1981 as
reported in MO I 1982
w Mimiz and Leivestad 1980
x Johnson 1975
y Brown 1981
z Kwain 1975
aa Beamish 1972
bb Robinson et al. 1976
cc McDonald et al. 1980
dd Swarts et al. 1978
ee Lloyd and Jordan 1964
ff Swartz et al. 1978
gg Edwards and Hjeldnes 1977
hh Mount 1973
ii Menendez 1976 (e from Table 5-22)
jj Baker and Schofield 1982
kk Johansson et al. 1977
11 Johansson and Milbank 1976
mm Carrick 1979
nn Peterson et al. 1980a
oo Trojnar 1977b
pp Trojnar 1977b
qq Peterson et al. 1980b (conf)
rr Daye and Garside 1976
ss Johansson and Kihlstrom 1975
tt Jacobsen 1977
uu Nelson 1982
vv Johnson and Webster 1977
ww Hoglund 1961
xx Ryan and Harvey 1977
yy 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).
5-135
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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 PH
pH RANGE OF SUCCESSFUL REPRODUCTION
pH LEVELS AT WHICH POPULATIONS OCCUR
VARIATIONS IN OBSERVED LOWER pH LIMITS
6.0
6.5
Figure 5-16.
Initial estimates of relationship between acidity and
fish response, based on references in Table 5-14.
5-136
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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
concentrations, 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 maculatum and A_. jeffersom'anum breed in shallow hilltop ponds
that have pH values 1.5 pH units less than nearby permanent ponds in New
York State (Rough and Wilson 1977). Spotted salamander (A_. maculatum)
egg mortality increased to > 60 percent in water less than pH 6.0, a
substantial rise from the normal mortality of < 1 percent at pH 7.0. In
contrast, the Jefferson salamander, A. jef fersonianum, was most
successful at pH 5.0 to 6.0 (Rough 1"976TT The preference for neutral
water by adult spotted salamanders may be a result of the absence of
their preferred prey, the stickleback (Eucalia), from acidic water
(Bishop 1941). When a stretch of stream was artificially acidified to
pH 4.0, "salamanders" were reported (Hall and Likens 1980a) to leave the
water. Elsewhere in its range 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 (A. maculatum) preferred neutral
substrates (Mushinsky and Brodie 1975).
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 salamanders, 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 sala-
manders were terrestrial, but all salamanders are aquatic as larvae. Not
only do amphibians provide energy for birds and mammals, but they repre-
sent the top predators in many temporary ponds (Orser and Shure 1972).
The species specific tolerance of amphibians to low pH was
confirmed by a survey of newts 1n England (Cooke and Frazer 1976).
Smooth newts (Triturus vulgaris) were rarely encountered 1n 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
5-137
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acid-stress is demonstrated by the american red-spotted newt,
Notophthalamus viridescens, which one of us (RS) has observed at 6 m in
acidic (pH 4.9) Woods Lake. This 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 sensitive to acidity, too. Calling densities (an
estimate of population size) of spring peepers (Hyla crucifer) were
positively correlated with the pH of water in whTcTTthey occurred (Clark
and Euler 1980). Bullfrogs (Rana catesbeiana) (Clark and Euler 1980,
Cecil and Just 1979, Saber and Dunson 1978), wood frogs (R. sylvatica)
(Clark and Euler 1980) the common frog (R_. temporaria) (rfagstrom 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
moldy 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 for a few hours suffered 85 percent mortality. Noble (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 R. pi pi ens to acidity in the
laboratory, one of us (RS) has seen aduTt leopard frogs in an acidic (pH
4.8) Adirondack lake. The larvae may have been breeding 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 (Hastrom 1977), aTuTthe natterjack toad
(Bufo calamita) was not found below pH 5.0 (Beebee and Griffin 1977) in
EngTaruT
The mechanism by which acidity effects 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 Smokey Mountains National Park. Another mechani sm may be the
inability to control ion fluxes across membranes against strong H+
gradients. This has been indicated in fish (Section 5.6), invertebrates
(Sections 5.3 and 5.5), and frogs (Fromm 1981).
5.7.2 Birds
5.7.2.1 Food Chain Alterations--Direct effects of acidity on birds are
not expected, but indirect effects by alterations in food resources and
bioaccumulation of toxic metals is possible. 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
5-138
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diet of the common loon, Gavia Imrner, 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) but habitat
restriction unrelated to acidification is important in. 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. 1980a). Birds are opportunistic
feeders. The alteration of a food resource from a number of lakes may
reduce the population but not cause a total loss of population as the
birds switch to other resources and to other lakes in the region.
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 the less acid
lakes studied (DesGranges and Houde 1981). Blancher (1982) observed
that weight gain of kingbirds was related to insect emergence, not lake
pH. Lake pH was not correlated with densities of red-winged blackbirds
(Agelaius phom'ceus) and barn swallows (Hirundo rustica).
5.7.2.2 Heavy Metal Accumulation—Al terations of food resources may not
be the only mechanism by which birds may be inhibited by acidity. The
mobilization of metals at low pH (Chapter E-4, Section 4.6) may result
in increased body burdens in higher trophic levels. Studies by Nyholm
and Myhrberg (1977) and Nyholm (1981) have implicated aluminum in the
impaired breeding of four species of passerines in Sweden. Aluminum is
quite insoluble in the alkaline conditions characteristics of vertebrate
intestines, but it might be actively transported across the intenstinal
barrier if calcium is in short supply. Effects were manifested by
reductions in breeding success; formation of thin, porous eggs; small
clutch size; and lower egg weight near acidic lakes. The causal link
with lake pH was suggested to be the high aluminum content of the
insects near the acidic lakes (Nyholm 1981). A laboratory study (Gilani
and Chatzinoff 1981) proved that aluminum is toxic to bird embryos but
5-139
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results from Al injected into eggs are not comparable to field responses
to dietary Al. Similar findings of decreased egg size and weight were
found for the eastern kingbird (Tyranus tyranus) in Ontario (Blancher
1982). Mercury levels were found to be elevated 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 more
research.
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 (Procypn 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
characteristic of North America. In remote areas of Sweden, however,
cadmium accumulated 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 potentially severe. 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/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 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 nutrients (Chapter E-2, Section
2.3.3.3). One mineral 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),
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and uptake by grasses is inhibited by SOX (Davies and Watkinson 1966,
Gissel-Nielsen 1973), so concentrations of selenium in forage are
reduced in areas of sensitive 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 All away 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 minerals, like
molybdenum (Chapter E-2, Section 2.3.3.3), may be equally important but
have not yet been considered.
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 summartized
in Table 5-15. Many of the effects are expected to take many 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.
Sensitivity 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.
o 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.
o Mammals that feed on plants and animals in acidic lakes may
accumulate higher than normal body burdens of heavy metals, but
population losses have not yet been demonstrated.
o 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.
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TABLE 5-15. SUMMARY OF EFFECTS OF ACIDITY ON NON-FISH VERTEBRATES
en
i
ro
Taxa
AMPHIBIA
Ambystoma maculatum
A. jeffersonianum
Trituris vulgaris
T. helveticus
Motophythalamus
viridescens
Hyla cruel fer
Rana catesbeiana
R. sylvatica
R. temporaria
R. pipiens
Acrls 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 nunber 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 manlp.
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 Brodle 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
Hagstron 1977
Noble 1979
Schlicter 1981
pers. obs. (RS)
Gosner and Black 1957
Hagstrom 1977
-------
TABLE 5-15. CONTINUED
en
i
co
Taxa
B. calamita
BIRDS
Gavia immer
Mergus merganser
Megaceryle alcyon
Iridoprocne bicolor
Anas rubripes
Eucephala clangula
Tyranus tyranus
Passerines
MAMMALS
Procyon lotor
Capreolus capreolus
Al ces alces
Common name
Natterjack toad
Common loon
Common merganser
Kingfisher
Tree swallow
Black duck
Goldeneye duck
Eastern kingbird
Songbirds (4 sp)
Raccoon
Roe deer
Moose
Observation
Not observed < pH 5.0
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
Decreased egg weight
near acidic lakes
Breeding failure, thin,
porous eggs
5 x normal (Hg)
Cd accumulation
Cd accumulation
Mechanism
?
Land use changes,
fish losses?
Fish losses
Fish losses
Fish losses
Aquatic Insect losses
Abundance of preda-
tory Insect food
1 terns
From Hg In Insects
Aluminum toxicity?
Aluminum in Insect
prey
Bioaccumulation
Bioaccumulation
Bioaccumulation
Evidence
Field obs.
Field obs.
Field obs.
Field obs.
Field obs.
Field obs.
Field obs.
Lab analysis
Lab analysis
Lab analysis
Lab analysis
Lab analysis
Lab analysis
References
Beebee and Griffin 1977
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. 1980a
Blancher 1982
Nyholm 1981, Nyholm and
Myhrberg 1977
Wren et al. 1980
Frank et al . 1981
Frank et al. 1981,
Rangifer sp.
Caribou
Loss of winter browse
over a long period
Sulfur sensitivity of Lab study
caribou lichen
Mattson et al. 1981
Lechowicz 1982
-------
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. J.
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 concern 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
(5.3.1.5). 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.
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
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FISH
ZOOPLANKTON
SMALL GRAZERS AND \
DETRITIVORES \
t "\\
MACROPHYTES AND \\
PERIPHYTON \N
DECOMPOSERS
I
DETRITUS
[NON-ACIDIFIED LAKE]
INVERTEBRATE
PREDATORS
V
SMALL GRAZERS AND
DETRITIVORES
\
MACROPHYTES AND
PERIPHYTON
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.
5-145
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is disputed (Allen and Starr 1982). Marmorek (1983) and Yan et al.
(1982) observed in field experiments that acidification indirectly
reduced the short-term stability and reslience 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 Nutrient Cycling--It 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 5.5.2.3.2), and also 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 theoretically 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 autochtonous 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 observations from field studies reviewed in Section
5.5.2.1.2 indicate, however, that in most cases acidification has no
consistent adverse effect on primary productivity. Adverse effects of
decreased nutrient availability on water column primary productivity may
5-146
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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
(including 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 examination of this phenomenon has been
limited. Fish have been observed in laboratory experiments to grow more
slowly at lower pH levels (Section 5.6.2.8). Primary producers in some
acidic waters (Sections 5.5.2.1.2 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
efficiencies are reduced in acidic environments, transfer of energy
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 5.5.3.3, 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 apparently decreases conversion of
primary production into secondary production (Section 5.3.2.5.4).
Processing of detrital particles may also be affected. Again, some
evidence suggests there may be inhibition of energy cycling and energy
transfer through the food chain.
One of the best documented changes associated with acidification is
the decline and loss of fish populations which represent major
components of upper trophic levels in aquatic ecosystems. Loss of fish
populations results in a shortened aquatic food chain (Section 5.7).
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. To
date, the limited evidence available suggests that ecosystem functions
5-147
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are relatively robust, but additional research is required before final
conclusions can be reached.
5.9 MITIGATIVE OPTIONS RELATIVE TO BIOLOGICAL POPULATIONS AT RISK
(C. T. Driscoll and G. C. Schafran)
The concept of surface water neutralization as a result of base,
carbon, and phosphorus additions is discussed in Chapter E-4, Section
4.4. The biological response to these additions and other mitigative
options for fish populations 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 concentrations have been observed to decline drastically
This phenomena 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 phytoplankton 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 note
-------
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).
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
(Van 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 organisms decrease substantially. Chironomids, which were
observed to be dominant prior to neutralization (Scheider et al. 1975,
Van and Dillon 1981), contributed 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 felt!ike structure of algal filaments, detritus,
and Sphagnum completely cover lake sediments and deplete 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 x
yr-1) as compared to acidic lake populations (8 to 10 cm x 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
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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 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.
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
(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 finger!ings (5.8
+ 0.8 g) in water synthetically acidified to pH values of 2.5, 3.0, and
T.O 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 finger!ings tested
at very low pH values and those determined 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
5-150
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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 heritabilities
(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.
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 LTso 195.5 hr) than those of the
non-resistant cohort (LT$Q 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 inferior performance 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) hypothesize 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)
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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 sublethal pH values and then tested for increased survival
at lethal pH values have produced negative results. Lloyd and Jordan
(1964) acclimated 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 two 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 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, Daye (1980) could find no difference in tolerance
between the different groups and thus concluded no acclimation had
occurred. In a similar 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 survival ship, was
not determined in this study.
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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 Jordan 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 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 £-!, 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 Techniquest to Improve Fish Survival—In 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 acidification is such that a natural fishery is no longer
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409-262 0-83-17
-------
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 since 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 re-establish 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.
5.9.2.5 Summary--All three techniques for producing the 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
acidification is sometimes conflicting, some effects have been well
documented. These are summarized below (Section 5.10.1). Emphasis is
placed on those biological changes which 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 laboratory 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.
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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
0 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.
o 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.
0 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.
o 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
0 Dominant macrophyte species are the same in both acidified (pH
less than 5.6) and nonacidified (pH 5.6 to 7.5) North American
lakes.
0 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,
soft-water lakes surveyed in North America.
o Standing crops of macrophytes vary widely (5 to 500 g dry wt
m-2) in soft-water, oligotrophic lakes, and acidification
produces no consistent changes in standing crop. In Lobe!ia
dortmanna, a common plant in soft-water, oligotrophic lakes,
oxygen production was reduced 75 percent at pH 4.0 vs pH 4.3 to
5.5 in one flow-through laboratory experiment.
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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
Benthos
Molluscs
{most species except
fingernail clams, family
sphaerlidae)
Crayfish
Amphi pods
(Gammarus)
Mayfly larvae
(Ephemeroptera)
Numerous
(Section 5.2 and
5.3)
Aimer et al. 1978 Mills 1982
K. Okland 1980c
Sutcliffe and
Carrlck 1973
Numerous (Section
5.3)
Hall et al. 1980
Mai ley 1980
Costa 1967
Borgstrom and
Hendrey 1976
Bell and
Nebeker 1969,
Bell 1971
Water striders (Gerridae), Numerous (Section
backswimmers (Notonectidac), 5.3)
water boatmen (Corixidae),
beetles (Dytiscidae,
Gyrinidae), dragonfl ies
(Odonata)
Benthos community structure Numerous (Section
5.2 and 5.3)
Hall et al. 1980
Benthlc algae
(periphyton)
Numerous (Section
5.3)
Bell and
Nebeker 1969,
Bell 1971,
Mai ley 1980
Hall et al. 1980 Hendrey 1976
Schindler 1980
The calcareous shell of these animals
is 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, in the laboratory
avoids pH 6.2 and lower.
Most species decline or are absent 1n
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, amphipobs, 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
Eriocaul on sp.
Lobelia sp.
Plankton
en
i
en
Zooplankton community
structure
Phytoplankton community
structure
Fishes
Fatheal Minnow
(Pimephales promelas)
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.
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 1982
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. Kellicottia,
Polyarthra); cladocerans
TBosmina); and copepods
(biaptonus) .
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 phytopl ankton they frequently
replace.
Mount 1973 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
en
00
Darters
(Etheestoma exile, £.
nigtum, Percina capnodes)
and Mi nnows (several
Notropis spp. Pimephales
notatus)
Smallmouth Bass
(Micropterus dolomieui)
Lake Trout
(Salvelinus namayeusch)
White Sucker
(Catostomus commersoni)
Rainbow Trout
(Salmo gairdneri)
Atlantic Salmon
(Salmo salar)
Brown Trout
(Salmo trutta)
Brook Trout
(Salvelinus fontinalis)
Sunfishes
(Ambloplites rupestris,
Micropterus salmoides,
Lepomis spp.l
Yellow Perch
(Perca flavescens)
Decomposition
Harvey 1980
Rahel and Magnuson
1983
Beamish 1976,
Harvey 1980, Rahel
and Magnuson 1983
Beamish 1976, Mills 1982
Beamish et al1. 1976
Rahel and
Magnuson 1983
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 1982
Hall et al. 1980
Smith 1957
Beamish 1972
Trojnar 1977a
Numerous
(Section 5.6)
Numerous
(Section 5.6)
Numerous
(Section 5.6)
Numerous
(Section 5.6)
Rahel 1983
Scheider et al. Leivestad et
1976, Gahnstrom al. 1976
et al. 1980, Hall
et al. 1980
Very acid-sensitive. Generally
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
Experiences reproductive failure near
pH 5.0. Generally absent below pH 5.0
in both naturally acidic and
anthropogenically acidifed waters.
Adversely affected by pHs below
5.0-5.5
Adversely affected by pHs below
5.0.
Lower pH limit between 4.5 to
5.0.
Lower pH limit between 4.2 to
5.0.
Lower pH limit near 4.5.
Lower pH limit 4.2 to 4.5. May
become very abundant after other
species have become extinct.
Bacterial decomposition is signifi-
significantly reduced in the pH
range 4.0 to 5.0. In many
cases, fungi replace bacteria as
the primary decomposers
-------
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 nonacidic 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
concentrations, 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 nonacidic lakes
within the 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 nonacidic 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 the LaCloche Mountain
range of Ontario, Nova Scotia, and southern Norway. 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.
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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
6.0) versus nonacidic waters, or through time with acidifica-
tion, typically indicate no change or increased growth 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
acidified aquatic habitats and include species declines, local
extinctions 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.
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5.10.2 Processes and Mechanisms by Which Acidification Alters Aquatic
Ecosystems
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 taxonomic 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
5.10.3.2, 5.10.3.4, 5.10.3.7) and biological interactions (Section
5.10.3.3).
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 nonacidified 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, bological 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 in fish
mercury concentrations include dissolved organic carbon, conductivity,
bioproductivity and watershed geology.
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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 been documented in Scandinavia, 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 zoopl ankton, can alter the rate of
primary production in lakes (Shapiro et al. 1975).
Changes in aquatic ecosystems following the loss of fish
populations are evident in nonacidified lakes where fish have been
intentionally removed (Stenson et al. 1978, Eriksson et al. 1980b,
Henrikson et al. 1980a,b). Large invertebrate predators (e.g.,
corixids, dytiscid beetles, Chaoburus) normally kept at low abundance by
fish predation become abunda"nT;Zoopl ankton 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—Water clarity typically increases with
increased acidity (Section 5.5.2.3.2 and Chapter E-4, Section 4.6.3.4).
This may be due to a reduction in algal biomass in the water column, the
precipitation 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 invertebrates. 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.
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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 1imnological 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. 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.
Alga 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.
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Contradictory responses should not be interpreted as evidence that
acidification has no effect, but rather as an indication that poorly
understood interactions among stresses may be involved. The infrequency
of manipulative, 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 re-
served. However, the long-term consequence of lake neutralization, pro-
vided 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, selective 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.
5.10.4 Summary
Biological effects due to acidification become apparent as
alkalinities decline to 65-35 yeq jr1 and pH's decline to between
6.5 to 6.0 (Table 5-16). Since the biological response to acidification
is a graded one, continuing pH declines below this range will result in
escalating biological changes. In Chapter E-4 it was concluded that,
under current rates of acidic deposition, a long-term pH of £ 4.9 can be
expected in low-alkalinity lakes and streams in the northeastern United
States and southeastern Canada that have pH levels in the mid 5's prior
to acidification (Section 4.4.4). Episodic depressions down to pH 4.3
to 4.9 will occur during periods of snowmelt and heavy rainfall and can
affect systems with a pH as high as 7.0 (Section 4.4.2). 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 4.9 or lower, 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.
5-164
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS
E-6. INDIRECT EFFECTS ON HEALTH
6.1 INTRODUCTION (T. W. Clarkson)
Indirect health effects 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 pheomenon 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
where its presence at elevated concentrations in water used in dialysis
therepy may be a cause of brain damage. Other elements and chemicals
will only be briefly mentioned as information is limited. These include
arsenic, asbestos, cadmium, copper, nickel, and the nitrosamines.
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 in contact with the contaminated water or consuming 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 considered 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, therefore, be limited to only a few toxic elements and
the major pathways of exposure.
6-1
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6.2.2 Availability and Bioaccumulation of Toxic Metals
Mercury and its compounds have been extensively studied in terms of
availability and bioaccumulation. The impetus for this work came from a
discovery made in the late I9601s (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 humal 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 interconversion of these three oxidations states
via the disproportionation reaction
Hg22+ Z 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).
6-2
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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 complexes are formed with chloride anions--HgCl + ,
HgCl2, 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 S2~, S-H and the sulfhydryl
anion in proteins and ami no 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), 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 remobilization 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." Examples of two organic forms of mercury are
depicted in Table 6-1. 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 fresh water 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 avidly
accumulated by 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
6-3
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TABLE 6-1. SOURCES OF MERCURY IN THE ENVIRONMENT 1971
(WHO 1976, NRIAGU 1979)
Source Amount
Metric tons yr-1
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
6-4
-------
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.
6.2.2.2 Concentrations and Speciations in Water (Mercury)--The early
findings of Stock and Cucuel (1934) that rain water 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 £~1 in Sweden, and Eriksson (1967) also in Sweden found most
samples of rain water 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
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 fresh water for which there is no known source of
contamination generally yield values less than 200 ng jT1 . Most
values fall in the range of 10 to 40 ng£-1 and drinking water
usually has values less than 30 ngji'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 precipitation is an important source of mercury to fresh water (see
next section), and second, that mercury in drinking water offers no
health threat. Concentrations on the order of a few hundred nanograrns
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 vg
Hg day-1 (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, NAS 1978, U.S. EPA 1980a) and
will be briefly summarized here. The subject is one of intensive
research, particularly 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 chanical species of mercury believed to be present at extremely low
concentrations in the atmosphere and in bodies of natural water.
6-5
-------
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 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,
oceans on the order of 2000 years, and 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, approximately 11.000 tons
yr-1, with land runoff accounting for about 5,000 tons yr~*.
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
fresh water 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 fresh waters
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 import fundamental discoveries
have been made in recent years and are summarized below.
The most important single discovery in understanding the chemical
and biochemical cycles of mercury in the environment was made by Swedish
investigators in the 1960s (for a review see NAS 1978, Nriagu 1979). An
intensive investigation into the source of the methyl mercury compound
6-6
-------
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
-------
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 612) that are produced as a result of bacterial synthesis.
However, other pathways both enzymatic and nonenzymatic 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 of various types of microbial 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 monmethyl
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 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,
occuring in water sediments and soils and in the gastrointestinal tract
of mammals, including humans. This biogeochemical cycle involving
bacterial methylation and demethyl ation 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.
6-8
-------
m
CH4 +
\
(CH,
J
FISH
i
1 +
^ CH~Hu — "^ (tH0
BACTERIA 3 3 BACTERIA ;
^^s^ '
°+ to.
Hn -f Ha
4H6- ;
5)2Hg
i
SHELLFISH
CH3SHgCH3
1
t
pHg CrUS-HgCHo
"+ + He
BACTERIA
\o
1 AIR
WATER
SEDIMENT
3
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
-------
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.
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 remitted 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
remission 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 of 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).
6-10
-------
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 considerable 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
distribution 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 (Nriagu 1979).
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, (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 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 statistically 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~l 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
6-11
-------
mercury in fish as noted above. Waters of low productivity
(oligotrophic lakes) and low alkalinity tend to be more sensitive 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 observation that low pH waters tend to be oligotrophic, soft
waters with low alkalinities. On the other hand, the association
between low alkalinity and elevated 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'M. 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.
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.
6-12
-------
CO
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METHYL MERCURY CONCENTRATION IN PIKE MUSCLE
Hg g-1 wet wt.)
-P>O
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T3 00
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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)
Change in water pH
one pH unit
two pH units
three pH units
Change in mercury concentration
mg Hg kg-1
0.14
0.28
0.42
Color increase
10 mg Pt
50 mg Pt £
100 mg Pt rj
-1
0.015
0.075
0.150
Change in specific conductivity
5 tnS nrl
10 mS nr1
20 mS m-1
0.075
0.150
0.300
6-14
-------
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
pg g~l wet weight,
Hgso = tne 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).
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 a question might be raised
on the appropriateness of linear regression analyses relating pH to
mercury concentration (e.g., in Figure 6-2 and the multilinear analysis
used for Table 6-3). The Hakanson formula has the general form of a
rectangular hyperbole:
FHg =
pH - 3
where Hg 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 (r2 = 0.81)
appreciably higher than that obtained by linear regression analysis
(r2 = 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
6-15
-------
200
~ 140-
cc
r>
<_>
a:
40-
4.5
PH
1. DUCK LAKE
2. LITTLE CLEAR LAKE
3. HARP LAKE
4. BIGWIND LAKE
5. NELSON LAKE
6. CHUB LAKE
7. CROSSON LAKE
8. DICKIE LAKE
9. LEONARD LAKE
10.
11.
12.
13.
14.
15.
16.
17.
HENEY LAKE
CRANBERRY LAKE
HEALEY LAKE
CLEAR LAKE
FAWN LAKE
BRANDY LAKE
MCKAY LAKE
LEECH LAKE
18. 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
-------
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,
Rutihis rutilis) adversely affected during acid episodes during spring
melt and a shift to predation on higher trophic levels (in this case
perch, Percas 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 significant (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.
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 (Stigostedion yitreum) of equal length, caught in Ontario
lakes with alkaline water (<_ 15 mg CaCOs &~1) had significantly
higher mercury levels than walleye caught in lakes with high alkalinity
(> 15 mg CaC03 £ )• Comparisons based on fish length may,
however, be somewhat misleading. If fish from waters with lower
alkalinity grow slower (possibly as a result of lower primary
prooductivity 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 fresh water fish caught in the United States. A
graph of mercury 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. From 1969 to 1972, over 3500 fish
from New York State lakes and streams were collected and analyzed for
mercury content. Two Adirondack Mountain reservoirs, Cranberry Lake and
Stillwater Reservoir, yielded fish with particularly high mercury
levels, despite the undeveloped nature of their watersheds (Blcornfield
et al. 1980) (Figure 6-4). The highest mercury levels in the New York
State survey were recorded for Onondaga Lake fish and caused, for the
most part, by mercury contributions from a chlor-alkali plant during the
1960s. Other areas, such as. Lake Ontario, Lake Champlain, and the St.
6-17
-------
250
2.25
2.00
^ 1.75
e-H
O>
o> 1.50
R 1.25
g 1.00
0.75
0.50
0.25
0
=20
LEGEND
.+ ADIRONDACK LAKES
-o— ST. LAWERENCE RIVER
-* ONONDAGA LAKE (No Fish>35 cm)
-• LAKE ONTARIO
-6 LAKE CHAMPLAIN
-A LAKE ERIE - NIAGRA RIVER
CRANBERRY LAKE and
STILLWATER RESERVOIR
20-25
25-30
30-35
35-40
LENGTH OF FISH (cm)
Figure 6-4. Average mercury concentration vs length class in New York
State smallmouth bass. Adapted from Bloomfield et al. (1980)
6-18
-------
Lawrence River also either have or have had in the past sources of
direct contamination, yet mercury levels found in fish from these areas
were lower than in fish from Stillwater Reservoir and Cranberry Lake.
Other Adirondack lake samples also tended to have lower fish mercury
levels. The source of mercury to Cranberry Lake and Stillwater
Reservoir remains unresolved. Stillwater Reservoir is an acidic (pH <
5.0) clear water lake. Cranberry Lake, on the other hand, had a mean pH
of 6.90 in the fall of 1978 (Bloomfield et al. 1980). Cranberry Lake
waters are, however, relatively highly colored with high concentrations
of humic organics.
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 frequently 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 (drainage 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 the pike and the 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-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 nonacidic 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 form most rapidly taken up by fish,
bioaccumulation presumably would be enhanced.
6-19
-------
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 nonacidic 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 6-5). Miller and Akagi (1979)
asserted that the effect of pH on the equilibrium 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
bioaccumulation 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 rl) 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 ta pike
increased with increased acidity. Harvey et al. (1982) determined
*
6-20
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150
%£> 100
OJ
El
1 50
LEGEND
H SAND
E3 SAND CHIP SEDIMENT
D WOOD CHIPS
MONOMETHYL MERCURY
•I
6
pH
Figure 6-5.
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).
6-21
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manganese concentrations in the vertebrae of 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 6-6). Lead (p < 0.01) and cadmium
(p < 0.05) were significantly correlated with lake pH level. No data
are available 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 yg Al g"1 wet weight of tissue on gills from dead ciscoe from
lakes Ransjon and Amten, Sweden, but only 6 yg Al g"1 for fish from
reference lakes without fish kills. Aluminum concentrations 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.
81oomfield 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 smallmouth 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.
6-22
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520
500
400
7 300
o>
en
c
.0
O.
200
100
J L
250
200
7 150
CO
CO
E
T3
O
100
50
0
co
CD
25
20
15
S 10
0
4.5 5.5 6.5 7.5 8.5
pH
25
20
^ 15
co
S 10
0
4.5 5.5 6.5 7.5 8.5
PH
Figure 6-6. Metal concentrations in yearling yellow perch and
epilimnetic pH in 1981 in lakes in the Muskoka-Haliburton
area of Ontario (U.S./Canada 1983).
6-23
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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 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
information on uptake, distribution, and excretion of methyl mercury In
man. Meth/l mercury is almost completely absorbed from the diet (90 to
100 percent, i.e., between 90 to 100 percent of the amount ingested is
absorbed). After absorption in the gastrointestinal tract, methyl
mercury passes into the bloodstream and is distributed to all organs in
the body. Approximately 5 percent of the absorbed dose goes to the
blood compartment and 10 percent to the brain—the target organ for
toxic effects.
After the initial distribution is completed, usually a matter of a
few days in man, the brain to blood concentration ratio is roughly
constant, having value between 5:1 to 10:1. Methyl mercury is
accumulated in growing hair. At the time of formation of the head hair,
the ratio of the concentration of mercury in hair to the simultaneous
concentration in blood is roughly constant and has an average value of
about 250:1. Once incorporated into the hair, the mercury concentration
remains constant. Because human head hair grows about 1 centimeter per
month, analyzing centimeter segments of hair can recapitulate average
monthly concentrations of methyl mercury in blood. Measurements of
mercury in samples of blood or hair are now routinely used to assess the
body burden of methyl mercury in humans and as an indicator of brain
concentrations.
Methyl mercury is excreted from the body mainly in feces. Before
excretion in the feces, methyl mercury is converted into inorganic
mercury. The site of this conversion is not known, but microflora in
the lower gut are known to possess this capability. The rate of
elimination from the body is directly proportional to the body burden.
It is well described by a single exponential function characterized by a
half-time of about 70 days. An important conclusion from this kinetic
information is that it will take about one year for humans to attain a
state of balance, i.e., to attain maximum steady body burden of methyl
mercury for any given daily intake in the diet. After cessation of
given exposure, it will take one year for the body burden to fall to
pre-exposure levels. Thus, dietary intake of methyl mercury from fish
should be evaluated over a matter of months. Intake on any one single
day does not normally make an important contribution to the overall body
burden.
6-24
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Considerable individual differences exist in biological half-times
in man although the average value is 70 days with a range of 30 to 180
days. The distribution is bimodal with 90 percent of the values
distributed about an average value of about 65 days and 10 percent
distributed about an average value of 120 days. The reasons for this
wide range of biological half-times are not known, except that lactating
women have a short half-time averaging about 40 days.
Methyl mercury readily crosses the placental 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 absorption of mercury. In assessing hazard to human
health, chronic exposure over weeks or months is important.
6.2.4.2 Toxicity in Man—Methyl 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.
6-25
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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 m -1 whole blood.
Corresponding hair concentrations 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 miximum blood level of 200 ng m -1 has been established to be
300 yg Hg. However, in the mother during pregnancy, the blood level
associated with the earliest damage to the fetus has not yet been
determined.
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 6-5 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 on 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 biological half-times, also obtained in the
Iraqi outbreak (Shahristani and Shihab 1974) to calculate the
relationship depicted in Figure 6-7.
Their calculations indicated that an intake of 50 yg day1 in
an adult gives a risk of about 0.3 percent of the symptom of
paresthesia, whereas an intake of 300 yg dayl would give a risk of
about 8 percent of symptoms of paresthesia. As pointed out by Nordberg
and Strangert (1976), the background 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 Iraq data the background frequency 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 epidemic!ogical
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
6-26
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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 mH) (yg g-1) (yg kg-1 body weight)
200 to 500 50 to 125 3 to 7
aThe 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-1, or hair levels between 50
to 125 yg g-1 would be expected to be affected (for further
details, see text).
6-27
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I/)
z;
o
o.
I/)
LU
a:
LU
a.
X
0.1
0.4 1.0 2
DAILY INTAKE (mg)
Figure 6-7.
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; Shahristani & Shihab 1974).
6-28
-------
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; therefore 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 below 200 ppb (Wheat!ey 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 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 concentrations of mercury in maternal
hair to risk of mental retardation in prenatally exposed infants (Figure
6-8). Berlin calculated a background frequency in the Iraqi children of
approximately 4 percent as compared to a background frequency of mental
retardation in Sweden of 2 percent. He also noted that in the case of
adults that 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
background 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 concentration of methyl mercury in the mothers' hair during
pregnancy. This association 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 g g~l.
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
6-29
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Q-
O
Q.
20-
300
500
LLJ
O
30
20
10
0
10
30
50
Figure 6-8.
MERCURY IN HAIR (ppm)
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 6A gives the complete dose-response curve and the
logit equation. Figure 6B gives the low frequency end of
the dose-response relationship, 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).
6-30
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estimation described in Figure 6-7 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.
(1979) 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 emerges. On this basis,
freshwater fishes dominate the rankings, with northern pike consumed at
17.4 g day"1, followed by freshwater trout at 12.3 g day"1, bass
(freshwater) and catfish at 12.1 g day"1. The highest user
consumptions of seafood are crabs and lobster at 10.6 g day"1, with
tuna down to 6.1 g day"1.
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 day"1.
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 approximately 25,000 individuals chosen to be representative
of the U.S. population. These data, along with additional information
on mercery concentration of 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 andfctrangert
(Figure 6-7) 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 6-8. 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 6-8
6-31
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TABLE 6-5. ESTIMATES OF AVERAGE AND MAXIMUM DAILY INTAKES OF
MERCURY BY THE "70 kg MAN" IN THE UNITED STATES POPULATION
(ADAPTED FROM U.S. EPA 1980a)
Media Mercury intake vg day-1 70 kg-1 Predominate
(average) form
Air 0.3 Hg°
Water 0.1 Hg2+
Food 3.0 CH3Hg+
6-32
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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. 1979)
Amount
Rank 1()6 lb yr-1
Total
Tuna (mainly
Canned)
Unclassified
(mainly
breaded,
Including fish
sticks)
Shrimp
Ocean Perchd
Fl ounder
Clams
Crabs/lobsters
Salmon
Oysters/scallops
Troutf
Codd
Bassf
Catfish^
Haddockd.
Pollock*1
Herring/smelt
Sardines
Pikef
Halibutd
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 day1)
18.7
6.1
10.0
8.3
9.7
8.€
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
0.07-0.14^
0.08
0.03
0.42
0.14
c
0.15
0.11
0.14
0.02
0.61
0.19-0.53
0.45-35g
c
c
aU.S. Chamber of Commerce (1978).
^Average values for skipjack, yellow fin, and white tuna, respectively.
cData not available.
dflainly imports.
CKing crab - all others, respectively.
fpresh Water.
9Red Snapper - other.
6-33
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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 freshwater 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 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 drinkinq
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 management 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.
6-34
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6.3.1.1 Direct Use of Precipitation (Cisterns)--The direct use of
precipitation 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 cisterns for the storage of
water, and 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 of the
world.
Roof catchments consist of an impervious surface, usually a house
or auxiliary 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. Since, in most systems, precipitation is used directly with
no treatment, the quality of precipitation and the amount 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 twofold. First, direct deposition of atmospheric
pollutants such as lead and copper may occur and, 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 (1983a) report that lead in atmospheric
deposition accumulates 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 (1983a) 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 are generally very low, indicating that some other
source such as corrosion of galvanized gutters and downspouts might be
present.
Young and Sharpe (1983a) 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
6-35
-------
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 impermeable 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.
Young and Sharpe (1983b) attempted to establish a relationship
between the corrosivity of precipitation and the concentration of copper
in the tapwater of cistern systems. Their basic premise was that
although atmospheric deposition is more so, and this increased
corrosivity has caused an increase in the concentration of copper in the
tapwater of the cistern systems studied. Table 6-7 contains a
comparison of pH, corrosivity (as indicated by the Ryznar Index) and
corresponding tapwater copper concentrations. The potential impact of
increasing the corrosivity of atmospheric deposition on tapwater copper
concentrations is obvious. If the relationship proposed by Young and
Sharpe is valid, increasingly polluted atmospheric deposition would
appear to be responsible for tapwater copper concentrations in excess of
drinking water standards. Unpolluted atmospheric deposition (pH 5.6)
would appear to result in tapwater copper concentrations within drinking
water standards for roof catchment cisterns.
The actual number of roof-catchment cisterns in those areas of the
country impacted by polluted atmospheric deposition is unknown. Kincaid
(1979) estimated that there were 67,000 cistern systems in the state of
Ohio, although there is good reason to believe that this figure may be
too high. 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 estimated costs of $500 to $1000 per household for water
treatment equipment and the necessary changes to plumbing systems
(Sharpe 1980).
Young and Sharpe (1983a) conclude that "The Presence of lead and
copper in the tapwater of cistern water supplies in western Pennsylvania
was sufficient 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 Mater Supplies--Very little work has been done on the
specific effects of atmospheric deposition on surface water supplies,
although quite a bit can be inferred from the surface water quality work
6-36
-------
TABLE 6-7. COMPARISON OF BEST AND WORST CASE AND MEAN RYZNAR
STABILITY INDEX (RI) OF BULK PRECIPITATION COLLECTED WEEKLY
DURING 1979, 1980 AND 1981 IN CLARION AND INDIANA COUNTIES, PA
(FROM YOUNG AMD SHARPE 1983b)
Parameter
PH
Alkalinity
(mg r1 CaCOs)
Best
case
5.29
2.22
Meana
3.87
0
Worst
case
3.40
0
Specific conductance
( mhos cnrl)
Calciun (mg £-1)
RI @ 20 C
Predicted Cu
cone, (yg jr1)
127
12.8
14.73
922
70
1.14
18.76
2724
220
0.23
19.58
3299
aBased on 138 samples collected over three years.
6-37
-------
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 aluminum concentrations in the stream directly above the water
supply intake increased from .05 mg jr1 to 0.70 mg £-1 in
response to a February rain and snowmelt event on the watershed. These
data are illustrated in Figure 6-9. High concentrations of aluminum
have been reported elsewhere by Cronan ahd Schofield (1979) and Herrman
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+ concen-
tration. As the H+ concentration increases so does the corrosivity of
the water.
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. The results of a review of the
drinking water standard for corrosivity, which was completed by U.S. EPA
in 1980, have not as yet been released. Numerous studies of surface
water chemistry have shown dramatic increases in the H+ concentration
of surface waters in response to acidification by atmospheric
deposition. In dilute surface waters such increases are almost certain
to produce corresponding increase in the corrosivity of that water. If
6-38
-------
CO
I
o?
o>
1.5
1.0
0.5
CARD MACHINE RUN
LEGEND
ALUMINUM
DISCHARGE
. j.^f^'L.j-nf. mmmpmm m^m in uu^-__-j_^_-J—A'
10 12 14 16 18
FEBRUARY
20
5 ^
4 5
i—i
3 ^
2 g
1 o
»—i
o
22
24
26
Figure 6-9. Aluminum concentration and discharge for Card Machine Run. Adapted from DeWalle et al
(1982).
-------
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 6-10) a strong relationship between the two
is identified. Linear regression techniqes were used to quantify the
relationship between pH and RI for this runoff event, and a correlation
coefficient of r = -1.00 was obtained. 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 6-11), 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^ = .80) for these two variables.
Although the data presented are limited, there would appear to be
strong indications that the corrosviity 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 the model for the relationship of pH and RI holds true for
all dilute surface waters, then increased corrosivity is likely anywhere
that the pH of such waters changes dramatically 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 Suppl 1_es—Acidification of groundwater as a
consequence of atmospheric (leposition has been reported in Sweden by
Hultberg and Wenblad (1980). Such changes have not as yet been well
documented in North America. Funs (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.
6-40
-------
0)
6.0
5.9
5.8
5.7
5.6
5.5
5.4
5.3
5.2
5.1
5.0
4.9
4.8
4.7
4.6
4.5
LEGEND
• pH
RYZNAR INDEX
17 18 19 20 21 22 23 24
FEBRUARY
25
26
27
28 29
18.9
18.8
18.7
18.6
18.5
18 A
18.3
18.2
18.1
18.0
17.9
17.8
17.7
17.6
X
LU
Q
CtL
o:
Figure 6-10. pH and Ryznar Index for Card Machine Run.
-------
Zfr-9
RYZNAR INDEX
N
3
CD
Q.
tn
x
O)
Q.
Q.
l/>
O
QJ
-s
05
O
-s
o
OJ
-s
a.
CD
DISCHARGE (£ s~l ha~l)
-------
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 6-12. 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 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 correlation 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 atmospheric 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 Waters—The U.S. national
interim primary drinking water standard for lead is 50 yg£-!.
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
s."1. In certain areas, such as Metropolitan Boston, it may contain
lead in excess of the 50 yg rl 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 State (U.S. EPA 1979a).
A recent national survey of Canadian drinking water supplies
involving 71 municipalities representing 55 percent of the population,
indicated a medial level of lead equal to or less than 1 yg £~
and values ranged from < 1 yg £-1 to 7 yg £~1.
Most natural ground waters have concentrations ranging from 1 to 10
yg r1.
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
6-43
-------
8.
8.0
7.5
en
i
7.0
6.5
6.0
21
18
15
ro
o
o
(O
T t
11
12
LEGEND
ALKALINITY (Spring)
pH (Spring)
DISCHARGE (Wildcat Run)
13
14
15
MARCH
16
17
18
(O
0.091
0.084
0.077
0.070
0.063
0.056 ~
0.049 ^
L±J
0.042 g
3C
0.035 S
Q
0.028
0.021
0.014
0.007
19
Figure 6-12. Alkalinity, pH, and discharge for Wildcat Run.
-------
pipes, the concentration may reach several hundred micrograms per liter
and even exceed 1000 yg r1 (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 tap water (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 Pb^+, 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 in recent years (WHO 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.
6-45
409-262 0-83-20
-------
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 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 accumulated 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 6-8).
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 6-8, the median lead level in U.S. drinking water (10
yg JT1) 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 6-13). The
inhibition of enzymes involved in synthesizing hemoglobin results in the
accumulation of precursor substances: s-aminolevulinic acid (s-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.
6-46
-------
TABLE 6-8. THE ESTIMATED RELATIONSHIP BETWEEN LEAD
CONCENTRATIONS IN RUNNING TAP MATER AND HUMAN
BLOOD LEAD LEVELS (MOORE ET AL. 1977 IN U.S. EPA 1980b)
Lead in running Total lead Lead in blood
tap water in blood (PbB) due to water
(ng A'1) (yg dl-1) (yg dl'1)
0 11 0
1 14.4 3.4
5 16.7 5.8
10 18.4 7.4
25 21.0 10.0
50 23.6 12.6
100 26.8 15.8
6-47
-------
ENZYMIC STEPS
INHIBITED
BY LEAD
NORMAL PATHWAYS
METABOLITES AND
ABNORMAL PRODUCTS ACCUMULATED
IN HUMAN LEAD POISONING
PROPHYRIN FORMATION
IRON UTILIZATION
1
2Pb
3
4
"7PK ..... .
MITOCHON
CYTOPLASM
~
o
3;
o
Pb
rKREBS CYCLE j Fe TRAf
__.. . J (C,EPUM
SUCCINYL CoA + GLYCINE ETICU,
ALAS 1
1 Fe
d-AMlNOLEVULINIC ACID (ALA) •
1 ALAD
PORPHOBILINOGCN
IURO I SYN
URO II COSYN
UROPORPHURINOGEN III
• UROGENASE
rflPDfiPfl[?nwYPTNn£FN T T T
COPROGENASE
PROTOPURRHYRIN S
HEMESYNTHETASE
«FERRIN
INTO
.OCYTES
t
+++
Pb
1 Pb
nrMr
^ ' . __ Pb
HEMOGLOBIN
Serum Fe
may be increased
ALA in urine (ALAU)
and serum increased
= 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 6-13.
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).
6-48
-------
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 6-14
shows the curvilinear 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
-------
FEP « 0.043 x (blood lead) +. 0.45(blood lead) - 2.14
r - 0.79
n * 1056
CO
10
o
o
o
o
O)
3.
1200
1080
960
840
720
600
480
360
240
120
0
0 15 30 45 60 75 90 105 120 135 150
BLOOD LEAD (yg 100 ml'1)
Figure 6-14. Free erythrocyte protoporphyrin (FEP) vs blood level
Shoshone County, Idaho, August 1974. Adapted from
Landrigan et al. (1976).
6-50
-------
TABLE 6-9. 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
Anemi aa
Peripheral neuropathy
Anemi aa
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.
6-51
-------
tQ
CT>
I
cn
PREVALENCE OF ABNORMAL VALUES (%.POPULATION)
I
on
ro
-s oo :> —• -h -H
fD CO CT fD -5 3"
00 .. 3 < fD fD
O 0) fD
fD
n
00 fD
fD
< CTl
fD CO <
• (Q C
fD
I—" 00
3=- O
0- O X
01 fD
T3 3 -S
r+ —i ft)
fD I
Q- I—O.
fD
-h -h -h
-S O -••
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3 fD
O Q.
70 3-
O -•• OJ
fD —' l/>
—' a.
(/> -s ~n
re m
fD 3 -a
3 ct- rt-
3" -••
3& -S O
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3" rh -"•
-i. fD -O
CT
fD
Q.-0
-S -S
fD O r+
3 rH «
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_i -i.-a
c+ 3 fD
-h Q> O
fD 3 fD
30-3
i O
3 c
fD fD
< —i O> 01
O> fD < CQ
—i 00 fD fD
^1 (a
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01 O
00 o> CQ -h
3 fD
Q^ Qj
cr cr
o —i 3
fD " O O
x o -s
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fD Q. O>
00
00
i fD — '
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-h 3
00 —'
ro ro
» 00
CQ
3-
o r
ro
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cn
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o
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a
(Q
o
o
ro
o
GO
o
-------
ami no aciduria, glycoseria, and hyperphosphoturia, usually do not occur
until blood levels exceed 70 yg dl'1. 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 arterioscelerotic 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 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 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 eposed 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 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
epidemiologial studies have failed to reveal an association between lead
exposure and human cancer. Measurement of precursor metabolite of heme
synthesis such as FEP or 6-ALA provide the earliest warning of the
effects of lead. If protected against, effect on heme synthesis will
protect against the more serious clinical effects of lead, such as
anemia and encephalopathy.
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 (WAS 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's 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
6-53
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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 sorne 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 MAS reported range of 100 to 3000 ml for children and a
U.S. median level of 10 yg £-1, the range of intake for children
would be 1 to 30 yg Pb and for adults 16 yg, assuming a water intake
of 1600 ml day-1 (Table 6-10). If average lead concentrations
attained the interim drinking water standard of 50 yg £-!, 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 proto-
porphyrin concentration of 50 to 250 yg 100 ml"1 of whole blood, the
child is thought to have undue lad absorption. Community 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 oT >_ 50 yg 100 ml"1 (Mahaffey et
al. 1982a). The review of human toxicity data in Section 6.3.2.4.2 also
indicate that blood lead levels in children >_ 30 yg 100 ml"1 indi-
cates 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 leads >_ 30 yg dl'1 (Table 6-11). The prevalence of elevated
blood lead values is 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
6-54
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TABLE 6-10. DAILY INTAKE OF LEAD FROM DRINKING WATER
Age Group Daily Water Intake3 Daily Lead Intakeb
ml yg Pb
Children 100 - 3000 1 - 30
Adults 1630 16
aNational Academy of Science (1980).
bAssumes U.S. median concentration of lead in drinking water to be 10
yg Pb A-l.
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TABLE 6-11. 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 1980a
Demographic variable
Estimated
population
(thousands)
No. of
persons
exami ned
Bloodb lead
yg 100 ml"1
Revalence of
blood lead
levels < 30
yg 100 ml"1
% persons
examined
Annual Family Incomec
< $6000 2465 448 20 +_ 0.6 10.9 +_ 1.4
$1000 - 14,999 7534 1083 16+0.5 4.2+0.7
> 15,000 6428 774 14 + 0.4 1.2 +~ 0.4
Degree of Urbanization
urban
urban
Rural
>
<
106
106
persons
persons
4344
6891
5627
544
944
884
18
16
14
+ 0
+ 0
TO
.5
.7
.6
7.2
3.5
2.1
+
T
-
0.7
0.6
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.
6-56
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concentrations in tap water (having stood overnight) in excess of the
drinking water standard of 50 yg Pb fc'1.
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 with children where
substantial numbers already have elevated blood levels. Drinking water
at the median concentration of 10 yg Pb &"1 already makes an appreciable
contribution to blood lead levels (approximately 30 percent add on to
other sources of lead; see Section 6.3.2.4.1). Thus the drinking water
standard of 50 yg Pb fc-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 USA subject to acidic
deposition and to check the extent to which charged corrosivity of this
water affects lead levels in tap water.
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 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 dissolved and suspended aluminum in water. He notes
that many investigators now use a 0.45 ym millipore filter to
distinguish between dissolved and parti oil ate 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
6-57
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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 r~l for the period
1971-73. A phosphate limestone lake in Florida had a mean value of 0.05
mg 5,"1 at a water pH 7.0 to 9.6. Tributaries to Lake Michigan had
mean values of 0.353 in 1972 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 Trojner (1980)
report that in a brook in the Adirondack Wilderness region of
New York State, aluminum concentrations rose from about 0.2 mg
r1 at pH 5.5 to 6.5 to 0.8 to 1.0 mg £-1 as the pH
fell to less than 5.0 during the spring snow melt.
2) Highly saline waters contain higher aluminum concentrations
than fresh waters.
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 A13+ 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).
6-58
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6.3.3.4.1 Dynamics of aluminum in humans. Data on absorption,
distribution, and excretion of aluminum compounds in 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.
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 change contamination 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 man. 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 accumulation 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 administered 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
6-59
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who undergo regular therapeutic dialysis. Driscoll et al. (1980) have
reported levels of aluminum in natural bodies of freshwater in the
Adirondack Region of New York State to attain values as high as 800 yg
Al &"1 under the influence of acidic deposition. A concentration of
50 yg A"! 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 tap water 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
heath 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 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 direct health effects and in doing so
mainly discusses lead and mercury availabilities as affected by acidic
deposition. The following statements summarize the content of this
chapter.
0 No adverse human health effects have been documented as being a
consequence 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).
o 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).
0 In virtually all studies published to date elevated methyl
mercury levels in fish muscle (most notably the pike and perch)
have been statistically associated with higher levels of
acidity in water. However, a number of factors influencing
6-60
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mercury levels in fish may also change in 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).
8 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 in acidic deposition of methyl mercury
concentrations 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.
0 Prenatal life is a more sensitive stage of the life cycle for
methyl mercury. More information is needed on fish consumation
patterns of women of child-bearing age in order to quantita-
tively 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.
0 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).
0 Increasing corrosivity is probably the most significant
potential impact of acidic deposition on surface water
supplies. Populations are at increased risk of being exposed
to higher concentrations of corrosive toxicants, such as lead
and possibly cadmium, where surface water storage facilities
6-61
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are small, necessitating the direct use of raw water during
storm flow 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 are used that have no particulate filters {Section
6.3.2).
o From the point of view of human health risks, any increases in
lead concentrations in drinking water should be viewed as an
additional burden of lead. This is especially important in
children where substantial numbers already have elevated blood
lead levels (Section 6.3.2.4).
° 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.
6-62
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THE ACIDIC DEPOSITION PHENOMENON AND ITS EFFECTS
E-7. EFFECTS ON MATERIALS
(J. E. Yocom and N. S. Baer)
7.1 INTRODUCTION
In the popular press many articles ascribe damaging effects to
acidic deposition (LaBastille 1981). Damage to non-living materials and
structures is invariably listed as one of the important effects of this
phenomenon. Furthermore, 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
anthropogenic air pollutants.
In general, the distinction between the effects on materials of
near or intermediate 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.
Damage to materials from acidic deposition takes a variety of forms
including the corrosion of metals, erosion and discoloration of paints,
decay of building stone, and the weakening and fading of textiles. All
of these effects occur to a significant degree as a result of natural
environmental conditions, even in unpolluted atmospheres. Moisture,
atmospheric oxygen, carbon dioxide, sunlight, temperature fluctuations,
and l'™ 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
Chapter A-9 of this document the following definitions for the
scales of acidic deposition transport are given: short range
(< 100 km), intermediate range (100 to 500 km), and long range
(> 500 km).
7-1
-------
pollutant transformation and contact processes (for example, acid
precipitation) becomes even more elusive.
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 N02
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 primary gases. While this chapter's
scope is extremely broad in concept, the literature describing research
on any one specific contact-and-effect scenario may be limited or even
non-existent.
There is a significant body of literature describing the effects of
primary air pollutants on materials as determined by both laboratory and
field experiments. This literature has been summarized in detail by the
U.S. Environmental Protection Agency in its Criteria Documents
supporting the establishment of air quality standards, for example, the
document on sulfur oxides and particulate matter (U.S. EPA 1981).
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 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 broadly be described as associated with acidic deposition.
Table 7-1 summarizes the potential damaging effects of air pollutants
and other environmental 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.
7.1.1 Long Range vs Local Air Pollution
Acidic pollutants whether they are present as primary pollutant
gases (e.g., S02 and NOX)» as fully oxidized acids or salts (e.g.,
sul fates 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 since phenomena associated with
long-range transport may be generated by local sources under the
appropriate conditions. For example, acidic precipitation may be
produced close to sources of primary pollutants under the proper
meteorological conditions. The distinction between different acidic
7-2
-------
TABLE 7-1. AIR POLLUTION DAMAGE TO MATERIALS
Type of Principal air Other
Materials impact pollutants environmental
factors
Methods of measurement
Mitigation measures
i
CjO
Metal s
Building
Stone
Ceramics
and Glass
Paints and
Organic
Coatings
Paper
Photo-
graphic
Materials
Textiles
Textile
Dyes
Leather
Rubber
Corrosion,
tarnishing
Surface erosion,
soiling, black
crust formation
Surface erosion,
surface crust
formation
Surface erosion
discoloration,
soiling
Embrittlement,
discoloration
[•11 c rob! emi she s
Reduced tensile
strength,
soiling
Fading, color
change
Weakening,
powdered surface
Cracking
Sulfur oxides
and other acid
gases
Sulfur oxides
and other acid
gases
Acid gases,
especially
fluoride-
containing
Sulfur oxides,
hydrogen
sulfide
Sulfur oxides
Sulfur oxides
Sulfur and
nitrogen
oxides
Nitrogen
oxides
Sulfur oxides
Moisture, air,
salt, parti cul ate
matter
Mechanical ero-
sion, parti cul ate
matter, moisture,
temperature
fluctuations,
salt, vibration,
COg, micro-
organisms
Mol sture
Moisture, sun-
light, ozone,
parti cul ate
matter, mechan
ical erosion,
microorganisms
Moisture, phys-
ical wear,
acidic materi-
al s introduced
in manufacture
Parti cul ate
matter,
moisture
Parti cul ate
matter,
moisture,
light, physical
wear, washing
Ozone, light,
temperature
Physical wear,
residual acids
introduced in
manufacture
Ozone, sun-
light, physical
wear
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 1n elasticity and
strength, measurement of crack
frequency and depth
Surface platinq 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
controlled atmosphere
deacidi fication, encapsula-
tion, impregnation with
organic polymers.
Removal to controlled
atmosphere
Replacement, use of substi-
tute materials, impregnation
with polymers
Replacements, use of
suhsti tute material s, removal
to controlled environment.
Removal to a controlled
atmosphere, consolidated with
polymers, or replacement
Add antioxidants to
formulation, replace with more
resistant materials
-------
TABLE 7-2. CHARACTERISTICS OF LONG-RANGE AND LOCAL AIR POLLUTION
Pollutant
or Effect
Long-range
Local
Pollutant Concen- Low concentrations and uniform
tration Patterns distribution.
Sulfur Oxides
Nitrogen Oxides
Participate Matter
(includes
aerosols)
Ozone and Other
Oxidants
Dry Acidic
Deposition
Acidic
Precipitation
Acidic Fog
(includes 1 iquid
aerosol s)
S02 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
sul fates.
Ozone and other oxidants are
produced from hydrocarbons and
NOX over moderate to long-range
transport in presence of
sun!ight.
Dry deposition of acidic
particles (for example, sul fates)
Is possible.
Acidic rain nechanisns may be
predominantly through droplet
condensation around acidic
particles.
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 S02;
however, under light winds and
stable atmospheric conditions
conversion to particulate sul fate
can occur.
Exist primarily as MO and M02,
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
-------
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.
7.1.2 Inaccurate Claims of "Acid Rain" Damage to Materials
The popular literature contains frequent references to "acid rain"
damage to cultural property. In most cases no attempt is made to
distinguish between local 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 Meedle," 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 attributed to advanced salt decay, high
humidity of the New York climate, and unfortunate attempts at
preservation (New York Times 1978b, Winkler 1980).
7.1.3 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 in the next section, 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 $03 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 is absorbed in moisture
(condensed 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 directly. An example would be acid-containing soot from
an oil-fired boiler.
7-5
-------
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 S02
contact, that some surface oxidation may take place within a relatively
dry adsorbed layer or that S02 may react directly with a reactive
surface to produce a sulfite salt. Nevertheless, 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
than in 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 deposition mechanisms are important because of the link between
pollutant levels and meteprological 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, Scenario 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 surface; 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.
7-6
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7.1.4 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 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 environmental
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 pollutants 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 surface 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
principle difficulties with this approach are
Materials exposed may not represent materials in actual use.
0 In normal use materials are found in combination. Field
studies may not include interactions of other materials in
contact with test materials.
0 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).
7.1.5 Measurement of Materials Damage
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 or reflectivity, chemical analysis,
and identification of corrosion products).
7.1.5.1 Metals—Corrosion of metals may be measured by weight changes
resulting from the accumulated corrosion products before and after a
predetermined exposure period. However, for long exposures, corrosion
products tend to spall or wear off. Thus corrosion products formed
during the exposure period are usually removed chemically to determine
7-7
-------
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 resistance. Mechanical tests involving
bending are frequently used to test for 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 nature of corrosion products.
7.1.5.2 Coatings—Paint and other coatings damaged by environmental
exposure usually erode, so measurement is most conveniently done by
measuring weight loss of painted panels. Surface darkening by deposits
of particulate matter or reactions between pigments and air pollutants
are usually measured photometrically.
7.1.5.3 Masonry—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. Discoloration 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.
7.1.5.4 Paper and Leather—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. Weakening of leather caused
by acidic deposition can be quantified by means of tensile strength
tests.
7.1.5.5 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 molecular weight of the exposed material.
Damage from acidic chemical attack 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.
7.2 MECHANISMS OF ACIDIC DEPOSITION
Acidic deposition damages a wide range of materials. This section
will cover some of the principal damage mechanisms for selected classes
of materials that are widely used and economically important.
7-8
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7.2.1 Metals
The atmospheric corrosion of metals is generally an electrochemical
process governed by diffusion of moisture, oxygen, and acidic pollutants
(e.g., S02) to the surface. The EPA Draft Criteria Document for
Sulfur Oxides and Particulate Matter (U.S. EPA 1981) provides a review
of the primary mechanisms governing the corrosion of metals in the
presence of S02 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.
The rusting of metals 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 $03 to sulfate
on iron surfaces:
S02 + 02 + 2e~ + S042- [7-1]
or
4 HS03" + 3 02 + 4e- •* 4 S042' + 2 H20. [7-2]
The electrons are provide by the oxidation of the metal (M):
M -> Mn+ + ne- [7-3]
Barton (1976) noted that rusting of iron occurs first at isolated
sites and then spreads across the entire surface. This phenomenon is
not well understood but may relate to a variety of factors including
differential deposition rates of S02 or acidic partial!ate matter, the
influence of rust deposits on subsequent corrosion, and variations in
"time-of-wetness" in relation to electrolyte concentrations at various
points on the surface. Rice et al. (1982) believe that moisture forms
in "clusters" on metal surfaces even in indoor environments and at the
site of these clusters, corrosion is initiated. While rust deposits
increase the absorption of S02, 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 non-ferrous metals (zinc, copper, and especially aluminum) tend
to suppress the absorption of S02.
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 1981) 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.
7-9
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7.2.2 Stone
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 S02 levels and resulting materials damage.
The general decay mechanism includes aerodynamic factors
controlling delivery of S02 to the stone surface, oxidation of S02
to sulfate and the subsequent 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 compounds, a comprehensive decay mechanism must include the roles
of nitrogen compounds, carbon dioxide, and water. For the carbonate
stone/sulfur compound system three general modes of attack pertain:
Gaseous SO?
S02 + CaCOa + CaSOa + C02 (Step 1) [7-4]
CaS03 + 1/202 ->CaS04 (Step 2) [7-5]
Wet Deposition
C02 [7-6]
Dry Deposition is exemplified by the reaction between sul fates in
particul ate matter and calcium carbonate either in the form of sulfuric
acid as in wet deposition, or in the form of ammonium sul fates (Stevens
et al 1980).
+ CaCOs -> CaSCty + (NH4)2C03 [7-7]
NH4HS04 + CaCOs -> CaSCH + NfyHCOa [7-8]
The anhydrous CaS04 is hydrated to form gypsum, which is highly
susceptible to surface erosion.
Humidity plays a key role in all aspects of the interactions of
S0v with carbonaceous stone. In autoradiographic experiments using
sulfur-35, Spedding (1969b) showed surface saturation of oolitic
limestone samples by SO? 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 factors contributing
to stone decay as shown in Table 7-3.
7-10
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TABLE 7-3. CLASSIFICATION OF MECHANISMS CONTRIBUTING TO STONE DECAY3
(ADAPTED FROM TOMBACH 1982)
Mechani sm
Temper-
Rainfall Fog Humidity ature
Solar
Insolation Wind
Gaseous
Pollutants Aerosol
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 SOg 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
circles denote principal atmospheric factors; open circles denote secondary factors.
-------
7.2.3 Glass
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 reaction with sulfur dioxide in the range 20 to 100
C in gas saturated with SOg involves the same process at approximately
the same rate as with water alone (Douglas and Isard 1949).
7.2.4 Concrete
Cement, concrete, and steel reinforced concrete structures are all
subject to complex actions reducing their durability. 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 (Browne 1981,
Volkwein and Springenschmid 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 02 through the concrete to the steel. The reaction of
SOo and 50^2- yith cement involves the formation of calcium
sulfate and calcium sulfate aluminum hydrate (ettringite).
7.2.5 Organic Materials
Most organic materials exposed boldly 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. Painted
surfaces are the most widely employed organic surfaces exposed to the
atmosphere and to some degree are susceptible to acidic deposition.
However, these effects are relatively minor when compared with natural
environmental factors such as sun and precipitation (Haynie et al 1976).
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
permeability of the surface. It has been shown that the presence of
extremely high concentrations of $03, a reducing gas, can interfere
with the drying process, which is an oxidation-polymerization reaction
(Hoibrow 1962). However, it is doubtful that S02 concentrations at
the present time 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)
7-12
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and extenders such as CaCCh. The long-term effect is the loss of
paint surface through erosion.
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).
7.2.6 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 , [7-9]
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 7-4 presents the deposition velocities for sulfur dioxide.
(More extensive discussion of deposition processes can be found in
Chapter A-7.)
7.3 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
that are judged to be economically and culturally most important. These
material classes are:
° metals
o masonry
° paint and other coatings
o cultural property (historically and culturally valuable
structures and objects)
* other materials (paper, photographic materials, textiles, and
leather)
7.3.1 Metals
The position of metals in the electromotive series determines their
relative reactivity. However, the solubility of the particular metal
7-13
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TABLE 7-4. MEASURED DEPOSITION VELOCITIES FOR SOo ON VARIOUS SURFACES
(COMPILED BY JUDEIKIS 1979)
Surface^ Vg (m m1n~l)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
> TJ.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
^Number in parentheses indicates the number of different
materials examined if greater than one.
bAs defined by Equation 7-9.
7-14
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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 atmospheres containing sea salt
or large, acidic particles.
Thermodynamic considerations governing electrochemical corrosion
are conveniently examined with the help of Pourbaix potential-pH
diagrams. By plotting electrical potential against solution pH, regions
of stability for various chemical species can be indicated. In
simplified form, when reactions to 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 subsequent attack,
the region is one of "passivation" or mitigation of corrosion. 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.
In using these diagrams to determine the effect of lowered pH on
corrosion one must determine the potential attained by the metal in the
natural environment. 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 may form which may be
considerably more soluble. This potentially synergistic problem is
sometimes overlooked in traditional writings on corrosion. The Pourbaix
diagram can give much insight into this process. However, caution must
be exercised in interpreting these diagrams because kinetic factors with
non-equilibrium behavior may govern corrosion.
7.3.1.1 Ferrous Metals--Corrosion of iron and steel in polluted
atmospheres has received a great deal of attention over the years.
Steel, unless it is an alloy designed for unprotected exposure is
usually coated by painting 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 containing 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 which he associated
with the deposition of particulate matter.
As pointed out in Section 7.2.1, many theories have been advanced
on the principle chemical reactions that describe iron and steel
7-15
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14
0 7 14
0 7 14
GOLD
SILVER
COPPER
LEAD
IRON
0 7 K
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 0? evolution line while the lower ine is that for H2
evolution. Adapted from Pourbaix (1966).
7-16
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LU UJ
o t/i
CL.
UJ
O >
O
o; >
H-*—•
O
LU
Figure 7-2. Pourbaix diagram for the system Fe, Fe2+f Fe3+,
Fe3d4, and F6203- The thin diagonal lines indicate
regions of water stability. Compare with Figure 7-1 for
designated regions of "corrosion," "immunity, and
"passivation." The reactions considered are:
1) Fe = Fe2+ + 2e-
2) 3Fe + 4H20 =
3) 3Fe2+ + 4H20 =
4) 2Fe2+ + 3H20 =
r- \ i- O± _ Ox
5) Fe^+ = Fe6 +
6) 2Fe3+ + 3H20 =
7) 2Fe304 + H20 = 3Fe203
8H+ + 8e-
+ 8H+ + 2e~
+ 6H+ + 2e-
+ 6H+
2H+ + 2e-
7-17
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corrosion In the presence of SO? and moisture. Various chemical
routes are possible and it is likely that several of them operate in
actual atmospheric exposures.
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 S02. 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 constituent of the corrosion products of steel. Mild
steel samples were exposed to two urban areas near Manchester. 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 products showed them to be
predominantly Y -Pe2Q3'^2Q» a -Fe203'H20 and a -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.
7.3.1.1.1 Laboratory studies. Exposing iron and steel samples to S02
and moisture under controlled laboratory conditions has two prinicipal
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 concentrations many times higher than those found in the ambient
atmosphere, or what are referred to as "reflux" conditions, where water
and excess S02 were continuously flushing the surface of the samples.
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 encompassing those found in ambient urban
atmospheres. Sunlight and the formation of dew were also simulated.
7-18
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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 relationship between steel
corrosion and S02 concentration, together with high humidity.
7.3.1.1.2 Field studies. A problem inherent 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 occurrance 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-(163.2/RH)] [7-10]
where
cor = depth of corrosion, ym,
t = time, years,
S02 = S02 concentration, yg nr3,
RH = average annual relative humidity, percent.
Figure 7-3, based on the above damage functions, shows the
relationship between pseudocorrosion rate (cor /t~l), relative
humidity, and S02 concentrations. This graph shows that the corrosion
rate is much more sensitive to humidity than to S02, especially at
levels of S02 normally experienced in urban areas.
For example, referring to Figure 7-3, if one were comparing
relative corrosion at 55 percent relative humidity in two areas with
average levels of 100 and 150 yg nr3, a very significant difference
in relative air pollutant levels, the difference in relative corrosion
7-19
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oc
z
o
o
cei
C£.
O
o
o
o
Z3
LU
I/O
D_
100
90
80
70
60
50
40
30
20
10
0
100
200
55% RH
300
400
AVERAGE S02 CONCENTRATION, jjg 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).
7-20
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would be approximately three pseudocorrosion units. On the other hand,
if one were comparing relative corrosion at a constant $03 level of
100 yg m~3 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 experienced 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 which have been summarized by EPA (1981) and Haagenrud et
al. (1982). It should be noted that nearly all metal corrosion damage
functions have been developed by regression analysis and 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 measure the length of time that a corrosion panel was wet
enough for electrochemical corrosion to take place (time-of-wetness).
His sample exposure array included weathering steel, galvanized steel,
house paint, and Georgia marble. Concentrations of $62 measured in
this study were an order of magnitude lower than those measured in
Upham1s 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 concentration. 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 approached 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.
Measurement of the effects of pollutants associated with long-range
transport (e.g., acid precipitation) as compared with locally generated
7-21
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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 year's 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
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 concentration also affected monthly
corrosion rates, but had little effect on long-term values because they
occurred so seldom.
7.3.1.2 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 $03,
confirming the lower sensitivity 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 conditions. 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 7.2.1. Further-
more, 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 non-ferrous
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 cm'2 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 NOX impacted areas (Harrison 1975).
Such damage to components in communications switch gear is an
insidious problem because a simple malfunction can put a large system
out of service.
7.3.1.2.1 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 $03 and
moisture. In view of the reductions in emissions of S02 and
particulate matter—especially larger particles or agglomerates that
7-22
-------
CM
3 -
u
CD
CVI
o
CO
o;
o
to
Q
1 —
I 1 I I
2 —
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).
7-23
-------
could act as centers for corrosion initiation--S02 related acidic
deposition and surface corrosion of aluminum does not appear to be a
significant problem (Fink et al. 1971).
7.3.1.2.2 Copper. Copper and copper alloys in most atmospheres develop
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 S0£ over many years. Nevertheless, both the sul fate
and chloride-based patinas are generally resistant to further attack
(Yocom and Upham 1977).
7.3.1.2.3 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
eletrolyte, the current flow protects the steel from corrosion at the
expense of some oxidation 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 an empirical
damage function for zinc corrosion in relation to SOg concentrations
and time-of-wetness. Time-of-wetness was measured by means of a dew
detector. S0j> was measured by lead peroxide sulfation candles and
conductions trie S02 measurements. His empirical damage function is
Y = 0.005461(A)0.8152x (B + 0.02889), [7_H]
where
Y = corrosion loss, mg for a 3 x 5 in. 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 where zinc panels were exposed, while concurrently
collecting data on SO?, temperature, and humidity. They developed the
following empirical damage function relating zinc corrosion to S02
levels and relative humidity:
y = 0.001028 (RH - 48.8) S02, [7-12]
7-24
-------
where
y = corrosion rate, m yr~l,
RH = average annual relative humidity.
S02 = average S02 concentration, yg m~3.
Note that in Equation 7-11 moisture is in terms of time of wetness
while in Equation 7-12 annual average relative humidity is used. Time of
wetness is a far more relevant indication of surface moisture than
average relative humidity when corrosion and other forms of
moisture-enhanced material damage are being considered. For example, if
Equation 7-12 is applied in an area that has annual average relative
humidity significantly less than 48.8 percent, 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.
The damage coefficients for these two functions plus functions
developed from other studies were compiled by EPA (1981). These
coefficients are compared in Table 7-5.
7.3.2 Masonry
The term "masonry" is applied to a large number of building and
decorative materials exhibiting a broad range of surface activities 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 concrete, 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 and then, the salts released by these reactions
diffuse into the brick, causing stress and subsequent spalling.
7.3.2.1 Stone--The accelerated decay of stone buildings and monuments
in highly industrialized areas has been documented by comparison of
current condition 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
in the past 60 years with 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 from the west frieze of the Parthenon (Plenderleith and Werner
1971). A detailed account of the restorations on the Acropolis and
measurements of the thickness of gypsum layers formed on its exposed
marble surfaces is presented by Skoulikidis (1982). The deteriorating
condition 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).
7-25
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TABLE 7-5. 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 coefficient3
(ym yr"1) (ym yr'Vyg m~l
Number
of
data
sets
Field Studies
CAMP (Haynie and Upham
1970)
ISP (Cavender et al.
1971)
Guttman 1968
Guttman and Sereda 1968
St. Louis (Mansfield
1980)
1.15 +_ 0.60
1.05 +_ 0.96
1.79
2.47 4-0.86
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 yg m'3 S02.
7-26
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Few quantative studies of air pollution damage to stone have been
reported although the Increased rate of erosion for marble tombstones in
the urban environment of Edinburgh was observed as early as 1880 (Geike
1880). A study of tombstones in U.S. National Cemeteries (Baer and
Berman 1981) has developed methodology for measuring damage to marble
headstones exposed to the environment for 1 to 100 years. Their data
base 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 below 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 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 S02 immission rate were obtained:
Baumberg sandstone U = 0.54 D; r2 = 0.92 [7-13]
Krehnsheim 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 = SOo immission rate (uptake rate) of the stone in (mg
m~2 d~l) by weight gain,
D = by weight gain S0£ immission rate, IRMA measured value (mg
itr2 d-1),
L = loss in weight, and
r = correlation coefficient.
The high contribution of the non-SOx factors for stones exposed
to rainfall suggests that damage functions for stone must specifically
address such variables as other pollutants and rainfall.
7-27
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TABLE 7-6. DETERIORATION OF SCHLAITDORF SANDSTONE EXPOSED FOR
100 YEARS IN WEST GERMANY (AFTER SCHREIBER 1982)
Monument
Location
Relative S02
immission Rate,3
mg m~2-day
Deterioration
Neuschwan stein
Castle
Ulm Cathedral
Cologne Cathedral
Fussen
Ulm
Cologne
6
48
111
Practically
Moderate
Very severe
none
Relative immission or uptake rate of $03, annual average (August
1973 - July 1974) measured by IRMA method. (See Baer et al. 1983 for
details of the technique.)
7-28
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A series of measurements made at St. Paul's Cathedral, London on
the Portland stone (biosparite limestone) balustrade, demonstrate a high
rate of weathering (Sharp, et al. 1982). Using lead plugs filled in
1718 in openings in the stone 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 micrometer 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 represented 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.
7.3.2.2 Ceramics and 61ass--Although enamels and glasses are quite
resistant 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).
Flourides, especially HF, are capable of attack on a wide variety
of ceramic materials and glasses. Restrictive legislation on flouride
emissions has, for the most part, eliminated fluoride induced damage.
Perhaps the most serious glass damage problems 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 below.
Properly fired brick is highly resistant to attack by air
pollutants while poorly fired brick is highly susceptible to chemical
attack. Acidic solutions 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).
7.3.2.3 Concrete—World production of concrete amounts to some 3
billion cubic meters per year. Many important structures, e.g., bridge
decjJk highways, military installations, and naval shore structures
sufWr from severe durability problems (NMAB 1980). Similarly, concern
has been expressed over leaching of possibly toxic components of cement
culverts transporting acidified water.
The highly alkaline nature of cement/concrete leaves such surfaces
vulnerable to acidic deposition. Spedding (1969b), reporting on the
contamination/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 deposition of S02 on concrete are water soluble. The
high volume of water flow through rain collecting and distribution
7-29
409-262 0-83-22
-------
culverts in drinking water systems raises questions about the possible
release of toxic materials leached from the concrete matrix.
Similar concerns have been expressed over the errosive effects of
acidified streams on concrete bridge piers. The literature reveals only
limited research on the effects of acidified water runoff on concrete
durability.
Cements used in dams and culverts require a special formulation for
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 resistance
to acid and salt attack and damage to freeze-thaw cycling (Sulphur
Institute 1979).
7.3.3 Paint
Paint consists of pigment and vehicle. Pigments, such as titanium
dioxide and zinc oxide, provide color, hiding power, and durability.
Sometimes fillers 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 staining. Chemically active large particles such as acid
smut (or soot) from oil-fired boilers, mortar dust near building
demolition sites, or iron particles 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 time for
various oil-based paints exposed to extremely high concentrations of
S02 (1 to 2 ppm) was 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.
7-30
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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 $63
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 house paint to S02 was attributed to the
use of extenders such as CaC03 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 sulfur dioxide 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 increases, 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 appeared to contradict the chamber
studies. However, the authors believed that differences in the
pollutant mix at the two locations and especially the presence of
7-31
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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
oil
latex
Coil coating
Automotive refinish
Industrial maintenance
Mean Erosion Rate (nm hr~* with 95
confidence limits) for unshaded
Clean air S02
control (1.0 ppn)
5.11 +_ 1.8 35.8 +_4.83a
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
panels
03
(1.0 ppn)
11.35 +_ 2.67a
2.16 jf 1.50°
3.78 +_ 0.64b
1.30 +_ 0.333
7.14 jv 3.56
PAINT EROSION RATES AND PROBABILITY DATA (T-TEST)
FOR FIELD EXPOSURES (ADAPTED FROM CAMPBELL ET AL. 1974)
Mean Erosion Rate (nm hr'l 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
+
+
+
+
+
191
13
20
28
41
Suburban
376+^
76 +_
254 +_
58 +_
208 +_
1243
183
483
18b
361b
Urban
(S02 dominant),
- 60 yg m"J
361
97
241
41
168
+_ 124b
±8»
+_ 203
± 10
+_ 99
Urban
(oxidant dominant),
- 40 wi m"3
533 +_
165 jf
223 +_
43 +_
198 +_
157a
142
433
10
613
Significantly different from control at p = 0.01.
^Significantly different from control at p = 0.05.
Note: 1 ppn S02 = 2620 yg m'3
7-32
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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 laboratory sulfur dioxide exposures. These coatings were the only
ones that contained a calcium carbonate extender—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 S02 or moisture.
A multiple regression relationship was developed for the joint
influence of S02 and relative humidity on the oil-based house paint:
E = 14.3 + 0.0151 S02 + 0.388 RH [7-17)
Where
E = erosion rate of pm yr-1,
S02 = concentration of S02 in v9 n>~3»
RH = means annual relative humidity in percent.
This relationship indicates that paint erosion is significantly
more sensitive to changes in humidity than to S02- 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 thickness 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. Since the field studies were carried out in the early 197'0s,
further laboratory 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 than those used earlier.
7-33
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7.3.4 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 EPA's Criteria Assessment Documents, so they
are considered here although the indoor locations in which they are
normally found dictate gaseous transport mechanisms for deposition.
Paper—The role of sulfur dioxide in the deterioration of paper
accepted since the 1930's. Early experiments (Langwell 1952,
7.3.4.1
has been
1953) relied on unrealistically 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 ppn,
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 century books in two English 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 S02 sorption. Spedding and Rowlands (1970) measured the
sorption characteristics 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.
7.3.4.2 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 sul fide, 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 of the 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 sensitive 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).
7-34
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7.3.4.3 Textiles and Textile Dyes—Sulfur oxides are capable of causing
deterioration to natural and synthetic fibers. Cotton, like paper, a
cellulosic fiber, is weakened by sulfur dioxide. 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 pprc S02. Loss in strength was 13 percent in the absence of S02
and 22 percent in the presence of S02- In a study of nylon fabrics
exposed to 0.2 ppm Sfy 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 tensile properties both reflect this damage
(Zeronian et al. 1973). Results demonstrated that the mode of
degradation is not changed although S02 accelerates the rate of
reaction.
Among proteinecous 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 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 when compared to exposure 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 appreciable 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.
7.3.4.4 Leather—Michael Farady is attributed (Parker 1955) with having
established 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)
7-35
-------
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 exposure to
an atmosphere containing S02« The mechanism is thought to involve the
metal ion catalyzed conversion to sulfuric acid of the 863 absorbed by
the collagen of the leather. Using sulfur-35 labelled SOg. 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.
7.3.5 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, oxidation, 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 component may require replacement
due to corrosion 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.
7.3.5.1 Architectural Monuments—Hi storic and artistic structures
represent 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 pollution and urbanization as well as by weathering cycles and other
natural phenomena (CCMS 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.
7.3.5.2 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
7-36
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relationship between 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 alumina impregnated with KMnCty) 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 $03. 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
examples of damage, the authors called for protective measures to
prevent air pollution damage to paintings.
7.3,5.3 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 (CaS04-K2S04-H20) are found in the weathering
crust. However, such crusts are found even in locations with low S02
concentrations, suggesting that background SOg levels are sufficient
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 lamination, 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.
7.3.5.4 Conservation and Mitigation Costs--Some indication of the
problem'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,
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TABLE 7-8. ESTIMATED COSTS ASSOCIATED WITH AIR POLLUTION DAMAGE TO
CULTURAL PROPERTY IN WEST GERMANY (AFTER SCHREIBER 1982)
co
CO
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 museuns
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
-------
France, Italy, and the United States. For example, the Italian
Parllment designated $200,000,000 In 1980 for a 5-year program to
restore and maintain the ancient monuments 1n Rome (Hofinann 1981) and It
1s 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).
7.4 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
1n 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 pollution 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:
0 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.
0 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.
° Incomplete knowledge of substitution scenarios where more
expensive material systems may replace more susceptible
materials.
• Inadequate knowledge of the exposure conditions of susceptible
materials, for example, coexistance 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 materials 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,
and concrete, and several other types of surfaces. Of the 357 areas
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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
S02 in the Boston area, together with humidity data and published 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
typical paint films. Then costs were assigned to the increase in
painting frequency, based on the S02-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 transmission 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 S02
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 annualized 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 to be 0.0028 mills/Kwh +_ 0.0011 to be added to customers'
electric bills. These estimates were based on an S02 concentration of
17 yg m~3. if $02 levels were allowed to reach the ambient air
quality standard of 80 yg m~3, the annualized extra maintenance cost
would rise to an estimated 0.0132 mills/Kwh +_ 0.0052.
Cost estimates for damage to galvanized roofing, siding, and
guttering required 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 S02 and
moisture on zinc, together with estimates for the thickness of zinc
coatings and various maintenance scenarios 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 nr3. At the primary standard
of 80 yg nr3, the best estimate of per capita costs would be $1.80.
7-40
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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.
7.5 MITIGATIVE MEASURES
Assuming that acidic deposition produces significant damage to
materials, the primary mitigative measure is to reduce the
concentrations of acidic deposition components in the ambient
atmosphere. However, as has been implied in the foregoing discussion,
the relative amount of damage and associated costs from acidic
deposition is not known. Therefore, at present there is no basis for a
given degree of control of acidic deposition components and precursors
that will eliminate or bring to acceptable levels the potential damage
from this source.
Given that there is some degree of damage to materials 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
0 Physical and chemical nature of the material,
0 Age and state of repair of the materials system,
0 Availability and cost of substitute materials,
0 Feasibility of isolating the object or surface of concern from
the ambient environment,
o 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
materials in question regarding the relative importance of the
damage.
As stated earlier, material damage from acidic deposition is
generally indistinguishable 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 themselves. 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.
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7.6 CONCLUSIONS
From a review of the available literature on the effects of acidic
deposition on materials the following conclusions are drawn:
° Several senarios and mechanisms exist for damage to materials
from acidic deposition including both long-range transport and
local source emissions (Section 7.1).
o Without question acidic deposition causes significant incremental
damage to materials beyond that caused by natural environmental
phenomena (Section 7.1).
o 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.,
$02) 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., condensed dew) is believed to be more due to
acidic deposition than to acidified rain produced from long-range
transport of pollutants and their reaction products (Section 7.2)
o Reliable cost estimates for material damage from acidic
deposition are at present fragmentary because they deal with only
selected material systems or linked 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 7.3 and 7.4).
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
developing methods to prevent damage (Section 7.3.5).
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 (Section
7.3).
o Studies that accurately assess damage costs associated with
acidic deposition are needed (Section 7.4).
0 Further research is needed in the development of 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 7.5).
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The Acidic Deposition Phenomenon and Its
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