Draft
Do Not Quote or Cite
External Review Draft No. 2
February 1981
Air Quality Criteria
for Participate Matter
and Sulfur Oxides
Volume
NOTICE
This document is a preliminary 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 and
policy implications.
Environmental Criteria and Assessment Office
Office of Health and Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Research Triangle Park, N.C. 27711
-------
NOTE TO READER
The Environmental Protection Agency is revising the existing criteria
documents for particulate matter and sulfur oxides (PM/SOX) under Sections 108
and 109 of the Clean Air Act, 42 U.S.C. §§ 7408, 7409. The first external
review draft of a revised combined PM/SO criteria document was made available
for public comment in April 1980.
The Environmental Criteria and Assessment Office (ECAO) filled more than
4,000 public requests for copies of the first external review draft. Because
all those who received copies of the first draft from ECAO will be sent copies
of the second external review draft, there is no need to resubmit a request.
To facilitate public review, the second external review draft will be
released in five volumes on a staggered schedule as the volumes are completed.
Volume I (containing Chapter 1), Volume II (containing Chapters 2, 3, 4, and 5),
Volume III (containing Chapters 6, 7, and 8), Volume IV (containing Chapters 9
and 10), and Volume V (containing Chapters 11, 12, 13, and 14) will be released
during January-February, 1981. As noted earlier, they will be released as
volumes are completed, not in numerical order by volume.
The first external review draft was announced in the Federal Register of
April 11, 1980 (45 FR 24913). ECAO received and reviewed 89 comments from the
public, many of which were quite extensive. The Clean Air Scientific Advisory
Committee (CASAC) of the Science Advisory Board also provided advice and
comments on the first external review draft at a public meeting of August 20-22,
1980 (45 FR 51644, August 4, 1980).
As with the first external review draft, the second external review draft
will be submitted to CASAC for its advice and comments. ECAO is also soliciting
written comments from the public on the second external review draft and
requests that an original and three copies of all comments be submitted to:
Project Officer for PM/SO , Environmental Criteria and Assessment Office, MD-52,
X
U.S. Environmental Protection Agency, Research Triangle Park, N. C. 27711. To
facilitate ECAO's consideration of comments on this lengthy and complex docu-
ment, commentators with extensive comments should index the major points which
they intend ECAO to address, by providing a list of the major points and a
cross-reference to the pages in the document. Comments should be submitted
during the forthcoming comment period, which will be announced in the Federal
Register once all volumes of the second external review draft are available.
-------
External Review Draft No. 2
Draft February 1981
Do Not Quote or Cite
Air Quality Criteria
for Particulate Matter
and Sulfur Oxides
Volume
NOTICE
This document is a preliminary 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 and
policy implications.
Environmental Criteria and Assessment Office
Office of Health and Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Research Triangle Park, N.C. 27711
-------
PREFACE
This document is a revision of External Review Draft No. 1, Air
Quality Criteria for Particulate Matter and Sulfur Oxides, released in
April 1980. Comments received during a public comment period from April
15, 1980 through July 31, 1980, and recommendations made by the Clean Air
Scientific Advisory Committee in August have been addressed here.
Volume III contains Chapters 6, 7,' and 8, which cover atmospheric
chemistry and dispersion modeling, acidic deposition, and vegetation
effects of sulfur oxides and particulate matter. A Table of Contents for
Volumes I, II, III, IV, and V follows.
11
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CONTENTS
VOLUMES I, II, HI, IV, AND V
Page
Volume I.
Chapter 1. Executive Summary 1-1
Volume II.
Chapter 2. Physical and Chemical Properties of Sulfur
Oxides and Particulate Matter 2-1
Chapter 3. Techniques for the Collection and Analysis of
Sulfur Oxides, Particulate Matter, and Acidic
Precipitation 3-1
Chapter 4. Sources and Emissions 4-1
Chapter 5. Environmental Concentrations and Exposure 5-1
Volume III.
Chapter 6. Atmospheric Transport, Transformation and
Deposition 6-1
Chapter 7. Acidic Deposition 7-1
Chapter 8. Effects on Vegetation 8-1
Volume IV.
Chapters. Effects on Visibility and Climate 9-1
Chapter 10. Effects on Materials 10-1
Volume V.
Chapter 11. Respiratory Deposition and Biological Fate
of Inhaled Aerosols and S0? 11-1
Chapter 12. Toxicological Studies 12-1
Chapter 13. Controlled Human Studies 13-1
Chapter 14. Epidemiology Studies on the Effects of Sulfur
Oxides and Particulate Matter on Human
Health 14-1
iii
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CONTENTS
6. ATMOSPHERIC TRANSPORT, TRANSFORMATION AND DEPOSITION 6-1
6.1 INTRODUCTION 6-1
6. 2 CHEMICAL TRANSFORMATION PROCESSES 6-3
6.2.1 Chemical Transformation of S0? and Particulate
Matter 7 6-3
6.2.2 Field Measurements on the Rate of S0? Oxidation 6-4
6.3 PHYSICAL REMOVAL PROCESSES 7 6-4
6.3.1 Dry Deposition 6-7
6.3.1.1 Sulfur Dioxide Dry Deposition 6-9
6.3.1.2 Particle Dry Deposition 6-10
6.3.2 Precipitation Scavenging 6-13
6.3.2.1 SO, Wet Removal 6-16
6.3.2.2 Pafticle Wet Removal 6-19
6.4 TRANSPORT AND DIFFUSION 6-20
6.4.1 The Planetary Boundary Layer 6-24
6.4.2 Horizontal Transport and Pollutant Residence Times. 6-26
6. 5 AIR QUALITY SIMULATION MODELING 6-29
6.5.1 Gaussian Plume Modeling Techniques 6-30
6.5.2 Long Range Air Pollution Modeling 6-31
6.5.3 Model Evaluation and Data Bases 6-35
6.5.4 Atmospheric Budgets 6-36
6.6 SUMMARY 6-36
6. 7 REFERENCES 6-38
7. ACIDIC DEPOSITION 7-1
7.1 INTRODUCTION 7-1
7.1.1 Overview of the Problem 7-1
7.1.2 Ecosystem Dynamics 7-5
7.2 CAUSES OF ACIDIC PRECIPITATION 7-12
7.2.1 Emissions of Sulfur and Nitrogen Oxides 7-12
7.2.2 Transport of Nitrogen and Sulfur Oxides 7-14
7.2.3 Formation 7-20
7.2.3.1 Composition and pH of Precipitation 7-21
7.2.3.2 Geographic Extent of Acidic Precipitation 7-28
7.2.4 Acidic Deposition 7-34
7. 3 EFFECTS OF ACIDIC DEPOSITION 7-36
7.3.1 Aquatic Ecosystems 7-36
7.3.1.1 Acidification of lakes and streams 7-36
7.3.1.2 Effects on decomposition 7-45
7.3.1.3 Effect on primary producers and primary
productivity 7-48
7.3.1.4 Effects on invertebrates 7-54
7.3.1.5 Effects on fish 7-57
7.3.1.6 Effects on vertebrates other than fish 7-64
7.3.2 Terrestrial Ecosystems 7-68
7.3.2.1 Effects on soils 7-68
7.3.2.2 Effects on vegetation 7-78
7.3.2.2.1 Direct effects on vegetation 7-79
7.3.2.3 Effects on Human Health 7-89
7.3.2.4 Effects of Acidic Precipitation on Materials.. 7-89
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7.4 ASSESSMENT OF SENSITIVE AREAS 7-92
7.4.1 Aquatic Ecosystems 7-92
7.4.2 Terrestrial Ecosystems 7-98
7.5 SUMMARY 7-102
7.6 REFERENCES 7-106
8. EFFECTS ON VEGETATION 8-1
8.1 GENERAL INTRODUCTION AND APPROACH 8-1
8.2 REACTION OF PLANTS TO S02 EXPOSURES 8-2
8.2.1 Introduction to Terminology 8-2
8.2.2 Wet and Dry Deposition of Sulfur Compounds on
Leaf Surfaces 8-3
8.2.3 Routes and Methods of Entry Into the Plant 8-3
8.2.4 Cellular and Biochemical Changes 8-6
8.2.5 Acute Foliar Injury 8-6
8.2.6 Chronic Foliar Injury 8-8
8.2.7 Classification of Plant Sensitivity to SO, 8-9
8.2.8 Beneficial "Fertilizer" Effects 8-12
8.2.9 Foliar Versus Whole Plant Responses 8-14
8.3 DOSE-RESPONSE RELATIONSHIPS - SO, 8-16
•8.4 EFFECTS OF MIXTURES OF SO, AND OTHER POLLUTANTS INCLUDING
PARTICULATE MATTER 8-42
8.4.1 Sulfur Dioxide and Ozone 8-42
8.4.2 Sulfur and Nitrogen Dioxide 8-46
8.4.3 Sulfur Dioxide and Hydrogen Fluoride 8-46
8.4.4 Sulfur Dioxide, Nitrogen Dioxide and Ozone 8-46
8.4.5 Summary 8-48
8.5 EFFECTS OF NON-POLLUTANT ENVIRONMENTAL FACTORS ON SO, PLANT
EFFECTS 7 8-48
8.5.1 Temperature 8-48
8.5.2 Relative Humidity 8-49
8.5.3 Light 8-49
8.5.4 Edaphic Factors 8-50
8.5.5 SO, and Biotic Plant Pathogen Interactions 8-50
8.6 PLANT EXPOSURE TO PARTICULATE MATTER 8-51
8.6.1 Deposition Rates 8-51
8.6.2 Routes and Methods of Entry Into Plants 8-52
8.7 REACTION OF PLANTS TO PARTICULATE EXPOSURE 8-53
8.7.1 Symptomatology of Particle-Induced Injury 8-53
8.7.2 Classification of Plant Sensitivity—Particles 8-56
8.8 DOSE-RESPONSE RELATIONSHIPS—PARTICULATES 8-57
8.9 INTERACTIVE EFFECTS ON PLANTS WITH THE ENVIRONMENT—
PARTICULATE MATTER 8-58
8.9.1 Biotic Interactions 8-58
8.10 EFFECTS OF SULFUR DIOXIDE AND PARTICULATES ON NATURAL
ECOSYSTEMS 8-58
8.10.1 Sulfur Dioxide in Terrestrial Ecosystems 8-59
8.10.2 Ecosystem Response to Sulfur Dioxide 8-62
8.10.3 Response of Natural Ecosystems to Particulate Matter.. 8-67
8.11 SUMMARY 8-69
8-12 REFERENCES 8-73
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LIST OF FIGURES
Figure Page
6-1 Pathway processes of airborne pollutants ^ 6-2
6-2 Predicted deposition velocities at 1 m fo$ |j*=30 cms and
particle densities of 1,4, and 11.5 g cm 6-14
6-3 Basic factors influencing precipitation scavenging 6-15
6-4 Abundance of dissolved S02 species as a function of pH (25°C)... 6-18
6-5 Relationship between rain scavenging rates and particle size 6-22
6-6 Percentages of aerosol particles of various sizes removed by
precipitation scavenging 6-23
6-7 Estimated residence times for select pollutant species and their
associated horizontal transport scale 6-27
6-8 Trajectory modeling approaches 6-33
7-1 The nitrogen cycle. Organic phase shaded 7-9
7-2 Law of tolerence 7-11
7-3 Historical patterns of fossil fuel consumption in the
United States 7-13
7-4 Forms of coal usage in the United States 7-15
7-5a Trends in emissions of sulfur dioxides 7-16
7-5b Trends in emissions of nitrogen oxides 7-16
7-6 Point sources of sulfur oxides emissions over 100 tons per year. 7-17
7-7 Point sources of nitrogen oxides emissions over 100 tons per
year 7-18
7-8 Trends in mean annual concentrations of sulfate, ammonium,
and nitrate in precipitation 7-22
7-9 Comparison of weighted mean monthly concentrations of sulfate
in incident precipitation collected in Walker Branch
Watershed, Tenn. (WBW) and four MAP3S precipitation chemistry
monitoring stations in New York (0,0), Pennsylvania (A), and
Virginia (n) 7-26
7-10 Seasonal variations in pH (A) and ammonium and nitrate con-
centrations (B) in wet-only precipitation at Gainesville,
Florida 7-27
7-11 Seasonal variation of precipitation pH in the New York
Metropol itan Area 7-30
7-12 History of acidic precipitation at various sites in and
adjacent to State of New York 7-31
7-13 Location of acidic precipitation monitoring stations 7-33
7-14 Annual mass transfer rates of sulfate expressed as a percentage
of the estimated total annual flux of the element to the
forest floor beneath a representative chestnut oak stand 7-35
7-15 Schematic representation of the hydrogen ion cycle 7-38
7-16 pH and calcium concentrations in lakes in norther and
northwestern Norway sampled as part of the regional survey of
1975, in lakes in northwestern Norway sampled in 1977 (o) and
in lakes in southernmost and southeastern Norway sampled in
1974 (o) 7-42
7-17 The pH value and sulfur loads in lake waters with extremely
sensitive surroundings (curve 1) and with slightly less
sensitive surroundings (curve 2) 7-43
vi
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7-18 Total dissolved Al as a function of pH level in lakes in
acidified areas in Europe and North America 7-44
7-19 pH levels in Little Moose Lake, Adirondack region of New York
State, at a depth of 3 meters and at the lake outlet 7-46
7-20 Numbers of phytoplankton species in 60 lakes having different
pH values on the Swedish West Coast, August 1976 7-50
7-21 Percentage distribution of phytoplankton species and their
biomasses 7-51
7-22 The number of species of crustacean zooplankton observed in
57 lakes during a synoptic survey of lakes in southern Norway... 7-55
7-23 Frequency distribution of pH and fish population status in
Adirondack Mountain lakes greater than 610 meters elevation 7-59
7-24 Frequency distribution of pH and fish population status in 40
Adirondack lakes greater than 610 meters elevation, surveyed
during the period 1929-1937 and again in 1975 7-60
7-25 Norwegian salmon fishery statistics for 68 unacidified and 7
acidified rivers 7-61
7-26 Showing the exchangeable ions of a soil with pH+7, the soil
solution composition, and the replacement of Na and
H from acid rain 7-69
7-27 Regions in North America with lakes that are sensitive
to acidification by acid precipitation by virtue of their
underlying bedrock characteristics 7-94
7-28 Equivalent percent composition of major ions in Adirondack
lake surface waters (215 lakes) sampled in June 1975 7-96
7-29 Percent frequency distribution of sulfate concentrations
in surface water from lakes in sensitive regions 7-97
7-30 Soils of the Eastern United States sensitive to acid rainfall... 7-100
8-1 A conceptual model of potential responses by plants following
exposure to various doses of sulfur dioxide 8-4
8-2 Exposure thresholds for minimum, maximum and average sensitivity
of 33 plant species to visible foliar injury by S0? 8-11
8-3 Map of the United States indicating major areas of sulfur
deficient soils 8-13
8-4 Conceptual model of the factors involved in air pollution
effects (dose-response) on vegetation 8-19
8-5 Percentage of plant species visibly injured as a function of
peak, 1-hr, and 3-hr S0» concentrations 8-21
8-6 Regression of yield response vs. transformed dose (ppm hr) for
controlled exposures using field chambers (zero and positive
effects excluded from regression analysis) 8-37
8-7 Regression of yield effects vs. transformed dose (ppm hr)
for laboratory and greenhouse studies using agricultural,
ornamental, and nati ve herbs 8-38
8-8 Yield responses vs. SO,, dose for Norway spruce (Keller, 1980)
and white pine (Linzon, 1971) 8-40
8-9 The sulfur cycle 8-60
vn
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LIST OF TABLES
Table
6-1 Field measurements on the rates of S02 oxidation in plumes 6-5
6-2 Average dry deposition velocity of S02 by surface type... 6-9
6-3 Laboratory measurements of deposition velocities of particles.... 6-11
6-4 Field measurements of deposition velocities of particles 6-12
6-5 Predicted particle deposition velocities 6-16
6-6 Field measurements of scavenging coefficients of particles 6-21
6-7 Summary of long range transport air pollution models 6-34
7-1 Composition of ecosystems 7-7
7-2 Deposition of sulfuric and nitric acids in precipitation in
eastern North America 7-23
7-3 Mean pH values in the New York metropolitan area 7-29
7-4 Storm type classification 7-29
7-5 Chemical composition (Mean ± standard deviation) of acid
lakes (pH <5) in regions receiving highly acidic
precipitation (pH <4.5), and of soft-water lakes in areas
not subject to highly acidic precipitation (pH >4.8) 7-39
7-6 pH levels identified in field surveys as critical to
long-term survival of fish populations 7-62
7-7 Changes in aquatic biota likely to occur with increasing
aci di ty 7-65
7-8 Summary of effects on aquatic organisms associated with a
range in pH 7-66
7-9 Potential effects of acid precipitation on soils 7-70
7-10 Types of direct, visible injury reported in response to acidic
wet deposition 7-80
7-11 Thresholds for visible injury and growth effects associated with
experimental studies of wet deposition of acidic substances
(after Jacobson, 1980a,b) 7-83
7-12 Lead and copper concentration and pH of water from pipes
carrying outflow from Hinckley Basin and Hanns and Steele
Creek Basin, near Amsterdam, New York 7-90
7-13 Composition of rain and hoarfrost at Headingley, Leeds 7-92
7-14 The sensitivity to acid precipitation baseiji on: buffer
capacity against pH-change, retention of H , and adverse
effects on soils 7-99
8-1 Relationship of biochemical response to visual symptoms of
plant injury 8-7
8-2 Sulfur dioxide concentrations causing visible injury to various
sensitivity groupings of vegetation 8-10
8-3 Summary of studies reporting results of S0» exposures using
exposure systems and/or chambers over plants under field
conditions 8-23
8-4 Summary of studies reporting results of SO- exposure under
laboratory conditions for agronomic and horticultural crops 8-25
8-5 The degree of injury of eastern white pine observed at various
distances from the Sudbury smelters for 1953-63 8-29
vm
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8-6 Summary of studies reporting results of S02 exposure under
laboratory conditions for various tree species 8-30
8-7 Dose-response information summarized from literature pertaining
to native plants as related to foliar, yield and specific
effects by increasing S0? dose 8-33
8-8 Effects of mixtures of S0? and 03 on plants 8-44
8-9 Effects of mixtures of SCn and N02 on plants 8-47
8-10 Plants sensitive to heavy metals, arsenic, and boron as
accumulated in soils and typical symptoms expressed 8-55
IX
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7. ACIDIC DEPOSITION
7.1 INTRODUCTION
The occurrence of acidic precipitation in many regions of the United Sates, Canada,
northern Europe, Taiwan and Japan has become a major environmental concern. Acidic precipita-
tion in the Adirondack Mountains of New York State, in the eastern Precambrian Shield area of
Canada, in southern Norway and in southwest Sweden has been associated with the acidification
of waters in ponds, lakes and streams with a resultant disappearance of animal and plant life.
Acidic precipitation (rain and snow), also is believed to have the potential for leaching
elements from sensitive soils, causing direct and indirect injury to forests. It also has the
potential for damaging monuments and buildings made of stone, for corroding metals and for
deteriorating paint.
The story of acidic precipitation is an ever-changing one. New information concerning the
phenomenon is forthcoming nearly every day. The sections that follow emphasize the effects of
wet deposition of sulfur and nitrogen oxides and their products on aquatic and terrestrial eco-
systems. Dry deposition also plays an important role, but contributions by this process have
not been quantified. Because sulfur and nitrogen oxides are so closely linked in the forma-
tion of acidic precipitation, no attempt has been made to limit the discussion which follows
to the main topic of this document, sulfur oxides.
7.1.1 Overview of the Problem
The generally held hypothesis is that sulfur and nitrogen compounds are largely respon-
sible for the acidity of precipitation. The emissions of the sulfur and nitrogen compounds
involved in acidification are attributed chiefly to the combustion of fossil fuels. Emissions
may occur at ground level, as from automobile exhausts, or from stacks of 1000 feet or more in
height. Emissions from natural sources are also involved; however, in highly industrialized
areas, emissions from man-made sources well exceed those from natural sources. In the eastern
United States the highest emissions of sulfur oxides are from electric power generators using
coal, while in the West, emissions of nitrogen oxides, chiefly from automotive sources, pre-
dominate.
The fate of sulfur and nitrogen oxides, as well as other pollutants emitted into the
atmosphere, depends on their dispersion, transport, transformation and deposition. Sulfur and
nitrogen oxides may be deposited locally or transported long distances from the emission
sources. Therefore, residence time in the atmosphere will be brief if the emissions are
deposited locally or may extend to days or even weeks if long range transport occurs. The
chemical form in which emissions ultimately reach the receptor is determined by the complex
chemical transformations that take place between the emission sources and the receptor.
Long range transport over distances of hundreds or even thousands of miles allows time
for a greater number of chemical transformations to occur.
Sulfates and nitrates are among the products of the chemical transformations of sulfur
and nitrogen oxides. Ozone and other photochemical oxidants are believed to be involved in
XDSX7B/A 7-1 2-9-81
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the chemical processes that form them. When sulfates and nitrates combine with atmospheric
water, dissociated forms of sulfuric (H2$04) and nitric (HN03) acids result. When these acids
are brought to earth in rain and snow, acidic precipitation occurs. Because of long range
transport, acidic precipitation in a particular state or region can be the result of emissions
from sources in states or regions many miles away, rather than from local sources.
To date, however, the complex nature of the chemical transformation processes has not made it
possible to demonstrate a direct cause and effect relationship between emissions of sulfur and
nitrogen oxides and the acidity of precipitation.
Acidic precipitation is arbitrarily defined as precipitation with a pH less than 5.6.
This value has been selected because precipitation formed in a geochemically clean environment
would have a pH of approximately 5.6 due to the combining of carbon dioxide with air to form
carbonic acid.
Acidity of solutions is determined by the concentration of hydrogen ions (H ) present and
is expressed in terms of pH units—the negative logarithm of the concentration of hydrogen
ions. The pH scale ranges from 0 to 14, with a value of 7 representing a neutral solution.
Solutions with values less than 7 are acidic, while values greater than 7 are basic. Because
pH is a logarithmic scale, a change of one unit represents a tenfold change in acidity, hence
pH 3 is ten times as acidic as pH 4. Currently the acidity of precipitation in the north-
eastern United States normally ranges from pH 3.0 to 5.0; in other regions of the United
States precipitation episodes with a pH as low as 3.0 have been reported. For comparison, the
pH of some familiar substances are: cow's milk, 6.6; tomato juice, 4.3; cola (soft drink)
2.8, and lemon juice, 2.3.
The pH of precipitation can vary from event to event, from season to season and from
geographical area to geographical area. Substances in the atmosphere can cause the pH to
shift by making it more acidic or more basic. Dust and debris swept up in small amounts from
the ground into the atmosphere may become components of precipitation. In the West and
Midwest soil particles tend to be more basic, but in the eastern United States they tend to be
acidic. Gaseous ammonia from decaying organic matter makes precipitation more acidic, so in
areas where there are large stockyards or other sources of organic matter, acidic precipitation
would be more likely to occur.
In the eastern United States sulfur oxide emissions are greater than nitrogen oxides,
therefore, sulfates are greater contributors to the formation of acids in precipitation in
this region. The ratio between the two emissions, however, has been decreasing. Sulfate con-
centrations are greater in summer than in winter in the eastern United States. In California,
however, around some of the larger cities, nitrates contribute more to the formation of acidity
in rainfall. In coastal areas sea spray strongly influences percipitation chemistry by con-
tributing calcium, potassium, chlorine and sulfates. In the final analysis, the pH of preci-
pitation is a measure of the relative contributions of all of these components.
XDSX7B/A 7~2 2-9-81
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*
The impact of acidic precipitation on lakes, streams, ponds, forests, fields and manmade
objects, therefore, is not the result of a single, or even of several precipitation events, but
the result of continued additions of acids or acidifying substances over time. When did
precipitation become acidic? Some scientists state that it began with the industrial
revolution and the burning of large amounts of coal; others say it began in the United States
with the introduction of tall stacks in power plants in the 1950's; other scientists disagree
completely and state that rain has always been acidic. In other words, no definitive answer to
the question exists at the present time, nor is there data to indicate with any accuracy pH
trends in precipitation. The pH of rain has not been continuously monitored in the United
States for any period of time, so no data exist. In Scandinavia, on the other hand, the pH of
rain has been monitored for many years, therefore a determination of the time of origin can be
made.
Though acidic precipitation (wet deposition) is usually emphasized, it is not the only
process by which acids or acidifying substances are added to bodies of water or to the land.
Dry deposition also occurs. During wet deposition substances such as sulfur and nitrogen
oxides are scavenged by precipitation (rain and snow) and deposited on the surface of the
earth. Dry deposition processes include gravitational sedimentation of particles, impaction
of aerosols and the sorption and absorption of gases by objects at the earth's surface or by
the soil or water. Gases, particles and solid and liquid aerosols can be removed by both wet
and dry deposition. Dew, fog and frost are also involved in the deposition processes but do
not strictly fall into the category of wet or dry deposition. Dry deposition processes are
not as well understood as wet deposition at the present time, however, all of the deposition
processes contribute to the gradual accumulation of acidic or acidifying substances in the
environment. In any event, percipitation at the present time is acidic and has been
associated with changes in ponds, lakes and streams that are considered by humans to be
detrimental to their welfare.
The most visible changes associated with acidic deposition, that is both wet and dry
processes, are those observed in the lakes and streams of the Adirodack Mountains in New York
State, the Pre-cambrian Shield areas of Canada and in the Scandinavian countries. In these
regions the pH of the fresh water bodies has decreased, causing changes in animal and plant
populations. The most readily observable has been the decrease in fish populations.
The chemistry of fresh waters is determined primarily by the geological structure (soil
system and bedrock) of the lake or stream catchment basin, by the ground cover and by land
use. Near coastal areas (up to 100 miles) marine salts also may be important in determining
the chemical composition of the stream, river or lake.
Sensitivity of a lake to acidification depends on the acidity of both wet and dry deposi-
tion plus the same factors—the soil system of the drainage basin, the canopy effects of the
ground cover and the composition of the waterbed bedrock—that determine the chemical composi-
tion of fresh water bodies. The capability, however, of a lake and its drainage basin to
neutralize incoming acidic substances is determined largely by the composition of the bedrocks.
XDSX7B/A 7-3 2-9-81
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Soft water lakes, those most sensitive to additions of acidic substances, are usually found
in areas with igneous bedrock which contributes few solids to the surface waters, whereas hard
waters contain large concentrations of alkaline earths (chiefly bicarbonates of calcium and
sometimes magnesium) derived from limestones and calcareous sandstones in the drainage basin.
Alkalinity is associated with the increased capacity of lakes to neutralize or buffer the in-
coming acids. The extent to which acidic precipitation contributes to the acidification pro-
cess has yet to be determined.
The disappearance of fish populations from freshwater lakes and streams is usually one of
the most readily observable signs of lake acidification. Death of fish in acidified waters has
been attributed to the modification of a number of physiological processes by a change in pH.
Two patterns of pH change have been observed. The first involves a sudden short-term drop in
pH and the second, a gradual decrease in pH with time. Sudden short-term drops in pH often
result from a winter thaw or the melting of the snow pack in early spring and the release of
the acidic constituents of the snow into the water. Fish may be killed at pH levels above
those normally causing death.
A gradual decrease in pH, particularly below 5, can interfere with reproduction and
spawning of fish until elimination of the population occurs. In some lakes, aluminum mobili-
zation in fresh waters at a pH below 5 has resulted in fish mortality .
Although the disappearance of and/or reductions in fish populations are usually empha-
sized as significant results of lake and stream acidification, changes of equal or greater
importance are the effects on other aquatic organisms ranging from waterfowl to bacteria.
Organisms at all tropic (feeding) levels in the food web appear to be affected. Species re-
duction in number and diversity may occur, biomass (total number of living organisms in a
given volume of water) may be altered and processes such as primary production and decompo-
sition impaired.
Primary production and decomposition are the bases of the two major food webs (grazing
and detrital) within an ecosystem by which energy is passed along from one organism to another
through a series of steps of eating and being eaten. Green plants, through the process of
photosynthesis, are the primary energy producers in the grazing web, while bacteria initiate
the detrital food web by feeding on dead organic matter. Disruption of either of these two
food webs results in a decrease in the supply of minerals and nutrients, interferes with their
cycling and also reduces energy flow within the affected ecosystems. Acidification of lakes
and streams affects both these processes when alteration of the species composition and struc-
ture of the pondweed and algae plant communities occurs due to a slowing down in the rate of
microbial decomposition.
At present there are no documented observations or measurements of changes in natural
terrestrial ecosystems that can be directly attributed to acidic precipitation. The informa-
tion available is an accumulation of the results of a wide variety of controlled research
XDSX7B/A 7-4 2-9-81
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*
approaches largely in the laboratory, using in most instances some form of "simulated" acidic
rain, frequently dilute sulfuric acid. The simulated "acid rains" have deposited hydrogen
(H ), sulfate (S0^~) and nitrate (N03) ions on vegetation and have caused nicrotic lesions in a
wide variety of plants species under greenhouse and laboratory conditions. Such results must
be interpreted with caution, however, because the growth and morphology of leaves under green-
house conditions are often not typical of field conditions. Based on laboratory studies, the
sensitivity of plants to acidic deposition seems to be associated with the wettability of leaf
surfaces. The shorter the time of contact, the lower the resulting dose and the less likeli-
hood of injury.
Soils may become gradually acidified from an influx of hydrogen (H ) ions. Leaching of
the mobilizable forms of mineral nutrients may occur. The rate of leaching is determined by
the buffering capacity of the soil and the amount and composition of precipitation. Unless
the buffering capacity of the soil is strong and/or the salt content of precipitation is high,
leaching will in time result in acidification. At present there are no studies showing this
process has occurred because of acidic precipitation.
Damage to monuments and buildings made of stone, corrosion of metals and deterioration of
paint can result from acidic precipitation. Because sulfur compounds are a dominant component
of acidic precipitation and are deposited during dry deposition also, the effects resulting
from the two processes cannot be distinguished. In addition, the deposition of sulfur com-
pounds on stone surfaces provides a medium for microbial growth that can result in deteri-
oration.
Human health effects due to the acidification of lakes and rivers have been postulated.
Fish in acidified water may contain toxic metals mobilized due to the acidity of the water.
Drinking water may contain toxic metals or leach lead from the pipes bringing water into the
homes. Humans eating contaminated fish or drinking contaminated water could become ill. No
instances of these effects having occurred have been documented.
Several aspects of the acidic precipitation problem remain subject to debate because
exisitng data are ambiguous or inadequate. Important issues include: (1) the rate at which
rainfall is becoming more acidic and the rate at which the problem is becoming geographically
more widespread; (2) the quantitative contributions of various acids to the overall acidity of
rainfall; (3) the relative extent to which the acidity of rainfall in a region depends on local
emissions of nitrogen and sulfur oxides versus emissions transported from distant sources; (4)
the relative importance of changes in total mass emission rates compared to changes in the
nature of the emission patterns (ground level versus tall stacks) in contributing to regional
acidification of precipitation; and (5) the relative contribution of wet and dry deposition to
the acidification of lakes and streams.
7.1.2 Ecosystem Dynamics
The emission of sulfur and nitrogen oxides into the atmosphere, their transformation,
transport and deposition, either as acidic precipitation or in dry form, as well as the
XDSX7B/A 7-5 2-9-81
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*
responses of aquatic and terrestrial ecosystems to acidic deposition are all natural phenomena
that have been in existence as long as humans can remember. Environmental problems arise
because the natural systems are being overloaded by emissions from the combustion of fossil
fuels from anthropogenic sources.
Life on the planet Earth depends on the movement of energy and minerals through the bio-
sphere, that thin layer of life surrounding the earth. The living systems (forest, grass-
lands, cultivated fields, lakes, rivers, estuaries and oceans) within the biosphere obtain
energy from the sun, nutrients from the earth's crust, the lithosphere, gases from the atmos-
phere and water from the hydrosphere. All of the living systems are interdependent. Energy
and nutrients move from one to another. The living systems together with their physical
environment, the lithosphere, hydrosphere and atmosphere, make up the ecosystem that is the
planet Earth (Billings, 1978; Boughey, 1971; Odum, 1971; Smith, 1980).
Ecosystems are basically energy processing systems "whose components have evolved
together over a long period of time. The boundaries of the system are determined by the
environment, that is, by what forms of life can be sustained by the environmental conditions
of a particular region. Plant and animal populations within the system represent the objects
through which the system functions." (Smith, 1980)
Ecosystems are composed of biotic (living) and abiotic (non-living) components. The
biotic component consists of: (a) producers, green plants that capture the energy of the sun;
(b) consumers that utilize the food stored by the producers for their energy; and (c) the
decomposers who break down dead organic matter and convert it into inorganic compounds again.
(See Table 7-1). The abiotic components are the soil matrix, sediment, particulate matter,
dissolved organic matter and nutrients in aquatic systems, and dead or inactive organic matter
in terrestrial systems (See Table 7-1) (Billings, 1978; Boughey, 1971; Smith, 1980).
Ecosystems are open systems that receive both abiotic (gases, nutrients and radient
energy) and biotic contributions from their environment and in turn contribute energy, water,
gases and nutrients. Energy flows through the system while water, gases and nutrients are
usually recycled and fed back into the system and to some extent control its functioning.
Populations are the structural elements of the ecosystem through which energy flows and nutri-
ents cycle (Smith, 1980).
Energy from the sun is the driving force in ecosystems. If the sun's energy were cut off
all ecosystems would cease to function. The energy of the sun is captured by green plants
through the process of photosynthesis and stored in plant tissues. This stored energy is
passed along through ecosystems by a series of feeding steps, known as food chains, in which
organisms eat and are eaten. Energy flows through ecosystems in two major food chains, the
grazing food chain and the detrital food chain. The amount of energy that passes through the
two food chains varies from community to community. The detrital food chain is dominant in
most terrestrial and shallow-water ecosystems. The grazing food chain may be dominant in deep-
water aquatic ecosystems (Smith, 1980). The two fundamental processes involved in these two
XDSX7B/A 7-6 2-9-81
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TABLE 7-1. COMPOSITION OF ECOSYSTEMS
Component
Description
Biotic (biological):
Individuals
Producers
Consumers
Decomposers
Populations
Communities
Abiotic (physical):
Energy
Water
Atmosphere
Fire
Topography
Geological
strata
Plants, animals (man), and microorganisms.
These are either producers, consumers, or
decomposers.
Green plants.
Herbivores, carnivores.
Macroorganisms (mites, earthworms, millipedes,
and slugs) and microorganisms (bacteria
and fungi).
Groups of interbreeding organisms of the same
kind, producers, consumers or decomposers,
occupying a particular habitat.
Interacting populations linked together by
their responses to a common environment.
Radiation, light, temperature, and heat flow.
Liquid, ice, etc.
Gases and wind.
Combustion.
Surface features.
Soil, a complex system. Nutrients. (Minerals)
XDSX7B/A
7-7
2-9-81
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*
food chains are photosynthesis, the capture of energy from the sun by green plants, and
decomposition, the final dissipation of energy and the reduction of organic matter into inor-
ganic nutrients.
In addition to the flow of energy, the existence of the living world depends upon the
circulation of nutrients through the ecosystems. Both energy and nutrients move through the
ecosystem as organic matter. It is not possible to separate one from the other. Both influ-
ence the abundance of organisms, the metabolic rate at which they live and the complexity and
structure of the ecosystem (Smith, 1980). Nutrients, unlike energy, after moving from the
living to the non-living return to the living components of the ecosystem in a perpetual
cycle. It is through the cycling of nutrients that plants and animals obtain the minerals
necessary for their existence.
The gaseous and sedimentary cycles are the two basic types of nutrient or biogeochemical
cycles. The gaseous cycles involve carbon, oxygen and nitrogen. Water, also, is sometimes
considered as belonging to the gaseous cycle. In the gaseous cycles, the main nutrient reser-
voir is the atmosphere and the ocean. In the sedimentary cycle, to which phosphorus belongs,
the soil and rocks of the earth's crust is the reservoir. The sulfur cycle is a combination
of the two cycles because it has reservoirs in the atmosphere and the earth's crust.
Nitrogen, sulfur and water cycles are involved in acidic deposition. Nitrogen, through
the agency of plants (chiefly legumes and blue green algae), moves from the atmosphere to the
soil and back (see Figure 7-1). Human intrusion into the nitrogen cycles include the addition
of nitrogen oxides to the atmosphere and nitrates to aquatic ecosystems. Sulfur enters the
atmosphere from volcanic eruptions, from the surface of the ocean, from gases released in the
decomposition processes and from the combustion of fossil fuels (see Chapter 8 for details).
Both the nitrogen and sulfur cycles have been overloaded by the combustion of fossil fuels by
man. For these cycles to function, an ecosystem must possess a number of structured relation-
ships among its components. By changing the amounts of nitrogen and sulfur moving through the
cycles, humans have perturbed or upset the structured relationships that have existed for
thousands of years and altered the movement of the elements through the ecosystems. The path-
ways the elements take through the system depend upon the interaction of the populations and
their relationships to each other in terms of eating and being eaten.
Change is one of the basic characteristics of our environment. Weather changes from day
to day, temperatures rise and fall, rains come and go, soils erode, volcanoes erupt, and winds
blow across the land. These are natural phenomena. Significant environmental changes also
result when human beings clear forests, build cities and factories, and dam rivers. All of
these environmental changes influence the organisms that live in the ecosystems where the
changes are occurring (Moran et al., 1980).
Existing studies indicate that changes occurring within ecosystems, in response to pollu-
tion or other disturbances, follow definite patterns that are similar even in different eco-
systems. It is, therefore, possible to predict the basic biotic responses of an ecosystem to
XDSX7B/A 7-8 2-9-81
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"if
NZ
lTROGEN
NITROGEN
ATMOSPHERE
BIOLOGICAL FIXATION j
VOLCANIC
ERUPTION
WEATHERING
OF ROCKS
FOREST & GRASSLAND FIRES
ELECTRICAL
AND
PHOTOCHEMICAL
FIXATION
STORAGE OF
NITROGENOUS
COMPOUNDS IN
SEDIMENTS. SOILS,
AND SEDIMENTARY
ROCKS
NITROGEN
OXIDES:
NO, NO2
ANIMALS IN GRAZING
FOOD CHAIN
AUTOTROPHS
DETRITUS FOOD CHAIN
AMINO
NITROGEN
R-NH2
/ 1 NITRITE/ 1 AMMONIA
\. 1 N02~ \ 1 NH
DENITRIFICATION
NITRIFICATION
i DENITRIFICATION
Figure 7-1. The nitrogen cycle. Organic phase shaded.
-------
TV
disturbances such as caused by environmental stress (Woodwell, 1970; Woodwell, 1962; Odum,
1965; Garrett, 1967). These responses to disturbance are (1) removal of sensitive organisms
at the species and subspecies level due to differential kill; (2) reduction in the number of
plants and animals (standing crop); (3) inhibition of growth or reduction in productivity; (4)
disruption of food chains; (5) return to a previous state of development; and (6) modification
in the rates of nutrient cycling.
Ecosystems can respond to environmental changes or perturbations only through the response
of the populations of organisms of which they are composed (Smith, 1980). Species of
organisms sensitive to environmental changes are removed. Therefore, the capacity of an eco-
system to maintain internal stability is determined by the ability of individual organisms to
adjust their physiology or behavior. The success with which an organism copes with envi-
ronmental changes is determined by its ability to produce reproducing offspring. The size and
success of a population depends upon the collective ability of organisms to reproduce and
maintain their numbers in a particular environment. Those organisms that adjust best contri-
bute most to future generations because they have the greatest number of progeny in the popu-
lation (Smith, 1980; Billings, 1978; Woodwell, 1970, 1962; Odum, 1971).
The capacity of organisms to withstand injury from weather extremes, pesticides, acidic
deposition or polluted air follows the principle of limiting factors (Billings, 1978; Odum,
1971; Moran et al., 1980; Smith, 1980). According to this principle, for each physical factor
in the environment there exists for each organism a minimum and a maximum limit beyond which
no members of a particular species can survive. Either too much or too little of a factor
such as heat, light, water, or minerals (even though they are necessary for life) can jeopar-
dize the survival of an individual and in extreme cases a species (Billing, 1978; Smith, 1980;
Boughey, 1971; Odum, 1971; Moran et al., 1980). The range of tolerance (see Figure 7-2) of an
organism may be broad for one factor and narrow for another. The tolerance limit for each
species is determined by its genetic makeup and varies from species to species for the same
reason. The range of tolerance also varies depending on the age, stage of growth or growth
form of an organism. Limiting factors are, therefore, factors which, when scarce or over-
abundant, limit the growth, reproduction and/or distribution of an organism (Billings, 1978;
Smith, 1980; Boughey, 1971; Odum, 1971; Moran et al., 1980). The increasing acidity of water
in lakes and streams is such a factor.
Organisms can exist only within their range of tolerance. Some populations of organisms,
annual plants, insects, and mice, for example, respond rapidly. They increase in numbers under
favorable conditions and decline rapidly when conditions are unfavorable. Populations of other
organisms, such as trees and wolves, fluctuate less in response to favorable or unfavorable
conditions. Ecosystems that contain both types of populations are more stable bacause they
are able to absorb changes and still persist because the structure of the ecosystem permits it
to persist even though populations within it fluctuate widely in response to environmental
changes (Smith 1980; Moiling, 1973). Other ecosystems are resistant; their structure enables
XDSX7B/A 7-10 2-9-81
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ZONE OF
INTOLERANCE
town LIMITS
OFTOLEft«»CE
ZONE OF
PHYSIOLOGICAL
STRESS
TOLERANCE RANGE
RANGE OF OPTIMUM
om« LIMITS
OF TOIEUIICE
ZONE OF
PHYSIOLOGICAL
STRESS
ZONE OF
INTOLERANCE
ORGANISMS
INFREQUENT
ORGANISMS
ABSENT
GREATEST
ABUNDANCE
ORGANISMS
INFREQUENT
ORGANISMS
ABSENT
LOW«-
-GRADIENT-
-fr-HIGH
Figure 7-2. Law of tolerance.
Source: Adapted from Smith (1980).
7-11
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them to resist changes. Typically, most resistant ecosystems have large living components,
trees for example, and store nutrients and energy in the standing biomass. Such resistant
systems, such as forests, once highly disturbed are very slow in returning to their original
state (Smith, 1980).
Aquatic ecosystems which lack components in which energy and nutrients may be stored for
long periods of time usually are not very resistant to environmental changes (Smith, 1980).
For example, an influx of pollutants such as effluents from sewage disrupts the system because
more nutrients enter the system than it can handle. However, since the nutrients are not
retained or recycled within the system it returns to its original state in a relatively short
time after the perturbation is removed.
No barriers exist between the various environmental factors or between an organism or
biotic community and its environment. Because an ecosystem is a complex of interacting com-
ponents, if one factor is changed, almost all will change eventually. "The ecosystem reacts
as a whole. It is practically impossible to wall off a single factor or organism in nature
and control it at will without affecting the rest of the ecosystem. Any change no matter-how
small is reflected in some way throughout the ecosystem: no 'walls' have yet been discovered
that prevent these interactions from taking place" (Billings, 1978).
Continued or severe perturbation of an ecosystem can overcome its resistance or prevent
its recovery with the result that the original ecosystem will be replaced by a new system. In
the Adirondack Mountains of New York State, in eastern Canada and parts of Scandinavia the
original aquatic ecosystems have been and are continuing to be replaced by ecosystems different
from the original due to acidification of the aquatic habitat. Forest ecosystems appear to
be more resistant because, thus far, changes due to stress from acidifying substances have not
been detected. The sections that follow discuss the response of aquatic and terrestrial eco-
systems to stressing or perturbation by acidic deposition. Sulfur and nitrogen oxide emissions,
their transformation, transport and deposition in acidic form is elucidated in the context of
the ecosystem processes that were discussed above.
7.2 CAUSES OF ACIDIC PRECIPITATION
7.2.1 Emissions of Sulfur and Nitrogen Oxides
The generally held hypothesis is that increased emissions of sulfur and nitrogen com-
pounds are largely responsible for the acidity of precipitation (Smith, 1872; Bolin et al.,
1972; Likens, 1976). The emissions of the sulfur and nitrogen compounds ^involved in the
acidification are attributed chiefly to the combustion of fossil fuels. Emissions from natural
sources can also be involved; however, in highly industrialized areas emissions from man-made
sources usually exceed those from natural sources (see Chapter 4). In the eastern United
States 90 percent of the sulfur oxides in the atmosphere are from anthropogenic sources (see
Figure 7-6).
Since 1900 there has been a nearly exponential increase in the consumption of coal, gas,
and oil in the United States (see Figure 7-3). Although the total consumption of coal has not
increased greatly since about 1925, the consumption of oil and gas has continued to rise
XDSX7B/A 7-12 2-9-81
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I I I I I I
1850
2050
Figure 7-3. Historical patterns of fossil fuel consumption in the
United States
Source: Adapted from Hubbert (1976).
7-13
-------
precipitously, thus overshadowing coal as the dominant fuel source during the past 50 years
(Hubert, 1976). Within this overall increase in fossil-fuel use there have been shifts in the
pattern of consumption. Whereas formerly a considerable proportion of coal was used for trans-
portation and heating, these functions have since been taken over by oil and gas. Coal is now
predominantly devoted to electric power generation (Figure 7-4). In fact, electric power
generation is the primary factor accounting for an absolute increase in coal consumption over
the past two decades. (The decline in the use of coal in the 1930s was due to the general
economic depression, and the decline in the 1950s was due to the availability of relatively
inexpensive oil and gas.) Approximately 550 MM tons (National Research Council, 1978) were
used per year during 1918-1928 compared to 672 MM tons/year during 1979 (Hamilton, 1980).
There was, however, a seasonal shift in the pattern of coal consumption. Summer coal consump-
tion has increased since 1960, while winter consumption has decreased due to increased summer
usage by the electric utilities.
These changes in the pattern of fuel use have been accompanied by changes in the pattern
of pollutant emissions. Figure 7-5A and 7-5B illustrate the rise since 1940 in emissions of
sulfur and nitrogen oxides, the primary gaseous pollutants resulting from the combustion of
fossil fuels. Although there has been a net increase in both categories, the more consistent
rise has been in emissions of nitrogen oxides. Almost all (93 percent) emissions of sulfur
oxides in the United States arise from stationary point sources, principally industrial and
power plant stacks. Nitrogen oxide pollutants, on the other hand, originate about equally
from transportation (mobile) sources and from stationary sources, which include not only
industrial and power plants, but residential and institutional heating equipment as well (U.S.
Environmental Protection Agency, 1978). (see Chapter 5 of Air Quality Criteria for Oxides of
Nitrogen for a more detailed discussion.)
The geographic distributions of sources of the gaseous precursors of acidic precipitation
are depicted in Figures 7-6 and 7-7. Clearly, the dominant sources of sulfur oxides in the
United States are in the eastern half of the country, particularly the northeastern quadrant.
Major nitrogen oxide sources also show a tendency to be concentrated somewhat in the north-
eastern quadrant of the country. Chapter 4 should be consulted for a more detailed account of
the sources and emissions of sulfur oxides.
7.2.2 Transport of Nitrogen and Sulfur Oxides
Among the factors influencing the formation as well as the location where acidic deposi-
tion occurs is the long-range transport of nitrogen and sulfur oxides. Neither the gases nor
their transformation products always remain near the sources from which they have been
emitted. They may be transported for long distances downwind (Altshuller and McBean, 1979;
Pack et al., 1978; Cogbill and Likens, 1974).
The geographic picture of the problem of acidic precipitation in North America can be
better understood in the light of some information on prevailing wind patterns. Winds trans-
port the precursors of acidic precipitation from their points of origin to areas where the
XDSX7B/A 7-14 2-9-81
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I
Q.
1
O
o
o
0
tr
LU
>
800
700
600
500
400
300
200
100
IIII "T
TOTAL
I I
- OTHER
-OVENCOKE
ELECTRIC
UTILITIES
! L
1900 10 20 30 40 50 60 70 80 90 2000
YEAR
Figure 7-4. Forms of coal usage in the United States. Electric
power generation is currently the primary user of coal. (Data
from U.S. Bureau of Mines, Minerals Yearbooks 1933-1974).
Source: U.S. Bureau of Mines (1954, 1976).
7-15
-------
35
30
I 25
c
o
CO
O
co
CO
O
20
15
CO
Z
2 10
CO
CO
ui 5
x
O
CO
1940
35
30
25
20
15
10
5
TRANSPORTATION
1950
1960
YEAR
1970
1980
TOTAL
TRANSPORTATION
1940
1950
1960
YEAR
1970
1980
Figure 7-5a. Trends in emissions of sulfur dioxides.
Figure 7-5b. Trends in emissions of nitrogen oxides.
Source: U.S. Environmental Protection Agency (1978).
7-16
-------
Figure 7-6. Point sources of sulfur oxides emissions over 100 tons per year.
Broken circles represent estimated locations of point sources.
Source: U.S. Environmental Protection Agency (1978).
-------
Figure 7-7. Point sources of nitrogen oxides emissions over 100 tons per year.
Broken circles represent estimated locations of point sources.
Source: U.S. Environmental Protection Agency(1978).
-------
*
acidified rain and snow eventually fall. Prevailing winds in the eastern United States tend
to be from the west and southwest. Atmospheric pollutants, therefore, are carried in a
generally northeasterly direction. Thus, pollution originating in the Ohio River valley can
be carried toward the New England states. Seasonal meteorological patterns, however, can
modify the direction of windflow, particularly in the summer. The Maritime Tropical air
masses from the Gulf of Mexico that occur in late summer have the greatest potential for the
formation and transport of high concentrations of sulfate into the northeastern United States
and into eastern Canada (Altshuller and McBean, 1979).
Cogbill and Likens (1974) associated acidic rainfall in central New York during 1972-73
with high altitude air masses transported into the region from the Midwest. They stated that
the NO and S0? that is involved in acidic rain formation may be transported distances of 300
to 1500 km. Reports by Miller et al. (1978), Wolf et al. (1979), and Galvin et al. (1973) all
support the concept that the trajectories of the air masses which come from the Midwest carry
sulfur and nitrogen compounds which acidify precipitation in New York State.
A significant though disputed factor in this transport picture is the height at which the
pollutants are emitted. Industrial and power plant smokestacks emit their effluents into the
atmosphere at higher elevations than do motor vehicles or most space heating equipment. In
fact, there has been a trend since the 1960s toward building higher stacks as a means of
dispersing pollutants and thereby reducing pollutant concentrations in the vicinity of power
plants, smelters, and similar sources (Grennard and Ross, 1974). The result has been that
sulfur and nitrogen oxides are carried by prevailing winds for long distances and allowed to
diffuse over greater areas through the atmosphere. Concomitantly, long-range transport allows
greater time for chemical reactions to convert these pollutant gases into particulate forms
which are more easily removed by wet processes (Eliassen and Saltbones, 1975; Smith and Jeffrey,
1975; Prahm et al., 1976). Chapter 6 discusses the chemical transformations of wet and dry
deposition and transport and diffusion of sulfur oxides in the atmosphere. Sulfates and
nitrates combine with atmospheric water to form dissociated forms of nitric (HN03) and sulfuric
(H?SO.) acids. These acids are considered to be the main components of acidic precipitation.
The mechanisms of these chemical reactions are quite complex and depend on a host of
variables ranging from physical properties of the pollutants to weather conditions and the
presence of catalytic or interacting agents (Fisher, 1978). Although these processes of atmos-
pheric chemistry are not well understood, it does appear that the long-range transport of sul-
fur compounds can cover 1000 to 2000 km over three to five days (Pack et al., 1978). Thus,
the impact of sulfur pollutants in the form of acidic precipitation may be far removed from
their points of origin. It is not yet clear whether the atmospheric transport of nitrogen
oxide pollutants is comparable to that of sulfur compounds (Pack, 1978), but in the northeast
nitrates are currently thought to contribute 15 to 30 percent of the acidity of polluted pre-
cipitation. This figure has increased over the past few years and is expected to increase
still further in the future (National Research Council, 1978).
XDSX7B/A 7-19 2-9-81
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Evidence from northern Europe also supports the idea that acidic rainfall is a large-
scale regional problem involving long distances between emission sources and deposition of
acidic precipitation. The acid rains that have received intensive study in southern Scandi-
navia have been shown to result primarily from emissions of nitrogen and sulfur oxides in
Great Britain and the industrial regions of continental Western Europe (e.g., Holland, Belgium,
West Germany) (Brosset, 1973).
7.2.3 Formation
Precipitation is that portion of the global water cycle by which water vapor from the
atmosphere is converted to rain or snow and then is deposited on the earth surfaces (Smith,
1974). Water moves into the atmosphere by evaporation and transpiration (water vapor lost by
vegetation). Once it reaches the atmosphere, the water vapor is cooled, then condenses on
solid particles and soon reaches equilibrium with atmospheric gases. One of the gases is car-
bon dioxide. As carbon dioxode dissolves in water, carbonic acid (H9CO,) is formed. Carbonic
£. O
acid is a weak acid and in distilled water only dissociates slightly, yielding hydrogen ions
and bicarbonate ions (HC03 ). When in equilibrium with normal atmospheric concentrations and
pressures of carbon dioxide, the pH of rain and snow is approximately 5.6 (Likens et al.,
1979).
The pH of precipitation may vary and become more basic or more acidic depending on sub-
stances in the atmosphere. Dust and debris may be swept from the ground in small amounts and
into the atmosphere where it can become a component of rain. Soil particles are usually
2+
slightly basic in distilled water and release positive ions, such as calcium (Ca ), magnesium
2+ + +
(Mg ), potassium (K ), and sodium (Na ) into solution. Bicarbonate usually is the corre-
sponding negative ion. Decaying organic matter adds gaseous ammonia to the atmosphere.
Ammonia gas in rain or snow forms ammonium ions (NH. ) and tends to increase the pH. In
coastal areas sea spray plays a strong role in the chemistry of precipitation. The important
ions entering into precipitation—sodium, magnesium, calcium, potassium, and the anions
2"
chloride (Cl ) and sulfate (SO. )—are also those most abundant in ocean water (Likens, 1976;
Likens et al., 1979).
Gases which enter precipitation, in addition to C02, are sulfur dioxide (S0?) and the ni-
trogen oxides (NO ). Some sulfur gases originate from natural sources, e.g. volcanoes and
swamps. Others originate from industrial emissions. In the wet atmosphere, both S02 and HLS can
be oxidized to sulfuric acid. Nitrogen oxides in the atmosphere are converted to nitric acid
(Likens, 1976; Likens et al., 1979). Strong acids dissociate completely in dilute aqueous
solutions and lower the pH to less than 5.6. Acidic precipitation has been arbitrarily con-
sidered by many scientists to be rain or snow with a pH below 5.6 (Galloway and Cowling, 1978;
Wood, 1975; Likens et al., 1979).
Additional acidic or potentially acidifying substances present in both wet and dry depo-
sition are sulfur trioxide (S03=), sulfate S04~), nitric oxide (NO), nitrogen dioxide (N02),
nitrite (N02 ), nitrate (N03 ), ammonium (NH.+), chlorine (Cl~) hydrochloric acid (HC1), and
XDSX7B/A 7-20 2-9-81
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Brrfnsted acids [e.g., dissolved iron (Fe) and ammonium (NH4+)] (Whelpdale, 1978).
The amounts of the various substances in the atmosphere originating from seawater, desert
sands, volcanic islands, or vegetated land influence the chemistry of natural precipitation.
In regions with calcareous soils, calcium and bicarbonate may enter precipitation as dust,
subsequently increasing the pH of rain or snow to 6.0 or above (Likens et al., 1979).
7.2.3.1 Composition and pH of Precipitation—Sulfur and nitrogen compounds are chiefly
responsible for the acidity of precipitation. Continuous measurement of pH in rain by Likens
et al. (1972) for the Hubbard Brook Experimental Forest in Hew Hampshire from 1964 to 1971
indicated the precipitation was acid with an annual weighted average pH range of 4.03 to 4.19.
(A weighted average takes into account the amount of rain as well as its composition.)
Cogbill and Likens (1974), using precipitation from the Ithaca area, and Hubbard Brook
reported that their analysis of precipitation which consistently had a pH of less than 4.4
showed that 65 percent of the acidity was due to H?SO,, 30 percent to HNCL, and less than 5
percent was due to HC1.
In 1976, Likens reported that the continued monitoring of precipitation at the Hubbard
Brook Forest through 1974 indicated the mean annual pH for the years 1964-1974 ranged from
4.03 to 4.21. No statistically significant trend was noted; however, pH values of 2.1 and 3.0
were observed for individual storms at various locations. The increased deposition of hydrogen
ion was due to an increase in nitric acid in the precipitation (rain and snow) falling there.
This change in the composition of acidic precipitation suggests that the sources of nitrogen
oxide emissions increased while those for sulfur oxides remained constant.
The acidity of precipitation is a reflection of the free hydrogen ions in precipitation.
The contribution of sulfate and nitrate anions has changed with time, and analysis indicates
that the nitrate anion makes up an ever-increasing fraction of the total negative ion equiva-
lents. Following the reasoning of Granat (1972), Likens et al. (1976) found [assuming 2H per
2~ + 2-
SO. ion as in H,,SO. or one H ion per SO. as in (NH..2SO.] that the contribution of sulfate
to acidity declined from 83 to 66 percent of the total acidity between 1964 to 1974 at Hubbard
Brook, and the contribution of nitrate increased from 15 to 30 percent of the total during the
same period. Furthermore, increased annual input of H was closely correlated with increased
input of nitrate, but there was little correlation between H input and sulfate input.
Data for nitrate, ammonium, and sulfate in rain at Ithaca and Geneva, New York, consti-
tute the longest record of precipitation chemistry in the United States (Likens, 1972). Data
are available from 1915 to the present, but long gaps exist in the measurements, especially at
the Geneva site. Figures 7-8 (A) to (C) show that marked changes in composition have occurred
at Ithaca: a gradual decline in ammonium, an increase in nitrate beginning around 1945, and a
marked decrease in sulfate starting between 1945 and 1950. Early data for Ithaca showed higher
concentrations of sulfate in winter than in summer, presumably because of greater local burn-
ing of coal in winter. Data for 1971 showed the reverse trend, however, with nearly half the
annual sulfate input occurring during the months of June to August. Likens (1972) concluded
XDSX7B/A 7-21 2-9-81
-------
CO
I
UJ~
K
u.
_J
D
I I I
1920 1930 1940 1950 1%0 1970
YEAR
1920 1930 1940 1950 1960 1970
YEAR
1920 1930 1940 1950 1960 1970
YEAR
0.6
0.5
e
Z | 0.4
01
w I 0.3
= 0.2
Z
0.1
0
1968 1970 1972 1974
YEAR
Figure 7-8. Trends in mean annual concentrations of sulfate, ammonium, and nitrate
in precipitation. (A), (B), and (C) present long-term data for Ithaca, New York; (D)
presents data for eight yearsaaveraged over eight sites in New York and one in Pennsylvania.
One point in (A), for 1946--47, is not plotted because it is believed to be an anomaly.
Source: (A), (B), and (C) modified from Likens (1972); (D) modified from Likens (1976).
7-22
-------
that, despite deficiencies in the historical data and questions concerning their reliability,
the trends are real and can be explained by changes in fuel consumption patterns, i.e., natural
gas began to replace coal for home heating near the time of the shifts in precipitation
chemistry. On the basis of United States Geological Survey data for nine stations, Likens
(1976) reported a sharp increase in nitrate concentrations in New York state during the past
decade (Figure 7-8 [D]).
Data for eastern North America indicate a roughly three-fold increase in nitrate in rain-
fall since 1955, whereas sulfate in rain has roughly doubled in this period. According to
Nisbet (1975), sulfate/nitrate ratios in rainfall averaged about 4 in the eastern United States
in 1955-1956, but the average ratio had fallen to about 3 in 1972-1973. Nisbet calculated that
the fraction of H deposition attributable to nitrate rose from 19 percent in 1955-1956 to 24
percent in 1972-1973 (Table 7-2), while the deposition attributable to H2S04 decreased from 80
to 73 percent.
TABLE 7-2. DEPOSITION OF SULFURIC AND NITRIC ACIDS IN
PRECIPITATION IN EASTERN NORTH AMERICA
Percent
1955-56 1972-73 change 1956-73
Total deposition of
acid (as H )
Estimated deposition
as sulfuric acid
(percent of total)
expressed as sulfate
Estimated deposition
4.0a
3.2 (80)b
0.76 (19)b
10. 8a
7.9 (73)b
2.60 (24)b
+170
+150
+240
as nitric acid
(percent of total)
expressed as nitrate
Total deposition 16.4 31.8 +94
of sulfates
Sulfuric acid as % 19.7 24.8 +27
of sulfates
Deposition rates are expressed as multiples of the chemical equivalent
weight, so that raj.es for different chemical species can be compared
directly. 1 ton H is equivalent to 49 tons sulfuric acid or to 63
tons nitric acid.
A small but increasing fraction of the acid in precipitation is attribut-
able to hydrochloric acid.
Source: Modified from Nisbet (1975).
XDSX7B/A 7-23 2-9-81
-------
*
Lindberg et al. (1979) noted that SO2' and H+ were by far the dominant constituents of
precipitation at the Walker Branch Watershed, Tennessee. Comparison with the annual average
concentration of major elements in rain at the Walker Branch Watershed on an equivalent basis
indicated that H+ constitutes approximately 50 percent of the cationic strength and trace
elements account for only 0.2 percent. Sulfate constituted approximately 65 percent of the
anionic strength and on an equivalent basis was 3.5 times more concentrated than NO.,, the next
most abundant anion. The incident precipitation for the 2-year (1976-1977) period was des-
cribed as "a dilute mineral acid solution", primarily H2$04, at a pH approximating 4.2 and
containing relatively minor amounts of various trace salts (Lindberg et al., 1979). In
Florida, Hendrey (1977) and Hendrey et al. (1980) found that sulfate contributed 69 percent,
nitrate 23 percent, and chloride 8 percent of the free acidity in rainfall at Gainesville,
Florida, during 1976.
Based on most reports, sulfate (S02~) appears to be the predominant anion in acidic pre-
cipitation in the Eastern United States. In the west in California, however, nitrate (N03)
seems to predominate. Liljestrand and Morgan (1978) reported that their analyses of acidic
rainfall collected from February 1976 to September 1977 in the Pasedena, CA, area showed 'that
the volume-weighted mean pH was 4.0, with nitric acid being 32 percent more important as a
source of acidity than sulfuric acid. The major cations present were H , NH., K , Ca and
Mg2+ while the major anions were Cl~, N0~ and SO2". McColl and Bush (1978) also noted the
strong influence of nitrate on rain in the Berkeley, CA, region. However, they note that in
2-
bulk precipitation (wet plus dry fall-out) that sulfate (SO, ) constituted 50 percent of the
total anions.
Nearly all of the nitrate in rainfall is formed in the atmosphere from NO . Little is
/\
derived from wind erosion of nitrate salts in soils. Similarly, nearly all of the sulfate in
rainfall is formed in the atmosphere from S0? (National Research Council, 1978). Thus, all
atmospherically derived nitrate and sulfate contribute to the acidification of precipitation,
since H is associated stoichiometrical ly with the formation of each. A second stoichiometric
process that affects the acidity of rain is the reaction of nitric and sulfuric acids with
ammonia or other alkaline substances (e.g., dust particles) in the atmosphere to form neutral
nitrate and sulfate aerosols. To the extent that such neutralization occurs, the acidity of
precipitation will be reduced (National Research Council, 1978). However, since much of the
ammonium ion reaching soil is converted to nitrate, these neutral salts still have an acidi-
fying effect on the soil.
Seasonal fluctuations in composition as well as pH of rainfall have been reported by many
workers. In addition, the composition of rainfall and pH fluctuates from event to event, from
locality to locality, and from storm to storm.
2~ +
In general S04 and H concentrations in precipitation in the eastern United States are
higher in the summer than in the winter. Wolff et al. (1979) found this to be true for the
New York Metropolitan Area. Hornbeck et al. (1976) and Miller et al. (1978) both stated that
XDSX7B/A 7-24 2-9-81
-------
a summer maximum for sulfate was associated with an increase in hydrogen ion concentration in
upstate New York, the Hubbard Brook Experimental Forest in New Hampshire, and in portions of
Pennsylvania. Pack (1978), using data (1977) from the four original MAP3S (Multistate Atmos-
pheric Power Production Pollution Study) precipitation chemistry networks, plotted the
weighted monthly sulfate ion concentrations (Figure 7-9). Maximum sulfate concentrations
occurred from June through August. Lindberg et al. (1979), studying wetfal1 deposition of
2~ +
sulfate in the Walker Branch Watershed, also noted summer maxima for SCL and H . Using the
same MAP3S data as did Pack, they plotted weighted mean concentrations of sulfate in rain
collected from November 1976 through November 1977. The concentrations at Walker Branch
Watershed, Tennessee, are lower than all of the stations except remote Whiteface Mountain, New
York. The regional nature of the wet deposition of sulfate is apparent. Reasons for the
existence of the high summer maxima of sulfate for the eastern United States are discussed in
some detail in Chapter 5, Section 5.3.4.
Seasonal variations of nitrogen compounds and of pH in precipitation have been reported
by several workers, but no simple trends are apparent (see U.S. Environmental Protection
Agency Air Quality Criteria for Nitrogen Oxides, 1980). Hoeft et al. (1972) found relatively
constant levels of nitrate in rain and snow collected in Wisconsin throughout the year, but
deposition of ammonia and organic nitrogen was lowest in winter and highest in spring, perhaps
because of the thawing of frozen animal wastes. Haines (1976) reported large random varia-
tions, but relatively small seasonal variations, for nitrogen forms in wet-only precipitation
at Sapelo Island, Georgia; nitrogen concentrations were lowest during the rainy months of July
and September. The highest nitrogen loadings occurred during July and were asosciated with
the lowest range in pH, 4.2-4.8. Hendrey (1977) and Hendrey and Brezonik (1980) found rela-
tively smooth seasonal trends in ammonia and nitrate concentrations in both wet-only and bulk
collections (wet- and dryfall) at Gainesville, Florida, with lowest concentrations in winter
(Figure 7-10). In addition, the pH of the bulk precipitation showed no seasonal trend. Wet-
only collections, however, showed the lowest pH value (4.0) during the spring and summer. This
historical record suggests there has been an increase in the concentration of inorganic nitro-
gen in Florida over the past 20 years.
Scavenging by rainfall produces large changes in atmospheric contaminant concentrations
during a given rainfall event. The decline in constituent levels is usually rapid, at least
in localized convective showers, and low, steady-state concentrations are usually reached
within the first half hour of a rain event.
Major ions [chloride (Cl ) and sulfate (S0.~)], inorganic forms of nitrogen [nitrate
(NO., ) and ammonium (NH. )], total phosphorus and pH were measured in rain collected in
5-minute segments within three individual rainstorms. Initially, rapid decreases were
observed for nitrate and ammonium and total phosphorus. There was also a decrease in pH from
4.65 to 4.4. Steady state concentrations were reached in 10 minutes. Two other storms
sampled in the same manner showed similar but less defined patterns.
XDSX7B/A 7-25 2-9-81
-------
46
12
40
J 8
0*
in
• WBW
o WHITEFACE "
o ITHACA
A PENN STATE
O VIRGINIA
I I I I I II I
?
n
MAP3S
PRECIPITATION
NETWORK
6 —
Figure 7-9. Comparison of weighted mean monthly concentrations of sulfate
in incident precipitation collected in Walker Branch Watershed, Tenn. (WBW)
and four MAP3S precipitation chemistry monitoring stations in New York
(O. 0), Pennsylvania (A), and Virginia (D).
Source: Pack (1978).
7-26
-------
z
o
<
tr
o
O
o
4.80
4.60
4.40
4.20
0.40
0.30
0.20
I
0.10
M
A S O
1976
M A
- 1977 —
M
MONTH
Figure 7-10. Seasonal variations in pH (A) and ammonium and
nitrate concentrations (B) in wet-only precipitation at Gainesvil
Florida. Values are monthly volume-weighted averages of level;
in rain from individual storms.
7-27
-------
Wolf et al. (1979) examined spatial meteorological and seasonal factors associated with
the pH of precipitation in the New York Metropolitan Area. Seventy-two events were studied
from 1975 through 1977. There was some site-to-site variability among the eight sites they
studied in the Manhattan area (Table 7-3). They also noted that the pH varied according to
storm type (Table 7-4). Storms with a continental origin have a lower pH than storms origi-
nating over the ocean. The storms with trajectories from the south and southwest had the
lowest pH's, while those from the north and east had the highest pH's (Wolff et al., 1979).
The mean pH of precipitation falling on the New York Metropolitan Area during a 2-year
(1975 to 1977) study was 4.28; however, a pronounced seasonal variation was observed (Figure
7-11). The minimum pH at all sites except Manhattan occurred during July to September, while
the maximum occurred during October to December. The minimum pH in Manhattan, however,
occurred January to March and then gradually increased through the year. The lowest pH of
4.12 for the New York Metropolitan area occurred during the summer months (Wolff et al.,
1979). In general, the pH of rain is usually lower in the summer than in the winter and is
associated with the high summertime sulfate concentrations. In addition, the lowest pH's were
associated with cold fronts and air mass type precipitation events. These events occur, more
frequently during the summer months. The lower pH's also occurred on westerly or south-
westerly winds (Wolff et al., 1979).
Seasonal variations in pH measured at several sites in New York State 70 km (45 mi.) apart
demonstrated a significant difference between seasons (Winter had an average pH of 4.2; summer,
3.9.) but no significant difference between sites. In New Hampshire, however, six summer
storms sampled at 4 sites less than 3 km (2 mi.) apart showed a significant difference (3.8 to
4.2) indicating considerable variation in pH may occur in the same storm.
Stensland (1978, 1980) compared the precipitation chemistry for 1954 and 1977 at a site
in central Illinois. The pH for the 1954 samples had not been measured, but were calculated and
compared with those measured in 1977. The corrected pH for 1954 was 6.05; the pH for 1977 was
4.1. The more basic pH in 1954, according to the author, could have resulted from low levels
of acidic ions (e.g. sulfate or nitrate) or from high amounts of basic ions (e.g. calcium and
magnesium). Stensland suggests that the higher pH in 1954 was due to calcium (Ca ) and
magnesium (Mg ) ions from the soil.
7.2.3.2 Geographic Extent of Acidic Precipitation—Acidic precipitation has been a reality in
New York State for an undetermined period of time. Data collected by the United States Geo-
logical Survey (Harr and Coffey, 1975) over a ten-year period are presentedjn Figure 7-12.
These curves represent the pH of precipitation at eight different locations in New York State
and one location in Pennsylvania. Each of these locations (Figure 7-13) represents an area
within a given watershed. The pH of precipitation has remained nearly at the same general
average during the entire ten-year period; therefore, since data for the years prior to 1965
are lacking, it is difficult to determine when the pH in precipitation first began to decrease
(Harr and Coffey, 1975).
XDSX7B/A 7-28 2-9-81
-------
TABLE 7-3. MEAN pH VALUES IN THE NEW YORK METROPOLITAN
AREA (1975-1977)
Site
Caldwell, N.J.
Piscataway, N.J.
Cranford, N.J.
Bronx, N.Y.
Manhattan, N.Y.
High Point, N.J.
Queens, N.Y.
Port Chester, N.Y.
All sites
Mean pH
4.32
4.25
4.34
4.31
4.29
4.25
4.63
4.60
4.28
SD
0.26
0.36
0.34
0.37
0.25
0.30
0.35
0.19
0.32
No. obsd
50
64
48
57
39
25
20
21
72
Range
3.35-5.60
3.57-5.50
3.44-5.95
3.42-5.75
3.80-5.50
3.74-4.90
3.98-5.28
4.00-5.10
3.50-5.16
From Wolff, et al., 1979
TABLE 7-4. STORM TYPE CLASSIFICATION
Type
Description of dominant storm
system
No.
obsd
Mean
pH
1
2
3
4
5
6
7
8
Closed low-pressure system which formed
over continental N. Amer.
Closed low-pressure system which formed in
Gulf of Mexico or over Atlantic Ocean
Closed low which passed to W or N of N.Y.C.
Closed low which passed to S or E of N.Y.C.
Cold front in absence of closed low
Air mass thunderstorm
Hurricane Bel le
Unclassified
22
21
26
17
16
5
1
6
4.35
4.43
4.39
4.39
4.17
3.91
5.16
4.31
From Wolff et al., 1979.
XDSX7B/A 7-29 2-9-81
-------
4.6
4.5
4.4
4.3
4.2
4.1
4.0
JFM AMJ JAS OND
MONTHS OF THE YEAR (1975 THROUGH 1977)
Figure 7-11. Seasonal variation of precipitation pH in the New York
Metropolitan Area.
7-30
-------
7.0
6.0
6.0
4.0
3.0
0.0*
ALBANY, NEW YORK
ni iiii
7.0
6.0
"
4.0
3.0
ao
6.0
6.0
4.0
,<^
ALLEGHENY STATE PARK. NEW YORK
3.0
0.0
6.0
5.0
4.0
0.0*
,-CL
it
ATHENS. PENNSYLVANIA
CANTON. NEW YORK
niniiii ill mini
Ji
J-U-t
1065
1966 1967
1968 1969 1970 1971 1972
YEAR
1973
Figure 7-12. History of acidic precipitation at various sites in and adjacent to State of New York.
7-31
-------
7.0
6.0
6.0
4.0
O.OUJ
HINCK LEY, NEW YORK,
MAYS POINT, NEW YORK
7.0
6.0
6.0
4.0
0.0^
7.0
6.0
5.0
4.0
fl.0^
MINEOLA. NEW YORK
6.0
5.0
4.0
0.0 I
ROCK HILL, NEW YORK
UPTON, NEW YORK
1065
1973
Figure 7-12 (continued). History of acidic precipitation at various sites in and adjacent to
State of New York.
7-32
-------
1. Albany, N.Y.
2. Allegheny State Park, N.Y.
3. Athens, Pa.
4. Canton, N.Y.
5. Hinckley, N.Y.
6. Mays Point, N.Y.
7. Mineola, N.Y.
8. Rock Hill, N.Y.
9. Upton, N.Y.
SYRACUSE
BUFFALO 6 •
ALBAN
1
BINGHAMTON
SCALE
j 3
0 50 100
MILES
Figure 7-13. Location of acidic precipitation monitoring stations.
7-33
-------
That precipitation is acidic in parts of the country other than the northeastern United
States is apparent. Average pH values around 4.5 have been reported as far south as northern
Florida (Likens, 1976; Hendry and Brezonik, 1980), from Illinois (Irving, 1978), the Denver
area of Colorado (Lewis and Grant, 1980), the San Francisco Bay area of California (McColl and
Bush, 1978; Williams, 1978), Pasadena, California (Liljestrand and Morgan, 1978), the Puget
Sound area of Washington (Larsen et al. , 1976), and from eastern Canada (Glass et al., 1979;
Dillon et al., 1978). Data from the San Francisco Bay area indicate that precipitation has
become more acidic in that region since 1957-1958 (McColl and Bush, 1978). The pH decreased
from 5.9 during 1957-1958 to 4.0 in 1974, and seems to be related to an increase in the N03
concentration (McColl and Bush, 1978). Another report, using data from the California Air
Resources Board (CARB) (Williams, 1978), states that acidic precipitation has been reported
from such widespread areas as Pasadena, Palo Alto, Davis, and Lake Tahoe.
Studies in the Great Smoky Mountain National Park (Lipske, 1980) indicate a downward
trend in pH has occurred there over the past twenty years. Over a period of 20 years, there
has been a drop in pH from a range of 5.3-5.6 to 4.3 in 1979.
The absence of a precipitation monitoring network throughout the United States in the
past makes determination of trends in pH extremely difficult and controversial. This short-
coming has been rectified recently through the establishment of the National Atmospheric
Deposition Program. Under the program, monitoring stations collect precipitation samples,
determine their pH and then send the samples to a Central Analytical Laboratory in Illinois to
be analyzed. This long-term network plans to have 75 to 100 collection sites throughout the
United States; 74 are already operational.
7.2.4 Acidic Deposition
The previous sections of this chapter have discussed the formation, composition and geo-
graphic distribution of acidic precipitation. Usually when the effects of acidic deposition are
discussed, emphasis is placed on the effects resulting from the scavenging of sulfur and
nitrogen compounds by precipitation. Dry deposition of gaseous and particulate and aerosol
forms of these compounds also occurs and is beginning to receive more emphasis in research
(Galloway and Whelpdale, 1980; Schlesinger and Hasey, 1980; Stensland, 1980; Schmel, 1980;
Chamberlain, 1980). Gaseous compounds reach the surface of the earth by turbulent transfer
while particulate sulfates and nitrates reach the earth's surface by gravitational sedimenta-
tion, turbulent transfer and impaction (Galloway and Whelpdale, 1980; Schmel, 1980; Hicks and
Wesely, 1980). A comparison of the relative significance of wet and dry deposition is diffi-
cult. Dry deposition, however, is always removing pollutants from the atmosphere, while
removal by wet deposition is intermittent (Schmel, 1980). Marenco and Fontan (1974) suggest
that dry deposition is more important than wet in removing air pollutants from manmade sources.
Lindberg et al. (1979) have calculated the annual mass transfer rates of sulfates to the
forest floor in Tennessee (Figure 7-14). Their calculations for S0.= suggest wet deposition
by incident precipitation to be 27 percent compared with a total dry precipitation of 13
XDSX7B/A 7-34
2-9-81
-------
400
IN CLOUD
PRECIPITATION
SCAVENGING
25%
BELOW CLOUD
PRECIPITATION
SCAVENGING
2%
TOTAL DRY
DEPOSITION
13%
TO GROUND
(DORMANT PERIOD)
2%
'.«' '" •» • i
« '."i i U«»»H'I '
INCIDENT PRECIPITATION
I i
TO LEAFY
CANOPY
10%
, 'ZTK'I'I" •
'"*'.{'•'•
TO BRANCHES
(DORMANT PERIOD)
1%
INTERNAL EXTERNAL
FLUX /
I m /
100%
RELATIVE ANNUAL MASS TRANSFER RATES
OF SOj-S TO THE FOREST FLOOR
Figure 7-14. Annual mass transfer rates of sulfate expressed as a percentage of the estimated
total annual flux of the element to the forest floor beneath a representative chestnut oak stand.
Source: Lindberg et al. (1979).
7-35
-------
percent. The dry deposition and foliar absorption of S0?, a very important component, is
missing from this calculation. The wet and dry deposition percentages are only an indication
of the relative magnitude of the two processes. The percentages do, however, point out that
the effects of acidic deposition usually attributed to precipitation scavenging alone are
probably a result of both wet and dry deposition. At the present time the accuracy with which
dry deposition can be measured is still under question.
The studies of McColl and Bush (1978), Hendrey and Brezonik (1980), and Schlesinger and
Hasey (1980) also point out that both wet and dry deposition are important when considering
the effects of H+, SO?", and NOl ions on aquatic and terrestrial receptors.
The effects of the dry deposition of S02 and particulate matter on vegetation and terres-
trial ecosystems is discussed in Chapter 8. The processes of wet and dry deposition of sulfur
oxides are discussed in Chapter 6 of this document; for nitrogen oxides in Chapter 6 of Air
Quality Criteria for Nitrogen Oxides.
7.3 EFFECTS OF ACIDIC DEPOSITION
Acidic precipitation has been implicated in the degradation of aquatic ecosystems, the
disintegration of stone buildings and monuments and as a potential source of harm to forests
and other terrestrial ecosystems. The sections that follow discuss these effects.
7.3.1 Aquatic Ecosystems—Acidification of surface waters is a major problem in regions of
southern Scandinavia (Oden, 1968; Aimer et al., 1974; Gjessing et al., 1976), Scotland (Wright
et al., 1980a), eastern Canada (Beamish and Harvey, 1972; Dillon et al., 1978), and the
eastern United States - in the Adirondack Region of New York State (Schofield, 1976; Pfeiffer
and Festa, 1980), in Maine (Davis et al., 1978), and in northern Florida (Crisman et al.,
1980). Damage to fisheries is the most obvious affect of acidification on freshwater life.
The disappearance of fish populations from acidified freshwater lakes and streams was first
noted in southern Norway in the 1920's. In 1959, Dannevig proposed that acidic deposition was
the probable cause for acidification and thus far the loss of fish populations (Leivestad et
al., 1976). Subsequent studies have verified this postulate. Declines in fish populations
have been related to acidification of surface waters in southern Norway (Jensen and Snekvik,
1972; Wright and Snekvik, 1978), southwestern Sweden (Aimer et al., 1974), southwestern
Scotland (Wright et al., 1980a), the Adirondack Region of New York State (Schofield, 1976),
and the LaCloche Mountain Region in southern Ontario (Beamish and Harvey, 1972). Acidifica-
tion may also have serious repercussions on other aquatic biota inhabiting these systems.
Changes in the acidity and chemistry of freshwater affect the communities of organisms living
there. Pertinent details of these effects are described in the following sections.
7.3.1.1 Acidification of lakes and streams. Precipitation enters lakes directly as rain or
snow or indirectly as runoff of seepage water from the surrounding watershed. The relative
magnitude of the influents from these two sources is dependent on the surface area and volume
of the lake, and the size of the watershed and its soil volume and type. In general, the
watershed plays a dominant role in determining the composition of water entering the lake. As
XDSX7B/A 7-36 2-9-81
-------
*
a result, the water will be strongly influenced by processes in the edaphic environment of the
watershed, such as weathering, ion exchange, uptake and release of ions by plants, carbon
dioxide production by vegetation, microbial respiration, and reduction and oxidation reactions
of sulfur and nitrogen compounds (Seip, 1980). Precipitation as a direct source of water to
the lake plays a relatively greater role when lake areas are large in comparison to the size
of the watershed.
Acidification of surface waters results when the sources of hydrogen ion exceed the
ability of an ecosystem to neutralize the hydrogen ion. In general, the soils and crust of
the earth are composed principally of basic materials with large capacities to buffer acids.
However, areas where bedrock is particularly resistant to weathering and soils are thin and
poorly developed have much less neutralizing ability. This inability to neutralize hydrogen
ion does not arise from a limited soil or mineral buffering capacity. Instead low cation
exchange capacity and slow mineral dissolution rates in relation to the relatively short
retention time of water within the soil system may result in incomplete neutralization of soil
waters and acidification of surface waters (Driscoll, 1980). Characteristics of regions
sensitive to surface water acidification are discussed in more detail in Section 7.4.1.
Sources of hydrogen ion to the edaphic-aquatic system include, besides acidic deposition,
mechanisms for internal generation of hydrogen ion - oxidation reactions (e.g., pyrite oxida-
tion, nutrification), cation uptake by vegetation (e.g., uptake of NH. or Ca ), or genera-
tion of organic acids from incomplete organic litter decomposition (Figure 7-15). The rela-
tive importance of the hydrogen ion content in acidic deposition to the overall hydrogen ion
budget of an ecosystem has been discussed by many researchers (Rosenqvist, 1976; SNSF Project,
1977).
The consensus is that changes in internal hydrogen ion generation related to land use or
other changes (e.g., Drablos and Sevaldrud, 1980) can not consistently account for the wide-
spread acidification of surface waters occurring in southern Scandinavia, the Adirondack
Region of New York, the LaCloche Mountain Area of Ontario, and elsewhere. Driscoll (1980)
developed a hydrogen ion budget for the Hubbard Brook Area in New Hampshire. Based on these
calculations, atmospheric hydrogen ion sources represent 48 percent of the total Hubbard Brook
ecosystem hydrogen ion sources.
As noted above, freshwater ecosystem sensitive to inputs of acids are generally in areas
of poor soil development and underlain by highly siliceous types of bedrock resistant to dis-
solution through weathering (Likens et al., 1979). As a result, surface waters in such areas
typically contain very low concentrations of ions derived from weathering. The waters are
diluted with low levels of dissolved salts and inorganic carbon, and low in acid neutralizing
capacity. The chemical composition of acid lakes is summarized in Table 7-5 for lakes in
southern Norway (Gjessing, et al., 1976), the west coast (Hb'rnstrom et al., 1976) and west-
central (Grahn et al., 1976) regions of Sweden, the LaCloche Mountains of southeastern Ontario
(Beamish, 1976), and the vicinity of Sudbury, Ontario (Scheider et al., 1976), as well as for
XDSX7B/A 7-37 2-9-81
-------
ALLOCHTHONOUS SOURCES OF HYDROGEN ION
PRECIPITATION.
DRY DEPOSITION.
DRAINAGE WATER
• ECOSYSTEM BOUNDARY
HYDROGEN ION
SOURCES
OXIDATION RxN
CATION UPTAKE
PYRITE
OXIDATION
NH4+ UPTAKE
HYDROGEN ION
SINKS
REDUCTION RxN
ANION UPTAKE
OXIDE
WEATHERING
STREAM EXPORTS
H+, HCOj. OH-LIGANDS.
ORGANIC ANIONS
Figure 7-15. Schematic representation of the hydrogen ion
cycle.
Source: Driscoll (1980).
7-38
-------
TABLE 7-5. CHEMICAL COMPOSITION (MEAN i STANDARD DEVIATION) OF ACID LAKES (pH <5) IN REGIONS RECEIVING HIGHLY
ACIDIC PRECIPITATION (pH <4.5), AND OF SOFT-WATER LAKES IN AREAS NOT SUBJECT TO HIGHLY ACIDIC PRECIPITATION
(pH >4.8)
Region
I. LAKES IN ACID AREAS
Scandinavia
^J Southernmost
' Norway
CO
^o Westcoast
Sweden
West-central
Sweden
North America
La Cloche Mtns,
Ontario
Sudbury,
Ontario
II. LAKES IN UNAFFECTED
Scandinavia
West-central
Norway
North America
Experimental
Lakes Area,
Ontario
No. of
Specific
lakes conductancett H (pH)
Measured: 26
Less s w*:
Measured: 12
Less s w*:
Measured: 6
Less s w*:
Measured: 4
Less s w*:
Measured: 4
Less s w*:
AREAS
Measured: 23
Less s w*:
Measured: 40
Less s w*:
27+10 18111 (4.76)
18
•70** 43** fd 37**^
43**
47+23 22+15 (4.66)
22
38+8 2019 (4.7)
20
120140 3615 (4.5)
36
13+3 612 (5.2)
6
19 0.2-2 (5.6-6.7)
0.2-2
Na
70+40
9
-50
1651120
20
2614
9
100130
50
50120
9
40
4
K
513
4
9(1
£.V
13
1518
12
10+3
10
40+10
40
3+1
3
10
10
Ca
56+35
50
-
75+10
70
150125
150
4501180
450
1819
16
80
80
Mg
41116
25
iftft
80140
50
7518
65
3101120
300
1615
7
75
65
ueq/1
HCOa
1H26
11
0
-
-
0
0
812
8
1318
13
60
60
HC1
711 45
0
0
170+90
0
22+6
0
50i20t
0
46121
0
40
0
S0«
100133
92
155
200170
180
290140
290
800+290
800
3318
30
60
55
N03
412
4
8
1914
19
-
-
-
-
512
5
<1. 5
<1.5
1 cations I anions Reference
189
106
-
360
175
.280
255
940
880
93
41
200
160
186
107
-
390
200
310
290
850
800
97
48
160
120
Gjessing
et al . , 1976
Hornstrb'm
et al . , 1976
Grahn
et al., 1976
Beamish,
1976
Armstrong,
1971
Gjessing
et al. , 1976
Armstrong,
1971
•Less s w = Concentrations after subtracting the seawater contribution according to the procedure explained by
Wright and Gjessing (1976).
**Data for 112 lakes
tMeasured after past liming of the lakes
ttuS/cm at 20°C
-------
lakes not yet affected by acidification but in regions of similar geological substrata in west-
central Norway (Gjessing et al., 1976) and the experimental lakes area of northwestern Ontario
(Armstrong and Schindler, 1971). Basic cation concentrations (Ca, Mg, Na, K) are low (e.g.,
calcium levels of 18-450 peg/liter or 0.4 - 9 mg/liter) relative to world-wide averages (15
mg/liter calcium, Livingstone, 1963). Bicarbonate is the predominant anion in most fresh-
waters (Stumm and Morgan, 1970). However, in acid lakes in regions affected by acidic de-
position, sulfate replaces bicarbonate as the dominant anion (Wright and Gjessing, 1976;
Beamish, 1976). With a decreasing pH level in acid lakes, the importance of the hydrogen ion
to the total cation content increases.
Surveys to determine the extent of effects of acidic deposition on the chemistry of lakes
have been conducted in Norway (Wright and Snekvik, 1978; Wright and Henriksen, 1978), Sweden
(Aimer et al., 1974; Dickson, 1975), Scotland (Wright et al., 1980a), the LaCloche Mountain
area of Ontario (Beamish and Harvey, 1972), the Muskoka-Haliburton Area of south-central
Ontario (Dillon et al., 1978), and the Adirondack Region of New York State (Schofield, 1976b),
Maine (Davis et al., 1978), and Pennsylvania (Arnold et al., 1980). In regions of similar
geological substrata not receiving acidic deposition, lake pH levels average 5.6-6.7
(Armstrong and Schindler, 1971). Of 155 lakes systematically surveyed in southern Norway in
October 1974, over 70 percent had pH levels below 6.0, 56 percent below 5.5, and 24 percent
below 5.0 (Wright and Henriksen, 1978). Of 700 lakes in the Stfrlandet Region of southern
Norway surveyed in 1974 to 1975 (May-November), 65 percent had pH levels below 5.0 (Wright and
Snekvik, 1978). On the west coast of Sweden, of 321 lakes investigated during 1968-1970, 93
percent had a pH level 5.5 or lower. Fifty-three percent had pH levels between 4.0 and 4.5
(Dickson, 1975). In the LaCloche Mountain Region of Ontario, 47 percent of 150 lakes sampled
in 1971 had pH levels less than 5.5, and 22 percent had pH levels below 4.5 (Beamish and
Harvey, 1972). In the Adirondacks, 52 percent of the high elevation (> 610 m) lakes had pH
values below 5.0 (Schofield, 1976b). In each of these studies, the pH level of an individual
lake could be related to, in most cases, the intensity of the acidic deposition and the geo-
logic environment of the watershed. Atmospheric contributions of sea salts are also important
in coastal regions.
Several methods have been developed to assess the degree of acidification in a lake and
relate it to inputs of hydrogen ion or sulfate. Henriksen (1979) utilized alkalinity-calcium
and pH-calcium relationships in lakes to estimate the degree of acidification experienced by a
surface water. This technique is based on the premise that when carbonic acid weathering
occurs one equivalent of alkalinity (acid neutralizing capacity) is released to the aquatic
environment for every equivalent of basic cation (Ca, Mg, K, or Na) dissolved. On the other
hand, if mineral acid weathering is occurring, for example as a result of acidic deposition,
one equivalent of hydrogen ion is comsumed for every equivalent of cation solubilized. There-
fore, for a given basic cation level, there is less aqueous acid neutralizing capacity in
lakes in systems experiencing strong acid weathering than in systems experiencing carbonic
XDSX7B/A 7-40 2-9-81
-------
acid weathering. When comparing alkalinity plots from two watersheds, one experiencing strong
acid contributions and the other undergoing largely carbonic acid weathering (assuming both
watersheds have similar edaphic environments), the difference in alkalinity between the two
plots for a given calcium level (the dominant basic cation) should be indicative of the amount
of strong acid the watershed receives and the degree of acidification of the surface water.
For waters with pH levels below 5.6, alkalinity is approximately equal to the negative of the
hydrogen ion concentration. Therefore, pH level can be substituted for alkalinity, and pH-
calcium plots developed (Figure 7-16). Data of this type for Norway indicate that significant
acidification of lakes has occurred in areas receiving precipitation with volume-weighted
average concentrations of H+ above 20-25 (jeq/liter (pH 4.7-4.6) and sulfate concentrations
above 1 nig/liter (20 peg/liter) (Henriksen, 1979).
Henriksen (1979) also utilized the concentration of excess sulfate in lake water (sulfate
in excess of that of marine origin) to estimate acidification. This suggests that bicarbonate
anions lost in acidified lakes have been replaced by an equivalent amount of sulfate. Aimer
et al. (1978) plot pH levels in Swedish lakes as a function of excess sulfur load (excess
sulfur in lake water multiplied by the yearly runoff) (Figure 7-17). Based on this relation-
ship, they estimate that the most sensitive lakes in Sweden may resist a load of only about
2 2
0.3 g/m of sulfur in lake water each year. At 1 g/m of sulfur, the pH level of the lake
will probably decrease below 5.0.
Elevated metal concentrations (e.g., aluminum, zinc, manganese, and/or iron) in surface
waters are often associated with acidification (Schofield and Trojnar, 1980; Hutchinson et
al., 1978; Wright and Gjessing, 1976; Beamish, 1976). Mobility of all these metals is in-
creased at low pH values (Stumm and Morgan, 1970). For example, an inverse correlation
between aluminum concentration and pH level has been identified for lakes in the Adirondack
Region of New York State, southern Norway, the west coast of Sweden, and Scotland (Wright,
1980b) (Figure 7-18). Aluminum appears to be the primary element mobilized by strong acid
inputs in precipitation and dry deposition (Cronan, 1978).
Aluminum is the third most abundant element by weight in the earth's crust (Foster,
1971). In general, aluminum is extremely insoluble and retained within the edaphic environ-
ment. However, with increased hydrogen ion inputs (via acidic deposition or other sources)
into the edaphic environment, aluminum is rapidly mobilized. Cronan and Schofield (1979)
suggest that input of strong acids may inhibit the historical trend of aluminum accumulation
in the B soil horizon. Consequently, aluminum tends to be transported through the soil pro-
file and into streams and lakes. Evidence from field data (Schofield and Trojnar, 1980) and
laboratory experiments (Driscoll et al., 1979; Muniz and Leivestad, 1980) suggest that these
elevated aluminum levels may be toxic to fish. Concentration of aluminum may be as or more
important than pH levels as a factor leading to declining fish populations in acidified lakes.
Aluminum toxicity to aquatic biota other than fish has not been assessed.
XDSX7B/A 7-41 2-9-81
-------
6.0
6.5
7.0
7.5
1
50
2
100
3
150
4
200
5 (mgV1)
250
-------
o.
o
o
CURVE 2
0123
EXCESS S IN LAKE WATER, g/m2/yaar
Figure 7-17. The pH value 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.)
Source: Aimer et al. (1978).
7-43
-------
1000
5 ioo
10
I
1000
1 100
10
t
- "T T T -
SOUTH NORWAY 1974
154 LAKES
x •* "• ' .*
. «• • • •
• . . V" < •
" . *
1 1 1
1 5 6 7 f
pH
— 1 1 1 —
WEST COAST SWEDEN
37 LAKES
. •*
*
• .
— :'. .: m —
• • • • * •
• •
1 1 ' ' 1
t 5 6 7 (
PH
1000
"5>
5 100
10
i '
1000
5 ioo
10
i t
-I I I -
SCOTLAND
72 LAKES
•
*". * " •• •
• •
• . ' :•'
•
I I I '
15678
pH
' .1 1 1
* ADIRONDACKSUSA
• 134 LAKES
.Y& •.*
~~ *i • ' ' * ' ~~
* * * *
• • • *
• ••
• • • •
• •
* •
•
1 I.I
• 5 6 7 I
PH
Figure 7-18. Total dissolved Al as a function of pH level in lakes in acidified areas ih Europe and
North America.
Source: Wright et al. (1980b).
7-44
-------
*
Surface water chemistry, particularly in streams and rivers, may be highly variable with
time. Since many of the neutralization reactions in soils are kinetically slow, the quality
of the leachate from the edaphic system into the aquatic system varies with the retention time
of water in the soil (Johnson et al., 1969). The longer the contact period of water with lower
soil strata, the greater the neutralization of acid contribution from precipitation and dry
deposition. Therefore, during periods of heavy rainfall or snowmelt, and rapid water dis-
charge, pH levels in receiving waters may be relatively depressed.
Many of the regions currently affected by acidification experience freezing temperatures
during the winter and accumulation of a snowpack. In the Adirondack Region of New York approxi-
mately 55 percent of the annual precipitation occurs during the winter months (Schofield, 1976b).
Much of the acid load deposited in winter accumulates in the snowpack, and may be released during
a relatively short time period during snowmelt in the spring. In addition, on melting, 50 to
80 percent of the pollutant load (including hydrogen ion and sulfate) may be released in the
first 30 percent of the meltwater (Johannessen and Henriksen, 1978). As a result, melting of
the snowpack and ice cover can result in a large influx of acidic pollutants into lakes and
streams (Figure 7-19) (Gjessing et al., 1976; Schofield and Trojnar, 1980; Hultberg, 1976).
The rapid flux of this meltwater through the edaphic environment, and its interaction with only
upper soil horizons, limits neutralization of the acid content. As a result, surface waters
only moderately acidic during most of the year may experience extreme drops in pH level during
the spring thaw. Basic cation concentrations (Ca, Mg, Na, K) may also be lower during this
time period (Johannessen et al., 1980). Similar but usually less drastic pH drops in surface
waters (particularly streams) may occur during extended periods of heavy rainfall (Driscoll,
1980). These short term changes in water chemistry may have significant impacts on aquatic
biota.
7.3.1.2 Effects on decomposition. The processing of dead organic matter (detritus) plays a
central role in the energetics of lake and stream ecosystems (Wetzel, 1975). The organic matter
may have been generated either internally (autochthonous) via photosynthesis within the aquatic
ecosystem or produced outside the lake or stream (allochthonous) and later exported to the aquatic
system. Detritus is an important food source for bacteria, fungi, some protozoa, and other
animals. These organisms through the utilization of detritus release energy, minerals and
other compounds stored in the organic matter back into the environment. Initial processing of
coarse particulate detritus is often accomplished by benthic invertebrate fauna. Among other
things, the particles are physically broken down into smaller units, increasing their surface
area. Biochemical transformations of particulate and dissolved organic matter occur via
microbial metabolism and are fundamental to the dynamics of nutrient cycling and energy flux
within the aquatic ecosystem.
In general, the growth and reproduction of microorganisms is greatly affected by hydrogen
ion concentration (Rheinheimer, 1971). Many bacteria can grow only within the range pH 4-9
and the optimum for most aquatic bacteria is between pH 6.5 and 8.5. There are more acidi-
philic fungi than bacteria; consequently in acid waters and sediments the proportion of fungi
XDSX7B/A 7-45 2-9-81
-------
I 6
F
1976/77
M
Figure 7-19. pH levels in Little Moose Lake, Adirondack region of New York State, at a depth of 3
meters and at the lake outlet.
Source: Adapted from Schofield and Trojnar (1980).
7-46
-------
in the microflora is greater than in waters or sediments with neutral or slightly alkaline pH
levels. Most aquatic fungi require free oxygen for growth (Rheinheimer, 1971).
Numerous studies have indicated that acidification of surface waters results in a shift
in microbial species and a reduction in microbial activity and decomposition rates. It should
be noted, however, that microorganisms in general are highly adaptive. Given sufficient time,
a given species may adapt to acid conditions or an acid-tolerant species may invade and colo-
nize acidified surface waters. Therefore, some caution is necessary in interpreting short-
term experiments on the effects of acidification on microbial activity and decomposition. On
the other hand, increased accumulations of dead organic matter (as a result of decreased
decomposition rates) are commonly noted in acidic lakes and streams.
Abnormal accumulations of coarse organic matter have been observed on the bottoms of six
Swedish lakes. The pH levels in these lakes in July 1973 were approximately 4.4 to 5.4. Over
the last three to four decades, pH levels appear to have decreased 1.4 to 1.7 pH units (Grahn
et al., 1974). In both Sweden and Canada, acidified lakes have been treated with alkaline
substances to raise pH levels. One result of this treatment has been an acceleration of
organic decomposition processes and elimination of excess accumulations of detritus (Andersson
et al., 1978; Scheider et al., 1975). Litterbags containing coarse particulate detrital mat-
ter have been used to monitor decomposition rates in acidified lakes and streams. In general,
the rates of weight loss were reduced in acidic waters when compared with more neutral waters
(Leivestad et al., 1976; Traaen and Laake, 1980; Petersen, 1980). Traaen and Laake (1980)
found that after 12 months of incubation dried birch leaves or aspen sticks showed a weight
loss of 50-80 percent in waters with pH levels 6 to 7 as compared to only a 30-50 percent
weight loss in waters with pH 4 to 5. Petersen (1980) likewise found reduced weight loss of
leaf packs incubated in an acidic stream when compared to leaf packs in either a stream not
affected by acidification or a stream neutralized with addition of lime. Petersen, however,
found no evidence of differences in microbial respiration between the streams. The acidic
stream did show a reduction in the invertebrate functional group that specializes in process-
ing large particles (shredders). Andersson et al. (in press) found no significant differences
in oxygen consumption by sediments from acidified and non-acidified lakes. Rates of glucose
decomposition were also studied in lake sediment-water systems adapted to pH values from 3 to
9. Glucose transformation increased at pH levels above 6. Lime treatment of acidic Lake
Hogsjon in Sweden also increased rates of glucose processing. However in a humic lake, the
maximum rate of glucose transformation occurred at the ui situ value pH 5 (Andersson et al.,
in press).
Laboratory and field experiments involving decomposition rates have fairly consistently
found decreasing microbial activity with increasing acidity. Traaen and Laake (1980) found
that litter decomposition at pH level 5.2 was only 50 percent of that at pH 7.0 and at pH 3.5,
only 30 percent that at pH 7.0. In addition, increasing acidity (pH 7.0 to 3.5) led to a
shift from bacterial to fungal dominance. Incubations of profundal lake sediments at pH 4, 5,
XDSX7B/A 7-47 2-9-81
-------
and 6 indicated a significant reduction in community respiration with increasing acidity, as
well as a possible inhibition of nitrification and a lowering of sediment redox potentials.
Bick and Drews (1973) studied the decomposition of peptone in the laboratory. With decreasing
pH, total bacterial cell counts and numbers of species of ciliated protozoans decreased, de-
composition and nitrification were reduced and oxidation of ammonia ceased below pH 5. At pH
4 and lower, the number of fungi increased.
Disruption of the detrital trophic structure and the resultant interference with nutrient
and energy cycling within the aquatic ecosystem may be one of the major consequences of acidifi-
cation. Investigations into the effects of acidification on decomposition have, apparently,
produced somewhat inconsistent results. However, many of these apparent inconsistencies arise
only from a lack of complete understanding of the mechanisms relating acidity and rates of
decomposition. It is fairly clear that in acidic lakes and streams unusually large accumula-
tions of detritus occur, and these accumulations are related, directly or indirectly, to the
low pH level. The processing of organic matter has been reduced. In addition, this accumula-
tion of organic debris plus the development of extensive mats of filamentous algae on lake
bottoms (discussed in Section 7.3.1.3) may effectively seal off the mineral sediments from
interactions with the overlying water. As a result, regeneration of nutrient supplies to the
water column is reduced both by reduced processing and mineralization of dead organic matter
and by limiting sediment-water interactions. Primary productivity within the aquatic system
may be substantially reduced as a result of this process (Section 7.3.1.3). These ideas have
been formulated into the hypothesis of "self-accelerating oligotrophication" by Grahn et al.
(1974).
7.3.1.3 Effect on primary producers and primary productivity. Organisms obtain their food
(energy) directly or indirectly from solar energy. Sunlight, carbon dioxide, and water are
used by primary producers (phytoplankton, other algae, mosses, and macrophytes) in the process
of photosynthesis to form sugars which are used by the plants or stored as starch. The stored
energy may be used by the plants or pass through the food chain or web. Energy in any food
chain or web passes through several trophic levels. Each link in the food chain is termed a
trophic level. The major trophic levels are the primary producers, herbivores, carnivores,
and the decomposers. Energy in an ecosystem moves primarily along two main pathways: the
grazing food chain (primary producers-herbivores-carnivores) and the detrital food chain
(Smith, 1980; Billings, 1978; Odum, 1971). Interactions between these two food chains are,
however, extensive. Green plants convert solar energy to organic matter and, as such, are the
base for both food chains. The grazing food chain involves primarily living ^organic matter;
the detrital food chain, dead organic matter. Any changes as a result of acidification in the
green plants and primary production within the aquatic ecosystem may therefore have a profound
effect on all other organisms in the aquatic food web. As noted in Section 7.3.1.2, a portion
of the detrital food chain is supported by dead organic matter imported into the aquatic-
system from external sources.
XDSX7B/A 7-48 2-9-81
-------
*
Extensive surveys of acidic lakes in Norway and Sweden (Leivestad et al., 1976; Aimer et
al., 1974) have observed changes in species composition and reduced diversity of phytoplankton
correlated with decreasing lake pH level (Figure 7-20). Generally at normal pH values of 6 to
8, lakes in the west coast region of Sweden contain 30 to 80 species of phytoplankton per
100-ml sample in mid-August. Lakes with pH below 5 were found to have only about a dozen
species. In some very acid lakes (pH<4), only three species were noted. The greatest changes
in species composition occurred in the pH interval 5-6. The most striking change was the dis-
appearance of many diatoms and blue-green algae. The families Chlorophyceae (green algae) and
Chrysophyceae (golden-brown algae) also had greatly reduced numbers of species in acidic lakes
(Figure 7-21). Dinoflagellates constituted the bulk of the phytoplankton biomass in the most
acidic lakes (Aimer et al., 1978). Similar phenomena were observed in a regional survey of 55
lakes in southern Norway (Leivestad et al., 1976) and in a study of nine lakes in Ontario
(Stokes, 1980). Changes in species composition and reduced diversity have also been noted in
communities of attached algae (periphyton) (Leivestad et al., 1976; Aimer et al., 1978).
Mougeotia, a green algae, often proliferates on substrates in acidic streams and lakes.
Shifts in the types and numbers of species present may or may not affect the total levels
of primary productivity and algal biomass in acidic lakes. Species favored by acidic condi-
tions may or may not have comparable photosynthetic efficiencies or desirability as a prey
item for herbivores. On the other hand, decreased availability of nutrients in acidic water
as a result of reduced rates of decomposition (Section 7.3.1.2) may decrease primary produc-
tivity regardless of algal species involved. In field surveys and experiments, relationships
between pH level and total algal biomass and/or productivity were not as consistent as the
relationship between pH and species diversity.
Kwiatkowski and Roff (1976) identified a significant linear relationship of decreasing
chlorophyll a concentrations (indicative of algal biomass) with declining pH level in six
lakes near Sudbury, Ontario, with a pH range of 4.05 to 7.15. In addition, primary productiv-
ity was reduced in the two most acid lakes (pH 4-4.6). Stokes (1980) also reports a decrease
in total phytoplankton biomass with decreasing pH level for nine lakes in the same region of
Ontario. Crisman et al. (1980) reported a linear decrease in functional chlorophyll a measure-
ments with declining pH for 11 lakes in northern Florida, pH range 4.5 to 6.9. On the other
hand, Aimer et al. (1978) note that in 58 nutrient-poor lakes in the Swedish west coast
region, the largest mean phytoplankton biomass occurred in the most acid lakes (pH <4.5). Van
and Stokes (1978) concluded that they have no evidence that the phytoplankton biomass in
Carlyle Lake, with a summer pH level about 5.1, is below that observed in circumneutral lakes
in the same region. In a continuing whole-lake acidification project (Schindler et al.,
1980), a lowering of the epilimnion pH level from 6.7-7.0 in 1976 to 5.7-5.9 in 1978 resulted
in no significant change in the chlorophyll concentration or primary production. Both ui situ
and experimental acidification have resulted in large increases in periphyton populations
(Muller, 1980; Hendrey, 1976; Hall et al., 1980). Hendrey (1976) andMuller (1980) observed
XDSX7B/A 7-49 2-9-81
-------
tf.
i—i—i—i—I—i—r
PHYTOPLANKTON SPECIES IN 60 LAKES
ON THE SWEDISH WEST COAST
AUGUST 1976
80
70
tn 60
ui
o
£ 50
tt.
UJ
OD
40
30
20
10
-71
i—i i i r
pH 4.1 4.3 4.5 4.7 4.9 5.1 5.3 55 5.7 5.9 6.1 6.3 6.5 6.7 6.9 7.1
NUMBER 1 104324331210331002035054123101 1
OF LAKES
Figure 7-20. Numbers of phytoplankton species in 60 lakes having different pH values on the Swedish
West Coast, August 1976.
Source: Adapted from Aimer et al. (1978).
7-50
-------
pH 4.60-5.45
pH 6.25-7.70
BIOMASS
SPECIES
DIATOMEAE
CHLOROPHYCEAE
CHRYSOPHYCEAE
CYANOPHYCEAE
PYRROPHYTA
SEPTEMBER 1972
Figure 7-21. Percentage distribution of phytoplankton species and their biomasses.
September 1972, west coast of Sweden. Biomass = living weight per unit area.
Source: Adapted from Aimer et al. (1978).
7-51
-------
A
carbon uptake by periphyton incubated in vitro. They found that, although the total rate of
photosynthesis increased with decreasing pH level due to the larger biomass at the lower pH,
the photosynthesis per unit biomass decreased with pH.
From the above discussion it is obvious that not only is there no clear correlation
between pH level and algal biomass or productivity, but the effects of acidification appear
inconsistent between systems. Again, these apparent inconsistencies probably reflect a lack
of knowledge about exact mechanisms relating acidification and lake metabolism, and also the
complexity of these mechanisms and interactions. Changes in the algal community biomass and
productivity probably reflect the balance between a number of potentially opposing factors;
those that tend to decrease productivity and biomass versus those that tend to increase pro-
ductivity and/or biomass when acidity increases. Factors working to decrease productivity and
biomass with declining pH levels may include: (1) a shift in pH level below that optimal for
algal growth, (2) decreased nutrient availability as a result of decreased decomposition rates
and a sealing-off of the mineral sediments from the lake water; and (3) decreased nutrient
availability as a result of changes in aquatic chemistry with acidification. For example,
despite the fact that the optional pH range for growth of label!aria flocculosa is between 5.0
to 5.3 (Cholonsky, 1968) or higher (Kallqvist et a!., 1975), this species dominated experi-
mentally acidified stream communities at pH level 4 in three out of five replicates (Hendrey
et al., 1980). As noted in Section 7.3.1.1, aluminum concentrations increase with decreasing
pH level in acidified lakes and streams. Aluminum is also a very effective precipitator of
phosphorus, particularly in the pH range 5 to 6 (Dickson, 1978; Stumm and Morgan, 1970). In
oligotrophic lakes, phosphorus is most commonly the limiting nutrient for primary productivity
(Wetzel, 1975; Schindler, 1975). Therefore, chemical interactions between aluminum and phos-
phorus may result in a decreasing availability of phosphorus with decreasing pH level, and, as
a result, decreased primary production.
Factors working to increase productivity and/or biomass with acidification of a lake or
stream may include: (1) decreased loss of algal biomass to herbivores; (2) increased lake
transparency; and (3) increased nutrient availability resulting from nutrient enrichment of
precipitation. Decreased population of invertebrates (as discussed in Section 7.3.1.4),
particularly herbivorous invertebrates, may decrease grazing pressure on algae and result in
unusual accumulations of biomass. Hendrey (1976) and Hall et al. (1980) include this
mechanism as one hypothesis to explain their observation of increased biomass of periphyton at
pH level 4 despite a decreased production rate per unit biomass.
Increases in lake transparency over time have been correlated with lake acidification in
Sweden (Aimer et al., 1978) and the Adirondack Region of New York (Schofield, 1976c). In
addition, after the second year of experimental lake acidification (pH 6.7-7.0 to 5.7-5.9) in
northwestern Ontario (Schindler etal., 1980), lake transparency increased by 1-2 m. These
increases in transparency have not been correlated with decreases in phytoplankton biomass.
XDSX7B/A 7-52
2-9-81
-------
Two mechanisms have been proposed. Aluminum acts as a very efficient precipitator for humic
substances. Dickson (1978) found that humic substances are readily precipitated in the pH
range 4.0 to 5.0. Dickson (1978) and Aimer et al. (1978) suggest that increases in aluminum
levels with lake acidification (Section 7.3.1.1) have resulted in increased precipitation of
humics from the water column and therefore increased lake transparency. Aimer at al. (1978)
provide data for one lake on the west coast of Sweden. The pH level declined from above 6 to
about 4.5 between 1940 and 1975. The secchi disc reading increased from about 3m to about 10m
over the same period. Organic matter in the water (as estimated by KMnO. demand) decreased
from 24 to 8 mg/liter from 1958 to 1973. Schindler et al. (1980), on the other hand, found no
change in levels of dissolved organic carbon with acidification. Instead, changes in hydro-
lysis of organic matter with declining pH level may affect the light absorbancy characteris-
tics of the molecules. Levels of particulate organic carbon, and changes with pH level, were
not reported by Schindler et al. (1980).
Acidification of precipitation (and dry deposition) has been accompanied by increases in
levels of sulfate and nitrate. Both of these are nutrients required by plants. However, as
noted above, the primary nutrient limiting primary productivity in most oligotrophic lakes is
phosphorus. Aimer et al. (1978) report that atmospheric deposition rates of phosphorus have
also increased in recent years. The world-wide extent of the correlation between acidic
deposition and increased atmospheric phosphorus loading, however, is not known. It is expect-
ed that changes in atmospheric phosphorus loading would be much more localized than changes in
acidic deposition. It is possible that in some areas increased atmospheric loading of phos-
phorus has occurred in recent years coincidently with increased acidic deposition. Increased
phosphorus nutrient loading into lakes may then increase primary production rates.
The effect of acidification on primary productivity and algal biomass of a particular
stream or lake system depends upon the balance of the above forces. Differences in the impor-
tance of these factors between systems may account for inconsistencies in the response of
different aquatic systems to acidic deposition. Acidification does, however, result in a
definite change in the nutrient and energy flux of the aquatic system, and this change may
eventually limit the total system biomass and productivity.
Acidification of lakes has also been correlated with changes in the macrophyte community.
Documentation for these changes comes mainly from lakes in Sweden. Grahn (1976) reported that
in five to six lakes studies in the last three to five decades the macrophyte communities
dominated by Lobelia and Isoetes have regressed, whereas communities dominated by Sphagnum
mosses have expanded. Acidity levels in these lakes apparently have increased approximately
1.3 to 1.7 pH units since the 1930-40's. In acid lakes where conditions are suitable the
Sphagnum peat moss may cover more than 50 percent of the bottom above the 4-m depth, and may
also grow at much lower depths (Aimer et al., 1978). The Sphagnum invasion may start at lake
pH levels just below 6 (Aimer et al., 1978). Similar growths of Sphagnum occur in Norwegian
lakes (National Research Council, 1978). Increases in Sphagnum as a benthic macrophyte have
XDSX7B/A 7-53 2-9-81
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been documented from one lake in the Adirondack Region of New York (Hendrey and Vertucci,
1980).
Under acid conditions the Sphagnum moss appears to simply outgrow flowering plant
aquatic macrophytes. In laboratory tests, the growth and productivity of the rooted macro-
phyte Lobelia was reduced by 75 percent at a pH of 4, compared with the control (pH 4.3-5.5).
The period of flowering was delayed by ten days at the low pH (Laake, 1976). At low pH levels
(pH<5), essentially all the available inorganic carbon is in the form of carbon dioxide or
carbonic acid (Stumm and Morgan, 1970). As a result, conditions may be more favorable for
Sphagnum, an acidophile that is not able to utilize the carbonate ion.
Besides the shift in macrophyte species, the invasion of Sphagnum into acid lakes may
have four other impacts on the aquatic ecosystem. Sphagnum has a very high ion-exchange
capacity, withdrawing basic cations such as Ca++ from solution and releasing H (Anschutz and
Gessner, 1954; Aimer et al., 1978). As a result, the presence of Sphagnum may intensify the
acidification of the system and decrease the availability of basic cations from other biota.
Second, dense growths of Sphagnum form a biotype that is an unsuitable substratum for many
benthic invertebrates (Grahn, 1976). Growths of Sphagnum in acidic lakes are also often
associated with felts of white mosses (benthic filamentous algae) and accumulations of
non-decomposed organic matter. In combination, these organisms and organic matter may form a
very effective seal. Interactions between the water column and the mineral sediments, and the
potential for recycling of nutrients from the sediments back into the water body, may be
reduced (Grahn, 1976; Grahn, et al., 1974). These soft bottoms may also be colonized by other
macrophytes. In Sweden, Aimer et al. (1978) report that growths of Juncus, Sparaganium,
Utricularia, Nuphar, and/or Nymphaea, in addition to Sphagnum, may be extensive in acidic
lakes. Thus primary production by macrophytes in lakes with suitable bottoms may be very
large. Increased lake transparency may also increase benthic macrophyte and algal primary
productivity.
7.3.1.4 Effects on invertebrates. In regional surveys conducted in southern Norway (Hendrey
and Wright, 1976), the west coast of Sweden (Aimer et al., 1978), the LaCloche Mountain Region
of Canada (Sprules, 1975), and near Sudbury, Ontario (Roff and Kwiatkowski, 1977) numbers of
species of zooplankton were strongly correlated with pH level (Figure 7-22). Changes in
community structure were most noticeable at pH levels below 5. Certain species (e.g., of the
genera Bosmina, Cyclops, Diaptomus, and rotiferans, of the genera Polyarthra, Keratella, and
Kellicottia) apparently have a high tolerance of acidic conditions and were commonly found in
the pH interval 4.4 to 7.9. Others, such as cladocerans of the Daphnia genus, apparently are
more sensitive and were only rarely found at pH <6 (Aimer et al., 1978).
Similar studies of the relationship between pH level and biomass or productivity of zoo-
plankton are not available. Proposed mechanisms for interactions between lake acidification
and zooplankton populations are therefore largely hypothetical.
The species, population size, and productivity of zooplankton are affected both by
changes in the quality and quantity of the food supply and shifts in predator populations.
XDSX7B/A 7-54 2-9-81
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o
Ul
a.
1 3
o
cc
UJ
00
S
D
>6.5 pH INTERVAL
12 NUMBER OF LAKES
Figure 7-22. The number of species of crustacean zooplankton observed in 57 lakes during a
synoptic survey of lakes in southern Norway.
Source: Levestad et al. (1976).
7-55
-------
Changes in zooplankton species and production in response to changes in fish populations have
been clearly demonstrated (Brooks and Dodson, 1965; Walters and Vincent^ 1973; Dodson, 1974).
Elimination of fish predators often results in dominance of the zooplankton community by large-
bodied species. Absence of invertebrate predators (e.g., large-bodied carnivorous zoo-
plankton) as a result of fish predation or other reasons often results in the prevalence of
small-bodied species (Lynch, 1979). Surveys of acidic lake waters often have shown the
dominance of small-bodied herbivores in the zooplankton community (Hendrey et al, 1980). Fish
also often are absent at these pH levels (Section 7.3.1.5). Different zooplankton species may
have different physiological tolerances to depressed pH levels (e.g., Potts and Fryer, 1979).
Food supplies, feeding habits, and grazing of zooplankton may also be altered with acidifi-
cation as a consequence of changes in phytoplankton species composition and/or decreases in
biomass or productivity of phytoplankton. Zooplankton also rely on bacteria and detrital
organic matter for part of their food supply. Thus an inhibition of the microbiota or a
reduction in microbial decomposition (Section 7.3.1.2) may also affect zooplankton popula-
tions. These alternate mechanisms postulate for changes in community structure and/or
production of zooplankton communities probably play an important role in zooplankton responses
to acidification.
Synoptic and intensive studies of lakes and streams have also demonstrated that numbers
of species of benthic invertebrates are reduced along a gradient of decreasing pH level
(Sutcliffe and Carrick, 1973; Leivestad et al., 1976; Conroy et al., 1976; Aimer et al., 1978;
Roff and Kwiatkowski, 1977). In 1500 freshwater localities in Norway studied from 1953-73,
snails were generally present only in lakes with pH levels above 6 (Okland, J. , 1980). Like-
wise Gammarus Lacustris, a freshwater shrimp and an important element in the diet of fish in
Scandinavia, was not found at pH levels below 6.0 (Okland, K. , 1980). Experimental investi-
gations have shown that adults of this species cannot tolerate two days of exposure to pH 5.0
(Leivestad et al., 1976). Eggs were reared at six different pH levels (4.0 to 6.8). At a pH
of 4.5 a majority of the embryos died within 24 hours. Thus the short-term acidification
which often occurs during the spring melt of snow could eliminate this species from small
lakes (Leivestad et al., 1976). Fiance (1977) concluded that ephemeropterans (mayflies) were
particularly sensitive to low pH levels and their populations were reduced in headwater
streams of the Hubbard Brook watershed in New Hampshire. In laboratory studies, Bell (1971),
Bell and Nebeke (1969), and Raddum (1978) measured the tolerance of some stream macroinverte-
brates to low pH levels. Tolerance seems to be in the order caddisflies > stoneflies > may-
flies (Hendrey et al., 1980).
2 2
Leivestad et al. (1976) reported on decreased standing crops (numbers/m and g/m ) of
benthic invertebrates in two lakes with pH levels near 4.5 as compared to five lakes with pH
near 6.0. Chironamids were the dominant group in all lakes. No fish were found in the acid
lakes. Lack of predation by fish should favor increases in benthic biomass, the opposite of
that observed. Hendrey et al. (1980), on the other hand, from data from eight Ontario lakes
(pH 4.3 - < 5.7) reported no reduction in abundance of benthos related to pH level.
XDSX7B/A 7-56 2-9-81
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*
Air-breathing aquatic insects (e.g., backswimmers, water boatmen, water striders) appear
very tolerant of acidic environments. Population densities are often greater in acidic lakes
and in the most acid lakes than in circumneutral lakes. Abundance of these large inverte-
brates may be related to reduced fish predation (Hendrey et al., 1980).
Hall et al. (1980) experimentally acidified a stream to pH 4 and monitored reactions of
macroinvertebrate populations. Initially following acidification there was a 13-fold increase
in downstream drift of insect larvae. Organisms in the collector and scraper functional
groups were affected more than predators. Benthic samples from the acidified zone of Norris
Brook contained 75 percent fewer individuals than those for reference areas. There was also a
37 percent reduction in insect emergence; members of the collector group were most affected.
Insects seem to be particularly sensitive at emergence (Bell, 1971). Many species of aquatic
insects emerge early in the spring through cracks in the ice and snow cover. These early-
emerging insects are therefore exposed in many cases to the extremely acidic conditions
associated with snowmelt (Hagen and Langeland, 1973).
Low pH also appeared to prevent permanent colonization by a number of invertebrate
species, primarily herbivores, in acidified reaches of River Dudden, England (Sutcliffe and
Carrick, 1973). Ephemeroptera, trichoptera, Ancylus (Gastropoda) and Gammarus were absent in
these reaches.
Damage to invertebrate communities may influence other components of the food chain.
Observations that herbivorous invertebrates are especially reduced in acidic streams, as
reported in Norris Brook and River Dudden, support the hypothesis (Hendrey, 1976; Hall et al.,
1980) that changes in invertebrate populations may be responsible for increased periphytic
algal accumulations in acidic streams and benthic regions of acidic lakes (Hendrey et al.,
1980). Benthic invertebrates also assist with the essential function of processing dead
organic matter. Petersen (1980) noted that decomposition of coarse particulate organic matter
in leaf packs was lower in an acidic stream than in two streams with circumneutral pH levels.
The invertebrate community also showed a reduction in the invertebrate functional group that
specializes in processing large particles (shredders). In unstressed aquatic ecosystems, a
continuous emergence of different insect species is available to predators from spring to
autumn. In acid-stressed lakes or streams, the variety and numbers of prey may be reduced.
Periods may be expected to occur in which the amount of prey available to fish (or other
predators) is diminished.
7.3.1.5 Effects on fish. Acidification of surface waters has had its most obvious, and
perhaps the most severe, impact on fish populations. Increasing acidity has resulted not just
in changes in species composition or decreases in biomass but in many cases in total elimina-
tion of populations of fish from a given lake or stream. Extensive depletion of fish stocks
has occurred in large regions of Norway, Sweden, and parts of eastern North America. Both
commercial and sport fisheries have been affected in these areas. However, precise assess-
ments of losses—in terms of population extinctions, reductions in yields, or economic and
XDSX7B/A 7-57 2-9-81
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social impacts—either have not been attempted or are still in the process of evaluation.
Potential damage to fish populations inhabiting other acid-sensitive aquatic ecosystems in New
England, the Appalachians, and parts of southeastern, north central, and northwestern United
States have not yet been assessed (National Research Council, 1978).
Declines in fish populations have been related to acidification of surface waters in the
Adirondack Region of New York State (Schofield 1976), southern Norway (Jensen and Snekyik,
1972; Wright and Snekvik, 1978), southwestern Sweden (Aimer et al., 1974), the LaCloche Moun-
tain Region in southern Ontario (Beamish and Harvey, 1972), and southwestern Scotland (Wright
et al., 1980a). Schofield (1976) estimated that in 1975 fish populations in 75 percent of
Adirondack lakes at high elevation (<610 m) had been adversely affected by acidification.
Fifty-one percent of the lakes had pH values less than 5, and 90 percent of these lakes were
devoid of fish life (Figure 7-23). Comparable data for the period 1929 to 1937 indicated that
during that time only about 4 percent of these lakes had pH values below 5 and were devoid of
fish (Figure 7-27). Therefore, entire fish communities consisting of brook trout (Salvelinus
fontinalis), lake trout (Salvelinus namaycush), white sucker (Catostomus commersoni), brown
bullhead (Icaturus nebulosus) and several cyprinid species were apparently eliminated over a
period of 40 years. This decrease in fish populations was associated with a decline in lake
pH level. A survey of more than 2000 lakes in southern Norway, begun in 1971, found that
about one third of these lakes had lost their fish population (primarily brown trout, Salmo
trutta L.) since the 1940's (Wright and Snekvik, 1978). Fish population status was inversely
related to lake pH level (Leivestad et al., 1976). Declines in salmon populations in southern
Norwegian rivers were reported as early as the 1920's. Catch of Atlantic salmon (Salmo salar,
L.) in nine acidified southern Norwegian rivers is now virtually zero (Figure 7-25). In
northern and western rivers not affected by acidification, no distinct downward trend in catch
has occurred (Leivestad et al., 1976; Wright et al., 1976; Jensen and Snekvik, 1972). Similar
changes have been observed in Sweden (Aimer et al., 1974) where it is estimated that 10,000
lakes have been acidified to a pH less than 6.0 and 5,000 below a pH of 5.0 (Dickson, 1975).
Populations of lake trout, lake herring (Coregonus artedii), white suckers, and other species
disappeared rapidly during the 1960's from a group of remote lakes in the LaCloche Mountain
Region of Ontario (Beamish et al., 1975).
It is difficult to determine at what pH level fish species disappear from lakes. Disap-
pearance of the fish is usually not due to massive fish kills, but is the result of a gradual
depletion of the population following reproductive failures (Leivestad et al., 1976). Field
surveys in Scandinavia and eastern North America (Wright and Snekvik, 1978, Aimer et al.,
1974; Schofield, 1976) suggest that many species do not occur in lakes with pH values below
5.0.
However, large spatial and temporal functions in pH, and the possibility for "refuge areas"
from acidic conditions during critical periods make it extremely difficult to generalize about
effects of acidification on fish populations based on grab samples or annual mean pH levels.
XDSX7B/A 7-58 2-9-81
-------
20
u.
O
E
ffi
s
D
10
PH
Figure 7-23. Frequency distribution of pH and fish population status in Adirondack Mountain
lakes greater than 610 meters elevation. Fish population status determined by survey gill netting
during the summer of 1975.
Source: Schofield (1976b).
7-59
-------
20
10
O
5
CD
10
~" I I
1975
NO FISH PRESENT
FISH PRESENT
1930s
rfl
6
PH
Figure 7-24. Frequency distribution of pH and fish pop-
ulation status in 40 Adirondack lakes greater than 610 meters
elevation, surveyed during the period 1929-1937 and again in
1975.
Source: Schofield (1976 b).
7-60
-------
STAVANGER
RIVER TOVDAL
300
250
V)
O 200
150
I I I I I
I I I \V I I I
1900
1920
1940
1960
1980
O
30
20
10
I I T T
7 ACID RIVERS
I I I I I I
1900
1920
1940
1960
1980
Figure 7-25. Norwegian salmon fishery statistics for 68 unacidified and 7 acidified
rivers.
Source: Adapted from Aimer et al. (1978).
7-61
-------
The pH levels identified in the literature as critical for reproduction of a species or corre-
lated with the absence of a species in lake surveys are summarized in Table 7-6. Values range
from pH 4.4 to over 6.0, and are highly species dependent.
TABLE 7-6. pH LEVELS IDENTIFIED IN FIELD SURVEYS AS CRITICAL TO
LONG-TERM SURVIVAL OF FISH POPULATIONS
Family
Salmonidae
Species
Brook trout (Salvelinus
fontinalis)
Lake trout (Salvelinus
namaycush)
Critical pH
5.0
5.1
5.2-5.5
Reference
Schofield, 1976c
Schofield, 1976c
Beamish, 1976
Brown trout (Salmo trutta) 5.0
Arctic char (Salvelinus alpinus)
5.2
Percidae
Catostomidae
Ictaluridae
Cyprinidae
Perch (Perca fluvi'ati'li's) 4.4-4.9
Yellow perch (Perca flavescens) 4.5-4.7
Walleye (Stigostedion vitreum) 5.5-6.0+
White sucker (Catostomus 4.7-5.2
commersoni) 5.1
Aimer et al., 1978
Aimer et al., 1978
Aimer et al., 1978
Beamish, 1976
Beamish, 1976
Beamish, 1976
Schofield, 1976c
Brown bullhead (Icaturus
nebulosus)
4.7-5.2 Beamish, 1976
5.0 Schofield, 1976c
Minnow (Phoxinus phoxinus) 5.5
Roach (Rutilus rutilus)5.5
Lake chub (Couesius plumbeus) 4.5-4.7
Creekchub (Semotilus atromaculatus) 5.0
Commonsniner (Notropis cornutasl5.5
Goldenshiner (Notemigonus 4.9
crysoleucas)
5.5-6.0+
Almer et al., 1978
Aimer et al., 1978
Beamish, 1976
Schofield, 1976c
Schofield, 1976c
Schofield, 1976c
Centrarchidae Smallmouth bass (Micropterus
dolomieui)
Rock bass (Ambloplites rupestris) 4.7-5.2
Esocidae
Pike (Esox lucius)
Beamish, 1976
Beamish, 1976
4.4-4.9 Aimer et al., 1978
Recent field and laboratory studies (Schofield and Trojnar, 1980; Dickson, 1978; Driscoll
et al., 1979; Baker and Schofield, 1980; Muniz and Leivestad, 1980) have indicated that aluminum
levels in acidic surface waters (Section 7.3.1.1, Figure 7-18) may be highly toxic to fish
(and perhaps other biota). Schofield and Trojnar (1980) analyzed survival of brook trout
stocked into 53 Adirondack lakes as a function of 12 water quality parameters. Levels of pH,
calcium, magnesium, and aluminum were significantly different between the two groups of lakes,
with and without trout survival. However, after accounting for the effects of aluminum
XDSX7B/A
7-62
2-9-81
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concentrations on differences between the two groups of lakes, differences in calcium,
magnesium, and pH levels were no longer significant. Aluminum, therefore, appears to be the
primary chemical factor controlling survival of trout in these lakes. Likewise, in laboratory
experiments with natural Adirondack waters and synthetic acidified aluminum solutions, levels
of aluminum, and not the pH level per se, determined survival and growth of fry of brook trout
and white suckers (Baker and Schofield, 1980). In addition, speciation of aluminum had a sub-
stantial effect on aluminum toxicity. Complexation of aluminum with organic chelates elimin-
ated aluminum toxicity to fry (Baker and Schofield, in press; Driscoll et al., 1979). As a
result, waters high in organic carbon, e.g., acidic bog lakes, may be less toxic to fish than
surface waters at similar pH levels but with lower levels of dissolved organic carbon.
Inorganic aluminum levels, and not low pH levels, may therefore be a primary factor
leading to declining fish populations in acidified lakes and streams. However, many labora-
tory or in situ field experiments have been conducted on the effects of pH on fish without
taking into account aluminum or other metal concentrations in naturally acidic waters. As a
result, many of the conclusions based on these experiments regarding pH levels critical for
fish survival are suspect. Therefore these experiments will not be reviewed here.
Sensitivity of fish and other biota to low pH levels has also been shown to depend on
aqueous calcium levels (Wright and Snekvik, 1978; Trojnar, 1977; Bua and Snekvik, 1972). In
southern Norway, the mean calcium level in lakes studied was approximately 1.1 mg/liter, as
compared to about 3 mg/liter in the LaCloche Mountain Region (Table 7-5) or 2.1 mg/liter in
the Adirondack Region (Schofield, 1976b). In Norwegian lakes, Wright and Snekvik (1978)
identified pH and calcium levels as the two most important chemical parameters related to fish
status.
Decreased recruitment of young fish has been cited as the primary factor leading to the
gradual extinction of fish populations (Leivestad et al., 1976; Rosseland et al., 1980; Wright
and Snekvik, 1978). Field observations (Jensen and Snekvik, 1972; Beamish, 1974; Schofield,
1976; Aimer et al., 1974) indicate changes in population structure over time with acidifi-
cation. Declining fish populations consist primarily of older and larger fish with a decrease
in total population density. Recruitment failure may result from inhibition of adult fish
spawning and/or increased mortality of eggs and larvae. Effects on spawning and decreased egg
deposition may be associated with disrupted spawning behavior and/or effects of acidification
on reproductive physiology in maturing adults (Lockhart and Lutz, 1976). Field observations
by Beamish et al. (1975) related reproductive failure in white suckers to an inability of
females to release their eggs. On the other hand, Amundsen and Lunder (1974) observed total
mortality of naturally spawned trout eggs in an acid brook a few weeks after spawning. A sum-
mary of Norwegian studies (Leivestad et al., 1976) concluded that egg and fry mortality is the
main cause of fish reproduction failure. Spawning periods and occurrence of early life
history stages for many fish species coincide with periods of extreme acidity, particularly
during and immediately after snowmelt in the spring.
XDSX7B/A 7-63 2-9-81
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In some lakes, fish population decreases are associated with a lack of older fish
(Rosseland et al., 1980). In Lake Tveitvatn on the Tovdal River in southern Norway, brown
trout mortality apparently occurs primarily after the first spawning. Since 1976, no fish
past spawning age have been found and population density has decreased steadily (Rosseland et
al., 1980). Fish kills of adult salmon in rivers in southern Norway have been recorded as
early as 1911 (Leivestad et al., 1976).
When evaluating the potential effects of acidification on fish, or other biotic, popula-
tions, it is very important to keep in mind the highly diversified nature of aquatic systems
spatially, seasonally, and year-to-year. As a result of this diversity, it is necessary to
evaluate each system independently in assessing the reaction of the population to acidifica-
tion. Survival of a fish population may depend more on the availability of refuge areas from
acid conditions during spring melt or of one tributary predominantly fed by baseflow and
supplying an adequate area for spawning than on mean annual pH, calcium, or inorganic aluminum
levels.
7.3.1.6 Effects on vertebrates other than fish. Certain species of amphibians may be the
vertebrate animals, other than fish, most immediately and directly affected by acidic deposi-
tion (Rough and Wilson, 1976). Their vulnerability is due to their reproductive habits. In
temperate regions, most species of frogs and toads, and approximately half of the terrestrial
salamanders, lay eggs in ponds. Many of these species breed in temporary pools formed each
year by accumulation of rain and melted snow. Approximately 50 percent of the species of
toads and frogs in the United States regularly breed in ephemeral pools; about one-third of
the salamander species that have aquatic eggs and larvae and terrestrial adults breed in
temporary pools. Most of these pools are small and collect drainage from a limited area. As
a result, the acidity of the eater in these pools is strongly influenced by the pH of the pre-
cipitation that fills them. Ephemeral pools are usually more acidic than adjacent permanent
bodies of water. Rough and Wilson (1976) report that in 1975, in the vicinity of Ithaca,
N.Y., the average pH of 12 temporary ponds was 4.5 (range 3.5 to 7.0), while the average pH of
six permanent ponds was 6.1 (range 5.5 to 7.0). Amphibian eggs and larvae in temporary pools
are exposed to these acidic conditions.
Rough and Wilson (1976) and Rough (1976) studied the effect of pH level on embryonic
development of two common species of salamanders: the spotted salamander (Ambystoma
maculatum) and the Jefferson salamander ( A. jeffersonianum). In laboratory experiments,
embryos of the spotted salamander tolerated pH levels from 6 to 10 but had greatest hatching
success at pH 7 to 9. The Jefferson salamander tolerated pH levels 4 to 8 and was most
successful at 5 to 6. Mortality of embryos rose abruptly beyond the tolerance limits. In a
four-year study of a large breeding pond (pH 5.0-6.5) 938 adult spotted salamanders produced
486 metamorphosed juveniles (0.52 juveniles/adult), while 686 adult Jefferson salamanders
produced 2157 juvenile (3.14 juveniles/adult). Based on these findings, Rough and Wilson
(1976) predict that continued acidic deposition may result in substantial shift in salamander
and other amphibian populations.
XDSX7B/A 7-64 2-9-81
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*
Gosner and Black (1957) report that only acid-tolerant species of amphibians can breed in
the acid (pH 3.6 to 5.2) sphagnoceous bogs in the New Jersey Pine Barrens.
Frog populations in Tranevatten, a lake near Gothenberg, Sweden, acidified by acidic
precipitation, have also been investigated (Hagstrom, 1977; Hendrey, 1978). The lake has pH
levels ranging from 4.0 to 4.5. All fish have disappeared, and frogs belonging to the species
Rana temporaria and Bufo bufo are being eliminated. At the time of the study (1977) only
adult frogs eight to ten years old were found. Many egg masses of Rana temporaria were
observed in 1974, but few were found in 1977, and the few larvae (tadpoles) observed at that
time died.
Frogs and salamanders are important predators on invertebrates, such as mosquitoes and
other pest species, in pools, puddles, and lakes. They also are themselves important prey for
higher tropic levels in an ecosystem. In many habitats salamanders are the most abundant
vertebrates. In a New Hampshire forest, for example, salamanders were found to exceed birds
and mammals in both numbers and biomass (Hanken et al., 1980).
The elimination of fish and vegetation from lakes by acidification may have an indirect
effect on a variety of vertebrates: species of fish-eating birds (e.g., the bald eagle, loon,
and osprey), fish-eating mammals (e.g., mink and otter), and dabbling ducks which feed on
aquatic vegetation. In fact, any animal that depends on aquatic organisms (plant or animal)
for a portion of its food may be affected.
Increasing acidity in freshwater habitats results in shifts in species, populations, and
communities. Virtually all trophic levels are affected. Summaries of the changes which are
likely to occur in aquatic biota with decreasing pH are listed in Tables 7-7 and 7-8.
TABLE 7-7. CHANGES IN AQUATIC BIOTA LIKELY TO OCCUR WITH INCREASING ACIDITY
1. Fish populations are reduced or eliminated.
2. Bacterial decomposition is reduced and fungi may dominate saprotrophic
communities. Organic debris accumulates rapidly, tying up nutrients and
limiting nutrient mineralization and cycling.
3. Species diversity and total numbers of species of aquatic plants and
animals are reduced. Acid-tolerant species dominate.
4. Phytoplankton productivity may be reduced due to changes in nutrient
cycling and nutrient limitations.
5. Biomass and total productivity of benthic macrophytes and algae may
increase due partially to increased lake transparency.
6. Numbers and biomass of herbivorous invertebrates decline. Tolerant
invertebrate species, e.g., air-breathing bugs (water-boatmen, back-
swimmers, water striders) may become abundant primarily due to reduced
fish predation.
7. Changes in community structure occur at all trophic levels.
XDSX7B/A 7-65 2-9-81
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TABLE 7-8. SUMMARY OF EFFECTS ON AQUATIC ORGANISMS ASSOCIATED WITH A RANGE IN pH*
1
cr>
en
8.0-6.0 • Long-term changes of less than 0.5 pH units in the range 8.0 to
6.0 are likely to alter the biotic composition of freshwaters to
some degree. The significance of these slight changes is, however,
not great.
• A decrease of 0.5 to 1.0 pH units in the range 8.0 to 6.0 may cause
detectable alterations in community composition. Productivity of
competing organisms will vary. Some species will be eliminated.
• Phytoplankton plentiful and well distributed but numbers of species
begin to decrease as pH decreases.
6.0-5.5 • Decreasing pH from 6.0 to 5.5 will cause a reduction in species
numbers and, among remaining species, alterations in ability
to withstand stress, and change in species dominance.
Reproduction of some salamander species is impaired.
5.5-5.0 • Below pH 5.5, numbers and diversity of species will be reduced.
Many species will be eliminated. Crustacean zooplankton, phy-
toplankton, molluscs, amphipods, most mayfly species, and many
stone fly species will begin to drop out. In contrast, several
pH-tolerant invertebrates will become abundant, especially the
air-breathing forms (e.g., Gyrinidae, Notonectidae, Corixidae),
those with tough cuticles which prevent ion losses (i.e.,
Si all's lutaria), and some forms which live within the sediments
(01igochaeta, Cniromomidae, and Tubificidae). Overall, inver-
tebrate biomass may be reduced.
5.0-4.5 • Below pH 5.0, decomposition of organic detritus will be severely
impaired. Organic matter accumulates rapidly. Some fungal
species increase (Hyphomycetes, basidomycetes). Many fish
species are eliminated, (see Table 7-7.)
Aimer et al., 1974;
Leivestad et al., 1976;
Aimer et al., 1978
Aimer et al., 1974;
Leivestad et al., 1976;
Conroy et al., 1976;
Aimer et al., 1978
Aimer et al., 1974;
Kwiatkowski and Roff, 1976;
Aimer et al., 1978
Aimer et al., 1974;
Leivestad, 1976;
Kwaitkowski and Roff, 1976;
Aimer et al., 1978
Rough and Wilson, 1976
Aimer et al., 1974;
Leivestad et al., 1976;
Hendrey et al., 1976;
Grahn et al., 1974;
Grahn, 1976;
Kwiatkowski and Roff, 1976;
Hagen and Langeland, 1973;
Henriksen and Wright, 1977
Hultberg, 1976;
Aimer et al., 1978
Leivestad et al., 1976;
Schofield, 1976b;
Aimer et al., 1978
Hall et al., 1980.
-------
Macrophytes, such as Lobelia, are replaced by Sphagnum moss.
Number of algal species decreases. Acid-tolerant forms remain.
4.5 and • Below pH 4.5, all of the above changes will be greatly exacerbated,
below and all fish will be eliminated. Lower limit for many algal
species.
Leivestad et al., 1976
Hendrey et al., 1976;
Grahn, 1974;
Aimer et al., 1978
Leivestad et al., 1976;
Hendrey et al., 1976;
Grahn et al., 1974;
Aimer et al., 1978
Aimer et al., 1974
Leivestad et al., 1976;
Schofield, 1976b;
Wright et al., 1976
Beamish et al. , 1975;
Menedex, 1976;
Trojnar, 1977;
Trojnar, 1977;
Schofield, 1975;
Schofield, 1979
Modified from Hendrey (1978).
-------
*
7-3.2 Terrestrial Ecosystems
Determining the effects of acidic precipitation on terrestrial ecosystems is not an easy
task. In aquatic ecosystems it has been possible to measure changes in pH that occur in acidi-
fied waters and then observe the response of organisms living in aquatic ecosystems to the
shifts in pH. In the case of terrestrial ecosystems the situation is more complicated since
no component of terrestrial ecosystems appears to be as sensitive to acidic precipitation as
organisms living in poorly buffered aquatic ecosystems. Nonetheless, soils and vegetation may
be affected, directly or indirectly, by acidic precipitation, albeit in complex ways.
7.3.2.1 Effects on soils—Acidity is a critical factor in the behavior of natural or agri-
cultural soils. Soil acidity influences the availability of plant nutrients and various
microbiological processes which are necessary for the functioning of terrestrial ecosystems,
therefore, there is concern that acidic precipitation over time could have an acidifying effect
on soils through the addition of hydrogen ions. As water containing hydrogen cations (usually
from weak acids) moves through the soil, some of the hydrogen ions replace adsorbed exchange-
able cations, such as Ca , Mg , K , and Na (see Figure 7-26). The removed cations are then
carried deep into the soil profile or into the ground water. In native soils hydrogen ions
are derived from the following sources (Wiklander, 1979):
1. nutrient uptake by plants—the roots adsorb cation
nutrients and desorb H ;
2. CO^ produced by plant roots and micro-organisms;
3. oxidation of NH4+ and S, FeS^, and hLS to HNO- and H^SO.;
4. very acid litter in coniferous forests, the main acidifying
source for the A and B horizons;
5. atmospheric deposition of H9SO. and some HNO,, NO , HC1 and
£ £f. Mg2+ > Ca2+ (Wiklander, 1979).
Norton (1977) cited the potential effects of acidic deposition on soils that are listed
in Table 7-9. Of those listed, only the increased mobility of cations and thsir accelerated
loss has been observed in field experiments. Overrein (1972) observed an increase in calcium
leaching under simulated acid rain conditions and increased loss by leaching of Ca*+, Mg*+,
and Al were observed by Cronan (1980) when he treated New Hampshire soils with simulated
acid rain at a pH 4.4.
Wiklander (1979) notes that in humid areas leaching leads to a gradual decrease of plant
nutrients in available and mobilizable forms. The rate of nutrient decrease is determined by
XDSX7A/D 7-68 2-9-81
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SOIL PARTICLES
ACID RAIN
1
SOIL SOLUTION
WEATHERING
-
-
-
NG »^
_
*
*
*
Ca"
Mg"
t/*
K
Na*— T
NH;
sof-
H2P07
A *
Mg2*
u*
K*
Na*
NOj
so?-
•
L ^
A & ki p»r- • F-A^LJr
I
Figure 7-26. Showing the exchangeable ions
of a soil with pH 7, the soil solution com-
position, and the replacement of Na+ by H+
from acid rain.
Source: Wiklander (1979).
7-69
-------
TABLE 7-9. POTENTIAL EFFECTS OF ACID PRECIPITATION ON SOILS
Effect
Comment
Increased mobility of
most elements
Increased loss of
existing clay minerals
A change in cation
exchange capacity
A general propor-
tionate increase in
the removal of all
cations from the soil
An increased flux in
nutrients through the
ecosystem below the
root zone
Mobility changes are essentially
in the order: monovalent,
divalent, trivalent cations.
Under certain circumstances may
be compensated for by production
of clay minerals which do not
have essential (stoichiometric)
alkalies or alkali earths.
Depending on conditions, this
may be an increase or a decrease.
In initially impoverished or
unbuffered soil, the removal
may be significant on a time
scale of 10 to 100 years.
Source: Norton (1977).
XDSX7A/D
7-70
2-9-81
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*
the buffering capacity of the soil and the amount and composition of precipitation (pH and
salt content). Leaching sooner or later leads to soil acidification unless the buffering
capacity of the soil is strong and/or the salt concentration of precipitation is high. Soil
acidification influences the amount of exchangeable nutrients and is also likely to affect
various biological processes in the soil.
Acidic precipitation increases the amounts of SO. and NO ~ entering the soils. Nitrate
is easily leached from soil; however, because it is usually deficient in the soil for both
plants and soil microorganisms, it is rapidly taken up and retained within the soil-piant
system (Gjessing et al., 1976; Abrahamsen et al., 1976; Abrahamsen and Do!lard, 1979). The
fate of sulfate is determined by its mobility. Retention of sulfate in soils appears to
depend on the amount of hydrous oxides of iron and aluminum present. The amounts of these
compounds present varies with the soil type. Insignificant amounts of the hydrated oxides of
iron (Fe) and aluminum (Al) are found in organic soils; therefore, sulfate retention is low
(Abrahamsen and Dollard, 1979). The presence of hydrated oxides of iron and aluminum, how-
ever, is only one of the factors associated with the capability of a soil to retain sulfur.
The capacity of soils to adsorb and retain anions increases as the pH decreases and with the
salt concentration. Polyvalent anions of soluble salts added experimentally to soils increases
adsorption and decreases leaching of salt cations. The effectiveness of the anions studied in
2
preventing leaching was in the following order: Cl ~ NO., < SO. < HUPO. (Wiklander, 1980).
Additions of sulfuric acid to a soil will have no effect on cation leaching unless the sulfate
is mobile, as cations cannot leach without associated anions (Johnson et al. 1980; Johnson,
1980; Johnson and Cole, 1980).
Leaching of soil nutrients is efficiently inhibited by vegetation growing on it. Plant
roots take up the nutrients frequently in larger amounts than required by the plants. Large
amounts of these nutrients will later be deposited on the soil surface as litter or as leach-
ate from the vegetation canopy (Abrahamsen and Dollard, 1979).
In lysimeter experiments in Norway, plots with vegetation cover were used. One plot had
a dense layer of the grass, Deschampsia flexuosa (L.) Trin. and the other a less dense cover.
2-
The soil retained 50 percent of the SO. added to it. The greatest amount was retained in
the lysimeters covered with grass; the relative retention increased with increasing additions
of sulfate (Abrahamsen and Dollard, 1979). Leaching of cations from the soil was reduced by
2- 2+ 2+
the retention of the SO. ; however, leaching of Ca and Mg increased significantly as the
acidity of the simulated rain increased. In the most acid treatment leaching of Al was highly
significant. The behavior of K , NO., , and NH. was different in the two lysimeter series.
These ions were retained in the grass-covered lysimeters whereas there was a net leaching of
K and NO, in the other series. Statistically significant effects were obtained only when
the pH of the simulated rain was 3.0 or lower (Abrahamsen and Dollard, 1979).
The Scandinavian lysimeter experiments appear to demonstrate that the relative rate of
adsorption of sulphate increases as the amounts applied are increased. In the control
XDSX7A/D 7-71 2-9-81
-------
lysimeters the output/input ratio was approximately one. These results are in agreement with
results of watershed studies which frequently appear to demonstrate that, on an annual basis,
sulfate outflow is equal to or greater than the amounts being added (Gjessing et al., 1976;
Abrahamsen and Dollard, 1979). Increased outflow may be attributed to dry deposition and the
weathering of sulfur-bearing rocks. The increased deposition of sulfate via acidic precipita-
tion appears to have increased the leaching of sulfate from the soil. Together with the
retention of hydrogen ions in the soil this results in an increased leaching of the nutrient
cations K+, Ca2+, Mg2+, Mn (Abrahamsen and Dollard, 1979). Shriner and Henderson (1978),
however, in their study of sulfur distribution and cycling in the Walker Branch Watershed in
eastern Tennessee noted the additions of sulfate sulfur by precipitation were greater than the
amount lost in stream flow. Analysis of the biomass and soil concentrations of sulfur indi-
cated that sulfur was being retained in the mineral soil horizon. It is suggested that leach-
ing from organic soil horizons may be the mechanism by which sulfur is transferred to the
mineral horizon. Indirect evidence suggests that vegetation scavenging of atmospheric sulfate
plays an important role by adding to the amounts of sulfur entering the forest system over wet
and dry deposition.
Studies of the nutrient cycling of sulfur in a number of forest ecosystems indicate that
some ecosystem accumulate (Johnson et al. 1980; Heinricks and Mayer, 1977, Shriner and Hender-
son 1978) while other ecosystems maintain a balance between the additions and losses of sulfur
or show a net loss (Cole & Johnson, 1977). Sulfur accumulation appears to be associated with
sulfate adsorption in subsoil horizons. Sulfate adsorption is strongly dependent on pH.
Little adsorption occurs above pH 6-7 (Harward and Reisenaur, 1966). The amount of sulfate in
a soil is a function of a soil's adsorption properties and the amount of sulfate that has been
added to the soil, integrated over time. Soil properties may favor the adsorption of sulfate;
however, the net annual accumulation of sulfate at any specific time will be influenced by the
degree of soil saturation (Johnson et al. 1980).
The effects of acidic precipitation on soils potentially could be long-lasting. Oden
(1971) has estimated that rainfall at pH 4.0 would be the cation equivalent of 30 kg Ca /ha,
which represents a considerable potential loss of cations essential for plant growth as well
as base saturation. McFee et al. (1976) calculated that 1000 cm of rainfall at pH 4.0 could
reduce the base saturation of the upper 6 cm of a midwestern United States forest soil by 15
percent and lower the pH of the A-l horizon (the surface layer in most agricultural soils) by
0.5 units if no countering forces are operating in the soil. They note; however, that many
counteracting forces could reduce the final effect of acidic precipitation, including the
release of new cations to exchange sites by weathering and nutrient recycling by vegetation.
Lowered soil pH also influences the availability and toxicity of metals to plants. In
general, potentially toxic metals become more available as pH decreases. Ulrich (1975)
reported that aluminum released by acidified soils could be phytotoxic if acid rain continued
for a long period. The degree of ion leaching increased with decreases in pH, but the amount
XDSX7A/D 7-72 2-9-81
-------
of cations leached was far less than the amount of acid added (Malmer, 1976). Baker et al.
(1976) found that sulfur dioxide in precipitation increased the extractable acidity and alumi-
num, and decreased the exchangeable bases, especially calcium and magnesium. Although dilute
sulfuric acid in sandy podsolic soils caused a significantly decreased pH of the leached
material, the amount of acid applied (not more than twice the yearly airborne supply over
southern Scandinavia) did not acidify soil as much as did nitrate fertilizer (Tamm et al.,
1977). Highly acidic rainfall, frequently with a pH less than 3.0, in combination with heavy
metal particulate fallout from smelters, has caused soils to become toxic to seedling survival
and establishment according to observations by Hutchinson and Whitby (1976). Very low soil
pH's are associated with mobility of toxic aluminum compounds in the soils. High acidity, high
sulfur, and heavy metals in the rainfall have caused fundamental changes in the structure of
soil organic matter. The sulfate and heavy metals were borne by air from the smelters in the
Sudbury area of Ontario and brought to earth by dry and wet deposition. Among the metals
deposited in rainfall and dustfall were nickel, copper, cobalt, iron, zinc, and lead. Most of
these metals are retained in the upper layers of soil, except in very acid or sandy soils.
The accumulation of metals is mainly an exchange phenomenon. Organic components of lit-
ter, humus, and soil may bind heavy metals as stable complexes (Tyler, 1972). The heavy metals
when bound may interfere with litter decay and nutrient cycling, and in this manner interfere
with ecosystem functioning (Tyler, 1972). Acidic precipitation, by altering the equilibria of
the metal complexes through mobilization, may have a negative effect upon the residence time
of the heavy metals in soil and litter (Tyler, 1972, 1977).
Biological processes in the soil necessary for plant growth can be affected by soil
acidification. Nitrogen fixation, decomposition of organic material, and mineralization,
especially of nitrogen, phosphorus and sulfur, might be affected (Abrahamsen and Dollard,
1979; Tamm et al., 1977; Malmer, 1976; Alexander, 1980). Nearly all of the nitrogen, most of
the phosphorus and sulfur as well as other nutrient elements in the soil are bound in organic
combination. In this form, the elements are largely or entirely unavailable for utilization by
higher plants (Alexander, 1980). It is principally through the activity of heterotrophic
microorganisms that nitrogen, phosphorus, and sulfur are made available to the autotrophic
higher plants. Thus, the microbial processes that lead to the conversion of the organic forms
of these elements to the inorganic state are crucial for maintaining plant life in natural or
agricultural ecosystems. The key role of these degradative processes is the fact that nitro-
gen is limiting for food production in much of the world and governs primary productivity in
many terrestrial habitats (Alexander, 1978).
Many, and probably most, microbial transformations in soil may be brought about by several
species. Therefore, the reduction or elimination of one population is not necessarily detri-
mental since a second population, not affected by the stress, may fill the partially or totally
vacated niche. For example, the conversion of organic nitrogen compounds to inorganic forms
is characteristically catalyzed by a number of species, often quite dissimilar, and a physical
XDSX7A/D 7-73 2-9-81
-------
*
or chemical perturbation affecting one of the species may not seriously alter the rate of the
conversion. On the other hand, there are a few processes that are in fact carried out, so far
as it is now known, by only a single species, and elimination of that species could have
serious consequences. Examples of this are the nitrification process, in which ammonium is
converted to nitrate, and the nodulation of leguminous plants, for which the bacteria are
reasonably specific according to the leguminous host (Alexander, 1980).
The nitrification process is one of the best indicators of pH stress because the
responsible organisms, presumably largely autotrophic bacteria, are sensitive both in culture
and in nature to increasing acidity (Dancer et al., 1973). Although nitrification will some-
times occur at pH values below 5.0, characteristically the rate decreases with increasing
acidity and often is undetectable much below pH 4.5. Limited data suggest that the process of
sulfate reduction to sulfide in soil is markedly inhibited below a pH of 6.0. (Connell and
Patrick, 1968) and studies of the presumably responsible organisms in culture attest to the
inhibition linked with the acid conditions (Alexander, 1978).
Blue-green algae have been found to be absent from acid soils even though there is both
adequate moisture and exposure to sunlight. Studies by Wodzinki et al. ( 1977) attest to the
sensitivity of these organisms to acidity. Inhibition of the rates of both C^ fixation and
nitrogen fixation was noted.
Studies concerned with the acidification of soil by nitrogen fertilizers or sulfur amend-
ments, as well as comparisons of the microbial populations in soils with dissimilar pH values,
attest to the sensitivity of bacteria to increasing hydrogen ion concentrations. Character-
istically, the numbers of these organisms decline, and not only is the total bacterial
community reduced in numbers, but individual physiological groups are also reduced (Alexander,
1980). The actinomycetes (taxonomically considered to be bacteria) also are generally less
abundant as the pH decreases, while the relative abundance of fungi increases, possibly due to
a lack of competition from other heterotrophs (Dancer et al., 1973). The pH of soil not only
influences the microbial community at large but also those specialized populations that
colonize the root surfaces (Alexander, 1980).
It is difficult to make generalizations concerning the effects of soil acidification on
microorganisms. Many microbial processes that are important for plant growth are clearly
suppressed as the pH declines; however, the inhibition noted in one soil at a given pH may not
be noted at the same pH in another soil (Alexander, 1980). The capacity of some microorganisms
to become acclimated to changes in pH suggests the need to study this phenomenon using environ-
ments that have been maintained at different pH values for some time. Typically the studies
have been done with soils maintained only for short periods at the greater acidity (Alexander,
1980). The consequences of increased acidity in the subterranean ecosystem are totally
unclear.
Adding nitrate and other forms of nitrogen from the atmosphere to ecosystems is an
integral function of the terrestrial nitrogen cycle. Higher plants and microorganisms can
XDSX7A/D 7-74 2-9-81
-------
assimilate the inorganic forms rapidly. The contribution of inorganic nitrogen in wet precipi-
tation (rain plus snow) is usually equivalent to only a few percent of the total nitrogen
assimilated annually by plants in terrestrial ecosystems; however, total nitrogen contribu-
tions, including organic nitrogen, in bulk precipitation (rainfall plus dry fallout) can be
significant, especially in unfertilized natural systems.
Atmospheric contributions of nitrate can range from less than 0.1 kg N/ha/yr in the North-
west (Fredricksen, 1972) to 4.9 kg N/ha/yr in the eastern United States (Likens et al., 1970;
Henderson and Harris, 1975). Inorganic nitrogen (ammonia-N plus nitrate-N) additions in wet
precipitation ranged from less than 0.5 kg/ha/yr to more than 3.5 kg/ha/yr in Junge's (1958)
study of rainfall over the United States. On the other hand, total nitrogen loads in bulk
precipitation range from less than 5 kg/ha/yr in desert regions of the West to more than 30
kg/ha/yr near barnyards in the Midwest. Total contributions of nitrogen from the atmosphere
commonly range from about 10 to 20 kg N/ha/yr for most of the United States (National Research
Council, 1978).
In comparison, rates of annual uptake by plants range from 11 to 125 kg N/ha/yr in eco-
systems selected from several bioclimatic zones (National Research Council, 1978). Since the
lowest additions are generally associated with desert areas where rates of uptake by plants
are low, and the highest additions usually occur in moist areas where plant uptake is high,
the contributions of ammonia and nitrate from rainfall to terrestrial ecosystems are equiva-
lent to about 1 to 10 percent of annual plant uptake. The typical additions of total nitrogen
in bulk precipitation, on the other hand, represent from about 8 to 25 percent of the annual
plant requirements in eastern deciduous and western coniferous forest ecosystems. Although
these comparisons suggest that plant growth in terrestrial ecosystems depends to a significant
extent on atmospheric deposition, it is not yet possible to estimate the importance of these
contributions by comparing them with the biological fixation and mineralization of nitrogen in
the soil. In nutrient-impoverished ecosystems, such as badly eroded abandoned croplands or
soils subjected to prolonged leaching by acidic precipitation, nitrogen additions from atmos-
pheric depositions are certainly important to biological productivity. In largely unperturbed
forests, recycled nitrogen from the soil organic pool is the chief source of nitrogen for
plants, but nitrogen to support increased production must come either from biological fixation
or from atmospheric contributions. It seems possible, therefore, that man-generated contribu-
tions could play a significant ecological role in a relatively large portion of the forested
areas near industrialized regions (National Research Council, 1978).
Sulfur, like nitrogen, is essential for optimal plant growth. Plants usually obtain
sulfur from the soil in the form of sulfate. The amount of mineral sulfur in soils is usually
low and its release from organic matter during microbial decomposition is a major source for
plants (Donahue et al., 1977). Another major source is the wet and dry deposition of atmos-
pheric sulfur (Donahue et al., 1977; Brady, 1974; Jones, 1975).
XDSX7A/D 7-75 2-9-81
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In agricultural soils crop residues, manure, irrigation water, and fertilizers and soil
amendments are important sources of sulfur. The amounts of sulfur entering the soil system
from atmospheric sources is dependent on proximity to industrial areas, the sea coast, and
marshlands. The prevailing winds and the amount of precipitation in a given region are also
important (Halsteand and Rennie, 1977). Near fossil-fueled power plants and industrial
installations the amount of sulfur in precipitation may be as much as 150 pounds per acre (168
kg/ha) or more (Jones, 1975). By contrast, in rural areas, based on the equal distribution of
sulfur oxide emissions over the coterminous states, the amount of sulfur in precipitation is
generally well below the average 15 pounds per acre (17 kg/ha). Approximately 5 to 7 pounds
per acre (7 to 8 kg/ha) per year were reported for Oregon in 1966 (Jones, 1975). Shinn and
Lynn (1979) have estimated that in the northeastern United States, the area where precipi-
tation is most acidic, approximately 5 x 10 tons of sulfate per year is removed by rain
(Brady, 1974). Hoeft et al. (1972) estimated the overall average sulfur as sulfate deposition
at 26 pounds of sulfur/acre per year (30 kg S/ha per year). Estimates for rural areas were 14
pounds of sulfur per acre per year (16 kg/ha/yr). Approximately 40 to 50 percent of the
sulfur additions occurred from November to February. Tabatabai and Laflin (1976) found that
SO.-S deposition in Iowa was greatest in fall and winter when precipitation was low. They also
estimated that the additions of sulfur by precipitation were the same for Ames in 1976 as were
reported for 1923, approximately 15 Ibs/acre. The average annual additions of sulfur by pre-
cipitation were similar to that reported for rural Wisconsin by Hoeft et al. (1972)
Experimental data have shown that even though plants are supplied with adequate soil sul-
fate they can absorb 25 to 35 percent of their sulfur from the atmosphere (Brady, 1974).
Particularly if the soil sulfur is low and atmosphere sulfur high, most of the sulfur required
by the plant can come from the atmosphere (Brady, 1974). Atmospheric sulfur would be of
benefit chiefly to plants growing on lands with a low sulfur content (Brezonik, 1975).
Tree species vary in their ability to utilize sulfur. Nitrogen and sulfur are biochem-
ically associated in plant proteins, therefore, a close relationship exists between the two in
plants. Apparently, nitrogen is only taken up at the rate at which sulfur is available. Pro-
tein formation is, therefore, limited by the amount of sulfur available (Turner and Lambert,
1980). Conifers accumulate as sulfate any sulfur beyond the amount required to balance the
available nitrogen. Protein formation proceeds at the rate at which nitrogen becomes avail-
able. Trees are not injured when sulfur is applied as sulfate rather than S02 (Turner and
Lambert, 1980).
When discussing the effects of acidic precipitation, or the effects of sulfates or
nitrates on soils, a distinction should be made between managed and unmanaged soils. There
appears to be general agreement that managed agricultural soils are less susceptible to the
influences of acidic precipitation than are unmanaged forest or rangeland soils. On managed
soils more than adequate amounts of lime are used to counteract the acidifying effects of
fertilizers in agricultural soils. Ammonium fertilizers, usually as ammonium sulfate
XDSX7A/D 7-76 2-9-81
-------
[(NH4)2S04] or ammonium nitrate, (NH4N03) are oxidized by bacteria to form sulfate (S042~)
and/or nitrate (N03) and hydrogen ions (H ) (Donahue et al. 1977; Brady, 1974). The release
of hydrogen ions into the soil causes the soil to become acidified. Hydrogen ions are also
released into the soil when plants take up mineral nutrients. Hence, substances (notably
various complexes of ammonium and sulfate ions), although of neutral pH, or nearly so, are
acidifying in their effects when they are taken up by plants or animals. Thus, the concept of
"acidifying precipitation" must be added to the concept of "acid precipitation."
The acidifying effects of fertilization or acidic precipitation is countered in managed
soils through the use of lime. Liming tends to raise the pH and thereby eliminate most major
problems associated with acidic soils (Donahue et al., 1977; Likens et al., 1977). Costs of
liming all natural soils would be prohibitive as well as extremely difficult to carry out.
Precipitation may add many chemicals to terrestrial, aquatic, and agricultural eco-
systems. In addition to sulfur and nitrogen, phosphorus and potassium are biologically most
important because they often are in limited supply in the soil (Likens et al., 1977). Other
chemicals of varying biological importance and varying concentration found in precipitation
over North America are the following: chlorine, sodium, calcium, magnesium, iron, nickel,
copper, zinc, cadmium, lead, manganese (Beamish, 1976; Hutchinson and Whitby, 1976; Brezonik,
1975), mercury (Brezonik, 1975), and cobalt (Hutchinson and Whitby, 1976). Rain over Britain
and the Netherlands, according to Gorham (1976), contained the following elements in addition
to those reported for North American precipitation: aluminum, arsenic, beryllium, cerium,
chromium, cesium, antimony, scandium, selenium, thorium, and vanadium. Again it is obvious
that many of these elements will be found in precipitation in highly industrialized areas and
will not be of biological importance until they enter an ecosystem where they may come into
contact with some form of life, as in the case of heavy metals in the waters and soils near
Sudbury, Ontario. Of chemical elements found in precipitation, magnesium, iron, copper, zinc,
and manganese are essential in small amounts for the growth of plants; however, at high con-
centrations these elements, as well as the other heavy metals, can be toxic to plants and
animals. Furthermore, the acidity of precipitation can affect the solubility, mobility, and
toxicity of these elements to the foliage or roots of plants and to animals or microorganisms
that may ingest or decompose these plants.
Wiklander (1979) has pointed out that based on the ion exchange theory, ion exchange
experiments, and the leaching of soil samples, the following conclusions can be drawn about
the acidifying effect on soils through the atmospheric deposition of mineral acids.
1. At a soil pH > 6.0 acids are fully neutralized by decomposition of CaCO, and other
unstable minerals and by cation exchange.
2. At soil pH < 5.5 the efficiency of the proton to decompose minerals and to replace
2+ 2+ + +
exchangeable Ca , Mg , K , and Na decreases with the soil pH. Consequently, the
acidifying effect of mineral acids on soils decreases, but the effect on the runoff
water increases in the very acid soils.
XDSX7A/D 7-77 2-9-81
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3. Salts of Ca2+, Mg2+, K+, and NH/ in the precipitation counteract the absorption of
protons and, in that way, the decrease of the base saturation. A proportion of the
acids percolate through the soil and acidify the runoff.
The sensitivity of various soils to acidic precipitation depends on the soil buffer
capacity and on the soil pH. Noncalcareous sandy soils with pH > 5 are the most sensitive to
acidification; however, acidic soils would be most likely to release aluminum.
Very acid soils are less sensitive to further acidification because they are already
adjusted by soil formation to acidity and are therefore more stable. In these soils easily
weatherable minerals have disappeared, base saturation is low, and the pH of the soil may be
less than that of precipitation. The low nutrient level is a crucial factor which limits pro-
ductivity in these soils. Even a slight decrease in nutrient status by leaching may have a
detrimental effect on plant yield (Wiklander, 1979). Fertilization appears to be the only
preventive measure.
In properly managed cultivated soils, acidic precipitation should cause only a slight
increase in the lime requirement, with the cost compensated for by the supply of sulfur,
nitrogen, magnesium, potassium, and calcium made available to plants (Wiklander, 1979).
7.3.2.2 Effects on vegetation. The atmosphere, as well as the soil, is a source of nutrients
for plants. Chemical elements reach the plant surface via wet and dry deposition. Nitrates
and sulfates are not the only components of precipitation falling onto the plant surface.
Other chemical elements (cadmium, lead, zinc, manganese), at least partially soluble in water,
are deposited on the surface of vegetation and may be assimilated by it, usually through the
leaves. An average raindrop deposited on trees in a typical forest washes over three tiers of
foliage before it reaches the soil. The effects of acidic precipitation may be beneficial or
deleterious depending on its chemical composition, the species of plant on which it is
deposited, and the physiological condition and maturity of the plant (Galloway and Cowling,
1978). Substances accumulated on the leaf surfaces strongly influence the chemical composi-
tion of precipitation not only at the leaf surface, but also when it reaches the forest floor.
The chemistry of precipitation reaching the forest floor is considerably different from that
collected above the forest canopy or a ground level where the canopy has no influence.
(Lindberg et al. 1979). Except for the hydrogen ion (H+) the mean concentrations of all
elements (lead, manganese, zinc and cadmium) studied in the Walker Branch Watershed in
Tennessee found by Lindberg et al. (1979) to be present in greater amounts in the throughfall
than in incident rain. The presence of trace elements was more variable than that of the
sulfate and hydrogen ions. Throughfall with a pH 4.5 appeared to be a more dilute solution
of sulfuric acid than rain not influenced by the forest canopy. The solution was found to
contain a relatively higher concentration of alkaline earth salts of sulfate and nitrate as
well as a somewhat higher concentration of trace elements (Lindberg et al. 1979).
Lee and Weber (1980) studied the effects of sulfuric acid rain on two model hardwood
forests. The experiment, conducted under controlled field conditions, consisted of the
XDSX7A/D 7-78
2-9-81
-------
application of simulated sulfuric acid rain (pH values of 3.0, 3.5, and 4.0), and a control
rain of pH 5.6 to the two model forest ecosystems for a duration of 3 and 1/2 years. Rainfall
applications approximated the annual amounts of areas in which sugar maple and red alder
communities normally occur.
In evaluating the results of the study, the authors conclude that a well developed forest
canopy and litter layer can increase the pH and concentration of bases (i.e., calcium and
magnesium) in rainwater. Such conditions would tend to decrease the acidification rate of
forest soils by acid rain. However, as bases are continually leached from the soil column
these cations could eventually be lost from the ecosystem and unavailable to influence the
acidification reactions. Changes in the ionic and pH balance of forest systems may impact the
productivity of forests through acidity-induced changes in the nutrient cycling process,
decomposition, reproduction, tree growth, and the structure of forest systems.
The additions of hydrogen, sulfate and nitrate ions to soil and plant systems have both
positive and negative effects. It has generally been assumed that the free hydrogen ion con-
centration in acidic precipitation is the component that is most likely to cause direct, harm-
ful effects on vegetation (Jacobson, 1980). Experimental studies support this assumption;
however, to date, there are no confirmed reports of exposure to ambient acidic precipitation
causing foliar symptoms on field grown vegetation in the continental United States (Jacobson,
1980) and Canada (Linzon personal communication).
7.3.2.2.2.1 Direct effects on vegetation. Hydrogen ion concentrations equivalent to that
measured in more acidic rain events (< pH 3.0) have been observed experimentally to cause
tissue injury in the form of necrotic lesions to a wide variety of plant species under green-
house and laboratory conditions. This visible injury has been reported as occurring between pH
3.0 and 3.6 (Shriner, 1980). The various types of direct effects which have been reported are
shown in Table 7-10. Such effects must be interpreted with caution because the growth and
morphology of leaves on plants grown in greenhouses frequently are atypical of field condi-
tions (Shriner, 1980).
Small necrotic lesions, the most common form of direct injury, appear to be the result of
the collection and retention of water on plant surfaces and the subsequent evaporation of
these water droplets once a lesion occurs. The depression formed by the lesion further
enhances the collection of water. A large percentage of the leaf area may exhibit lesions
after repeated exposures to simulated acid rain at pH concentrations of 3.1, 2.7, 2.5 and 2.3
(Evans et al. 1977a, 1977b). In leaves injured by simulated acidic rain, collapse and distor-
tion of epidermal cells on the upper surface is frequently followed by injury to the palisade
cells and ultimately both leaf surfaces are affected (Evans et al., 1977b). Evans et al.
(1978) using six clones of Populus spp. hybrids found that leaves that had just reached full
expansion were more sensitive to simulated acid rain at pH 3.4, 3.1, 2.9, and 2.7 than were
unexpanded or those which were fully expanded. On two of the clones, gall formation due to
abnormal cell proliferation and enlargement occurred. Other effects attributed to simulated
XDSX7A/D 7-79 2-9-81
-------
TABLE 7-10. TYPES OF DIRECT, VISIBLE INJURY REPORTED IN RESPONSE TO ACIDIC WET DISPOSITION
Injury Type
Species
pH Range
Reference
Remarks
00
o
Pitting, curl
shortening, death
1-mm necrotic lesions,
premature abscission
Cuticular erosion
Chlorosis
(A) small, shallow
circular depressions:
slight chlorosis
(B) larger lesions,
chlorosis always present
palisade collapse
(C) 1-mm necrotic lesions
general distortion
(D) 2-mm bifacial necrosis
sue to coalescence of
smaller lesions, total
tissue collapse.
Wrinkled leaves, excessive
adventitious budding,,, pre-
mature abscission
Yellow birch
Kidney bean,
soybean,
loblolly pine,
E. white pine,
willow oak
Willow oak
Sunflower,
bean
Sunflower,
bean
Sunflower,
bean
Sunflower,
bean
Sunflower,
bean
Bean
2.3-4.7 Wood and Bormann (1974)
3.2
3.2
Shriner et al. (1974)
Shriner (1978a),
Lang et al. (1978)
2.3-5.7 Evans et al. (1977)
2.7 Evans et al. (1977)
2.7 Evans et al. (1977)
2.7 Evans et al. (1977)
2.7 Evans et al. (1977)
1.5-3.0 Ferenbaugh (1976)
More frequent near
veins. (A) - (D)
represent sequential
stages of lesion
development, through
time, up to 72 h (one
6-min rain event daily
for 3 d)
-------
TABLE 7-10 (Continued).
oo
Injury Type
Incipient bronzed spot
Bifacial necrotic pitting
Species
Bean
Bean
pH Range
2.0-3.0
2.0-3.0
Reference
Hindawi et al.
Hindawi et al.
(1980)
(1980)
Remarks
After first
After 24 h (
few ho
^report
Necrotic lesions,
premature abscission
Marginal and tip necrosis
Galls, hypertrophy,
hyperplasia
Dead leaf cells
Necrotic lesions
E. white pine,
scotch pine,
spinach,
sunflower,
bean
Bean, poplar,
soybean, ash
birch, corn,
wheat
Hybrid poplar
Soybean
Citrus
2.6-3.4 Jacobson and
van Leuken (1977)
Submicron Lang et al. 1978
H2S04
aerosol
2.7-3.4 Evans et al. (1978)
3.1 Irving (1979)
0.5-2.0 Heagle et al. (1978)
pooling of drops =
more injury)
Injury associated with
droplet location
within 24-48 h.
Shriner, 1980
-------
acid rain include the modification of the leaf surface, e.g. epicuticular waxes, and altera-
tion of physiological processes such as carbon fixation and allocation.
Lee et al. (1980) studied the effects of simulated acidic precipitation on crops.
Depending on the crop studied, they reported positive, negative or no effects on crop yield
when exposed to sulfuric acid rain at pH volumes of 3.0, 3.5 and 4.0 when compared to crops
exposed to a control rain of pH 5.6. The yield of tomatoes, green peppers, strawberries,
alfalfa, orchard grass and timothy were stimulated. Yields of radishes, carrots, mustard
greens and broccoli were inhibited. Potatoes were ambiguously affected except at pH 3.0 where
their yield as well as that of beets was inhibited. Visible injury of tomatoes could possibly
have decreased their market value. In sweet corn, stem and leaf production was stimulated,
but no statistically significant effects on yield were observed for 15 other crops. Results
suggest that the possibility of yield's being affected by acid rain depends on the portion of
the plant being utilized as well as the species. Plants were regularly examined for foliar
injury associated with acid rain. Of the 35 cultivars examined, the foliage of 31 was injured
at pH 3.0; 28 at pH 3.5; and 5 at pH 4.0. Foliar injury was not generally related to effects
on yield. However, foliar injury of swiss chard, mustard greens, and spinach was severe
enough to adversely affect marketability. These results are from a single growing season and
therefore considered to be preliminary.
Studies indicate that wet deposition of acidic or acidifying substances may result in a
range of direct or indirect effects on vegetation. Environmental conditions before, during
and after a precipitation event affect the responses of vegetation. Nutrient status of the
soil, plant nutrient requirements, plant sensitivity and growth stage and the total loading or
deposition of critical ions e.g. H , NO., and SO. all play a role in determining vegetational
response to acidic precipitation.
Wettability of leaves appears to be an important factor in the response of plants to acid
deposition. This has been demonstrated in the work of Evans et al. (1977), Jacobson and Van
Leuken (1977), and Shriner (1978a), who variously report a threshold of between pH 3.1 and 3.5
for development of foliar lesions on beans. The cultivars of Phaseolus vulgaris L. used in
the above studies are all relatively non-waxy and therefore fairly easily wettable. By
comparison, studies with the very waxy leaves of citrus (Heagle et al., 1978) reported a
threshold for visible symptoms to be near pH 2.0. Waxy leaves apparently minimize the contact
time for the acid solutions, thus accounting for the <400X increase in H+ ion concentration
required to induce visible injury. Table 7-11 summarizes the thresholds, including range,
species sensitivity, concentration, and time, for visible injury associated with experimental
studies of wet deposition of acidic substances.
Leaching of chemical elements from exposed plant surfaces is an important effect rain,
fog, mist, and dew have on vegetation. Substances leached include a great diversity of
materials. All of the essential minerals, ami no acids, carbohydrate growth regulators, free
sugars, pectic substances, organic acids, vitamins, alkaloids, and alleopathic substances are
XDSX7A/D 7-82 2-9-81
-------
TABLE 7-11. THRESHOLDS FOR VISIBLE INJURY AND GROWTH EFFECTS ASSOCIATED WITH EXPERIMENTAL
STUDIES OF WET DEPOSITION OF ACIDIC SUBSTANCES (AFTER JACOBSON, 1980a,b)
oo
CO
Effect
Foliar lesions, decrease
in growth
Foliar aberrations,
decrease in growth
Foliar lesions
Foliar lesions
Foliar lesions
Foliar lesions
Foliar symptons, no
reduced growth
Increased growth,
i ncreased/decreased
nutrient content
Reduced growth
Reduced yield
Reduced growth
Reduced yield
Species
Yellow birch
Bean
Bean, sunflower
Bean
Hybrid poplar
Sunflower
Soybean
Lettuce
Pinto bean
Pinto bean
Soybean
Soybean
Threshold
pH 3.1
pH 2.5
pH 3.1
pH 3.2
pH 3.4
pH 3.4
pH 3.0
pH 3.0, 3.2
pH 3.1
pH 2.7
pH 3.1
pH 2.5
Reference
Wood and Bromann (1974)
Ferenbaugh (1976)
Evans et al. (1977a)
Shriner (1978a)
Evans et al. (1978)
Jacobson and
van Leuken (1977)
Jacobson (1980b)
Jacobson (1980b)
Jacobson (1980b)
Remarks
greenhouse
greenhouse
greenhouse
greenhouse
greenhouse
greenhouse
greenhouse
greenhouse (varied
with S04 & N03")
greenhouse
-------
TABLE 7-11 (Continued).
I
00
-pi
Effect
Species
Threshold Reference Remarks
Increased yield
Foliar symptoms
Reduced growth
Reduced yield
Reduced quality
No foliar symptoms, or
effects on growth
No foliar symptoms, but
a) decreased growth, yield
b) increased yield
No effect on growth, yield
Reduced quality
Soybean
Tomato
Tomato
Tomato
Tomato
Soybean
Soybean
Soybean
Soybean
Tomato
Tomato
pH 3.1
pH 3.0 Jacobson (1980b) greenhouse
pH 3.0
pH 3.0
pH 3.0
pH 3.1 Irving (1979) field
pH 2.8 Jacobson (1980b) field, low ozone
pH 2.8 field, high ozone
pH 2.8 field, low ozone
pH 3.0 Jacobson (1980b) field
pH 3.0 field
Highest pH to elicit a negative response, or lowest pH to elicit a positive response
Shriner, 1980.
-------
among the materials which have been detected in plant leachates (Tukey, 1970). Many factors
influence the quantity and quality of the substances leached from foliage. They include fac-
tors associated directly with the plant as well as those associated with the environment. Not
only are there differences among species with respect to leaching, but individual differences
also exist among individual leaves of the same crop and even the same plant, depending on the
physiological age of the leaf. Young, actively growing tissues are relatively immune to
leaching of mineral nutrients and carbohydrates, while mature tissue which is approaching
senescence is very susceptible. The stage of plant development, temperature, and rainwater
falling on foliage and running down plant stems or tree bark influences leaching. Rainwater,
which naturally has a pH of about 5.6, washing over vegetation may become enriched with
o
substances leached from the tissues (Nihlgard, 1970).
Leaching of organic and inorganic materials from vegetation to the soil is part of the
normal functioning of terrestrial ecosystems. The nutrient flow from one component of the
ecosystem to another is an important phase of nutrient cycling (Comerford and White, 1977;
Eaton et al., 1973). Plant leachates have an effect upon soil texture, aeration, permea-
bility, and exchange capacity. Leachates, by influencing the number and behavior of soil
microorganisms, affect soil-forming processes, soil fertility, and susceptibility or immunity
of plants to soil pests and plant-chemical interactions (Tukey, 1970).
It has been demonstrated under experimental conditions that precipitation of increased
acidity can increase the leaching of various cations and organic carbon from the tree canopy
(Abrahamsen et al., 1976; Wood and Bormann, 1975). Foliar losses of potassium, magnesium, and
calcium from bean plants and maple seedlings were found to increase as the acidity of an arti-
ficial mist was increased. Below a pH of 3.0 tissue damage occurred; however, significant
increases in leaching were measured at pH 3.3 and 4.0 with no observable tissue damage (Wood
and Bormann, 1975). Hindawi et al. (1980) also noted that as the acidity of acid mist
increased so did the foliar leaching of nitrogen, calcium, phosphorous, and magnesium.
Potassium concentrations were not affected, while the concentration of sulfur increased.
Abrahamsen and Do!lard (1979), in experiments using Norway spruce (Picea abies L. Karst),
observed that despite increased leaching under the most acid treatment, there was no evidence
of change in the foliar cation content. Wood and Bormann (1977), using Eastern white pine
(Pinus strobus L.), also noted no significant changes in calcium, magnesium or potassium con-
tent of needles. Tukey (1970) states that increased leaching of nutrients from foliage can
accelerate nutrient uptake by plants. No injury will occur to the plants as long as roots can
absorb nutrients to replace those being leached; however, injury could occur if nutrients are
in short supply. To date, the effects, if any, of the increased leaching of substances from
vegetation by acidic precipitation remain unclear.
Some experimental evidence suggests that acidic solutions affect the chlorophyll content
of leaves and the rate of photosynthesis. Sheridan and Rosenstreter (1973) reported marked
reduction of photosynthesis in a moss exposed to increasing H ion concentrations. Sheridan
XDSX7A/D 7-85 2-9-81
-------
and Rosenstreter (1973), Ferengaugh (1976), and Hindawi et al. (1980) reported reduced chloro-
phyll content as a result of tissue exposure to acid solutions. In the case of Ferenbaugh
(1976), however, the significant reductions in chlorophyll in the leaves of Phaseolus vulgaris
at pH 2.0 were associated with large areas of necrosis. A significant aspect of this study
was the loss of capacity by the plant to produce carbohydrates. The rate of respiration in
these plants showed only a slight but significant increase while the rate of photosynthesis at
pH 2.0 increased nearly fourfold as determined by oxygen evolution. Ferenbaugh concluded that
due to a reduction in biomass accumulation and sugar and starch concentrations, photophosphory-
lation in the treated plants was in some way being uncoupled by the acidic solutions.
Irving (1979) reported a higher chlorophyll content and an increase in the rate of photo-
synthesis in field-grown soybeans exposed to simulated rain at pH 3.1. She attributed the
increases to improved nutrition due to the sulfur and nitrogen components of the simulated
acid rain overcoming any negative effects.
Vegetation is commonly exposed to gaseous phytotoxicants such as ozone and sulfur dioxide
at the same time as acidic precipitation. Little information is available upon which to
evaluate the potential for determining the effects of the interaction of wet-and dry-deposited
pollutants on vegetation. Preliminary studies by Shriner (1978b), Irving (1979), and Jacobson
et al. (1980) suggest that interactions may occur. Irving (1979) found that simulated acid
precipitation at pH of 3.1 tended to limit the decrease in photosynthesis observed when field-
grown soybeans were exposed 17 times during the growing season to 0.19 ppm of S0?. Shriner
(1978b), however, reported no significant interaction between multiple exposure to simulated
rain at p 4.0 and four SOp exposures (3 ppm peak for 1 hr.) upon the growth of bush beans.
Shriner (1978b) also exposed plants to 0.15 ppm ozone (4 3-hour exposures) in between 4 weekly
exposures to rainfall of pH 4.0, and observed a significant growth reduction at the time of
harvest. Jacobson et al. (1980), using open-top exposure chambers with field-grown soybeans,
compared growth and yield between three pH levels of simulated rain (pH 2.8, 3.4, and 4.0) and
two levels of ozone (<0.03 and <0.12 ppm). Results demonstrated that ozone depressed both
growth and yield of soybeans with all three rain treatments, but that the depression was
greatest with the most acidic rain. Ozone concentrations equal to or greater than those used
in the studies are common in most areas of the northeastern United States where acidic deposi-
tion is a problem (Jacobson et al. , 1980); therefore, the potential for ozone-acidic deposi-
tion interaction is great.
Shriner (1978a) studied the effect of acidic precipitation on host-parasite interactions.
Simulated acid rain with a pH of 3.2 inhibited the development of bean rust and production of
telia (a stage in the rust life cycle) by the oak-leaf rust fungus Cronartium fusiforme. It
also inhibited reproduction of root-knot nematodes and inhibited or stimulated development of
halo blight of bean seedlings depending on the time in the disease cycle during which the
simulated acid rain was applied. The effects which inhibited disease development could
XDSX7A/D 7-86 2-9-81
-------
result in a net benefit to plant health. Shriner (1976, 1980) also observed that root nodula-
tion by Rhizobium on common beans and soybeans was inhibited by the simulated acid rain, suggest-
ing a potential for reduced nitrogen fixation by legumes so effected.
Plants such as mosses and lichens are particularly sensitive to changes in precipitation
chemistry because many of their nutrient requirements are obtained directly through precipita-
tion. These plant forms are typically absent from regions with high chronic S02 air pollution
and acidic precipitation (Denison et al., 1977; Sheridan and Rosenstreter, 1973). Gorham
(1976) and Giddings and Galloway (1976) have written reviews concerning this problem. Most
investigations on the effects of air pollution on epiphytes have dealt with gaseous pollut-
ants. Very few studies have considered acidic precipitation. Denison et al. (1977), however,
did observe that the nitrogen-fixing ability of the epiphytic lichen Lobaria oregana was re-
duced when treated with simulated rainfall with a pH of 4.0 and below. Investigations con-
cerning the effects of acidic precipitation on epiphytic microbial populations are very few
(Abrahamsen and Dollard, 1979).
Limited fertilization could occur in the bracken fern Pteridium aquilinum under condi-
tions of acidic precipitation (pH and sulfate concentrations) that prevail in the northeastern
United States. Evans and Bozzone (1977), using buffered solutions to simulate acidic precipi-
tation, observed that flagellar movement of sperm was reduced at pH levels below 5.8. Fertili-
zation was reduced after exposure to pH's below 4.2. Sporophyte production was also reduced
by 50 percent at pH levels below 4.2 when compared to 5.8. Addition of sulfate (86 mM)
decreased fertilization at least 50 percent at all pH values observed. In another study,
Evans and Bozzone (1978) observed that both sperm motility and fertilization in gametophytes
of Pteridium aquilinum were reduced when anions of sulfate, nitrate, and chloride were added
to buffered solutions.
Sulfur and nitrogen in precipitation have been shown to play an important role in vegeta-
tional response to acidic deposition. Jacobson et al. (1980) investigated the impact of simu-
lated acidic rain on the growth of lettuce at acidities of pH 5.7 and 3.2. At pH 3.2, solu-
tions were compared with NO,:SO, mass ratios of 20:1, 2:1, and 1:7.5. The high nitrate at pH
3.2 showed no difference from the treatments controls at pH 5.7 for those growth parameters
(root dry weight, apical leaf dry weight) that responded to treatment; however, the results
were significantly less than those from the low nitrogen, high sulfur, treatment. These
observations suggest that sulfur was possibly a limiting factor in the nutrition of these
plants, with the result that the plant response to sulfur overwhelmed the hydrogen ion effect.
Other studies also have cited the beneficial effects of simulated acidic precipitation.
Irving and Miller (1978) observed that an acidic simulant had a positive effect on producti-
vity of field-grown soybeans as reflected by seed weight. Increased growth was attributed to
a fertilizing effect from sulfur and nitrogen delaying senescence. Irving and Miller (1978),
in the same study, also exposed soybeans to S02 and acidic precipitation. No visible injury
was apparent in any of the plots; however, a histological study revealed significant increases
XDSX7A/D 7-87 2-9-81
-------
in the number of dead mesophyll cells in all plots when compared to the control. The propor-
tion of dead mesophyll cells of plants exposed to acid rain and S02 combined was more than
additive when compared to the effects of each taken singly. Wood and Bormann (1977) reported
an increase in needle length and the weight of seedlings of Eastern white pine with increasing
acidity of simulated precipitation where sulfuric and nitric acid were used to acidify the
mist. Increased growth was attributed to increased N03" application. Abrahamson and Dollard
(1979) also presented data suggesting positive growth responses in forest tree species result-
ing from nitrogen and sulfur in simulated rain. Simulated acidic precipitation was observed
to increase the growth of Scots pine saplings in experiments conducted in Norway. Saplings in
plots watered with acid rain of pH 3.0, 2,5. and 2.0 grew more than the control plots. The
application of acid rain increased the nitrogen and sulfur content of the needles. As the
acidity of the artificial rain was adjusted using sulfuric acid only, the increased growth was
probably due to increased nitrogen mineralization and uptake. Turner and Lambert (1980)
reported evidence indicating a positive growth response in Monterey pine from the deposition
of sulfur in ambient precipitation in Australia.
Acidifying forest soils that are already acid by acidic precipitation or air pollutants
is a slow process. Growth effects probably could not be detected for a long time. To iden-
tify the possible effects of acidification on poor pine forests, Tamm et al. (1977) conducted
experiments using 50 kg and 100 kg of sulfur per hectare as dilute sulfuric acid (0.4 percent)
applied annually with and without NPK (nitrogen, phosphorous, potassium) fertilizer. Nitrogen
was found to be the limiting factor at both experimental sites. Acidification produced no
observable influence on tree growth. Lysimeter and soil incubation experiments conducted at
the same time as the experiments described above suggest that even moderate additions of sul-
furic acid or sulfur to soil affect soil biological processes, particularly nitrogen turnover.
The soil incubation studies indicated that additions of sulfuric acid increased the amount of
mineral nitrogen but lowered the amount of nitrate.
Soil fertility may increase as a result of acidic precipitation as nitrate and sulfate
ions, common components of chemical fertilizers, are deposited; however, the advantages of
such additions are possibly short-lived as depletion of nutrient cations through accelerated
leaching could eventually retard growth (Wood, 1975). Laboratory investigations by Overrein
(1972) have demonstrated that leaching of potassium, magnesium, and calcium, all important
plant nutrients, is accelerated by increased acidity of rain. Field studies in Sweden corre-
late decreases in soil pH with increased additions of acid (Oden, 1972).
Major uncertainty in estimating effects of acid rain on forest productivity is the capac-
ity of forest soils to buffer against leaching by hydrogen ions. Forest canopies have been
found to filter 90 percent of the hydrogen ions from rain (pH 4.0) falling on the landscape
during the growing season (Eaton et al., 1973). As a result, solutions reaching the forest
floor are less acidic (pH 5.0). Mayer and Ulrich (1977), however, point out that their studies
XDSX7A/D 7-88 2-9-81
-------
*
suggest that for most elements the addition by precipitation (wetfal1 plus dryfall) to the
soil beneath the tree canopy is considerably larger than that by precipitation to the canopy
surface as measured by rain gauges on a non-forested area. The leaching of metabolites,
mainly from leaf surfaces, and the washing out from leaves, branches, and stems of airborne
particles and atmospheric aerosols intercepted by trees from the atmosphere, are suggested as
the reason for the mineral increase.
Forest ecosystems are complicated biological organizations. Acidic precipitation will
cause some components within the ecosystem to respond even though it is not possible at pre-
sent to evaluate the changes that occur. The impact of the changes on the ecosystem can only
be determined with certainty after the passage of a long period of time.
7.3.2.3 Effects on Human Health—One effect of acidification that is potentially of concern
to human health is the possible contamination by toxic metals of edible fish and of water sup-
plies. Studies in Sweden (Landner and Larsson, 1975; Turk and Peters, 1977), Canada (Tomlin-
son, 1979; Brouzes et al., 1977), and the United States (Tomlinson, 1979) have revealed high
mercury concentrations in fish from acidified regions. Methylation of mercury to monomethyl
mercury occurs at low pH while dimethyl mercury forms at higher pH (Fagestrom and Jernelo'v,
1972). Monomethyl mercury in the water passing through the gills of fish reacts with thiol
groups in the hemoglobin of the blood and is then transferred to the muscle. Methyl mercury
is eliminated very slowly from fish; therefore, it accumulates with age.
Tomlinson (1979) reports that in the Bell River area of Canada precipitation is the
source of mercury. Both methyl mercury and inorganic mercury were found in precipitation.
The source of mercury in snow and rain was not known at the time of the study.
Zinc, manganese, and ajuminum concentrations also increase as the acidity of lakes
increases (Schofield, 1976b). The ingestion of fish contaminated by these metals is a dis-
tinct possibility.
Another human health aspect is the possibility that, as drinking-water reservoirs acidify,
owing to acidic precipitation, the increased concentrations of metals may exceed the public-
health limits. The increased metal concentrations in drinking water are caused by increased
watershed weathering and, more important, increased leaching of metals from household plumbing.
Indeed, in New York State, water from the Hinckley Reservoir has acidified to such an extent
that "lead concentrations in water in contact with household plumbing systems exceed the maxi-
mum levels for human use recommended by the New York State Department of Health." (Turk and
Peters, 1977) The lead and copper concentrations in pipes which have stood over night (U) and
those in which the water was used (F) are depicted in Table 7-12 (see following page).
7.3.2.4 Effects of Acidic Precipitation on Materials—Acidic precipitation can damage the
abiotic as well as the biotic components of an ecosystem. Of particular concern in this sec-
tion are the deteriorative effects of acidic precipitation on materials and cultural artifacts
of manmade ecosystems. At present in most areas, the dominant factor in the formation of
acidic precipitation is sulfur, usually as sulfur dioxide (Likens, 1976; Cowling and Dochinger,
1978). Because of this fact, it is difficult to isolate the effect of acidic precipitation
XDSX7A/D 7-89 2-9-81
-------
TABLE 7-12. LEAD AND COPPER CONCENTRATION AND pH OF WATER FROM PIPES
CARRYING OUTFLOW FROM HINCKLEY BASIN AND HANNS AND STEELE CREEK BASIN,
NEAR AMSTERDAM, NEW YORK
Col lection site
and date
Hi nek ley Dam
Nov. 21, 1974
Nov. 21, 1974
Nov. 7, 1974
Nov. 7, 1974
Oct. 1, 1974
Oct. 1, 1974
Aug. 15, 1974
Aug. 15, 1974
Amsterdam
Jan. 6, 1975
Jan. 6, 1975
Pipe ,
condition
U
F
U
F
U
F
U
F
U
F
Copper
(ug/i)
600
20
460
37
570
30
760
40
2900
80
Lead
(ug/1 )
66
2
40
6
52
5
88
2
240
3
PH
—
7.4
6.3
6.3
6.8
7.1
6.3
6.3
4.5
5.0
U, unflushed, (water stands in pipes all night); F, flushed
From Turk and Peters (1979)
from changes induced by sulfur pollution in general. (The effects of sulfur oxides on mate-
rials are discussed in Chapter 10.) High acidity promotes corrosion because the hydrogen ions
act as a sink for the electrons liberated during the critical corrosion process (Nriagu,
1978). Precipitation as rain affects corrosion by forming a layer of moisture on the surface
of the material and by adding hydrogen (H+) and sulfate (SO2') ions as corrosion stimulators.
Rain also washes out the sulfates deposited during dry deposition and thus serves a useful
function by removing the sulfate and stopping corrosion (Kucera, 1976). Rain plays a critical
role in the corrosive process because in areas where dry deposition predominates the washing
effect is greatest, while in areas where the dry and wet deposition processes are roughly
equal, the corrosive effect is greater (Kucera, 1976). The corrosion effect, particularly of
certain metals, in areas where the pH of precipitation is very low may be greatly enhanced by
that precipitation (Kucera, 1976). In a Swedish study the sulfur content of precipitation,
XDSX7A/D 7-90 2-9-81
-------
* 2
expressed as meq/m per year, was found to correlate closely with the corrosion rate of steel.
The metals most likely to be corroded by precipitation with a low pH are those whose corrosion
resistance may be ascribed to a protective layer of basic carbonates, sulfates, or oxides, as
used on zinc or copper. The decrease in pH of rainwater to 4.0 or lower may accelerate the
dissolution of the protective coatings (Kucera, 1976).
Materials reported to be affected by acidic precipitation, in addition to steel, are:
copper materials, linseed oil, alkyd paints on wood, antirust paints on steel, limestone,
sandstone, concrete, and both cement-lime and lime plaster (Cowling and Dochinger, 1978).
Stone is one of the oldest building materials used by man and has traditionally been con-
sidered one of the most durable because structures such as the pyramids, which have survived
since antiquity, are made of stone. What is usually forgotten is that the structures built
with stone that was not durable have long since disappeared (Sereda, 1977).
Atmospheric sulphur compounds (mainly sulfur dioxide, with subsidiary amounts of sulfur
trioxide and ammonium sulfate) react with the carbonates in limestone and dolomites, calcar-
eous sandstone and mortars to form calcium sulfate (gypsum). The results of these reactions
are blistering, scaling, and loss of surface cohesion, which in turn induces similar effects
in neighboring materials not in themselves susceptible to direct attack (Sereda, 1977).
Sulfates have been implicated by Winkler (1966) as very important in the disintegration
of stone. The surface flaking on the Egyptian granite obelisk (Cleopatra's Needle) in Central
Park, New York is cited as an example. The deterioration occurred within two years of its
erection in 1880.
A classic example of the effects of the changing chemical climate on the stability of
stone is the deterioration of the Madonna at Herten Castle, near Recklinghausen, Westphalia in
Germany. The sculpture of porous Baumberg sandstone was erected in 1702. Pictures taken of
the Madonna in 1908 shows slight to moderate damage during the first 206 years. The features
of the Madonna—eyes, nose, mouth and hair—are readily discernable. In pictures taken in
1969 after 267 years, no features are visible (Cowling and Dochinger, 1978).
It is not certain in what form sulfur is absorbed into stone, as a gas (SOO forming sul-
furous or sulfuric acid or whether it is deposited in rain. Rain and hoarfrost both contain
sulfur compounds. Schaffer (1932) compared the sulfate ion in both rain and hoarfrost at
Heachingley, Leeds, England in 1932 (Table 7-13) and showed that the content of hoarfrost was
approximately 7 times greater than rain. Wet stone surfaces unquestionably increase the con-
densation or absorption of sulfates. Stonework kept dry and shielded from rain, condensing
dew, or hoarfrost will be damaged less by S02 pollution than stone surfaces which are exposed
(Sereda, 1977).
Acid rain may leach ions from stonework just as acidic runoff and ground water leaches
ions from soils or bedrock; however, at the present time it is not possible to attribute the
deleterious effects of atmospheric sulfur pollution to specific compounds.
XDSX7A/D 7-91 2-9-81
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TABLE 7-13. COMPOSITION OF RAIN AND HOARFROST AT HEADINGLEY, LEEDS
Average rain Hoarfrost
parts per million parts per million
Suspended matter
Tar
Ash
Acidity
Sulphur as S03
Sulphur as S0?
Total sulphur
Chlorine
Nitrogen as NH3
Nitrogen as N?0r
Nitrogen as albuminoid
115
15
28
1.9
22
5.7
27.7
7.3
1.98
0.196
0.434
4620
158
67
102.9
148
41.0
189.0
94.6
8.57
0.0
1.618
Schaffer (1932)
Microbial action can also contribute to the deterioration of stone surfaces. Tiano et
al. (1975) isolated large numbers (250 to 20,000 cells per gram) of sulfate-reducing bacteria
from the stones of two historical buildings of Florence, Italy. The majority of the bacteria
belonged to the genus Thiobacillus. This genus of chemosynthetic aerobic microorganisms
oxidizes sulfide, elemental sulfur, and thiosulphate to sulfate to obtain energy (Anderson,
1978). Limestone buildings and particularly mortar used in the construction of brick and
stone buildings are particularly susceptible to when Thiobacillus can convert reduced forms of
sulfur to sulfuric acid. Sulfate in acidic precipitation as well as other sulfur compounds
deposited in dry deposition could permit the formation of sulfur compounds utilizable by micro-
organisms. (For more information concerning the effects of sulfur oxides on materials, please
consult Chapter 10).
7.4 ASSESSMENT OF SENSITIVE AREAS
7.4.1 Aquatic Ecosystems
Why do some lakes become acidified by acidic precipitation and others not? What deter-
mines susceptibility? Are terrestrial ecosystems likely to be susceptible; if so, which ones?
XDSX7A/D 7-92 2-9-81
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The sensitivity of lakes to acidification is determined by: (1) the acidity of both wet
deposition (precipitation) and dry deposition, (2) the hydrology of the lake, (3) the soil
system, geology, and canopy effects, (4) the surface water. Given acidic precipitation, the
soil system and associated canopy effects are most important. The hydrology of lakes includes
the sources, amounts, and pathways of water entering and leaving a lake. The capability of a
lake and its drainage basin to neutralize acidic contributions as well as the mineral content
of its surface water is largely governed by the composition of the bedrock of the watershed.
The chemical weathering of the watershed strongly influences the salinity (ionic composition)
and the alkalinity (hardness and softness) of the surface water of a lake (Wetzel, 1975;
Wright and Gjessing, 1976; Wright and Henriksen, 1978). The cation exchange capacity and wea-
thering rate of the watershed and the alkalinity of the surface water determine the ability of
the system to neutralize the acidity of precipitation.
Lakes vulnerable to acidic precipitation have been shown to have watersheds whose geolo-
gical composition is highly resistant to chemical weathering (Wright and Gjessing, 1976;
Galloway and Cowling, 1978; Wright and Henriksen, 1978). In addition, the watersheds of the
vulnerable lakes usually have thin, poor soils and are poorly vegetated. The cation exchange
capacity of such soils is low and, therefore, its buffering capacity is low (Schofield, 1979;
Wright and Henriksen, 1978).
Wright and Henriksen (1978) point out that the chemistry of Norwegian lakes could be
accounted for primarily on the basis of bedrock geology. They examined 155 lakes and observed
that 59 of them lay in granite or felsic gneiss basins. Water in these lakes was low in most
major ions and had low electrical conductivity. [The fewer the minerals in water the lower
its conductivity (Wetzel, 1975)]. The waters in the lakes surveyed were "among the softest
waters in the world." (Wright and Henriksen, 1978) Sedimentary rocks generally weather
readily, whereas igneous rocks are highly resistant. The Adirondacks, as pointed out by Scho-
field (1975; 1979) have granite bedrock with much of the area covered with a mantle of mixed
gneisses. Shallow soils predominate in the area. Thus, areas are susceptible to acidifica-
tion.
Limestone terrains, on the other hand, are capable of buffering intense concentrations of
acids. Glacially derived sediment has been found to be more important than bedrock in assimi-
lating acidic precipitation in the Canadian Shield area (Kramer, 1976). Generally, however,
bedrock geology is the best predictor of the sensitivity of aquatic ecosystems to acidic preci-
pitation (Hendrey et al., 1980).
Areas with aquatic ecosystems that have the potential for being sensitive to acidic preci-
pitation are shown in Figure 7-27. In Figure 7-27, the shaded areas on the map indicate that
the bedrock is composed of igneous or metamorphic rock while in the unshaded areas it is of
calcareous or sedimentary rock. Metamorphic and igneous bedrock weathers slowly; therefore,
lakes in areas with this type of bedrock would be expected to be dilute and of low alkalinity
[<0.5 meq HCO~/liter (Galloway and Cowling, 1978)]. Galloway and Cowling verified this
XDSX7A/D 7-93 2-9-81
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Figure 7-27. Regions in North America with takes that are sensitive to acidification
by acid precipitation by virtue of their underlying bedrock characteristics.
Source: Galloway and Cowling (1978).
7-94
-------
hypothesis by compiling alkalinity data. The lakes having low alkalinity existed in regions
having igneous and metamorphic rock (Galloway and Cowling, 1978). Hendry et al. (1980) have
developed new bedrock geology maps of the eastern United States for predicting areas which
might be impacted by acidic precipitation. The new maps permit much greater resolution for
detecting sensitivity than has been previously available for the region.
Henriksen (1979) has developed a lake acidification "indicator model" using pH-calcium
and calcium-alkalinity relationships as an indicator for determining decreased surface water
acidification. The indicator is based on the observation that in pristine lake environments
(e.g., Northwest Norway or the Experimental Lakes Area in Northwest Ontario, Canada) calcium
is accompanied by a proportional amount of bicarbonate because carbonic acid is the primary
chemical weathering agent. The pH-calcium relationship found for such regions is thus defined
as the reference level for unacidified lakes. Acidified lakes (e.g., Southeast Norway and the
Adirondack region) will exhibit lower pH or lower alkalinity than the reference lakes, at com-
parable calcium levels, due to the replacement of bicarbonate by strong acid anions.
Schofield (1979) has illustrated the use of Henriksen's model with data from Norway, the
Adirondacks, and the Experimental Lakes Area of Ontario, Canada. In the acidified lakes sul-
fate replaces bicarbonate as the major anion present (Figures 7-28 and 7-29) and is derived
primarily from precipitation. Since the bicarbonate lost in acidified lakes has been replaced
by an equivalent amount of sulfate, the concentration of excess sulfate serves as an index of
the amount of acidification that has taken place. Henriksen (1979) compared estimated acidi-
fication in Norwegian lakes to the pH and sulfate concentrations in the prevailing precipita-
tion and concluded that significant lake acidification had occurred in areas receiving preci-
pitation with an annual average (volume weighted) pH below 4.6 to 4.7 and sulfate concentra-
tions above 1 mg/1. This approximate threshold of precipitation acidity may be applicable to
sensitive lake districts in other regions as well. For reference, the estimated annual bulk
deposition sulfate for the acidified lake districts in the Adirondacks and southern Norway are
approximately 30 to 60 kg SO./ha, as compared with only 5 to 10 kg SO./ha in the reference
areas of northern Norway and the Experimental Lakes Area in Ontario. A comparison of lake pH
with regional sulfate loading levels in Sweden suggests that critical loading levels for
sensitive lakes are in the range of 15 to 20 kg SO^/ha/yr. The amount of precipitation must
also be considered since it affects total sulfate additions.
The report by Hendrey et al. (1980) compared pre-1970 data with post-1975 data. A marked
decline in both alkalinity and pH of sensitive waters of North Carolina and New Hampshire were
tested. In the former, pH and alkalinity have decreased in 80 percent of the streams and in
the latter pH has decreased 90 percent since 1949. These areas are predicted to be sensitive
by the geological map on the basis of their earlier alkalinity values. Detailed county by
county maps of other states in the eastern United States suggest the sensitivity of these
regions to acidic precipitation.
XDSX7A/D 7-95 2-9-81
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CJ
cr
LLI
0.
>
o
2
UJ
a
UJ
cc
40 60
EQUIVALENT PERCENT
60
Figure 7-28. Equivalent percent composition of major ions in Adirondack lake
surface waters (215 lakes) sampled in June 1975.
Source: Schofield (1979).
7-96
-------
30
10
10
O 0
CC
10
0
30
20
10
NW Norway
(58)
SE Norway
(57)
JZL
Adironclacks
(184)
J±
ELA
(102)
100 150
$04, \i eq/liter
200
250
Figure 7-29. Percent frequency distribution of sulfate concentrations in surface
water from lakes in sensitive regions. (ELA refers to Experimental Lake Area
of Ontario.)
Source: Schofield (1979).
7-97
-------
Though bedrock geology generally is a good predictor of the susceptibility of an area to
acidification due to acidic precipitation, other factors also have an influence. Florida, for
example, is underlaid by highly calcareous and phosphate rock, suggesting that acidification of
lakes and streams is highly unlikely. Many of the soils, however, (particularly in northern
Florida) are very mature, have been highly leached of calcium carbonate, and, as a result, some
lakes in which groundwater inflow is minimal have become acidified (Hendrey, 1980). Con-
versely, there are areas in Maine with granitic bedrock where lakes have not become acidified,
despite receiving precipitation with an average pH of approximately 4.3, because the drainage
basins contain lime-bearing till and marine clay (Davis et al., 1978). Small amounts of lime-
stone in a drainage basin exert a strong influence on water quality in terrain which would
otherwise be vulnerable to acidification. Soils in Maine in the areas where the pH of lakes
has decreased due to acidic precipitation are immature, coarse, and shallow and are derived
largely from granitic material and commonly have a low capacity for assimilating hydrogen ions
from leachate and surface runoff in lake watersheds (Davis et al., 1978). The occurrence of
limestone outcroppings in the Adirondack Mountains of New York state are highly correlated with
lake pH levels. The occurrence of limestone apparently counteracts any effects of acidic
precipitation. Consequently, when predicting vulnerability of a particular region to acidifi-
cation, a careful classification of rock mixtures should be made. Rock formations should be
classified according to their potential buffering capacity, and the type of soil overlying the
formations should be noted. Local variations in bedrock and soils are very important in
explaining variations in acidification between lakes within an area.
7.4.2 Terrestrial Ecosystems
Predicting the sensitivity of terrestrial ecosystems to acidic precipitation is much more
difficult than for aquatic ecosystems. With aquatic ecosystems it is possible to compare
affected ecosystems with unaffected ones and note where the changes have occurred. With ter-
restrial ecosystems, comparisons are difficult to make because the effects of acidic precipi-
tation have been difficult to detect. Therefore, predictions regarding the sensitivity of
terrestrial ecosystems must, as much as possible, use the data which link the two ecosystems,
i.e., data on bedrock geology. Since, in most regions of the world, bedrock is not exposed
but is covered with soil, it is the sensitivity of different types of soil which must be
assessed. Therefore, the first step is to define "sensitivity" as it is used here in relation
to soils and acidic precipitation. Sensitivity of soils to acidification alone, though it may
be the most important long-term effect, is too narrow a concept. Soils influence the quality
of waters in associated streams and lakes and may be changed in ways other than simple pH-base
saturation relationships, e.g., microbiological populations of the surface layers, accelerated
loss of aluminum by leaching. Therefore, criteria need to be used that would relate soil
"sensitivity" to any important change brought about in the local ecosystem by acid precipita-
tion (McFee, 1980).
XDSX7A/D 7-98 2-9-81
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*
All soils are not equally susceptible to acidification. Sensitivity to leaching and to
loss of buffering capacity varies according to the type of parent material from which a soil
is derived. Buffering capacity is greatest in soils derived from sedimentary rocks, especially
those containing carbonates, and least in soils derived from hard crystalline rocks such as
granites and quartzites (Gorham, 1958). Soil buffering capacity varies widely in different
regions of the country (Figure 7-30). Unfortunately, many of the areas now receiving the most
acidic precipitation also are those with relatively low natural buffering capacities.
The buffering capacity of soil depends on mineralogy, texture, structure, organic matter,
pH, base saturation, salt content, and soil permeability. Above a pH of 5.5 virtually all of
the H ions, irrespective of source, are retained by ion exchange and chemical weathering.
Below pH 5.5, the retention of the H ion decreases with the soil pH in a manner determined by
the composition of the soil (Donahue et al., 1977). With a successive drop in the soil pH
below 5.0, an increasing proportion of hydrogen ions (H+) and deposited sulfuric acid will pass
through the soil and acidify runoff water (Donahue et al., 1977). The sensitivity of different
soils based on pH, texture, and calcite content is summarized in Table 7-14.
TABLE 7-14. THE SENSITIVITY TO ACID PRECIPITATION BASED ON: BUFFER
CAPACITY AGAINST pH-CHANGE, RETENTION OF H , AND ADVERSE EFFECTS ON SOILS
Noncalcareous
Buffering
H retention
Adverse
effects
Calcareous
soils
Very high
Maximal
None
clays
pH > 6
High
Great
Moderate
sandy soils
pH > 6
Low
Great
Considerable
Cultivated
soils
pH > 5
High
Great
None -
slight
Acid
soils
pH < 5
Moderate
Slight
Slight
Reference: Wiklander (1979).
Soils are the most stable component of a terrestrial ecosystem. Any changes which occur
in this component would probably have far-reaching effects. McFee (1980) has listed four
parameters which are of importance in estimating the sensitivity of soils to acidic precipita-
tion. They are:
1. The total buffering or cation exchange capacity which is provided
primarily by clay and soil organic matter.
2. The base saturation of that exchange capacity which can be estimated
from the pH of the soil.
3. The management system imposed on the soil; is it cultivated and
amended with fertilizers or lime or renewed by flooding or by other
additions?
4. The presence or absence of carbonates in the soil profile.
XDSX7A/D 7-99 2-9-81
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REGIONS WITH SIGNIFICANT
AREAS OF SOILS THAT ARE
D NON SENSITIVE
SLIGHTLY SENSITIVE
SENSITIVE
WITHIN THE EASTERN I) S
Figure 7-30. Soils of the Eastern United States sensitive to acid rainfall.
Source: McFee (1980). 7-100
-------
In order that the factors listed above could be used in broad scale mapping of soils,
McFee evaluated them for wide applicability and ready availability. In natural soils the most
serious effects would be caused by changes in pH by leaching of soil minerals. Susceptibility
of soils to changes in either of these categories is most closely associated with the cation
exchange capacity (CEC). Soil with a low CEC and a circumneutral pH is likely to have the pH
rapidly reduced by an influx of acid. Soils with a high CEC, however, are strongly buffered
against pH changes or changes in the composition of the leachate. Acidic soils with a pH near
that of acidic precipitation will not rapidly change pH due to acidic precipitation, but will
probably release Al + ions into the leachate (McFee, 1980). Soils having a low CEC are
usually low in plant nutrients; therefore, significant changes in their productivity could
occur with only a slight loss of nutrients (McFee, 1980).
Even though CEC or buffering capacity does not completely define soil sensitivity to pos-
sible influents of acid, for the reasons given above it was the primary criterion used by McFee
for the regional mapping of soil sensitivity to acidic precipitation in the eastern United
States. Further, though it is frequently stated in much of the literature that soils with low
CEC or sandy soils having low organic matter are likely to be most susceptible to effects of
acidic precipitation, the "low CEC" values are not quantified. To develop a working set of
classes, it was necessary to make certain assumptions and "worst case" calculations. Since
soils in general are rather resistant to change due to additions of acid, a fairly high addi-
tion of acid was assumed and the question asked, "What is the maximum effect that it can have
on soil, and how high would the CEC have to be to resist that effect?" (McFee, 1980)
To determine sensitivity of a soil, McFee arbitrarily chose a span of 25 years. It was
hypothesized that a significant effect could occur if the maximum influx of acid (100 cm of
precipitation at pH 3.7 per annum) during that period equaled 10-to 25 percent of the cation
exchange capacity in the top 25 cm of soil. Soils are considered slightly sensitive if the
top 25 cm of soil has an average CEC of 6.2 to 15.4 meq/100 g (assuming a bulk density of 1.3
g/cc). If the same influx of acid exceeds 25 percent of the CEC in the top 25 cm, i.e., when
the CEC is less than 6.2 meq/100 g, the soils are considered sensitive.
Based on the above concepts, the soils of the eastern United States including effects of
cultivation were mapped (Figure 7-31) by McFee. The areas containing most of the soils poten-
tially sensitive to acidic precipitation are in the upper Coastal Plain and Piedmont regions
of the southeast, along the Appalachian Highlands, through the east central and northeastern
areas, and in the Adirondack Mountains of New York (McFee, 1980). The present limited state
of knowledge regarding the effects of acidic precipitation on soils makes a more definitive
judgment of the location of areas with the most sensitive soils difficult at the present time.
The capacity of soils to absorb and retain anions also important in determining whether
soils will become acidified was not discussed by McFee (1980). The capacity for anion
absorption is great in soils rich in hydrated oxides of aluminum (Al) and iron (Fe). Reduced
leaching of salt cations is of great significance not only in helping to prevent soil
XDSX7A/D 7-101 2-9-81
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acidification but in geochemical circulation of nutrients, fertilization in agriculture and
preventing water pollution (Wilklander, 1980, Johnson et al. 1980, Johnson, 1980). (See
Section 7.3.2.1) This parameter, as well as those listed by Me Fee (1980) should be used in
determining the sensitivity of soils to acidification by both wet and dry deposition.
7.5 SUMMARY
Occurrence of acidic precipitation (rain and snow) in many regions of the United
States, Canada, and Scandanavia has been implicated in the disappearance or reduction of
fish, other animals, and plant life in ponds, lakes, and streams. In addition, acidic pre-
cipitation possesses the potential for impoverishing sensitive soils, degrading natural
areas, injuring forests, and damaging monuments and buildings made of stone.
Sulfur and nitrogen oxides, emitted through the combustion of fossil fuels are the
chief contributors to the acidification of precipitation. The fate of sulfur and nitrogen
oxides, as well as other pollutants emitted into the atmosphere, depends on their disper-
sion, transport, transformation and deposition. Emissions from automobiles occur at ground
level, those from electric power generators from smoke stacks 1000 feet or more in height.
Transport and transformation of the sulfur and nitrogen oxides are in part associated with
the height at which they are emitted. The greater the height, the greater the likelihood
of a longer residence time in the atmosphere and a greater opportunity for chemical trans-
formation of the oxides to sulfates, nitrates or other compounds. Ozone and other photo-
chemical oxidants are believed to be involved in the chemical transformations. Because of
long range transport, acidic precipitation in a particular state or region can be the result
of emissions from sources in states or regions hundreds of miles away rather than local
sources. To date the complex nature of the chemical transformation processes has not made
the demonstration of a direct cause and effect relationship between emissions of sulfur and
nitrogen oxides and the acidity of precipitation possible.
Natural emissions of sulfur and nitrogen compounds are also involved in the formation
of acidic precipitation; however, in industrialized regions anthropogenic emissions exceed
natural emissions.
Precipitation is arbitrarily defined as being acidic if its pH is less than 5.6.
Currently the acidity of precipitation in the northeastern United States, the region most
severely impacted, ranges from pH 3.0 to 5.0. Precipitation episodes with a pH as low as
3.0 have been reported for other regions of the United States. The pH precipitation can
vary from event to event, from season to season and from geographical area to geographical
area.
The impact of acidic precipitation on aquatic and terristrial ecosystems is not the
result of a single or several precipitation events, but the result of continued additions
of acids or acidifying substances over time. Wet deposition of acidic substances on fresh-
water lakes, streams, and natural land areas is only part of the problem. Acidic substances
exist in gases, aerosols, and particulate matter transferred into the lakes, streams, and
XDSX7A/D 7-102 2-9-81
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*
land areas by dry deposition. Therefore all the observed biological effects should not be
attributed to acidic precipitation alone.
Sensitivity of a lake to acidification depends on the acidity of both wet and dry
deposition, the soil system of the drainage basin, canopy effects of ground cover and the
composition of the watershed bedrock.
Ecosystems can respond to environmental changes or perturbations only through the
response of the populations of organisms of which they are composed. Species of organisms
sensitive to environmental changes are removed. Therefore, the capacity of an ecosystem to
maintain internal stability is determined by the ability of individual organisms to adjust
their physiology or behavior. The success with which an organism copes with environmental
changes is determined by its ability to yield reproducing offspring. The size and success
of a population depends upon the collective ability of organisms to reproduce and maintain
their numbers in a particular environment. Those organisms that adjust best contribute most
to future generations because they have the greatest number of progeny in the population.
The capacity of organisms to withstand injury from weather extremes, pesticides, acidic
deposition or polluted air follows the principle of limiting factors. According to this
principle, for each physical factor in the environment there exists for each organism a
minimum and a maximum limit beyond which no members of a particular species can survive.
Either too much or too little of a factor such as heat, light, water, or minerals (even
though they are necessary for life) can jeopardize the survival of an individual and in
extreme cases a species. When one limiting factor is removed another takes its place. The
range of tolerance of an organism may be broad for one factor, narrow for another. The
tolerance limit for each species is determined by its genetic makeup and varies from species
to species for the same reason. The range of tolerance also varies depending on the age,
stage of growth or growth form of an organism. Limiting factors are, therefore, factors
which, when scarce or overabundant, limit the growth, reproduction and/or distribution of
an organism. The increasing acidity of water in lakes and streams appears to be such a
factor. Significant changes that have occurred in aquatic ecosystems with increasing
acidity include the following:
1. Fish populations are reduced or eliminated.
2. Bacterial decomposition is reduced and fungi may dominate saprotrophic communi-
ties. Organic debris accumulates rapidly, tying up nutrients, and limiting
nutrient mineralization and cycling.
3. Species diversity and total numbers of species of aquatic plants and animals are
reduced. Acid-tolerant species dominate.
4. Phytoplankton productivity may be reduced due to changes in nutrient cycling and
nutrient limitations.
5. Biomass and total productivity of benthic macrophytes and algae may increase due
partially to increased lake transparency.
XDSX7A/D 7-103 2-9-81
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6. Numbers and biomass of herbivorous invertebrates decline. Tolerant invertebrate
species, e.g., air-breathing bugs (water-boatmen, back-swimmers, water striders)
may become abundant primarily due to reduced fish predation.
7. Changes in community structure occur at all trophic levels.
Studies indicate that pH concentrations between 6.0 and 5.0 inhibit reproduction of
many species of aquatic organisms. Fish populations become seriously affected at a pH lower
than 5.0.
Disappearance of fish from lakes and streams follows two general patterns. One results
from sudden short-term shifts in pH, the other arises from a long-term decrease in the pH
of the water. A major injection of acids and other soluble substances occurs when polluted
snow melts during warm periods in winter or early spring. Fish kills are a dramatic conse-
quence of such episodic injections.
Long-term increases in acidity interfere with reproduction and spawning, producing a
decrease in population density and a shift in size and age of the population to one consist-
ing primarily of larger and older fish. Effects on yield often are not recognizable until
the population is close to extinction; this is particularly true for late-maturing species
with long lives. Even relatively small increases (5 to 50 percent) in mortality of fish
eggs and fry can decrease yield and bring about extinction.
Aluminum is mobilized at low pH values. Concentrations of aluminum may be as or more
important than pH levels as factors leading to declining fish populations in acidified lakes.
Certain aluminum compounds in the water upset the osmoregulatory function of the blood in
fish. Aluminum toxicity to aquatic biota other than fish has not been assessed.
An indirect effect of acidification potentially of concern to human health is possible
heavy metal contamination of edible fish and of water supplies. Studies in Canada and Sweden
reveal high mercury concentrations in fish from acidified regions. Lead and copper have
been found in plumbing systems with acidified water, and persons drinking the water could
suffer from lead or copper poisoning.
Acidic precipitation may indirectly influence terrestrial plant productivity by alter-
ing the supply and availability of soil nutrients. Adidification increases leaching of
plant nutrients (such as calcium, magnesium, potassium, iron, and manganese) and increases
the rate of weathering of most minerals. It also makes phosphorous less available to
plants. Acidification also decreases the rate of many soil microbiological processes such
as nitrogen fixation by Rhizobium bacteria on legumes and by the free-living Azotobacter,
mineralization of nitrogen from forest litter, nitrification of ammonium compounds, and
overall decay rates of forest floor litter.
At present there are no documented observations or measurements of changes in natural
terrestrial ecosystems that can be directly attributed to acidic precipitation. This does
not necessarily indicate that none are occurring. The information available on vegetational
effects is an accumulation of the results of a wide variety of controlled research
XDSX7A/D 7-104 2-9-81
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*
approaches largely in the laboratory, using in most instances some form of "simulated" acidic
rain, frequently dilute sulfuric acid. The simulated "acid rains" have deposited hydrogen
(H+), sulfate (S04~) and nitrate (N03) ions on vegetation and have caused necrotic lesions in
a wide variety of plants species under greenhouse and laboratory conditions. Such results
must be interpreted with caution, however, because the growth and morphology of leaves under
greenhouse conditions are often atypical of field conditions. Based on laboratory studies,
sensitivity of plants to acidic depositions seems to be associated with the wettability of
leaf surfaces. The shorter the time of contact, the lower the resulting dose, and the less
likelihood of injury.
Erosion of monuments and buildings made of stone and corrosion of metals can result from
acidic precipitation. Because sulfur compounds are a dominant component of acidic precipita-
tion and are deposited during dry deposition also, the effects resulting from the two pro-
cesses cannot be distinguished. In addition, the deposition of sulfur compounds on stone
surfaces provides a medium for microbial growth that can result in deterioration.
XDSX7A/D 7-105 2-9-81
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7.6 REFERENCES
Abrahamsen, G. , K. Bjor, R. Horntvedt, and B. Tveite. Effects of acid precipitation on coni-
ferous forest. In: F. H. Braekke, ed., Research Report FR-6. SNSF Project, NISK, Aas,
Norway, 1976. pp 37-63.
Abrahamsen, G., Horntvedt, R., and Tveite, B. Impacts of acid precipitation on coniferous
forest ecosystems, Water, Air and Soil Pollution, 8:57-73, 1977.
Abrahamsen, G. , and G. J. Dollard. Effects of Acid Deposition on Forest Vegetation. U:
Wood, M. J. (ed.). Ecological Effects of Acid Precipitation. Report of workshop held at
Calley Hotel, Gatehouse-of-Fleet, Galloway, U.K., 4-7 Sept. 1978. EPRI SOA77-403, Elec-
tric Power Res. Institute, 3412 Hillview Ave., Palo Alto, California 94303, 1979.
Alexander, M. Effects on acidity on microorganisms and microbial processes in the soil. p.
341-362. _In: Effects of Acid Precipitation on Terrestrial Ecosystems. T. C. Hutchinson
and M. Havas (eds.) Plenum Press, NY, 1978.
Alexander, M. Effects of Acidity on Microorganisms and Microbial Porcesses in a Soil. In:
Effects of Acid Precipitation on Terrestrial Ecosystems. T. C. Hutchinson and M. Havas,
eds., 1980. pp. 363-374.
Aimer, B. W., C. Dickson, C. Ekstrom, and E. Hornstrom. Chapter 7. Sulfur Pollution and the
Aquatic Ecosystem. Li: Sulfur in the Environment, Part II: Ecological Impacts. J.
Nriagu, ed. John Wiley and Sons, NY, 1978. pp. 271-311.
Aimer, B. , W. Dickson, C. Ekstrom, E. Hornstrom, and U. Miller. Effects of acidification on
Swedish lakes. Ambio 3:30-36, 1974.
Altshuller, A. P., and G. A. McBean, Co-Chairman, 1979. The LRTAP Problem in North America: a
preliminary overview. Prepared the United States-Canada Research Consultation Group on
the Long-Range transport of Air Pollutants, 1979. pp. 48.
Amundsen, T. , and K. Lunder. Report on fishery-biological surveys in Tjagevatn and Sonstevatn
in Grandsherod, Notodden in Telemark. Fiskerikonsulenten i Ost-Norge, Direktoratet for
vitt og ferskvannsfisk, Oslo, 1974. 13 pp.
Andersson, I., 0.0Grahn, H. Hultberg, and L. Landner. Jamforande undersokning av olika
tekniker for aterstallande av forsurade sjdar. STU Report 73-3651. Stockholm: Insti-
tute for Water and Air Research, 1975. Cited in: National Research Council, Sulfur
Oxides, National Academy of Sciences, Washington, DC, 1978. 118 pp.
Anschutz, J. , and F. Gessner. Der loneraustausch bei torfmoosen (Sphagnum). Flora
141:178-236, 1954. Cited in: Aimer et al., 1978.
Armstrong, F. A. J. , and D. W. Schindler. Preliminary Chemical Characterization of Waters in
the Experimental Lakes Area, Northwestern Ontario. Jour. Fish. Research Board, Canada
28:171-, 1971.
Arnold, D. E. , R. W. Light, and V. J. Dymond. Probable Effects of Acid Precipitation on
Pennsylvania Waters. EPA 600/3-80-D12. January 1980. EPA Corvallis, OR. 19 p.
Baker, J. , D. Hocking, and M. Nyborg. Acidity of open and intercepted precipitation in
forests and effects on forest soils in Alberta, Canada. In: Proceedings of the First
International Symposium on Acid Precipitation and the Forest Ecosystem, May 12-15, 1975,
Columbus, Ohio, L. S. Dochinger and T. A. Seliga, eds., U.S. Forest Service General
Technical Report NE-23, U.S. Department of Agriculture, Forest Service, Northeastern
Forest Experiment Station, Upper Darby, PA, 1976.pp. 779-790.
SOXJ7A/B 7-106 2-10-81
-------
Baker, J. , and C. Schofield. Aluminum toxicity to fish as related to acid precipitation and
Adirondack surface water quality. In: Proc. International Conference on the Ecological
Impact of Acid Precipitation, March 1980, SNSF-project report, Oslo, Norway, (in press).
Beamish, R. J. , and H. H. Harvey. Acidification of the La Cloche Mountain Lakes, Ontario and
resulting fish mortalities. Jour. Fisheries Res. Board Canada. 29:1131-1143, 1972.
Beamish, R. J. Growth and survival of white suckers (Catostomus commersoni) in an acidified
lake. J. Fish Res. Board Can. 31:49-54, 1974.
Beamish, R. J. , W. L. Lockhart, J. C. Van Loon, and H. H. Harvey. Long-term acidification of
a lake and resulting effects on fishes. Ambio 4:98-102, 1975.
Beamish, R. J. Acidification of lakes in Canada by acid precipitation and the resulting
effects on fishes, pp. 479-498. In: L. S. Dochinger and T. A. Seliga, eds. Proc. First
International Symposium on Acid Precipitation and the Forest Ecosystem, 12-15 May 1975,
USDA Forest Service General Technical Report NE-23, Upper Darby, Pennsylvania, 1976.
Beamish, R. J. Acidification of Lakes in Canada By Acid Precipitation and the Resulting
Effects of Fish. Water, Air and Soil Poll. 6:501-514, 1976.
Bell, H. L. , and A. V. Nebecker. Preliminary studies on the tolerance of aquatic insects to
low pH. J. Kansas Entomological Soc. 42:230-236, 1969.
Bell, H. L. Effect of low pH on the survival and emergence of aquatic insects. Water Res.
5:313-319, 1971.
Bick, H., and E. F. Drews. Selbstreiningung und ciliatenbesiedlung in saurem milieu
(modelIversuche). Hydrobiologia 42:393-402, 1973, Cited in Leivestad et al., 1976.
Billings, W. D. Plants and the Ecosystem. 3rd ed. pp. 1-62. Wadsworth Publishing Company,
Inc. Belmont, PA., 1978. 177 p.
Bolin, B., et al. Sweden's Case Study for the United Nations Conference on the Human Environ-
ment: Air Pollution Across National Boundaries. J_n: The Impact on the Environment of
Sulfur in Air and Precipitation. Norstadt and Sons, Stockholm, 1972. 97 pp.
Boling, R. H. , Jr., E. D. Goodman, J. A. VanSickle, J. 0. Zimmer, K. W. Cummins, R. C.
Petersen, and S. R. Reice. Toward a model of detritus processing in a woodland stream.
Ecology 56:141-151, 1975.
Borgstrb'm, R. , J. Brittain, and A. Lillehammer. Evertebrater og surt vann: Oversikt over
innsamlingslokaliteter. Research Report IR 21/76. 1432 Aas-NLH, Norway: SNSF Project
Secretariat, 1976. Cited in: National Research Council, Sulfur Oxides, National Academy
of Sciences, Washington, DC, 1978. 33 pp.
Boughey, A. S. Fundamental Ecology. Scranton, Pa., Intex Educational Publishers, 1971. p.
11-50.
Brady, N. C. The Nature and Property of Soils. McMillan Publishing Co., New York. p.
464-472, 1974.
Braekke, F. H., ed. Impact of Acid Precipitation on Forest and Freshwater Ecosystems in
Norway. SNSF Project, Research Report FR 6/76, Oslo, Norway, 1976.
SOXJ7A/B 7-107 2-10-81
-------
Brezonik, P. L. Nutrients and Other Biologically Active Substances in Atmospheric Precipita-
tion. Proc. First Specialty Symposium on Atmospheric Contribution to the Chemistry of
Lake Waters. Internat. Assoc. Great Lakes Res. Sept. 28-Oct. 1, 1975. p. 166-186.
Brock, Thomas D. Lower pH Limit for the Existence of Blue-Green Algae: Evolutionary and
Ecological Implications. Science 179:480-483, 1973.
Brooks, J. L. , and S. I. Dodson. Predation, body size and composition of plankton. Science
150:28-35, 1965.
Brosset, C. Air-borne acid. Ambio 2:2-8, 1973.
Brouzes, R. J. P., R. A. N. McLean, and G. H. Tomlinson. Mercury - the Link Between pH of
Natural Waters and the Mercury Content of Fish. Paper presented at a meeting of the
Panel on Mercury of the Coordinating Committee for Scientific and Technical Assessments
of Environmental Pollutants, National Academy of Sciences, National Research Council,
Washington, DC, May 3, 1977. Montreal, Quebec: Domtar Research Center, 1977.
Bua, B. , and E. Snekvik. Hatching experiments with roe of salmonid fish 1966-1971. Effects
of acidity and salt content of hatchery water. Van 7: 86-93. Cited in Wright and
Snekvik, 1978, 1972.
Chamberlain, A. C. Dry deposition of sulfur dioxide. J.n: Atmospheric Sulfur Deposition,
Environmental Impact and Health Effects. Proceedings of the Second Life Sciences
Symposium, Potential Environmental and Health Consequences of Atmospheric Deposition. D.
E. Shriner, C. R. Richmond, and S. E. Lindberg, eds. , Ann Arbor Science, 1980. pp.
185-197.
Cholonky, B. J. Die Okologie der diatomeen in Binnengewasser. Cramer, Weinheim, 1968. 699
pp. Cited in Hendrey et al., 1980.
Cogbill, C. V. The effect of acid precipitation on tree growth in eastern North America,
Water, Air and Soil Pollution, 8:89-93, 1977.
Cogbill, C. V., and G. E. Likens. Acid precipitation in the northeastern United States.
Water Resour. Res. 10:1133-1137.
Cole, D. W. , and D. W. Johnson. Atmospheric sulfate additions and cation leaching in a Doug-
las fir ecosystem. Water Resour. Res. 13:313-317, 1977.
Comerford, N. b. and White, E. H. , 1977, Nutrient content of throughfall in paper birch and
red pine stands in northern Minnesota, Can. J. For. Res., 7, (4), pp. 556-551.
Connell, W. E., and W. H. Patrick, Jr., Science 159. 86, 1968.
Conroy, N. , K. Hawley, W. Keller, and C. Lafrance. Influences of the atmosphere on lakes in
the Sudbury area. J. Great Lakes Res. 2(Suppl.1):146-165, 1976.
Cowling, E. B. , and L. S. Dochinger. The Changing Chemistry of Precipitation and Its Effects
on Vegetation and Materials. AICHE Symposium Series. Control and Dispersion of Air
Pollutants: Emphasis on NO and Particulte Emissions. 7:134-142, 1978.
A —
Crisman, T. L., R. L. Schulze, P. L. Brenzonik, C. D. Hendrey, and S. A. Bloom. The biotic
response in Florida lakes. In: Proc. International Conf. on the Ecological Impact of
Acid Precipitation, March 1980, SNSF-project report, Oslo, Norway, (in press).
SOXJ7A/B 7-108 2-10-81
-------
Cronan, C. S. Consequences of Sulfuric Acid Inputs to a Forested Soil. .In: Atmospheric
Sulfur Deposition, Environmental Impact and Health Effects. D. S. "Shriner, C. R.
Richmond, and S. E. Lindberg, eds., Ann Arbor Science, 1980. pp. 325-343.
Cronan, C. S. Solution chemistry of a New Hampshire subalpine ecosystem: biochemical
patterns and processes. Ph.D. Thesis, Dartmouth College, Hanover, NH, 1978. 248 pp.
Cronan, C. S. , and C. L. Schofield. Aluminum Leeching Response to Acid Precipitation:
Effects on High-Elevation Watersheds in the Northeast. Science 204:304-305, 1979.
Dancer, W. S. , L. A. Peterson, and G. Chesters. Ammonification and nitrification of N as
influenced by soil pH and previous N treatments. Soil Sci. Soc. Am. Proc. 37:67-69,
1973. ~~
Dannevig, A. Influence of precipitation on the acidity of water courses and on fish stocks.
Jeger og Fisker. 3:116-118, 1959. Cited in Leivestad et al., 1976.
Davis, R. B. , M. 0. Smith, J. H. Baily, and S. A. Norton. Acidification of Maine (U.S.A.)
Lakes By Acidic Precipitation. Verh. Internat. Verein Limnol. 20:532-537, 1978.
Denison, R. , Caldwell B.,, Bormann, B. , Eldred, L. , Swanberg, C. , and Anderson, S. The
effects of acid rain on nitrogen fixation in western Washington coniferous forests.
Water, Air and Soil Pollution 8:(1), 1977. pp. 21-34.
Dickson, W. The acidification of Swedish lakes. lr\: Fishery Board of Sweden, Institute of
Freshwater Research, Drottningholm, Sweden. Report No. 54, Lund, Sweden: Carl Bloms
Boktryckeri A.-B, 1975. pp. 8-20.
Dickson, W. Some effects of the acidification of Swedish lakes. Verh. Internat. Verein.
Limnol. 20:851-856, 1978.
Dillon, P. J., D. S. Jefferies, W. Snyder, R. Reid, N. D. Van, D. Evans, J. Moss, and W. A.
Scheider. Acidic precipitation in south-central Ontario: recent observations. J. Fish.
Res. Board Can. 35:809-815, 1978.
Dodson, S. I. Zooplankton competition and predation: an experimental test of the size-
efficiency hypothesis. Ecology 5_5:605-613, 1974.
Donahue, R. L. , R. W. Miller, and J. C. Shickluna. Soils. An introduction to soils and plant
growth. 4th ed. p. 244-248. Prentice-Hall, Inc. Englewood Cliffs, NJ 07632, 1977.
Donnevig, A. Influence of precipitation on the acidity of water courses and on fish stocks.
Jeger og Fisker. 3:116-118, 1959 (in Norwegian cited in Wright #12).
Drablefs, D., and T. Sevaldrud. Lake acidification, fish damage, and utilization of outfields.
A comparative survey of six highland areas, southeastern Norway. In: Proceedings of the
International Conference on the Ecological Impact of Acid Precipitation, March 1980. D.
Drab Ms and A. Tollan, eds., SNSF Project Report, Oslo, Norway, 1980. pp, 354-357.
Driscoll, C. T., Jr., J. P. Baker, J. I. Bisogin, Jr., and C. Schofield. Aluminum Speciation
in Dilute Acidified Waters and Its Effects on Fish. Nature 284:161-163, 1979.
Driscoll, C. T. Chemical characterization of some dilute acidified lakes and streams in the
Adirondack region of New York state. Ph.D. Thesis, Cornell University, Ithaca, New York,
1980. 309 pp.
SOXJ7A/B 7-109 2-10-81
-------
Eaton, J. S. , G. E. Likens, and F. H. Bormann. Throughfall and Stemflow Chemistry in a
Northern Hardwood Forest. J. Ecol. 61:495-508, 1973.
Eliassen, A., and J. Saltbones. Decay and transformation rates of S02 as estimated from
emission data, trajectories and measured air concentrations. Atmos Environ. 9:425-429,
1975.
European Inland Fisheries Advisory Committee (EIFAC). Water quality criteria for European
freshwater fish. Report on extreme pH values and inland fisheries. Water Res.
3:593-611, 1969.
Evans, K. S., N. F. Gmur, and J. J. Kelsch. Perturbations of upper leaf surface structures by
simulated acid rain. Environ and Exptl. Bot. 17:145-149, 1977.
Evans, L. S. , and D. M. Bozzone. Effect of buffered solutions and sulfate on vegetative and
sexual development in gametophytes of Pteridium aquilinum. Amer. J. Bot. 64:897-902,
1977.
Evans, L. S. , N. F. Gmur, and F. DaCosta. Leaf surface and histological perturbations of
leaves of Phaseolus vulgaris and Helianthus annuus after exposure to simulated acid rain.
Amer. J. Bot. 64:903-913, 1977.
Evans, L. S. , and D. M. Bozzone. Effect of buffered solutions and various anions on vegeta-
tive and sexual development in gameophytes of Pteridium aquilinum. Canadian Jour. Bot.
56:779-785, 1978.
Evans, L. S. , N. F. Gmur, and F. DaCosta. Foliar response of six clones of hybrid poplar to
simulated acid rain. Phytopathology 68:847-856, 1978.
Evans, L. S., and T. M. Curry. Differential responses of plant foilage to simulated acid
rain. Amer. J. Bot. 66:953-962, 1979.
Fagestrb'm, T. , and A. Jernelb'v. Some Aspects of the Quantitative Ecology of Mercury. Water
Res. 6:1193-1202, 1972.
Ferenbaugh, R. W. Effects of simulated acid rain on Phaseolus vulgaris L. (Fabaceae). Amer.
J. Bot. 63:283-288, 1976.
Fiance, S. B. Distribution and biology of mayflies and stoneflies of Hubbard Brook, New
Hampshire, MS. Thesis, Cornell University, Ithaca, NY, 1977.
Fisher, B. E. A. Long-range transport and deposition of sulfur oxides. |ri: Sulfur in the
Environment, Part I: The Atmospheric Cycle. J. 0. Nriagu, ed. , John Wiley and Sons, New
York, NY, 1978. pp. 243-295.
Foster, R. J. Geology. Charles E. Merrill Publishing Co., Columbus, Ohio, 1971. 162 pp.
Fredricksen, R. L. Nutrient budget of a Douglas-fir forest on an experimental watershed in
western Oregon. Pages 115-131, Research on Coniferous and R. H. Waring. Portland,
Oreg.: Pacific Northwest Forest and Range Experiment Station, 1972.
Galloway, J. N. , and E. B. Cowling. The effects of precipitation on aquatic and terrestrial
ecosystems: A proposed precipitation network. JAPCA 28:229-235, 1978.
Galloway, J. N., and D. M. Whelpdale. An Atmospheric Sulfur Budget for Eastern North America.
Atmos. Environ. 14:409-417, 1980.
SOXJ7A/B 7-110 2-10-81
-------
Galmstrb'm, G. , G. Andersson, and S. Fleischer. Decomposition and exchange processes in
acidified lake sediment. |n: Proceedings of the International Conference on the
Ecological Impact of Acid Precipitation, March 11-14, 1980, D. Drablos and A. Tollan,
eds., SNSF Project Report, Oslo, Norway, 1980. pp. 306-308.
Galvin, P. J. , P. J. Samson, P. E. Coffey, and D. Romeno. Transport of sulfate to New York
State. Environ. Sci. and Tech. 12:580-584, 1973.
Galvin, P. J. , and J. A. Cline. Measurement of anions in the snow cover of the Adirondack
Mountains. Atmos. Environ. 12:1163-1167, 1978.
Garrett, A. Compositional changes of ecosystem during chronic gamma irradiation. ]_n:
Symposium on Radioecology. Proc. Second National Symposium, Ann Arbor, Mich., May 15-17,
1967.
Gatz, D. F. Comment on 'Acid precipitation in the northeastern United States' by Charles V.
Cogbill and Gene E. Likens. Water Resour. Res. 12:569-570, 1976.
Giddings, J. , and J. N. Galloway. The Effects of Acid Precipitation on Aquatic and Terres-
trial Ecosystems. Literature Reviews on Acid Precipitation, EEP-2. Cornell University,
Ithaca, NY, 1976.
Gjessing, E. T. , A. Henriksen, M. Johannessen, and R. F. Wright. Changes in the Chemical
Composition of Lakes. J_n: Impact of Acid Precipitation on Forest and Freshwater
Ecosystems in Norway. F. H. Braekke, ed., Research Report FR 6/76, 1432 Aas-NLH, Norway,
SNSF Project Secretariat, 1976. pp. 65-85.
Glass, G. E. , and 0. L. Loucks (eds). Impacts of Air Pollutants on Wilderness Areas of
Northern Minnesota. pp. 188. Environmental Research Laboratory-Duluth, ORD, U.S.
Environmental Protection Agency, Duluth, Minnesota, October 1979.
Glass, N. R. , G. E. Glass, and P. J. Rennie. Effects of acid precipitation. Environ. Sci.
Tech. 13:1350-1355, 1979.
Glover, G. M. , and A. H. Webb. Weak and Strong Acids in The Surface Waters of The Tovdal
Region in S. Norway. Water Res. 13:781-783, 1979.
Gorham, E. On the acidity and salinity of rain. Geochim. Cosmochim. Acta. 7:231-239, 1955.
Gorham, E. The influence and importance of daily weather conditions in the supply of
chloride, sulfate and other ions to fresh waters from atmospheric precipitation. Phil.
Trans. R. Soc. London, Ser. B 241:147-178, 1958.
Gorham, E. Acid precipitation and its influence upon aquatic ecosystems—an overview. In:
Proceedings of the First International Symposium on Acid Precipitation and the Forest
Ecosystem, May 12-15, 1975, Columbus, Ohio, L. S. Dochinger and T. A. Selinga, eds., U.S.
Forest Service General Technical Report NE-23, U.S. Department of Agriculture, Forest
Service, Northeastern Forest Experiment Station, Upper Darby, PA, 1976. pp. 425-458.
Gosner, K. L. , and I. H. Black. The effects of acidity on the development and hatching of New
Jersey frogs. Ecology 38:256-262, 1957.
Grahn, 0. Macrophyte succession in Swedish lakes caused by deposition of airborne acid sub-
stances, pp. 519-530. J_n: Proceedings of the First International Symposium on Acid
Precipitation and the Forest Ecosystem, L. S. Dochinger and T. A. Seliga, eds., Ohio
State University, May 12-15, 1975. U.S.D.A. Forest Service General Technical Report
NE-23. Upper Darby, Pennsylvania. Forest Service, U.S. Department of Agriculture,
Northeastern Forest Experiment Station, 1976.
SOXJ7A/B 7-111 2-10-81
-------
Grahn, 0., H. Hultberg, and L. Landner. Oligotrophication--a self-accelerating process in
lakes subjected to excessive supply of acid substances. Ambio 3:93-94, 1974.
Granat, L. On the relationship between pH and the chemical composition in atmospheric preci-
pitation. Tell us 24:550-560, 1972.
Grennard, A., and F. Ross. Progress report on sulfur dioxide. Combustion 45:4-9, 1974.
Hagen, A., and A. Langeland. Polluted snow in southern Norway and the effect of the meltwater
on freshwater and aquatic organisms. Environ. Pollut. 5:45-57, 1973.
Hagstrom, T. 1977. Grodornas fb'rsvinnande i an forsurad sjo.
Sveriges Natur 11/77:367-369. Cited in Hendrey, 1978.
Haines, E. B. Nitrogen content and acidity of rain on the Georgia coast. Water Resour. Bull.
12:1223-1231, 1976.
Hall, R. J., G. E. Likens, S. B. Fiance, and G. R. Hendrey. Experimental acidification of a
stream in the Hubbard Brook Experimental Forest, New Hampshire. Ecology 61:976-989,
1980.
Halstead, R. L. , and P. J. Rennie. Effects of Sulphur on Soils in Canada. 181-220. In:
Sulphur and its Inorganic Derivatives in the Canadian Environment. Nat. Res. Council of
Canada. NRC Assoicate Committee on Scientific Criteria for Environmental Quality.
Ottawa, Canada, 1977.
Hamilton, W. C. Estimate by the National Coal Association. Personal communication. Jan.
1980.
Hanken, I., J. F. Lynch, and D. B. Wake. Salamander invasion of the tropies. Natural Hist.
89:47-53, 1980.
Harr, T. E., and P. E. Coffey. Acid Precipitation in New York State. Technical Paper No. 43.
New York State Department of Environmental Conservation, Albany, NY. pp. 21-29, 1975.
Harward, M. E. , and H. H. Reisenaur. Movement and reactions of inorganic soil sulfur. Soil
Sci. 101:326-335, 1966.
Heagle, A. S. , W. W. Heck, W. M. Knott, J. W. Johnston, E. P. Stahel , and E. B. Cowling.
Responses of Citrus to Acidic Rain from Simulated SRM Fuel Exhaust Mixtures and Exhaust
Components, Internal report, 1978. 11 pp.
Heinrichs, H. , and R. Mayer. Distribution and cycling of major and trace elements in two
central forest ecosystems. J. Environ. Qua!. 6:402-407, 1977.
Henderson, G. S., and W. F. Harris. An ecosystem approach to characterization of the nitrogen
cycle in a deciduous forest watershed. Pages 179-193, Forest Soils and Forest Soils
Conference, edited by B. Bernier and C. H. Winget. Quebec, Canada: Les Presses de
1'Universite Laval, 1975.
Hendrey, C. D., and P. L. Brezonik. Chemistry of precipitation at Gainesville, Florida.
Environ. Sci. and Tech. 14:843-849, 1980.
Hendrey, G. R. , K. Baalstrud, T. S. Traaen, M. Laake, and G. Raddum. Acid precipitation:
Some hydrobiological changes. Ambio 5:224-227, 1976.
SOXJ7A/B 7-112 2-10-81
-------
Hendrey, G. R. , and R. F. Wright. Acid precipitation in Norway: Effects on aquatic fauna.
J. Great Lakes Res. 2(Suppl 1):192-207, 1976.
Hendrey, G. R. Effects of low pH on the growth of periphytic algae in artificial stream
channels. SNSF Project, 1432 Aas-NLH, Norway, 1976.
Hendry, C. Chemical composition of rainfall at Gainesville, Florida. M.S. Thesis. Depart-
ment of Environmental Engineering Sciences, University of Florida, Gainesville, 1977.
Hendrey, G. R. Aquatics Task Force on Environmental Assessment of the Atikokan Power Plant:
Effects on Aquatic Organisms. Land and Fresh Water Environmental Sciences Group,
Department of Energy and Environment. Brookhaven National Laboratory Associated Univer-
sities, Inc. Lipton, NY BNL 50932. November 1978. 16 p.
Hendrey, G. , and F. W. Barvenik. Impact of Acid Precipitation on Decomposition and Plant
Communities in Lakes. j_n: Scientific Papers from the Public Meeting on Acid Precipita-
tion, May 4-5, 1978, Lake Placid, New York, published by Science and Technology Staff,
New York State Assembly, Albany, NY March, 1979. p. 92-103.
Hendrey, G. R. , J. N. Galloway, S. A. Norton, C. L. Schofield, P. W. Shaffer and D. A. Burns.
Geological and Hydrochemical Sensitivity of the Eastern United Stated to Acid Precipita-
tion. Report prepared for Ecological Effects Division, Corvallis Environmental Research
Laboratory, U.S. Environmental Protection Agency under EPA-DOE Interagency Agreement
79-D-X0672. EPA-600/3-80-024. January 1980.
Hendrey, G. R. , N. D. Yan, and K. J. Baumgartner. Responses of freshwater plants and inverte-
brates to acidification. EPA/OECD International Symposium for Inland Waters and Lake
Restoration, September 1980, Portland, Maine, 1980.
Hendrey, G. R. , and F. Vertucci. Benthic plant communities in acidic Lake Colden, New York:
Sphagnum and the algal mat. I_n: Proc. Internat. Conf. on the Ecological Impact of Acid
Precipitation. March 1980, SNSF-project report, Oslo, Norway, (in press).
Henriksen, A., and R. F. Wright, Effects of acid precipitation on a small acid lake in south-
ern Norway. Nord. Hydrol. 8:1-10, 1977.
Henriksen, A. Acidification of Freshwaters: A Simple Approach for Identification and Quanti-
fication. Nature 278:542-545, 1979.
Hicks, B. B. , and M. L. Wesely. Turbulent transfer processes to a surface and interaction
with vegetation. I_n: Atmospheric Sulfur Deposition, Environmental Impact and Health
Effects. D. E. Shriner, C. R. Richmond, and S. E. Lindberg, eds. , Proceedings of the
Second Life Sciences Symposium, Potential Environmental and Health Consequences of Atmos-
pheric Deposition, Ann Arbor Science, 1980. pp. 199-207.
Hindawi, I. J., and H. C. Ratsch. Growth Abnormalities of Christmas Trees attributed to
Sulfur Dioxide and Particulate Aerosol. Presented at the 67th Annual Meeting of the Air
Poll. Control Assoc. Abstract Number 74-252, 1974.
Hindawi, I. J., J. A. Rea, and W. L. Griffis. Response of bush bean exposed to acid mist. J.
Bot. 67:168-172, 1980.
Hornstrom, E. , C. Ekstrdm, U. Miller, W. Dickson, Forsurnings inverkan pa vastkustsjbar
(Effects of the Acidification on Lakes in the Swedish West Coast Region) Statens
Naturvardsverk, Slona, Sweden, Publikationer 1973: 7 (in Swedish, summary in English).
Cited in: Wright, R. F., and Engil. T. Gjessing. Changes in the chemical composition of
lakes. Ambio. 5: 219-223, 1976.
SOXJ7A/B 7-113 2-10-81
-------
Hoeft, R. G., D. R. Kieney, and L. M. Walsh. Nitrogen and Sulphur in Precipitation and Sulfur
Dioxide in the Atmosphere in Wisconsin. J. Environ. Quality 1:203-208, 1972.
Rolling, C. C. Resilience and stability of ecological systems. Ann. Rev. Ecol. Syst. 4:1-23,
1973.
Hornbeck, J. W., G. E. Likens, and J. S. Eaton. Seasonal patterns in acidity of precipitation
and their implications for forest stream ecosystems. .In: Proceedings of the First
International Conference on Acid Precipitation and the Forest Ecosystem, May 12-15, 1975,
Columbus, Ohio. L. S. Dochinger and T. A. Seliga, eds. U.S. Forest Service General Tech-
nical Report NE-23, U.S. Department of Agriculture, Forest Service, Northeastern Forest
Experimental Station, Upper Darby, PA, 1976. pp.
Hubbert, M. K. Outlook for fuel reserves. I_n: McGraw-Hill Encyclopedia of Energy, D. N.
Lapedes, ed., McGraw-Hill Inc., New York, NY, 1976. pp 11-23.
Hultberg, H. Thermally stratified acid water in late winter - A key factor inducing self-
accelerating processes which increase acidification p. 503-517. _In: Proceedings of the
First International Symposium on Acid Precipitation and the Forest Ecosystem, May 12-15,
1975, Columbus, Ohio, L. S. Dochinger and T. A. Seliga, eds., U.S. Forest Service General
Technical Report NE-23, Upper Darby, Pennsylvania, Forest Service, U.S. DOA, Northeastern
Forest Experiment Station, 1976.
Hutchinson, T. C. , and L. M. Whitby. The effects of acid rainfall and heavy metal particu-
lates on a boreal forest ecosystem near the Sudbury smelting region of Canada. In:
Proceedings of the First International Symposium on Acid Precipitation and the Forest
Ecosystem, May 12-15, 1975, Columbus, Ohio, L. S. Dochinger and T. A. Seliga, eds., U.S.
Forest Service General Technical Report NE-23, U.S. Department of Agriculture, Forest
Service, Northeastern Forest Experiment Station, Upper Darby, PA, 1976. pp.
Hutchinson, T. C. , W. Gizym, M. Havas, and V. Zoberns. Effects of long-term lignite burns on
Arctic ecosystems at the Smoking Hills, N.W.T., pp. 317-332. In: D. D. Hemphill, ed.,
Trace Substances in Environmental Health-XII, University of Missouri, Columbia, Missouri,
1978.
Irving, P. Rainfall Acidity at Argonne. RER Division Annual Report, Argonne National Labora-
tory, Argonne, IL, ANL-78-65-III, 1978.
Irving, P. M. Induction of Visible Injury in Chamber-Grown Soybeans Exposed to Acid Precipi-
tation. RER Division Annual Report, ANL-78-65-III, Argonne, IL, 1978.
Irving, P., and J. E. Miller. The Effects of Acid Precipitation Alone and in Combination with
Sulfur Dioxide on Field-Grown Soybeans. RER Division Annual Report, ANL-78-65-III,
Argonne, IL, 1978.
Irving, P. M. Response of Field-Grown Soybeans to Acid Precipitation Alone and in Combina-
tion with Sulfur Dioxide. PhD. Thesis, University of Wisconsin, Milwaukee, WI, 1979.
Jacobson, J. S. , L. J. Heller, and P. Van Leuken. Acidic precipitation at a site within the
northeastern concentration. Water, Air and Soil Poll. 6:339-349, 1976.
Jacobson, J. S. , and P. van Leuken. Effects of acidic precipitation on vegetation. Proc.
Fourth Intern. Clean Air Congress, pp. 124-127. 1977.
SOXJ7A/B
7-114
2-10-81
-------
Jacobson, J. S. J. Trojano, L. J. Colavito, L. I. Heller, and D. C. McCune. Polluted Rain and
Plant Growth. In: Polluted Rain, 12th Rochester International Conference on Environ-
mental Toxicity. Plenum Publishing Co., New York, NY, 1980.
Jacobson, J. S. Experimental Studies on the Phytotoxicity of Acidic Precipitation: The
United States Experience. In: Effects of Acid Precipitation on Terrestrial Ecosystems.
T. C. Hutchinson and M. Havas, eds. , Plenum Press, New York, NY, 1980a. pp. 151-160.
Jacobson, J. S. The influence of rainfall composition on the yield and quality of agricul-
tural crops. |n: Proceedings of the International Conference on the Ecological Impact
of Acid Precipitation, Sandefjord, Norway, March 11-14, 1980. SNSF project report Oslo,
Norway, D. Drableis and A. Tollan, eds., 1980b. pp. 41-46.
Jensen, K. W. , and E. Snekvik. Low pH Levels Wipe Out Salmon and Trout Populations in South-
ernmost Norway. Ambio 1:223-225, 1972.
Johannessen, M. , and A. Henriksen. Chemistry of snow meltwater: changes in concentration
during melting. Water Resources Res. 14:615-619, 1978.
Johannessen, M. , R. F. Wright, and A. Skartvect. Streamwater chemistry before, during and
after snowmelt. _In: Proc. International Conference on the Zoological Impact of Acid
Precipitation, SNSF-project, Oslo, Norway, (in press).
Johnson, N. M. , G. E. Likens, F. H. Bormann, D. W. Fisher, and R. S. Pierce. A working model
for the variation in stream water chemistry at the Hubbard Brook Experimental Forest, New
Hampshire. Water Resources Res. 5: 1353-1363, 1969.
Johnson, D. W. Site Susceptibility to Leaching by H«SO. in Acid Rainfall. I_n: Effect of
Acid Precipitation on Terrestrial Ecosystems. T. C. Hutchinson and M. Havas, eds.,
Plenum Press, New York, NY, 1980. pp. 525-535.
Johnson, D. W., and D. W. Cole. Anion mobility in sols: Relevance to nutrient transport from
forest ecosystems. Environ. International 3:79-90, 1980.
Johnson, D. W. , J. W. Hornbeck, J. M. Kelly, W. T. Swank, and D. E. Todd. Regional patterns
of soil sulfate accumulation: Relevance to ecosystem sulfur budgets. In: Atmospheric
Sulfur Deposition, Environmental Impact and Health Effects. D. E. Shriner, S. R.
Richmond, and S. E. Lindberg, eds., Proceedings of the Second Life Science Symposium,
Potential Environmental Health Consequences of Atmospheric Deposition, Ann Arbor Science,
1980. pp. 507-520.
Jones. M. B. Sulfur in Agricultural Lands, p. 146-158. Iji: Sulfur in the Environment.
Missouri Botanical Garden in cooperation with Union Electric Company, St. Louis, Mo.,
July, 1975.
Jonsson, B. , and R. Sundberg. Has the acidification by atmospheric pollution cause a growth
reduction in Swedish forests? Rep. Notes No. 20. Stockholm, Sweden: Royal College of
Forestry, 1972.
Junge, C. E. The distribution of ammonia and nitrate in rainwater over the United States.
Trans. Amer. Geophys. Union 39:241-248, 1958.
Kallqvist, T. , R. Romstad, and J. Kotai. Orienterende undersizikelse av pH-effekter pa
algesamfunn i forstfksrenner. (Preliminary study of pH-effects on algal communities in
experimental channels.) SNSF-Project IR 7/75 (in Norwegian), 1975. Cited in Hendry et
al., 1980.
SOXJ7A/B 7-115 2-10-81
-------
Kramer, J. R., Geochemical and Lithological Factors in Acid Precipitation, p. 611-618. In:
Proceedings of the First International Symposium on Acid Precipitation and the Forest
Ecosystem, May 12-15, 1975, Columbus, Ohio, L. S. Dochinger and T. A. Seliga, eds. , U.S.
Forest Service General Technical Report NE-23, Upper Darby, Pennsylvania, Forest Service,
U.S. Department of Agriculture, Northeastern Forest Experiment Station, 1976.
Krupa, Sagar. Acidic Rain: Its Chemistry and Effects on Terrestrial Vegetation. Proc. 3rd
International Congress of Plant Pathology, Munich, West Germany, August 1978. p. 345
(Abstract).
Kucera, V. Effects of Sulfur Dioxide and Acid Precipitation on Metals and Anti-Rust Painted
Steel. Ambio 5:243-248, 1976.
Kwiatkowski, R. E. , and J. C. Roff. Effects of acidity on the phytoplankton and primary pro-
ductivity of selected northern Ontario lakes. Can. J. Bot. 54:2546-2561, 1976.
Laake, M. Effekter av lav pH pa produksjon, nedbryting og stoffkret sleip i 1ittoralsonen.
Research Project IR 29/76, 1432 Aas-NLH, Norway. SNSF-project, 1976. 75 pp. Cited in
National Research Council, 1978.
Landner, L. , and P. 0. Larsson. Biological Effects of Mercury Fall-Out Into Lakes from the
Atmosphere, IUL Report B115. Stockholm: Institute for Water and Air Research, 1972. 18
pp. (in Swedish, translated by H. Altosaar, Domtar Research Centre, December 23, 1975)
Lang, D. S. , D. S. Shriner, and S. V. Krupa. Injury to vegetation incited by sulfuric acid
aerosols and acidic rain. Paper 78-7.3, 71st Annual Meeting, Air Pollution Control
Association, Houston, Texas, 1978.
Larson, T. V., R. J. Charlson, E. J. Knudson, G. D. Christian, and H. Harrison. The influence
of a sulfur dioxide point source on the rain chemistry of a single storm in the Puget
Sound region. Water Air Soil Pollut. 4:319-328, 1975.
Lee, J. J. , and D. E. Weber. Effects of Sulfuric Acid Rain on Two Model Hardwood Forests:
Throughfall, Litter Leachate, and Soil Solution. EPA-600/3-80-014, U.S. Environmental
Protection Agency, Corvallis, OR, January 1980.
Lee, J. J. , G. E. Neely, and S. C. Perrigan. Sulfuric Acid Rain Effects on Crop Yield and
Foliar Injury. January EPA-600/3.-80-016, U.S. Environmental Protection Agency, Cor-
vallis, OR, January 1980.
Leivestad, H., and I. P. Muniz. Fish Kill at Low pH in a Norwegian River. Nature 259:391-392,
1976.
Leivestad, H. , G. Hendrey, I. P. Muniz, and E. Snekvik. Effects of acid precipitation on
freshwater organisms. Itn: Impact of Acid Precipitation of Forest and Freshwater Eco-
systems in Norway. F. J. Braekke, ed. , SNSF Project Report FR 6/76, 1432 Aas-NLH,
Norway, 1976. pp. 87-111.
Lewis, W. M. , Jr., and M. C. Grant. Acid Precipitation in the Western U.S. Science
207:176-177, 1980.
Likens, G. E. , G. H. Bormann, N. M. Johnson, D. W. Fisher, and R. S. Pierce. Effects of
cutting and herbicide treatment on nutrient budgets in the Hubbard Brook Watershed -
ecosystem. Ecol. Monogr. 40:23-47, 1970.
Likens, G. W. , F. H. Bormann, and N. M. Johnson. Acid rain. Environment 14:33-40, 1972.
SOXJ7A/B 7-116 2-10-81
-------
Likens, G. E. The Chemistry of Precipitation in the Central Finger Lakes Region. Technical
Report No. 50. Water Resources and Marine Science Center, Cornell University Ithaca
NY, 1972.
Likens, G. E. , and F. H. Bormann. Acid rain: A serious regional environmental problem.
Science 184:1176-1179, 1974.
Likens, G. E. , and F. H. Bormann. Linkages Between Terrestrial and Aquatic Ecosystems.
BioScience 24:447-456, 1974.
Likens, G. E. Acid precipitation. Chem. Eng. News 54: (48)29-44, 1976.
Likens, G. E. , F. H. Bormann, J. S. Eaton, R. S. Pierce, and N. M. Johnson. Hydrogen ion
input to the Hubbard Brook Experimental Forest, New Hampshire, during the last decade.
I_n: Proceedings of the First International Symposium on Acid Precipitation and the
Forest Ecosystem, May 12-15, 1975. Columbus, Ohio. L. S. Dochinger and T. A. Seliga,
eds. , U.S. Forest Service General Technical Report NE-23. U.S. Department of Agricul-
ture, Forest Service, Northeastern Forest Experiment Station, Upper Darby, PA, 1976.
pp. 397-407.
Likens, G. E. , J. S. Eaton, and J. N. Galloway. Precipitation as a Source of Nutrients for
Terrestrial and Aquatic Ecosystems. Precipitation Seevenging (1974). Proc. of a
Symposium held at Champaign, 111. Oct. 14-18, 1974. Illinois State Water Survey. Publ.
Tech. Information Center, ERA. CONF-741003, 1977. p. 552-570.
Likens, G. E. , R. F. Wright, J. N. Galloway, and T. J. Butler. Acid Rain. Scientific
American. 241:43-51, 1979.
Liljestrand, H. , and J. J. Morgan. Chemical composition of acid precipitation in Pasadena,
California. Environ. Sci. Tech. 12:1271-1273. 1978.
Lindberg, S. E. , R. C. Harriss, R. R. Turner, D. S. Shriner, and D. D. Huff. Mechanisms and
rates of atmospheric deposition of selected trace elements and sulfate to a deciduous
forest watershed. ORNL/TM 6674, Environmental Sciences Division Pub. No. 1299, 1979.
pp. 514.
Lipske, M. Sour Rain, Deadly Rain. Defenders. 55:2-5, 1980.
Livingstone, D. A. Chemical composition of rivers and lakes. U. S. Geological Survey Profes-
sional Paper 440-6, Washington, DC, 1963.
Lockhart, W. L. , and A. Lutz. Preliminary biochemical observations of fishes inhabiting an
acidified lake in Ontario, Canada, pp. 545-569. In: L. S. Dochinger and T. A. Seliga,
eds. , Proc. First International Symposium on Acid Precipitation and the Forest Ecosystem,
USDA Forest Service General Technical Report NE-23, Upper Darby, Pennsylvania, 1976.
Lynch, H. Predation, competition and zooplankton community structure: an experimental study.
Limnol. Oceanogr. 24:253-272, 1979.
Malmer, N. Acid precipitation: Chemical changes in the soil. Ambio 5:231-234, 1976.
Marenco, A, and J. Fontan. Influence of drt deposition on the residence time of particulate
pollutants in the troposphere. I_n: Proceedings, Atmosphere-Surface Exchange of Particu-
late and Gaseous Pol lutants-1974 Energy Reasearch and Development Administration
Symposium Series, CONF-740921, National Technical Information Service, U.S. Department of
Commerce. Springfield, VA. pp. 54-61.
SOXJ7A/B 7-117 2-10-81
-------
Mayer, R. , and B. Ulrich. Acidity of Precipitation as Influenced by the Filtering of
Atmospheric Sulphur and Nitrogen Compounds-Its Role in the Element Balance and Effect on
Soil. Water, Air, and Soil Poll. 7:409-416. 1977.
McColl, J. G. , and D. S. Bush. Precipitation and throughfall chemistry in the San Francisco
Bay Area. J. Environ. Qual 7:352-357, 1978.
McFee, W. W. , J. M. Kelly, and R. H. Beck. Acid precipitation: Effects on soil base pH and
base saturation of exchange sites. In: Proceedings of the First International Symposium
on Acid Precipitation and the Forest Ecosystem, May 12-15. 1975, Columbus, Ohio, L. S.
Dochinger and T. A. Seliga, eds. , U.S. Forest Service General Technical Report NE-23,
U.S. Department of Agriculture, Forest Service, Northeastern Forest Experiment Station,
Upper Darby, PA, 1976. pp. 725-735.
McFee, W. W. Sensitivity of Soil Regions To Long Term Acid Precipitation, p. 131. Corvallis
Environmental Research Laboratory, U.S. Environmental Protection Agency, Corvallis, OR.
EPA-600/3-80-013, January 1980.
Menendex, R. Chronic effects of reduced pH on brook trout (Salvelinus fontinalis). J. Fish.
Res. Board Can. 33:118-123, 1976.
Miller, J. M. , J. N. Galloway, G. E. Likens. Origin of air masses producing acid precipita-
tion at Ithaca, NY, A preliminary report. Geophysical Research Letters 5:757-769, 1978.
Minshall, G. W. Role of Allochtronous detritus in the trophic structure of a woodland spring-
brook community. Ecology 48:139-149, 1967.
Moran, J. M. , M. D. Morgan, and J. H. Wiersma. Introduction to Environmental Science. W. H.
Freeman and Co., San Francisco, CA, 1980. pp. 7-75.
Miiller, P. Effects of artificial acidification on the growth of periphyton. Can. J. Fish.
Aquat. Sci. 37:355-363, 1980.
Muniz, I. P., and H. Leivestad. Toxic effects of aluminum on brown trout (Salmo trutta, L.).
In: Proc. International Conference on the Ecological Impact of Acid Precipitation, 11-14
March 1980, SNSF-project, Oslo, Norway, (in press).
National Research Council. Scientific and Technical Assessments of Environmental Pollutants.
Nitrates: An Environmental Assessment. National Academy of Sciences, 1978. p. 275-317.
National Research Council. Sulfur Oxides. National Academy of Sciences. Washington, D. C.
Chapter 5. Effects of Sulfur Oxides on Aquatic Ecosystems, 1978.
National Research Council. Sulfur Oxides. Chapter 4. National Academy of Sciences.
Washington, D.C. 1978.
o
Nihlgard, B. Precipitation, its chemical composition and effect on soil water in a beech and
a spruce forest in south Sweden, Oikos, 21:208-217, 1970.
Nisbit, I. C. T. Sulfates and acidity in precipitation: Their relationship to emissions and
regional transport of sulfur oxides. Ijj: Air Quality and Stationary Source Emission
Control. A report prepared by the Commission on Natural Resources, National Academy of
Sciences for the Committee on Public Works, U.S. Senate, 94th Congress, 1st Session.
Committee Serial No. 94-4, U.S. Government Printing Office, Washington, D.C., 1975. pp.
276-312.
SOXJ7A/B 7-118 2-10-81
-------
Norton, S. A. Changes in Chemical Processes in Soils Caused by Acid Precipitation Water
Air and Soil Poll. 7:389-400, 1977.
Nriagu, J. Deteriorative Effects of Sulfur Pollution on Materials. In: Sulfur in the
Environment Part II: Ecological Impacts, Nriagu, J. ed. John Wiley and Sons, New York,
1978. pp. 2-59.
0akland, J. Distribution and ecology of the freshwater snails Gastropoda) of Norway.
Malacologia 9:143-151, 1969.
0akland, K. A. On the distribution and ecology of Gammarus lacustris, G. 0. Sars in Norway,
with notes on its morphology and biology. Norw. J. Zool. 17:111-152, 1969.
0kland, J. A group of fish-food organisms (snails. Gastropoda) and environmental parameters
in 1000 lakes in Norway, with emphasis on low pH habitats. JJK Proc. Internat. Conf.
The Ecological Impact of Acid Precipitation, March 1980, Oslo Norway, (in press).
0kland, K. Some fish-food organisms in Norway, Asellus aquaticus, Gammarus lacustris
(Crustacea) and small mussels (Sphaeriidae): ecology and distribution. I_n: Proc.
Internat. Conf. on the Ecological Impact of Acid Precipitation, March 1980, Oslo Norway.
(in press).
Oden, S. The acidification of air and precipitation and its consequences on the natural
environment. Swedish Nat. Sci. Res. Council, Ecology Committee Bull. 1, 1968. 68 pp.
Oden, S. Acidification by Atmospheric Precipitation—A General Threat to the Ecosystem. J_n:
Acidification of the Biological Environment. I. Mysterus, ed., Translation Series No.
2564, Fisheries Research Board of Canada, Ottawa, Canada, 1971.
Oden, S. , R. Andersson, and M. Barting. The long-term changes in the chemistry of soils in
Scandanavia due to acid precipitation. In: Supporting Studies to air Pollution Across
National Boundaries. The impact on the environment of sulfur in air and precipitation,
B. Bolin, ed. , Sweden's case study for the United Nations conference on the human
environment, Royal Ministry of Foreign Affairs., Royal Ministry of Agriculture,
Stockholm, Sweden, 1972. p. 20.
Oden, S. Acid precipitation: A world concern. I_n: Proceedings of the Conference on
Emerging Environmental Problems, Acid Precipitation, May 19-20, 1975, The Institute on
Man and Science, Rensselaerville, N.Y. U.S. Environmental Protection Agency, Region II.
Water Resources and Marine Sciences Center, Cornell Univeristy and Center for Environ-
mental Quality Management, Cornell University, EPA-902/9-75-001, November 1975. pp.
5-44.
Oden, S. The acidity problem—An outline of concepts. Jji: Proceedings of the First Inter-
national Conference on Acid Precipitation and the Forest Ecosystem. May 12-15, 1975,
Columbus, Ohio, L. S. Dochinger and T. A. Seliga, eds. , U.S. Forest Service, General
Technical Report NE-23, U.S. Department of Agriculture, Forest Service, Northeastern
Forest Experiment Station, Upper Darby, PA, August 1976. pp. 1-36.
Odum, E. P. Summary. In: Ecological Effects of Nuclear War. G. M. Woodwell (ed.). Brook-
haven National Lab. 917 (C-43), 1965.
Odum, E. P. Fundamental of Ecology 3 ed. W. B. Saunders Co. Philadelphia, 1971. p. 5,
8-139.
SOXJ7A/B 7-119 2-10-81
-------
Oliver, D. R. Life history of the Chironomidae. Ann. Rev. Entomol. 16:211-230, 1971.
Overrein, L. N. Sulphur pollution patterns observed; leaching of calcium in forest soil
determined. Ambio. 1:145-147, 1972.
Pack, D. H., G. J. Ferber, J. L. Heffter, K. Telegadas, J. K. Angel 1, W. H. Hoecker, and L.
Machta. Meterology of Long-Range Transport. Atmos. Environ. 12:425-444, 1978.
Pack, D. W. Sulfate behavior in eastern U.S. precipitation. Geophysical Res. Letters.
5:673-674, 1978.
Peterson, R. C. Dynamics of coarse particulate organic material processing in diversified pH
regimes. |n: Internat. Conf. The Ecological Impact of Acid Precipitation, March 1980,
SNSF-project, Oslo, Norway, (in press).
Pfeiffer, M. List and Summary of Acidified Adirondack Waters Based on Data Available as of
April 1979. New York State Department of Environmental conservation. Albany, New York,
May 1979.
Pfeiffer, M. , and P. Festa. Acidity status of lakes in the Adirondack region of New York in
relation to fish resources. New York State Department of Environmental Conservation
Publication FW-P168, 1980. 36 pp.
Potts, D. T. W., and C. Fryer. The effect of pH and salt content on sodium balance in Daphnia
magna and Acantholeberis curvirostris (Crustacea; Cladocera). J. Comp. Physiol. Series B
129:289-294, 1979.
Potts, W. T. W. , and D. J. A. Brown. Study Group 5 Discussions. D.5.1-D.5.5. In: Wood, M.
J. (ed). Ecological Effects of Acid Precipitation. Report of workshop held at Cally
Hotel, Gatehouse-of-Fleet, Galloway, U.K. 4-7 Sept. 1978. EPRI SOA77-403, Electric Power
Res. Institute, 3412 Hillview Ave., Palo Alto, California 94303, 1979.
Pough, F. H. Acid precipitation and embryonic mortality of spotted salamanders, Ambystoma
maculatum. Science 192:68-70, 1976.
Pough, F. H. , and R. E. Wilson. Acid precipitation and reproductive success of Ambystoma
salamanders, pp. 531-544. In: Proceedings of the First International Symposium on Acid
Precipitation and the Forest Ecosystem, Ohio State University, May 12-15, 1975. U.S.
For. Serv. Gen. Tech. Rep. NE-23, 1976.
Prahm, L. R. , U. Trop, and R. M. Stern. Deposition and transformation rates of sulfur oxides
during atmospheric transport over the Atlantic. Tellus 18:355-372, 1976.
Raddum, G. Invertebrates: quality and quantity as fish food In: Hendrey, G. R. , ed. ,
Limnological Aspects of Acid Precipitation, Rep. Internat. Workshop. Brookhaven National
Laboratory Report No. 51074, 1978.
Reid, G. K. , and R. D. Wood. Ecology of Inland Waters and Estuaries. 2nd ed. D. Von Nostrand
Co. New York, 1976. p. 261-319.
Rheinheimer, G. Aquatic Microbiology. John Wiley & Sons, New York, 1971. 184 pp.
Roff, J. R. , and R. E. Kwiatkowski. Zooplankton and zoobenthos communities of selected
northern Ontario lakes of different acidities. Can. J. Zool. 55:899-911, 1977.
Rosenqvist, I. T. A contributor towards analysis of buffer properties of geological materials
against strong acids in precipitation water, report written for the Council for Research
in Natural Sciences, Norwegian General Research Council, 1976.
SOXJ7A/B 7-120 2-10-81
-------
Rosseland, B. 0., 1. Sevaldrun, D. Svalastog, and I. P. Muniz. Studies on freshwater fish
populations - effects of acidification or reproduction, population structure, growth and
food selection. Ln: Proc. International Conference on the Ecological Impact of Acid
Precipitation, 11-14 March 1980, SNSF-project. (in press).
Schaffer, R. J. The weathering of natural building stones. Spec. Ref. No. 18, London, HMSO.
149 pp. Cited in: Sereda, P. J. Effects of Sulphur on Building Materials.
Scheider, W. , J. Adamski, and M. Paylor. Reclamation of Acidified Lakes Near Sudbury,
Ontario. Rexdale, Ontario. Canada: Ontario Ministry of the Environment, 1975. 129 pp.
Scheider, W. , J. Adamski, and M. Paylor. Reclamation of Acidified Lakes Near Sudbury, Ontario
(Ontario Ministry of the Environment, Rexdale, Ontario, Canada, 1975. Cited in: Wright,
R. F. , and Egil. T. Gjessing. Changes in the chemical composition of lakes. Ambio.
5:219-223, 1976.
Scheider, W. , J. Adamski, and M. Paylor. Reclamation of Acidified Lakes Near Sudbury, Ontario
(Ontario Ministry of the Environment, Rexdale, Ontario, Canada, 1975. Cited in: Wright,
R. F. , and Egil. T. Gjessing. Changes in the chemical composition of lakes. Ambio.
5:219-223, 1976.
Schindler, D. W. Whole lake eutrophication experiments with phosphorus, nitrogen and carbon.
Verb. Internat. Verein. Limnol. 19:577-582, 1975.
Schindler, D. W., R. Wagemann, R. B. Cook, T. Ruszczynski, and J. Prokopowich. Experimental
acidification of Lake 223, experimental lakes area: background data and the first three
years of acidification. Can. J. Fish. Aquat. Sci. 37:342-354, 1980.
Schlesinger, W. H. , and M. M. Hasey. The Nutrient Content of Precipitation, Dry Fallout, and
Intercepted Aerosols in the Chaparral of Southern California. Am. Midland Naturalist
103:114-122, 1980.
Schmel, G. A. Particle and gas dry deposition: a review. Atmos. Environ. 14:983-1011, 1980.
Schofield, C. L. Acid precipitation: Effects on fish. Ambio. 5:228-230, 1976.
Schofield, C. L. Lake acidification in the Adirondack Mountains of New York: Causes and
consequences. In: Proceedings of the First International Symposium on Acid Precipita-
tion and the Forest Ecosystem, May 12-15, 1975. Columbus, Ohio, L. S. Dochinger and T. A.
Seliga, eds. , U.S. Forest Service General Technical Report NE-23, U.S. Department of
Agriculture, Forest Service, Northeastern Forest Experiment Station, Upper Darby, PA,
1976. p. 477. (Abstract)
Schofield, C. L. Acidification of Adirondack lakes by atmospheric precipitation: extent and
magnitude of the problem. Final Rep. D. J. Project F-28-R, NYS Dept. Env. Cons., 1976b.
11 pp.
Schofield, C. L. Dynamics and management of Adirondack fish populations: acidification of
Adirondack lakes by atmospheric precipitation - long-term and seasonal trends and effects
of lake acidification on fish populations. Research Report F-28-R-4, N.Y.S. Dept. of
Env. Cons., 1976c.
Schofield, C. L. Acid precipitation's destructive effects on fish in the Adirondacks. N.Y.
Food Life Sci. Q. 10(3):12-15, 1977.
SOXJ7A/B 7-121 2-10-81
-------
Schofield, C. L. The Acid Precipitation Phenomenon and Its Impact in the Adirondack_Mountains
of New York State. In: Scientific Paper from the Public Meeting on Acid Precipitation,
May 4-5, 1978. Lake Placid, New York, published by Science and Technology Staff, New York
State Assembly, Albany, NY, March, 1979. p. 86-91.
Schofield, C. Effects of Acid Rain on Lakes. Presented at the American Society of Civil Eng.
Convention, Boston. April 2, 1979. 13 pp.
Schofield, C. , and J. R. Trojnar. Aluminum toxicity to brook trout (Salvelinus fontinalis)
in acidified waters, pp. 341-365. Iji: T. Y. Toribara, M. W. Miller and P. E. Morrow,
eds., Polluted Rain. Plenum Press, New York, 1980.
Seip, H. M. Acidification of freshwater-sources and mechanisms. Iji: Proceedings of the
International Conference on the Ecological Impact of Acid Precipitation, D. Drabltfs and
A. Tollan, eds., SNSF Project Report, Oslo, Norway, 1980.
Sereda, P. J. Chapter 8. Effects of Sulphur on Building Materials, p. 359-426. In: Sulphur
and its Organic Derivatives in the Canadian Environment. Nat. Res. Council of Canada.
NRC Associate Committee on Scientific Criteria For Environmental Quality. Ottawa,
Canada, 1977.
Sheridan, R. P. , and R. Rosenstreter. The effect of hydrogen ion concentration in simulated
rain on the moss Tortula rural is (Hedw) Sm. Bryologist 76:168-173, 1973.
Shinn, J. H. , and S. Lynn. Do man-made sources affect the sulfur cycle of northeastern
states? Environ. Sci. and Tech. 13:1062-1067, 1979.
Shriner, D. S. , M. DeCot, and E. B. Cowling. Simulated acid rain caused direct injury to
vegetation. Proc. Am. Phytopath. Soc. 1:112, 1974.
Shriner, D. S. Effects of simulated rain acidified with sulfuric acid on host-parasite inter-
actions. In: Proceedings of the First International Symposium on Acid Precipitation and
the Forest Ecosystem, May 12-15, 1975, Columbus, Ohio, L. S. Dochinger and T. A. Seliga,
eds., U.S. Forest Service General Technical Report NE-23, U.S. Department of Agriculture,
Forest Service, Northeastern Forest Experiment Station, Upper Darby, PA, 1976. pp.
919-925.
Shriner, D. S. , and G. S. Henderson. Sulfur distribution and cycling in a deciduous forest
watershed. J. Environ. Qua!. 7:392-397, 1978.
Shriner, D. S. Effects of simulated acidic rain on host-parasite interactions in plant
diseases. Phytopathology 68:213-218, 1978a.
Shriner, D. S. Interactions between acidic precipitation and S09 or 0,: Effects on plant
response. Phytopathol. News, 12:153, 1978b. L 6
Shriner, D. S. Chapter 11. Atmospheric deposition: monitoring the phenomenon; studying the
effects. |n: Handbook of Methodology for the Assessment of Air Pollutant Effects on
Vegetation. W. W. Heck, S. V. Krupa, and S. N. Linzon, Editors. Air Pollution Control
Association, Pittsburgh, Pennsylvania, 1979.
Shriner, D. S. Terrestrial vegetation - air pollutant interactions: Nongaseous pollutants,
wet deposition. Paper presented at the International Conference on Air Pollutants and
Their Effects on Terrestrial Ecosystems, Banff, Alberta, Canada, May 10-17, 1980.
Sinclair, W. A. Polluted air: Potent new selective force in forest. J. Forestry.
67:305-309, 1969.
SOXJ7A/B 7-122 2-10-81
-------
Smith, F. B. , and G. H. Jeffrey. Airborne transport of sulfur dioxide from the United
Kingdom. Atmos. Environ. 9:643-659, 1975.
Smith, R. A. Air and Rain: The Beginnings of Chemical Climatology. Longmans, Green, London,
1872. 600 pp.
Smith, R. L. Ecology and Field Biology, 3rd ed. , p. 170. New York, Harper and Row, 1980.
pp. 11-199.
SNSF Project. Acid precipitation and some alternative sources as the cause of the acidifying
water courses. Norway's Agrarian Science Research Board, Norway's Technical Nature
Science Research Board, Dept. of Environmental Protection, ed., A. Tollan, 1977.
Sprules, G. W. Midsummer crustacean zooplankton communities in acid-stressed lakes. J. Fish
Res Board Can. 32:389-395, 1975.
Stensland, G. J. A Comparison of Precipitation Chemistry Data At A Central Illinois Site in
1954 and In 1977. Presented at 71st Annual Meeting of the Air Poll. Control. Assoc. ,
Houston, TX. June 25-26, 1978.
Stensland, G. J., Precipitation chemistry trends in the northeastern United States. j_n:
Proceedings twelfth Annual Rochester International Conference on Environmental Toxicity:
Polluted Rain, May 21-23, 1979, Rochester, N. Y. T. Y. Toribara, M. W. Miller, P. E.
Morrow, eds. Plenum Press, NY, 1980. pp. 87-103.
Stokes, P. Phytoplankton of acidic lakes in Killarney Ontario: community structure related
to water chemistry. In: Proc. International Conference on the Ecological Impact of Acid
Precipitation, March 1980, SNSF-project report, Oslo, Norway, (in press).
Stumm, W., and J. J. Morgan. Aquatic Chemistry, Wiley-Interscience, New York, NY, 1970. 583
pp.
Sutcliffe, D. W. , and T. R. Carrick. Studies on mountain streams in the English Lake District
I. pH, calcium, and the distribution of invertebrates in the River Duddon. Freshwater
Biol. 3:437-462, 1973.
Tabatabai, M. A., and J. M. Laflen. Nitrogen and Sulfur Content and pH of Precipitation in
Iowa. J. Environ. Quality. 5:108-112, 1976.
Tamm, Co. 0., Wiklander, G. , and Popovie, B. Effects of application of acid to poor pine
forests, Water, Air and Soil Pollution, 8:75-87, 1977.
Tomlinson, G. H. Acidic Precipitation and Mercury In Canadian Lakes and Fish. Scientific
Paper From the Public Meeting on Acid Precipitation, May 4-5, 1978. Lake Placid, N.Y.
Sponsored by The Committee on Environmental Conservation, New York State Assembly. Pub.
by Science and Technology Staff, New York State Assembly, p. 104-118. March 1979.
Traaen, T. , and M. Laake. Microbial decomposition and community structure as related to
acidity of fresh waters - an experimental approach. J_n: Proc. Internat. Conf. on Ecolo-
gical Impact of Acid Precipitation, March 1980, SNSF-project report, Oslo, Norway, (in
press).
Trojnar, J. R. Egg and larval survival of white suckers (Catostomus commersoni) at low pH.
J. Fish. Res. Board Can. 34:262-266, 1977.
Trojnar, J. R. Egg hatchability and tolerance of brook trout (Salvelinus fontinalis) fry at
low pH. J. Fish. Res. Board Can. 34:574-579, 1977.
SOXJ7A/B 7-123 2-10-81
-------
Tukey, H. B. , Jr. The leaching of substances from plants. Ann. Review of Plant Physiology.
21: 305-324, 1970.
Turk, J. T. , and N. E. Peters. Acid-rain weathering of a metasedimentary rock basin, Herkimer
County, New York. U.S. Geol. Surv. Open file Report 77-538, 1977. 14 pp.
Turk, J. T. , and N. Peters. Acid-rain Weathering of a Metasedimentary Rock Basin, Herkimer
County, New York. Scientific Paper From the Public Meeting on Acid Precipitation, May
4-5, 1978. Lake Placid, N.Y. Sponsored by The Committee on Environmental Conservation,
New York State Assembly. Pub. by Science and Technology Staff, New York State Assembly.
p 136-145. March 1979.
Turner, J. , and M. Lambert. Sulfur nutrition of forests. I_n: Atmospheric Sulfur Deposition
Environmental Impact and Health Effects. Proceedings of the Second Life Science
Symposium, Potential Environmental Health Consequences of Atmospheric Deposition, D. E.
Shriner, S. R. Richmond, and S. E. Lindberg, eds. , Ann Arbor Science, 1980. pp. 321-333.
Tyler, G. Heavy metals pollute nature, may reduce productivity, Ambio, 1, (2), pp. 52-59,
1972.
Tyler, G. Leaching rates of heavy metal ions in forest soil. Water, Air and Soil Pollution,
9, pp. 137-148, 1977.
U.S. Bureau of Mines. Minerals Yearbook 1952. Volume II. Area Reports: Domestic. U.S.
Department of the Interior, Washington, DC, 1954. p. 117.
U.S. Bureau of Mines. Minerals Yearbook 1974. Volume II. Area Reports: Domestic. U.S.
Department of the Interior, Washington, DC, 1976. p. 392.
U.S. Department of Commerce. Climatic Atlas of the United States. National Climatic Center,
Asheville, NC, June 1968.
U.S. Department of the Interior, Fish and Wildlife Service. Impacts of Cool-Fired Power
Plants on Fish, Wildlife, and their Habitats. FWS/OBS-78/29, March, 1978. p. 64-70.
U.S. Environmental Protection Agency. National Air Quality, Monitoring, and Emissions Trends
Report, 1977. EPA-450/2-78-052, Office of Air Quality, Standards, and Planning. U.S.
Environmental Protection Agency, Research Triangle Park, NC, December, 1978.
Ulrich, B. Die Umweltbeeinflussung des Nahrstoffhaushaltes eines bodensauren Buchenwalds.
Forstwiss. Centralbl. 94:280-287, 1975.
Walters. C. J. , and R. E. Vincent. Potential productivity of an alpine lake as indicated by
removal and reintroduction of fish. Trans. Amer. Fish Soc. 102:675-697, 1973.
Wetzel, R. G. Limnology, p. 287-621, W. B. Saunders Co., Philadelphia, PA, 1975. 743 pp.
Whelpdale, D. M. Acidic Deposition. Sect. 4.1. Precipitation Chemistry and Deposition. I_n:
Ecological Effects of Acid Precipitation. M. J. Wood, ed. , report of workshop held at
Calley Hotel, Gatehouse-of-Fleet, Galloway, U.K., September 4-7, 1978. EPRI SOA77-403,
Electric Power Res. Institute, Palo Alto, CA, 1979. pp. 2-12.
Wiklander, L. Interaction between cations and anions influencing adsorption and leaching.
Jji: Effects of Acid Precipitation on Terrestrial Ecosystems. T. C. Hutchinson and H.
Havas, eds. Plenum Press, New York, NY, 1980.
SOXJ7A/B 7-124 2-10-81
-------
Wiklander, L. Leaching and Acidification of Soils. In: Wood, M. J. (ed. ). Ecological
Effects of Acid Precipitation. Report of workshop held at Cally Hotel, Gatehouse-of-
Fleet, Galloway, U.K. 4-7 Sept. 1978. EPRI SOA77-403, Electric Power Res. Institute,
3412 Hillview Ave., Palo Alto, California 94303, 1979.
Williams, Wayne T. Acid Rain: The California Context. Citizens for Better Environment.
Environmental Review, May 1978. p. 6-8, 10.
Winkler, E. M. Important Agents of Weathering For Building and Monumental Stone. Eng. Geol
1:381-400, 1966.
Wodzinski, R. S. , D. P. Labeda, and M. Alexander. Toxicity of S0? and NO : Selective inhibi-
tion of blue green algae by bisulfite and nitrite. J. Air Pollx Control Assoc 27-
891-893, 1977. —
Wolff, G. T. , P. J. Lioy, H. Golub, and J. S. Hawkins. Acid precipitation in the New York
Metropolitan Area: Its relationship to meteorological factors. Env. Sci. and Tech.
13:209-212, 1979.
Wood, T. Acid Precipitation. In: Sulfur in the Environment. Missouri Botanical Garden in
cooperation with Union Electric Company, St. Louis, MO. 1975. pp. 39-50.
Wood, T. , and F. Bormann. Increases in foliar leaching by acidification of an artificial
mist. Ambio 4:169-171, 1975.
Wood, T., and F. H. Bormann. Short-term effects of a simulated acid rain upon the growth and
nutrient relations of Pinus strobus L. Water, Air Soil Pollut. 7:479-488, 1977.
Wood, T., and F. H. Bormann. The effects of an artificial acid mist upon the growth of Betula
alleghaniensis. Brit. Environ. Pollut. 7:259-268, 1974.
Woodwell, G. M. Effects of ionizing radiation on terrestrial ecosystems. Science.
138:572-577, 1962.
Woodwell, G. M. Effects of pollution on the structure and physiology of ecosystems. Science.
168:429-433, 1970.
Woodwell, G. M. The ecological effects of radiation. Scientific American. 208:40-49, 1963.
Wright, R. F. , A. Henriksen, R. Harriman, B. Morrison, and L. A. Caines. Acid lakes and
streams in the Galloway area, southwestern Scotland. In: Proc. International Conference
on the Ecological Imapct of Acid Precipitation, 11-14 March 1980, SNSF-project, 1980a.
(in press).
Wright, R. F. , and A. Henriksen. Chemistry of small Norwegian lakes with special reference to
acid precipitation. Limnol. Oceanog. 23:487-498, 1978.
Wright, R. F. , and E. Snekvik. Acid precipitation: Chemistry and fish populations in 700
lakes in southernmost Norway. Verh. Int. Ver. Theor. Angew. Limnol. 20:765-775, 1978.
Wright, R. F. , and E. T. Gjessing. Changes in the chemical composition of lakes. Ambio.
5:219-223, 1976.
Wright, R. F. , N. Conroy, W. T. Dickson, R. Harriman, A. Henriksen, and C. L. Schofield.
Acidified lake districts of the world: a comparison of water chemistry in southern
Norway, southern Sweden, southwestern Scotland, the Adirondack Mountains of New York, and
southeastern Ontario. JJK Proc. International Conference on the Ecological Impact of
Acid Precipitation, 11-14 March 1980, SNSF-project, Oslo, Norway, 1980b. (in press).
SOXJ7A/B 7-125 2-10-81
-------
Wright, R. F. , T. Dale, E. T. Gjessing, G. R. Hendry, A. Henriksen, M. Johannessen, and I. p.
Muniz. Impact of Acid Precipitation on Freshwater Ecosystems in Norway. Water, Air and
Soil Poll. 6:483-499, 1976.
Van, N. D. , and P. Stokes. Phy to plankton of an Acidic Lake, and Its Responses to
Experimental Alterations of pH. Environ. Conservation 5:93-100, 1978.
SOXJ7A/B 7-126
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8. EFFECTS ON VEGETATION
8.1 GENERAL INTRODUCTION AND APPROACH
The objective of this chapter is to review available data relating atmospheric concentra-
tions of S02 and participates to effects on terrestrial vegetation. Many reviews of the
general effects of S02 (Daines, 1968; Guderian, 1977; Jacobson and Hill, 1970; Linzon, 1978;
Mudd, 1975; MAS, 1978; Treshow, 1970; U.S. Environmental Protection Agency, 1973, 1978; Van
Haut and Stratmann, 1970) and, to a lesser extent, of particulate matter (Guderian, 1981)
exist in the literature. Additional reports of S02 effects on plants for use in diagnostic
situations have also been prepared (Davis, 1972a,b; Lacasse and Moroz, 1969; U.S. Environ-
mental Protection Agency, 1976).
This chapter has a more specific purpose. It addresses factors which influence our
ability to quantify relationships between exposure dose and plant response. Principal focus
is placed on quantifying specific concentrations of SO^ and particulates associated with vege-
tation responses at levels ranging from biochemical to that of plant populations. In this
process, information of historical interest has been kept to a minimum and emphasis has been
placed rather on more recent studies that have employed modern monitoring, experimental, and
statistical techniques. Since air quality standards are formulated around effects of single
pollutants, only studies in which S0? or particulates were determined to be the major cause of
measured effects have been included.
As a backdrop against which to consider pollutant effects on plants, it is important to
recognize that the many factors which play a major role in determining the likelihood that a
given quantity of pollutant will produce a known level of effect vary tremendously in nature.
These factors include the type of exposure (acute or chronic), influences of stress from other
biotic (plants, disease) or abiotic (edaphic or climatic) factors, the type of response mea-
sured, and the species or population under study. These factors and associated terminology
have been addressed in Sections 8.2 and 8.4 for SO,,, and Sections 8.6 and 8.7 for particulate
matter.
While a broad variety of responses measured following exposure of vegetation to SO,, or
particulates are discussed, it should be noted that not all responses are negative, and that
all short-term negative responses do not ultimately result in negative effects on plant growth
and development. These concepts are developed more fully in Sections 8.2.7 and 8.3.
The end-point of this presentation of concepts, components, and modifiers of pollutant
dose and plant response can be found in attempts to generalize dose-response relationships.
This is done in Sections 8.3 and 8.8 for S02 and particulates, respectively. With SO,, dose-
response, mathematical models are presented for visible injury to foliage of plants, for num-
bers of species in a plant population, and for degree of effects on plant growth and yield.
The latter effort attempts to synthesize both average and "upper limit" conditions from a
broad variety of species, sites, and exposure conditions.
XRD8A/A 8-1 2-9-81
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The concluding section on ecosystem responses (Section 8.10) derives information
from a much more limited number of studies in this area. Here reliance on the broadly
based concepts of ecosystem analyses, in many cases, form the basis for strong inference
rather than proof of effects on a more subtle scale. In this area definitive data to
evaluate the degree or extent of ecosystem changes over broad regions do not at present
exist.
8.2 REACTION OF PLANTS TO S02 EXPOSURES
8.2.1 Introduction to Terminology
As with all plant stress-inducing agents, SO- may initiate changes within plant meta-
bolic systems that may substantially lead to extensive physiological dysfunctions; and if
sufficient physiological modification occurs, visible symptoms may be manifested. In some
instances, the addition of low S0? concentrations to a plant's environment may induce a
fertilizer-like response, but only relatively few studies of this phenomenon in agronomic
crops have been completed to date; none have shown beneficial effects on natural ecosys-
tems.
The following definitions as relevant to this document have generally been acceptable
to plant scientists working on air pollutant induced effects to plants (American Phytopath.
Soc., 1974).
Chronic Injury (effects) - injury which develops only after long-term or repeated
exposure to an air pollutant and is expressed as chlorosis, bronzing, premature senescence,
reduced growth, etc., can include necrosis.
Acute Injury (effects) - injury, usually involving necrosis, which develops within
several hours to a few days after short-time exposure to a pollutant, and is expressed
as fleck, scorch, bifacial necrosis, etc.
Injury - a change in the appearance and/or function of a plant that is deleterious
to the plant.
Damage - a measure of decrease in economic or aesthetic value resulting from plant
injury by pollutants.
Plant death may result from continual exposure to low or high pollutant doses and, if
such is the case, other mitigating factors may also be involved, i.e., abiotic or biotic
disease-inducing agents or insect attack. Depending upon the plant species, exact condi-
tions of the seasonal stage of crop growth, pollutant dose and environmental conditions,
many forms of injury may take place and their relative impact may vary. Symptoms of acute
and chronic injury may occur on a given plant simultaneously.
Injury does not necessarily imply damage, i.e., economic loss. Timing of pollutant
exposure in relation to the stage of crop development often determines the relationship
of foliar injury to subsequent yield losses.
The influence of S02 as affecting plant health is a complex process that involves not
only pollutant concentration and duration of exposure but also environmental factors and
XRD8A/A 8-2 2-9-81
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*
then influence on the overall response of the plant itself as a biological entity under
stress. In simplistic overview, this process involves pollutant ingress through gas phase
interaction at leaf and stomatal surfaces, contact with wet cellular membranes and subse-
quent liquid phase reactions, induced perturbations of metabolic and/or physiological sys-
tems, plant responses through homeostasis (repair) mechanisms, and some resultant disposi-
tion of plant health. Figure 8-1 has been modified from a similar description as prepared
by Tingey and Taylor (1980) for plant response to ozone.
There are several possible plant responses to S02 and related sulfur compounds: (1)
fertilizer effects appearing as increased growth and yields, (2) no detectable responses,
(3) injury manifested as growth and yield reductions without visible symptom expressions
on the foliage or with very mild foliar symptoms that would be difficult to perceive as
air pollution incited without the presence of a control set of plants grown in pollution-
free conditions, (4) injury exhibited as chronic or acute symptoms on foliage with or
without associated reductions in growth and yield, and (5) death of plants and plant com-
munities (Figure 8-1).
8.2.2 Wet and Dry Deposition of Sulfur Compounds on Leaf Surfaces
Deposition processes limit the lifetime of sulfur compounds in the atmosphere, con-
trol the distance traveled before deposition, and limit the atmospheric pollutant concen-
trations (Garland, 1978).
There have been several studies of the deposition of particulate material to natural
surfaces (Chamberlain, 1975; Little and Wiffen, 1977; Sehmel and Hodgson, 1974). Very
large particles are chiefly deposited by sedimentation. Particles in the range of 1 to 100 |jm
are also borne towards the surface by turbulence where sedimentation is supplemental to
impaction on rough surfaces. Submicron-sized particles (e.g., sulfuric acid aerosols) diffuse
by Brownian motion through the thin laminar layers close to the plant surface. This may
be followed by active uptake by plants. The mean SO,, deposition velocities are surpris-
ingly similar for a wide range of plant leaf surfaces (Garland, 1978).
Dry deposition results in the removal of significant amounts of the larger particles
from the atmosphere within 2 or 3 days following emission, but several weeks are required
to remove the submicrometer fraction.
8.2.3 Routes and Methods of Entry Into the Plant
Stomata of leaves have been demonstrated to be the major avenue of SO- entrance into
plants. Although this is a widely accepted conclusion that has been presented in numerous
reviews (Guderian, 1977; Katz, 1949; Thomas and Hendricks, 1951, 1956; U.S. Environmental
Protection Agency, 1973), there is still controversy as to the importance of stomatal move-
ment relative to plant biochemistry in determining plant sensitivity. Many factors that
govern the mechanism of stomatal opening and closing have been determined to be indepen-
dent of SO- exposure concentrations. Physical factors such as light, leaf surface mois-
ture, relative humidity, and soil moisture availability influence stomatal opening and
XRD8A/A 8-3 2-9-81
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CONDUCTANCE
GAS PHASE LIQUID PHASE
EFFECT
FERTILIZER
NON^E
< YIELD: NO SYMPTOMS
SY MPTOMS_
ACUTE_SY^PTOMS
DEATH OF POPULATIONS
HOMEOSTASIS
Figure 8-1. A conceptual model of potential responses by plants following exposure to various doses of
Sulfur dioxide. Model modified from similar treatment of plant response to ozone.
Source: Tingey and Taylor, 1980.
8-4
-------
closing and play a major role in plant sensitivity by limiting passive entry of S02 into the
leaf (Domes, 1971; Meidner and Mansfield, 1968; Setterstrom and Zimmerman, 1939; Spedding,
1969; Mclaughlin and Taylor, 1981). These factors must therefore be considered when determin-
ing plant sensitivity or tolerance to entry of SCL.
Internal resistances to flux of gases into plant leaves may also be substantial and may
exceed those imposed by stomata under some conditions. Barton et al. (1980) found that photo-
synthetic depression in kidney beans (Phaseolus vulgaris) during SCL exposure was explained
predominantly by increases in mesophyll resistance. Stomatal resistances changed only slightly
and were a minor component of total leaf resistance to CCL at both high (71 percent) and low
(33 percent) relative humidity. Winner and Mooney (1980) also found that differences in non-
stomatal components of leaf resistance to SOp uptake were associated with differences in
resistance of deciduous and evergreen shrubs to S0?.
The absorption rate of S02 into plants varies not only among species but also with
previous exposure to SOp. Bigtooth aspen (Populus grandidentata) had a higher S0? absorption
rate during exposure to high concentrations (2.75 pmm) of S0? for 2 hr than did white ash
(Fraxinus americana) or yellow birch (Betula alleghaniesis) when they were exposed 0 to 6 hr
prior to fumigation with 0.75 ppm S0? (Jensen and Kozlowski, 1975). However, after 20 to 36
hr of pre-fumigation treatment, the rate of SOp absorption during the 2.75 ppm SO- exposure
was greater for birch and ash. The sulfur content of foliage increased in all species. Eight
days after fumigation with SOp, varying amounts of the labelled S were translocated through-
out the plants including roots (Jensen and Kozlowski, 1975). Subsequent effects were not
indicated.
Sulfur dixoide has been shown to increase or decrease stomatal resistance and thus affect
potential photosynthetic performance (Hallgren, 1978). Sulfur dioxide induced the closure of
"Pelargonium x hortorum" stomata especially when they had been fully opened, and necrosis was
not averted (Bonte et al., 1975). Kodata and Inoue (1972) demonstrated that S0? entered
leaves of Pinus resinosa through stomata and was accumulated in the cells around stomata for
some time before diffusing inward through the leaf; i.e., internal diffusion was slower than
diffusion into the leaf.
Once SOp has entered, it may induce stomata to remain open for longer periods of time or
to open wider than before fumigation. Exposure to SOp (0.5 ppm) at relative humidities above
40[SOp] percent caused an increase in stomatal opening (Majernik and Mansfield, 1970; Mansfield
and Majernik, 1970). A 3-minute fumigation with 2.5 ppm SO,, increased carbon dioxide uptake
and stomatal opening in Si napsis alba plants. However, with the same concentration, suppressed
carbon dioxide uptake and stomatal closure have also been noted (Buron and Cornic, 1973).
XRD8A/A 8-5 2-9-81
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*
8.2.4 Cellular and Biochemical Changes
Based on the available literature, it is difficult to assess the relationship of S02-
induced biochemical and/or physiological changes at the cellular level to subsequent
effects on photosynthetic activity or resultant growth and yield. Numerous studies have
utilized detached leaves and/or isolated chloroplasts in culture solutions for evaluation
of effects, but their use for field estimations under ambient conditions remains limited.
Recent studies have also shown a variety of SCL-induced biochemical effects: enzyme
inhibition (Pahlich, 1971, 1973; Ziegler, 1972); interference with respiration (Haisman,
1974); interference with energy transduction (Ballentyne, 1973); interference with lipid
biosynthesis (Malhotra and Kahn, 1978); alterations in amino acid content and quality
(Godzik and Linskens, 1974); and chlorophyll loss (Rao and LeBlanc, 1965). Pahlich (1975)
has rationalized some of this diverse list of effects in terms of sulfite and sulfate
accumulation by exposed plant tissues.
Vogl et al. (1965) attempted to integrate biochemical responses with the type and
magnitude of resultant plant effects (Table 8-1). Development of models is necessary to
relate changes in physiology (biochemical responses) of specific plant species to altered
growth and productivity.
Horsman and Wellburn (1976) prepared the most complete listing of reported metabolic
or enzymatic effects of S09 on plants or plant tissues. In only one of eleven studies
14
(5 references) was an increase in photosynthesis ( C02 fixation) in response to exposure
to S02 or its derivatives reported.
With SO™, which upon absorption is further oxidized to S03 and SO. and subsequently
incorporated into S-containing amino acids and proteins, the rate of entry is particularly
important to determining toxicity. Plants have an inherent, and apparently species-
dependent, capacity to absorb, detoxify, and metabolically incorporate S02 and may absorb
low concentrations of S0? over long time periods without damage. Thomas et al. (1943),
for example, exposed alfalfa to S0? continuously at 0.20 ppm for 8 weeks without adverse
effects. Toxicity to S02 may occur during short episodes when the S0?-S0. conversion rate
is exceeded and the extremely toxic sulfite SO., form, a partially oxidized metabolic inter-
mediate, accumulates (Ziegler, 1975). During longer exposures at lower S02 levels, SO^
may accumulate as the rate of metabolic incorporation of SO. is exceeded, and chronic
symptoms may appear.
8.2.5 Acute Foliar Injury
This type of injury occurs following rapid absorption of a toxic dose of S02 and
results at first in marginal and interveinal areas having a dark-green, watersoaked appear-
ance. After desiccation and bleaching of tissues, the affected areas become light ivory
to white in most broadleaf plants. Some species show darker colors (brown or red), but
there is characteristically an exact line of demarcation between symptomatic and asympto-
matic portions of leaf tissues. Bifacial necrosis is common. In monocotyledons,
XRD8A/A 8-6 2-9-81
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TABLE 8-1. RELATIONSHIP OF BIOCHEMICAL RESPONSE TO VISUAL SYMPTOMS OF PLANT INJURY
Degree
of
injury
A
B
C
Visual
symptoms
none
not detectable
loss of assimi-
lation capacity
through:
1) premature death
of assimilation
organs (leaves,
Symptoms of biochemical
Injury in leaf cells
stress on
buffer systems
photosynthesis
adversely affected,
diminished assimi-
lation rate
diminished activity
of enzymes
effect upon
chlorophyll
Description of
Injury To:
Assimilation
organs
none
temporary
Impedance of
gaseous exchange
prolonged imped-
ance of gaseous
exchange
Injury
Whole plant
none
not detectable
reduced growth
(deficiency
conditions)
Ability to
Assimilation
organs
very quick,
completely
slowly, com-
pletely
recover in:
Whole plant
slowly, completely
for perennials
oo
needles)
2) diminished
growth of new
tissues (short-
er needles, etc.)
D
E
necrosis of the
assimilating and
active plant tis-
sues
destruction of
all important
assimilate ry
plant tissues
death of cells
through protein
and enzyme degrada-
tion
death of organs
irreversible
injury: necro-
sis of some
assimilation
organs or parts
thereof
irreversible
injury to all
assimilation
organs
loss of assimi-
lation capabil-
ity
destruction
of assimila-
tion capability
quick, not com-
pletely, some-
times (for iso-
lated tissues)
not any more
not any more
slowly, completely
for perennials
sometimes (for
isolated tissues)
-------
(corn, grasses) foliage injury occurs at the tips and in lengthwise strips along the
parallel veins (U.S. Environmental Protection Agency, 1976).
In conifers, acute injury on foliage usually appears as a bright orange red tip
necrosis on the current-year needles, often with a sharp line of demarcation between the
injured tips and the normally green bases. Occasionally, the injury may occur as bands at
the tip, middle, or base of the needles (Linzon, 1972).
Recently incurred injury is light colored; but later, bright.orange or red colors are
typical for the banded areas and tips. As needle tips die, they become brittle and break
or whole needles drop from the tree. Pine needles are most sensitive to SO- during the
period of rapid needle elongation but injury may also occur on mature needles (Davis,
1972a).
8.2.6 Chronic Foliar Injury
Plant injury that is visible but does not involve collapse and necrosis of tissues is
termed chronic injury. This type of visible injury is usually the result of variable
fumigations consisting of both short-term, high-concentration or long-term, low-concen-
tration exposures to S02- It has also been referred to as "sulfate injury" since a slow
accumulation of sulfate is the end result of such exposures (Daines, 1968). Within sub-
stomatal cavities, SO^ reacts with intercellular water to quickly form sulfite and bisul-
fite, which are slowly oxidized to sulfate which is approximately 30 times less toxic than
sulfite and bisulfite (Thomas, 1951). The capacity of the plant tissues to convert sulfite
and bisulfite to sulfate may never be exceeded, and visible expression of symptoms will not
occur. In some studies, sulfate levels in plants exposed to S02 have been demonstrated to
be several times greater than those in controls (Linzon, 1958). As sulfite and bisulfite
ions are formed, and as sulfate accumulates to phytotoxic levels, then chronic symptoms first
appear as various forms of chlorotic (yellowing) patterns. As sulfite and bisulfite ions
continue to accumulate, destruction of individual chloroplast membranes or a reduction of
chlorophyll production ensues resulting in reddening or bleaching of cells without necro-
sis (Thomas, 1951). Following such accumulations, there is a fine distinction between
chronic and acute symptom expressions.
In broadleaf plants, chronic injury is usually expressed in tissues found between the
veins, with various forms of chlorosis predominating. Chlorotic spots or chlorotic
mottle may persist following exposure or may subside and disappear following pollutant
removal or as a result of changing environmental conditions (Jacobson and Hill, 1970).
Chronic effects of S02 in conifers are generally first expressed on older needles
(Linzon, 1966). Chlorosis of tissues starting at the tips progresses down the needle
towards the base, i.e., symptoms progress from the oldest to youngest tissues. Advanced
symptoms may follow, involving reddening of affected tissues (Linzon, 1978).
XRD8A/A 8-8
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*
8.2.7 Classification of Plant Sensitivity to S0a
Because of space limitation, it is not possible to list all plants that are known to be
sensitive to various doses of S02. Furthermore, in a listing of sensitive plants, the
evidence collected should also indicate the environmental, genetic, and cultural considera-
tions that may in fact determine such sensitivities. It has also been demonstrated that plant
response to air pollutants varies at the genus, species, variety, and cultivar levels. Lists
of plant sensitivities have been prepared on the basis of the expression of visible symptoms
by any given plant. Injury expressed by growth or yield losses has not been considered in the
preparation of such lists.
Jacobson and Hill (1970) included a listing of plants sensitive to the major phytotoxic
air pollutants. Linzon (1972) has listed 36 tree species as being tolerant, intermediate, and
sensitive to S02- Many of these sensitivity lists have not attempted to identify the dose
required to induce visible injury on indicator species. However, Jones et al. (1974) have
published such details based upon observations over a 20-year period of 120 species growing in
the vicinity of coal-fired power plants in the southeastern United States (Table 8-2).
Mclaughlin (1980) used symptom data as collected by Dreisinger and McGovern (1970) on 31
species of forest and agricultural plants following SO- exposure and plotted the average,
maximum, and minimum tolerances of individual species (Figure 8-2). The injury threshold for
most sensitive plants was 0.41, 0.37, 0.28, and 0.12 ppm S0? at averaging intervals of 1, 2,
4, and 8 hours, respectively.
Other compilations have been presented. The report of Davis and Wilhour (1976) provides
information on an international basis. Specific reports have been prepared for vegetation
native to the southwestern deserts of the United States (Hill et al., 1974).
Extensive efforts have been made to develop certain plant species as bioindicators.
Perhaps the most extensively examined plants for this use are eastern white pine (Pinus
strobus) and numerous species of lichens. The white pine literature has been reviewed
(Gerhold, 1977), and the most recent review of lichen bioindicators was prepared by LeBlanc
and Rao (1975).
Other recent reports have been prepared for various ornamentals (Daessler et al., 1972);
Heggestad, 1973; Pelz, 1962), bluegrass cultivars (Murray et al., 1975), scotch pine (Demeritt
et al., 1971), hybrid poplar (Dochinger and Jensen, 1975), and trembling aspen (Karnosky,
1977). These represent examples of the continued efforts to identify sensitive plants suit-
able for use as bioindicators.
Other incitants such as drought, nutrient imbalances and other pollutants may induce
injury symptoms that mimic those of S02, and several bioindicators are desirable for
evaluation in any given area. In addition, individual species and more complicated plant
bioindicator systems are not as effective for detecting SOp at low concentrations as are
sophisticated instruments.
XRD8A/A 8-9 2-9-81
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TABLE 8-2. SULFUR DIOXIDE CONCENTRATIONS CAUSING VISIBLE INJURY TO
VARIOUS SENSITIVITY GROUPINGS OF VEGETATION3
(ppm S0£)
Maximum Sensitivity grouping
concentration Sensitive
ppm SOy
Peak 1.0-1.5
1-hr 0.5-1.0
3-hr 0.3-0.6
Ragweeds
Legumes
Blackberry
Southern pines
Red and black oaks
White ash
Sumacs
Intermediate
ppm SOy
1.5-2.0
1.0-2.0
0.6-0.8
Maples
Locust
Sweetgum
Cherry
Elms
Tuliptree
Many crop and
garden species
Resistant
ppm SOy
>2.0
>2.0
>0.8
White oaks
Potato
Upland cotton
Corn
Dogwood
Peach
Based on observations over a 20-year period of visible injury occurring on
over 120 species growing in the vicinities of coal-fired power plants in the
southeastern United States. Source: Jones et al., 1973.
XRD8A/A
8-10
2-9-81
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1.0
o
E
u 0.6
o"
ta
0.2
MOST SENSITIVE
1
1
I I
2 4
S
AVERAGING INTERVAL (H)
Figure 8-2. Exposure thresholds for minimum, maximum and average sensitivity of 33 plant
species to visible foliar injury by SO2.
Source: Dreisinger and McGovern (1970) as applied by Mclaughlin (1980).
8-11
-------
8.2.8 Beneficial "Fertilizer" Effects
Under certain conditions, atmospheric S02 can have beneficial effects to agronomic
vegetation (Noggle and Jones, 1979). Sulfur is one of the elements required for plant growth
and Coleman (1966) reported that crop deficiencies of S have been occurring with increasing
frequency throughout the world. Sulfur requirements to maintain high crop production range
from 10 to 40 kg/ha per year. Figure 8-3 presents a map of sulfur deficient soils of the
United States (The Sulphur Institute, 1979).
Cowling et al. (1973) found beneficial effects of S02, such as increases in yield and
sulfur content, in perennial ryegrass that was grown with an inadequate supply of sulfur to
the roots. Faller (1970) conducted a series of experiments to determine effects of varying
atmospheric concentrations of S0? on sunflower, corn, and tobacco. In these studies, plants
were grown in nutrient media containing adequate supplies of all essential elements except
sulfur, which was low. Plants grown in the atmosphere without SOp developed sulfur- deficiency
symptoms within a few days. In other treatments, total plant yield increased to some extent
with increasing concentrations of S02 added to the atmosphere during plant growth. For tobacco,
the total dry weight increased by up to 40 percent. Yields of leaves and stems alone increased
by 80 percent while dry weights of tobacco increased even at the highest SOp concentration used
(0.57 ppm); sunflower and corn had their highest biomass at SOp concentrations of 0.40 ppm and
0.20 ppm, respectively. Beyond these concentrations, visible injury was observed. Additional
studies by Faller (1970) with S suggest that up to 90 percent of plant sulfur requirements
may originate from the atmosphere under the specific experimental conditions.
Although no monitoring or handling procedures for SOp delivery were presented in a study
by Thomas et al. (1943), its results indicated that SOp could serve as a required source of
nutrient S to alfalfa fumigated with approximately 0.10 ppm S0? for 6 to 7 hr/day, 6 days/week
for the growing season. An additional caveat for plants being grown under S-deficient soil
nutrient status must also be considered for this study.
Recently, Noggle and Jones (1979) reported the results of a 2-year study using potted soil
and S as S nutrient addendum to determine the contributions of soil and atmospheric sulfur
to the sulfur requirements in cotton and fescue. Cotton was more efficient than fescue in
accumulating sulfur from the atmosphere. The amount of sulfur accumulated from the atmosphere
was apparently influenced by the amount of sulfur supply in the soil relative to the sulfur
requirements of the plant. A crop grown in a sulfur-deficient soil will accumulate more sulfur
from the atmosphere than the same crop grown in a soil that has an adequate supply of sulfur.
Noggle and Jones (1979) also showed that cotton grown in specifically designed growth con-
tainers in the vicinity of certain coal-fired power plants accumulated significant amounts of
atmospheric sulfur (as S02) and produced significantly more biomass than those grown at a
location remote to the industrial source of sulfur. Thus, under appropriate conditions,
XRD8A/A 8-12 2-9-81
-------
CO
i—>
OO
Figure 8-3. Map of the United States indicating major areas of sulfur-deficient soils.
Source: The Sulfur Institute (1979).
-------
such as with sulfur-deficient soils, the atmosphere can be an important source of sulfur for
plant requirements; however, S02 monitoring data were not presented in their study report.
Similar stimulatory responses of Oryzopsis hymenoides (a desert grass) were noted follow-
ing exposure to S02 concentrations of 0.03 and 0.06 ppm continuously for 6 weeks (Ferenbaugh,
1978). The noted increases in productivity expressed as mg dry wt/plant were not statistically
different from the control plants, but the trends were stable for these doses. Exposure to
0.13 and 0.25 ppm SO- under identical conditions resulted in foliar symptoms and decreased pro-
ductivity.
The interpretation of studies that have demonstrated beneficial effects must be evaluated
in light of their single influence to one crop (as has been done for studies demonstrating
detrimental effects). There is little doubt that direct application of S as a nutrient to
certain crops grown under borderline or S-deficient conditions, may result in increased produc-
tivity for that crop. However, long-term natural ecosystem studies showing similar positive
effects for the entire ecosystem have not been accomplished. Since these agronomic and natural
ecosystems are often physically proximal to one another, further research on the potential
influence of S compounds to each singly and collectively is greatly warranted.
Numerous conditions may actually determine if S delivered to the plant as SO- will termi-
nate as a nutrient or pathogenic agent. Those factors discussed in relation to Figure 8-2 will
significantly influence the final disposition. Growth rates of leaf tissues, available nutri-
ent supplies, and environmental factors affecting stomatal opening and closing should all be
considered as influencing the rate of S accumulation in plant tissues (Bell and Clough, 1973).
The addition of nutrient forms of nitrogen being delivered to the plant along with S0? in
plumes has not been investigated.
Cowling and Lockyer (1976) demonstrated that S 23 ryegrass, when grown under S deficiency
and at low nitrogen, did not respond to 0.02 ppm SO- for 85 days; however, plants grown at high
N under the same exposure conditions responded with a 227 percent yield increase and S-defici-
ency symptoms were alleviated. Data presented by Cowling and Jones (1971) previously showed
that at high levels of N, and inadequate levels of S, nitrate-nitrogen accumulated in ryegrass,
thus indicating protein synthesis was inhibited. A review of crop response to sulfur has been
published by The Sulphur Institute (1971).
8.2.9 Foliar Versus Whole Plant Responses
The presence of acute or chronic foliar injury is not necessarily associated with growth
or yield effects. Furthermore, when present, the degree of foliar injury may not always be a
reliable indicator of subsequent growth or yield effects.
A prediction of two-thirds of one percent crop loss for each percent of leaf area
destroyed was generated from 30-minute exposures of soybeans (King) in the field to S02 con-
centrations ranging from approximately 0.5 to 6.0 ppm (Davis, 1972b). The exposures were
XRD8A/A 8-14 2-9-81
-------
conducted to generate varying amounts of injury. Usually, no injury resulted at concentra-
tions below 0.5 ppm, and all symptomatic tissue (necrosis, chlorosis, bronzing) was estimated.
Therefore, all leaves with any degree of symptoms were considered as non-functional in plant
productivity as completely destroyed. Soybean (Dare) plants exposed to 0.10 ppm S0? 6 hr/
day for 133 days in closed field chambers failed to exhibit significant yield reductions,
even in the presence of foliar injury (Heagle et al., 1974).
Yield effects in the absence of foliar injury have recently been reported for soybean
in field fumigations (Sprugel et al., 1980) and chamber exposures (Reinert and Weber, 1980).
Sprugel et al. (1980), utilizing the Zonal Air Pollution System, reported significant yield
reductions in soybeans (Wells) exposed to mean S02 concentrations of 0.09, 0.10, 0.19, 0.25,
or 0.36 ppm for an average 4.2 hr/day intermittently for 18 days during July and August
1978. Visual leaf injury was noted only at the highest S02 concentration. It must be noted
that the characteristics of the ZAPS allow for significant variation in the pollutant con-
centrations (although this does relate to ambient conditions near a source). For example,
in 1977, within experiments reporting a mean S02 concentration of 0.30 ppm, the actual
concentration ranged from 0.00 to 1.20 ppm S02 (Miller et al, 1980). The greenhouse chamber
study by Reinert and Weber (1980) was an 03+S02 interaction study with Dare soybeans
exposed 4 hr/day, 3 times per week, for 11 weeks. They reported significant growth reduc-
tions in the absence of visible injury for 0.25 ppm S0? when the treatment sums of squares
were partitioned.
Similar ambiguities are found in some of the literature dealing with grasses. In a
preliminary study, S 23 ryegrass exhibited significantly reduced growth when exposed to
0.12 ppm S02 for 9 weeks or 0.067 ppm S02 for 26 weeks (Bell and Clough, 1973). The only
foliar injury noted was a slight chlorosis and an enhanced rate of leaf senescence.
Ashenden (1978) noted similar significant growth reductions for cocksfoot when exposed to
0.11 ppm SOp for 4 weeks. Reductions ranged from 32 to 52 percent for various parameters
while foliar necrosis was only 5 percent. On the other hand, exposure of S 23 ryegrass
to 0.02 or 0.14 ppm S0? in two successive growth periods of 29 and 22 days resulted in
foliar injury at the high concentration, but no yield effects at either concentration
(Cowling and Koziol, 1978). Net photosynthesis and dark respiration were also not sig-
nificantly affected.
Different plant species differ in tolerance to S02 injury. Leaf injury and radial
growth were evaluated on Douglas fir and ponderosa pine growing in nursery plots exposed
to various doese of SOp in controlled fumigations (Katz and McCallum, 1952). Slightly
injured ponderosa pine (10 percent foliar symptoms) exhibited no significant deviations
in growth while slightly injured Douglas fir (10 percent foliar symptoms) showed a defi-
nite growth retardation when compared with control plants. The growth retardations were
evident for 3 years after S0? exposure, followed by substantial recovery.
XRD8A/A 8-15 2-9-81
-------
Increasing SOp concentrations have been negatively correlated with annual ring width
in Norway spruce (Keller, 1980). Exposures to 0.05, 0.10, and 0.20 ppm S02 were continuous
for 10 weeks in the spring. Some injury was noted at 0.10 and 0.20 ppm, but a distinct
decline of wood production was also found in cases where no visible injury occurred. When
dormant seedlings of beech were exposed to the same SO- concentrations (0.05, 0.10, and
0.20 ppm) for about 16 weeks, there were adverse effects to the terminal buds (Keller, 1978).
There was an increase in the number of terminal buds which failed to "break" in the spring
for the 0.10 and 0.20 treatments.
The literature concerning S0»-induced growth and yield effects and correlations with
visible foliar injury is ambiguous. No studies consider all the potential variables that
can affect plant response. This is a virtual impossibility for a single study and is espe-
cially true for field studies (which are most relevant) where many environmental variables
cannot be controlled. Also, many of the studies that have demonstrated negative effects of
S0? at low concentrations have utilized sensitive cultivars of plant species (which may or
may not be representative of plant populations as a whole) and maintained exposure condi-
tions conducive to maximum plant sensitivity. However, from the data available, we can con-
clude that growth and yield effects are not necessarily related to foliar injury. Depending
upon the plant affected, the environmental conditions, and the pollutant exposure conditions,
one may observe yield effects without injury, injury without yield effects, or more direct
correlations between injury and yield.
8.3 DOSE-RESPONSE RELATIONSHIPS - S0£
The primary focus of dose-response studies should be to develop useful generalizations
of the relationship between meaningful parameters of plant response and measurable indices
of exposure dose. This section will examine this relationship both from the perspective of
deficiencies in our present knowledge of dose-response relations and from the perspective
deficiencies in our present knowledge of dose-response relationships and from the perspective
of what generalizations are presently possible with existing data sets.
The dose of S02 to which vegetation may be exposed is conventionally designated as the
product of concentration in the plant's environment times the duration of exposure. Response
may be characterized by a measurable change in any parameter such as biochemical pathways,
gas exchange rates, photosynthetic rates, physiological functions, degree of visibly recog-
nizable leaf injury, or subsequent growth and yields. Plant responses may be beneficial
(see Section 8.2.8) or detrimental. They may involve the expression of growth and yield
effects without foliar symptoms (see Section 8.2.9) or lead to overt symptoms that seldom
become more serious than those associated with acute injury (see Section 8.2.5) or chronic
injury (Section 8.2.6).
In interpreting dose-response studies wherein a measured plant response is correlated
with exposure dose, it is important to realize that the relationship between exposure dose
and the amount of pollutant entering the plant may be influenced significantly by environmental
XRD8A/A 8-16 2-9-81
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*
factors which control rates of pollutant flux into plant leaves and by plant factors which
determine the metabolic fate of the pollutant within leaf tissues (see Figure 8-2).
The role of short-term fluctuations in S02 may be particularly significant where
impacts of point sources are under study (Mclaughlin et al., 1976). Here concentrations
may fluctuate widely during exposure and damage to vegetation may be most closely associ-
ated with short-term averages (1 hr) or even peak concentrations. Laboratory experiments
by Zahn (1961, 1970) have demonstrated the greater relative toxicity of short-term expo-
sures at high concentrations of S02 than longer-term exposures with the same total expo-
sure dose. More recently, Mclaughlin et al. (1979) studied the effects of varying the
peak to mean S02 concentration ratio on kidney beans in short-term (3 hr) exposures to S0?.
Working at or below the present National Secondary Air Quality Standard (NSAQS) they
found that increasing the peak:mean ratio from 1.0 (steady state exposure at 0.50 ppm for
3 hours) to 2.0 (3 hr exposure with peak = 1.0 ppm) did not alter post fumigation photo-
synthetic depression. Further increasing the ratio to 6.0 (1 hour exposure with peak =
2.0 ppm), however, tripled the post fumigation photosynthetic depression. Total dose
delivered in the three exposures was 1.5, 1.8, and 1.1 ppm, respectively. Clearly the
quantity of S02 to which the plants are exposed may have a very different effective poten-
tial as the kinetics of the exposure are changed.
Another important aspect of exposure dose is the frequency and duration of periods of
low SOp stress. Zahn (1970) emphasized that periods of low S0? concentration may be criti-
cal to the recovery potential of plant systems following exposure to elevated levels of
S0?. Thus, continuous exposure systems probably over-estimate the toxicity of the delivery
dose in many cases because physiological recovery is not permitted. Such recovery would be
expected under most exposure regimes in the field where fluctuating synoptic or local
meteorological conditions strongly influence exposure patterns.
Equally critical to definition of the biological significant features of exposure
regimes is the definition of significant parameters of plant response. Sections 8.2.1
through 8.2.9 have emphasized the many types of responses which may be elicited by exposure
to S02< In interpreting or predicting plant response to S02, it is important to keep in
perspective the fact that plant growth and development represents an integration of cellu-
lar and biochemical processes just as community behavior is an integration of the perfor-
mance of component species. The internal allocation of resources (carbon, water, and
nutrients) to growth is an integrative and in many cases resilient process which plays a
major role in determining how both individual plants and plant communities respond to
environmental stress (Mclaughlin and Shriner, 1980); (see also Section 8.10). The fact
that a response is measured following exposure to a given dose of S02 may be of interest
in understanding the mechanism of action and in identifying the biologically significant
features of dose; however, it does not necessarily mean that an effect will be measured
at a subsequent higher level of organization. Responses at higher levels of organization,
however, must be viewed within the perspective of the increasingly complex biotic and
XRD8A/A 8-17 2-9-81
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abiotic factors which control plant response (see Section 8.5) and influence our attempts
to move the focus of our studies from processes to plants and from plants to communities.
These factors similarly limit our capacity to easily extend experimental protocols beyond
the confines of our carefully controlled laboratory studies to more natural, and more
variable, field situations.
Several attempts have been made to characterize dose-response relationships in a
mathematical sense using monitored concentrations, exposure times, and injury thresholds
as modified by physical and biotic factors expressed as constants (O'Gara, 1922; Thomas
and Hill, 1935; Zahn, 1963a,b,). However, their consistency and usefulness are limited
due to numerous physical and biotic factors which must be considered in evaluating dose-
response data. Changes in exposure conditions, differences in exposure methodology and
efficiency of monitoring equipment, ad consistency of measurements within a study and
between studies on the same plant will directly influence results (Figure 8-4).
According to Larsen and Heck (1976), (1) a constant percentage of leaf surface will
be injured by an air pollutant concentration that is inversely proportional to exposure
duration raised to an exponent, and (2) for a given exposure duration, the percent leaf
injury as a function of pollutant concentration tends to fit a log-normal frequency
distribution:
c = mg 1 hrsgZtP
where
c = the SCL concentration in ppm
m , . = the geometric mean concentration, the concentration that injures
^ 50 percent of the leaf surface after a 1-hr exposure
s = the standard geometric deviation, ratio of the concentration that
" injures 50 percent of the leaf surface to the concentration that
injures (16 percent)
z = the number of standard deviations the percent leaf injury is from
the median (50 percent)
t = the exposure duration in hr
p = the slope of the percent leaf injury isopleths on logarithmic
graph paper.
Parameters for the above equation have been determined from leaf injury observations
on four tree species. The concentration expected to cause threshold (1) leaf injury in
Norway maple exposed to S02 for 8 hr is determined by substituting its parameters into the
above equation:
c = 6.76(1.29)"2'3V°-38
c = 1.7 ppm
XRD8A/A 8-18 2-9-81
-------
NUMBER OF.
EXPOSURES
CLIMATIC FACTORS
EDAPHIC FACTORS
BIOTIC FACTORS
POLLUTANT
CONCENTRATION
I
DOSE-4-
PLANT RECEPTOR
MECHANISM OF ACTION
DURATION OF
"EACH EXPOSURE
• GENETIC MAKEUP
STAGE OF PLANT
•DEVELOPMENT
EFFECTS
ACUTE
\
CHRONIC
SUBTLE
Figure 8-4. Conceptual model of the factors involved in air pollution effects (dose-response) on vegetation.
Source: Heck and Brandt, 1977.
8-19
-------
Similar calculations indicate that threshold injury is expected after an 8-hr exposure to
2.2 ppm for ginkgo and 3.7 ppm for pin oak. The injury data are too far from threshold to
calculate an accurate threshold for the fourth tree species, Chinese elm.
As discussed in preceding sections, plant response to SCL may occur at many levels. For
purposes of air quality control, however, two parameters, visible damage to foliage, and plant
productivity, provide the most functional basis for evaluating response. Both can be quanti-
fied as a "cost" to economic or ecological performance of many plant species.
Dose-response relationships involving visible injury may be expressed in terms of the
level of injury (percent leaf area destroyed) produced for a single species, or as the upper
and lower limits of sensitivity of a group of species. The latter approach is presented here
because it provides data more applicable to responses of plant populations and because of the
difficulties in quantifying a dependable relationship between degree of visible injury and
growth responses (see Section 8.2.7). Data on generalized concentrations at which sensitive,
intermediate, and resistant species may be injured by SCL were presented earlier in Table 8-2.
In Figure 8-2 (Section 8.2.7) the data of Dreisinger and McGovern (1970) were graphed to show
upper and lower concentration limits of susceptibility of 31 species of herbs, trees, and
shrubs to visible foliar injury. Plotted as a function of S02 concentration and exposure
time, these data demonstrate a number of important points: first, the most sensitive plants
at each concentration were injured at S02 levels 6 to 7 times lower than the most resistant
plants. Secondly, the dose (ppm x h) required to cause injury was 30 to 60 percent lower for
1-hr than 8-hr exposures, emphasizing the importance of exposure kinetics. Finally, exposure
to SCL at 0.5 ppm for 3 hr represents a rather close estimate of the injury threshold for
"average" plants.
A second approach to defining dose-response relations focuses on the numbers of indivi-
duals in a plant population which may be injured as function of exposure concentration. As an
example of a "worst-case" situation of S02 exposures and vegetation effects near a rural coal-
fired power plant, TVA's Widows Creek Plant in northern Alabama provides some interesting data
(Mclaughlin and Lee, 1974). During the period 1970-1973 (before partial sulfur scrubbing and
stack elevation by TVA improved surrounding air quality), surveys of vegetation in the
vicinity of this plant documented foliar injury to 84 plant species growing in the vicinity of
continuous SO- monitoring stations. Tabulation of these data (Figure 8-5) as a function of
exposure concentration provides an index of probability of injury of species in a plant com-
munity as a function of S02 concentration. If one arbitrarily sets a limit of acceptable risk
as 10 percent of the plant population (here 8 of 84 species), then peak, 1-hr, and 3-hr con-
centrations of (1.00, < 0.50, and < 0.30 ppm should be avoided. At the present National
Secondary Ambient Air Quality Standard (3 hr at 0.5 ppm), approximately 30 percent of the
species present sustained foliar injury. This form of data representation allows interconr
parisons between short-term concentrations and risk of foliar injury to be readily made.
XRD8A/A 8-20 2-9-81
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100
SPECIES
AFFECTED SO CONCENTRATION
SO2 CONCENTRATION, ppm
Figure 8-5. Percentage of plant species visibly injured as a function of peak, 1h and 3h SC>2 concentra-
tions.
Source: McLaughlin and Lee, 1974.
-------
Data on S02 effects of plant growth and yield in most cases provide the most relevant
basis for studing dose-response relationships. As a whole-plant measurement, plant produc-
tivity is an integrative parameter which considers the net effect of multiple-factors over
time. Productivity data are presently available for a wide range of species under a broad
range of experimental conditions. Because results would not be expected to be closely
comparable across these sometimes divergent experimental techniques, data have been tabu-
lated separately for controlled field exposures (Table 8-3), laboratory studies with agro-
nomic and horticultural crops (Table 8-4) and tree species (Tables 8-5 and 8-6), and a
variety of studies with native plants (Table 8-7).
Relatively few crops of economic importance have been studied under field conditions
utilizing various field exposure systems. Of the twelve "studies" reviewed in Table 8-3,
seven reported induced yield effects following SCL exposures at varying doses. The lowest
dose exposure to induce a yield effect was 0.09 ppm S02 for 4.2-hr average fumigation
period on 18 days scattered from July 19 through August 27 of the soybean growing season
(Sprugel et al., 1980). Five studies indicated no effect following various exposure
regimes, and one study (Neely and Wilhour, 1981) reported increased yields (27 percent and
8 percent) of winter wheat cs. Yamhill following exposure dose of 0.03 and 0.06 ppm S02 for
24 hr/day for 30 days, respectively.
Table 8-4 presents a summary of studies that have investigated effects of SO- on
agronomic and horticultural crops grown and fumigated within artificial exposure chamber
or growth chamber systems. It is difficult to determine the significance of the results
of such studies in relation to actual similar fumigations under field conditions. Addi-
tionally, with the exception of a relatively few studies, doses used for exposure treat-
ments would be considered as in excess of expected doses for ambient field exposures.
As indicated in Table 8-4, acute foliar effects have not been reported in long-term
studies using less than 0.15 ppm SOy for 24 hr/day for 7 days (Mandl et al., 1975).
A few major investigations of the effects of SO- on tree species growing under
natural conditions have been reported (Linzon, 1971, 1980; Dreisinger, 1965;
Dreisinger and McGovern, 1970; Materna et al., 1969; and Vins and Mrkva, 1973). Table
8-5 illustrates the degree of injury to eastern white pines (Pinus strobus) over a 10-
year period (1953-1963) in the sulfur fume effects area near Sudbury, Ontario, CAN.
(Linzon, 1971). Linzon (1971, 1978) has indicated that a pollution (S0?) gradient
existed within the designated study area and effects correlated well with this gradient.
Chronic effects on forest growth were prominent where SO- air concentrations averaged
annually 0.017 ppm S02, and such effects were not reported in areas receiving 0.008 ppm
S02 annually. Although monitoring of S02 was conducted in these studies, other air
pollutants nor their potential effects were evaluated.
The studies of Vins and Mrkva (1973) and Materna et al. (1969), although reporting
foliar and growth increment losses to forest trees as being due to S0?, were done in areas
XRD8A/A 8-22 2-9-81
-------
TABLE 8-3. SUMMARY OF STUOUS REPORTING RESULTS OF SOj EXPOSURES USING EXPOSURE SYSTEMS AND/OR CHAMBERS OVER PLANTS UNDER FIELD CONDITIONS
00
I
ro
CO
Ffifia- '
bone.
PC"
0.04
0.02
o.os
0.10
GOO*.
0.022;
0.03-
0.10
0.03-
0.06
0.10
O.IS
0.20
0.01
0.06
0.10
O.IS
0.20
0.06
0.09
0.10
0.19
0.25
0.36
0.10
0.12
0.30
0.79
EMfwsum exposure
TlM Condition
3 tir for S ««p. F/CC
graving tMton
Growing season F-ZAP
leans
0.038; 0.068, respectively
72 hr/w* for
graving season F/CC
24 hr/day F/CC
for 30 with > SO, cone. ; digesti-
bility dry utter wal < by 2 years of
treatment; crude protein content In
u. wheat < significantly.
No effect on yield
27X yield grain vt.
« yield grain wt.
ITS yield grain wt.
17X yield grain wt.
70S yield grain wt.
No e feet
22X yield grain wt.
36X yield grain wt.
27X yield grain wt.
56X yield grain wt.
28S < foliage stubble, 4SX < root dry
wt. ; 21X < total protein, aitno acids.
{unstructured carbohydrates, and syep-
tlotlcally fixed nitrogen content.
6.4X yield reduction
S.2% yield reduction
12. 2X yield reduction
19. 2X yield reduction
15.9 X yield reduction
No significant effect on foliar Injury,
fresh wt. seeds/plant or wt. of seeds/
plant 92nd day defoliation was 12* >
control ; 135th day seed wt. only IX <
control .
12. » yield reduction
20. SX yield reduction
45. » yield reduction
Caveat*
Potted plants set
on soil surface.
grown hydroplnlcally
Sd of fualgatlon
conj. ranged 44-4SX
of x.
Sg^standerd geoMtrlc
deviation
Sd of fualgatlon
cone. r|nged 41-
64X of «
Sg^standard geometric
deviation
Reference
Stj et al..
1974
Preston and
Lewis, 1978
Hllnour, 1980
Neely and
Wilhour, 1981
Neely et al.,
1977
Sprugel et al.,
1980 and
Miller et al...
1980
Heagle et *!..
1974
Sprugel et al.,
1980 and
Miller et el..
1980
-------
TABLE 8-3 (continued)
Co
ro
Cone.*
PP«
0.15
0.40
0.25
0.40
0.80
1.20
•0.45
0.80-
2.00
Exposure3 Exposure
Ti«e Condition
72 hr/wk for F/CC
growing season
Once every week F/CC
(3hr) to once
In 5 wks (3hr)
3 hr for 7 exp. F/CC
growing season
4 hr 20 Bin F-ZAP
Effects onc
Plant Foliage Yield
Barley X
Durum wheat X
Spring wheat
Alfalfa
Barley
Durum wheat
Spring wheat
Wheat
Soybean X
X
X
X
Species effectd Caveat6
44% < yield In Barley
42% < yield In Durum wheat
No effect on Spring wheat
No effect on yield
No accumulative effect on yield, no
effect on avg. head length or * grains/head
4.5% < yield at 1.4 ppm Pollutant avg.;
11* < yield at 1.7 ppm no est. on range
15% < yield at 2.0 ppm of exposure doses
Epidermal and mesophyll cell death; the
# of dead mesophyll cells highly corre-
lated with > S02
Reference
Wilhour, 1981
Wilhour, 1981
Sij et al.,
1974
Miller et al.,
1979
Irving et al. ,
1979
Table arranged by > SO- concentration as first order and exposure time as second order divisions. Doses within a single study that did not induce
.specifically different effects are listed along with the lowest SO, concentration that induced said effect.
F/CC = field, closed chambers; F/OT = field, open top chambers, F-ZAP = field, zonal air pollution system.
J( indicates study found bo11age and/or yield effects.
"Host prominent or significant effect reported.
Caveats for consideration about proper study design and interpretations.
-------
TABLE 8-4. SUMMARY OF STUDIES REPORTING RESULTS OF SO, EXPOSURE UNDER LABORATORY CONDITIONS
FOR AGRONOMIC AND HORTICULTURAL CROPS
Cone.
ppm
0.035
0.175
0.05
0.05
0.10
0.25
0.05
0.20
0.06-
0.08
0.07-
0.53
0.10
0.11
0. 125-
0.15
Exposure
Time
8 hours
5h/d; 5d/wk
for 4 wk
4 hour
8hr/day
5 days/wk
for 18 days
103.5hr/wk
for 20 wks
24hr/day
9-20 days
18 days
24hr/day
4 wks
1-3 hours
18 days
Exposure
Condition Plant
EC/SD Broadbean
EC/SO Alfalfa
Tobacco, Bel W3
Tobacco, Burley 21
EC/SD Oats
Radish
Soybean
Tobacco
EC/SD Soybean
GC Cocksfoot
Meadowgrass
GC Tobacco
Sunflower
Corn
GC Pea
GC Italian ryegrass
EC/SD Oats
Radish
Sweet pea
Swiss chard
GC Pea
Effects
Foliage
X
X
X
X
X
X
X
X
X
X
X
d
Yield Species effect0
Depressed net photosynthesis
26% < in foliage dry wt. at final harvest
493! < of root dry wt. at final harvest
22* < of leaf dry wt. at final harvest
No effect
No foliar injury
No effect on top fresh or dry wt. , root
fresh or dry wt. ; plant height, shoot/root
fresh or dry wt. ratio
40% < total dry wt.
28% < total dry wt.
Increased dry wt. yield up to 0.53 ppm
(44% > control)
Greatest dry wt. yield at 0.35 ppm (44%
> control); 27% increased yield over con-
trol at 0.53 ppm
Greatest dry wt. yield at 0.17 ppm (24%
> control) 7% increased yield over con-
trol at 0.53 ppm
3% < fresh wt. shoot
5% < dry wt. shoot
4% < to^al nitrogen
30% < H (buffer capacity)
10% > glutamate dehydrogenase activity
110% > inorganic sulfur content
No difference from control at low wind;
17%-40% < total dry wt. high wind
No foliar injury at concentrations < 0.50 ppm
2% maximum foliar injury experienced at all
doses.
3% < fresh wt. of shoot
8% < dry wt. of shoot
2% < total nitrogen
35% < H (buffer capacity)
32% > glutamate dehydrogenase activity
140% > inorganic sulfur content
Caveat Reference
Black and
Unsworth
, 1979
Tingey and
Reinert,
Sensitive plant
Tingey et
1971a
Tingey et
1973b
Ashenden,
1975
al. ,
al. ,
1979
No monitoring Faller, 1970
methods pre-
sented; Low S
in soil medium
Jager and
1977
Klein,
Ashenden and
Mansfield, 1977
Bennett et
1975
Water Culture Jager and
1977
al. ,
Klein,
-------
TABLE 8-4 (continued)
CO
Cone. Exposure9
ppm Ti»e
0.15- 24 h/day
0.30 7 days
0.17 2 hour
0.20- 2 hour
0.30
0.20 30, 78, 100
hours
0.20 15 days
0.20 Continuous
to maturation
0.25 4 hours
0.25 18 days
0.25 4 hr 3 times/wk
11 weeks
Exposure
Condition Plant
EC/SD Barley
Bean
Corn
EC/SD Broadbean
GC Alfalfa
Barley
EC/SO Wheat
GC Tomato
EC/SD Kidney bean
EC/SD Broccol i
Tobacco, Bel B
Alfalfa
Onion
Soybean
Lima bean
Bromegrass
Cabbage
Radish
Spinach
Tomato
GC Pea
EC/SD Soybean
Effects onc
Foliage Yield Species effect Caveat
X Severe foliar injury
X No injury
X Severe foliar Injury
< photosynthetic rate, < stomatal re-
sistance if RH > 40%, > stomatal resis-
tance if RH < 40%
Threshold dose for inhibition of photo-
synthesis, reversible effect
X X Trend of increased dry wt. for 19 or 21 Trend, not
exposures; small amount foliar injury significant
from control
X Treshold dose for initial symptom of tis-
sue death, < or change vitamin B,, B,,
and nicotinic acid content
X < 15% of total yield; no change in protein
content
X 6% leaf injury
X 1% leaf injury
No effects
No effects
No effects
No effects
No effects
No effects
No effects
No effects
No effects
X 32% < fresh wt. of shoot Water Culture
26% < dry wt. of shoot
24% < total nitrogen
42% < H (buffer capacity)
80% > glutamate dehydrogenas activity
150% > inorganic sulfur content
X No foliar injury; significant < in plant
ht. at 5,7,9,11 wks; significant shoot
dry wt. < at 7,11 wks; significant root
dry wt. < at 9,11 wks; significant total
dry wt. < at 11 wks
Reference
Handl et al. ,
1975
Black and
Unsworth, 1979
Bennett and
Hill, 1973
Laurence, 1979
Unzicker et al. ,
1975
BerigaH et al. ,
1974
Tingey et al. ,
1973a
Jager and Klein,
1977
Re inert and
Weber, 1980
-------
TABLE 8-4 (continued)
oo
PO
Cone.
w»
it. JO
0.35
0.40
O.SO
0.60
0.40
1.00
1.00
1.00
1.00
1.00
1.00
1.50
l.SO
2.00
Exposure
T*M
Shr/day
taay/wk
U days
2C days
1 hour
4 hours
6 hours
2 hours
3 hour*
2 hours
4 hours
6 hr/day
for 3 days
1.5 hour
3 hours
.75-3 hours
3 hours
2 hours
Expotur*
Condition
£C/4S
EC/SO
EC/ SO
EC/SO
GC
CC
EC
EC/SO
EC/SO
EC/SO
EC/SO
EC/SO
EC
Plant
Barley
(•an
Sunflower
BarUy
Bon
Sunflotttr
Alfalfa
Tou to
Apples
BarUy
Polnsettla Bcv's
Begonia
Pttunii
Coleus
Snapdragon
Broccoli
BroMgrass
Cabbage
Liu bean
Radish
Spinach
To« to
Strawberry
Soybtan
Soybean
Alfalfa
Begonia
Petunia
Coleus
Snapdragon
Effects
Foliage
X
X
X
X
X
X
X
X
X
K
X
X
X
X
X
X
X
X
«
X
onc
Yield
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
Spades effect11 Caveat
11X foliar Injury; 38X < dry wt. shoot Hontorlng systeei
<1X foliar Injury: 2SX < dry wt. shoot explained In
SX foliar Injury; 4 » < dry wt. shoot unavailable
publication
2Ut foliar Injury; 2bX < dry wt. shoot
ZX foliar Injury; 15X < dry wt. shoot
16X foliar Injury; 2« < dry wt. shoot
8X < In apparent photosynthesis
> accumulation total and soluble S content
2X leaf Injury
Threshold dose for foliar necrosis; 30-60X
< net photosynthesis
Foliar Injury 1 cultlvar
No effect
30X < flower *'s; 19X < shoot wt.
27X < flower ft; 19X < shoot wt.
14X < flower ft; 16X < shoot wt.
38X leaf Injury
(5X leaf Injury
70X leaf Injury
25X leaf Injury
46X leaf Injury
49X leaf Injury
33X leaf Injury
No effect on growth and development; necro-
t1c lesions, lower leaf surface
9X < shoot fresh wt. , 4X leaf Injury Short-tens
growth response
21-29X < shoot fresh wt. only
24-94* < shoot fresh wt. , 63-93X foliar Short-tens
Injury response
only
Leaf necrosis et 31S pp> CO. was 2.5x that In-
duced under 64S ppn O>2
HX < flower *'s; 22X < shoot wt.
32X < flower f's; 24X < shoot wt.
30X < flower »'s; 20X < shoot wt.
15X < flower »'s; 1SX < shoot wt.
Reference
harkowskl et •!..
1975
White et al.. 1974
Bennett and Hill,
1973
Kender and
Splerlngs, 197S
Bennett and Hill.
1973a
Heggestad et al..
1973
Adedlpe tt a).,
1972
Tlngey et al. ,
1973a
Rajput et al..
1977
Heagle and John-
ston. 1979
Heagle and John-
ston, 1979
Hou et al . , 1977
Adedpte et al.,
1972
-------
TABLE 8-4 (continued)
Conc.a Exposure8
PfMI TIM
2.00 3 hours
2.50 6 hours
3.00 1 hour
2 hour
3 hour
4.00 2 hour
r\> 0.40 30, 78. 100
00 hours
0.50 1.5 hour
0.50 100 hours
0.60 6 hours
0.60 30, 78, 100
hours
0.75 3 hour
0.80 2 hour
Exposure
Condition Plant
GC
EC/SO
GC
GC
GC
EC
EC/SO
EC/SO
EC/SO
EC/SO
EC/SO
EC/SO
GC
Poinsettia 8cv's
Apples
Poinsettia Scv's
Poinsettia Scv's
Poinsettia Scv's
Begonia
Petunia
Coleus
Snapdragon
Marigolds
Celosia, Salvia
Impatiens
Wheat 7cv's
Soybean
Corn
Apples
Wheat 7cv's
Alfalfa
Alfalfa
Effects onc
Foliage Yield
X
X X
X
X
X
X X
X X
X
X
X
X
X
X
X X
X
X X
X
X
Species effect*1 Caveat
Foliar injury, 2 cultivars
> foliar injury; 62% > leaf abscission;
19% < shoot growth
Foliar injury, 5 cultivars
Foliar injury, 7 cultivars
Foliar injury, 8 cultivars
27% < flower #'s 33% < shoot wt. ; severe necrosis
42% < flower #'s 32% < shoot wt. ; slight injury
30% < flower #'s 21% < shoot wt. ; no foliar injury
20% < flower *'s 19% < shoot wt. ; slight injury
Slight injury
Slight injury
Slight injury
No effect on yield, small amount foliar injury
7% < in short fresh wt. , trace foliar injury Short-term
growth
response only
Minimal foliar injury, no effect on dry mass
7.3% > foliage injury; 5% > leaf abscission
Trend of decreased dry wt. in 17 or the 21 Trend not sig-
exposureSj small amount of foliar injury nif leant from
control
No injury developed
Threshold dose for foliar necrosis; 25-50X
< in net photosynthesis
Reference
Heggestad et al. ,
1973
Kender and
Spierings, 1975
Heggestad et
al., 1973
Adedipe et al. ,
1972
Laurence, 1979
Heagle and John-
ston, 1979
Laurence, 1979
Kender and
Spierings, 1975
Laurence, 1979
Hou et al . , 1977
Bennett and Hill,
1973
'Table arranged by > SO. concentration as first order and exposure time as second order division. Doses within a single study that did not induce
specifically different effects are listed along with the lowest SO. concentration that induced said effect.
GC - Growth chambers, EC = Exposure chambers, EC/SO = Exposure chamBer, special design.
*ix indicates study found foliage and/or yield effects.
Most proninent or significant effect reported.
'Caveats for consideration about proper study design and interpretation.
-------
TABLE 8-5. THE DEGREE OF INJURY OF EASTERN WHITE PINE OBSERVED AT VARIOUS DISTANCES FROM THE SUDBURY SMELTERS FOR 1953-63
00
Trees with
Current
Year1 s
Trees with 1- Year-
Old (1962) Foliage
Injured
Forest Sampling Foliage
Station Injured in
(Distance and
Direction fro*
Sudbury)
West Bay
(19 Biles NE)
Portage Bay
(25 Biles NE)
Grassy to E*erald Lake
(40-43 Biles NE)
Lake Matinenda
(93 ailes W)
Correlation
Coefficient (r)
August
1963
(X)
2.0
1.1
0.4
0.6
0.96*
June
1963
(X)
38.0
21.5
2.5
0.3
0.96*
August
1963
(X)
77.9
55.6
16.7
2.1
0.93**
Trees with 2-Year
Old Foliage
Injured
In June
1963
(X)
96.0
77.0
37.5
10.1
0.90**
Lacking
in
August
1963
(X)
20.6
15.2
9.1
3.9
0.94**
Net Annual
Average
Gain or
Loss in
Total
Volume,
1953-1963
(X)
-1.3
-0.5
+1.8
+2.1
0.90**
Annual
Average
Mortality
1953-1963
(X)
2.6
2.5
1.4
0.5
0.81
Degree of SO,
Damage
Acute and chronic
Injury
Mostly chronic
and little
acute injury
Average SO.
Concentration
for Total
Measurement
Period 1954-
1963°
(ppm)
0.045
0.017
Very little chronic 0.008
Injury
Control: no SO,
Injury c
0.001C
(Sturgeon Falls)
rv> 'Linzon (1971)
10 °Dreisinger (19
Data for 5-Bonth growing season-1971
*p < 0.05
**p < 0.10
-------
TABLE 8-6. SUMMARY OF STUDIES REPORTING RESULTS OF S02 EXPOSURE UNDER LABORATORY CONDITIONS FOR VARIOUS TREE SPECIES
Conc.a
ppm
0.025
0.05-
0.15
0.05
0.05
0.10
0.20
0.18-
0.20
0.25
0.25
0.35
0.45
0.45
0.50
0.50
0.50
0.50
Exposure3
Time
6 hour
6 hour
16 weeks
Winter
10 weeks
24 hour
2 hour
2 hour
3 hour
6 hour
9 hr/day
for 8 wks
2 hour
3 hour
5 hour
24 hr/day
up to 30 day
Exposure
Condition Plant F<
EC/SD E. white pine
EC/SO E. white pine
EC/SD Beech
EC/SD Norway
spruce
EC/SO Jack pine
EC/SD E. white pine
Jack pine
Red pine
EC/SD Loblolly pine
Shortleaf pine
Slash pine
Virginia pine
EC/SD Trembling aspen
EC/SD E. white pine
EC Ponderosa pine
EC/SO E. white pine
Jack pine
Red pine
EC/SD Trembling aspen
GC Austrian pine
Ponderosa pine
Scotch pine
Balsam, Fraser fir
White fir
Blue, white spruce
Douglas fir
GC Chinese elm
Gingko
Norway maple
Pin oak
Effects onc
oliage Yield
X
X
X X
X X
X X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
Species effect*1 Caveat6
Threshold dose for needle damage; most
sensitive clones only
60% of tolerant clones developed foliar Sensitive clones
injury
Number of dead buds in Spring > at 0.10 ppm
and higher; 50% > kill at 0.20.
No foliar effect 25% < vol. growth (Avg. ) 2 cones only
Foliar injury 38% < vol. growth (Avg.)
Foliar injury 53% < vol. growth (Avg.)
Inhibit foliar lipid synthesis, inhibition
reversible; > dose = > recovery time
6.5% foliar injury Plants maintained
4.5% foliar injury in sensitive
0.5% foliar injury condition
All equally sensitive; most sensitive Plants maintained
period 8-10 weeks of age or older in sensitive
condition
2% foliar injury
All tolerant clones developed foliar injury
Severe needle tip chlorosis and necrosis
12% foliar injury
11% foliar injury
2% foliar injury
115! foliar injury
No injury
7 days to chlorosis
14 days to chlorosis
12 days to chlorosis
30 days to chlorosis
Reference
Houston, 1974
Houston, 1974
Keller, 1978
Keller, 1980
Malhotra and
Kahn, 1978
Berry, 1971
Berry, 1974
Karnosky, 1976
Houston, 1974
Evans and
Miller, 1975
Berry, 1972
Karnosky, 1976
Smith and
Davis, 1978
Temple, 1972
-------
Conc.a
ppm
0.50
0.50
0.65
1.00
1.00
1.0
2.00
2.00
2.00
2.00
2.00
TABLE
b Effects onc
Time Condition Plant Foliage yield
24 hr/day EC/SC Sugar maple
1/wk Black oak
White ash
24 hr/day EC/SO Sugar maple
Black oak
White ash
3 hour EC/SO Trembling aspen X
4 hour GC Austrian pine X
Ponderosa pine
Scotch pine
Balsam, fraser fir
White fir
Blue, white spruce
Douglas fir
8 hour EC American elm
1 hour EC Scots pine X
3 hour X
5 hour X
2 hour GC Austrian pine X
Ponderosa pine
Scotch pine
Balsam, fraser fir
While fir
Blue, white spruce
Douglas fir
6 hour GC American elm X X
6 hour GC American elm
6 hour GC Chinese elm X
6.5 hour EC Gingko
8-6 (continued)
Species effect Caveat8
48% < rate of photosynthesis, no symptoms
54% < rate of photosynthesis, no symptoms
20% < rate of photosynthesis, no symptoms
74% < rate of photosynthesis, no symptoms
33% < rate of photosynthesis, no symptoms
T% < rate of photosynthesis, no symptoms
23% foliar injury
Less than 4% foliar injury all species
Inhibition of stomatal closing
No injury, primary needles; slight injury,
secondary needles
14% maximum injury primary needles;
52% maximum injury secondary needles
37% maximum injury primary needles,
60% maximum injury secondary needles
No foliar injury on Douglas fir, firs,
spruce
Pine foliar injury threshold, necrotic tips
Induce severe foliar injury; defoliation in
older leaves; significant reduced expansion
of new leaves; # of emerging leaves and
root dry wt. reduced
No significant reduction in lipid content;
significant < in new leaf protein content;
significant < leaf, stem root carbohydrate
content
100% leaf necrosis
Water stressed plant > uptake of 50^
Reference
Carlson, 1979
Carlson, 1979
Karnosky, 1976
Smith and
Davis, 1978
No land and
Kozlowski, 1979
Smith and
Davis, 1977
Smith and
Davis, 1978
Constantinidou
and Kozlowski,
1979a
Constantinidou
and Kozlowski ,
1979b
Temple, 1972
Noland and
Kozlowski, 1979
-------
TABLE 8-6 (continued)
Conc.a
ppm
2.00
3.00
Exposure
Time
12 hour
6 hour
Exposure
Condition
GC
GC
Effects onc
Plant Foliage Yield
American elm
Gingko X
Norway maple
Species effect*1 Caveat*
Induced stomatal closing; S content > in
plants fumigated in light
50% leaf necrosis
Reference
Temple, 1972
Temple, 1972
Table arranged by SO. concentration as first order and exposure time as second order divisions. Doses within a single study that did not induce
specifically different effects are listed along with the lowest SO,, concentration that induced said effects.
L e.
GC = growth chambers, EC = exposure chambers, EC/SD = exposure chamber, special design.
X indicates study found foliar and/or yield effects.
Most prominent or significant effect reported.
Caveats for consideration about proper study design and interpretation.
-------
TABLE 8-7. DOSE-RESPONSE INFORMATION SUMMARIZED FROM LITERATURE PERTAINING TO NATIVE PLANTS
AS RELATED TO FOLIAR, YIELD AND SPECIFIC EFFECTS BY INCREASING S0? DOSE
Cone.3
ppm
0.024
0.02
0.14
0.03
0.06
0.067
0.08
0.11
0.11
0.11
(weekly
mean
0.067)
Exposure3 Exposure
Time Condition
85 days EC/SD
29 days and EC/SD
22 days
later
6 week EC/SD
26 week EC/SD
13 hr/day EC
for
28 days
4 week EC/SD
4 week EC/SC
103.5 hr/wk EC/SD
for 20 wk
Effects onc
Plant Foliage Yield Species effectd Caveat
523 Ryegrass X Plants at high nitrogen, SO. exp. allev.
S deficiency of plants not provided SO^
227% > in yield. No effect on plants
grown with adequate SO^
X Plants at low nitrogen, no effect of SO^
or S02
523 Ryegrass X No significant effects at 0.02 or 0.14 at
first harvest (29 days);
X at 0.14, significant reduction (16%) in
specific leaf area at second harvest
(22 days later). No significant effect
on dry wt. , tillers, dark respiration
or transpiration coef.
Indian X Non-significant increase in productivity (19%);
ricegrass Non-significant increase in productivity (21%);
significant decrease in chlorophyll content
(43%)
S23 ryegrass X Significant increase in number (88%) abd dry
wt. (78%) of dead leaves; significant de-
crease in number of tillers (41%); leaf
area (51%), dry wt. of stubble (55%), and
number (45%) and dry wt. (51%) of living
leaves; significant decrease in yield (52%)
Wild ryegrass X None
Foxtail grass X X Foliar injury as caused by heavy metals was
increased by SO, exposure; yield not sig-
nificantly affected
Cocksfoot X X 5% foliar necrosis; significant (30%) de-
crease in leaf area, dry wt. (45%), til-
lers, green leaves, and root/shoot ratios
Ryegrass X Significant (20%) decease in leaf area.
dry wt. (40%), root/shoot ratio at a
wind speed of 25m min (.93mph)
No effect at a wind speed of 10m min"
(,37mph)
Smooth- stalked X Significant decrease in leaf area (28%), all
meadowgrass dry wt. fractions (44%), leaves (37%) and
tillers (27%)
Preliminary
study mean
weekly cone.
wintertime
exposure
No technical
SO- monitor-
Reference
Cowling and
Lockyer, 1978
Cowling and
Koziol, 1978
Ferenbaugh, 1978
Bell and
Clough, 1973
Krause and
Kaiser, 1977
ing information
Wind tunnel
exposures
Wind tunnel
exposures
Ashenden, 1978
Ashenden and
Mansfield, 1977
Ashenden, 1978
-------
TABLE 8-7 (continued)
Conc.a
ppm
0.13
0.13
0.25
0.50
1.00
0.15
0.30
0.60
0.20
0.25
0.27
0. SO-
IL 00
11.00
0.71
1.0
Exposure9 Exposure" Effects onC
Time Condition Plant Foliage Yield Species effect
9 week EC/SO 523 ryegrass X Significant decrease in dry wt. (46%) and
number (34%) of living leaves, tillers
(42%), leaf area (44%), dry wt. of stubble
(47%); significant increase in number (46%)
and dry wt. (46%) of dead leaves; significant
decrease in yield
Wild ryegrass X None
6 week EC/SO Indian ryegrass X X Necrotic foliar lesions noted; non-signifi-
cant decrease in productivity (6%), signi-
ficant decrease in chlorophyll content (51%)
Necrotic foliar lesions noted; significant
decrease in productivity (35%) and chloro-
phyll content (6US)
Plants mostly dead after 4 weeks
Plants dead after 4 weeks
6 weeks Duckweed X Decrease in diameter of fronds, no dry wt.
effects
Decrease in diameter of fronds
Decrease in starch content, no dry wt. effects,
no irreversible damages up to 0.60 ppm SOj
2 hour GC Kentucky bluegrass X Visible foliar injury, high degree of
variation among 17 cultivators
5 week EC/SD Ryegrass X Significant decrease in yield (17%), no
effect on number of tillers
8 week Significant decrease in green wt. (38%),
total dry wt. (30%), no reduction in number
of tillers, senescence rate doubled
2 hour F/CC 87 desert species X Most plants required more than 2.00 ppm
S0_ to produce foliar injury
Sol to produce foliar injury
1 hour EC/SD Lily Significant pollen tube elongation,
2 hour inhibition at 1 and 2 hours
5 hour
6 hour EC/SD Eucalyptus X >40% foliar necrosis, 32 of 131 species
of Australian trees and shrubs were
rated as sensitive to acute (>1 ppm)
exposure to SO.
Caveat
Wind tunnel
exposures
Ambient air +
SO. exposure
system not
described
No SO, monitor-
ing information,
plants previously
exposed to SO-
Wind tunnel
exposure
Field plants
watered heavily
watered heavily
and exposed to
ambient air
before and after
fumigation
Pol len on agar,
relationship
those effects
have to ambient
conditions is
unknown
Reference
Bell and
Clough, 1973
Ferenbaugh,
1978
Frankhauser
et al. , 1976
Murray et al. ,
1975
Horsman
et al, 1978
Hill et al. ,
1974
1974
Masaru et al .
1976
0' Conner
et al. , 1974
aTable arranged by > SO. concentration as first order and exposure time as second order divisions. Doses within a single study that did not induce
-------
*
of fluoride contamination or used only sporadic monitoring schedules, respectively. Pollution
gradients were evident and S02 exposure was most probably involved, but conclusive proof of
losses was not presented as in the Sudbury area.
Table 8-6 summarizes the results of tree studies that have utilized artificial exposure
chamber systems under laboratory conditions. Only two studies (exposures) used doses close to
ambient concentrations (Houston, 1974); however, the use of selected clones of known sensitiv-
ity to S02 hinders further field speculation from this study. The remainder of the studies
presented in Table 8-6 have used doses above expected occasional exposures under field condi-
tions. Concentrations of greater than 0.25 ppm S02 for 2 hours were required to induce slight
injury to several pine species (Berry, 1972), but overall trends for increasing foliar injury
do not follow increasing dose for conifers per se. Smith and Davis (1978) exposed several con-
ifers (pine, spruce, fir and Douglas fir) to doses of 1.0 ppm S02 for 4 hours or 2.0 ppm S02
for 2 hours and only pines developed necrotic tips at the 2.0 ppm dose.
Studies done with tree seedlings under artificial conditions are difficult to extrapolate
to expected yield loss effects under field conditions. Of the studies reviewed and accepted
without caveats, none demonstrated significant height or annual increment growth effects.
In an evaluation of SO™ effects to photosynthesis, Keller (1977) used field chambers to
expose potted white fir, Norway spruce and Scotch pine to 3 concentrations of S02 (0.05, 0.10,
and 0.20 ppm and a control 0.0 ppm) for 10 weeks each during the spring, summer, and fall,
respectively. Several types of photosynthetic responses rates were obtained; however, trends
of decreasing rate occurred as dose increased especially when administered during the 10-week
spring period. Effects were less during the summer and fall periods and spruce responded
positively to S0? exposures of 0.05 ppm during the initial part of the fall period. Keller,
also reported that even with the most severe depression of photosynthesis, there were not
visible foliar symptoms in evidence. Keller (1980) utilized a similar exposure system and S0?
doses to study effects to the annual ring width in two clones of Norway spruce. He reported
significantly depressed C0? uptake with higher doses (0.10 and 0.20 ppm S02, 10 weeks), but
only a trend was noted for the 0.05 ppm S0? treatment; visible symptoms and trends of reduced
cambial growth occurred at only the higher dose.
On the limited data available regarding non-woody components of native ecosystems, there
appear to be no adverse yield effects below 0.06 ppm S02 for 6 weeks (Table 8-7). Most of the
acceptable literature deals with long-term exposures (several weeks) and results in S02 doses
well above, or not comparable, with the current standards. There is some indication of bene-
ficial yield effects below 0.06 ppm S02 on one species (i.e., Indian ricegrass). As reviewed
in Section 8.2.8, low doses of S0? may result in an increase of productivity within certain
crops. Concentrations of 0.03 to 0.06 continuously for 6 weeks of exposure increased the
XRD8A/A 8-35 2-9-81
-------
productivity of desert grass by 8 percent over the control plants were grown in indigenous
soils (Ferenbaugh, 1978). Other studies have demonstrated beneficial effects, but conditions
included a sulfur-deficient growing medium.
In spite of differences due to exposure regimes, techniques, and species, certain genera-
lizations can be made with respect to average and outer-limit responses of the plants under
study. These have been made in the form of correlations of yield response with total exposure
dose in part-per-million hours (ppmh). The latter data were calculated as the product of
exposure time and SCL concentration and transformed to log values. For experiments employing
controlled exposures under field conditions, data are graphed in Figure 8-6. For the 36 data
points shown, exposure dose ranged from 0.24 to 259 ppmh. No effects on yield were detected
in any of the six studies at doses < 6 ppmh. Yield losses occurred in 26 cases at levels > 6
ppmh while no effects and positive effects were noted in two cases each at levels > 6 ppmh. A
linear regression of yield on dose for all studies reporting yield losses showed strong posi-
tive correlation (r = .75) of yield with dose and took the form:
Yield loss = -13.6 + 23.8 (log dose)
r2 = 0.53 (Significance = > 0.001)
This correlation excludes four data points, two with no effect and two with positive responses.
All were studies with wheat reported by Wilhour (1978). Data from studies reporting no effect
or a positive effect are all plotted in Figure 8-6, however.
Data from Tables 8-4 through 8-7 are graphed together in Figure 8-7. These data were
derived from 59 yield responses involving 24 species or cultural varieties. Of these responses
no effect was found in 10 cases, a positive effect was noted in 6 cases (4 under low-sulfur
fertility), and a negative response was found in the remaining 33 cases. The linear fit to
all data points for which yield losses were detected gave the formula:
Yield loss = 13.7 + 11.1 (log dose)
2
r = 0.22 (Significance = > 0.003)
Yield loss showed a strong positive correlation (r = 0.48) with exposure dose. An alternate
approach to deriving a dose-response surface for these data involves describing the upper limit
of responses over the range of exposure doses studied. Such an approach, termed boundary-line
analysis (Webb, 1972), defines the yield losses under "worst-case" conditions (i.e., sensitive
species, sensitive conditions).
A boundary line for exposures below 500 ppmh in Figure 8-7 shows a threshold for effects
of 0.6 ppmh and maximum potential losses of approximately 12, 39, and 50 percent at 1, 10, and
100 ppmh, respectively. Average losses as estimated from the linear regression were 14, 25,
and 36 percent at these same dosage levels.
XRD8A/A 8-36 2-9-81
-------
I
t
HI
O
0.
C/J
01
cc
80
60
40
20
20
40
-|- - | -
REGRESSION LINE:
% YIELD LOSS = - 13.6 + 23.8 (LOG DOSE)
r~ = 0.53 P>F<0.001
22 DATA POINTS
0.1
1.0 10.0
EXPOSURE DOSE (ppmh)
100.0
Figure 8-6. Regression of yield response vs. transformed dose (ppmh) for controlled exposures
using field chambers (zero and positive effects excluded from regression analysis). See table 8-3 for
details of exposures.
3-37
-------
60
ui
VI
1
cc
Q
I
+
% YIELD LOSS = 13.7 + 11.1 (LOG DOSE)
R2 = 0.22. P>F<0.003
40 —
10.0
EXPOSURE DOSE (ppmh)
100.0
Figure 8-7. Regression of yield effects vs. transformed dose (ppmh) for laboratory and greenhouse
studies using agricultural, ornamental, and native herbs. (Zero and positive effects excluded from regres-
sion). See tables 8-4 and 8-7 for details of exposure.
8-38
-------
Only two studies in which quantitative data on S02 concentrations were related to
effects on tree growth were found. These represent the work of Keller (1980) with Norway
spruce and Linzon (1971) with white pine, and are shown in Figure 8-8. These data are
interesting in that they demonstrate a linear, and very different, response surface for
these two coniferous species over a rather broad range of log dose exposures (4-fold for
spruce and 50-fold for white pine). These data further confirm that plant response is not
equally affected by equal increments of S02 exposure dose.
In summary, our present dose-response data sets are heavily weighted towards controlled
exposures in the laboratory or in field chambers. In spite of the variety of species studied
and experimental protocols utilized, it is possible to derive potentially useful generaliza-
tions from these data:
(1) The concentration threshold for visible injury is generally lower than the thres-
hold for efforts on growth and yield. Doses causing visible injury to 10 percent
of a variety of southeastern plant species were 0.75 ppm for a 3-hr exposure and
0.50 for a 1-h exposure. The present NAAQS is 1.5 ppmh for I hr, and represents
an approximate average threshold for most plant species.
(2) Visible injury data emphasize the greater relative biological effectiveness of
short-term higher concentrations than longer exposures with the same total dose.
(3) Plant responses to SO^ may be positive, neutral, or negative over a rather wide
range of exposure dose. Positive responses were generally restricted to a very
few species or conditions when plants were known to have been grown in S-
deficient soils. Negative responses constituted approximately 85 percent of all
responses noted above threshold levels.
(4) Data derived from controlled exposures of six species or cultivars in field
chambers provided the most reliable basis for estimating yield responses from
total logarithmically transformed exposure dose. Regression analysis of these
data provided a no-effects limit of approximately 4.5 ppmh. Yield losses of
10 percent and 20 percent were similarly estimated at 10 and 27 ppmh, respectively.
(5) Data derived from laboratory and greenhouse exposures with 23 species or cultivars
indicated generally greater sensitivity of test plants and were described both
by a boundary line which delimited the maximum observed response over the range of
concentrations employed and a regression line. Upper-limit yield losses deter-
mined by this approach were approximately 10 percent and 20 percent for exposure
doses of 0.9 and 17 ppmh, respectively. (Average responses determined by regres-
sion analysis indicated that 10 and 20 percent yield losses would be produced by
exposures of 0.6 ppmh and 45 ppmh, respectively.)
In interpreting the dose-response information presented above, it should be noted that
responses of plants to S0? in the field may occur as a consequence of one or more short-term
episodes or as a result of the cumulative dose experienced over an entire growing season.
XRD8A/A 8-39 2-9-81
-------
50
40
V)
O
30
T 20
+
10
NORWAY SPRUCE (LAB)
WHITE PINE (FIELD)
10
100
1000
Figure 8-8. Yield responses vs. SC>2 dose for Norway Spruce (Keller, 1980) and White Pine (Linzon,
1971). See table 8-6 for details of exposure.
8-40
-------
The regression lines developed in Figures 8-6 and 8-7 were derived from predominantly low-dose
exposures in which a wide variety of exposure times were used and are intended to be used to
describe risks associated with cumulative exposures. These relationships do not infer that
single recent exposures with a total dose below the boundary line will not produce effects.
The risks associated with these short-term exposures are probably best delimited by the
visible injury data (Figures 8-2 and 8-5) which were developed around episodic exposure
conditions.
Under field conditions, exposure kinetics may be dominated by episodic contributions from
one or more point sources or they may be primarily a function of regional S02 loading and
atmospheric stagnation patterns. Calculations of the probabilities of growth impacts around
point sources should consider the average length and frequency of exposures as well as the
exposure concentration. Data collected in 1973 in the vicinity of a 1900 megawatt power plant
in Alabama demonstrate one useful approach to describing episodic exposures (Mclaughlin and
Lee, 1974). At this site S02 exposures occurred on a total of 40 percent of the days, lasted
an average of 3 hr each, comprised approximately 10 percent of the total daylight hours, and
provided approximately 230 h of exposure to S02 during the growing season. One-hr average
concentrations > 0.50 ppm occurred on an average of 12 hr at monitoring stations in the
vicinity of this plant and provided a total of 6 ppmh of "high-level" exposure.
Calculation of the phytotoxic potential for regional scale exposures involves many
assumptions regarding toxic and non-toxic components of the total dose to which vegetation is
exposed. Obviously not all, and probably most, exposures to S0? on a regional scale are below
levels producing phytotoxic reactions. An important aspect of evaluating the likelihood that
plants will be negatively influenced by S0? exposures is the determination of what components
of a plant's total exposure history are phytotoxic. Mclaughlin (1980) recently examined
Environmental Protection Agency (1978) data on regional SO- concentration averages. Using the
assumption that only the upper 10 percent of all SO, exposure days would have S09 concentra-
-1
tions high enough to cause stress to vegetation, and that only daylight exposures (8 hr/day )
during the active growing season (6 mo/yr ) would be effective, he calculated that the aver-
age potentially phytotoxic dose within designated air quality control regions would range from
0.9 ppmh (Region IX) to 5.5 ppmh (Region VIII). Maximum doses (highest reporting stations
within regions) ranged from 2.6 ppmh to 27 ppmh.
More definitive dose-response studies both with and without the addition of other pollut-
ants (see Section 8.4) are needed before the biologically significant features of typical
regional scale exposure regimes can be positively delineated. However, the above calculations
represent the type of data reduction which will eventually be needed to place air quality data
into biological perspective. An examination of the minimum levels of SO^ associated with
yield depression (regardless of time of exposure) indicates that this level is approximately
0.05 ppm for long-term exposures.
XRD8B/A 8-41 2-9-81
-------
8.4 EFFECTS OF MIXTURES OF S02 AND OTHER POLLUTANTS INCLUDING PARTICULATE MATTER
Ambient atmospheres usually contain more than one pollutant. Long-distance transport of
photochemical oxidants and oxidant precursors (Husar et al., 1977; U.S. Environmental Protec-
tion Agency, 1971), the demonstration of acidic rainfall over large areas of the eastern
United States (Cogbill, 1976), and atmospheric monitoring have documented that emissions
from specific sources are mixed with ambient concentrations of one or more associated pollu-
tants (Shriner et al., 1977). Extrapolation from results of single pollutant effects on vege-
tation under ambient field conditions must be approached with caution. Reactions to pollu-
tant combinations may be additive (sum of effects), less than additive (antagonistic), or
more than additive (synergistic). In addition to pollutant combinations under controlled
conditions, the interaction of constantly changing environmental factors and fluctuating
pollutant doses must be further evaluated before a conclusive statement of the importance of
such interactions can be made. Reinert (1975) and Reinert et al. (1975) have prepared the
most recent reviews of this area of investigation. Some examples of the available literature
follow.
8.4.1 Sulfur Dioxide and Ozone
A more than additive reaction on vegetation was first noted with ozone and S0? (Menser
and Heggestad, 1966). Tobacco were severely injured by 0.03 ppm ozone and 0.24 ppm SO^ when
the pollutants were combined for either 2 or 4 hours, whereas when used alone, neither pol-
lutant produced foliar symptoms.
Since that first report, the effects of mixtures of ozone and S0? have been studied
using a variety of plant species. Radish and alfalfa plants showed more than additive foliar
injury from a 4-hr exposure to a mix of 0.10 ppm 03 + 0.10 ppm S02 (Tingey et al., 1973a), but
less than additive growth reduction (top and root weights) from an 8-hr, 5 day/wk, 5-week
exposure of radish (alfalfa total exposure time unknown) to a mix of 0.05 ppm 0., + 0.05 ppm
«J
SOp (Tingey et al., 1971a; Tingey and Reinert, 1975). Greater than additive foliar injury
effects have also been reported for broccoli and tobacco, while additive or less than addi-
tive effects have been noted for cabbage, tomato, lima bean, bromegrass, spinach, onion,
and soybean (Tingey et al., 1973a). Soybean has exhibited non-significant less than additive
foliar injury effects (Tingey et al., 1973a) while exhibiting significantly greater than
additive growth effects (Tingey et al., 1973b).
Most research examining the effects of pollutant mixtures has utilized standard means
comparisons to express the responses. These tests usually do not adequately evaluate the
interaction: the failure of one pollutant to be consistent at different concentrations of
the second pollutant. Reinert and Nelson (1980) utilized sums of squares partitioning by
factorial analysis to examine the effects of 0.5 ppm S02 and 0.25 ppm 0, (4-hr exposures,
4 times, 6 days apart) on Begonia. A significantly less than additive effect was found
for flower weight of 1 to 5 cultivars. The same technique was utilized by Reinert and
Weber (1980) to evaluate the effects of 0.25 ppm 03 and 0.25 ppm S02 (exposed 4 h/day,
XRD8B/A 8-42 2-9-81
-------
*
3 days/wk, for 11 wks) on soybean (Dare); an additive effect of the pollutant mixture was
demonstrated.
Field-grown soybeans (cv. Dare) exposed to 0.10 ppm 03 alone or 0.10 ppm 03 + 0.10 ppm
S02 for 6 hr/day for 133 days in field chambers exhibited injury and defoliation. Injury
and yield due to the mixture were increased (9 percent) and decreased (19 percent), respec-
tively, compared to the ozone-alone treatment, but the differences were not significant
(Heagle et al., 1974). Two cultivars of bean exposed to sulfur dioxide and ozone showed
interactive effects between these two gases, but the magnitude and direction of the effects
depended on the cultivar and on the pollutant concentrations (Jacobson and Colavito, 1976).
Alfalfa exposed in closed field chambers to low levels of ozone and sulfur dioxide
single and in combination for varying periods of time exhibited significant reductions in
yield, quality, and nitrogen fixation compared with the control plants, but there were no
significant interactive effects (Neely et al., 1977).
Many studies have been conducted on the effects of mixtures of sulfur dioxide and
ozone on eastern white pine (Pinus strobus L.) (Costonis, 1973; Dochinger and Heck, 1969;
Houston, 1974; Houston and Stairs, 1973). Genetic control of sulfur dioxide and ozone
tolerance in this species has been demonstrated for low concentrations of SCL (0.025 ppm)
and Op (0.05 ppm) for only 6 hr with consistent injury to the exposed sensitive clones
(Houston and Stairs, 1973). Houston (1974) later used mixtures of sulfur dioxide and ozone
and doses to simulate actual field conditions and reported that even the lowest concentra-
tions of 03 (0.05 ppm) and SOp (0.05 ppm) for 6 hr in mixture caused more serious damage
than that resulting from either pollutant alone at similar concentrations. A less than
additive effects on foliar injury was noted when Scotch pine trees were exposed to 0.25 ppm
SOp and/or 0.14 ppm or 0.29 ppm 0,, 6 hr/day for varying time periods (Neilson et al., 1977).
Exposure of aspen clones to 0.05 ppm 03 and/or 0.20 ppm SOp for 3 hr resulted in a
more than additive number of plants in the mix exhibiting foliar injury (Karnosky, 1976).
Table 8-8 lists some selected 0, + S0? combination studies.
Oshima (1978) conducted a field experiment using constant-stirred, round, closed
chambers to assess the effect of SOp (0.10 ppm, 6 hr/day, total 335 hrs) on yield of potted
red kidney beans, Phaseolus vulgaris exposed to a gradient of ozone doses; that is, the
ozone present in 0, 25, 50, 75, and 100 percent charcoal-filtered air at Riverside, Cali-
fornia. The temperature, light, and humidity approximated that of ambient air; plants
were in the chambers for 78 days. An interaction with ozone and SOp was documented in the
50 percent carbon filtered treatment (5144 ppm-hrs) and produced a significant reduction
in yield (37 percent) and plant biomass. The data also indicated the suggestion of an
interaction in the 75 percent filtered treatment (2822 ppmh-hrs) (yield reduced 17 per-
cent), but at an unacceptable level of significant (p = .20 level). Sulfur dioxide at
10 pphm did not produce detectable plant or yield responses alone and did not have an
interactive effect at ozone doses exceeding 5144 pphm-hrs.
XRD8B/A 8-43 2-9-81
-------
TABLE 8-8. EFFECTS OF MIXTURES OF SOj and
ON PLANTS
Cone.3 Exposure9
ppm Time
0.025 SO, 6 hr
+ 0.05 0^
b
Exposure
Condition
EC/SO
Plant Foliage Productivity Species effect Caveat6
E. white X X No effect to needle elongation.
pine Foliar injury on sensitive clones
only; 10 trees with 75-100% of needles
with tip necrosis.
SO, alone caused tip necrosis on 75-
,100% of the needles on 1 tree.
0.. alone caused no injury.
Reference
Houston, 1974
0.05 SO, 8 hr/d, 5 d/wk EC/SO
* 0.05 63 5 wks
0.05 SO, 8 hr/d, 5 d/wk EC/SO
+ 18 days
0.05 0,
Radish X
Soybean
Plant weight reductions additive (leaf
fresh and dry weight) or significantly
less than additive (plant fresh wt.,
root fresh and dry weight).
Additive foliar injury effects, greater than
additive root dry weight
Tingey et al.,
197 Ib
Tingey et al.,
1973b
0.05 SO, 8 hr/d, 5 d/wk EC/SO
+ 4 wk
0.05 0,
Tobacco
Additive growth reductions
Tingey and
Reinert, 1975
8 hr/d, 5 d/wk
until control
plants 40-45 cm
high
m 0.06 SO, 7 hr/day
-k +0.05 6, 68 days
F/CC
0.075- 5 or 10 days EC/SO
0.60 SO,
+ 0.15 6,
Alfalfa
Alfalfa
White bean
X X
Less than additive growth reductions
No significant alteration of plant responses Potted plants Neely et al.,
(carbohydrate, protein, dry weight) compared set on soil 1977
to the effects of single pollutants surface; grown
hydroponically
Less than additive growth reductions and Hofstra and
foliar injury Ormrod, 1977
0.10 SO, 6 hr/d
+0.10 133 days
Soybean X Less than additive foliar injury
F/CC Soybean X X SO, alone and in the mix did not significantly
affect the yield and injury responses
Heagle et al.
1974
0.10-
0.50 SO,
t 0.05-0.10
o,
4 hr
EC/SO
Alfalfa
Broccoli
Cabbage
Radish
Tomato
Tobacco W,
Greater than additive foliar injury at 0.10
ppm of each gas for alfalfa, brocolli, and
radish. Less than additive effect for tomato.
At 0.25 ppm, SO. + 0.10 ppm 0, greater than
additive injury noted on alfalfa, radish, and
tobacco. At 0.50 ppm, SO, and 0.05 ppm 0,
greater than additive injury on broccoli
and tobacco and less than additive injury on
alfalfa. At 0.50 ppm, SO- and 0.10 ppm 0,
greater than additive effects on alfalfa,
cabbage, radish, tobacco.
Tingey et al.
1973a
-------
TABLE 8-8. Continued
Conc.a
ppm
0.20 SO,
+ 0.05 63
0.35 SO.
* 0.05 03
0.24 SO,
+ 0.27 Oj
0.28 SO,
+ 0.28 63
0.25 SO,
+ 0.05 03
1.00 SO,
? + 0.10 Z
°3
0.25 SO,
+ 0.25 i
°3
0.45 SO,
+ 0.15 6r
0.45 Oj
0.25 SO,
«• 0.14 6,
0.25 SO,
+ 0.29 6,
0.50 SO,
+ 0.25 6,
d
Exposure3 Exposure
Time Condition
3 hr EC/SO
3 hr EC/SD
2 hr GC
4 hr GC
4 hr EC/SD
4 hr/d, 3 d/wk EC/SD
11 wks
4 hr EC/SD
10 6hr/day EC/SO
68hr 5day/wk
154
164
10 6hr/day EC/SD
68hr 5day/wk
154
164
4hr/day EC/SD
4 times.
6 days apart
Effects onc
Plant Foliage Yield
Trembling X
Aspen
(5 clones)
Trembling X
Aspen
(5 clones)
Tobacco X
Bel-W3
Bel-B
Consolidation 402
Tobacco X
Bel-W3
Bel-B
Consolidation 402
Alfalfa X
Onion
Soybean
Tobacco Bel-B
Tobacco
White gold
Tobacco Bel-W3
Lima bean
Broccoli
Bromegrass
Cabbage
Radish
Spinach
Tomato
Soybean X X
Radish X
Scotch Pine X
Scotch Pine X
Begonia X X
(5 cultivars)
Species effectd Caveat6
Greater than additive injury to 3 clones
no injury due to SO. alone
Greater than additive injury to 4 clones
9-38% foliar injury—no injury due to either
pollutant singly
23-76% foliar injury—no injury due to either
pollutant singly
Only tobacco Bel Wj showed greater than Bel W3 tobacco
additive foliar injury very sensitive
At 1.00 ppm SO, tobacco Bel B and Bel W,
exhibited greater than aditive effects,
and there were less than additive effects
for Bromegrass, cabbage, spinach, and tomato.
Additive growth effects
Additive growth effects
Less than additive effects— no effects
due to 0, alone
Less than additive effects — no effect due to
0, alone
Less than additive effects for flower weight
of one cultivar. 0.50 SO- alone significantly
reduced flower production in the absence of
foliar injury for one cultivar
Reference
Karnosky, 1976
Karnosky, 197J6
Menser and
Heggestad,
1966
Menser and
Heggestad,
1966
Tingey et al . ,
1973a
Reinert and
Weber, 1980
Tingey and
Reinert, 1975
Nielsen et al. ,
1977
Nielsen et al . ,
1977
Reinert and
and Nelson,
1980
-------
*
8.4.2 Sulfur and Nitrogen Dioxide
The occurrence of S02 and nitrogen dioxide (N02) has been associated with power plant
plumes as well as mobile sources. However, ambient concentrations of N02 seldom reach the
injury threshold, and the literature for N02 suggests that any injury associated with N02
results from interactions with other pollutants (Jacobson and Hill, 1970).
No injury occurred to oats, beans, soybeans, radish, tomato, or tobacco following
exposure for 4 hr to up to 2 ppm N02 or 0.50 ppm S02- However, at 0.10 ppm of each gas
for 4 hr, injury was noted on all species; at 0.05 ppm of each gas, slight injury was
noted on all species except tomato (Tingey et al., 1971a). A greater than additive suppres-
sion of the apparent photosynthetic rate of alfalfa was obvious when exposed to 0.25 ppm
of S02 and/or N02 for 2 hr (White et al., 1974). At 0.15 ppm of each gas singly, there
were no measurable effects, but a 7-percent suppression of apparent photosynthetic rates
was noted in the mixture (White et al., 1974).
Field exposure of seven different species of plant indigenous to the cold desert
areas of the southwestern United States to 0.50-11.0 ppm S02 singly, or 0.50-11.0 ppm
S02 and 0.10-5.00 ppm N02 combined in 2-hr fumigations resulted in no evidence of more
than additive foliar injury (Hill et al., 1974).
More than additive foliar injury was noted on radish leaves exposed for 1 hr to
0.50 ppm SOp and/or 0.50 ppm N02_ No interactive effects were found for other plants
tested (oats, Swiss chard, and pea) (Bennett et al., 1975). More than additive effects
have been noted for the enzyme activity in pea plants exposed to 0-0.20 ppm S02 and/or
0-0.10 ppm N02 for 6 days. Peroxidase activity was increased and ribulose-l,5diphosphate
carboxylase activity was decreased (Horsman and Wellburn, 1975). Some selected S0? + N0?
combination studies are shown in Table 8-9.
8.4.3 Sulfur Dioxide and Hydrogen Fluoride
Linear growth and leaf area suppressions (in the absence of foliar injury) of Koethen
orange plants exposed to S0? (0.80 ppm) and/or hydrogen fluoride (2.3-19.4 ppb) for 23
days were no greater than additive. Satsuma mandarin plants exposed to the same condi-
tions for 15 days exhibited only additive foliar injury effects, and no growth suppres-
sions at all (Matsushima and Brewer, 1972). Greater than additive foliar injury was
exhibited by barley and sweet corn exposed to 0.06-0.08 ppm S0? and/or 0.60-0.90 ppb
hydrogen fluoride for 27 days. Using higher concentrations of S09 for only 7 days
resulted in simply additive foliar injury effects. Pinto beans were not injured in any
of the treatments (Mandl et al., 1975).
8.4.4 Sulfur Dioxide, Nitrogen Dioxide and Ozone
Fujiwara et al. (1973) combined So2, N02, and 03 at concentrations ranging from 0 to
0.2 ppm in an artificially controlled environment and exposed peas and spinach for 5 hr.
Ozone was the most injurious, S02 was next, and N02 elicited only minor injury. More
than additive foliar injury followed exposure to S02 + 03, but only additive effects were
XRD8B/A 8-46 2-9-81
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TABLE 8-9. EFFECTS OF MIXTURES OF S02 AND N02 ON PLANTS
Cone.8 Exposure8 Exposure
ppm Time Condition
0.05-0.25 4 hr EC/SO
SO, + 0.05-
25*N02
0.15-0.25 4 hr EC/SD
SO. + 0. 10-
0.20 N02
0.11 SO, 103.5 h/wk EC/SD
+ 0.11 N02 20 wk
0.125-1.0 Ihr or 3hr EC/SD
SO ,+0.125
-ifO N02
0.15-0.50 Ihr and 2hr EC/SD
S0,+0.15-
O.*0 N02
0.2 SO, 6 days EC/SD
* 0.1 6r
1.0 N02
0.5-11.0 2 hr F/CC
SO +
0.1-5.0 N02
0.8 SO, 2 hr EC/SD
+ 0.3 N02
Plant
Tobacco
Pinto Bean
Tomato
Radish
Oats
Soybean
Pinto Bean
Tomato
Radish
Oats
Cocksfoot
Meadow-grass
Radish
Swiss chard
Oats
Sweet pea
Alfalfa
Pea
87 desert
species
Alfalfa
aT,kl j . , „ cn, concentration as first order
J^c?f"ESf?Sle8i^ereS?2effects are listed along with
Effects onc
Foliage Yield
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
Species effect*1
0-2X foliar injury at 0.05 ppm S0,+
0.05 ppm NO,. 1-35% foliar injury
at 0.10 ppm SO, and 0.10 or 0.15
ppm N0?. Injufy less at 0.20-0.25
ppm SO, than at 0.10 ppm SO,. Thres-
holds--S02=0.50 ppm, N0=2.00 ppm
No foliar injury
Greater than additive decreases in
# of tillers, # leaves, and leaf
area
Greater than additive foliar injury
to radish at higher concentrations.
Thresholds for radish were 0.50 ppm
of each gas, and for the other species
0.75 ppm of each gas.
Slight but significantly greater than
additive depression of photosynthesis
at S02 concentrations of 0.15 and 0.25.
No greater than additive effects at
0.35 or 0.50 ppm S02.
Significantly greater than additive
increase in peroxidase and RuDPC
enzyme activity at 0.20 ppm SO, and
and 0.10 ppm W>2 *•
No evidence of SO, + HO, synergism at
an N02/S02 ratio of 0.28. Most
species required over 2.00 ppm S02
to cause injury.
Apparent photosynthesis reduced at
315 ppm CO, taut increased at 645
ppm C02
and exposure time as second order divisions. Doses within a
the lowest SO, concentration that induced said effect.
Caveat6
Recirculating air
Experimental condi-
tions approximated
those of Tingey, et
1971b, but used dif-
ferent cultivars of
tomato, radish, and
oats
Wintertime exposures
Recirculating air
Reversible effects
Reversible effect?
Plants exposed to
ambient air before
and after fumigation
Reversible effects
Reference
Tingey, et al. ,
1971a
Bennett et al. ,
1975
al.,
Ashenden,
1979
Bennett et al . ,
1975
White et al. ,
1974
Horsman and
Wellburn,
1975
Hill et al. ,
1974
Hou et al. ,
1977
single study that did not induce
jX indicates study found foliar and/or yield effects.
Most prominent or significant effect reported.
Caveats for consideration about proper study design and interpretation.
-------
observed with S02 + N02 or N02 + 0.,. The addition of N02 to the 03 + S02 had little
effect on foliar injury. Reinert and Gray (1981) examined the effects of 0.2 or 0.4 ppm
each of 0-., S09, and NO, (3- or 6-hr exposures) on the growth of radish. Through parti-
*5 c. C.
tioning, the main effects of each pollutant and the potential interactive effects of each
mixture were examined. Sulfur dioxide depressed the root/shoot ratio at both 0.2 and
0.4 ppm; however, when N0? and S0? were both present there was a greater than additive
depression of the root/shoot ratio at 0.4 ppm.
8.4.5 Summary
As can be seen from the preceding research, plant species vary in their responses to
pollutant mixtures, and the type of response (additive, less than additive, greater than
additive) may depend on the parameter measured. Our understanding of how pollutant
combinations influence plant growth and development, and how environmental factors can
modify those responses is still fragmentary. Insufficient information exists to deter-
mine the influence of pollutant sequencing during combination exposures, meteorological
influences, the effect of various cultural practices, and many other variables in rela-
tion to vegetation effects induced by S0? combined with other pollutants.
There is a need to determine the best technique for evaluating the effects of pol-
lutant mixtures. Only recently has the partitioning technique been utilized to express
the effects of pollutant mixtures. This technique allows for separation of the main
effects of each pollutant, and also provides a statistical test of the significance of
potential interactions between pollutants.
The data have demonstrated that interactions can occur between pollutants and, due
to the occurrence of pollutant mixtures in the ambient situation, knowledge of interactive
effects is very important. However, the nature of the effects of pollutant mixtures is
extremely complex. Most research studies have necessarily taken a rather simplistic
approach to this complex problem. It is, therefore, difficult to relate these relatively
few results to "real world" situations.
8.5 EFFECTS OF NON-POLLUTANT ENVIRONMENTAL FACTORS ON S02 PLANT EFFECTS
The physical environment plays an extremely important role in determining response
to S0«. Most evidence has accumulated on factors leading to or inhibiting the ingress of
SO™ into stomates and immediate plant reactions as determined by the metabolism of the
plant at the time of exposure; the metabolic state of the plant is likewise affected by
the physical environment. As illustrated in the following sections, the response of any
given plant species may also be quite different from any other given species grown under
identical physical conditions.
8.5.1 Temperature
Temperature plays an important part not only in determining the metabolic rate of
the plant, but in determining (with moisture, fertility, and light) species diversity
and richness of a given ecosystem (NAS, 1978). The primary path of entry of S0« into the
XRD8B/A 8-48 2-9-81
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leaf is through stomata. Temperature exerts an effect on the guard cells that control the
stomatal opening and closing and thus the entry of SCL. Temperature regimes that increase the
physiological activity of the plant may also increase the plant response to S02 (Heck and
Dunning, 1978). It is generally believed that plant sensitivity increases with temperature
over a wide range, from about 4 to 35°C (Guderian, 1977; Rist and Davis, 1979). Several
studies suggest greater resistance of conifers to SOp in the winter, attributed to lower rates
of physiological activity (MAS, 1978). However, according to Guderian and Stratmann (1968),
in areas with SC^ emissions, winter wheat and winter rye are more severely injured than the
summer varieties. Guderian (1967) interpreted this to be due to gas exchange taking place
through stomata at temperatures as low as -2°C.
8.5.2 Relative Humidity
Relative humidty exerts important control over plant sensitivity to S0? both by affecting
stomatal opening and closing (Bonte et a!., 1975; Majernik and Mansfield, 1970; Mansfield and
Majernik, 1970; Buron and Comic, 1973) as well as by affecting the internal leaf resistance
to S02 flux (Mclaughlin and Taylor, 1981). Although plant sensitivity increases with relative
humidity, Setterstrom and Zimmerman (1939) found rather large changes (> 20 percent) were
required to cause change in plant sensitivity once RH levels become > 40 percent. Mclaughlin
and Taylor (1981), in laboratory studies, found 2 to 3 fold increases in S02 uptake by kidney
beans over a range of S0? concentrations (0.16-0.64 ppm) as relative humidity was raised from
35 percent to 78 percent during exposure times of < 3 hours. According to Zimmerman and
Crocker (1954), although relative humidity is important in governing sensitivity and conse-
quently the sensitive plant population, it is not as important as the tissue turgidity as may
be influenced by soil moisture as well as relative humidity. Based on the water relations in
certain trees, Halbwachs (1976) has rated plants as sensitive, intermediate, and tolerant at
relative humidities of over 75 percent, 50 to 75 percent, and below 50 percent, respectively.
8.5.3 Light
Light also controls stomatal opening and thus plant sensitivity. Plants are more tolerant
when fumigated in darkness with S0? or when held in the dark for several hours before exposure
(Zimmerman and Crocker, 1954). This relationship is complex, since injury is greater if the
night exposure follows a daylight exposure (NAS, 1978).
Setterstrom and Zimmerman (1939) observed that buckwheat grown at a light intensity of 35
percent or less of full sunlight was more sensitive to S0? than when grown under full sunlight.
Other investigators have found that injury was more severe when tomato stems and foliage were
fumigated on clear days than it was on cloudy days (MAS, 1978).
Plants seem to be more sensitive from midmorning to midafternoon, in spite of a high light
intensity that might continue after midafternoon (Rennie and Halstead, 1977; Thomas and
Hendricks, 1956). At the same time, plants may be more sensitive in the morning during good
XRD8B/A 8-49 2-9-81
-------
weather, but may become more sensitive if temperature and light increase in late afternoon (Van
Haut and Stratmann, 1970).
8.5.4 Edaphic Factors
Soil factors influence directly and indirectly the responses of plants to Sf^. Soil
fertility, moisture, and soil physics directly influence plant sensitivity to SC^ (NAS,
1978). Adequate soil moisture and the resultant stomatal opening have been shown to
increase the degree of plant sensitivity, whereas wilting conditions confer tolerance
(Setterstrom and Zimmerman, 1939; Zahn, 1963; Zimmerman and Crocker, 1954). As long as
plants are grown with an adequate supply of water, they are much more sensitive to SC^ than
are plants grown with an adequate supply, even though the moisture content of the soil is
the same at the time of fumigation (Setterstrom and Zimmerman, 1939). Withholding water
from some crops during periods of high pollution risk has been suggested (Brandt and Heck,
1968).
Soil fertility has a significant influence on plant response to SG^. Some plants
become more tolerant to SCL upon fertilization (Enderlein and Kastner, 1976; Zahn, 1963).
However, with eastern white pine, increased nitrogen, phosphorus, and potassium concentra-
tions in the greenhouse raised tolerance (decreased needle necrosis) in sensitive clones,
but did not prevent chlorotic banding in the field (Cotrufo and Berry, 1970). Nitrogen
and sulfur deficiencies were correlated with increased tolerance to S0« in tobacco and
tomato (Leone and Brennan, 1972). Conversely, nutrient deficiencies increased S0? sensi-
tivity in alfalfa (Setterstrom and Zimmerman, 1939). Fertilization of several dicotyledons
with a complete fertilizer has been effective in decreasing their sensitivity to S0?) but
similar treatment of monocotyledons like oats and barley have been ineffective (Van Haut
and Stratmann, 1970; Zahn, 1963).
8.5.5 S0? and Biotic Plant Pathogen Interactions
Plant disease is caused by the interaction of a plant and a pathogen acting under
suitable environmental conditions. The influence of S0? directly or indirectly on the
interrelations of a given plant and its possible biotic pathogens has been difficult to
investigate. Additionally, whenever the variables of the physical environment would
be considered within such experimental sequences, the subject becomes even more difficult
to examine. Heagle (1973) and Laurence (1978) have provided the most recent reviews of
the interaction between air pollutants and plant parasites.
In seven of nine plant diseases reported in S0?-related studies as reviewed by
Laurence (1978), there was no effect or a reduction in disease development demonstrate;
disease increased only in needle case of pine (Chiba and Tanaka, 1968) and increased
virus titer of southern bean mosaic virus has been reported by Laurence et al. (1981).
In a recent study by Laurence et al. (1979) maize and wheat were exposed to 0.10
or 0.15 ppm SO,, for either 2 or 10 days and innoculated at various times with
Helminthosporium maydis or Puccini a graminis. The ability of these fungi to infect either
XRD8B/A 8-50 2-9-81
-------
*
corn or wheat, respectively, was inhibited by SOp exposures; greater inhibition occurred if
plants were fumigated prior to inoculation attempts.
Studies done under ambient conditions without monitoring of other pollutants have sug-
gested a decrease in disease incidence in areas of higher S0» pollution with the possible
exception of those pathogens which are able to better invade weakened plants. If S02 expo-
sure has resulted in an overall weakened condition, other agents such as root invading
fungi may be able to gain entrance into an otherwise resistant host. Such is the suspected
reason for increased incidence and severity of attack by the root pathogen Armillaria mellea
in trees weakened by S02 (Donaubauer, 1968; Jancarik, 1961; Kudela and Novakova, 1962).
The effects of SO,, on infection by organisms other than fungi have also been studied.
Abies concolor and A. vertchi were severely attacked by plant lice in an environment of
high SO,,, but Pinus strobus was attacked less and P. griffithi and P. sylvestris were not
attacked (Stewart et al., 1973).
A direct influence of S0? on plant pathogenic fungi has been demonstrated and a review
has been presented by Saunders (1973); no direct effects of S0? on plant pathogenic bacteria
have been reported.
8.6 PLANT EXPOSURE TO PARTICULATE MATTER
8.6.1 Deposition Rates
Deposition of particles is strongly dependent on particle size. Most sulfates and
nitrates are found in the size range of 0.1 to 1.0 urn, and very little information if avail-
able on the deposition rate for these particles. Shinn (1978) divided particulate deposi-
tion into three categories based on particle size:
CATEGORY 1. Particles more than 10 urn in diameter; includes dust
and spores.
CATEGORY 2. Particles between 1 (jm and 10 urn in diameter where the
collection efficiency is highly dependent on the parti-
cle diameter.
CATEGORY 3. Submicron particles between 0.1 and 1.0 urn in diameter,
which have a nearly constant collection efficiency.
Current experimental data suggest that collection efficiencies in Category 2 are at
least 10 times less than in Category 2 (Shinn, 1978). According to Clough (1975), in the
range of wind speeds normally encountered, the larger particles in the atmosphere are much
more efficiently collected than the smaller fraction.
Little (1977) evaluated the effects of leaf surface texture on the deposition of mono-
disperse polystyrene aerosols on the leaf surfaces. Rough and hairy leaf discs collected
5.0 urn particles up to seven times more efficiently than did smooth leaves. Very large
differences in particle deposition velocities were observed between the laminas, petioles,
and stems of each species. The velocity of deposition of particles to plant surfaces
varies according to both wind speed and particle size.
XRD8B/A 8-51 2-9-81
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Further information on atmospheric transport transformation and deposition of particu-
late matter may be found in Chapter 6 of this document.
8.6.2 Routes and Methods of Entry Into Plants
Direct Entry Through Foliage—Foliage is continuously subjected to natural and man-
made coarse particles that are insoluble or sparingly soluble in water. Coarse particles
in general are too large to enter leaves through stomata. Based upon a review by Meidner
and Mansfield (1968) which presented stomatal data for 27 species of plants (e.g., pine,
oak, corn, soybean, and tobacco) the overall average pore (opening) width is 6 microns which
accounted for 0.15 to 2.0 percent of the average stomate total area. In certain cases,
such as with cement kiln dusts (Lerman and Darley, 1975) and other types of aggregate parti-
cles (Smith, 1977), a limited amount of stomatal clogging can occur. This is apparently
dependent on the statistical probability of the particulate matter falling on the stomata,
the size of the particle, and the stomatal aperture. In many plants, the stomatal opening
is on the lower surface. Cement kiln dust forms a crust on leaves, twigs, and flowers.
According to Czaja (1961) crusts of this type form because some portion of the settling
dust consists of calcium aluminosilicates typical of the clinker from which cement is made.
Hydration of the dust on the leaf surface results in the formation of a gelatinous calcium
aluminosilicate hydrate which later crystallizes and solidifies to a hard crust.
When coarse particles are water soluble or have some water-soluble components, plant
uptake of ions from the leaf surface does occur. Because of analytical difficulties, the
exact magnitude of the uptake is difficult to measure. Since it is not possible to predict
the efficiency of any washing procedure used to remove particles from the leaf surface, it
is difficult to delineate and separate the concentration of a given element on the leaf sur-
face from its concentration inside the tissue. In addition, leaching of elements from
within the tissue is known to occur during the washing (Little, 1973).
Smith (1973) evaluated metal contamination of urban woody plants by using a variety
of washing procedures. Indirect evidence from all these studies suggests that a variable
concentration of metals originating from coarse particulates can accumulate in plant
foliar tissue through direct uptake.
At this time, one of the significant problems in deriving conclusions concerning the
magnitude of direct foliar deposition and uptake of atmospheric particulates is the lack
of coordinated data on size and frequency distribution of the particle and their chemistry,
rates of deposition, and dose in conjunction with changes in tissue -concentrations over
time relative to background conditions.
Indirect Entry Through Roots—Many of the inorganic constituents of particulate air
pollutants occur naturally in the soil. Deposition of these pollutants may increase the
soil concentrations of the chemical species in question. Some of the chronic effects
caused by particulate air pollutants may result from changes in soil physics and chemistry
and from increased plant uptake of either the added elements associated with the particles
XRD8B/A 8-52 2-9-81
-------
themselves or some other soil-borne elements made more available by the influence of the
deposited particles.
It should be recognized that only a portion of the total elemental content of the soil
is available at a given time for plant absorption (Brady, 1974). As uptake of elements pro-
ceeds, there may be a redistribution of nutrients or toxicants in the soil.
The availability of nutrients or other chemical elements from the soil is strongly
influenced by type, chemical composition, and acidity of the soils. Plant nutrients,
when present in optimal amounts, may usually be available at a neutral pH; however, when
the soil becomes acidic, toxic elements such as aluminum become available.
8.7 REACTION OF PLANTS TO PARTICULATE EXPOSURE
8-7.1 Symptomatology of Particle-Induced Injury
Particulate-induced injury to plants has most often been associated with sustained
accumulation of particles such as dust or fly ash. Few investigations have dealt with
direct or indirect chemical interactions at the plant surface or subsequent effects. The
toxicity of accumulated heavy metals in soils has been established for several plant
species.
The various forms of particulates and their associated impacts on plants have been
reviewed (Darley, 1966; Lerman and Parley, 1975; U.S. Department of Health, Education, and
Welfare, 1969; U.S. Environmental Protection Agency, 1977a). Krupa et al. (1976) and
Linzon (1973) have also prepared an extensive review of various forms of heavy metal
depositions and impact. Tolerance of plants for heavy metals and fine particles and
their bioenvironmental impacts have also been reviewed (U.S. Environmental Protection
Agency, 1975).
Dusts directly affect plants by coating exposed plant parts, including leaves, stems,
flowers, and fruits (Jennings, 1934; Katz, 1967; Linzon, 1973). Depending on the chemical
nature of the particles and environmental conditions, deposits may accumulate as dry dusts,
as encrustations in the presence of free moisture, or as greasy films or tars. Encrusta-
tions of particulates on leaves result in reduced gas exchange, increased temperature,
reduced photosynthesis, and eventual yellowing and tissue desication (Daessler et al., 1972;
Parish, 1910).
Terminal growth reduction and chlorosis of 2-year needles of hemlocks coated with
heavy dust deposits has been reported. Manning (1971) also reported that fungal propa-
guels increased and bacterial numbers decreased on such needles. Brandt and Rhoades
(1972) reported long-term changes in plant community structure and species composition,
and later indicated that radial growth rates were reduced in the tree species involved
(Brandt and Rhoades, 1973). On the exposed site they demonstrated a reduction in radial
growth of red maple (18 percent, chestnut oak (29 percent), and red oak (23 percent)
but a 76-percent increase in radial growth of tulip poplars as compared with representa-
tives of these species growing on a similar but nonexposed site. Reduction in growth of
XRD8B/A 8-53 2-9-81
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the dominant species (oak, maple) most probably gave a competitive advantage to tulip
poplar and greater than expected increase may have occurred.
The deposition of limestone dusts has caused substrate pH changes followed by lichen
community changes, namely, replacement of acid-loving communities of lichens by more
alkaline-loving species (Gilbert, 1976). A reversal of this trend occurred in areas where
S02 was of importance prior to limestone dust emission. No exact pollutant dose of either
limestone dusts or S0? were reported. Winter S0? levels were estimated to average 65 ug/m .
Cement kiln dusts have been collected from precipitators applied to vegetation. Visi-
ble effects were demonstrated on beans following application of particles of > 10 urn at
rates of 0.05 mg/cm /day to 0.38 mb/cm /day for 2-3 days. The lower dose induced a slight
reduction in carbon dioxide exchange, and the two higher doses reduced carbon dioxide
uptake by 16-32 percent (Darley, 1966).
The accumulation of dust caused increased reflection of solar radiation in wavelengths
of 400 to 760 nm and has been demonstrated to reduce photosynthesis (Ricks and Williams,
1974). Conversely, increased absorption of solar radiation by dusted leaves at wavelengths
750-1350 nm has been demonstrated to lead to heat stresses within the leaf tissues (Spinka,
1971).
Growth and yield effects induced by the accumulation of dust have recently been
reviewed (U.S. Environmental Protection Agency, 1975). Conflicting reports of yield
increases and decreases from such accumulation appear to be caused by variations in doses
applied, substrate nutrient balances and pH, and other specific physiological interferences
with processes such as pollination of fruit trees (Anderson, 1914).
Dusts, therefore, have only been considered of importance to vegetation growing near
emission sources. Accumulation of dusts has been demonstrated to reduce photosynthesis
and radial-increment growth of some forest tree species but has increased them in other
species.
The phytotoxicity of heavy metals, arsenic, and boron has been demonstrated after
accumulations in soils and subsequent uptake by various plants. Table 8-10 presents a
summation of toxic effects of individual elements (Krupa et al., 1976). Published reports
of direct effects on plants from specific sources are discussed in the following paragraphs.
Arsenic—Only one study was available to show foliar uptake of airborne arsenic
(Linzon, 1977). Phytotoxicological studies in the vicinity of gold smelters in Ontario,
Canada revealed the occurrence of several injuries to vegetation (primarily fireweed,
Epi1 obi urn angustifolium) as induced by airborne arsenic compounds in a sulphurous plume.
Chemical tissue analysis of affected leaves revealed arsenic at 200 ppm as compared to
1 ppm arsenic in leaves collected 50 miles from the source area. Linzon (1977) suggested
10 points of evidence leading to the conclusion of the airborne nature of the toxicant
including elevated concentrations on sides of plants closer to the source and correlation
of arsenic emissions (variation) and corresponding changes in tissue concentrations on an
annual basis.
XRD8B/A 8-54 2-9-81
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TABLE 8-10. Plants sensitive to heavy metals, arsenic, and boron as accumulated in soils and typical symptoms expressed.
Metal
Plant
Symptoms
Reference
Arsenic Snap bean, lima bean, onion, pea, cucumber, alfalfa, legumes, sweet
corn, strawberry (on light and sandy soils).
Boron Barley, var. Atlas 46; lima bean, var. Henderson; kidney bean, var.
Navel; oats, var. Riverside; onion, var. Cabot; pea, var. Alaska;
peach, var. J.H. Hale; persimmon, var. Kaki; rose, var. Snow White;
soybean, var. Wilson, var O'lootan; wheat, var. Opal; yellow zinnia.
Reduced germination of seeds, rotting of roots, Liebig (1966)
leaf wilt, brown to red coloration of leaves, Linzon (1977)
reduced yield in fruit trees, death.
Yellowing of leaf tips, necrosis between lateral Bradford (1966)
veins and midrib of monocotyledons, marginal leaf Krupa and Kohut
scorch, downward cupping of leaves, reduced flower- (1976)
ing, fruit lesions. Yopp et al. (1974)
Cadmium Red oak, birch, trembling aspen, beet, carrot, celery, green pepper,
lettuce, radish, soybean, Swiss chard, tomato, winter wheat.
Copper Bean, citrus fruits, corn, mustard.
Lead Bean dwarf French, var. Carters; beet, corn, fescue, lettuce,
lupine, lobblolly pine, red maple.
Reduced root elongation, general growth retardation.
Stunted root development, chlorotic leaves, reduced
vegetative growth.
Stunted root growth, shoot retardation, increased
leaf abscission, reduced yields.
Yopp et al. (1974)
Jordan (1975)
Reuther and Laban-
auskas (1966)
Yopp et al. (1974)
Orange (only case known from published literature).
Alfalfa, broadbean, cabbage, cauliflower, cereals, citrus, clover,
lespedeza, pineapple, potato, tobacco, tung, barley, var Atlas 46,
var. Herta; yellow birch, cranberry, peanut, potato, var. Kesweck;
alfalfa, apple, apricot, barley, bean, brussels sprout, carrot,
clover, cotton, lettuce, medic, orange, peas, potato, sugar beet,
vetch, wheat.
Leaf rolling, spiraling inhibition of leaf emer- Embleton (1966)
gence. Narrow leaf development.
Necrotic spots on leaves, necrosis of internal Yopp et al. (1974)
bark,marginal leaf yellowing, incurling of Labanauskas (1966)
leaf margins. ' HAS (1973)
Broadbean, oxalis, sunflower, bean, butterfly weed, cinquefoil,
fern, Hydrangea, Mimosa, Oxalis, privet, sunflower, willow.
Nickel Citrus fruit, alfalfa, oats, var. Victory; pear.
Potas-
sium-
Orange, tung (only case known for published literature).
Possible reduced growth.
Repression of vegetative growth, leaf chlorosis,
white or light yellow and green striping.
Fruit coarseness and leaf necrosis, leaves curl
downward, marginal leaf necrosis, intervenal
chlorosis, plant dieback.
Yopp et al. (1974)
Lagerwerff,(1972)
Jacobson and
Hill (1970)
Vanselow (1966)
Yopp et al. (1974)
Ulrich and
Ohki (1966)
Zinc Oats, orange, tung, barley, var. Trail; corn, var.
Whatley's Pro- Univorm chlorosis, reduced terminal growth, twig
lific, var. Ida Hybrid 330; cowpea, var. Suwannee; wheat, var.
Gaines; barley, citrus, oats, sugarbeet.
Chapman (1966)
dieback, chlorotic striping of leaves, stems stiff Yopp et al. (1974)
and erect.
Source: Adapted from Krupa et al. (1976)
-------
Arsenic sprays have been applied to the foliage of many plants to hasten fruit matura-
tion by causing premature defoliation and chemical changes in the fruit. For example, lead
arsenate sprayed on grapefruit trees caused a "fruit gumming" reminiscent of boron defi-
ciency (Liebig, 1966). Boertitz et al. (1976) reported that arsenic deposited at 22 mg/kg
soil reduced the yield of wheat, rye, winter rape, and red clover by 25, 8, 0, and 6 per-
cent, respectively.
Cadmium—Most biologically active cadmium enters plants through root uptake (Jordan,
1975). Small oxide particles (0.01 to 0.03 urn) may enter leaves through stomata, but it
is thought that the oxides remain largely inert. Cadmium accumulated by apple leaves may
be translocated and incorporated into fruit as they develop (Yopp et al., 1974).
Copper--Wu and Bradshaw (1972) demonstrated a selection of individual plants of
Agrostis stolom'fera growing near metal smelters, thus indicating an indirect effect of
within-species simplification within a population through selection.
Lead—Davis and Barnes (1973) reported reduced growth of loblolly pine and red maple
-4 -3
seedlings in pots of two forest soils treated with 2 x 10 to 2 x 10 M lead chloride.
Lead toxicity symptoms may include fewer and smaller leaves, reduced plant size, leaf
yellowing, and necrosis of elder, sugar beet, squash, and bush bean (Schoenbeck, 1973).
Plants growing in soil already high in these metals tended to be more sensitive to the
addition of metals by air pollution.
Nickel—Plants can absorb and translocate airborne nickel salts (NAS, 1975). Once
inside' the plant, nickel affects photosynthesis and other processes such as stomatal
function (Bazzaz and Govindjee, 1974). In cases of incipient nickel toxicity to vegeta-
tion, no definite symptoms have been observed other than growth repression. In cases of
moderate or acute nickel toxicity, chlorosis-resembling symptoms of iron deficiency is
common (Anderson et al., 1973; Anderson, 1974).
8.7.2 Classification of Plant Sensitivity—Particles
Coarse particulates have not been shown to elicit responses in plants in a manner to
allow plants to be placed into sensitivity classes similar to those developed for gaseous
pollutants. Accumulations of particulate matter such as roadside dusts, cement, quarry
particle emissions, or other forms of deposits such as fly ash, are deposited on all
surfaces and induce responses discussed under the symptom portion of this chapter. How-
ever, heavy metals do elicit differential responses in plants, and therefore it is possi-
ble to develop lists of particularly sensitive plants.
Heavy metals are constituents of many coarse particles emitted from various sources.
To our knowledge, there has not been an organized effort to establish the toxicity of
specific chemical constituents of particulates in relation to sensitivity groups of
vegetation under field conditions. Table 8-10 has listed plants that may be sensitive
to heavy metals following deposition and various symptoms as expressed following their
respective accumulations.
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8.8 DOSE-RESPONSE RELATIONSHIPS—PARTICULATES
Review of the published literature suggests that it is not possible at present to
give even generalized dose-response relationships for the effects of particulate air pollu-
tants on plants. Many reports deal only with gross visible effects or tissue accumulation
of one or more constituents of the particulates. The emphasis of research has been on
settleable coarse particles. Since these are conglomerates of several pollutants, their
chemical constitutions are frequently ill-defined although their sources have often been
identified. Little information could be found on the effects of fine particles on vegeta-
tion.
Where cause-and-effect relationships have been suggested, generally no information is
presented on the actual concentration, particle size, and frequency distributions.
Deposition rates and plant effects vary significantly with particle size. Few studies are
available where two independent scientists have evaluated the effects of particles on vege-
tation with closely comparable physical and chemical properties under reproducible condi-
tions.
Much of the literature refers to particulates from point and line sources, and their
accumulation in or on soils and vegetation. Tissue accumulation of a given element must
be considered as a plant response. Soil scientists have contributed most of the informa-
tion on plant toxicity symptoms that has been obtained under laboratory conditions. Since
many of the plant effects observed are due to the accumulation of elements up to toxic
concentrations, tissue concentrations prior to actual exposure will affect the amount
and elemental composition of particulates that plants can tolerate. Dose-response rela-
tionship are therefore determined and predetermined by the background concentrations of
various elements in the soil where the plant is growing.
Effects of surface accumulation of cement kiln dust on bean leaves have been inves-
tigated by Darley (1966). Doses of 0.6-3.8 g/m per day were applied for 2 or 3 days,
and foliar injury and reductions in carbon dioxide exchange were observed. Reductions
in carbon dioxide exchange of up to 33 percent were noted in the absence of visible
2
foliar injury. Bean leaves dusted with cement kiln dust at the rate of 4.7 g/m /day
for 2 days and then exposed to dew developed leaf rolling and interveinal necrosis
(Lerman and Darley, 1975). Leaves not exposed to dew following the dust treatment
remained asymptomatic.
Reduced yields and injury to leaves and flowers of several plant species were
observed when the plants were exposed once a week for 4 weeks to a dust containing
cadmium, lead, copper, and manganese (Krause and Kaiser, 1977). Yield reductions of
up to 36 percent were noted.
Plants accumulate different elements at differing rates. Tissue concentrations of
some elements (particles) are known to be significantly higher in the vicinity of a
source for those elements (particles) in comparison with background or baseline concen-
trations. This elevated tissue concentration may be due to direct foliar uptake or uptake
XRD8B/A 8-57 2-9-81
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from the pollutant accumulations in the soil. In many cases, elevated tissue concentrations
of a given metal or metalloid are not paralleled by visible injury.
Demonstration of injury symptoms on vegetation under field conditions as a result
of accumulation of metals or metalloids is rare. These demonstrated cases are for such
cases as strip mine wastes. Predicted effects from atmospheric deposition include plant
community changes, chronic long-term physiological changes, and indirect effects through
modification to response to other types of stress. Thus, the state of our knowledge
concerning the effects of particles on vegetation is inadequate at this time and does not
allow the development of accurate dose-response curves.
8.9 INTERACTIVE EFFECTS ON PLANTS WITH THE ENVIRONMENT—PARTICIPATE MATTER
8.9.1 Biotic Interactions
Few studies have exmained the influence of dusts or heavy metal containing particles
on the interactions between organisms capable of causing disease and the predisposition
of the host plant to the disease process.
Infection due to Cercospora spp. increased on sugar beet leaves exposed to cement
dust of 36 percent calcium oxide and 15 percent silica (Schoenbeck, 1960). Increased
occurrence of fungus-induced leaf spots on wild grape and sassafras have been observed
near a source of heavy emissions of limestone dust (Manning, 1971). After examining
40 leaves in each of five locations exposed and not exposed to the dust accumulations,
disease development was two to three and six to seven times greater, respectively, for
the two diseases in the exposed areas.
Natural exposure to combustion nuclei from automobile exhaust which supplied
increased levels of Aitken nuclei and atmospheric lead reduced germination of uredospores
of Puccini a striiformis (stripe rust of wheat); i_n situ development of disease was pro-
longed about 4 days. Similar studies at a non-exposed site did not result in decreased
spore germination (Sharp, 1967, 1972).
8.10 EFFECTS OF SULFUR DIOXIDE AND PARTICULATES ON NATURAL ECOSYSTEMS
The previous sections of this chapter have discussed the effects of sulfur dioxide and
particulate matter on individual plants. This section discusses the effects of these sub-
stances on natural ecosystems since, due to their complexity, these systems respond to
environmental perturbations differently from individuals or populations of organisms.
Ecosystems are basically energy processing systems whose components have evolved together
over a long period of time. They are composed of living organisms together with their physical
environment (see Chapter 7, Table 7-1). The boundaries of the system are determined by the
environmental conditions that determine the kinds of life forms that can exist in a particular
habitat or region. The plant and animal populations within the system are the objects through
which the system functions. Ecosystems respond to environmental changes or perturbations only
through the response of the organisms of which they are composed (Smith, 1980).
XRD8B/A 8-58 2-9-81
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Relationships among the various ecosystem components are structured, not haphazard. The
living (biotic) and the non-living (abiotic) units are linked together by functional inter-
dependence. Processes necessary for the existence of all life, the flow of energy and cycling
of nutrients are based on the functional relationships that exist among the organisms within
the system (Odum, 1971; Smith, 1980; Billings, 1978). Because of these relationships, unique
attributes emerge when ecosystems are studied that are not observable when individuals, popu-
lations or communities are studied. For a more detailed account of ecosystems see Chapter 7,
Section 7.1.2.
The discussion that follows emphasizes the response of terrestrial ecosystems to sulfur
dioxide and particulate matter. Natural ecosystems are seldom, if ever, exposed to a single
air pollutant. Therefore, the responses observed under ambient conditions cannot conclusively
be attributed to a single substance such as sulfur dioxide or particulate matter alone.
8.10.1 Sulfur Dioxide In Terrestrial Ecosystems
Sulfur is an element that is essential for the normal growth and development of plants
and animals. It is a basic constituent of protein and is required in large amounts by some
plants. Under normal circumstances sulfur in rainwater and in soil organic matter is suffi-
cient to meet plant requirements. Excessive sulfur in the form of sulfur dioxide however,
can be toxic to plants. The phytotoxic forms of sulfur, routes of entry into plants, and the
symptomatology of SO- injury to plants have been discussed in the preceding sections.
Within any ecosystem, nutrients sources are in the atmosphere, living and dead organisms,
and available and unavailable salts in the soil and rocks. The nutrients are cycled from the
living to the non-living elements and back again. Air pollution, however, can disrupt nutrient
cycling by altering the amounts in the various compartments and the rate of flow among them.
(Smith, 1980; Chadwick, 1975).
The biogeochemical cycle of sulfur is both sedimentary and gaseous. Sulfur enters the
atmosphere from the combusiton of fossil fuels, volcanic eruptions, the surface of oceans and
gases released by decomposition processes (see Sources and Emissions, Chapter 4). Anaerobic
decomposition of organic matter releases hydrogen sulfide (H^S) into the atmosphere where it
is quickly oxidized into sulfur dioxide. Sulfur dioxide is soluble in water and is carried
back to earth in rainwater as weak sulfuric acid (HLSO.) (Smith, 1980). Regardless of the
source, sulfur in soluble form is taken up by vegetation and is incorporated through a series
of metabolic processes including photosynthesis into sulfur-containing amino acids. Sulfur is
transferred from the producers to the consumers and through excretion and death back to the
soil and to the sediments in the bottoms of ponds, lakes, and seas where bacterial action
releases it as hydrogen sulfide or sulfate. Sulfur in the long-term sedimentary phase is tied
up in organic and inorganic deposits in the soil and sulfur is added to ecosystems through
geological weathering and meteorological processes with the latter being the predominant
source. Weathering and decomposition permit sulfur to enter into solution and to be carried
into aquatic and terrestrial ecosystems. In its gaseous state, sulfur is circulated on a
global scale (Figure 8-9).
XRD8B/A 8-59 2-9-81
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Photochemical
Oxidation
Weathering
of Rocks
Storage of Sulfur or
Sulfur Compounds in
Sediments, Fuels, Soils,
and Sedimentary Rocks
Combustion
of
Sulfur-Containing
Fuels
Figure 8-9. The sulfur cycle. Organic phase shaded.
Source: Clapham (1973).
8-60
-------
*
Based on empirical watershed studies (Likens et al., 1977; Shriner and Henderson, 1978)
and modelling (Coughenour, 1978), the soil is a major reservoir for atmospherically derived
sulfur within in ecosystems. The majority of soil sulfur is unavailable to vegetation and is
organically bound in the humus (May and Downes, 1968). Microbial activity oxidizes organically
bound sulfur to sulfates. Sulfates may be taken up by plants or leached from the soil. The
rate of sulfur released from the organic to the inorganic compartment is the major factor con-
trolling the movement of sulfur between the soil and vegetation (May et al., 1972; Moss, 1976).
A distinction between natural and agroecosystems is that soils in agroecosystems through con-
tinued cropping are depleted of their supply of organic sulfur and it is not renewable; there-
fore, sulfur must be added as fertilizer. Sulfur dioxide brought down in precipitation is also
added to the soil. The amount of sulfur added to soils through precipitation will depend on
the industrial activity of the surrounding area (Kamprath, 1972).
The influence of anthropogenic sources of sulfur on the sulfur cycle is most pertinent
when addressed on a regional basis (Granat et al., 1976) as Shinn and Lynn (1979) have done
for the northeastern United States. Comparing global versus regional sulfur cycling, atmos-
pheric sulfur additions are not equally distributed over the global land areas, and the north-
eastern U.S. experiences anthropogenic atmospheric additions of sulfur that are 28.4 times
that expected if additions were distributed uniformly over the globe. The most notable con-
trast between the global and regional sulfur cycle is the importance of atmospheric sulfur
sources. Globally, natural processes far exceed anthropogenic contributions, whereas in the
northeast man generates 12.5 times the amount of sulfur released by nature. A total of 27 x
r c
10 tons of SO. enters the northeastern regional atmosphere annually and 13 x 10 tons are
deposited within this region by wet and dry deposition; the remaining sulfur is exported to
other areas of the globe.
The conclusion that S0? emitted into the atmosphere through anthropogenic activity is
ultimately transferred to terrestrial and aquatic ecosystems is well documented (Meszaro et
al., 1978). Unfortunately, the fate of sulfur in the ecosystem after deposition is not fully
resolved. The issue is critical since ecosystems subject to excess nutrients or toxic materi-
als do not commonly distribute them uniformly throughout the system but rather preferentially
sequester them in specific pools or compartments. In addition, sulfur dioxide as a gas can
cause injury to the vegetative components of an ecosystem so that energy flow and the cycling
of other nutrients as well as sulfur may be disrupted.
Within the U.S. man is a major source of atmospheric sulfur and in the northeast alone
anthropogenic sources exceed all others by a factor of 12.5. Within this region S02 levels
annually average 16 ug/m (Shinn and Lynn, 1979) which is several times that recorded in
pristine areas. The immediate fate of approximately 60 percent of this atmospheric sulfur is
deposition on terrestrial and aquatic ecosystems; however, the subsequent fate of sulfur within
the ecosystem is not fully known. Experimental evidence from forested watersheds coupled with
results from simulation models indicates that sulfur in ecosystems is highly mobile. Although
sulfur levels in the soil and vegetation compartments, in aggrading and mature ecosystems
XRD8B/A 8-61 2-9-81
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impacted by SCL increase with time the majority of sulfur deposited annually is exported out
of the system in stream flow.
8.10.2 Ecosystem Response to Sulfur Dioxide
The Kaybob gas plants (Fox Creek, Alberta, Canada), which emit S02 during the removal of
hydrogen sulfide from natural gas, are located within transition montane-boreal forest domi-
nated by a mixed assemblage of deciduous and coniferous trees; however, white spruce (Picea
glauca) stands predominate (Winner and Bewley, 1978a,b; Winner et al., 1978). Since these white
spruce forests have less species variation than other sites, they were selected for analysis
along a transect showing decreasing SCL stress. The facility began operation in 1968, and the
field study was completed in 1976. Field measurements of ambient atmospheric conditions were
not made although the nature of the technological process would necessitate S0? emission pro-
ducts (Legge et al., 1976). From 1973 to 1975, it is estimated that the Kaybob facility
emitted approximately 71,000 kg/day of S0?.
Relative species diversity showed no gradient pattern of response to SCL; however, per-
cent coverage for all understory plants, including vascular species and mosses, showed a marked
increase with distance from the source (Winner and Bewley, 1978b). Numbers of white spruce
seedlings close to the refinery were reduced. Changes in moss communities were conspicuous
and included decreasing values for moss canopy coverage, moss carpet depth, dry weight, capsule
number, and frequency of physiologically active versus inactive moss plants. Close to the
source, there were no mosses at all (Winner and Bewley, 1978a). These results indicate that
species diversity, particularly in the mosses, has changed as a consequence of sulfur gas emis-
sions.
In subsequent study of the ecological fate of sulfur emissions in the same white spruce
forests, Winner et al. (1978) found an association between sulfur accumulation in foliage and
the vertical location of the organism in the forest's stratification. Specifically, the
sensitivity of mosses versus understory and canopy species was attributed to the greater sulfur
accumulation derived solely from sulfur emitted from the Kaybob facility. This conclusion was
based on tracer experiments ( S: S) which assessed the fate of emitted sulfur (Krouse, 1977).
This condition of enhanced moss sensitivity has ecological consequences since mosses serve
several unique functions within a forest. These included water storage, soil formation, eoro-
sion prevention, and nutrient cycling.
A second study of the effects of chronic sulfur gas emissions on a forested ecosystem
were conducted in a region experiencing effluents from a natural gas-processing facility (West
Whitecourt Gas Plant, Whitecourt, Alberta, Canada)(Legge et al., 1976). Ecological observa-
tions were coupled with physiological investigations, the goal being to explain changes in
ecosystem structure and function by alterations in physiology. The processing plant began
operation in 1961, and field studies were conducted in 1975 within a 120-year-old lodgepole x
jackpine forest (P. contorta x banksiania) characterized by a bearberry-blueberry understory.
Results of an air-monitoring program in 1975 within the pine forest showed S09 levels to be
variable but not exceeding the Alberta air quality criteria standard of 0.2 ppm/0.5 hr/24 hr.
XRD8B/A 8-62 2-9-81
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Measurable quantities of S02 (0.01 ppm detection level of the instrument) were recorded on 40
out of 46 days. On an average day, detectable concentrations of S02 remained in the forest
for 4 hours; however, over 46 percent of all hours monitored were characterized by pollutant
concentrations below 0.05 ppm. The highest instantaneous level was 0.45 ppm; pollution
concentrations were greater during daylight hours. These data depicting the ambient S0?
exposure for the forest are specific for 1975.
Two locations were selected based upon their similarity in many edaphic, climate, and
biological indices, but differing in their proximity to the source. Studies of structural
changes induced by S02 at the ecosystem level were limited to studies of basal-area increments
and biomass of the dominant pines in both a reference and S0?-stressed location. Using basal-
area increments as an example (Legge et al., 1976), trees nearer the source showed smaller
incremental additions, beginning in 1964 and continuing through 1975; statistically signifi-
cant differences were obtained in 1971 through 1975. In addition to these data, tracer tech-
niques ( S: S) coupled with atmospheric profile, micrometeorological studies showed that the
pine canopy was acting as a sink for the sulfur emitted from the processing facility.
The movement of materials within the forest also showed signs of S0»-induced modifica-
tions (Legge et al., 1976). This phase of the study compared indices of changes in a series
of sites selected for their similarity in vegetation and soil properties but varying in their
exposure to S0?. As expected, sulfur distribution in the soils and vegetation varied with
distance and direction from the source. In lodgepole x jackpines, sulfate levels in leaves in
three successive years decreased with increasing distance along the corridor of S0? stress.
Seasonal variation in sulfate in the leaves of tamarisk (Larix larincinia), aspen (Populus
tremuloides), and lodgepole x jackpine was recorded, with progressive sulfur loading from June
through September annually.
The status of soil sulfur also reflected the direction and distance of the sample site
from the source. Total soil sulfur generally decreased with distance along the downwind
corridor of S0? stress. Along the corridor, consistently higher sulfur levels were found in
the upper 2 cm of the soil, which also had the highest organic content; higher soil acidity
was recorded proximal to the source of S0?.
Pollutant-incuded modifications of the biogeochemical cycles were not restricted to
sulfur alone but were also evident in other inorganic elements. The most striking and consis-
tent pattern was recorded with manganese distribution. Over four successive years, the manga-
nese content of lodgpole x jackpine needles decreased with increasing distance away from the
source. The potential importance of this effect lies in the fact that elevated manganese
levels elicit iron deficiencies which, in turn, are known to affect the regeneration of pine
forest. The distribution of other elements along the corridor also exhibited patterns associ-
ated with the level of S0? exposure: lodgepole x jackpine foliar concentrations of potassium,
zinc, phosphorus, and iron were consistently lower in S02-exposed versus reference locations.
In the near future, the grasslands of the upper plains will be subject to S02 emissions
from new coal-burning power facilities that are being constructed in areas rich in coal
XRD8B/A 8-63 2-9-81
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reserves (Durran et al., 1979). To address this problem, plots of Montana grasslands were
exposed to SO, during growing seasons of successive years. The monthly median exposure levels
3 ^ 3
were approximately 0, 0.02 (52 ug/m ), 0.04 (106 ug/m ), and 0.07 ppm (185 ug/m ) S02 and were
delivered by a zonal air pollution system or ZAPS (Lee et al., 1978). Field observations over
four years verified that these concentrations were not sufficient to elicit any leaf lesions
characteristic of acute S02 (Heitschmidt et al., 1978). Ambient pollutant concentrations were
typically greater at night, and the concentration decreased rapidly from the interface of tur-
bulent air and grass canopy downward to the soil (Preston, 1979).
The most prevalent producer species within the grassland is a perennial, Agropyron
smithii. In populations sampled over the growing season in each of the exposure regimes, SO-
induced a variety of changes in biochemical indices of plant performance. Monthly samples of
tillers and leaves showed a positive correlation of foliar sulfur with time of exposure and
canopy-level SO- concentrations (Lauenroth et al., 1980). This relationship was most conspi-
cuous with the two higher exposure regimens, and total foliar sulfur in the highest exposure
plot was three times greater than that in vegetation sampled from control locations (Lauren-
roth and Heasley, 1980). As the sulfur content of leaf tissue increased, the ratio of nitro-
gen to sulfur decreased (Laurenroth and Heasley, 1980).
These biochemical changes in the major producer species were mirrored by other modifica-
tions in plant performance. In A. smithii populations exposed to 52 ug/m S0? over the grow-
ing season, the functional leaf life (the period of active photosynthesis) was increased by
several weeks, while the same index of plant performance was shortened by two weeks at 105 and
185 ug/m S0? (Laurenroth and Heasley, 1980). Parallel increases and decreases in chlorophyll
content at the low and high S0? levels, respectively, were also recorded. Finally, with
increasing S0? exposure, plants stored less photosynthate in rhizome tissue (Heitschmidt et
al., 1978).
Dominant producers were not the only flora exhibiting sensitivity to S0?. In simulated
pollutant exposure using a bisulfate solution, Sheridan (1979) showed that nitrogenase activity
in a major component of the lichen flora (Pollema tenex) was reduced. Although the applica-
bility of the data must be validated through field studies, the potential for such an effect
must be recognized, particularly in the light of the importance of soil lichens in regulating
nitrogen fixation in the grasslands (Sheridan, 1979).
Further evidence of S0?-associated effects on grasslands is recorded in both consumer and
decomposer populations. The density of grasshoppers, a major consumer of P±_ smithii foliage,
decreased with increasing SOp stress in two successive seasons (Laurenroth and Heasley, 1980).
Decomposition rates were apparently also altered with less litter disappearance in 50,-exposed
plots. The mechanism involves a direct pollutant effect on decomposer activity rather than an
indirect effect, such as elevated sulfur levels in the litter (Laurenroth and Heasley, 1980).
Larger consumers also exhibited responses reflecting the presence of S07 in the atmos-
phere; however, the responses were not dose dependent (Chilgren, 1978). Peromyscus
maniculatus, prairie deer mouse, is a common and active vertebrate in grassland communities.
XRD8B/A 8-64 2-9-81
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*
Over one exposure season, the frequency of P^ manlculatus in control plots increased, impli-
cating an S02-induced behavioral response (habital preference) whereby individuals seek habi-
tats free of the pollutant.
In summary, at S02 levels above 0.02 ppm (52 ug/m3), S02 induced changes in the perform-
ance of producers, consumer, and decomposers. Many of the responses are individually small,
but collectively over time they are gradually modifying the structure and function of the
grasslands. The significance of these changes to the long-term persistence of the ecosystem
remains controversial (Preston, 1979).
The results of these studies, particularly the West Whitecourt and Montana grasslands
studies, document the usefulness of addressing ecosystem-level responses to S02 from a multi-
disciplinary approach incorporating investigations of physiology, autecology, synecology, geo-
chemistry, meteorology, and modeling. The results confirm that producers are sensitive to
direct S02 effects as evidenced by S02-associated changes in cell biochemistry, physiology,
growth, development, survival, fecundity, and community composition. Such responses are not
unexpected. An equally important point of agreement among the different research efforts is
the potential for ecological modification resulting from either direct SO- effects on nonpro-
ducer species or direct changes in habitat parameters, which in turn affect an organism's per-
formance. Changes in biogeochernistry, particularly in the soil compartment, are notably
responsive to chronic SO- exposure.
The influence of prolonged S0? exposures on plant communities is not well documented;
however, a theoroetcal basis from which to evaluate effects is emerging. This conceptual
effort is exemplified by the development of a generalized forest growth model (Botkin et al.,
1972) that was designed to assess the consequences of perturbation on plant communities
(Botkin, 1976; Shugart and West, 1977). This model has been applied to the response of a
mixed species, deciduous forest in the southeastern U.S. to differential levels of growth
reduction (0, 10, and 20 percent) following simulated air pollution stress (West et al.,
1980). Over several decades, simulated pollution stress altered the biomass importance of the
major tree species within the forest; some species populations increased while others decreased
in importance. These results suggest that competitive interactions between species may signi-
ficantly modify both the level and direction of change in growth rate of individual species in
response to air pollution stress. Stand age was also shown to strongly influence the role of
competition in modifying responses of individual species within the forest community. Since
community composition is determined in part by species interactions (e.g., competition, symbi-
osis), the ecological importance of resistant species, whose prominence in the community is
determined by interaction with sensitive species, can be expected to be enhanced under stresses
such as air pollution which does not affect all species equally (West et al., 1980). An under-
standing of this governing role of species interactions is essential to predicting how eco-
systems respond to low levels of pollution (Botkin, 1976). This is also the justification for
not extrapolating freely the results from intensely managed forest and agroecosystems to
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predict how a mixed species community (e.g., natural forests or grasslands) will respond to a
comparable perturbation (Miller and McBride, 1975; Kickert and Miller, 1979).
The results from community level studies in areas experiencing chronic levels of SC^ lend
credence to the modelling effort. Using communities comprised of only 2 or 3 different
species, Guderian (1967) analyzed community level responses to S02 and their underlying causes.
Changes in community composition were a function of pollutant dose; the higher the does the
more rapidly the community changed. Altered community composition was attributed to both
direct SO^ effects on sensitive species' populations and indirect changes in species' inter-
actions. Community biomass exhibited little quantitative change but striking differences in
species composition. Similar conclusions have been reached in studies of natural plant
communities experiencing prolonged S0? exposure (Guderian and Stratmann, 1966; Rosenberg et
al., 1979). Rosenberg et al. (1979) assessed species composition in 27 stands of natural
regrowth, Northern hardwood forest dominated by oaks (Quercus spp.), white pine (P. strobus),
and hemlock (Tsuga canadensis). The stands, which varied in their distance from a 25-year-old
coal-consuming power plant, exhibited no obvious a priori compositional differences. Atmos-
pheric pollutant levels were not reported although foliar symptoms typical of S0? toxicity
were recorded on several occasions. In both upwind and downwind directions, the number of
vascular plants (canopy, understory and ground) per unit area (species richness) increased
with distance from the source. A similar distance dependent response was recorded for species
diversity (Shannor-Weiner). In spite of these S0?-induced changes in community composition,
an index of aboveground biomass (basal area of overstory species) exhibited no variation among
stands. Among the vascular plants, shrub and ground vegetation was more responsive (diversity
and richness) than the overstory to S0? stress, and this enhanced susceptibility of the lower
strata was attributed to the intense competition, unique population biology attributes, and
microhabitat factors that tend to increase S0? levels close to the ground.
Some of the most notable examples of SCL affecting plant communities are the responses of
cryptogamic flora (lichens and mosses), and several reviews are available (DeSloover and
LeBlanc, 1968; Hawksworth, 1971). A map of epiphytic lichen communities for England and Wales
has been devised which associates progressive shifts in species composition with S0? levels
(Hawksworth and Rose, 1970). In general, the higher ambient S02 levels were consistently
associated with fewer species and an increasing relative frequency of crustose versus foliose
or fruticose forms. The fidelity with which community composition changes in accordance with
S0? has led to the suggestion that analyses of lichen communities be used as a bioassay to
estimate ambient SCL levels.
Similar mapping efforts are reported for several regions of North America. In a rural
area of Ohio surrounding a coal-consuming power station (emitting 1025 tons S0?/day), the
distribution of two corticolose lichens, Parmelia caperata and P. ruderta, was markedly
affected by elevated S02 levels (Showman, 1975). In regions experiencing an annual S0?
average exceeding 0.020 ppm, both species were absent. The distribution of more resistant
lichens was not noticeably affected until SO^ levels exceeded 0.025 ppm (annual average).
XRD8B/A 8-66 2-9-81
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*
Somewhat lower levels were projected by LeBlanc and Rao (1973) to affect the ability of sensi-
tive lichen species to survive and reproduce; acute and chronic symptoms of S0? toxicity in
epiphytic lichens occurred when annual averages of S02 exceeded 0.03 and 0.006-0.03 ppm,
respectively.
The susceptibility of crytogamic flora to elevated levels of S0? may influence the move-
ment of materials within the ecosystem. In the northwestern coniferous forests, lichens fix
2-11 kg/ha of nitrogen, which represents 5-20 percent of the total nitrogen requirement for
the dominant producer, Douglas fir (Denison, 1973).
The network of biotic-abiotic interactions, which is characteristic of managed and
natural ecosystems, leads to the hypothesis that S02 effects on producers must have reper-
cussions to other trophic levels. Demonstration of such responses, however, is difficult
experimentally, and an accurate assessment of the specific importance of SO- in eliciting
these responses is complicated by the often complex relationships between producers, con-
sumers, and decomposers.
Consumer and decomposers may respond to SOp via a direct, adverse effect of the
pollutant. The presence of elevated atmospheric levels of S0? is particularly relevant to
soil organisms (Babich and Stotzky, 1979) since soils are preferential sulfur accumulation
sites. This focus on soil-borne organisms takes on even more relevance since the rhizosphere
is not only biologically active but also the major site for sulfur accumulation within the
ecosystem (Legge et al., 1976). In a forested area experiencing atmospheric S02 levels
averaging 0.048 ppm, the species composition of soil microflora shifted toward a greater
number and frequency of species capable of utilizing the soil sulfur additions (Wainwright,
1979). Specifically, the levels of thiobacilli and sulfur-oxidizing fungi were positively
correlated with levels of S0? stress and soil depth.
The edaphic and climatic environment strongly influences the community of plants, animals
and microorganisms that develop at a given site. In natural ecosystems in sulfur deficient
soils, communities have evolved within the constraints imposed by a limited supply of sulfur.
Although atmospherically derived sulfur may not be sufficient to cause injury, the prolonged
input of sulfur may relax the constraints of a limited sulfur supply thereby inducing shifts
in species composition.
8.10.3 Response of Natural Ecosystems to Particulate Matter
Particulate matter originating from both natural and anthropogenic emission sources is a
common component of the atmosphere. As discussed in Section 8.7, the heterogeneous physical
and chemical nature of particulates presents problems in addressing the significance of ele-
vated atmospheric particulate levels for natural and agroecosystems.
Wet and dry deposition are the two processes by which particulates are transferred from
the atmosphere to terrestrial ecosystems. The fate of particulates deposited on foliar sur-
faces depends on the solubility of the constituents (chemicals and elements), the occurrence
of precipitation, and the sorptive capacity of the leaf (Little, 1973). Furthermore, many
elements commonly associated with particulates are essential for plant metabolism (e.g., zinc
XRD8B/A 8-67 2-9-81
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*
and phosphorus), and as a consequence, absorption may be a means by which the plant can
supplement its nutrient supply (Klake, 1974). Leaf surfaces may be only a transitory site for
particulate matter and its associated constituents. If not retained by the leaf, material is
ultimately transferred to the" forest floor through washoff in rain events or litterfall. The
net effect of these processes is to funnel leaf surface deposits to the litter-soil complex.
This conclusion is verified for many atmospherically derived heavy metals deposited in natural
ecosystems (e.g., Coughtrey et al., 1979; Thompson et al., 1979).
Given the regional character of particulate emissions, particularly along the east coast
(National Air Quality, Monitoring and Emissions Trend Report, 1978), the fate of particulates
in terrestrial ecosystems experiencing low levels particulate pollution needs to be assessed.
In a deciduous forest in the Southeast, wet and dry deposition of aerosols, gases, precipita-
tion and large particulates were major sources of trace element input to the forest floor,
including 99 percent for lead, 44 percent for zinc, 42 percent for cadmium, 39 percent for
sulfate and 14 percent for manganese (Linberg et al., 1979). These seemingly large percent-
ages are typical for rural or remote areas even though 3 major coal consuming power facilities
(total coal consumption of 7 x 10 tons/year) were within 20 km of the forest.
Irrespective of the source of particulates deposited in the forest, the atmosphere con-
tributed a major portion of the trace element inputs (Lindberg et al., 1979). Water solu-
bility was critical since insoluble constituents associated with the particulate were not
readily mobilized within the forest. Any event promoting solubilization (e.g., aerosol forma-
tion, rainfall scavenging, moisture formation on leaves) enhanced an element's mobility.
The leaf surface is not the only accumulation site for particulates and their associated
constituents within the ecosystem. Through precipitation scavenging of particulates in air,
washoff of surface deposits or litterfall, particulates are transferred to the soil where they
are tightly bound to decaying organic matter. The upper soil horizon, including the decaying
organic material, is a region of intense biological activity as a result of the physical
degradation of litter, remineralization of the bound materials and root uptake of the plant-
available nutrients. Consequently, particulate emissions that interfere with microbial
activity can have delayed effects on primary production (Tyler, 1972) and soil consumer
species.
In summary, even though the impact of particulate matter on terrestrial ecosystems is
most apparent proximal to large emission sources, ecosystems within the same geographic region
may be the site of deposition. Foliar surfaces are the most common site for initial dry and
wet deposition; however, most material is eventually transferred to the soil. Particulate
matter alone constitutes an ecological problem only where deposition rates are high. However,
concern for terrestrial ecosystems must also address elements and chemicals that may be asso-
ciated with the particulates. Solubility of these particulate constituents is a critical
factor since insolubility limits mobility within the ecosystem. One common behavior of parti-
culates is their tendency to selectively accumulate within a given component of the landscape.
Soils are long-term sites for the retention of many constituents found in particulates. While
XRD8B/A 8-68 2-9-81
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*
this accumulation in the soil-litter layer has had demonstrable adverse consequences for eco-
logical processes such as decomposition, mineralization, nutrient cycling and primary produc-
tion around some point sources, the effects of the much lower levels chronically deposited
over large regions have not yet produced documented adverse impacts to natural ecosystems.
8.11 SUMMARY
The widespread occurrence of particulate matter and sulfur dioxide in the atmosphere
results frequently in terrestrial vegetation being exposed simultaneously to these two pollut-
ants and other phytotoxic pollutants. More is known about the effects of sulfur dioxide than
about the effects of particulate matter on plant life. Studies of the effects of particulate
matter have generally focused on the effects of heavy accumulations and the reduction in the
photosynthesis resulting from these accumulations. The more subtle effects of particulate
matter on vegetation have not been extensively investigated and are, therefore, not well
understood. Even less information is available concerning plant response following exposure
to sulfur oxides and particulate matter in combination.
Sulfur dioxide must enter a plant through leaf openings termed stomata to cause injury.
Sulfur dioxide after entering plant cells from the stomata is converted to sulfite and bisul-
fite, which may then be oxidized to sulfate. Sulfate is about 30 times less toxic than sulfite
and bisulfite. As long as the absorption rate of S02 in plants does not exceed the rate of
conversion to sulfate, the only effects of exposure may be changes in opening or closing of
stomata or insignificant changes in the biochemical or physiological systems. Such effects
may abate if S0? concentrations are reduced. Both negative and positive influences on crop
productivity have been noted following low-dose exposures.
Symptoms of S0?-induced injury in higher plants may be quite variable since response is
governed by pollutant dose (concentration x duration of exposure), kinetics of the exposure
(e.g., day vs night, peak y_s long-term); physiological status of the plant, phenological stage
of plant growth, environmental influences on the pollutant/plant interaction, and the environ-
mental influences on the metabolic status of the plant itself.
There are several possible plant responses to SCL and related sulfur compounds: (1) fer-
tilizer effects appearing as increased growth and yields, (2) no detectable responses, (3) in-
jury manifested as growth and yield reductions without visible symptom expressions on the
foliage or with very mild foliar symptoms that would be difficult to perceive as air pollution
incited without the presence of a control set of plants grown in pollution-free conditions,
(4) injury exhibited as chronic or acute symptoms on foliage with or without associated reduc-
tion in growth and yield, and (5) death of plants and plant communities.
A member of species of plants are sensitive to low levels of SO,,. Some of these plants
may serve as bioindicators in the vicinity of major sources of S02- Even these sensitive spe-
cies may be asymptomatic, however, depending on the environmental conditions before, during,
and after S02 exposure. Various species of lichens appear to be among the most sensitive plants.
As the dose of S02 increases, plants develop more predictable and more obvious visible
symptoms. Foliar symptoms advance from chlorosis, or other types of pigmentation changes, to
XRD8B/A 8-69 2-9-81
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necrotic areas, and the extent of necrosis increases with exposure. Studies of the effects of
SCL on growth and yield have demonstrated a reduction in the dry weight of foliage, shoots,
roots, and seeds, as well as a reduction in the number of seeds. At still higher doses,
reductions in growth and yield increase. Extensive mortality has been noted in forests
continuously exposed to SOp for many years.
The amount of sulfur accumulated from the atmosphere by leaf tissues is influenced by the
amount of sulfur in soil relative to the sulfur requirement of the plant. After low-dose
exposure to SO-, plants grown in sulfur deficient soils have exhibited increased productivity.
Sulfur dioxide and particulate sulfate are the main forms of sulfur in the atmosphere,
and a plant may be exposed to these pollutants in several different ways. Dry deposition of
particulate matter and wet deposition of gases and particles bring sulfur compounds into con-
tact with plant surfaces and soil substrates. The effects of such exposures are more diffi-
cult to assess than those associated with the entry of SO,, through plant stomata. Plant
response to dynamic physical factors such as light, leaf surface moisture, relative humidity,
and soil moisture may influence pollutant uptake through internal physiological changes as
well as stomatal opening and closing and hence play a major role in determining sensitivities
of species and cultivars or the time of sensitivity of each on a seasonal basis. Dose-reponse
relationships are significantly conditioned by environmental conditions before, during, and
following exposure to S0?.
Analysis of existing data sets relating SCL exposure concentration to both visible injury
of foliage and reduced growth and yield indicate that both thresholds of reponse and general-
ized average reponse can presently be described mathematically. For visible injury, the
threshold varied with duration of exposure ranging from 0.4 ppm for 1 hr (0.4 ppm/hr) to 0.15
ppm for 8 hr (1.2 ppm/hr). The response threshold was 0.5 ppm for 3 hr (1.5 ppm/hr). For
plant species growing in the vicinity of a coal-fired power plant approximately 10 percent of
the species visibly injured by S0? experienced that injury at peak, 1-hr, and 3-hr average
concentrations of 1.00, 0.45, and 0.28 ppm, respectively.
Since, under many conditions, visible injury may not lead directly to measurable losses
in plant productivity, correlations of productivity with exposure dose represent the most
direct link to the effects of S02 on economic and ecological parameters of plant response.
Regression analysis of data from both controlled exposures in field chambers and from labora-
tory and greenhouse studies showed positive and statistically significant correlations between
degree of yield loss and logarithmically transformed exposure dose in ppm/hr. The plants and
conditions utilized in field studies, which were heavily oriented toward crop plants, provided
generally lower yield losses for the same exposure dose than did laboratory and greenhouse
studies. In field studies a 20 percent yield loss was associated with approximately 27 ppm/hr
of accumulative exposure dose. In laboratory studies with agronomic, ornamental, and native
species a 20 percent yield loss occurred after only 9 ppm/hr of cumulative exposure dose.
A critical need in evaluating the likelihood of adverse effects occurring in association
with longer-term SOp exposures is the indentification of the fraction of the total SO exposure
XRD8B/A 8-70 2-9-81
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which may constitute a stress to plant growth and development. A review of dose-response data
indicates that this level may be approximately 0.05 ppm for sensitive species. Additional data
from studies involving SO^ alone and in combination with other pollutants may provide a more
accurate basis for determining this level.
At present, data concerning the interactions of SO- with other pollutants indicates that,
on a regional scale, SO^ occurs at least intermittently at concentrations high enough to pro-
duce significant interactions with other pollutants, principally Or A major weakness in the
approach to pollutant interactions, however, is the lack of in-depth analysis of existing
regional air quality data sets for the three principal pollutants (S02, 03, and N0?). These
data should determine how frequently and at what concentrations the pollutants occur together
both spatially and temporally within regions of major concern. The relative significance of
simultaneous versus sequential occurrence of these pollutants to effects on vegetation is also
not well documented and is critical in evaluating the likelihood and extent of potential pol-
lutant interactions under field conditions.
A few studies have reported that combinations of particulate matter and SO-, or particu-
late matter and other pollutants, increase foliar uptake of S0?, increase foliar injury of
vegetation by heavy metals, and reduce growth and yield. Because of the complex nature of
particulate pollutants, conventional methods for assessing pollutant injury to vegetation, such
as dose-response relationships, are poorly developed. Studies have generally reported vegeta-
tional responses relative to a given source and the physical size or chemical composition of
the particles. For the most part, studies have not focused on effects associated with specific
ambient concentrations. Coarse particles such as dust directly deposited on the leaf surfaces
result in reduced gas exchange, increased leaf surface temperature, reduced photosynthesis,
chlorosis, reduced growth, and leaf necrosis. Heavy metals deposited either on leaf surfaces
or on the soil and subsequently taken up by the plant can result in the accumulation of toxic
concentrations of the metals within the tissue.
Natural ecosystems are integral to the maintenance of the biosphere and disturbances of
stable ecosystems may have long-range effects which are difficult to predict. Within the
United States anthropogenic contributions to atmospheric sulfur exceed natural sources. In
the Northeast these contributions exceed natural sources by a factor of 12.5 and approximately
60% of the anthropogenic emissions into the atmosphere are deposited (wet and dry deposition)
on terrestrial and aquatic ecosystems; its subsequent fate and distribution in these systems
is not well understood. Wet deposition of sulfur compounds is discussed in Chapter 7.
Data relating ecosystem responses to specific doses of SO,, and other pollutants are
difficult to obtain and interpret because of the generally longer periods of time over which
these responses occur and because of the many biotic and abiotic factors which modify them.
Vegetation within terrestrial ecosystems is sensitive to S02 toxicity, as evidenced by
changes in physiology, growth, development, survival, reproductive potential and community
composition. Indirect effects may occur as a result of habitat modification through influences
on litter decomposition and nutrient cycling or through altered community structure. At
XRD8B/A 8-71 2-9-81
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the community level chronic exposure to SCL, particularly in combination with other pollutants
such as CU, may cause shifts in community composition as evidenced by elimination of
individuals or populations sensitive to the pollutant. Differential effects on individual
species within a community can occur through direct effects on sensitive species and through
alteration of the relative competitive potential of species within the plant community.
Particulate emissions have their greatest impact on terrestrial ecosystems near large
emission sources. Particulate matter in itself constitutes a problem only in those few areas
where deposition rates are very high. Ecological modification may occur if the particles con-
tain toxic elements, even though deposition rates are moderate. Solubility of particle con-
stitutents is critical, since water-insoluble elements are not mobile within the ecosystem.
Most of the material deposited by wet and dry deposition on foliar surfaces in vegetated areas
is transferred to the soil where accumulation in the litter layer occurs.
XRD8B/A 8-72 2-9-81
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8.12 REFERENCES
Adepipe, N. 0., R. E. Barrett, and D. P. Ormrod. Phototoxicity and growth response of
ornamental bedding plants to ozone and sulfur dioxide. J. Am. Soc. Hortic. Sci
97:341-345, 1972.
American Phytopath. Soc. Glossary of air pollution terms and selected reference list
Phytopathology News 8:5-8, 1974.
Anderson, P. J. The effect of dust from cement mills on the setting of fruits. Plant World
17:57-68, 1914.
Anderson, A. J., D. R. Meyer, and F. K, Mayer. Heavy metal toxicities: levels of nickel,
cobalt, and chromium in the soil and plants associated with visual symptoms and variation
in growth of an oat corp. Aust. J. Agric. Res. 24:557-571, 1973.
Anderson, T. J. The effect of dust from cement mills on setting of fruits. Plant World
17:57-68, 1974.
Ashenden, T. W. , and T. A. Mansfield. Influence of wind speed on the sensitivity of ryegrass
to S02. J. Exp. Bot. 28:729-735, 1977.
Ashenden, T. W. Growth reductions in Cocksfoot (Dactylis golmerata L.) as a result of S09
pollution. Environ. Pollut. 15:161-166, 1978.
Ashenden, T. W. The effects of long-term exposures to S0? and N02 pollution on the growth of
Dactyl is golmerata L. and Poa pratensis L. Environ. Pollut. 18:249-258, 1979.
Babich, H. , and G. Stotzky. Air pollution and microbial ecology. CRC Crit. Rev. Environ.
Control 4:353-421, 1974.
Ballentyne, D. J. Sulfite inhibition of ATP formation in plant mitrochondria. Phytochemistry
12:1207-1209, 1973.
Bazzaz, M. B. , and Govindjee. Effects of lead chloride on chloroplast reactions. Environ.
Lett. 6:175-191, 1974.
Bell, J. N. B. , and W. S. Clough. Depression of yield in'ryegrass exposed to sulfur dioxide.
Nature 241:47-49, 1973.
Bennett, J. H. , and A. C. Hill. Inhibition of apparent phytosynthesis by air pollutants. J.
Environ. Qual. 2:526-530, 1973a.
Bennett, J. H. , A. C. Hill, A. Soleimani, and W. H. Edwards. Acute effects of combination of
sulfur dioxide and nitrogen dioxide on plants. Environ. Pollut. 9:127-132, 1975.
Berigari, M. S. , C. F. Jordan, and C. A. Feickert. The Effect of Sulfur Dioxide on Yield and
Growth of Kidney Beans (Phaseolus vulgaris L.). Argonne Nat. Lab. Radiol. Environ. Res.
Div. Annu. Rep. (Ecol.), January-December 1974. pp. 51-61.
Berry, C. R. Relative sensitivity of red, jack, and white pine seedlings to ozone and sulfur
dioxide. Phytopathology 61:231-232, 1972.
Billings, W. D. Plants and the Ecosystem. 3rd ed. pp. 1-62. Wadsworth Publishing Company,
Inc. Belmont, PA., 1978. 177 p.
SOX8B/B 8-73 2-10-81
-------
Black, V. J., and M. H. Unsworth. Effects of low concentrations of sulphur dioxide on net
photosynthesis and dark respiration of Vicia faba. J. Exp. Bot. 30:473-483, 1979.
Boertitz, S. , H. G. Daessler, and E. Friedrich. Effect of metallurgical dust on agricultural
plants. Tech. Umweltschutz 15:247-254, 1976.
Bonte, J. , L. DeCormis, and P. Louguet. Effect of sulfur dioxide pollution on the degree of
opening of Pelargonium x hortorum stomata. C. R. Hedb. Sciences Acad. Sci. Ser. D
280:2377-2380, 1975.
Botkin, D. B. , J. f. Janak, and J. R. Wallis. Some ecological consequences of a computer
model of forest growth. J. Ecol. 60:849-872, 1972.
Botkin, D. B. The role of species interactions in the response of a forest ecosystem to
environmental perturbation. In: Systems Analysis and Simulation in Ecology, vol. IV.
B. C. Patten, ed., Academic Press, New York, NY, 1976. pp. 147-173.
Bradford, G. R. Boron. Iri: Diagnostic Criteria for Plants and Soils. H. G. Chapman, ed.,
University of California, 1966. pp. 33-61.
Brady, N. C. The Nature and Properties of Soils, 8th edition. MacMillan Publishing Co., New
York, NY, 1974. 639 pp.
Brandt, C. S., and W. W. Heck. Effects of Air Pollutants on Vegetation. _In: Air Pollution,
vol. I: Air Pollution and Its Effects. A. C. Stern, ed., Academic Press, New York, NY,
1968. pp. 401-443.
Brandt, C. J. , and R. W. Rhoades. Effects of limestone dust accumulation on composition of a
forest community. Environ. Pollut. 3:217-225, 1972.
Brandt, C. J., and R. W. Rhoades. Effects of limestone dust accumulation on lateral growth of
forest trees. Environ. Pollut. 4:207-213, 1973.
Buron, A., and G. Cornic. Effect of sulfur dioxide on gas exchange in white mustard (Sinapsis
alba). Bull. Soc. Vaudoise Sci. Natur. 71:451-461, 1973.
Carlson, R. W. Reduction in the photosynthetic rate of Acer Quercus and Fraxinus species
caused by sulphur dioxide and ozone. Environ. Pollut. 18:159-170, 1979.
Chadwick, 1975.
Chamberlain, A. C. The Movement of Particles in Plant Communities. In: Vegetation and the
Atmosphere, Vol. 1: Principles. J. L. Monteith, ed. , Academic Press, London, England,
1975. pp. 155-203. -
Chiba, 0. , and V. Tanaka. The effect of sulphur dioxide on the development of pine needle
blight caused by Rhizosphaera kalkhoffii Bubak (I). J. Jap. For. Soc. 50:135, 1968.
Chilgren, J. D. The response of prairie deer mice to a field S02 gradient. In: Proceedings
of the 4th Joint Conf. Sensing of Environmental Pollutants. American Chemical Society,
Washington, DC, 1978. pp. 61-65.
Clapham, W. B. J. Natural Ecosystems. The Macmillan Company, N.Y., 1973. p. 43.
Clough, W. S. The deposition of particles on moss and grass surfaces Atmos Environ
9:1113-1119, 1975.
SOX8B/B 8-74 2-10-81
-------
Cogbill, C. V. The history and character of acid precipitation in eastern North America. _In:
Proceedings of the First International Symposium on Acid Precipitation and the Forest
Ecosystem, U.S. Department of Agriculture and Ohio State University, Columbus, Ohio, May
12-15, 1975. L. S. Dochinger and T. A. Selige, eds., USDA Forest Service General
Technical Report NE-23, U.S. Department of Agriculture, Upper Darby, PA 1976 DP
363-370.
Coleman, R. The importance of sulfur as a plant nutrient in world crop production. Soil
Science 101:230-239, 1966.
Constantinidou, H. A., and T. T. Kozlowski. Effects of sulfur dioxide and ozone on Ulmus
americana seedlings. I. Visible injury and growth. Can. J. Bot. 57:170-175, 1979a.
Constantinidou, H. A., and T. T. Kozlowski. Effects of sulfur dioxide and ozone on Ulmus
americana seedlings. II. Carbohydrates, proteins, and lipids. Can. J. Bot. 57:176-184,
1979b. ~
Costonis, A. C. Injury to eastern white pine by sulfur dioxide and ozone alone and in
mixtures. Eur. J. For. Pathol. 3:50-55, 1973.
Cotrufo, C. , and C. R. Berry. Some effects of a soluble NPK fertilizer on sensitivity of
eastern white pine to injury from S02 air pollution. Forensic Sci. 16:72-73, 1970.
Coughenour, M. B. Grasslands Sulfur Cycle and Ecosystem Response to Low-Level SO,,. Ph.D.
Thesis, Colorado State University, Fort Collins, CO, 1978. 353 p.
Coughtrey, P. J. , C. H. Jones, M. H. Martin, and S. W. Shales. Litter accumulation in
woodlands contaminated by Pb, Zn, Cd and Cu. Oecologia 39:51-60, 1979.
Cowling, D. W. , and L. H. P. Jones. Sulphur deficiency on two forage plants in England.
Sulphur Institute J. 6:11, 1971.
Cowling, D. W. , L. H. P. Jones, and D. R. Lockyer. Increased yield through correction of
sulfur deficiency in ryegrass exposed to sulfur dioxide. Nature 243:479-480, 1973.
Cowling, D. W. , and D. R. Lockyer. The effect of S0? on Lo1iurn perenne L. grown at different
levels of sulfur and nitrogen nutrition. J. Expf Bot. 29:257-265, 1978.
Cowling, D. W. , and M. J. Koziol. Growth of ryegrass (Lolium perenne L.) exposed to S0?. I.
Effects on photosynthesis and respiration. J. Exp. Bot. 29:1029-1036, 1978.
Czaja, A. T. Zementstaubwir-Kungen auf pflanzen: Die enstehung der Zementkrusten. Qual.
Plant Mater. Veg. 8:201-238, 1961.
Daessler, H. G., J. Ranft, and K. H. Rehn. The susceptibility of woody plants exposed to
fluorine compounds and sulfur dioxide. Flora (Jean) 161:289-302, 1972.
Daines, R. H. Sulfur dioxide and plant response. J. Occup. Med. 10: 516-534, 1968.
Darley, E. F. Studies on the effect of cement-kiln dust on vegetation. J. Air Pollut.
Control Assoc. 16:145-150, 1966.
Davis, D. D. Sulfur Dioxide and Deciduous Plants. Air Pollution and Plants Series. PA. Coop.
Ext. Ser. U. Ed. 3-17, 1972a. 2 pp.
Davis, D. D. Sulfur Dioxide and Evergreens. Air Pollution and Plants Series. Pa. Coop. Ext.
Serv. U. Ed. 3-526, 1972b. 2 pp.
SOX8B/B 8-75 2-10-81
-------
Davis, J. B., and R. L. Barnes. Effects of soil-applied fluoride and lead on growth of
loblolly pine and red maple. Environ. Pollut. 5:35-44, 1973.
Davis, D. D. , and R. G. Wilhour. Susceptibility of woody plants to sulfur dioxide and
photochemical oxidants: a literature review. EPA 600/3-76-102, U.S. Environmental
Protection Agency, Corvallis, OR, September 1976.
Demeritt, M. E. , Jr., W. M. Chang, J. D. Murphy, and H. D. Gerhold. Selection system for
evaluating resistance of Scots pine seedlings to ozone and sulfur dioxide. j_n:
Proceedings of the 19th Northeastern For. Tree Improvement Conf., 1971.
Denison, W. E. Life in tall trees. Sci. Am. 228:74- , 1973.
DeSloover, J. , and F. LeBlanc. Mapping of atmospheric pollution on the basis of lichen
sensitivity. In: Proceedings of the Symp. Recent. Adv. Trop. Ecol. R. Misra and B.
Gopal eds. , The International Society for Tropical Ecology, Varanashi-5, India, 1968.
pp.42-56.
Dochinger, L. S. , and W. W. Heck. An ozone-sulfur dioxide synergism produces symptoms of
chlorotic dwarf disease of eastern white pine. Phytopathology 59:399, 1969.
Dochinger, L. S. , and K. F. Jensen. Effects of chronic and acute exposure to sulfur dioxide
on the growth of hybrid poplar cuttings. Environ. Pollut. 9:219-229, 1975.
Domes, W. Uterschiedliche C0?-Abhangigkeit des Gasaustaushes beider Blattseiten von Zea Mays.
Planta 98:186-189, 1971.^
Donaubauer, E. Sekundarshaden in Osterreichischen Rauchschadensgegieten. Schwierigkeiten der
Diagnose and Bewertung. Niedzynarodoweg Konf. WpTyw Zanieczyszczen Powietrza na Lasy,
6th., Katowice, Poland, 1968.
Dreisinger, B. P. Sulfur dioxide levels and the effects of the gas on vegetation near
Sudbury, Ontario. Presented at the 58th Annual Meeting, APCA, Toronto, Ontario, 1965.
Paper no. 65-121.
Dreisinger, R. B. , and P. C. McGovern. Monitoring Atmospheric Sulfur Dioxide and Correlating
its Effects on Crops and Forests in the Sudbury Area. In: Impact of Air Pollution on
Vegetation Conf. S. N. Linzon, ed., Ontario Department of Energy and Resource Management,
Toronto, 1970. 122 pp.
Durran, D. R. , M. J. Meldgin, M. Liu, T. Thoem, and D. Henderson. A study of long-range air
pollution problems related to coal development in the northern Great Plains. Atmos. Environ.
13: 1021-1037, 1979.
Embleton, T. W. Magnesium. _In: Diagnostic Criteria for Plants and Soils. H.D. Champman,
ed. , University of California Division of Agricultural Sciences, Riverside, CA. 1966.
pp. 225-263.
Enderlein, H. , arid W. Kastner. What effect has a nutrient deficiency .on the resistance of
one-year-old conifers to SOp? Arch. Forstwes. 16:431-435, 1976.
Faller, N. Effects of atmospheric S02 on plants. Sulfur Inst. J. 6:5-7, 1970.
Fankhauser, N. C. Brunold, and K. N. Erisman. The influence of sublethal concentrations of
sulfur dioxide on morphology, growth and production yield of the duckweed Lemne minor L.
Oecologia 23:201-208, 1976.
SOX8B/B 8-76 2-10-81
-------
Ferenbaugh, R. W. Effects of prolonged exposure of Oryzopsis hemenoides to S00. Water Air
Soil Pollut. 10:27-31, 1978. — 2
Garland, J. A. Dry and wet removal of sulfur from the atmosphere. Atmos. Eviron. 12:349-362,
1978. —
Gerhold, H. D. Effect of air pollution on Pinus strobus L. and genetic resistance.
EPA-600/3-77-002, U.S. Environmental Protection Agency, Corvallis, OR, January 1977.
Gilbert, 0. L. An alkaline dust effect on epiphytic lichens. Lichenologist 8:173-178, 1976.
Godzik, S. , and H. F. Linskens. Concentration changes of free amino acids in primary bean
leaves after continuous and interrupted S00 fumigation and recovery. Environ. Pollut.
7:25-28. 1974. 2
Granat, L. , H. Rehe, and R. 0. Hallberg. The global sulfur cycle. Ecol. Bull. (Stockholm)
22:89-134, 1976.
Guderian, R. Reaktione von Fflanzer Gemeinschaften des Feslufferbaues auf
Schwefeldioxideinwirkungen. Essen: Girardet-Verlag, Schriftenr. Landesnst.
Immisions-Bodennutsungssch. d. Landes Nordrheim-Westfalen 4:80-100, 1967.
Guderian R. , and H. Stratmann. Freilandversuche zur Ermittlung von Schwefeldioxidwirkungen
auf die Vegetation. III. Grenzewerte schadlicher SOp-Immissionen fur Obst-und
Forstkulturen Sowie fur land wirtschaftliche and gratnerische Pflanzenarten.
Forschunsber. Landes Nordrhein-West-falen Nr. 1920. Westdeutscher Verlag, Cologne,
1968. 113 pp.
Guderian, R. Air Pollution: Phytotoxicity of Acidic Gases and its Significance in Air
Pollution Control. Springer-Verlag, Berlin, 1977. 127 pp.
Guderian, R. Terrestrial Vegetation--Air Pollutant Interactions: Non-Gaseous Air Pollutants.
Presented at the International Conference on Air Pollutants and Their Effects on the
Terrestrial Ecosystem. Banff, Alberta, Canada, May 10-17, 1980.
Haisman, D. R. The effect of sulfur dioxide on oxidizing enzyme systems in plant tissue. J.
Sci. Food Agric. 25:803-810, 1974.
Halbwachs, G. Der Wasserhaushalt rauchgeschadigter Holzgewasche. Allg. Forstz. 78:196-197,
1976.
Hallgren, J-E. Physiological and Biochemical Effects of Sulfur Dioxide on Plants. I_n:
Sulfur in the Environment, vol. 2. J. 0. Nriagu, ed. , John Wiley & Sons, New York, NY,
1978. pp. 163-210.
Hawksworth, D. L. Lichens as litmus for air pollution: a historical review. Int. J.
Environ. Stud. 1:281-296, 1971.
Hawksworth, D. L., and F. Rose. Qualitative scale for estimating sulfur dioxide air pollution
in England and Wales using epiphytic lichens. Nature 227:145-148, 1970.
Heagle, A. S. Interactions between air pollutants and plant parasites. Annu. Rev.
Phytopathol. 11:365-388, 1973.
Heagle, A. S. , D. E. Body, and G. E. Neely. Injury and yield responses of soybean to chronic
doses of ozone and sulfur dioxide in the field. Phytopathology 64:132-136, 1974.
SOX8B/B 8-77 2-10-81
-------
Heagle, A. S. , and J. W. Johnston. Variable responses of soybeans to mixtures of ozone and
sulfur dioxide. J. Air Pollut. Control Assoc. 29:729-732, 1979.
Heck, W. W. , and C. S. Brandt. Effects on Vegetation: Native, Crops, Forest, in: Air
Pollution. A. C. Stern, ed. , Academic Press, New York, NY, 1977. pp. 157-229.
Heck, W. W. , and J. A. Dunning. Response of oats to sulfur dioxide: interactions of growth
temperature with exposure temperature or humidity. J. Air Pollut. Control Assoc.
28:241-246. 1978.
Heggestad, H. E. , K. L. Tuthill, and R. N. Stewart. Differences among poinsettias in
tolerance to sulfur dioxide. Hort. Sci. 8:377-338, 1973.
Heitschmidt, R. K. , W. K. Lauenroth, and J. L. Dodd. Effects of controlled levels of sulfur
dioxide on western wheatgrass in a southeastern Montana grassland. J. Appl. Ecol.
14:859-868, 1978.
Hill, A. C., S. Hill, Lamb, and T. W. Barrett. Sensitivity of native desert vegetation to SC>2
and SO, and N00 combined. J. Air Pollut. Control Assoc. 24:153-157, 1974.
L. C. ~
Hofstra, G. , and D. P. Ormrod. Ozone and sulfur dioxide interaction in white bean and
soybean. Can. J. Plant Sci. 57:1193-1198, 1977.
Horsman, D. C. , and A. R. Wellburn. Synergistic effect of S02 and N02 polluted air upon
enzyme activity in pea seedlings. Environ. Pollut. 8:123-133, 1975.
Horsman, D. C. , and A. R. Wellburn. Appendix II. Guide to Metabolic and Biochemical Effects
of Air Pollutants on Higher Plants. In: Effects of Air Pollutants on Plants. T. A.
Mansfield, ed. , Cambridge University Press, Cambridge, 1976. pp. 185-199.
Horsman, D. C. , T. M. Roberts, and A. D. Bradshaw. Evolution of sulfur dioxide tolerance in
perennial ryegrass. Nature 276:493-494, 1978.
Hou, L-Y. , A. C. Hill, and A. Soleimani. Influence of C09 on the effects of S09 and N09 on
alfalfa. Environ. Pollut. 12:7-16, 1977. ^ ^ ^
Houston, D. C. , and G. R. Stairs. Genetic control of sulfur dioxide and ozone tolerance of
eastern white pine. For. Sci. 19:267-271, 1973.
Houston, D. B. Response of selected Pinus strobus L. clones to fumigation with sulfur dioxide
and ozone. Can. J. For. Res. 4:65-68, 1974.
Husar et al., 1977
Irving, P. M., D. J. Martinson, and J. E. Miller. A histological study of soybeans injured by
exposure to sulfur dioxide. Radiol. Environ. Res. Div. Ann. Rep., Argonne National Lab.,
ANL-78-65. Part III, 1979.
Jacobson, J. S. , and L. J. Colavito. The combined effect of sulfur dioxide and ozone on bean
and tobacco plants. Environ. Exp. Bot. 16:177-185, 1976.
Jacobson, J. S. , and A. C. Hill. Recognition of Air Pollution Injury to vegetation: A
Pictorial Atlas. Air Pollution Control Association, Pittsburgh, PA, 1970.
Jager, J. H. , and H. Klein. Biochemical and physiological detection of sulfur dioxide injury
to pea plants (Pisum sativum). J. Air Pollut. Control Assoc. 27:464-466, 1977.
SOX8B/B 8-78 2-10-81
-------
Jancarik, V. Vyskyt drevokaznych hub v hourem poskozovani oblasti Krusnych hor. Lesnictvi
7:677-692, 1961.
Jennings, 0. E. Smoke injury to shade trees. Natl. Shade Tree Conf. 10:44-48, 1934.
Jensen, K. R. , and T. T. Kozlowski. Absorption and translocation of sulfur dioxide by
seedlings of four forest tree species. J. Environ. Qual. 4:379-382, 1975.
Jones, H. D., D. Weber, and D. Balsillie. Acceptable limits for air pollution dosages and
vegetation effects: sulfur dioxide. Presented at the 67th Annual Meeting, Air Pollution
Control Association, Denver, CO, 1974. Paper No. 74-225.
Jordan, M. Effects of zinc smelter emissions and fire on a chestnut oak woodland. Ecology
56:78-91, 1975.
Kamprath, E. J. Possible effects from sulfur in the atmosphere. Combustion 44:16-17, 1972.
Karnosky, D. F. Threshold levels for foliar injury to Populus tremuloides by sulfur dioxide
and ozone. Can. J. For. Res. 6:166-169, 1976.
Karnosky, D. F. Evidence for genetic control of response to sulfur dioxide and ozone in
Populus tremuloides. Can. J. For. Res. 7:437-440, 1977.
Katz, M. Effect of contaminants other than sulfur dioxide on vegetation and animals. In:
Pollution and Our Environment Conf., Background Paper A4-2-2, Vol. 1, Council Res.
Ministers, Montreal, Can. 1967. pp. 1-18.
Katz, M. Sulfur dioxide in the atmosphere and its relation to plant life. Ind. Eng. Chem.
41:2350-2465, 1949.
Katz, M. , and A. W. McCallum. The effect of sulfur dioxide on conifers. Proceedings of the
U.S. Tech. Conf. Air Pollution, 1952. pp. 84-96.
Keller, T. The effect of long duration, low S02 concentrations upon phytosynthesis of conifers.
Jjn: Proceedings of the 4th International Clean Air Congress, 1977. pp. 81-83.
Keller, T. Wintertime atmospheric pollutants—do they effect the performance of deciduous trees
in the ensuing growing season. Environ. Pollut. 16:243-247, 1978.
Keller, T. The effect of a continuous springtime fumigation with SO,, and CO^ uptake and
structure of the annual ring in spruce. Can. J. For. Res. 10:1-6, T.980.
Kender, W. J. , and F. H. F. G. Spierings. Effects of sulfur dioxide, ozone, and their inter-
actions on "golden delicious" apple trees. Neth. J. Plant Pathol. 81:149-151, 1975.
Kickert, R. N. , and P. R. Miller. Responses of Ecological Systems. I_n: Methodology for the
Assessment of Air Pollution Effects on Vegetations. W. W. Heck, S. V. Krupa, and S. N.
Linzon, eds. , Air Pollution Control Association, 1979.
Make, 1974.
Kodata, M. , and T. Inoue. Invading path of sulfur dioxide into pine leaves as revealed by
microradioauthography. J. For. Soc. 54:207-208, 1972.
Krause, G. H. M. , and H. Kaiser. Plant response to heavy metals and sulfur dioxide. Environ.
Pollut. 12:63-71, 1977.
SOX8B/B 8-79 2-10-81
-------
Krouse, H. R. Sulphur isotope abundance elucidate uptake of atmospheric sulphur emissions by
vegetation. Nature 265:46, 1977.
Krupa, S. V., and R. J. Kohut. Impact of stark emissions from the NSP-SHERCO power plant on
terrestrial vegetation. Annual Report, Northern States Power Co., Minneapolis, MN, 1976.
Kudela, M. , and E. Novakova. Lesnf skudci a skody zeri v lesich poskozovanych Kourum.
Lesnictvi 6:493-502, 1962.
Labanauskas, C. K. Manganese. Ijr Diagnostic Criteria for Plants and Soils. H. D. Chapman,
ed., University of California, 1966. pp. 264-285.
Lacasse, N. L., and W. J. Moroz. Handbook of effects assessment; vegetation damage. CAES.
Pennsylvania State University, University Park, PA, 1969.
Lagerwerff, J. V. Lead, Mercury, and Cadmium as Environmental Contaminants. In:
Micronutrients in Agriculture. J. J. Mortuedt, P. M. Giodana, and W. L. Lindsay, eds. ,
Soil Science Society of America, Madison, WI, 1972. pp. 593-636.
Larsen, R. I., and W. W. Heck. An air quality data analysis system for interrelating effects,
standards, and needed source reduction: Part 3. Vegetation injury. J. Air Pollut.
Control Assoc. 26:325-333, 1976.
Lauenroth, W. K. , and J. E. Heasley. Impact of atmospheric sulfur deposition on grassland
ecosystems. In: Atmospheric Sulfur Deposition: Environmental Impacts and Health
Effects. D. S. Shriner, A. Richmond, and S. E. Lindbergh, eds., Ann Arbor Science
Publishers, in press, 1980.
Laurence, J. A. Effects of air pollutants on plant pathogen interaction. In: Proceedings of
the 71st Annual Meeting, Air Pollution Control Association, Houston, TX, 1978. pp. 3-24.
Laurence, J. A. Response of maize and wheat to sulfur dioxide. Plant Dis. Reptr. 63:468-471,
1979.
Laurence, J. A., L. G. Weinstein, D. C. McCune, and A. L. Aluisio. Effects of sulfur dioxide
on southern corn leaf blight maize and stem rust of wheat. Plant Dis. Reptr. 63:975-978,
1979.
Laurence, J. A., A. L. Alvisio, L. H. Weinstein, and D. C. McCune. Effects of sulphur dioxide
on southern bean mozaic and maize dwarf moziac. Environ. Pollut., 1981. (in press)
LeBlanc, F. , and D. N. Rao. Effects of air pollutants on lichens and bryophytes. In:
Responses of Plants to Air Pollution. J. B. Mudd and T. T. Kozlowski, eds., Academic
Press, New York, NY, 1975. 383 pp.
LeBlanc, F. , and D. N. Rao. Effects of sulfur dioxide on lichen and moss transplants.
Ecology 54:612-617, 1973.
Lee, J. J. , E. M. Preston, and R. A. Lewis. A system for the experimental evaluation of the
ecological effects of sulfur dioxide. Proceedings of the 4th Joint Conference on Sensing
of Environmental Pollutants, American Chemical Society, Washington, DC, 1978. pp.
49-53.
Legge, A. H. , C. W. Harver, P. F. Lester, D. R. Jaques, H. R. Krouse, J. Mayo, A. P.
Hartgerink, R. G. Amundson, and R. B. Walker. Quantitative assessment of the impact of
sulfur gas emissions on a forest ecosystem. Final report submitted to Whitecourt
Environmental Study Group. Environmental Sciences Centre, Kananskis, The University of
Calgary, Alberta, 1976.
SOX8B/B 8-80 2-10-81
-------
Leone, I. A., and E. Brennan. Modification of sulfur dioxide injury to tobacco and tomato by
varying nitrogen and sulfur nutrition. J. Air Poll. Control Assoc. 22:544-547, 1972.
Lerman, S. L. , and E. F. Darley. Particles, J.n: Responses of Plants to Air Pollution. J.
B. Mudd and T. T. Kozlowski, eds. , Academic Press, New York, NY, 1975. pp. 141-158.
Liebig, G. F. Arsenic. |n: Diagnostic Criteria for Plants and Soils. H. D. Chapman, ed. ,
University of California, 1966. pp. 13-23.
Likens, G. E. , F. H. Bormann, R. S. Pierce, J. S. Eaton, and N. M. Johnson. Biogeochemistry
of a Forested Ecosystem. Springer-Verlag, New York, 1977.
Lindberg, S. E. , A. C. Harriss, R. R. Turner, D. S. Shriner, and D. D. Huff. Mechanisms and
rates of atmospheric deposition of selected trace elements and sulfate to a deciduous
forest watershed. Pub. No. 1299, Oak Ridge National Laboratory, Oak Ridge, TN, 1979.
Linzon, S. N. The influence of smelter fumes on the growth of white pine in the Sudbury
region. Canadian Department of Agriculture Publication, Ontario Dep. Lands Forests,
1958. 45 pp.
Linzon, S. N. Damage to eastern white pine by sulfur dioxide, semimature tissue needles
blight and ozone. J. Air Pollut. Control Assoc. _16:140-144, 1966.
Linzon, S. N. Economic effects of sulphur dioxide on forest growth. J. Air Pollut. Control
Assoc. 21:81-86, 1971.
Linzon, S. N. Effects of sulfur oxides on vegetation. Forest Chronicle 48:182-186, 1972.
Linzon, S. N. Some effects of particulate matter on vegetation. J_n: Ontario Proceedings of
the 3rd International Clean Air Congress, Dusseldorf, West Germany, 1973. pp. A118-A120.
Linzon, S. N. Vegetation injury by airborne arsenic and sulphur dioxide emissions from gold
smelters. Proceedings of the Fourth International Clean Air Congress, Tokyo, Japan,
1977.
Linzon, S. N. Effects of Airborne Sulfur Pollutants on Plants. lr\: Sulfur in the
Environment: Part II, Ecological Impacts. J. 0. Nriagu, ed. , John Wiley & Sons. New
York, NY, 1978. pp. 109-162.
Linzon, S. N. Acute and chronic effects of sulfur dioxide on natural vegetation. Proceedings
of the Air Pollut. Control Assoc. Met., Atlanta, GA, September 1980.
Little, P. A study of heavy metal contamination of leaf surfaces. Environ. Pollut.
5:159-172, 1973.
Little, P. Deposition of 2.75, 5.0 and 8.5mm particles to plant and soil surfaces. Environ.
Pollut. 12:293-305, 1977.
Little, P., and R. D. Wiffen. Emission and deposition of petrol engine exhaust Pb-1.
Deposition of exhaust Pb. Atmos. Environ. 11:437-447, 1977.
Majernik, 0., and T. A. Mansfield. Direct effect of So? pollution on the degree of opening of
stomata. Nature 227:377-378, 1970.
Malhotra, S. S. , and A. A. Kahn. Effects of sulfur dioxide fumigation on lipid biosynthesis
in pine needles. Phytochemistry 17:241-244, 1978.
SOX8B/B 8-81 2-10-81
-------
Mandl, R. H., L. H. Weinstein, and M. Keveny. Effects of hydrogen fluoride and sulfur dioxide
alone and in combination on several species of plants. Environ. Pollut. 9:133-143, 1975.
Manning, W. J. Effects of limestone dust on leaf condition, foliar disease incidence, and leaf
surface microflora of native plants. Environ. Pollut. 2:69-76, 1971.
Mansfield. T. A., and 0. Majernik. Can stomata play a part in protecting plants against air
pollution? Environ. Pollut. 1:149-154, 1970.
Markowski, A., S. Grzesiak, and M. Schramel. Indexes of susceptibility of various species of
cultivated plants to sulfur dioxide action. Bulletin de L'Academie Polanaise des
Sciences 23(9):637-646, 1975.
Masaru, N. , F. Syozo, and K. Saburo. Effects of exposure to various injurious gases on
germination of lily pollen. Environ. Pollut. 11:181-187, 1976.
Materna, J. , J. Jirgle, and J. Kucera. Vysledky merceni koncentraci lyclicniky siricitheo v.
lesich krusnych hor. (Measurement results of sulfur dioxide concentrations in the Ore
Mountain Forests.) Ochrana Oyzdusi. 1:84-92, 1969.
Matsushima, J. , and R. R. Brewer. Influence of sulfur dioxide and hydrogen fluoride as a mix
or reciprocal exposure on citrus growth and development. J. Air Pollut. Control Assoc.
22:710-713, 1972.
May, P. F. , A. R. Till, and M. J. Cumming. Systems analysis of sulfur kinetics in pastures
grazed by sheep. J. Appl. Ecol. 9:25-49, 1972.
Mclaughlin, S. B., and N. T. Lee. Botanical Studies in the Vicinity of the Widows Creek Steam
Plant. Review of Air Pollution Effects Studies, 1952-1972, and Results of 1973 Surveys.
Internal Report I-EB-74-1, TVA, 1974.
Mclaughlin et al., 1976
Mclaughlin, S. B. , D. S. Shriner, R. L. M'Coutley, and L. K. Mann. The effects of S02 dosage
kinetics and exposure frequency on photosynthesis and transpiration of kidney beans
(Phaseolus vulgaris L). Env. Exp. Bot. 19:179-191, 1979.
Mclaughlin, S. B. S0? vegetation effects and the air quality standard: Limits of interpre-
tation and application. Proceedings of the Air Pollut. Control Assoc. Mtg. , Atlanta, GA,
September 1980.
Mclaughlin, S. B. , and D. S. Shriner. Allocation of Resources to Defense and Repair. Chapter
in Plant Disease: An Advanced Treatise. J. G. Horsfall and E. B. Cowling, eds. ,
Academic Press, 1980.
Mclaughlin, S. B. , and Taylor. Relative humidity: Important modifier of pollutant
uptake by plants. Science 211:167-169, 1981.
Meidner, H. , and T. A. Mansfield. Physiology of stomata. McGraw-Hill, Great Britian, 1968.
128 pp.
Menser, H. A., and H. E. Heggestad. Ozone and sulfur dioxide synergism: injury to tobacco
plants. Science 153:424-425, 1966.
Meszaros, E. , G. Varhelyi, and L. Haszpra. On the atmospheric sulfur budget over Europe.
Atmos. Environ. 12:2273-2277, 1978.
SOX8B/B 8-82 2-10-81
-------
Miller, P., and J. R. McBride. Effects of Air Pollutants on Forest. In: Response of Plants
to Air Pollution. J. B. Mudd and T. T. Kozlowski, eds. , Academic Press, New York, NY,
1975. pp. 195-235.
Miller, J. E. , H. J. Smith, P. G. Sprugel, and P. B. Xerikos. Yield response of field-grown
soybeans to an acute SO, exposure. Radiol. Environ. Res. Div. Annu. Rep., Argonne
National Laboratory, ANL-78-65, Part III, 1979.
Miller, J. E. , D. G. Sprugel, R. N. Muller, H. J. Smith, and P. B. Xerikas. Open-air
fumigation system for investigating sulfur dioxide effects on crops. Phytopathology 70,
1980 (in press).
Moss, M. R. Biogeochemical cycles as integrative and spatial modes for the study of
environmental pollution (the example of the sulphur cycle). Int. J. Environ. Studies
9:209-216, 1976.
Mudd, J. B. Sulfur Dioxide. In: Responses of Plants to Air Pollution. J. B. Mudd and T. T.
Kozlowski, eds., Physiological Ecol. Mono. Series, Academic Press, Inc., New York, NY,
1975. pp. 9-22.
Murray, J. J. , R. K. Howe 11, and A. C. Wilton. Differential response of 17 Poa pratensis
cultivars to ozone and sulfur dioxide. Plan Dis. Rep. 59:852-854, 1975.
National Academy of Science. Series on Medical and Biological Effects of Environmental
Pollutants. National Academy of Sciences, Washington, DC, 1973. 191 pp.
National Academy of Science. Principles for Evaluating Chemicals in the Environment.
National Academy of Sciences, Washington, DC, 1975.
National Academy of Sciences. Sulfur oxides. Board on Toxicology and Environ. Health
Hazards. U.S. Environmental Protection Agency Contract, Report No. 68-01-4655, 1978.
National Air Quality, Monitoring and Emissions Trends Report, 1977. 1978. U. S. Environmental
Protection Agency, Research Triangle Park, NC.
Neely, G. E., D. T. Tingey, and R. G. Wilhour. Effects of ozone and sulfur dioxide singly and
in combination on yield, quality and N-fixation of alfalfa. Proceedings of the
International Conference on Photochemical Oxidant Pollution and Its Control.
EPA-600/3-77-001b, 1977. pp. 663-673.
Neely, G. , R. G. Wilhour, D. Weber, and L. Grothaus. The Response of Selected Small Grains,
Range Grasses, and Alfalfa to S0?. Bioenvironmental Impact of a Coal-fired Power Plant:
An Interm Report. E. M. Preston and T. L. Gallett, eds., EPA-600/3-79-044, U.S.
Environmental Protection Agency, Ecol. Res. Ser., Colstrip, MT, 1980.
Neely, G. E. , and R. G. Wilhour. Yield response of Yamhill and Hyslop winter wheat cultivars
to chronic S0? exposures. J. Air Pollut. Control Assoc., 1981. (in press)
Neilson, D. G. , L. E. Terrell, and T. C. Weidensaul. Phytotoxicity of ozone and sulfur
dioxide to laboratory fumigated Scots pine. Plant Dis. Rep. 61:699-703, 1977.
Noggle, J. C., and H. C. Jones. Accumulation of Atmospheric Sulfur by Plants and Sulfur-
supplying Capacity of Soils. EPA-600/7-79-109, U.S. Environmental Protection Agency,
1979.
Noland, T. L. , and T. T. Kozlowski. Effect of S02 on stomatal aperature and sulfur uptake of
woody angiosperm seedlings. Can. J. For. Res. 9:57-62, 1979.
SOX8B/B 8-83 2-10-81
-------
O'Connor, J. A., D. G. Parbery, and W. Strauss. The effect of phytotoxic gases on native
Australian plant species: Part I. Acute effects of sulfur dioxide. Environ. Pollut.
7:7-23, 1974.
Odum, E. P. Fundamentals of Ecology, 3rd ed. W. B. Saunders, Co., Philadelphia, PA, 1971.
570 pp.
O'Gara, P. J. Sulfur dioxide and fume problems and their solutions. I_n: 14th Semiannual
Meeting of the American Institute of Chemical Engineers. J. Ind. Eng. Chem. 14:744,
1922.
Oshima, R. J. The impact of sulfur dioxide on vegetation: a sulfur dioxide-ozone response
model. Final Report of the Cal. Air Res. Board Agree. No. A6-162-30, 1978.
Pahlich, E. Allosterische Regulation der Aktivitat der Glutamat-dehydrogenase aus Erbsenkeim-
lingen durch das Substrat L-Ketoglutarsaure. Planta (Berline) 100:222-227, 1971.
Pahlich, E. Uber den Hemm-Mechanismus mitochondriater Glutamat-Oxalacetat-transaminase in
S02-begaster Erbsen. Planta 110:267-278, 1973.
Pahlich, E. Effect of S0?-pollution on cellular regulation. A general concept of the mode of
action of gaseous air contamination. Atmos. Environ. 9:261-263, 1975.
Parish, S. B. The effect of cement dust on citrus trees. Plant World 13:288-291, 1910.
Pelz, E. The individual smoke resistance of spruce. Wiss. 2. Tech Univ. , Dresden
11:3:595-600, 1962.
Preston, E. M. , and R. A. Lewis. The Bioenvironmental Impact of a Coal Fired Power Plant,
Colstrip, Montana, December, 1977. EPA-600/3-78-021, U.S. Environmental Protection
Agency, 1978.
Preston, E. M. The ecological implications of chronic sulfur dioxide exposure for native
grasslands. Presented at the 72nd Annual Meeting Air Pollution Control Association,
Cincinnati, OH, 1979.
Rajput, C. B. S. , D. P. Ormrod, and D. W. Evans. The resistance of strawberry to ozone and
sulfur dioxide. Plant Dis. Rep. 61:76-79, 1977.
Rao, D. N. , and F. LeBlanc. Effect of sulfur dioxide on the lichen algae with special
reference to chlorophyll. Bryologist 69:69-75, 1965.
Reinert, R. A. Pollutant interactions and their effects on plants. Environ. Pollut.
9:115-116, 1975.
Reinert, R. A., A. S. Heagle, and W. W. Heck. Plant Response to Pollutant Combinations. I_n:
Responses of Plants to Air Pollution. J. B. Mudd and T. T. Koslowski, eds. , Academic
Press Inc., 1975. pp. 159-177.
Reinert, R. A., and D. E. Weber. Ozone and sulfur dioxide-induced changes in soybean growth.
Phytopathology 70:914-916, 1980.
Reinert, R. A., and P. V. Nelson. Sensitivity and growth of five Elation Begonia cultivars to
S02 and 0.,, alone and in combination. J. Amer. Soc. Hort. Sci. 105:721-723, 1980.
Reinert, R. A., and T. N. Gray. The response of radish to nitrogen dioxide, sulfur dioxide,
and ozone alone and in combination. J. Environ. Qual., 1981. (in press)
SOX8B/B 8-84 2-10-81
-------
Rennie, P. J., and R. L. Halstead. The Effects of Sulfur on Plants in Canada. Jji: Sulfur
and its Inorganic Derivatives in the Canadian Environment. National Research Council
Canada, NRCC No. 15015:69-180, 1977.
Reuther, W., and C. K. Labanauskas. Copper. |n: Diagnostic Criteria for Plants and Soils.
H. D. Chapman, ed., University of California, 1966. pp. 157-179.
Ricks, G. R. , and J. H. Williams. Effects of atmospheric pollution on deciduous woodland.
Part 2: Effects of particulate matter upon stomatal diffusion resistance in leaves of
Quercus petraea (Mattuschka) Liebl. Environ. Pollut. 6:87-109, 1974.
Rist, D. L. , and D. D. Davis. The influence of exposure temperature and relative humidity on
the response of pinto bean foliage to S02. Phytopathology 69:231-235, 1979.
Rosenberg, C. R., R. J. Hutnik, and D. D. David. Forest composition of varying distances from
a coal-burning power plant. Environ. Pollut. 14:307-317, 1979.
Saunders, P. J. W. Effects of atmospheric pollution of leaf surface microflora. Pestic Sci.
4:589-594, 1973.
Schoenbeck, H. Beobachtungen zur Frage des einflusses von industriellen immissionen auf die
drankheit bereitschaft der Pflanze. Benchte der Landesanstalt fur Bodennutzungschertz.
1:89-98, 1960.
Schoenbeck, H. Detection of heavy metal-containing air pollutants by selected plant indicators.
Ver. Dtsch. Ing. 203:75-85, 1973.
Sehmel, G. A., and W. A. Hodgson. Predicted dry deposition velocities. In: Atmosphere-
surface exchange of particulate and gaseous pollutants symposium. Conf-74092. Richland,
WA, 1974. pp. 399-422.
Setterstrom, C. , and P. W. Zimmerman. Factors influencing susceptibility of plants to sulfur
dioxide injury. Contrib. Boyce Tompson Inst. 10:155-186, 1939.
Sharp, E. L. Atmospheric ions and germination of uredospores of Purcinia striiformis.
Science 156:1359-1360, 1967.
Sharp, E. L. Relation of air ions to air pollution and some biological effects. Environ.
Pollut. 3:227-239, 1972.
Sheridan, R. P. Impact of emissions from a coal-fired electricity generating facilties on
N2-fixing in lichens. Bryologist 82:54-58, 1979.
Shin, J. H. A critical survey of measurements of foliar deposition of airborne sulfates and
nitrates. Ann. Air Pollut. Control Assoc. 78:1-16, 1978.
Shinn, J. H., and S. Lynn. Do manmade sources affect the sulfur cycle of northeastern states?
Environ. Sci. Technol. 13:1062-1067, 1979.
Showman, R. E. Linchens as indicators of air quality around a coal-fired power generating
plant. Bryologist 78:1-6, 1975.
Shriner, D. S. , and G. S. Henderson. Sulfur distribution and cycling in a deciduous forest
watershed. J. Environ. Qual. 7:392-397, 1978.
Shugart, H. H. , and D. C. West. Development of an Appalachian deciduous forest succession
model and its application to assessment of the impact of the Chestnut Blight. J.
Environ. Manag. 5:161-179, 1977.
SOX8B/B 8-85 2-10-81
-------
Sij, J. W. , E. T. Kanemasu, and S. M. Goltz. Some preliminary results of sulfur dioxide
effects on phytosynthesis and yield in field-grown wheat. Tans Kans. Acad. Sci.
76:199-297, 1974.
Smith, W. H. Metal contamination of urban woody plants. Environ. Sci. Technol. T_:631-636,
1973.
Smith, H. J. , and D. D. Davis. The influence of needle age on sensitivity of Scotch pine to
acute doses of S02. Plant Dis. Reptr. 61:870-874, 1977.
Smith, H. J. , and D. D. Davis. Susceptibility of conifer cotyledons and primary needles to
acute doses of sulfur dioxide. Hortic. Sci. 13:703-704, 1978.
Smith, R. L. Ecology and Field Biology, 2rd ed., p. 170. New York, Harper and Row, 1980. pp
11-199.
Spedding, D. J. Uptake of sulfur dioxide by barley leaves at low sulfur dioxide concentrations.
Nature 224:1229-1231, 1969.
Spinka, J. Effects of polluted air on fruit trees and legumes. Ziva 19:13-15, 1971.
Sprugel, D. G. , J. E. Miller, R. N. Muller, H. J. Smith, and P. B. Xerikos. Sulfur dioxide
effects on yield and seed quality in field-grown soybeans. Phytopathology 70, 1980 (in
press).
Stewart, D. , M. Treshow, and F. M. Harner. Sulfur dioxide and hydrogen fluoride emissions in
regard to phytopathology. Can. J. Bot. 51:983-988, 1973.
Sulphur Institute. Crop Responses to Sulphur in North America. Vol. 18. The Sulphur
Institute, Washington, DC, 1971. 38 pp.
Sulphur Institute. Sulphur in Agriculture Series. Vol. 3. J. S. Platu, ed. , The Sulphur
Institute, Washington, DC, 1979.
Temple, P. J. Dose-response of urban trees to sulfur dioxide. J. Air Pollut. Control Assoc.
22:271-274, 1972.
Thomas, M. D. , and G. R. Hill, Jr. Absorption of sulfur dioxide by alfalfa and its relation
to lead injury. Plant Physiol. 10:291-307, 1935.
Thomas, M. D., R. H. Hendrick, T. R. Collier, and G. R. Hill. The utilization of sulphate and
sulphur dioxide for the sulphur nutrition of alfalfa. Plant Physiology 18:345-371, 1943.
Thomas, M. D. Gas damage to plants. Ann. Rev. Plant Physiol. 2:293-322, 1951.
Thomas, M. D. , and R. H. Hendricks. Effect of air pollution on plants i_n Air Pollution.
World Health Organization Monograph Ser. No. 46., Columbia University Press, New York,
NY, 1951. pp. 233-278.
Thomas, M. D., and R. H. Hendricks. Effect of Air Pollution on Plants. In: Air Pollution
Handbook. P. L. Magill et al. eds., New York, NY, 1956. pp. 9:1-44.
Thompson, L. K. , S. S. Sidhu, and B. A. Roberts. Fluoride accumulations in soil and
vegetation in the vicinity of a phosphorus plant. Environ. Pollut. 18:221-234, 1979.
Tingey, D. T. , R. A. Reinert, J. A. Dunning, and W. W. Heck. Vegetation injury from the
interaction of nitrogen dioxide and sulfur dioxide. Phytopathology 61:1506-1511, 1971a.
SOX8B/B 8-86 2-10-81
-------
Tingey, D. T. , W. W. Heck, and R. A. Reinert. Effect of low concentrations of ozone and
sulfur dioxide on foliage, growth, and yield of radish. J. Am. Soc. Hort. Sci.
96:369-371, 1971b.
Tingey, D. T. , R. A. Reinert, J. A. Dunning, and W. E. Heck. Foliar injury responses of 11
plant species to ozone/ sulfur dioxide mixtures. Atmos. Environ. 7:201-208, 1973a.
Tingey, D. T. , R. A. Reinert, C. Wickliff, and W. W. Heck. Chronic ozone or sulfur dioxide
exposures or both affect the early vegetative growth of soybean. Can. J. Plant Sci.
53:875-879, 1973b.
Tingey, D. T. , and R. A. Reinert. Effect of ozone and sulfur dioxide singly and in combina-
tion on plant growth. Environ. Pollut. 9:117-125, 1975.
Tingey, D. T. , and G. E. Taylor, Jr. Variation in plant response to ozone. Proceedings of
the 32nd Conference of the School of Agriculture, University of Nottingham, England,
1980. 45 pp.
Treshow, M. Environment and Plant Response. McGraw-Hill Book Co., New York, NY, 1970. 422
pp.
Tyler, G. Heavy metals pollute nature, may reduce productivity. Ambio 1:53-59, 1972.
U.S. Department of Health, Education, and Welfare. Air Quality Criteria for Particulate
Matter. Public Health Service, EHS, National Air Pollution Control Administration,
Washington, DC, 1969.
U.S. Environmental Protection Agency. Mount Storm, West Virginia -Gorman, Maryland, and Luke,
Maryland - Keyser, West Virginia Air Pollution Abatement Activity, 1971. 67 pp.
U.S. Environmental Protection Agency. Effects of Sulfur Oxides in the Atmosphere on Vegeta-
tion. EPA-R3-73-030, National Environmental Research Center, U.S. Environmental Protec-
tion Agency, Raleigh, NC, 1973. 43 pp.
U.S. Environmental Protection Agency. The Bioenvironmental Impact of Fine Particulates: A
Critical Review and Summary. USEPA/CERL Report, 1975. Mimeo, 49 pp.
U.S. Environmental Protection Agency. Diagnosing Vegetation Injury Caused by Air Pollution.
D. R. Hicks, ed., EPA Cont. Pub. No. 68-02-1344, 1976.
U.S. Environmental Protection Agency. Airborne Particles. Environmental Health Effects
Series, Report No. EPA-600/1-77-053, 1977.
U.S. Environmental Protection Agency. Airborne Particles. EPA-600/1-77-053, Environmental
Health Effect Series, 1977a. 555 pp.
U.S. Environmental Protection Agency. Office of Air Quality Planning and Standards. National
Air Quality, Monitoring and Emissions Trends Report, 1977, Research Triangle Park, NC,
Publ. No. EPA-450/2-78-052, 1978.
U.S. Environmental Protection Agency. Emission of Sulfur-Bearing Compounds from Motor Vehicle
and Aircraft Engines. A Report to Congress. EMSL Report EPA-600/9-78-028, 1978. 437
pp.
Ulrich, A., and K. Ohki. Potassium. In: Diagnostic Criteria for Plants and Soils. H. D.
Chapman, ed. , University of California, 1966. pp. 157-179.
SOX8B/B 8-87 2-10-81
-------
Unzicker, H. J. , H. J. Jager, and L. Steubing. Influence of S00 on von Pfanzen. Agnew. Bot.
49:131-139, 1975. i
Van Haul, H. , and H. Stratmann. Farbtfelatlas uber Schwefeldioxidwirkugen and Pflanzen.
(Color-Plate Atlas of the Effects of Sulfur Dioxide on Plants.) Verlag W. Girardet,
Essen, West Germany, 1970. 206 pp.
Vanselow, A. P. Nickel. Jn: Diagnostic Criteria for Plants and Soil. H. D. Chapman, ed. ,
University of California, 1966. pp. 302-309.
Vins, B. , and R. Mrkva. Thr diameter increment losses of pine stands as a result of injurious
emissions. Acta Univ. Agric-Brno Ser. C 42:25-46, 1973.
Vogl, M. , S. Bortitz, and H. Polster. Physiological and biochemical contributions to research
of fume damage. 6th report. Definitions of degrees of damage and forms of resistance
against the S02 Component of injurious fumes. Biol. Zentralb. 84:763-777, 1965.
Wainwright, M. Microbial S-oxidation in soils exposed to heavy atmospheric pollution. Soil
Biol. Biochem. 11:95-98, 1979.
Webb, R. A. Use of the boundary line in analysis of biological data. J. Hort. Sci.
47:309-319, 1972.
West, D. C. , S. B. Mclaughlin, and H. H. Shugart. Simulated forest response to chronic air
pollution stress. J. Environ. Qual. 9:43-49, 1980.
White, K. L. , A. C. Hill, and J. H. Bennett. Synergistic inhibition of apparent
photosynthesis rate of alfalfa by combinations of sulfur dioxide and nitrogen dioxide.
Environ. Sci. Techno 1. 8:574-576, 1974.
Wilhour, R. G. Growth response of selected small grain and foliage crops to sulfur dioxide.
U.S. Environmental Protection Agency-CERL Report, 1978. Mimeo. 4 pp.
Winner, W. E., J. D. Bewley, H. R. Krouse, and H. M. Brown. Stable sulfur isotope analysis of
S0? pollution impact on vegetation. Oecologia 36:351-361, 1978.
Winner, W. E., and J. D. Bewley. Terrestrial mosses as bioindicators of SO,, pollution stress.
Synecological analysis and the index of atmospheric purity. Oecologia 35:221-230, 1978a.
Winner, W. E., and J. D. Bewley. Contrasts between bryophyte and vascular plant synecological
responses in an S0?-stressed white pine association in central Alberta. Oecologia
33:311-325, 1978b.
Wu, L. , and A. D. Bradshaw. Aerial pollution and the rapid evolution of copper tolerance.
Nature 238:167-169, 1972.
Yopp, J. H. , W. E. Schmid, and R. W. Hoist. Determination of maximum permissible levels of
selected chemicals that exert toxic effects on plants of economic importance in Illinois.
Illinois Institute for Environmental Quality, 1974. 272 pp.
Zahn, R. Effects of sulfur-dioxide on vegetation: results of gas exposure experiments.
Staub 21:56-60, 1961.
Zahn, R. Uber den Einfluss verschiedener Umweltfaktoren auf die Pflanzenempfindlich-keit
gegenuber Schwefeldioxyd. Z. Pflanzenkr. (Pflanzenpathol.) Pflanzenschutz 70:81-95,
1963a.
SOX8B/B 8-88 2-10-81
-------
Zahn, R. Utersuchungen uber die Bedeutung Kontinuierlicher and interim'ttierender
Schwefeldioxideinwirkung fuer die Pflanzenreaktion. (Investigations on plant reaction to
continuous and/or intermittent sulfur dioxide exposure.) Stub. 23:334-352, 1963b.
Zahn, R. The effect on plants of a combination of subacute and toxic sulfur dioxide doses.
Staub 30:20-23, 1970.
2-
Ziegler, I. The effect of SO- on the ability of ribulose-l,5-disphosphate carboxylase in
insulated spinach chloroplasts. Planta (Berlin) 103:155-163, 1972.
Ziegler. The effect of air polluting gases on plant metabolism. Environ. Qual. Safety
2:182-208, 1975.
Zimmerman, P. W. , and W. Crocker. Toxicity of air containing sulfur dioxide gas. Contrib.
Boyce Thompson Inst. 6:445-470, 1954.
SOX8B/B 8-89 2-10-81
------- |