United Stawes 

Department of uation of the RO e of Ozone, 

Agriculture 

sewice position, and other Airborne 

ts in the Forests of 

Eastern North America 

Southeastern Forest 

Experiment Station J. H. B. Garner, Terry Pagano, and Ellis B. Cowling 

General Technical 

Repoft SE-59 
This paper was submitted by the authors in camera-ready form, 

and authors are responsible for its content and accuracy. 

December 1989 

Southeastern Forest Experiment Statican 

PO. Box 2680 

klaeville, Nofih Carolina 28802 
An Evaluation of the Role of Ozone, Acid 

Deposition, and Other Airborne Pollutants in the 

Forests of Eastern North America 

J.H.B. Garner 

Ecol ogi st 

Environmental Criteria and Assessment Office 

U. S. Environmental Protection Agency 

Research Triangle Park, North Carolina 27711 U.S.A. 

Terry Pagano 

Research Assistant 

Envi ronmental Pol 1 uti on 

Control Program 

Pennsylvania State University 

University Park, Pennsylvania 16802 U.S.A 

Ellis B. Cowling 

Associate Dean for Research 

School of Forest Resources 

North Carolina State University 

Raleigh, North Carolina 27695 U. S.A. 

ACKNOWLEDGEMENTS 

The authors especially acknowledge the continuing assistance of Dr. David 

f. Shriner of the Environmental Sciences Division, Oak Ridge National 

Laboratory, Oak Ridge, TM and Ms. Kimberly C. Joyner, North Carolina State 

University, Atmospheric Impacts Research Program, Raleigh, NC for their 

valuable comments, suggestions and recommendations throughout the writing of 

this document. The authors also express their appreciation to Dr. Paul A. 

Altshuller, Atmospheric Research and Exposure Assessment Laboratory, U.S. 

Environmental Protection Agency, Research Triangle Park, N. C. ; Dr. David A. 

Bennett, Office of Solid Waste and Emergency Response, U.S. Environmental 

Protection Agency, Washington, DC; Dr. Dan Binkley, Department of Forest and 

Wood Sciences, Colorado State University, Ft. Collins, CO; Dr. Ronald Bradow, 

Visiting Professor, Department of Marine Earth and Atmospheric Sciences, North 

Carol ina State University, Raleigh, NC; Mr. Bruce Hicks, National Oceanic and 

Atmospheric Administration, Air Resources Laboratory, Atmospheric Turbulence 

and Diffusion Division, Oak Ridge, TN; Dr. Beverly A.H. Marie, Department of 

Horticultural Science, University of Guelph, Ontario, Canada; and Dr. Wi 11 iam 

E. Winner, Department of General Sciences, Oregon State University, Corvallis, 

OR for their comments, suggestions, recommendations and, in particular, their 

carefully annotated reviews of the document. 

The writing of this document was supported by funds provided by the 

Southeastern Forest Experiment Station, Southern Commercial Forest Research 

Cooperative of the Forest Response Program. The Forest Response Program, part 

of the National Acid Precipitation Assessment Program, is jointly sponsored by 

the USDA Forest Service, U.S. Environmental Protection Agency, and the National 

Council of the Paper Industry for Air and Stream Improvement. 

iii 
DISCLAIMER 

This document has not been subject to Forest Service policy review and 

should not be construed to represent the policies of the Forest Service or of 

the National Council of the Paper Industry for Air and Stream Improvement. 

This document has, however, been reviewed in accordance with U.S. Environmental 

Protection Agency policy and approved for publication. Mention of trade names 

or commercial products does not constitute endorsement or recommendation for 

use. 
CONTENTS 

Page 

INTRODUCTION ........................................................ 1. EXECUTIVE SUMMARY ............................................... 

1.1 Introduction .............................................. 1.2 Purpose of This Document .................................. 1.3 Airborne Chemicals of Interest ............................ 1.4 Airborne Chemical Pollutant Exposure of Forests in Eastern 

North America ............................................. 1.5 Critical Loadings of Airborne Pol lutant Chemicals ......... 1.6 General Conclusions ....................................... 

2. EMISSIONS OF AIRBORNE SULFUR- AND NITROGEN-DERIVED CHEMICALS 

OVER EASTERN NORTH AMERICA ...................................... 

Introduction ............................................. What Sulfur- and Nitrogen-Containing Compounds are 

Emitted into the Atmosphere of Eastern North America ...... What are the Principal Natural and Man-made Sources of 

These Compounds ........................................... In What Amounts and in What Chemical and Physical Forms 

are These Compounds Emitted in the Various Regions 

within North America? ..................................... What Temporal Changes in Emissions of Sulfur- and 

Nitrogen-derived Compounds Have Occurred During the 

Industrialization and Electrification of Various Regions 

of Eastern North America? ................................. In What Directions and Over What Distances are Major 

Airborne Nitrogen and Sulfur Compounds Usually Dispersed 

from Their Sources? ....................................... What Chemical and Physical Transformations of Sulfur, 

Nitrogen, and Volatile Organic Compounds Occur During 

Transport from Natural and Man-made Sources and 

Deposition into Forest Ecosystems? ........................ 

3. CHEMICAL EXPOSURE OF FORESTS IN EASTERN NORTH AMERICA ........... 

3.1 Introduction .............................................. 3.2 What are the Most Important Airborne Sulfur- and 

Nitrogen-derived Pollutant Chemicals to which Forests 

are Exposed in Eastern North America? ..................... 3.3 By What Physical Processes are Airborne Pollutant 

Chemicals Transferred from the Atmosphere into the 

Forests of Eastern North American ......................... 
Page 

3.4 What Physical and Meteorological Features of the 

Terrain on which Forests Grow and the Physical Form 

of Forest Canopies Influence the Amounts of Airborne 

Pollutant Chemicals that are Delivered into Forest 

Ecosystems in Various Parts of Eastern North America ...... 3.5 In What Concentrations and During What Times of Day 

and Seasons of the Year are Sulfur- and Nitrogenderived Compounds Deposited in Forests in Various 

Regions of Eastern North America .......................... 

4 . EFFECTS OF AIRBORNE SULFUR- AND NITROGEN-DERIVED POLLUTANT 

CHEMICALS ON FOREST ECOSYSTEMS OF EASTERN NORTH AMERICA ......... 

4.1 How do Forest Ecosystems Respond to Changes in the 

Physical and Chemical Climate? ............................ 4.2 What is the Likelihood That Acid Deposition and Other 

Airborne Sulfur- and Nitrogen-Derived Chemicals Affect 

the Forests of Eastern North America Through FoliageMediated Response Mechanisms? ............................. 4.2.1 By What Physical, Chemical. and Biochemical 

Processes Are These Compounds Taken up by Plants. 

and Microorganisms ................................ 4.2.2 What is the Mode of Action of the Different 

Airborne Chemical Pollutants? ..................... a) Sulfur Compounds .............................. b) Nitrogen Compounds ............................ C) Ozone and Hydrogen Peroxide ................... d) Mixtures of Gaseous Pollutants ................ e) Acidic Deposition ............................. 4.2.3 Is Photosynthesis in Forest Trees Altered by 

Exposure to Airborne Chemical Pollutants? ......... a) Sulfur Compounds .............................. b) Nitrogen Compounds ............................ c) Ozone ......................................... d) Mixtures of Gaseous Pollutants ................ 4.2.4 Is Respiration in Forest Trees Altered by 

Exposure to Airborne Chemical Pollutants? ......... 4.2.5 Are Carbon Uptake and Allocation Altered by 

Exposure to Airborne Chemical Pollutants? ......... a) Sulfur Compounds .............................. b) Nitrogen Compounds ............................ c) Ozone ......................................... d) Mixtures of Gaseous Pollutants ................ 4.2.6 Is Increased Fol iar Leaching in Forest Trees a 

Response to Airborne Chemical Pollutant 

Exposure? .......................................... a) Ozone ......................................... b) Acidic Deposition ............................. 
4.2.7 Does Exposure to Airborne Chemical Pollutants 

Disrupt the Processes of Regeneration and 

Reproduction in Trees? ............................ a) Sulfur Compounds .............................. b) Nitrogen Compounds ............................ c) Ozone ......................................... 4.2.8 Is Host Susceptibility to Pests, Pathogens, and 

Mycorrhizal Colonization Altered by Exposure to 

Airborne Chemical Pollutants? ..................... a) Sulfur Compounds .............................. b) Ozone ......................................... 4.3 Do Sulfur- and Nitrogen-Derived Pollutant Stresses Affect 

the Growth of Forest Trees? ....................... a) Sulfur Compounds .............................. b) Nitrogen Compounds ............................ c) Ozone ......................................... d) Mixtures of Gaseous Pollutants ................ 4.4 What is the Likelihood That Airborne Chemical Pollutants 

Affect Forest Ecosystems Through Soil-Mediated Response 

Mechanisms? ............................................... 4.4.1 Airborne Gases, Fine Aerosols, and Coarse 

Particles? ........................................ 4.4.2 Acidic Deposition ................................. 4.5 What Effects Are Airborne Chemical Pollutant Stress 

Having on Forest Ecosystems of Eastern North America? ..... 

5 . REFERENCES ...................................................... 
INTRODUCTION 

Purpose of this Document 

In recent years, there has been increasinq scientific and public concern 

that airborne pollutant chemicals might be causing some of the visible injuries 

and decreases in growth of forests that have been reported in various parts of 

North America. This concern was further stimulated by reports of widespread 

visible symptoms and increased mortal i ty of forests in central Europe. Sul fur 

dioxide, ozone, nitrogen oxides and acid deposition are all suspected to be 

among the airborne pollutant chemicals that could be involved. 

One of the major objectives of current forestry research is to determine 

if wet and dry deposition of acidic and acidifying substances derived from 

short- and long-distance transport of airborne sulfur and nitrogen oxides 

could account for the changes in the condition of forests observed in the 

eastern United States. These changes i ncl ude decreased growth, increased 

visible symptoms of injury, and mortality of high-elevation spruce-fir forests; 

increased visible injury and mortality of sugar maple and other eastern hardwood forests; and decreased radial growth of southern pine forests without 

visible symptoms of injury. This objective has it origins in the public and 

scientific concern about "acid rain. " Acid deposition is known to be causing 

detrimental effects in aquatic ecosystems; however, convincing scientific 

evidence that acid deposition is also harming forests is lacking. 

The Forest Response Program (FRP) of the National Acid Precipitation 

Assessment Program (NAPAP) is a major interagency effort initiated speci f ical ly 

to investigate effects of acidic deposition and associated pollutants on 

forests and forest tree species. The Southern Commercial Forest Research 

Cooperative functions under this umbrella. This document was conceived to 

assist the Cooperative in resolving questions relating to the prevai 1 ing 

scientific judgement that ozone and other photochemical oxidants, either alone 

or in combination with acid deposition, are the airborne pollutant chemicals 

most likely to affect southern commercial forests, eastern hardwood forests, 

and possibly also eastern spruce-fir forests; and the prevai 1 ing publ ic 

perception that emissions of sulfur and nitrogen oxides leading to increased 

deposition of acidic and acidifying substances are the ai rborne pol 1 utant 

chemicals most likely to adversely affect these same forests. 

Approach 

The approach used in preparing this document has been to develop a . . 

critical review of publ ished scientific information that is pertinent -to a 

series of specific scientific and policy questions suggested to us by leaders 

in the Southern Commercial Forest, Eastern Spruce-Fir and Eastern Hardwood 

Research Cooperatives within the Forest ~espohse Program or that we developed 

ourselves. This document was begun in 1986 and submitted during the last half 

of 1988. The references and information used in this document, therefore, 

reflect this time frame. No data from the Forest Response Program were yet 

available although many research projects were underway. 

This document concentrates on the effects of airborne pollutant chemicals 

on trees and in forest ecosystems of eastern North America by way of foliagemediated response mechani sms. Many control 1 ed fumigation studies have exposed 

viii 
seedlings in chambers to determine the exposure-responses of a variety of trees 

to different pollutant chemicals. With the exception of some of the studies 

dealing with physiological responses such as photosynthesis and carbon allocation, the authors have based their statements on currently available field 

data. This document is being published at the present time because it provides 

a summary of the state of knowledge of the responses of forest trees to aboveground chemical pollutant exposures and the possible forest ecosystem effects 

resulting from them. This information concerning pollutant effects on the 

forests of eastern North American has not been previously presented in a single 

document. 

Though this document concentrates on the responses of trees to exposure 

to ai rborne chemical pol 1 utants entering through the 1 eaves, trees are a1 so 

exposed through the soil and roots. Exposures to acid deposition and its 

sulfate and nitrate components via soil processes are discusssed in this 

document only briefly. Binkley and others (1989) have discussed in detail the 

impact of acid deposition on forest soils in the southeastern United States. 

Whenever possible, the responses to each question have been formulated 

as simple statments that concisely present the scientific understanding about 

the question. Each statement is followed by references from the peer-reviewed 

scientific literature from which the statement was derived, except in those 

few i nstances where such peer-reviewed pub1 ications were not avai lab1 e. The 

authors have sought critical review of this document and each of its component 

sections by scientists whose knowledge and expertise is greater than our own. 

Our hope is that this document wi 11 provide an understanding of the probable 

role of acid deposition, photochemical oxidants, and other airborne sulfurand nitrogen-derived pollutant chemicals in the growth and development of 

forests in eastern North America and will be of assistance to the Forest 

Response Program and other divisions of the National Acidic Precipitation 

Assessment Porgram who are currenty developing a comprehensive assessment of 

air pollutant effects on forest ecosystems. 

Airborne Chemicals of Interest 

The airborne chemicals addressed in this document are those known to be 

or presumed to be injurious to forest trees or ecosystems. Primary air 

pollutant chemicals are emitted as gases or other volatile waste products from 

mobi 1 e or stationary sources di rectly into the atmosphere where they may be 

subsequently transformed chemically into secondary pollutants. Ecosystems may 

be exposed to either primary or secondary pollutants or both when deposited 

from the atmosphere as gases and aerosols, and as both dissolved and suspended 

substances in precipitation, fog, and cloud water. Determination of ecosystem 

exposure and response, therefore, requires knowledge of the source, chemical 

form, transport and deposition of the primary airborne pol 1 utants emitted 

into the atmosphere. When primary pol 1 utants , are converted into secondary 

pollutants, knowledge of their transport and deposition also is required. 

Concentration, residence time in the atmosphere and time of year the exposures 

occur are also important when determining the effects of ecosystem exposure. 

The airborne chemicals addressed in this document include gases, aerosols, and 

both dissolved and suspended substances in precipitation. The specific 

compounds of interest include: 
gaseous sulfur dioxide, nitric oxide, and nitrogen dioxide; 

.I aerosol s and precipitation containing sul fate, nitrate, 

ammonium, and hydrogen ions; 

gaseous ammonia and nitric acid vapor; 

* gaseous ozone, hydrogen peroxide, peroxyacetyl nitrate (PAN), 

peroxypropionyl nitrate (PPN), and other photochemical oxidants. 

Organization of the Document 

This paper is organized in four distinct chapters: 

Chapter 1 -- Executive Summary 

Chapter 2 -- Emissions of Airborne Sulfur- and Nitrogen-Derived 

Chemicals over North America 

Chapter 3 -- Chemical Exposure of Forests in Eastern North America 

Chapter 4 -- Effects of Acid Deposition, Photochemical Oxidants, and 

Other Airborne Sulfur- and Ni trogen-Derived Pol 1 utant 

Chemicals on Forest Ecosystems. 

Chapters 2 to 4 are organized around the series of specific scientific 

questions about the role of acid deposition, photochemical oxidants, and other 

airborne sul fur- and ni trogen-derived pol 1 utant chemical s in the forests of 

eastern North America. These questions are designed to show the possible 

relationship between the increasing emissions of airborne pollutant chemicals 

and the effects observed in the forest ecosystems. 

These questions are covered in Chapter 2: 

What sul fur- and ni trogen-contai ni ng compounds are emitted into 

the atmosphere of eastern North America? 

- What are the principal natural and man-made sources of these 

compounds? 

In what amounts and chemical and physical forms are these 

compounds emitted in various regions within North America? 

What temporal changes in emissions of sulfur- and nitrogenderived compounds have occurred during the industrialization and 

electrification of various regions within eastern North America? 

In what directions and over what distances are major airborne 

nitrogen and sulfur compounds usually dispersed from their 

sources? 
* What chemical and physical transformations of sulfur and nitrogen compounds and volatile organic compounds occur during 

transport from natural and man-made sources and deposition into 

forest ecosystems? 

These questions are covered in Chapter 3: 

* What is the chemical exposure of forests in various regions of 

eastern North America? 

* In what chemical and physical forms and by what physical processes are sul fur- and ni trogen-derived pol 1 utants deposited 

into the forests of eastern North America? 

What is the extent of geographical and elevational variation in 

deposition of these compounds in various types of forests across 

eastern North America? 

- In what concentrations and during what times of day and seasons 

of the year are these compounds deposited in forest ecosystems? 

These questions are examined in Chapter 4: 

How do forest ecosystems respond to changes in the physical and 

chemical climate? 

* By what physical, chemical, and biological processes are these 

compounds taken up by plants, and microorganisms in forest 

ecosystems, and what interactions exist among them? 

What is the likelihood that acid deposition and other airborne 

sul fur- and ni trogen-derived chemical s affect the forests of 

eastern North America through foliage-mediated response 

mechanisms? 

a What is the likelihood that airborne chemical pollutants affect 

forest ecosystems through soil-mediated response mechanisms? 

* What effects are airborne chemical pollutant stresses having on 

forest ecosystems of eastern North America? 

1. EXECUTIVE SUMMARY 

I. 1 Introduction 

Natural forest ecosystems in the United States have been exposed t o t o 

ai rborne pol 1 utants for several decades as i ndustri a1 i zati on, el ectri f i cation, urbanization and automotive transportation have increased. The chemical 

pollutant emissions associated with these sources have increased as the number 

of sources increased. The effects on forest trees of exposures to the chemical 

emissions have not been evaluated or quantified except in Southern California 

and on white pine in the East. Changes in the condition of forests in various 

parts of eastern North America, however, have been observed. Decreased growth, 

dieback, and mortality of certain tree species have been reported for highelevation forest ecosystems from Maine, New Hampshire, Vermont and New York 

south to North Carolina and Tennessee. Decreased growth in natural stands of 

ye1 low pines in the Piedmont of Georgia, and of North and South Carolina has 

also been reported. Dendroecological studies suggest that the decreases in 

growth of forest trees in the eastern United States began during the 1950's or 

early 1960's. The extent of forest decline and the factors that precipitated 

them, however, are questions of controversy. The problem is determining 

causation. Many hypotheses have been advanced as to the possible cause or 

causes (see table 1 and chapter 4). 

. Documentation of effects on forest trees is lacking particularly because 

thei r size and perennial nature are not conducive to short-term manipulative 

experiments that make inference of a causal relationship possible. A causal 

relationship may be inferred when there is a strong pattern of consistency, 

responsiveness, and a proven biological mechanism. Any two, of the above three 

linkage patterns may be sufficient to infer cause in simple systems, however, 

more compl icated systems may require a1 1 three. To determine-consistency, the 

symptoms of injury or disfunction must be associated consistently with the 

presence of the causal factor. To determine mechanism, the suspected cause 

must be related to an observed effect through one or more biological processes. 

Responsiveness is the most difficult to prove in mature ecosystems as it 

requires development of an exposure-response relationship based on the exposure 

of healthy trees to various known concentrations of airborne pollutants under 

controlled conditions that simulate those in the forest (see chapter 4). 

Responses of seedlings exposed to known concentrations of air pollutants 

under controlled conditions cannot be readily extrapolated to mature trees 

(see chapter 4). Responses of mature forest trees to stresses resulting from 

exposure to ambient air pollutants are difficult to determine because they are 

usually subtle growth responses that are the cumulative result of many years 

of multiple exposures. Altered growth dynamics of trees over time may, 

depending on the severity of the response (e. g. , mortal i ty, 1 oss in competitiveness), bring about changes in forest community growth and development. 

Such stress responses can also be expected to be strongly influenced by site 

variables, such as soil chemistry, the developmental state of the stand, and by 

cl imatic factors such as temperature and rainfall. These factors have made it 
Table 1. Some Hypotheses to Explain Forest Decline in Eastern North America 

Category Hypothesis 

Natural process Synchronous canopy dieback 

Biotic pathogens 

Assorted abiotic stress agents 

Forest stand history Landscape conversion 

Forest management practices (e. g. , genetic stock) 

Forest fire frequency and i ntensi ty 

Anthropogenic stress Nitrogen subsidy (e. g. , nitric acid vapor, 

nitrate deposition) 

Direct ozone effects (e.g. , inhibition of 

photosynthesis) 

Indirect ozone effects (e. g. , leaching) 

Direct effects of hydrogen ion deposition 

Indirect effect of hydrogen ion deposition 

(e. g. , rhi zosphere a1 umi num toxicity) 

Deposition of sulfur 

Deposition of organics 

Deposition of multiple pollutants 

(wet and dry deposition) 

Deposition of trace elements 

Acid cloud moisture 

Mu1 tiple factors Site-specific combination of assorted natural 

processes, forest stand history, and anthropogenic stress 

Adapted from Taylor and Norby (1985). 

difficult to determine the effects of air pollutants on forest trees and on the 

ecosystems of which they are a part (see chapter 4). 

The following information explains the purpose of this document and 

presents the analyses and conclusions of the authors based on the data 

available. 

1.2 Purpose of This Document 

This document was developed to help resolve the apparent conflict between: 

a) The prevailing scientific judgment that ozone and other photochemical oxidants, either alone or in combination with acid 

deposition, are the ai rborne pol 1 utant chemical s most 1 i kely to 

adversely affect southern commercial forests, eastern hardwood 

forests, and possibly eastern spruce-f i r forests ; 

b) the prevai 1 ing pub1 ic perception that emissions of sulfur and 

nitrogen oxides leading to increased deposition of acidic and 

aci di fyi ng substances are the ai rborne chemical s most 1 i kely to 

adversely affect these same forests. 
1.3 Airborne Chemicals of Interest 

Forest ecosystem responses to stresses, because they are the summation of 

multiple responses of individual organisms and their interactions, usually 

require at least a generation (30-50 years) before their expression can be 

observed unless they are the result of a catastrophe. Therefore, determining 

the effects of airborne pollutant chemicals on forest ecosystems requires a 

knowledge of the sources, chemical composition and volume of emissions and 

whether these have changed over time. It also requires an understanding of 

the transformation, transport, atmospheric duration, deposition and uptake 

phenomena associated with the movement of airborne chemical pol 1 utants into 

leaves of forest trees. Meteorological features of the terrain on which 

forests grow and time of year when exposures occur also are important factors 

when determining responses. 

For these reasons, the airborne chemicals addressed in this document 

include the primary and secondary pollutants that are transferred from the 

atmosphere into forest ecosystems in the form of gases and aerosols (dry 

deposition), and both di ssol ved and suspended substances in precipitation, fog and cloud water (wet deposition). Primary pollutants are gases or other 

volatile waste products emitted directly from stationary or mobile sources. 

Secondary pollutants are formed in the atmosphere by chemical transformation 

of primary pol 1 utants. 

The three most important primary air pollutants are sulfur dioxide (SO ), 

nitrogen oxides (NO and NO), and volatile organic compounds (VOC). ~mal lgr 

amounts of gaseous immonia, sulfuric acid, sulfate aerosol, and hydrochloric 

acid also are emitted as primary pollutants (see chapter 2). 

The most important secondary pol 1 utants i ncl ude two general cl asses of 

compounds -- photochemical oxidants and acid deposition. Photochemical 

oxidants include ozone ((I3), hydrogen peroxide (H202), peroxyacetyl nitrate 

(PAN), and peroxypropionyl nitrate (PPN). 

The constituents of acidic and acidifying deposition include 

gaseous sulfur dioxide, nitrogen oxides, nitric acid, and 

ammonia; 

fine and coarse aqueous aerosols containing sulfuric and nitric 

acids and all the major nutrient cations and anions including 

sulfate, nitrate, phosphate, chloride, potassium, sodium, 

magnesium, ammonium, and hydrogen ions; and 

precipitation and cloud water containing dissolved or suspended 

matter including all of these major nutrient cations and anions 

-- sulfate, nitrate, phosphate, chloride, potassium, sodium, 

magnesium, ammonium, and hydrogen ions (see chapter 2 and 

chapter 3). 
1.4 Airborne Chemical Pollutant Exposure of Forests in Eastern North 

Amer i ca 

Forests cover more than 50 percent of the total land area of eastern 

North America. The average distance between major stationary sources and urban 

centers in the eastern United States is considerably less than the average 

dai 1 y di stance (100-200 km) that both primary and secondary pol 1 utants are 

transported. Thus, most forests in eastern North America (with the exception 

of the most northerly portions of Ontario and Quebec) are within the area 

where exposure to either primary and secondary pollutants or both could occur. 

Proximity to pollutant sources is particularly important in determining 

the exposure of forests to sulfur dioxide, nitrogen oxides and pollutant 

mixtures. Ozone, however, can be transported for hundreds of miles and be 

enriched as the air masses move over urban and other areas where additional 

ozone or its precursors are present. Sulfates and nitrates, secondary pollutants formed from SO and NO , also may be transported long distances before 

they are deposited &to fores? ecosystems. 

With the possible exception of ozone and hydrogen peroxide, a1 1 of the 

primary and secondary airborne chemicals of interest in this document, depending 

on circumstances, can have either stimul atory or inhibitory growth effects on 

forests: 

- - stirnulatory effects at low concentrations or loadings; and 

- - inhibitory effects at high concentrations or loadings. Unfortunately, 

the demarcation between growth-stimulating and growth-inhibiting 

chemical exposures is not known for most airborne pollutant chemicals 

(see chapter 3). 

In characterizing the chemical exposure of forests of eastern North 

America to atmospheric pollutants, we found it useful to recognize five general 

types that, because they differ appreciably in proximity to major sources of 

pol lutants, or are subject to distinctive chemical meteorology, could differ 

significantly in terms of pollutant exposure. These types include the 

fol lowing: 

1) Urban forests such as the 30,000 hectares of forest land within 

the city limits of Atlanta, GA. Although urban forests make up 

only a very small percentage of the total forest area of eastern 

North America, these areas contain about 30 percent of the 

people of the United States and Canada and thus are of interest 

to many citizens in both countries; 

2) Rural forests near major point sources (for example, within 

approximately 30 km of a major power plant or metal smelter). 

We estimate that these areas make up about 5 percent of the 

total forest area of the eastern United States and southeastern 

Canada; 
3) Near-urban forests within about 50 km of major area sources of 

pollution. We estimate that these near-urban areas could 

include as much as about 50 percent of the total forest area of 

the eastern United States and southeastern Canada; 

4) Rural forests more than about 50 km from major urban areas or 

other large area sources of pollution. We estimate that these 

areas make up about 25-30 percent of the total forest area of 

the eastern United States and southeastern Canada; 

5) High elevation forests such as those on Mount Mitchell in North 

Carolina or Whiteface Mountain in New York (see fig. 3). We 

estimate that these areas make up less than 1 perc6nt of the 

total forest area of the eastern United States and southeastern 

Canada (see chapter 3). 

The exposure of high-elevation forests (>1,000 meters above sea level) to 

pollutant chemicals is quite different from that of low elevation forests. The 

reasons for this are listed below. 

1) Precipitation inputs are much greater than at low elevation; 

2) Cloud-water inputs at high elevations significantly augment 

preeipitaLion inputs; 

3) Concentrations of dissolved and suspended substances in fog and 

cloud water are typically 6-10 times greater than in precipitation; 

4) Ozone and hydrogen peroxide concentrations increase directly 

with altitude. Maximum ozone concentrations observed at elevated 

mountainous sites , as we1 1 as at many non-mountai nous rural and 

remote sites, often occur at night. This is particularly 

important for species such as eastern white pine whose stomata 

remain open at night (see chapter 4). 

5) Sites at higher elevations are often exposed to sustained or 

multiple peak concentrations of ozone within a given 24-hour 

period as a result of conditions such as (a) trapping inversions; 

(b) the successive transport of ozone from multiple urban 

sources upwind, either aloft or across terrain devoid of 

sufficient ozone scavengers; and (c) the occurrence of mountainvalley and upslope-downslope flows, such that the trajectory of 

an air parcel passes back over the same forest or stands of 

trees (see chapter 4). 

6) Wind speeds are greater than at low elevation and thus augment 

aerosol impaction rates; and 

7) Evaporation of fog and cloud water further increase the concentrations of airborne chemicals to which these forests are 

exposed (see chapter 3). 
Tables 2 and 3 show the range in amounts of the various pollutants to 

which forests of the five types listed above may be exposed in the southeastern 

and northeastern States of the United States and southeastern provinces of 

Canada (see chapter 3). 

In developing the data on chemical exposure of forests as summarized in 

tables 2 and 3 and in drawing certain of the General Conclusions 1 isted later 

in this Executive Summary, we have found it useful to combine the data on 

chemical exposure of urba6 forests and rural forests near major point sources, 

and also the data on exDosure of near urban and rural forests. 

In addition, analyses of data on pollutant exposure indicate the 

fol l owi ng: 

1) Gaseous sulfur dioxide rarely occurs in concentrations sufficient 

to cause visible injury to vegetation (50 to 5,000 ppb for 

8 hours or 500 to 3,000- ppb for 1 hour) in - remote, - rural, or 

near-urban forests in eastern North America. (The occurrence 

of vi si bl e i njury does not necessari ly i n d i c a t e that decreases 

in growth or significant changes in the health or productivity 

of whole forest stands will occur). In high-elevation forests 

sulfur dioxide concentrations are usually near the limits of 

detection, Injurious concentrations do occur in some urban and 

in rural forests near major pollution sources. Maximum concentrations of 500 to 1,000 ppb or even higher were measured 

near industrial sources for several days durjng the 1970's and 

early 1980's (see chapter 3). 

2) Nitrogen oxides are rarely, if ever, found in concentrations 

sufficient to cause visible injury to vegetation (2,000 to 

5,000 ppb for 8 hours or 4,000 to 8,000 ppb for 1 hour) in 

eastern North America (see chapter 3). This is true in part 

because nitrogen oxides are often consumed in photochemical 

oxidant formation. 

33 Ozone occurs in concentrations sufficient to cause visible 

injury to vegetation in most of eastern North America. For 

example, episodes lasting from hours to weeks and exposing 

eastern North America from the Midwest to the Atlantic Coast 

with concentrations ranging from 60 to a maximum of 300 ppb have 

been reported. The highest 1-hour peak concentrations of ozone 

in the eastern United States occur in Mew England, the midAt1 antic States and the Gul f Coast. (Peak concentrations 

) From 1981 to 

1983, peak concentrations of 170, 180, 250 and 280 ppb were 

reported for Raleigh, NC; Boston, MA; Newark, NJ; and Houston, 

TX, respectively (see chapter 3). 

4) Co-occurring and sequential exposure to sulfur dioxide, nitrogen 

oxides, and ozone have been investigated using 30 monitoring 

sites in the eastern United States. Sulfur dioxide and ozone 

co-occurred approximately 30 percent of the time. Sul fur 

dioxide and ozone co-occurred because ozone was frequently 
Table 2. Typical range of chemical exposure in forests in the Southeastern 

United States reported in refereed publications 

Urban forests & 

Rural Forests 

Near Major Near-Urban & High-Elevation 

Pollutant Chemical Sources Rural Forests Forests 

Sulfur dioxide (ppb) ~5-800 (5-200 

Nitrogen oxides (ppb) (5-200 < 5- 100 

Photochemical Oxidants 

Ozone (ppb) (20-250 (20-200 

Hydrogen peroxide in 

cloud water 

(pmol/L) 

Unknown Unknown 

Nitric acid (ppb) (1-10 tl-5 

PAN (ppb) Not reported Not reported 

PpN (PP~) Not reported Not reported 

Sulfate ion (kg/ha/yr): 

from sulfur dioxide Unknown 

in precipitation 10-20 

in cloud water - - 

in sulfate aerosol 10- 15 

Nitrate ion (kg/ha/yr): 

from nitrogen oxides Unknown 

in precipitation 5-15 

in cloud water - - in nitrate aerosol 5-15 

Ammonium ion (kg/ha/yr): 

from ammonia Unknown 

in precipitation 1-2 

in cloud water - - 

in ammonium aerosol 1-2 

Hydrogen ions (g/ha/yr) : 

from HNO, & HC1 Unknown 

in precipitation 100-200 

in cl oud water - - 

in aerosol 100-200 

Unknown 

10-20 

- - 

Unknown 

5- 10 

- - 

5- 10 

Unknown 

1-3 

Unknown 

100-200 

- - 

100-200 

< 1- 3 

Mot reported 

Not reported 

Unknown 

10-20 

10- 15 

10- 15 

Unknown 

10-40 

5-20 

5- 10 

Unknown 

1-2 

1-2 

1-2 

Unknown 

300-400 

300-400 

300-400 
Table 3. Typical range of chemical exposure in forests in the Northeastern 

United States and Southeastern Canada reported in refereed publications 

Urban Forests & 

Rural Forests 

Near Major Near-Urban & High-Elevation 

Pollutant Chemical Sources Rural Forests Forests 

Primary Pollutants (1-hr average values): 

Sulfur dioxide (ppb) t5-1,000 

Nitrogen oxides (ppb) (5-400 

Photochemical Oxidants (1-hr average values): 

Ozone (ppb) (20-200 

Hydrogen peroxide in 

cloud water 

(pmol /L) 

Unknown 

Nitric acid (ppb) (1-10 

PAN (ppb) Not reported 

PPN (ppb) Not reported 

Acidic and Acidifying Deposition: 

Sulfate ion (kg/ha/yr): 

from sulfur dioxide 

in precipitation 

in cloud water 

in sulfate aerosol 

Unknown 

10-35 

Nitrate ion (kg/ha/yr): 

from nitrogen oxides Unknown 

in precipitation 10-20 

in cloud water - - 

in nitrate aerosol 5-15 

Ammoni um i on (kg/ha/yr): 

from ammonia Unknown 

in precipitation 2-5 

in cloud water - - 

in ammonium aerosol 2- 5 

Hydrogen ions (g/ha/yr): 

from HNO, & HC1 Unknown 

in precipitation 400-600 

in cloud water - - 

in aerosol 400-600 

(20- 150 

Unknown 

(1-5 (1-3 

Not reported Not reported 

Not reported Not reported 

Unknown 

10-35 

Unknown 

10-15 

- - 

5-10 

Unknown 

2-5 

Unknown 

400-600 

Unknown 

10-35 

10-20 

10-15 

Unknown 

10-40 

5-25 

5-15 

Unknown 

1- 2 

1- 2 

1- 2 

Un known 

400-600 

300-400 

400-600 
present at concentrations greater than 50 ppb, and concentrations of sulfur dioxide greater than 50 ppb occurred 

periodically. When SO concentrations were 30 ppb or lower, no 

co-occurrences were obzerved (see chapter 3). 

5) Exposure of forests or their components to the photochemical 

oxidants listed below have not been studied, Controlled 

experiments have shown that most woody plants are not sensitive 

to PAN. 

Gaseous hydrogenperoxide; 

Hydrogen peroxide dissolved in cloud water and precipitation; . Nitric acid; 

PAN; and 

PPN. 

6) Exposure of forest trees to gasesous chemical pollutant stresses, 

whether severe acute stresses or long-term chronic stresses 

occurs chiefly through the stomata, the same openings in leaves, 

needles or young shoots through which plants exchange gases with 

the atmosphere during the processes of photosynthesis, respiration and transpiration (see chapter 4). 

7) Transfer of gases from the atmosphere to the physiological sites 

of action within the leaf interior involves both a gas phase and 

a liquid phase. The processes of deposition are the crucial 

links between the transport of atmospheric pollutants to forest 

canopies, their uptake by trees and the manifestation of effects 

Any factor that influences pollutant deposition will change the 

relationship between the concentrations of the pollutant in the 

atmosphere and the corresponding concentration at the sensitive 

site within the leaf cells (see chapter 4). 

8) Gases moving into the leaf encounter concentrations of transpired 

water vapor and a variety of hydrocarbons, (e.g., terpenoids) 

that are produced within the leaf in secondary metabolism, 

volatilized into the gas phase of the leaf interior, and emitted 

through the stomata. These substance are more highly concentrated within the leaf boundary layer and may significantly 

affect the surface and uptake of gases and particles (see 

chapter 4). 

9) Stornatal control of diffusion into the leaf is a major factor 

influencing the exchange of gaseous pollutants between the 

atmosphere and internal leaf tissues. All studies indicate that 

stornatal resistance is the most dynamic and most influential 

resistance to transfer of gaseous ozone, sulfur dioxide and 

nitrogen oxides when the canopy is dry. Stomata1 functioning is 

controlled by both physiological and bioenvironmental processes 

(see chapter 4). 
1.5 

Total 1 oadi ng of sul fate, nitrate, ammoni urn, and hydrogen ions (amount 

deposited into a forest or aquatic ecosystem) is the result of wet and dry 

deposition. The effects of these chemicals on vegetation and forest ecosystems, 

regardless of whether they are added through acid precipitation or in dry form, 

is chiefly through soil. The soil-mediated responses of forests in the 

Southeast to acid deposition are discussed in detail by Binkley and others 

(1989) and are, therefore, not emphasized in this document. 

Evidence now accumulating suggests that the addition of nitrate and 

ammonium to the soil plays a significant role in the growth of forest trees. 

Concern has arisen regarding the possible effects on both terrestrial and 

aquatic ecosystems of these additions. Attempts are being made to determine 

a 'critical lbad. A critical load is defined as the highest load that will not 

cause chemical changes leading to long-term harmful effects on the most 

sensitive ecosystems (see chapter 4). The definition carries with it the 

notion that there is a load at which no long-term effects occur. This concept 

has been used in attempts to determine the effects of sulfate and nitrate 

deposition on ecosystems. Though the effects of sulfate on aquatic ecosystems 

have been studied for many years, critical loadings have not been determined 

for any of the sulfur- or nitrogen-derived constituents of acid deposition. 

Certain important facts concerning the effects of deposition into forest 

ecosystems are listed below. 

1) Exposure of forests is through wet and dry deposition. In 

general, sulfate concentrations are greater in summer than in 

winter. However, in a broad band extending from New England 

southwestward along both sides of the Applachian Mountains as 

far western North Carolina, the ratio of wintertime to summertime sulfate concentration is about two to one (see chapter 3). 

2) Wet deposition of sulfate and nitrate ions in warm periods is 

highly episodic. About 50 to 70 percent of the total annual 

deposition of sulfate and nitrate, occurs in approximately 

20 percent of the total days; however, nitrate deposition is not 

as seasonally dependent. 

3) Ambient deposition of nitrogen is much greater in high- than in 

low- elevation forests, therefore, detrimental effects of total 

nitrogen are more likely in high-elevation forests (chapter 3). 

4) Ambient deposition of ai rborne sul fur and nitrogen compounds 

(including gases, aerosols, and dissolved and suspended substances in precipitation and cloud water) provides amounts of 

nutrient sulfur and nitrogen that probably increase growth of 

forest trees in rural, near-urban, urban, and both high- and 

low-elevation forests in some parts of eastern North America. 

This is true because the total nutrient input from atmospheric 

sources of the several compounds involved provides an important 

part (10 to 30 percent) of the annual ecosystem demand for total 

nutrient sulfur and nitrogen (see chapter 3 and chapter 4). 
5) The total supply of airborne nutrient nitrogen includes the 

fol lowing compounds: gaseous nitrogen oxides + gaseous nitric 

acid, + gaseous ammonia; ammonium + nitrate aerosols; ammonium + 

nitrate ions in precipitation; and ammonium + nitrate ions in 

fog and cloud water; and possibly PAN and PPN. 

6) The total supply of airborne nutrient sulfur includes the 

following compounds: gaseous sulfur dioxide + sulfuric-acid and 

sulfate aerosols + sulfate ions in precipitation and fog and 

cloud water. 

7) Present evidence is not sufficient to accept or reject any of 

the hypothetical mechanisms for action of acid deposition on 

forest tree growth by either foliage- or soil-mediated responses. 

Such evidence does suggest, however, that direct injury to 

foliar organs by ambient acid deposition is much less likely 

than indirect effects mediated through interference with normal 

foliar and root nutrient uptake and leaching processes (see 

chapter 3 and chapter 4). 

1.6 General Goncl usions 

At ambient concentrations and rates of deposition observed in 

the eastern United States, neither acid deposition in the form 

of aerosols nor in the form of dissolved or suspended substances 

in precipitation or cloud water have been shown to induce acute 

detrimental effects on forests or crops via foliage-mediated 

response mechanisms. Deposition of sulfate and nitrate onto 

the soil during acid precipitation, however, can affect tree 

growth (see chapter 4). 

2. The pH of the ambient precipitation (pH 4.0-6.0) in most of the 

eastern United States is substantially higher than the pH of 

simulated acid precipitation necessary to induce visible injury 

to forest trees or commercial crops (pH 2.0-3.5), with the 

exception of one cultivated variety of soybeans (see chapter 3 

and 4). 

3. Gases, fine and coarse aerosols (including fog and cloud water 

droplets), and coarse particulate matter are transferred into 

forests by dry deposition processes. Dry deposition occurs 

under all meteorological conditions, at all times of the day and 

night, and in all seasons of the year. Dry deposition is more 

difficult to quantify than wet deposition (chapter 3). 

4. Toxic gases at ambient concentrations are the only physical 

forms of ai rbsrne pol 1 utant chemicals known to cause vi si b1 e 

symptoms of injury, decreased growth, or mortal i ty of forest 

trees. Toxic gases enter plants mainly through leaf stomata. 

5. In eastern North America, sulfur dioxide rarely occurs in high 

enough concentrations (above 50 ppb for 8 hours) to be injurious 

to forests on a broad regional scale. In addition, in northern 
areas where residential and commercial space and water heating 

are major sources, sulfur dioxide concentrations are usually 

higher during the winter months when most trees are not actively 

growing. 

6. Sulfur dioxide injury to vegetation is found most frequently in 

the vicinity of strong point sources such as metal smelters and 

power plants. Vegetation exposure to phytotoxic concentrations 

in these areas is usually of short duration. Exceptions are 

Wawa and Sudbury, Ontario, Trai 1, Brjtish Columbia, and Copper 

Hill Basin near Ducktown, TN, where catastrophic emissions 

resulted in long-term injury. (see chapters 3 and 4). 

. Based on long-term trends in emissions, it may be inferred that 

sulfur (sulfur dioxide and sulfate) loadings of forest ecosystems 

have remained relatively constant or decreased in the northeastern United States and southeastern Canada since the 1920is, 

but have increased steadily in the southeastern United States 

since 1940's (see chapter 3). 

8. Gaseous nitrogen oxides at ambient concentrations (50 ppb) have 

not been proven to cause visible injury, decreased growth, or 

mortality of either forest trees or agricultural crops at any 

location in North America (see chapter 3 and 4). 

9. Gaseous ozone occurs in eastern North America in concentrations 

sufficient to cause fol iar injury, inhibit photosynthesis, a1 ter 

carbon allocation, and in tree roots, interfere with mycorrhizal 

formation. Disruption of these important physiological and biochemical processes can suppress the growth of trees, shrubs, 

and herbaceous vegetation by decreasing their capacity to form 

carbon compounds and their ability to absorb water and needed 

nutrients from the soil. Ozone has been shown to cause foliar 

injury, decreased photosynthesis and reduced growth in a number 

conifer and hardwood species (see chapter 4). 

10. Ambient ozone concentrations between 40 and 70 ppb, or higher, 

for 1 to 7 days during the growing season have reduced the 

growth of sensitive vegetation, intermediately-sensitive vegetation and some tolerant vegetation. Growth of sensitive tree 

species has been reduced without causing foliar injury. Present 

evidence is not sufficient, however, to determine if ozone is 

causing economically significant injury to the health or 

productivity of whole stands of eastern white pine, other 

eastern conifers or hardwoods (see chapter 4). 

11. Emission trends data indicate a sharp increase for SOX and NOx 

beginning in the late 1940's. Ozone concentrations, fueled by 

large-scale NOx emissions that began during World War 11, 

undoubtably also increased in the late 1940's even though data 

indicating this does not exist. Circumstantial evidence suggests 

that the ecosystems of eastern North America have been subject 

to phytotoxic episodic ozone concentrations, and sulfate and 

nitrate deposition, for at least 30, possibly 40 years (see 

chapter 3 and 4). 
12. Field studies and observations support the view that ozone has 

been a important gaseous pollutant in the eastern United States 

for many years. Vegetation injury , resulting ozone exposure 

was first observed in New Jersey on many cultivated crops, 

ornamentals and indigenous vegetation beginning in 1944, though 

the cause of the injury was not known until 1950 when Heggestad 

and Middleton reported ozone injury of tobacco (see chapter 4). 

In the 19501s, the United States Forest Service reported 

oxidant-induced injury of eastern white pine on the Cumberland 

Plateau of East Tennessee. However, at that time the injury was 

attributed to unidentified atmospheric constituents. Needle 

blight (also called emergence tipburn) of eastern white pine, 

the result of acute and chronic oxidant exposure, was reported 

from West Virginia and western North Carolina in the late 1950's 

and early 1960's (see chapter 4). 

14. Data from studies made since 1976 indicate that ozone episodes 

in eastern North America are a common phenomenon. Typical 

ozone episodes are associated with high pressure systems that 

originate in Canada and move south into the Midwest enroute to 

the Atlantic Ocean and northward into New England. Other 

episodes that affected the entire southeastern United States 

from western Texas northeastward through Illinois, and finally 

east to the Atlantic Ocean have been associated with high 

pressure systems over the Gulf of Mexico. In addition, some 

episodes originating over New York City and moving westward to 

Washington, DC, and Philadelphia to Pittsburgh and south a1 ong 

the Appalachian Mountains to the Gulf Coast of Florida have been 

associated with high pressure systems over the Atlantic Ocean 

(see chapter 3 and 4). 

15. Ozone is ephemeral. Unlike residuals of sulfur or nitrogen 

oxides, f 1 uorides, heavy metal s, or radio-nucl ides, ozone or 

its by-products, are not selectively retained or accumulated in 

any particular components of forest ecosystems (see chapter 4). 

16. Ozone is a relatively homogenous regional pollutant, (chiefly 

due to the sources of its precursors and the processes that 

contribute to its formation). It is the one gaseous airborne 

chemical which is capable of having exposed to phytotoxic 

concentrations without leaving a permanent trace the entire 

eastern United States where visible injury and growth reductions 

of forest trees, crops and other vegetation have been observed. 

It is probably, therefore, that the mu1 tiple episodic ozone 

exposures have been a source of major above-ground stresses on 

forests and other vegetation for many years (see chapter 3 and 

4). 

17. Experiments using sulfuric acid aerosols suggest that exposure 

to acid aerosols may predispose trees to injury by ozone (see 

chapter 4). 
18. Mixtures of toxic gases, fine and coarse aerosols, and coarse 

particulate matter are known to occur occasionally in the 

atmosphere near major point and area sources of pollution -- 

in or near cities or near power plants or metal smelters. 

But actual field measurements indicate that co-occurrence of 

phytotoxic concentrations of gaseous ozone, sulfur dioxide, or 

nitrogen oxides is rare. Thus though present evidence is far 

from conclusive, the risks to forests in eastern North America 

from mixtures of gaseous pollutants and aerosols does not appear 

to be severe (see chapter 3 and 4). 

19. Hydrogen peroxide has been detected in cloud water collected in 

high-el evation forests in eastern North America in concentrations that were sufficient to experimentally cause injury to 

spruce seedlings in The Netherlands. Attempts in North American 

to demonstrated simi 1 ar injurious effects at these same 

concentrations on spruce and pine seed1 ings have failed to 

confirm the observations made in The Netherlands (chapter 3 and 

4) 

20. Tree growth is the culmination of many biochemical and 

physiological processes. Virtually every perturbation of a 

plant community results in stress and affects the performance 

and surivival of individual plants. The age of the tree, the 

complex interactions among pollutant and natural stresses 

combined with various environmental and genetic factors and the 

action these have on physiological and biochemical processes 

within the tree, determine the impact that ambient chemical 

pollutants have on growth (see chapter 4). 

21. Trees are perennial plants. Thus, they must cope with the 

cumulative effects of both short- and 1 ong-term stresses 

encountered year after year. They may respond to some stresses 

rapidly as, for example, when needles of eastern white pine 

exhibit visible injury symptoms within days of exposure to 

acute ozone concentrations. In many cases, however, response to 

stresses is through differential growth, resulting from changes 

in carbon allocation over many years. Therefore, attributing 

growth reductions observed today to air pollution exposures that 

began 30-50 years ago is difficult, because long-term data 

indicating the presence of phytotoxic concentrations of airborne 

chemical pollutants is lacking (see chapter 3 and 4). 

22. Tree vigor is extremely important in determining the response 

of trees to insects and pathogens. Sulfur dioxide and ozone, by 

inhibiting photosynthesis and altering carbon allocation, can 

predispose to attack by insects (especially bark beetles) and 

various fungal pathogens (especial ly root rotting fungi). 

23. Changes within forest ecosystems result from altered energy flow 

and nutrient cycling within the forest communities that compose 

them. These a1 terations begin with the responses of individual 

trees to stresses. Stresses sufficient to cause mortality, to 
severely after growth patterns, or to change the ability of 

trees to compete can set the stage for ecosystem changes 

especially if they lead to changes in species composition (see 

chapter 4). 

24. Mature forest ecosystems maintain themselves in an oscillating 

steady state. Trees are continuously undergoing severe competition for space, light, water, and mineral nutrients. Those 

trees most susceptible to natural stresses, are eliminated more 

rapidly by additional and severe periodic stresses such as 

drought, flooding, fire and attacks by insects or disease-causing 

organisms. Continuous elimination of old trees and those 

susceptible to stresses ultimately produces a mature forest with 

trees and other plant species capable of tolerating the competitional stresses. This usually results in changes in species 

composition of the ecosystem over time. Unless they are 

catastrophic (e. g. , fire, flood, or windstorm), the effects of 

natural stresses upon ecosystems are frequently difficult to 

determine (see chapter 4). 

Transport Transformations Primary and 

(mtxing 

wind speed, 

herght, 

etc.) 

- (sunlight, catalysts) - secondary pollutants - 

t 

Prtmary em~ssrons Dry and wet 

(NO, SO2, HC, C02, HF, C=C, deposit ion 

part~culate, etc.) 

t I 

NO, SO2, 

Air 

Dry and wet 

deposrtion 

I 

Sources HF, C=C, etc. 0,. SO:;NO; etc, 

(Anthropogenic, 

natural) 

Vegetation 

(exposure dose) 

vegetation 

(exposure dose) 

Effects 

(uptake dose) 

Effects 

(uptake dose) ------------ 

Soil 

Processes t(1 stu* : Abiotic Biotic 

Absorption Uptake Biochemicat Organismai 

Adsorption Accumulation Cellular Community 

Impaction Sedimentation Organ Ecosystem 

Figure 1. The emission, transport, transformation, and deposition subsystems. The atmospheric portion of biogeochemical cycling. (c = c is ethylene) 

Source: Heck (1 982). 
1) Gaseous sulfur dioxide and nitrogen oxides are the most abundant 

sulfur and nitrogen compounds emitted into the atmosphere of 

eastern North America from primary sources (tables 4 and 5) 

(U. S. Envi ronmental Protection Agency 1982a, 1982b). 

2) Sulfur dioxide accounts for about 90 percent of the total sulfur 

from man-made sources; the balance consists of various sulfate 

aerosols (table 4) (U. S. Environmental Protection Agency 1982a). 

Gaseous hydrogen sul f i de (H S) , carbonyl sul f i de (COS) , carbon di sul f ide 

(CS ), and dimethyl sulfide [(CH 5 S] comprise about 98 percent of the airborne 

sulfur compounds emitted by natdril sources ; a1 1 are reduced sul fur compounds 

(Sze and KO 1980; Andreae and Raemdonck 1983). 

3) Gaseous nitrogen oxides (NOx) and ammonia (NH3) are the most 

abundant nitrogen compounds emitted into the atmosphere of 

eastern North America. Nitric oxide (NO) and nitrogen dioxide 

(NO2) make up about 96 percent of total emissions from man-made 

sources. Nitric oxide is so quickly oxidized to NO2 in the 

atmosphere that it can be assumed for most biological purposes 

that all nitrogen oxides are NO (table 5) (U. S. Environmental 

Protection Agency 1982b). Ammon a (NH3) makes up almost a1 1 of 

the remaining 4 percent of man-made nitrogen compounds emitted 

into the atmosphere. 

Other nitrogen compounds i ncl udi ng nitrous oxide (N20) and ami nes 

(-NH2) are emitted by man-made sources, but in relatively smaller 

amounts (U. S. Envi ronmental Protection Agency 1982b). 

Nitrous oxide, nitric oxide, and ammonia comprise almost all the airborne 

nitrogen compounds emitted by natural sources (U. S. Environmental Protection 

Agency 1982b). 

1) Sul fur compounds: 

The major natural source of airborne SO, (98 percent) in 

eastern North America is the atmospheric oxidation of 

reduced sulfur compounds. The reduced sulfur compounds of 

greatest importance are hydrogen sul f ide (H2S), carbonyl 

sulfide (COS), dimethyl sulfide [(CH3),S], dimethyl disulfide [(CH3),S, J, and carbon disulfide (CS2) (table 4). 

These compounds are found in sea spray and are also 

produced by decomposition of organic matter by soil and 
Table 4. Airborne sulfur compounds and their sources in eastern North America 

Compound Formula Source Date Reference 

Hydrogen sul f ide W2S biogenic via 1981 Aneja and others 

1982 

A1 tshul 1 er and 

Li nthurst 1984 

Dimethyl sulfide (CH3)2S microbial process 1981 A1 tshul ler and 

Li nthurst 1984 

Dimethyl sulfide (CH3)2S microbial process 1982 A1 tshul ler and 

Li nthurst 1984 

Carbonyl sul f ide COS microbial process 1982 A1 tshul ler and 

Li nthurst 1984 

Carbon disulfide cs2 microbial process 1982 A1 tshul ler and 

Li nthurst 1984 

Methyl mercaptan CH3SH microbial process 1982 A1 tshul 1 er and 

ii nthurst 1984 

Sulfur oxides: Sox fossil fuel 

(gas phase) combustion 

Sul fur monoxide SO fossil fuel 

combustion 

Sulfur dioxide soz fossil fuel 

combustion 

1982 U. S. Environmental 

Protection Agency 

1982a 

1982 U. S. Environmental 

Protection Agency 

1982a 

1982 U. S. Envi ronmental 

Protection Agency 

1982a 

Sulfur trioxide so3 fossil fuel 1982 U. S. Environmental 

combustion and Protection Agency 

atmospheric reactions 1982a 

Disul fur monoxide 520 fossil fuel 

combustion 

1982 U. S. Environmental 

Protection Agency 

1982a 

Sulfate (primaryld S042- fossil fuel 1984 A1 tshul ler and 

combustion Linthurst 1984 

a Nearly 100 percent of aerosol sulfur is in the form of sulfate. 
Table 5. Ni trogen-contai ni ng compounds and thei r sources 

Compound Formula Source Date Reference 

Nitrogen oxides Nox biogenic; anthropogenic 

Nitric oxide NO small amounts of micro- 1976 Soderlund & Svenson 

bial dentrification 1976 

A1 tshul 1 er and 

Li nthurst 1984 

fossil fuel 

combustion 

Nitrogen di oxide NO2 fossil fuel 

combustion; and 

oxidation of NO 

1946 U.S. Environmental 

Protection Agency 

1982b 

1982 U. S. Environmental 

Protection Agency 

1982 b 

Nitrous oxide NzO mainly microbial 1972 A1 exander 1977 

deni tri fication U.S. Environmental 

but some microbial Protection Agency 

nitrification 1982b 

Ammonia 

Ammon i urn 

NH3 microbial and insect 1982 U. S. Envi ronmental 

associated micro- Protection Agency 

bial deami nation and 1982b 

ammoni fication; Altshul ler and 

fertilizers; and Li nthurst 1984 

feedlots 

NH4+ biogenic; atmospheric 1982 A1 tshul ler and 

Li nthurst 1984 

U. S. Environmental 

Protection Agency 

1982b 
>"+!a 

ZV-E! -5mPt 

;*: 2 

Ptmn 

3 

-dCz 20 5. 

d-hn 7 

Q " 

G5 

Q-h mczj! igT m 

,Am; 3. 8" -2.g 

* OLS: 

3 7 m 0 

m 

--crW 

0 

-5 73 

m "5 

t2 z 

ctc: 

m (7 "5 m 13 

If Pt 

0 a- 

"li 0 

c+c: 

3-c+ 
total of all volatile sources of nitrogen compounds in 

eastern North America but have been unable to locate 

published data either to support or refute this judgment. 

d) Man-made emissions of nitrogen compounds are about 10 times 

larger than emissions from all naLural sources in eastern 

North America (U. S. Envi ronmental Protection Agency 1982b). 

1) Each year about 25 million tons of gaseous sulfur dioxide, 

about 23 million tons of gaseous nitrogen oxides, and about 

3.7 million tons of gaseous ammonia are emitted from man-made 

sources into the atmosphere over North America (Husar 1986; 

Husar and Holloway 1983; Husar and Patterson 1980). 

2) About 75 percent of these emissions occur in the eastern and 

only about 25 percent occur in the western parts of the continent (using as demarcation a line running from the TexasLouisiana border to Manitoba in Canada) (Husar and Hol loway 

1983). 

3) The various states of the United States and the provinces of 

Canada differ markedly in emissions of both sulfur and nitrogen 

oxides (see figs. 2 and 3) (National Research Council 1983). 

The total emissions by region for the year 1980 are shown in 

table 6 (National Research Council 1983). 

4) Emissions of sulfur oxides in most regions, States, and 

provinces of eastern North America range from about one to three 

times those of nitrogen oxidesl In the States of Texas, 

Louisiana, Vermont, Connecticut, Rhode Island and New Jersey, 

however, emissions of nitrogen oxides exceed those of sulfur 

dioxide. But the regions, States, and provinces with high (or 

low) sulfur dioxide emissions usually are also high (or low) in 

emissions of nitrogen oxides (compare figs. 2 and 3) (National 

Research Counci l 1983). 

2.5 

In our analysis of regional emissions trends, we found it convenient to 

fol low Husar? (1986) lead and divide eastern North America into southeastern 

Canada and the northeastern, upper and lower midwestern, and southern portions 

of the United States (fig. 4). 

1) From 1880 to 1980, total man-made emissions of sulfur dioxide in 

eastern North America increased from about 2 mi 11 ion to the 

present 30 million tons per year (fig. 5a), while total man-made 

emissions of nitrogen oxides increased from about 1 million to 

the present 23 mil 1 ion tons per year (fig. 6a) (Husar 1986). 
Figure 2. Representative values of SO, emissions in the United States 

and Canada in 1980 (thousands of metric tons). 

Source: National Research Council (1 983). 
Figure 3. Representative values of NOx emissions in the United States 

and Canada in 1980 (thousands of metric tons). 

Source: National Research Council (1 983). 
Table 6. Distribution of man-made emissions of sulfur and nitrogen compounds 

in eastern North America 

Sulfur Compounds 

Netric 

Nitroaen Com~ounds 

Metric 

Source Region Tons Percent Source Region Tons Percent 

Total Emissions: 

Eastern Canada 3,271 14 Eastern Canada 1,023 7 

Northeast U. S. 3,932 16 Northeast U. S. 2,731 17 

Midwest U. S. 8,671 36 Midwest U. S. 4,803 30 

Southern U. S. 8,319 34 Southern U. S. 7,276 46 

TOTAL 24,193 100 TOTAL 15,833 100 

Emi ssions Density (metric tons per hectare): 

Eastern Canada 1.2 x l0I5 

Northeast U. S. 8.5 x 

Midwest U. S. 7.3 x lo_5 

Southern U. S. 5.0 x 10 

Eastern Canada 0.31 x lom5 

Northeast U.S. 5.8 x low5 

Midwest U. S. 3.9 x lom5 

Southern U. S. 3.5 x 10 

Source: National Research Council (1983). 

Total man-made emissions of sulfur dioxide in eastern North 

America rose rapidly between 1880 and 1920, reached about 

20 million tons per year in the early 1920ts, then increased 

again about 50 percent, reaching 30 million tons per year by 

1970, after which there was about a 15 percent decrease to about 

25 mi 11 ion tons per year, due mainly to increased use of low 

sulfur fuels (see fig. 5a) (Gschwandtner and others 1986; Husar 

1986; National Research Council 1986). 

3) These several subregions of eastern North America have differed 

greatly in temporal patterns of change in man-made emissions of 

sulfur and nitrogen oxides (see figs. 5 and 6, which show 

emissions of sulfur per se rather than of sulfur dioxide; please 

note that 1 ton of sulfur equals approximately 2 tons of 

sul fate). 

For example, with regard to sulfur emissions: 

a) In southeastern Canada (Region A), emissions of sulfur 

dioxide rose from negligible in 1880 to about 2 million 

tons per year in the 1930Ls, further rose to about 

8 million tons per year in the mid 1970Ls, and decreased to 

about 4 mi 11 ion tons per year by 1980 (fig. 5b) (Husar 

1986). 
Figure 4. Regions in eastern North America selected for the study of regional 

trends in emissions. 

Source: Husar (1 986). 

Region 

A (Eastern Canada): Eastern Ontario, New Brunswick, Southern Quebec 

B (Northeastern U.S): Connecticut, Delaware, Maine, Maryland, Massachusetts, 

New Hampshire, New Jersey, New York, Pennsylvania, 

Rhode Island, Vermont 

C (Southeastern U.S.): Alabama, Arkansas, Florida, Georgia, Kentucky, Louisiana, 

Mississippi, North Carolina, South Carolina, Tennessee, 

Virginia, West Virginia 

D (Lower Midwest): Illinois, Indiana, Michigan, Missouri, Ohio 

E (Upper Midwest): Iowa, Minnesota, Wisconsin 

F (Western U.S.): All states in the contiguous U.S. not included in Regions B to E 
"" .I I 

18 1 ACGREGATE OF REGIONS A-E 

REGION B C 1 

A 

I 

REGION O 

1 -- I 

1880 1gDC 1920 1940 1960 1980 2000 

YEAR 

1 REGION E 

4 1 

1880 1900 9920 I340 1960 1980 20Cf0 

YEAR 

Figure 5. Regional sulfur emission trend estimates for (a) eastern North 

America (the aggregate of Regions A, B, C, D, and E); (b) Region A; (c) 

Region 6; (d) Region C; (e) Region D; (f) Region E. (SO, = tons sulfur 

multiplied by 2). (Shading indicates range of uncertainty.) 

Source: Husar (1 986). 
i880 1900 1920 1940 1960 1980 2000 

YEAR 

imo i goo 1920 I 940 1960 1980 2000 

YEAR 

Figure 6. (a) Trends in emissions of nitrogen oxides in the Eastern 

United States (the aggregate of region 6, C, D, and E); (b) Region B: 

densities of regions north (Regions B, D, and E) and south (Region C) of 

the Ohio River. (Shading indicates range of uncertainty.) 

Source: Husar (1 986). 
b) In the Northeastern United States (Region 6), emissions of 

sulfur dioxide rose from negligible in 1880 to about 

6 mi 11 ion tons per year in the 19201s, and have remained 

between 4 and 6 mi 11 ion tons per year since that time. As 

in southeastern Canada, emissions decreased from about 6 to 

about 4 million tons per year between 1970 and 1980 

(fig, 5c) (Husar 1986). 

c) In the Southern United States (Region C), emissions of 

sulfur dioxide rose from negligible in 1880 to nearly 

4 mil l ion tons per year in the 1920fs, remained at about 

4 mill ion tons per year until the mid 19601s, and then 

further rose to about 8 million tons per year by 1980 

(fig. 5d) (Husar 1986). 

d) In the lower Midwest (Region D), emissions of sulfur 

dioxide rose from negligible in 1880 to about 3 and a half 

mil l ion tons per year in the early 19201s, varied between 4 

and 6 million tons per year until about 1970, and decreased 

to about 4 million tons per year by 1980 (fig. 5e) (Husar 

1986). 

e) In the upper Midwest (Region E), emissions of sulfur 

dioxide remained below 2 million tons per year from 1880 to 

1980 (fig. 5f) (Husar 1986). 

Similarly, with regard to nitrogen oxide emissions: 

a) In the Northeastern and lower Midwestern United States 

(Regions 6 and D), nitrogen oxide emissions rose steadily 

from negligible between 1880 and 1900 to about 4 million 

tons per year in 1970 and then decreased slightly by 1980 

(figs. 6b and 6d) (Husar 1986). 

b) In the upper Midwest (Region E), nitrogen oxide emissions 

rose steadily from 1900 to 1980, but even then reached only 

about 1 mi 11 ion tons per year (fig. 6e) (Husar 1986). 

c) In the Southern United States (Region C), emissions of 

nitrogen oxides remained below 0.5 million tons per year 

until the 1920s, but then rose nearly exponentially to 5 to 

8 million tons per year by 1980 (fig. 6c) (Husar 1986). 

4) In North America as a whole, total man-made emissions of nitrogen oxides rose slowly from 1880 to 1920, reached about 10 

million tons per year in the mid 19509, and 20 million tons per 

year in the early 1980's (fig. 6a) (Husar 1986). This upward 

trend is primarily the result of increased use of coal, gasoline, diesel fuel, and natural gas (fig. 7) (Gschwandtner and 

others 1986). 
1 ton = 907.1 8 kg 

0 

YEAR 

Figure 7. Overall trend in NOx emissions from 1900 to 1980 for the 

United States and by fuel type for each study year. 

Source: Gschwandtner and others (1 986). 
5) The increase in total SO2 emissions caused by the electrification of North American society between 1940 and 1980 is about 

four times greater than that caused by the industrial izaticn of 

North American society between 1880 and 1980 (ARshuller and 

tinthurst 1984). In fact, industrial emissions of sulfur 

dioxide in the United States decreased from about 7 ml"lfion 

tons per year -in the late 1920's to about 4 million tons per 

year in the mid 1950's to 1980; by contrast, electric utility 

emissions increased from about 3 million tons per year in the 

early 1340% to about 18 mi 11 ion tons per year since 1970 

(fig. 8) (Gschwandtner and others 1986; Husar and Wo1 loway 

1983). 

6) Total man-made emissions of sulfur dioxide in North America are 

predicted to remain between 24 and 30 million tons per year for 

the next 50 years (fig. 9) (U. S. Congress 1984; U. S. Environmental Protection Agency 1984). 

7) Total man-made emissions of nitrogen oxides in North America are 

predicted to continue to increase from about 20 million tons per 

year to between 22 and 32 million tons per year during the next 

50 years (fig. 9) (U.S. Congress 1984; U.S. Environmental 

Protection Agency 1984). 

1) Once emitted, sulfur and nitrogen compounds are carried wherever 

the wind blows (National Research Council 1983, 1386). 

The average zone of influence of any particular emission source 

extends in a1 l directions from the source, and is roughly 

symmetrical--wi th the center of the deposition field general 1 y 

displaced from the source by a distance of only a few tens or 

perhaps hundreds of km, Thus, air pollutants generally have 

their greatest effects within a few hundred km from their source 

of emission (Bolin and others 1971; National Research Council 

1983, 1986). 

3) The concentration and deposition of primary air pollutants 

decreases progressively with increasing distance and time afLer 

emission from any source (Altshuller and Linthurst 1984). 

4) The concentration and deposition of secondary pollutants is a 

complex function of meteorological, seasonal, altitudinal, 

temporal, geographical, and other factors (National Research 

Gounci 1 %983). 

5) The elapsed time between emission and deposition varies with the 

airborne pollutant chemical in question but ordinarily ranges 

from a few minutes or hours Lo a maximum of about 5 days for 
35 

1 ton = 907.1 8 kg 

YEAR 

Figure 8. Overall trend in SO, emissions from 1900 to 1980 for the 

United States and by source category for each study year. 

Source: Gschwandtner and others (1 986). 
Year 

Figure 9. Sulfur dioxide and nitrogen oxides emissions trends-national totals, 

1 900-2030. The graph above displays estimates of historical emissions, and projections of future emissions of sulfur dioxide and nitrogen oxides. Pre-1940 estimates and 

post-1 990 projections are subject to considerable uncertainty. Projections of future 

emissions incorporate a wide range of assumptions about future economic growth, 

energy mix, and retirement of existing facilities; they assume no change in current air 

pollution laws and regulations. 

Source: U.S. Congress (1984). 
sulfur dioxide and about 4 days for nitrogen oxides and ammonia 

(Husar and Holloway 1983; National Research Council 1983; U.S. 

Environmental Protection Agency 1982a, 1982b). 

6) These intervals are sufficient for dispersal over both short 

distances (0 to 500 km) and long distances (>500 km) (National 

Research Council 19831, 

7) Short-distance transport is favored by low wind speeds, stable 

atmospheric condi ti ons , and mountai nous terrai n. Long-di stance 

transport is favored by high wind speeds over flat terrain and 

unstable atmospheric conditions (National Research Council, 

1983, 1986). 

8) The distance pollutants travel also depends on the height at 

which they are emitted into the atmosphere. Generally, the 

lifetime of sulfur and nitrogen compounds in the atmosphere is 

increased somewhat by the height at which they are emitted. 

Very tall stacks (>250 meters) facilitate transport over 

1,000 km (U. S. Environmental Protection Agency 1982b). 

1) During transport, some of the gaseous sulfur dioxide, nitrogen 

oxides, and ammonia is transformed both chemically and physical ly into a complex mixture of gases, fine aerosol particles, 

coarse particles, and dissolved and suspended substances in 

rain, snow, dew, hail, fog, cloud water, and rime ice 

(Altshuller and Linthurst 1984; National Research Council 1983). 

2) The predomi nant chemical reactions occurri ng duri ng transport 

through the atmosphere are shown in figs. 10 and 11 and may be 

summarized as fol lows (A1 tshul ler and Linthurst 1984; National 

Research Counci 1 1983) 

a) Gaseous sulfur dioxide (SO,) is further oxidized by 

airborne hydroxyl radicals (OH* ) to sul f i te ion (SO3-), 

which in turn reacts very rapidly with water or airborne 

particles, producing H,SO, or sulfate particles such as 

ammonium sulfate [(NH4)2S04]. 

b) Gaseous nitric oxide (NO) is quickly oxidized by airborne 

hydroxyl radicals (OH') to gaseous NO,. This reaction 

occurs so rapidly that almost all gaseous nitrogen oxides 

occur as NO2. 

c) Gaseous NO2 is further oxidized by airborne hydroxyl 

radicals (OH-) to form nitrate (NO3-), which combines 

with water or airborne particles to form HNO, or nitrate 

particles such as ammonium nitrate (NH4N03). 
Oxibtlon 

Hydroxy radical (HO) 

G~seous oxidas Hydroperoxy radical (H02) L Sulfur dioxide (SO2) Ozone to3) Gaseous Acids 

0 Nitric oxide (NO) - %Sulfuric Acid (H2S04) 

Nitrogen dioxide (NO2) Nitric acid (HN03) 

Hydrogen peroxide (H202) 

Neutralization 

Catalysis of oxygen (02) 

reaction wtth manganese Metal oxides (MxOy) Aquews or Dry Salts g & Aqueous oxides 

(Mn), iron (Fe), etc. Aqueous Strong Acids ~~~~l urbonater [M~(co~)~ j Ammonium bisulfate (NHIHS04) S02(aq)P " g - ~2~04~ Ammonium sulfate [(MM4)2S04! 5, NO (ad HNO,C Ammonium nitrate (NH,NO,) % 3 N02(aq) 

*I s Metal nitrates [M,(N03)yl 

aS02(aq) reacts to form hydrogen ions (H') and bisulfite ions (HSO;). 

b~queous H2S04 dtssociates into hydrogen ions. bisulfate ions IHSOj). and sulfate ions (SO:). 

CAqueous HMO3 dissociates into H+ and nitrate ions (N05). 

Figure 10. Pathways for the formation of atmospheric sulfate and nitrate. 

Source: National Research Council (1 983). 
Figure 11. Sample processes in SO, oxidation. This figure illustrates 

the complexity of the atmospheric processes. 

Source: Bennett and others (1 985). 
d) Gaseous nitrogen dioxide (NO,) reacts with volatile hydrocarbons [volatile nonmethane organic compounds (VNMOC)] in 

the presence of sunlight to give a variety of highly 

reactive photochemical oxidants including ozone (03)$ 

hydrogen peroxide (H,O,), formaldehyde (H2CO), peroxyacetyl 

nitrate (PAN) and peroxypropionyl nitrate (PPN). 

e) Gaseous ammonia (NH3) reacts with water to form ammonium 

ion (NH,+), which may react with sulfate or nitrate to form 

(NH4)2S04 or NH4N03. 

3) Some of the more important physical transformations occurring 

during transport through the atmosphere include the following 

(Altshuller and Linthurst 1984; National Research Council 

1983) : 

a) Formation of liquid and solid suspended particles by 

condensation of gases; 

b) Coalescence of fine liquid and solid suspended particles 

into larger aerosols; 

c) Adsorption and absorption of gases onto liquid and solid 

suspended particles; 

d) Formation of clouds, as water vapor becomes supercooled; 

e) Formation of raindrops, snowflakes, and hail; 

f) Dissolution of gases in raindrops, snowflakes, and hail 

particles; and 

g) Evaporation. 

4) The end result of all the natural and man-made emissions, and 

the chemical and physical transformation processes during 

transport, is formation of a complex mixture of airborne 

chemicals that are present in the atmosphere as (Altshuller and 

Linthurst 1984; National Research Council 1983): 

a) Gases; 

b) Suspensions of fine aerosol particles (<2 microns in 

diameter) and larger aerosol particles (2-10 microns in 

di ameter) ; 

c) Suspensions of fly ash, soot, and other coarse particulate 

matter (10 microns to 1 millimeter or larger particles); 

d) Dissolved and suspended substances in cloud water, fog, and 

precipitation including ral'n, snow, hail, fog, dew, and 

rime ice. 
3. CHEMICAL EXPOSURE OF FORESTS IN EASTERN NORTH AMERICA 

3.1 Introduction 

Chapter 2 of this document describes the major sources, magnitude of 

emissions, principal chemical transformations, and the usual d-irec"ions and 

distances over which sulfur- and nitrogen-derived compounds are dispersed 

through the atmosphere over eastern North America. Many meteorological, 

cl imatological , chemical, and biological factors influence the transfer of 

these compounds and their derived products from the atmosphere into forest 

ecosystems. These factors a1 so cause 1 arge-scal e spati al and temporal 

variation in concentration, residence time, and deposition of these substances 

across the continent and small er-sca1 e vari ati on5 associated with differences 

in elevation. 

The major purpose of the present chapter is to define and quantify the 

nature and magnitude of the chemical exposure of forests in various regions 

of eastern North America. The chemical form of the pollutant, concentration, 

duration, time of day and season of year are all important in determining 

exposure and response of forests. 

3.2 

Pollutant Chemicals to which Forests are Ex~osed in Eastern North America? 

1) The airborne sul fur- and ni trogen-derived chemicals presently 

known or suspected to cause growth-stimulating or growthinhibiting effects on forests through foliage-mediated response 

mechanisms include the fol lowing (U. S. Environmental Protection 

Agency 1982a, 1982b, 1986) : 

Gaseous sulfur dioxide (SO2); 

Gaseous nitrogen oxides (NOx), mainly NO2 but also NO; 

Gaseous peroxyacetyl nitrate (PAN); 

Gaseous peroxypropionly nitrate (PPN); 

Gaseous nitric arid; 

Gaseous ammonia (NH3) ; 

- Gaseous ozone (03); 

Gaseous hydrogen peroxide (H202); 

* Hydrogen peroxide dissolved in clud water precipitation; 
2) The next four ionic materials occur mainly as fine aqueous aerosol 

particles and as dissovled and suspended materials in cloud 

water and precipitation. Their main effect on forest growth is 

through depostion on the soil. 

Sul fate ions (SO,=) ; 

Nitate ions (NO,-) ; 

Ammoni um ions (NH,') ; 

Hydrogen ions (H'); 

All of these last four ionic materials occur mainly in two forms: 

fine aqueous aerosol particles; and dissolved and suspended materials 

in cloud water and precipitation. 

3) Gaseous peroxyacetyl nitrate (PAN) ; measurement methods for the 

sulfur- and ni trogen-derived pol 1 utants are described in the 

Criteria Documents for nitrates, sul fates, and ozone (U. S. 

Envi ronmental Protecti on Agency 1973, 1982b, 1986). The 

majority of air pollution monitoring devices in eastern North 

America have been placed in urban areas. The only monitors 

located specifical ly in forests were the ozone monitoring sites 

of the Environmental Protection Agency-Forest Service and 

National Park Service monitoring programs. 

The rate of transfer of airborne pollutant chemicals into forests is 

determined by a variety of chemical, physical, and meteorological features of 

the pollutants and their dispersal in the atmosphere of the forest, and by 

various physical, chemical, and biological features of plants, animals, and 

microorgani sms i n the ecosystems themsel ves. These several atmospheric and 

bio1ogi;al factors are imp&tant in assessing exposure and are disc*ussed in the 

following sections. 

I) Pollutant chemicals are dispersed in the atmosphere in four 

physical states (Hales 1984; National Research Counci 1 1983) : 

e as gases; 

as fine aerosol particles (<2 microns in diameter) and 

coarse aerosol particles (2 to 10 microns in diameter). In 

the humid environment of eastern North America, most 

aerosols occur as tiny liquid particles, very few as sol id 

particles; 

as fly ash and other coarse particulate matter (10-micron 

to I-mm or larger particles); and 

as dissolved and suspended substances in cloud water, fog 

water, and precipitation. 
2) These four types of substances are transferred from the atmosphere into forest ecosystems where they are deposited onto two 

types of surfaces (Hales 1984; Hicks 1984): solid plant, 

animal, microbial, soil, and stone surfaces; and liquid water 

surfaces within forest ecosystems. 

The following physical mechanisms of transfer and deposition are involved: 

* Gases are transferred to solid surfaces by adsorption, 

absorption, condensation, and diffusion processes. They 

also are transferred to dew and other 1 iquid water films, 

water droplets, and pools of liquid water by condensation, 

dissolution, and diffusion processes. 

+ Fine and coarse aerosol particles are transferred mainly by 

physical impaction processes onto both the solid and liquid 

surfaces in forests. 

* Fly ash and other coarse particulate matter are transferred 

into forests by gravitational settl ing. 

Dissolved and suspended substances in precipitation are 

transferred into forests by physical impaction in the case 

of cloud water, fog water, and rime ice; and by gravitational settling in the case of raindrops, snowflakes, and 

hail and sleet particles. 

3) Gases, fine and coarse aerosols (including fog and cloud water 

droplets), and coarse particulate matter are transferred into 

forests by dry deposition processes. These dry deposition 

processes occur under all meteorological conditions, at all 

times of the day and night, and in all seasons of the year 

(Hicks 1984). 

4) The rate of transfer of gases and aerosols from the atmosphere 

into forests (deposition velocity) is a complex function of the 

fol 1 owi ng (Fowl er 1980) : 

* concentration of gases and aerosols in the atmosphere which 

in turn depend on: 

- proximity to and strength of major pollution sources; 

- duration and frequency of air stagnation events (see 

fig. 12); 

- time since the most recent precipitation event; and 

- thoroughness of atmospheric cleansi nb achieved during that 

event; 

speed and turbulence of wind; 

* size and aerodynamic properties of aerosol particles; 
Figure 12. Incidence of air stagnation advisories issued over a 1 0-year 

period. Adapted from Lyons (1 975). 

Source: Bennett and others (1 985). 
chemical and physical properties of gases and aerosol s ; 

chemical and physical properties of the surfaces to which 

the gases and aerosols are transferred; 

temperature and relative humidity of the free almosphere, 

and of the atmosphere near the surface to which the gases 

and aerosols are being transferred; and 

* temperature of the receiving surface, 

Since these many factors vary greatly in both time and space (especially in 

complex terrain), the rate and total amount of dry deposition of gases and 

aerosols in eastern North America varies greatly geographically, temporally, 

altitudinally, and with regard to the chemical nature of the gases and aerosols 

involved (Wales 1984; Wicks 1984). 

5) Major cations and anions dissolved and suspended in precipitation are transferred into forest ecosystems by wet deposition 

processes e during rain, snow, hail, and sleet storms). 

Wet deposition episodes typically last from a few minutes or 

hours to a Few days. The time between precipitation events 

varies greatly, but typically is less than 1 week in most parts 

of eastern North America. Of course, drought periods of up to 7 

or 8 weeks or even longer have occurred in exceptionally dry 

years. Wet deposition events in eastern North America generally 

occupy about 4 to 8 percent of the total hours per year 

(A1 tshul ler and Linthurst 1984). 

6) The rate of transfer of dissolved and suspended substances in 

precipitation into forests is determined mainly by the concentration of these substances in the atmosphere and the amount of 

precipitation that occurs per unit time. These substances are 

removed from the free atmosphere by cloud nucleation processes 

(rai n-out) , and by impaction processes as raindrops , snowf 1 akes , and hail and sleet particles fall through the atmosphere below 

clouds (wash-out) (Altshuller and Linthurst 1984). 

3.4 

Airborne Pollutant Chemicals that are Delivered into Forest Ecosystems in " 

Various Parts of Eastern North America? 

1) The magnitude of pollutant load received by a forest is 

generally an inverse function of the distance from major sources 

of pollution (A1 tshul ler and Linthurst 1984; Fowler 1980). 

2) The magnitude of pollutant load received by a forest on sloping 

terrain is generally greater on slopes facing predominant winds 

than on slopes facing in other directions, especially those 

facing away from predominant winds (Altshuller and Linthurst 

1984). 
3) Forests with very uneven canopy heights (such as the mixed 

conifer-hardwood forests of Mew England and southeastern Canada) 

generally receive heavier pollutant loads than forests that have 

very even canopy heights (such as the beech-birch-maple forests 

of New England and southeastern Canada or the naturally regenerated pine-oak-hickory forests or planted stands of southern pine 

in the Piedmont and Coastal Plain regions of the Southern United 

States) because they have a larger number of impaction surfaces. 

4) Mixed-species forests generally receive heavier pollutant loads 

than forests of single tree species because they have very 

uneven canopy heights. 

5) Trees near the edge of forest stands generally receive heavier 

loads of pollutants than trees within forest stands. 

6) Trees with stomata closed (high stomatal resistance) take up 

airborne gases much more slowly than forests with stomata open 

(low stomatal resistance). Typical ly, 4- to 20-fold differences 

in rate of uptake of gases are observed, for example, during 

night and day in the case of tree species that close stomata at 

night, and during droughts compared to wet periods (Mudd and 

Kozl ows ki 1975). 

7) Forests on high mountains (> 1000 meters) are exposed to much 

heavier 1 oads of airborne pol l utant chemicals than simi 1 ar 

forests at low elevation (Lovett and others 1982; Wolff and 

others 1987). This is caused by the following: 

* Wind speeds are much greater at high than at low elevation 

(typically 2 to 10 times greater); 

Amounts of precipitation are generally much greater at high 

than at low elevation (typically 2 to 3 times greater); 

Fog and clouds occur much more frequently at high than at 

low elevation (typically 2 to 10 times greater); 

Fog and cloud water significantly augment the total amount 

of water delivered to forest canopies through precipitation 

processes (typically by 20 to 200 percent); 

Concentrations of dissolved and suspended substances in fog 

and cloud water are typically 5 to 10 times greater than 

those in precipitation; 

* Concentrations of ozone and other photochemical oxidants 

are often greater at high than at low elevation (typically 

about 10 to 50 percent greater); and 

Diurnal periodicity in concentration of photochemical 

oxidants, especially ozone, is much less at high than at 

low elevations. 
3.5 

Proximity to pol 1 utant sources is ~mportant i n determi ni ng the chemical 

exposure of forests in various regions of North America. The general contrast 

in' concentrations of sulfur dioxide. nitrogen oxides, and ozone in urban, 

suburban, rural, and remote regions of the -northeastern States is shown in 

fig. 13 (Barchet 1987). 

In our efforts to characterize the chemical exposure of forests in various 

parts of eastern North America, we found it useful to recognize five general 

types of forests based on proximity to sources: 

1) Urban forests such as the 30,000 hectares of forest land within 

the city limits of Atlanta, GA. Although urban forests probably 

constitute a very small percentage of the total area of eastern 

North America, urban areas are home to about 30 percent of the 

human population of eastern North America and thus are of great 

interest to many people in the United States and Canada; 

2) (for example, within about 30 km of a major power plant or metal 

smelter). We estimate that these areas make up about 5 percent 

of the total forest area of the eastern United States and 

southeastern Canada; 

3) Near-urban forests within about 50 km of major sources of 

pollutant areas. We estimate that these areas include about 

37 percent of the total forest area of the eastern United States 

and southeastern Canada; 

4) Rural forests more than 50 km distant from major urban areas or 

from major point sources of pollution. We estimate that these 

areas make up about 25 to 30 percent of the total forest area of 

the eastern United States and southeastern Canada; and 

5) High-elevation forests such as those on Mount Mitchell, NC, and 

Whiteface Mountain in New York (see fig. 14). These areas 

are characterized by frequent exposure to clouds and concentrations of ozone that show little or no diurnal variability. We 

estimate that these areas make up about 2 percent of the total 

forest area of the eastern United States and southeastern 

Canada. 

In making these distinctions we have followed the lead of the personnel 

of the National Vegetation Survey of the Forest Response Program who have been 

developing the so-called Southern Forest Atlas/Geographic Information System, 

in particular the "Chemical Atlas". They have used aerial photographs in 

assessing the proximity of the U.S. Environmental Protection Agency (EPA), 

Electric Power Research Institute (EPRI), and Tennessee Valley Authority (TVA) 

monitoring sites to forests and to major point and area sources of airborne 

pollutant chemicals. Although we have been unable to do so thoroughly in the 
Urban Suburban Aur Qi Asmote 

Figure 13. Mean annual air quality for SO,, NO,, and 0, by station class in 

the three-state region of New Jersey, New York, and Pennsylvania for the 

period 1 976 to 1 983. The number on the vertical bar indicates the number 

of stations in that class. 1 ppm SO, = 2600 pg/m3; 1 ppb NO, = 1900 p/m3. 

Source: Barchet (1 987). 
time available, we believe aerial-photograph interpretation of the amounts and 

types of forests near monitoring sites offers a useful means by which to 

increase further our understanding of the chemical exposure of forests of the 

five types listed above. 

In drawing certain general inferences about the chemical exposure of these 

five types of forests, we found the chemical exposure of near urban and 

rural forests to be very similar. 

1) Suf fur dioxide (SO2): As di scussed more f ul1 y be1 ow, gaseous 

sulfur dioxide is rarely found in concentrations sufficient to 

cause visible injury to vegetation (50 to 5,000 ppb for 8 hours 

or 500 to 3,000 ppb for 1 hour, see table 3) in remote forests, 

rural forests, or near-urban forests in eastern North America. 

It does occur' in injurious concentrations in some urban forests 

and in rural forests near major pollution sources, however. 

These effects on forests in the immediate vicinity of major 

pol 1 ution sources are we1 1 documented in the pub1 i shed 1 i terature (U. S. Envi ronmental Protection Agency 1982a). 

a) In high-elevation forests (fig. 14), gaseous sulfur dioxide 

concentrations are usually very near the detection limits 

for available monitoring equipment. At Whiteface Mountain, 

for example, average l-hr concentrations in 1982 were below 

1 ppb; maximum concentrations were about 20 ppb at both 

high and low elevation locations on this mountain (Kelly 

and others 1984). These values are far below the threshold 

for foliar injury -- about 50 ppb for 8 hours (Lefohn and 

Ti ngey 1984). 

b) In most near-urban and rural forests of eastern North 

America, sul fur dioxide concentrations seldom exceed 

50 ppb. For example, 39 percent of 922 rural monitoring 

site years (number of specific sites analyzed over a 

mu1 ti -year period) across the Eastern Uni ted States in the 

Envi ronmental Protection Agency SAROAD, EPRI , and TVA 

moni tori ng networks for 1978-81 showed no 1- hour average 

sulfur dioxide concentrations greater than or equal to 

50 ppb (Lefohn and Tingey 1984). Similarly, 90 percent of 

a1 1 1-hour average values for sulfur dioxide at 20 sites 

classified as rural or remote in the SAROAD data base for 

1979-83 in the States of Arkansas, Louisiana, Mississippi, 

Georgia, and Florida were less than 10 ppb. The highest 

l-hour average concentration observed at any of these 

20 sites was 822 ppb. Only two sites showed l-hour average 

values greater than 500 ppb. Eleven sites showed no values 

greater than 200 ppb even for a single hour in 1979-83 

(Lefohn and others 1987). This js true in part because 

sulfur dioxide is rapidly transformed to sulfate aerosol 

duri ng transport from major sources (Husar and Hol 1 oway 

1983). 
Figure 14. Regions of the United States with high-elevation forests (boreal, 

sub-alpine, and montane; mixed boreal and deciduous; mixed boreal, lake, 

and deciduous). 

Source: Bennet and others (1985). 
c) Some urban forests and rural forests near maior sollution Y 

sources in eastern North America are exposed to phytotoxic 

concentrations of gaseous sul fur dioxide. Maximum concentrations of 500 to 1,000 ppb or even higher have been 

reported near some industrial sources for several days at a 

time during the 1970's and early 1980's (U.S. Environmental 

Protection Agency 1982a). 

In Electric Power Research Institute's Sulfate Regional Experiment in 

1977-78, the highest annual average concentrations of sulfur dioxide observed 

in rural areas was 20 to 30 ppb (see fig. 15). These relatively high values 

were observed in the Ohio River Valley and in western Pennsylvania. The 

highest monthly 1-hour average values of about 30 ppb occurred during January 

and February 1978. The highest concentrations in near-urban areas were found 

in the vicinity of Richmond, VA, Philadelphia, PA, New York, NY, and Boston, MA 

(Mueller and others 1980). 

d) Regional variability. The region with the highest annual 

average sulfur dioxide concentrations at rural sites in 

eastern North America is in the Ohio River Valley where 

these values approach 10 ppb. Average concentrations 

decrease with increasing distance from this region, 

reaching approximately 2 ppb in New England and less than 

1 ppb in the Southern United States (Muel ler and Hidy 

1983). 

In the Environmental Protection Agency SAROAD data base for 1980-85, the 

ranges of sulfur dioxide concentrations at representative low-elevation rural 

and remote sites in the Northeastern and Southeastern United States were as 

follows: 

Number of sites 

24-hr arithmetic mean 

24-hr maximum concentration 

1-hr maximum concentration 

Northeast Southeast 

11 16 

7 - 13 ppb 7 - 8 ppb 

14 - 57 ppb 3 - 52 ppb 

29 - 210 ppb 5 - 200 ppb 

e) Di urnal variabi 1 i ty. The highest 1- hour concentrations of 

sulfur dioxide in near-urban and rural forests usually 

occur at night and during the early-morning hours if 

emissions are predominantly from residential and industrial 

space-heating sources with low stacks (fig. 16) (U. S. 

Environmental Protection Agency 1982a). When emissions are 

from tall stacks, however, highest concentrations usually 

occur in the morning just after sunrise with a broad 

diurnal maximum between 4 p.m. and 6 a.m. (U. S. Environmental Protection Agency 1982a). 
Figure 15. Sulfur dioxide (arithmetic mean) and sulfate (geometric mean) 

concentrations. Data obtained during 5 months between August 1977 and 

July 1978. Note: 18 ppb SO, is approximately 28 pg/m3 SO,. 

Source: Bennett and others (1 985). 
HOUR 

Figure 16. Composite diurnal pattern of hourly sulfur dioxide concentrations 

are shown for Watertown, MA, for December 1978. 

Source: U.S. Environmental Protection Agency (1 982a). 
f > . Several-day episodes with atypical ly high concentrations of sul fur dioxide sometimes 

result from the accumulation of pol lutants in slow-moving 

air masses. Accumulations over large areas sometimes 1 ast 

3 to 10 days, or when the area involved is smal 1, 1 to 

7 days (Mueller and Hidy 1983). 

Seasonal variability. Both the atmospheric residence time 

and the average distance of transport for sulfur dioxide is 

greater in winter than in summer months (Husar and Holloway 

1983). As a consequence, highest concentrations of sul fur 

dioxide usually occur during the winter months in northern 

areas where residential and commercial space and water 

heating are major sources. In southern areas where air 

conditioning is common, however, a small secondary peak 

during the summer may also be evident (U.S. Environmental 

Protection Agency 1982a). Summer concentrations are are 

presumed to be of greater biological importance than winter 

concentrations, since forest trees are more active during 

the growing season (Kozlowski and Constantinidou 1986a). 

h) Major sources. Approximately 70 percent of sulfur dioxide 

emissions in the Eastern United States are from power 

plants, many with tall stacks and located in rural areas. 

Residence time in summer in the midwestern States is 

approximately 2 days. During that time, the plumes from 

power plants are likely to have been transported up to 

1000 km and to have broadened considerably, and the sulfur 

dioxide to have been transformed into sulfate or to have 

been removed by wet or dry deposition (Husar and Holloway 

1983). 

i) Long-term trends. Judging from long-term trends in emissions, it is reasonable to infer that sulfur dioxide 

loadings of forest ecosystems have remained relatively 

constant or decreased in the Northeastern United States and 

southeastern Canada since the 1920' s, but have increased 

steadily in the Southeastern United States since the 1940's 

(Husar 1986). 

2) Nitrogen oxides (NOx): As discussed more fully below, gaseous 

nitrogen oxides are found rarely, if ever, in concentrations 

sufficient to cause visible injury to vegetation (2,000 to 

5,000 ppb for 8 hours or 4,000 to 8,000 ppb for 1 hour, (see 

table 3) in high-elevation forests, rural forests, coastal and 

maritime forests, and near-urban forests or even in urban or 

rural forests near major pollution sources. In fact, injury to 

trees or other forest vegetation by exposure to gaseous nitrogen 

oxides occurs so rarely in eastern North America that there are 

no reports of injury in the published literature (National 

Research Counci 1 1977b; U. S. Envi ronmental Protection Agency 

1982b) (see also chapter 4). 
a) In concentrations of gaseous 

nitroaen oxides usuallv are below the detection limits for 

avai 1 able monitoring equipment. At Whi teface Mountain, for 

example, average 1-hour concentrations in 1982 were below 

1 ppb, and maximum concentrations were about 3 ppb (Kelly 

and others 1984). At a series of remote sites in the 

Eastern United States in 1978-81, l-hour NOx concentrations 

typically exceeded 50 ppb fewer than 10 times per year 

(Lefohn and Tingey 1984). These values are far below the 

threshold for visible injury -- about 2000 ppb for 8 hours 

(U.S. Environmental Protection Agency 1982b). 

b) In most near-urban and rural forests of eastern North 

America, concentrations of nitrogen oxide seldom exceed 

10 ppb (U. S. Environmental Protection Agency 1982b). Thi s 

is true in part because nitrogen oxides are often consumed 

in the photochemical formation of ozone (see sec. 2.6). 

Annual average l-hour NOx concentrations for 9 rural EPRI 

sites in the Eastern United States were 7 ppb from 1978 to 

1982. The only EPRI site where significant nitrogen oxides 

were recorded at a rural site was near Philadelphia, PA 

(Mueller and Hidy 1983). There, the maximum nitrogen 

oxides concentration was 150 ppb. 

c) Significant nitrogen oxides concentrations are usually 

confined to urban areas and rural areas near major 

ollution sources. For example, in 1976-1980, peak concenfrations of 270-ppb nitrogen oxides were reported at St. 

Louis, MO; New York, NY; Springfield, IL; Cincinnati, OH; 

and Sagi naw and Southf ield, MI (U. S. Environmental Protection Agency 1982b). Similarly, during the years 1976-1980, 

peak NOx concentrations equaling or exceeding 400 ppb were 

measured at rural industrial areas near Ashland, KY, and 

Port Huron, MI. Annual average concentrations in urban 

areas of eastern North America are typically about 70 ppb 

(U.S. Environmental Protection Agency 19826). 

Urban nitrogen oxide concentrations between 6: 00 and 9: 00 a.m. are very 

important in the formation of ozone and other photochemical oxidants. Average 

nitrogen oxides concentrations during these hours ranged from about 50 to 

150 ppb in nine urban areas (U.S. Environmental Protection Agency 1982b). 

d) Regional variability. The regions with the highest annual 

average nitrogen oxide concentrations at rural and remote 

sites in eastern North America are in the Ohio River Valley 

and in the southern Piedmont regions of the southeastern 

States; here these values approach 12 ppb. Average concentrations decrease with increasing distance from these 

regions, reaching approximately 1 ppb in remote areas of 

New England and elsewhere in the Southern United States 

(Mueller and others 1980). 
In the Environmental Protection Agency SAROAD data base for 1980-85, the 

ranges of nitrogen oxide concentrations at representative low-elevation rural 

sites in the Northeastern and Southeastern United States were as follows: 

Northeast Southeast 

Number of sites 8 16 

24-hr arithmetic mean 10 - 20 ppb 3 - 17 ppb 

24-hr maximum concentration 27 - 62 ppb 12 - 75 ppb 

1-hr maximum concentration 27 - 338 ppb 21 - 100 ppb 

e) Diurnal variability. No typical diurnal patterns of 

nitrogen oxide concentrations exist. A1 though a recurrent 

diurnal pattern is discernible in some areas of the United 

States, in many other areas peak concentrations can occur 

at almost any time of day or night. Concentrations of 

nitrogen oxides in urban areas, like those of volatile 

hydrocarbons, tend to increase rapidly in the morning as a 

result of NO emissions and photochemical conversion to NOZ 

when atmospheric dispersion is limited and automobile 

traffic is dense. Photochemical generation of nitrous 

oxides is followed by a concomitant rapid increase in ozone 

concentration, which depresses the nitrous oxide concentrations until later afternoon hours, when decreasing 

radiant energy and increasing NO emissions overwhelm the 

ozone-generating mechanisms. High NO, concentrations then 

occur after photochemical activity has ceased (fig. 17). 

Short-term peaks of NO2 can occur at night and are not 

associated with traffic emissions or photochemical oxidation (U.S. Environmental Protection Agency 1982b). 

f) Dai ly to weekly variabi 1 i ty. Several-day episodes with 

atypically high concentrations of nitrogen oxide sometimes 

result from the accumulation of pollutants in slow-moving 

ai r masses. Accumul ations over 1 arge areas sometimes 1 ast 

3 to 10 days, or when the area involved is small, 1 to 

7 days (Mueller and Hidy 1983). 

g) Seasonal variability. Just as there is no typical diurnal 

pattern for peak nitrogen dioxide concentrations, there is 

also no typical seasonal pattern. Peak monthly averages 

occur at different times of the year in different locations. In remote areas, nitrogen oxide and nitrous oxide 

seasonal peaks occurred any time between October and April 

(U. S. Environmental Protection Agency 1982b). Power plants 

emit about 15 percent more nitrogen oxides in summer than 

in winter. Also, seasonal variation in automobile travel 

leads to seasonal variation in emissions of nitrogen oxides 

-- 18 percent more vehicle miles are traveled in summer 

than in winter (U. S. Environmental Protection Agency 1986). 
----- 

--NO 

NO FORMAT ION 4- OZOME LCAVCNGIXG PHOTOCHEMICAL FORMATtON OF ma - FORM*710NOfN01 

SUNRISE 

TIME OF OAV lCSTl 

Figure 17. Pollutant concentrations in Central City, St. Louis, October 1, 

1976, average of RAMS sites 101, 102, 106, and 107. illustration of photochemical and ozone scavenging formation of NO,. 

Source: U.S. Environmental Protection Agency (1 982b). 
k) . Approximately 48 percent of nitrogen oxide 

emissions are from transportation vehicles, about 40 percent from power plants, and the remaining 20 percent 

from other industrial, commercial, and domestic sources 

(U. S. Envi ronmental Protection Agency 1982b). 

i > . Judging from long-term trends in ernissions, it is reasonable to infer that nitrogen oxide 

load? ngs in forest ecosystems have increased markedly in 

most parts of eastern North America (Husar 1986) and will 

probably increase still further well into the 21st century 

(U. S. Congress 1984). 

3) : As discussed more fully below, gaseous ozone is 

found in concentrations sufficient to cause visible injury to 

vegetation in many parts of eastern North America. For example, 

20 to 40 ppb for 8 hours or 90 to 150 ppb for 2 hours (see . . 

table 3) have been reported in many urban forests, rural 

oint sources, near-urban forests, rural 

forests, coastal an1 maritime forests, and in high-elevation 

forests. Ozone-induced visible injuries on forest trees in 

various parts of the continent are well documented in the 

pub1 i shed 1 i terature (U. S. Environmental Protection Agency 

1986). 

a) In average 7-hour summertime concentrations of ozone are only somewhat lower than those near 

major urban and rural pdl lution sources in the southern 

United States--for example, 30 to 32 ppb at monitoring 

sites in the Croatan National Forest in North Carolina and 

DeRidder, LA, compared to 37 to 57 ppb at 26 other sites 

listed in table 7 (Evans and others 1983; Pinkerton and 

Lefohn 1986). At some other remote sites, particularly at 

high elevations, however, l-hour peak concentrations of 

gaseous ozone often are as high as those in some urban and 

near-urban forests. In the Shenandoah National Park, for 

example, peak summertime ozone concentrations in excess of 

100 ppb were reported in the early 1960's (Berry 1964). 

Simi 1 arly high l-hour peak concentrations have a1 so been 

observed in most years between 1975 and 1984 at Whiteface 

Mountain (Lefohn and Mohnen 1986; Mohnen 1987). Furthermore, ozone shows little diurnal periodicity at high 

elevation sites; thus, high-elevation forests are exposed 

to high concentrations of ozone for many more consecutive 

hours than forests in low elevation locations. Seven-hour 

average ozone concentrations at high-elevation sites are 

frequently about double the concentrations of low elevations in the same geographical areas (Lefohn and Jones, 

1986; Lefohn and Mohnen 1986). These values are substantially greater than the threshold for visible injury by 

ozone--20 to 40 ppb for 8 hours (Lefohn and Tingey 1984; 

U. S. Environmental Protection Agency 1986). 
Table 7. Ozone concentrations in 1983 at rural sites located more than 50 km from a city with a population over 

Number of hourly 

7-hr mean average concentrations 

0900- 1559 

Monitoring April- Cumulative Frequency Percent i 1 el! 

Site October 90 95 99 Peak PPR PPm PPm PPm 

VIRGINIA 

Fauquier Co. 

Marion 

Loves Mill 

B rowntown 

Waynesboro 

Skyland 

NORTH CAROLINA 

Lenoi r 

Alexander Co. 

Edgecornbe Co. 

Farmvi 1 le 

Martin Co. 

Johnston Co. 

Montgomery Co. 

Croatan NF 

SOUTH CAROLINA 

Chester Co. 

TENNESSEE 

Giles Co. 

KENTUCKY 

Calloway Co. 

Land between Lakes 

Fulton Co. 

Russellville 

Paradise 

Metcalf Go. 

Pulaski Co. 

Clay Go. 

Prestonburg 

LOUISIANA 

OeR i dder 

Gal 1 i ano 

Morgan City 

Source: Pinkerton and Lefohn (1986). 
b) In many rural forests more than 50 km from a city with a 

popul ation over 50,000 i n the Southeastern United States, 

7-hour average ozone concentrations in the summer months of 

1983 ranged from 50 to 70 ppb with I-hour peak concentrations sometimes exceeding 100 or even 120 ppb (see table 7) 

(Pinkerton and Lefohn 1986). 

c) Although the 7-hour average concentrations in many urban 

forests, near-urban forests, and some rural forests 

near major point sources (see table 8) were about the 

same as those reported for rural forests (see table 7), 

this is not true of the highest 1-hr peak concentrations. 

For example, the data of table 8 show that 7-hour average 

summertime ozone concentrations for a selection of southeastern urban areas ranged from about 40 to 60 ppb--about 

the same as the 50 to 70 ppb shown for southeastern rural 

forests in table 8. But these same urban areas showed 

1-hour peak concentrations in 1983 between 150 and 200 ppb 

-- substantially higher than the 102 to 143 ppb shown for 

rural forests in table 8. The data in table 8 also show 

that the six monitoring sites in Washington, DC, had ozone 

concentrations that exceeded 100 ppb for as long as 42 to 

250 hours. Similarly, the two sites in Atlanta, GA, and 

Norfolk, VA, and the three sites in Louisville, KY, and 

Charlotte, NC, had ozone concentrations that exceeded 

100 ppb for as long as 122 to 134 hours, 74 to 97 hours, 47 

to 167 hours, and 26 to 133 hours, respectively (Pinkerton 

and Lefohn 1986). The 1980-85 Environmental Protection 

Agency SAROAD data for rural sites in the northeastern and 

midwestern States showed average and maximum concentrations 

that were similar to those measured in the southeastern 

States. Unfortunately, however, the number of sites in the 

southern States was much smaller than those in the northeastern and midwestern States (Evans and others 1983). 

d) Regional variability. The highest 1-hour peak concentrations of ozone in the Eastern United States occur in New 

England, the mid-Atlantic States, and the Gulf Coast. From 

1981 to 1983, peak concentrations of 170 ppb, 180 ppb, 

250 ppb, and 280 ppb were reported for Raleigh, NC; Boston, 

MA; Newark, NJ; and Houston, TX, respectively (U.S. 

Environmental Protection Agency 1986). 

The high concentrations of ozone that accumul ate in urban, near-urban, 

rural, and even high-elevation areas are mainly associated with significant 

emissions of volatile organic compounds and nitrogen oxide precursors. 

Secondary pollutants such as ozone and other photochemical oxidants often are 

not formed until approximately 1 day after emission of their precursors. These 

high concentrations of ozone can be transported over distances of 100 to 

300 km, during which time such air masses can pick up additional ozone as they 

travel over other urban areas. For example, transport of high concentrations 
t 

0 

~'ds 

3 rll .I$ E xcea 

sara 

= E 

0 

U 

E 

L a 

n 0 

3 6 0 .. C 

Ub ca 

OI Y . L .-- Er *P . I 

L (Oh 

P 

cr 0 

0 

m 

cc m 0 c 

-0- 

LLVl 

worn 

act ar 

5 .E & z 0 

r 

C 

.- 8 Vl c Qcta 

ua rn-- m-Fpl =? - 

f5iE 

6% 

C 

(O 

.F 4 

I-- m Q aJ a L e": 

cr 

+: 

dddddddddddddd ddddddddd d d IittIIlll~lfltlllllltllltltfl 

I1 fill 811 llllllllll1111111111 

dddddddddddddd ddddddddd d d 

0 

D w yl 

V) i--- Qi tar: 3C t QI O, =a u& w o mC 2wQ1C B a - =rer roo o~ z r cr aJ .-F-tO 3 &OF 0- L4OQa03 8 mc m- v).-- mar c ocr K.--ECYF- o V) re- ac+~ ccd c 04 L.- 3.- t v,- om o o w 0; .F aw ma at 0 L o ~ZI V-cr car cn >--- e 6 v, 6- E r- c >- WIF-4 E- L 0 cn r c v, m a.-r E area ca LC w Q, o X.F- w~ct =I& o~ :s a.-~-- #S-ZE3MLatLU-rllUW-&Q5Ktatac&Lrrj%UX 

(O ocr EU w D (o.- L o ta L.C.- L w rn t o.-er o-s: a~ (B e, 31ur~~rm~z70u~a~m~~~0~~~~1as%
of ozone from the western Gulf Coast area to the midwestern and the northeastern States has been observed (Wolff and Lioy 1980). Under similar meteorological conditions, transport of high concentrations of ozone from the New 

York, Philadelphia, and Washington, DC, corridor into Virginia, West Virginia, 

western Pennsylvania, and Ohio has been reported (Fankhauser 1976). 

For all of the above reasons, transport of ozone and other pollutants from 

urban and industrial ized areas to near-urban, rural, and even high-elevation 

forests is the most important factor determining ozone exposure of forests 

throughout eastern North America. The highest ozone concentrations occur i n 

summer and fall when sunlight is most intense and stagnant meteorological 

conditions augment the conditions necessary for formation and accumulation of 

ozone. 

e) Diurnal variability. The patterns of ozone occurrence in 

ambient air depend on daily and seasonal variations in the 

concentrations of precursor chemicals in the air, the 

ambient temperature, and the intensity of sunlight. In 

most urban areas in eastern North America, the typical 

diurnal pattern (observed at urban, near-urban, and rural 

monitoring sites) has a minimum ozone concentration near 

zero. Thi s occurs around sunri se, then i ncreases through 

the morning to an early afternoon peak concentration, and 

finally decreases toward minimal concentration again in the 

evening (fig. 18). Thus, high concentrations of ozone 

may occur for 6 to 10 hours per day. 

An important variation on this basic diurnal theme occurs when a secondary 

peak appears in addition to the primary early afternoon peak. This secondary 

peak occurs only in some localities. It may occur any time from mid-afternoon 

to the middle of the night. It is attributed to ozone transported from an 

upwind area where high concentrations accumulated earlier in the day. Secondary peak concentrations may be higher than those in the primary peak. For 

example, at one rural site in Massachusetts, primary peak concentrations of 

approximately 110 ppb, 140 ppb, and 140 ppb occurred at noon, or from noon to 

4 p.m. , respectively, on 3 successive days. Secondary peaks at those same 

sites for the same 3 days were 150 ppb, 157 ppb, and 130 ppb; these occurred 

at 6: 00 p.m. , 8: 00 p.m. , and 8: 00 p.m. , respectively (fig. 19) (U. S. Environmental Protection Agency 1986). 

High-elevation forests often experience no important diurnal variability. 

This is in distinct contrast to the pattern in low-altitude forests. One study 

at a rural site in New Jersey characterized ozone concentrations along an 

elevation gradient (see fig. 20) (Wolff and others 1987). Peak ozone concentrations were similar at each elevation, but the lower concentrations increased 

with elevation. Ozone concentrations at this site varied during the day by as 

much as 100 ppb. 

f 1 . During episodes of high ozone 

concentration, diurnal patterns may vary from day to day 

with peaks occurring at any time during the day or night. 

At many monitoring sites throughout eastern North America, 
0.19 

0.18 

0.17 

0.16 % 0.16 

0.14 8 0.13 

p 0.12 

0.11 

0.10 

y 0.0s 

z 0.m 

0 0.07 (: 0.08 

0.06 8 0.04 

O 0.03 

0.02 

0.01 

0 

121 2 3 4 5 6 7 8 91011 t 1 2 3 4 5 6 7 8 91011 

NOON 

+-a.m. HOUR OF DAY- p.m. 4 

Figure 18. Diurnal pattern of I -hour ozone concentrations on July 13, 1979, 

Philadelphia, PA. 

Source: SAROAD (1 985). 
a.m. NOON p.m. 

SATURDAY, 27 AUGUST 

a.m. NOON p.m. 

SUNDAY, 28 AUGUST 

a.m. NOON p.m. 

MONDAY, 29 AUGUST 1977 

Figure 1 9. Three-day sequence of hourly ozone concentrations at 

Montague, MA, SURE station showing locally generated midday peaks 

and transported late peaks. 

Source: U.S. Environmental Protection Agency (1 986). 
71 122 7123 7124 

Time (EDT) 

Figure 20. Ozone concentrations from July 21-24, 1975, at three 

altitudes at High Point. NJ. 

Source: Wolff and others (1 987). 
episodes sf high csncentrations of ozone occur more frequently than episodes of high sulfur dioxide or nitrogen 

oxide concentrations. 

4) 

of SO,, 

NOx, and 0, have been investigated at 30 monitoring sites in the 

Eastern United States. During 1978-81, co-occurrences of 

concentrations sf > 58 ppb of all three pollutants were observed 

very rarely, i .e. , < 1.5 percent of the monitoring hours 

analyzed (tefohn and lingey 1984). This was also true for 

co-occurrences of NOr and O3 and NOx and SO2. But SO2 and O3 

co-occurred about 30 percent of the time. Sulfur dioxide and 

O3 co-occurred because O3 was very often present at concentrations greater than 50 ppb, and SO2 concentrations greater than 

50 ppb occurred periodically (Lefohn and Tingey 1984). 

Scavenging of NOx during the formation of ozone explains why 

NOx and ozone rarely occur together--NOx concentrations are 

usually lowest during the day when ozone concentrations are 

usual ly highest (see fig. 19) (U. 5. Envi ronmental Protection 

Agency 4986). 

Concentrations of PAN, PPN, hydrogen peroxide, ammonia, 

nitrogen oxides, and other oxidant gases usually are lower 

than concentrations of ozone and sulfur dioxide but in some 

areas may equal them, Concentrations of these other 

oxidants including PAN and PPN are usually below the 

detection l imi ts of most available monitoring instruments. 

The reason for considering these pollutants in this chapter 

is the possibility that they co-occur with other airborne 

pollutant chemicab with which they may cause additive or 

synergistic effects on vegetation. 

PAN and PPN concentrations typically occur at about 4 to 

12 percent and 1 to 2 percent, respectively, of the ozone 

concentrations ad any given loca%ion (Altshuller 1986; 

U.S. Environmental Protection Agency 1986). There are no 

available measurements of PAN or PPM for the Southeastern 

United States. If we use the above percentages to convert 

measurements of ozone concentration do estimate PAN and 

PPN concentrations however, PAM and PPN in combination 

may occur at ?-hour average concentrations of 8 to 21 ppb. 

c) Only since 1984 have gas-phase concentrations of H202 been 

measured reliably. Available data suggest that maximum 

hydrogen peroxide concenlrat?'ons occur in summer in the 

EasLern United States. Measurements at Whiteface Mountain, 

NY, and Whitetop Mountain, VA, indicate l-hour peak concentrations of 60 and 80 ppb, respectively (fig. 21) (Hei kes 

and others 1987; Mohnen 1989). 
24 

22 IIVhitafsca Mt. 

20 

e 18 

0 

.I 

16 L: 

= 14 8 

.c, 12 c. 

C 

'" 10 

g a8 

6 

4 

2 

0 

0 0.3 0.7 1 .I 1.6 1.9 2.3 2.7 

Hai Concentration, ppbv 

Figure 21. Distribution of hourly, summertime concentrations at two high 

elevation sites. 

Source: Heikes and others (1 987). 
d) Only very sparse data for ammonia and nitric acid exposure 

are available, Measurements of ammonia and nitric acid at 

three southeastern sites averaged approximately 1 to 3 ppb 

and tl ppb, respectively; in many other studies, concentrations were usually below the detection lislits for the 

instruments used (Cadle and others 1982; Kelly and others 

1984). Possi bl e co-occurrence of these conupczunds with 

other airborne pollutant chemicals cannot be described frm 

these few measurements. 

Acid deposition: The acidity of precipitation and cloud water 

(hydrogen ion concentration, usually expressed as pH) is determined by the balance among a1 1 the airborne cations and anions 

dissolved in these components of wet deposition. Similarly, the 

acidifying effects on forest ecosystems of all the airborne 

gases, aerosols, and dissolved and suspended substances in 

precipitation and cloud water are determined by the balance 

among the airborne cations and anions transferred from the 

atmosphere, and the exchange and transformation processes that 

take place with the above- -and below-ground parts- of the plants 

and soi 1 s of the fqrest., Thus+ the~rincip$l acidity d-eterrnio- - ing, ions include K , Na , WH4 , Ca , Mg , C1 , NO3 , SO4 , PO4-. Among these, ammonium and hydrogen are the most important 

cations, while sulfate, nitrate, and occasionally chloride are 

the most important anions. Ammonium ion has a distinctive role 

since it is its effects in precipitation and 

cloud water in its effects when taken up by 

plants, animals, and microorganisms in forest ecosystems. 

a) Total loading. The total loading of sulfate, nitrate, 

ammonium, and hydrogen ions into forest ecosystems is 

determined by the wet and dry deposition of five types of 

substances, the first four of which undergo conversion and 

dissolution processes as soon as they contact the moist 

surfaces of plants, animals, microorganisms, soi 1 s, rocks, 

and water films, droplets, and pools: 

Gaseous sulfur dioxide is further oxidized to sulfate 

ions in solution; 

Gaseous nitrogen oxides (NO, and NO), PAN, PPN, and 

nitric acid vapor are all further oxidized to nitrate 

ions in solution; 

* Gaseous ammonia is first dissolved in water and then 

di ssoci ated i nto ammoni um ions i n sol uti on; 

Airborne aqueous or dry aerosols (which consist mainly of 

ammoni um nitrate, ammoni um sul fate, and ammoni urn acid sulfate) 

dissolve and then dissociate into a mixture of sulfate, 

nitrate, ammonium, and hydrogen ions in solution; and all 

of the above are mixed with 

cloud water and precipitation that already contain a 

mixture of sulfate, nitrate, ammonium, and hydrogen ions in 

solution. 
If the gases and aerosols listed above are deposited on truly moisturefree surfaces, it will stiff be only a short time (usually only a few hours or 

days) before these surfaces will be wetted by dew, by precipitation, or by fog 

or cloud water. Thus, the end-products of all these dissolution, dissociation, 

and mixing reactions will be incorporated into the soil solution or into the 

moisture that bathes the above-ground portions of the plants, animals, microorganisms, soil particles, and rocks of forest ecosystems. The sulfate, nitrate, 

ammonium, hydrogen, and other cations and anions in these solutions will then 

be free to diffuse: 

* through stomata, cuticle, cell walls, and cell 

membranes on their way into the cytoplasm of leaves of 

trees and other Porest planls, or 

* into the soil pores, soil colloids, cell walls, cell 

membranes, and on into the cytoplasm of the fine 

feeder roots or wcorrhizae of forest vegetation. 

cations and anions in precipitation are made weekly -at 

wet-deposi ti on moni tori ng sites in various parts of North 

America (NADP 1985). At present there are about 200 such 

sites in the United States and about 60 in Canada. Thus, 

reliable maps are now available describing the geographical, seasonal, and multi-year variability in the concentration and deposition of sulfate, nitrate, ammonium, 

hydrogen, and other ions in precipitation (see figs. 22 

through 25). Year-to-year variability in these maps for 

both concentration and deposition i s fai rly small ; thus, 

the maps shown in these figures for 1985 are representative 

of the concentration and deposition of these ions in 

precipitation in various parts of eastern North America in 

the years between 1980 and 1986. 

These maps show the following regional trends in concentration and 

deposition in precipitation across eastern North America: 

.) Sul fate concentration in precipitation varied from a 

maximum of about 3 milligrams per liter in portions of 

Ohio, Pennsylvania, and Ontario and decreased progressively 

to less than 1 mg/L in portions of Quebec and Ontario to 

the novth and North Dakota and Texas to the west, and to 

between 1 and 1.5 milligrams per liter across the Southeastern States (see fig. 22a). 

(r in precipitation varied from a maximum 

of about 35 kiloqrams per hectare in northern Ohio, and 

decreased progre;sively to l ess than 10 ki l ograms per 

hectare in portions of Quebec and Ontario to the north and 

the States between North Dakota and Texas to the west, and 

to between 10 and 20 kilograms per hectare across the 

southern States (see fig. 22b). 
1986 Annual 

Sulfate Concantration 

1988 Annual 

Sulfate Depollition 

Figure 22. The 1985 annual distribution of SO,": (a) concentration and 

(b) deposition. Contour intervals are (a) 0.5 rng/L and (b) 0.5 g/rn2. 

Source: Barchet (1 987). 
19815 Annual 

Nitrate Concentration 

1986 Annual 

Nitrate Deposition 

Figure 23. The 1985 annual distribution of NO;: (a) concentration and 

(b) deposition contour intervals are (a) 0.5 rng/L and (b) 0.5 g/m2. 

Source: Barchet (1 987). 
Figure 24. The 1985 annual distribution of NH,+: (a) concentration and 

(b) deposition. Contour intervals are (a) (0.1 mg/L and (b) 0.1 g/rn2. 

Source: Barchet (1 987). 
1986 Annual 

Figure 25. The 1985 annual distribution of (a) precipitation - weighted 

pH and (b) H+ deposition. Contour intervals are (a) 0.2 pH units and (b) 

20 mg/m2. 

Source: Barchet (1 987). 
. Nitrate concentration in precipitation varied from a 

maximum of about 2 mg/L in portions of Michigan, Ontario, 

and New York, and decreased progressively to less than 

I mg/L in portions of Quebec and Ontario to the north, the 

States between Minnesota and Kansas to the west, and the 

States between Arkansas and North Carolina to the south 

(see fig. 233). 

Nitrate deposition in precipitation varied from a maximum 

of about 20 kilograms per hectare in northern Ohio, and 

decreased progressive1 y to 1 ess than 10 ki 1 ograms per 

hectare in portions of Quebec and Ontario to the north, the 

States between Minnesota and Oklahoma to the west, and the 

States between Arkansas and North Carolina to the south 

(see Fig. 23b). 

. Ammoni um concentration in precipitation varied from a 

maximum of about 0.5 milligrams per liter in portions of 

Michigan and Ontario, and decreased progressively to 1 ess 

than 0.2 mil 1 igrams per liter in por'tions of Quebec and 

Ontario to the north, the States between North Dakota and 

Colorado to the west, and the States between Texas and 

North Carolina to the south (see fig. 24a). 

Ammonium deposition in precipitation varied from a maximum 

of about 5 kilograms per hectare in portions of Ohio and 

Ontario, and decreased progressively to less than 

1 kilogram per hectare in portions of Quebec and Ontario to 

the north, the States between North Dakota and west Texas 

to the west, and to less than 2 kilograms per hectare 

across the Southeastern States (see fig. 24b). 

The volume-weighted average pH of precipitation varied from 

a minimvm of about 4.2 in portions of Pennsylvania, Ohio, 

and West Virginia to a maximum of above 4.5 in portions of 

Ontario and Quebec to the north, a maximum of 5.0 in the 

States between Minnesota and Texas to the west, and a 

maximum between 4.4 and 4.8 across the southern States (see 

fig. 25a). 

in precipitation varied from a 

maximum of about 600 grams per hectare in portions of 

Pennsylvania, Ohio, and west Virginia, and decreased 

progressively to less than 200 grams per hectare in 

portions of Quebec and Ontario to the north, less than 

IOQ grams per hectare in the States between Minnesota and 

Texas to the west, and to between 200 and 400 grams per 

hectare across the southern States (see fig. 25b). 

In general, sulfate concentrations are greater in summer than in winter. 

In a broad band extending from New England southwestward along both the eastern 

and western sides of the Appalachian Mountains as far as western North 
Carol ina, the ratio of summertime to wintertime sulfate concentration is 

greater than two to one. 

Wet deposition of sulfate in precipitation in warm periods is highly 

episodic -- 50 to 70 percent of the total annual deposition of sulfate occurs 

in about 20 percent of the total days (Summers and others 1986). 

Wet deposition of nitrate ions in precipitation is also highly episodic -- 50 to 70 percent of the total annual deposition of nitrate occurs on about 

20 percent of the total days per year. High nitrate deposition occurs more 

often in warm than in cool periods, but does not show as strong seasonal 

dependence as sulfate deposition (Summers and others 1986). 

Sulfate and nitrate concentrations in cloudwater have been measured at a 

few high-elevation sites in the northeastern States and in Virginia. In 1984, 

sulfate concentrations ranged from about 930 to 1850 microequivalents per 

liter, while nitrate concentrations in 1984 ranged from about 190 to 900 microequivalents per 1 iter (Weathers and others 1986). High elevation forests in 

the northern and southern Appalachian Mountains are immersed in clouds on more 

than ha1 f the total days per year (Mohnen 1987). 

The pH of cloudwater in both the northeastern and southeastern States 

typically averaged 3.5, with highly acidic events of around pH 2.6 (Mohnen 

1987; Weathers and others 1986). 

Regional variability in dry deposition of sulfate, nitrate, 

ammonium, and hydrogen ions. In contrast to the situation 

for wet deposition, no reliable measurements of regional 

variability in dry deposition of gases and aerosols leading 

to the accumulation of sulfate, nitrate, ammoni um, and 

hydrogen ions in forest ecosystems have been made in North 

America. The fragmentary information that is available is 

based mainly on short-term studies in a few specific 

locations and on ambient air qua1 i ty monitors establ ished 

mainly in urban areas. The amounts of gases and aerosols 

present in the atmosphere of most rural, high-elevation - 

remote, and even many near-urban locations are so low that 

they are at or below the detection limits for the monitors 

that are used. These instruments were designed and 

instal 1 ed mainly for the purpose of insuring compl i ance 

with human-health-related air quality monitoring in cities. 

Thus, these data cannot be used to give reliable estimates 

of air concentrations or dry deposition of sulfate, nitrate, 

ammonium, and hydrogen ions. 

The only direct measurements of the regional distribution of gaseous 

sulfur dioxide and sulfate aerosol were made in 1977-78 during the Sulfate 

Regional Experiment conducted by the Electric Power Research Institute. As 

shown in figure 15: 
Gaseous sulfur dioxide concentrations varied from an 

arithmetic 1-hour average of about 20 ppb in portions of 

West Virginia and Pennsylvania, to less than 10 ppb in 

Ontario and New England to the north, portions of Michigan 

and Illinois to the west, and almost all of the southern 

States to the south (see fig. 15a); 

Sulfate aerosol concentrations decreased from a maximum 

geometric 24-hour average of about 8 micrograms per cubic 

meter i n portions of A1 abama, Kentucky, and Pennsylvania, 

to a minimum of less than 4 micrograms per cubic meter in 

New England and Ontario to the north, the States between 

Wisconsin and Missouri to the west, and to less than 

6 micrograms per cubic meter in the Gulf and Atlantic Coast 

States (see figure 15b) (Mueller and Hidy 1983). 

Thus, sulfate aerosol is more uniformly distributed, on the average, than 

sul fur dioxide, its gaseous precursor. The more uniform distribution results 

from the dispersion, mixing, and long-distance transport of sulfur dioxide 

during its conversion to sulfate (Mueller and Hidy 1983). High concentrations 

of sulfate aerosol frequently accumulate in stagnant air masses as sulfur 

dioxide is converted to sulfate aerosol. These high sulfate episodes sometimes 

persist. for several days and extend over large areas of eastern North America 

(Muel l er and Hidy 1983). 

Unfortunately, similarly reliable data for gaseous nitrogen oxides and 

nitrate aerosol are not available. 

In theory, it should be possible to calculate the amounts of sulfate, 

nitrate, ammonium, and hydrogen ions transferred from the atmosphere into 

forest ecosystems in various regions of North America by multiplying the values 

for regional measurements of the air concentration of gases and aerosols by the 

highest and lowest reasonable estimates of the deposition velocity for these 

substances. Such broad regional estimates of dry deposition of gases and 

aerosols have not found wide acceptance, however (Hicks 1984). 
4. EFFECTS OF AIRBORNE SULFUR- AND NITROGREN-DERIVED POLLUTANT 

CHEMICALS ON FOREST ECOSYSTEMS OF EASTERN NORTH AMERICA 

4.1 

Cl imate? 

Mature forest ecosystems are seldom stable, therefore, assessing ecosystem 

responses to stresses i s extremely di ff icul t. They mai ntai n themsel ves i n an 

oscillating steady state characterized by the continuous elimination of 

suppressed and old trees and the addition of new ones. Intense competition 

among trees for light, water, nutrients, and space, as well as recurrent 

natural climatic and biological stresses (such as those cited below), alter the 

species composition of mature forests by eliminating those individuals sensitive to specific stresses. Competition among trees of the same species does 

not affect the composition of stands and does not influence species succession. 

Competition among different species, however, results in succession and ultimately produces a mature forest with forest trees and other plant species 

having a capacity to tolerate the competitional stresses (Kozlowski 1980; 

Kozlowski and Constantinidou 1986a). Succession, unless there is a catastrophic disturbance, can take 100 to 1000 years. 

In a mature forest, a mild disturbance has little effect on its oscillating 

steady state. However, if injury or disturbance, whether from air pollution or 

other stresses, leads to changes in the species composition of the forest 

severe enough to disrupt food chains and modify the rates of energy flow and 

nutrient cycling, succession may be returned to an earlier, less complex, 

successional stage (Bormann 1985; Wosdwell 1970). 

Ecosystems respond to changes in their environment through the organisms 

that comprise them. Forest ecosystem responses to stress begin with the 

interaction of the individual organism with its environment. Three levels of 

interaction are involved: (1) between the individual tree and its environment; 

(2) between the population and its environment; and (3) between the forest (the 

biological community composed of many species) and its environment (Billings 

1978). 

In most ecosystems, the principal interaction among species is competition 

for resources. In their struggle for existence, forest trees are continually 

undergoi ng one or more of the fol 1 owing natural environmental stresses (Woodman 

and Cowling 1987): 

1. Natural stresses. These occur whenever trees 

compete with each other or with other plants for limited supplies of growing space, solar radiation, water and/or essential 

nutrients from the same soi 1 or microcl imate. Competi tional 

stresses are the dominant cause of low vigor, poor growth, and 

death of trees in most forests. More trees die of competitional stresses than from all other stresses combined (Woodman 

and Cowling 1987); 
2. Natural climate stresses. These include extremes of high and 

low temperature, drought, flooding, low humidity, and high winds. 

Water stress is the most common cause of decreased growth, low 

vigor, premature loss of fol iage, and mortal i ty (Kramer 1983). 

Temperature and the availability of water have the qreatest 

impact on growth in the northern hemisphere (Osmond and others 

1987) ; 

3. Natural biological stresses. These include parasitic and 

pathogenic fungi, insects, nematodes, bacteria, viruses, 

viroids, plasmids, mycoplasmas, parasitic seed plants, and 

herbivorous animals such as porcupines. In forests, fungi and 

insects are the most important biotic pathogens. Any of these 

biotic stresses can cause impairment of normal physiological 

processes. The effects of these biotic agents vary over time 

because their activities depend upon the environmental requirements of the predators and pathogens (Osmond and others 1987); 

4. Natural chemical stresses. These are induced by deficiencies 

(andcasional ly excesses) of essential nutrients or toxic soi 1 

chemicals such as aluminum. Nitrogen is the nutrient most 

commonly deficient in forests. Virtually all of the nitrogen 

and sulfur entering a forest ecosystem comes from the atmosphere 

(Waring and Schlesinger 1985). 

Changes that occur in the physical or chemical climate of plant populations apply new selection pressures (Treshow 1980a; Woodwell 1970). The 

subtle and indirect effects of pollutant dosages can set the stage for changes 

in community structure that may possibly have irreversible consequences 

(Guderian and Kueppers 1980). Increasi ng pol l utant stress provides a 

selective force that favors some genotypes, suppresses others, and eliminates 

those species that lack sufficient genetic diversity to survive. Thus, the 

occurrence and distribution of plants is influenced, and community composition 

and species interactions are changed such that the basic structure of the 

ecosystem is ultimately changed (Treshow 1980a). This succession may take 

years, decades, or longer, depending on the pollutant concentration, period of 

exposure, and plant species involved. 

Man-made stresses that change the chemical and physical climates of plants 

are of two main types: 

1. - Air pollution stresses. These are chemical stresses that occur 

whenever forest trees are exposed to in.jurious concentrations of 

toxic gases such as sulfur or nitrogen oxides, ozone, or 

f 1 uoride; toxic aerosol particles or coarse particulate matter; 

or dissolved or suspended chemicals in air, cloud water, or 

precipitation. Pollutant stresses can also occur as a result of 

accumulation or mobilization of toxic substances in soils often 

after wet or dry deposition from the atmosphere (Woodman and 

Cowl i ng 1987). 
2. Mechanical stresses. These result from various human activities 

i ncl udi ng l oggi ng, control l ed burni ng , drai ni ng , f 1 oodi ng, and 

physical disturbance of soils leading to compaction, erosion, 

leaching of nutrients, or accumulation of toxic substances 

(Woodman and Cowling 1987). 

The above stresses, both natural and anthropogenic, can occur at the same 

time or in any order, one after another. They can act independently, additively, synergistically, or antagonistically; they can be periodic, either 

short- or long-term, and can increase or decrease with the age of the forest 

stand (Cowling 1985). 

Air pol l ution stresses general ly result not because gases and chemical s 

are unique substances never before encountered by plants, but rather because 

the amounts of these substances are beyond the limits of tolerance or avoidance 

of the organism. Acute air pollution stress tends to be episodic. Trees 

experience high pollutant exposures for minutes, hours, or for a few days to a 

week at the most. These acute exposures may occur several times during a year. 

In chronic air pollution stress, however, low concentrations are experienced 

continuously for a considerable portion of the life of the plant. 

In theory, airborne chemicals in the form of gases or aerosols, when suspended in air or dissolved in cloud water or precipitation, could cause, 

either alone or with other stress factors, many different types of effects in 

forests. These include the following: 

Effects on Individual Trees: 

1) Visible symptoms of injury to individual trees- 

* Change in shape and size of leaves, 

Change in normal patterns of growth and development (shape and 

timing of development of leaves, flowers, buds, branches, 

roots, etc, ) , Change i n normal senescence patterns of 1 eaves, sel f-pruni ng 

of branches, or turn-over of feeder roots, 

Change in life expectancy of individual trees; 

2) Decreased annual height, growth, or radial increments (with or 

without visible symptoms of injury to individual trees); 

3) A1 teration of physiological processes (with or without visible 

symptoms)- 

* Photosynthesis, 

* Respiration, 

* Transpi ration, 

Mineral nutrition, 

Transport and allocation of photosynthate, 

* Reproduction, 

Hormonal control of growth, 

Symbiotic relationships with other organisms; 
4) Changes in susceptibility to other stress factors; 

5) Changes in reproductive behavior of individual trees; 

Effects on Forest Stands: 

5) Decreased productivity of whole forest stands; 

7) Change in stand stocking, such as age-class distribution; 

8) Changes in normal patterns of competition and mortality within 

forest stands (both among individual trees and between trees 

and other vegetation in forest stands); 

9) Changes in normal patterns of successian within forest stands; 

10) Changes in species composition of forest; 

11) Changes in nutrient cycling in forest stands (such as excessive 

leaching of nitrate or other mobile anions or cations from 

soils); 

12) Changes in hydrologic behavior and watershed functions of forest 

stands ; 

Effects on Tree S~ecies: 

13) Changes in the genetic diversity within tree species (alteration 

of gene pools within tree species); 

14) Changes in fitness within and between tree species; 

Effects on Reaional and Social Values of Forests: 

15) Changes in the regional productivity, distribution, or economic 

value of forests; and 

16) Changes in aesthetic qua1 ity or other ecosystem values for 

humans. 

Fifteen of these 16 theoretical effects have been confirmed by observation 

and experiments in forests of North America. In both North America and Europe, 

studies of the effects of air pollutants on forests and agricultural systems 

have been made in the vicinity of strong point sources of sulfur dioxide 

(Legge and others 1981; Linzon 1978; Miller and McBride 1975; Scheffer and 

Hedgcock 1955; Winner and Bewley 1978a,b; Winner and others 1978); fluoride, 

(Miller and McBride 1975; Treshow and others 1967); and chronic radiation 

(Woodwell 1970). Much less is known about the effects of regionally dispersed 

airborne chemicals. Ozone has been the most studied regional pollutant 

(Mi 1 ler 1973; Miller and others 1982). Only number 15 1 isted above has not 

been associated with regional ozone exposure in the San Bernardino Forest of 

Cal i fornia. 
Air pollutants may affect trees directly through the foliage or indirectly 

via the soil. Regardless of whether the air poglution stress is acute or 

episodic, forest trees or perennial shrubs must cope with the cumulative 

effects of both short- and long-term stresses. Tree response may appear 

rapidly as when, for example, sensitive white pine needles show visible injury 

from exposure to high concentrations of ozone. In other instances, however, 

responses are often subtle and may not be observable for many years because 

adaptation and response to stress is expressed by the differential growth that 

results from changes in carbon allocation (Waring and Schlesinger 1985). 

Tree populations play a critical role in mature forest ecosystems. They 

represent the later stages of succession and are adapted for high competitive 

ability (Brown 1984). As the dominant producers, trees influence the structure (species composition and trophic relationships), energy flow, and nutrient 

re1 ati onships (Ehrl ich and Mooney 1983). The ecosystem processes of energy 

flow and nutrient cycling are directly related to the plant physiological 

processes of photosynthesis, nutrient uptake, biosynthesis, carbon allocation, 

respiration, and translocation. The alteration of these physiological processes 

is the fundamental cause of all other ecosystem effects (U.S. Environmental 

Protection Agency 1986). 

Changes in ecosystems begin with the death of individual plants. Temporal 

dynamics at the level of individual organisms can, therefore, upset the equilibrium and be disorganizing in ecosystems such as forests that are dominated 

by nonmobile organisms (Shugart 1987). In forests with large canopy trees, a 

canopy tree dominates the space where it is growing, reduces the amount of 

light reaching the forest floor and alters the microclimate, conditions that 

help to determine the plant species that can survive beneath the canopy. Death 

of a tree opens the canopy and changes the microclimate and substantially 

increases the resources of light, nutrients, water and energy available to 

other organisms. This change initiates a struggle for dominance among the 

understory trees and seedlings. In time a new canopy becomes established 

(Shugart 1987). 

The mode of death of a tree is also ecologically important as it determines the regeneration success of trees that form the next forest generation. 

Trees may die catastrophically as when high winds or ice storms break off the 

crown or branches or when they are blown over, exposing the roots; or they may 

die slowly and tend to waste away as in the case of those injured by pathogens 

or insects (Shugart 1987). Tree death, as influenced by gaseous air pollutants 

such as ozone, is usually gradual rather than catastrophic unless the tree is 

extremely sensitive or the pollutant concentration is extremely high. Growth 

responses require time, Therefore, because growth responses are the result of 

cumulative stresses and trees are continual ly being subjected to many other 

stresses, the cause of death is difficult to determine. 

Pollutant stresses, as well as natural and biological stress as indicated 

above, have important implications for forest succession because they provide 

forces that favor some genotypes, affect others adversely, and eliminate 

species that lack genetic diversity (fig. 26). A population tolerant to the 

dominant stresses becomes established. However, studies using heavy metals and 

herbicides indicate that once the dominant stress is removed, the pollutanttolerant plants tend to decline in number. Competition increases selection for 
PROCESS 

Figure 26. Pathways of air pollutant impact in forest ecosystems. 

Source: McClenahen (1 984). 
- 

tolerance under polluted conditions and against tolerance under less polluted 

conditions (Roose and others 1982). This evidence is corroborated by observations of ecosystems functioning under natural conditions. Certain terrestrial 

ecosystems require major disturbances (e.g., fire, drought, windstorms) to 

retain their characteristics (Smith 1980; Vogl 1980). The development of 

tolerance depends upon the genotype being present in the plants when they are 

growing in environments free from air pollution or other selective stresses 

(Roose and others 1382). 

Responses at the forest ecosystems level, because they are the summation 

of mu1 tiple responses of individual organisms and their interaction, usual ly 

take longer to develop, are more diffuse, and are usually of longer duration 

than responses at the individual or population level. Therefore, the cause and 

the response are more difficult to determine. In addition, forest stands 

differ greatly in age, species composition, stability, and capacity to recover 

from disturbance. Therefore, data dealing with the responses of one forest 

type may not be applicable to another forest type (Kozlowski 1980). 

Explanations for changes observed at the ecosystem level often can be 

determined by understanding processes occurring at physiological, population, 

and community 1 eve1 s (Wari ng and Schlesi nger 1985). Responses at the i ndividual tree level are determined by the extent that physiological processes within 

the tree are altered or inhibited by exposure to air pollution (McClenahen 

1984). It is now generally accepted that air pollution injury to individual 

organisms is initiated at the biochemical level. Changes induced in cells lead 

sequentially to changes in metabolic pathways of the whole organism, a disorganization of cellular membranes resulting in a loss of control of diffusion of 

water and solutes, and a breakdown of cellular components that finally results 

in visible symptoms such as chlorosis and necrosis of foliar tissue. Growth 

reduction, loss in yield (i .e. , reproduction) and in plant quality can occur 

without visible injury to the plant (U. S. Environmental Protection Agency 

1986 ; Koz 1 ows ki 1980). 

Studies of the responses at the cellular level of trees most sensitive to 

specific stresses, both natural and anthropogenic, could lead to the identification of those physiological processes most likely to be affected by changes 

in the physical and chemical climate of forest ecosystems. Concentrations of 

the chemical compounds, the length of exposure, and the amount of the compound 

taken up are critical aspects of each individual plant response. Responses at 

the cellular level are determined in part by these factors. In the next 

section, the concepts of exposure, concentration, and uptake are discussed. 

42 What is the Likelihood That Acid De~osition and Other Airborne Sulfurand Mitroaen-Derived Chemicals Affect the Forests of Eastern North America J 

? 

Plant response to air pollution stress involves pollutant concentration, 

length of exposure, deposition, uptake, and alteration of biochemical 

processes. Gaseous ai rborne chemical s (SO NO and mixtures of them) 

cause phytotoxicity of trees and other pflnts xd y '2' ntering the leaves and 

disrupting the biochemical processes that occur within the cells. The severity 

of injury usually is determined by the extent physiological processes are 

altered by the amount of pollutant entering the leaves. Plant response cannot 

be related solely to the concentration of the pollutant gas in the ambient air 
(Unsworth and others 1976). The factors influencing chemical deposition, 

concentration, uptake, and the manifestation of injury are discussed in the 

following section. Plant response may be characterized by a measurable change 

in any plant function such as respiration, photosynthesis, rate of gas exchange, 

biochemical pathways, physiological reactions, degree of visible leaf injury, 

or decrease in growth or reproduction (yield) (U. S. Environmental Protection 

Agency 1986). 

Visible injury to foliage does not necessarily mean that decreases in 

growth have occurred or are occurring. It is, however, an indication that 

phytotoxic concentrations are present and that studies should be conducted to 

assess the risk to vegetation from the concentrations present. Many studies, 

primarily of crop plants, have been made to assess the effect of a particular 

concentration and period of exposure. Table 9 presents the time period and 

concentration of three gaseous airborne chemicals (SO NO , 03) that have been 

reported to induce visible injury to vegetation in N&th her~ca. Most of the 

data have been obtained from chamber, greenhouse, or field experiments, not 

from ambient exposures. The table may be used as a guideline to determine 

whether concentrations reported in the ambient.air are likely to cause visible 

injury to vegetation or, more specifically, trees. Few exposure statistics are 

available for nitric acid, nitrate, ammonia, the ammonium ion, sulfate, hydrogen peroxide, or PPN (peroxy propionylnitrate) since these substances have not 

been widely identified as causing visible fol iar injury. PAN (peroxyacetylnitr-ate), a regional pollutant in southern California, has not been widely 

reported in eastern North America. Though PAN is more toxic to crops than 

ozone, studies indicate trees are not as sensitive (U.S. Environmental Protection Agency 1986). Ambient air concentrations of nitrogen oxides (NO and 

NO ) high enough to cause vegetational injury in the field have not been 

re6orted in eastern North America (National Research Counci 1 1977b; U. S. 

Environmental Protection Agency 1982b). 

Research has not yet clearly defined which components of pollutant exposure (pollutant concentration, time of exposure, time since previous exposure, 

etc.) are most important in causing plant responses. The characterization and 

representation of exposures to air pollutants has been and continues to be a 

major problem. An appropriate summary statistic for one exposure duration 

usually cannot be easily transformed to describe a different duration of 

exposure without access to the original aerometric data. In addition, statistics used to represent extremely short exposures cannot be readily aggregated 

to provide a representative summary statistic for plant responses resulting 

from an extended exposure, for exampl e, a growi ng season (U. S. Envi ronmental 

Protection Agency 1986). In addition, the exposure statistics currently 

available were developed chiefly for crop plants, not for forest trees. The 

absence of reliable exposure statistics makes it extremely difficult to develop 

a cause-and-effect relationship between air pollutants and forest trees. 

To document a cause-and-effect relationship it must be possible to infer 

a strong pattern of consistency, responsiveness and a proven biological 

mechanism (Woodman and Cowling, 1987). In simple systems, any two of the 

above three linkage patterns may be sufficient to infer cause. In more 

compl icated systems, however, a1 1 three patterns may be required. Consi stency 

requires that the symptoms of injury and disfunction be consistently associated 

with the presence of the causal factor; the devel opkent of 
Table 9. Critical Concentrations of Gaseous Pol lutants That Induce Visible Fol iar Injury to Vegetation in Eastern North America* 

Time of SO2 NO2 

Exposure Sensitive lntermedi ate Tolerant Sensi Live Intermediate Tolerant 

'U. S. Environmental Protection Agency. Air Q u a l i t y C r i t e r i a for Sulfur Oxides. Revised Chapter 5: Effects of Sulfur Oxides in the 

Atmosphere an Vegetation. EPA-R3-73-030. National Envi ronmental Research Center, U. S. Envi ronmental Protection Agency, Research 

Triangle Park, NC, 1973. 

2U. S. Environmental Protection Agency. Air Qua1 i ty Criteria for Nitrogen Oxides. EPA-600/8-82-026. U. S. Envi ronmental Protection 

Agency, Research Triangle Park, NC, September, 1982b. 

3U. S. Environmental Protection Agency. Air Qua1 i ty Criteria for Ozone and Other Photochemical Oxidants. EPA 600/8-84/0200F, 

Environmental Criteria Assessment Office, Research Triangle Park, NC, August, 1986. Draft Final. 

*The majority of the responses cited in the table are based on laboratory experiments. 
exposure-response relationships by exposing healthy trees to various known 

concentrations of airborne pol 1 utants under conditions simulating those found 

in forests and mechanisms requires that the observed effect be related to the 

suspected causal factor by one or more biological mechanisms. Responsiveness 

is the most difficult to prove because exposing large forest trees using 

controlled conditions is extremely difficult. Extrapolation from exposure 

using seedlings is not satisfactory either (Pye, 1988). 

Sulfur Dioxide.--The adverse effects of SO on vegetation have been 

observed in cities and around smelters for more ?han 100 years. The areas in 

the vicinity of smelters such as those around Sudbury, Ontario, Canada, and 

Copper Hill, TN, are denuded of vegetation due in large part to SO emissions. 

Research concerning 50 effects on vegetation, for reasons of hfstorical 

precedent, has been e&ensively and recently reviewed (U. S. Environmental 

Protection Agency 1982a; Winner and others 1985;). 

Gaseous sulfur dioxide has been rigorously proven to cause visible fol iar 

injury, decreased growth, and mortality of many different species of forest 

trees in concentrations greater than approximately 50 ppb for 8 hours or 

longer (see table 9). This knowledge has been accumulated through many years 

of field studies near strong point sources as well as field and greenhouse 

experiments (U.S. Environmental Protection Agency 1982a). The majority of the 

studies dealing with the effects of SO exposure on growth and yield have used 

crop plants (i. e. , annuals). Crop pla%ts have been emphasized because of the 

possibly greater economic impact if they are injured. 

In the eastern United States, the highest sulfur dioxide concentrations 

have been measured in the Ohio River Valley (see chapter 3, fig. 15). In the 

southeastern States, exposure to high concentrations is most likely to occur in 

the vicinity of power plants, metal smelters, and near large cities. 

Sulfur is an essential nutrient for plant growth and development. Forest 

ecosystems require about 4-22 kg/ha/year for normal functions and growth -- 

mainly synthesis of enzymes, structural and storage proteins, and amino acids. 

There is a close biochemical relationship between sulfur and nitrogen due to 

their mutual involvement in amino acid and protein synthesis (Kelly and 

Lambert 1972). Uptake of sulfur by forest trees, mainly in the form of water 

soluble sulfate ions, is by absorption through the roots, and through foliar 

organs from air, cloud water, and precipitation (Turner and Lambert 1980). 

Thus, low concentrations of gaseous sulfur dioxide (tapproximately 40 ppb) , and 

small amounts of airborne sulfate aerosol (up to about 3 kg/ha/year) may 

enhance growth and uptake of all elements (Turner and Lambert 1980). Atmospheric sulfur additions in pol 1 uted areas (10-80 kg/ha/year), however, 

frequently exceed not only the requirement of the forest ecosystem, but also 

its ability to accumulate sulfur biologically (Johnson 1984). 

Nitrogen Oxides.--Injury to vegetation by NO (NO or NO ) compounds, if it 

does occur in nature, is presumed to be restrictei to the imiediate vicinity of 

industrial point sources -(e. g. , power plants, arsenals, or nitric acid factories) or to large metropolitan areas. Ambient concentrations of NO seldom 

reach the magnitude (around 2000 ppb) necessary to cause vegetati on31 injury 

(see table 9). Most vegetational injury associated with NO is indirect -- 

that is, it is the result of photochemical oxidant reactio%s that produce 
ozone, PAN, and PPN (Kozl ows ki and Constanti ni dou 1986a; National Research 

Council 19776). Visible injury to forest trees or other forest plants from 

exposure to concentrations of nitrogen oxides occurs so rarely in eastern 

North America that there have been no published reports of injury in the field 

due to NO or NO (National Research Counci 1 1977b; U. S. Environmental Protection Agency 198 

Nitrogen, like sulfur, is an essential nutrient for plant growth and 

development. Uptake, mostly as water-sol uble nitrate and ammoni um ions, is 

chiefly through the soil. Each year, forest ecosystems in the Southern United 

States require about 30-75 kg/ha/year of nitrogen for their normal functions 

and growth -- mainly for synthesis of enzymes, structural proteins, nucleic 

acids, and many secondary metabolites. These ions, obtained chiefly from soils 

by absorption through the roots, also can be assimilated through foliar organs 

from air, cloud water, and precipitation. Approximately a third to a half of 

the current need for nitrogen (20-60 kg/ha/year) is met by recycl ing nitrogen 

through litter-fall. Most of the other half (28-85 kg/ha/year) comes from 

atmospheric sources and by mineralization of soil nitrogen (Binkley and others 

1989). 

Concentrations of ai rborne ni trogen-contai ni ng chemical s are bei ng conti nuously added to forest ecosystems today. The effect of these chemicals on 

ecosytems has not been clearly determined. Plants and the communities in which 

they 1 ived were adapted to soi 1s with low nitrogen concentrations and were 

thriving before anthropogenic nitrogen deposition began. In many areas of 

North America, current additions of atmospheric nitrogen are in excess of growth 

requirements, particularly of subalpine and boreal coniferous forests. Only a 

small fraction of the inorganic nitrogen added to the soil from new sources is 

absorbed and enhances growth. Most of the nitrogen added is leached from the 

soil in the form of nitrate (Waring and Schlesinger 1985). 

Ammonia (NH ) and Ammonium (NH +). Ammonia has been known to be a 

phytotoxic air pallutant since the l%te 18001s, chiefly because of injury to 

vegetation in the vicinity of accidental releases of gaseous or 1 iquefied 

ammonia (National Research Council 1977a; Treshow 1970). Vegetational injury 

was most commonly associated with release of ammonia from refrigeration 

systems; however, freons are now being used as heat transfer fluids, so injury 

from this source has declined. 

Agricultural use of anhydrous ammonia and urea as fertilizers is a 

present-day source of injury to vegetation. In addition, ammonia is added to 

the atmosphere from a variety of combustion sources, including domestic incineration and automobile engines, as well as volatilization from cattle feedlots 

and from natural sources (National Research Council 19786). 

Nitrates.--Nitrates are present in the atmosphere as gases (PAN and nitric 

acid vapor), dissolved in droplets, and also in the form of particulate matter 

such as ammonium nitrate (NH NO3). Nitrates are secondary pollutants formed in 

the atmosphere from NO and a0 and are an important component of acid deposition (Lovett and Lindberg 19863. 

Ozone. --Ozone is the gaseous pol 1 utant most injurious to forests. This 

conclusion is based upon the massive number of studies involving the responses 
of crops to ozone and the somewhat smaller number of studies dealing with tree 

responses to ozone. All fumigation studies, whether in the laboratory or 

field, whether on crops or trees, suggest that plants will respond to ozone 

concentrations above 0.06 ppm within hours (see table 9). Concern stems from 

the fact that ozone concentrations across most of North America exceed 60 ppb 

during most of the growing season. Much of our knowledge about plant responses 

to ozone is summarized in recent volumes (Guderian 1985; U. S. Environmental 

Protection Agency 1986). 

Mixtures of Gaseous Pol lutants. --In the vicinity of major pol 1 utant 

sources such as cities and power plants, the air contains mixtures of gases, 

aerosols, and coarse particulate matter. Near power plants SO usually is the 

predominant gas, but NO and O3 can also be present. The gase$ may be present 

simul taneously , sequent?al ly , or intermittently ; however, the exposure of 

forests to more than one of the pollutants simultaneously at concentrations of 

30 ppb or higher is unlikely to occur with any appreciable frequency in rural 

areas farther than 50 km from a point source. The frequency of simultaneous 

occurrences decreased as the concentration of the gas increased for the months 

of May through September during the 5-year period, 1978-1982, of the study by 

Lefohn and his co-workers (1987). Mixtures of SO and NO are experienced most 

frequently near point sources; mixtures of SO a6d NO a?e next, and mixtures 

of O3 and NO2 occur least frequently (Lefohn a6d other$ 1987). 

Concern regarding pollutant combinations exists because controlled exposure experiments with certain mixtures indicate plants respond to exposures of 

pollutant mixtures differently from the way they respond to the single pollutant (e.g., mixtures of O3 and SO together can cause injury at concentrations 

lower than can either gas by itsglf). Unfortunately, however, very few of 

these experiments have included woody plants. 

The effects of simultaneous co-occurrences of pollutants (0 SO , NO ) on 

forest trees may not be important if co-occurrences are as rar8'as ?epor5ed 

by Lefohn and Tingey (1984) and Lefohn and others (1987); however, the sequence 

to which vegetation is exposed to multiple pollutants may be important in 

causing growth effects. 

Mixtures of gases can cause effects that are 1) significantly greater 

than the additive effects of the gases singly (synergism); 2) significantly 

less than the additive effects of the gases (antagonism); or 3) not significantly different from the additive effects of the gases acting alone (additive) 

(Heagle and Johnston 1979). 

The 1 i kel ihood that acid deposition and other airborne sulfur- and 

nitrogen- derived chemical s wi 11 affect the forests of eastern North America 

depends on the extent that the chemicals enter the trees. 

The alteration of physiological processes (photosynthesis, respiration, 

carbon uptake and allocation), foliar leaching, disruption of regeneration and 

reproduction, and changes in host susceptibility to pests, pathogens, and 

mycorrhizal colonization resulting from the exposure of trees to airborne 

chemical pollutants are discussed in the sections that follow. 
4.2.1 By What Physical, Chemical, and Biological Processes Are These Compounds 

Taken up by Plants, and Microorganisms? 

The life-sustaining processes of photosynthesis, respiration, and transpiration involve the exchange of gases between plants and the atmosphere. 

Exchange of gases during these processes occurs primarily through the stomata 

on the leaves because the epidermis is coated with a waxy cuticle that is 

relatively impervious to gases. Exposure to airborne chemical pollutants 

occurs during the above natural processes. 

Uptake of airborne chemical pol 1 utants involves: 1) exposure, i . e. , the 

presence of the chemical pollutant in the vicinity of the forest; 2) deposition 

into the forest; and 3) movement into the trees through the leaves or from the 

soi 1 into the roots. 

The processes of deposition are the crucial link between transport of 

atmospheric pollutants to forest canopies, their ultimate uptake and the manifestation of effects on vegetation (Lindberg and McLaughlin 1986). The major 

problems of how to quantify pollutant dose so that both chemical species and 

exposure kinetics are identified, and how to characterize plant response so 

that actual or potential economic or ecologic losses can be quantified, are 

related to gas exchange at the atmosphere-leaf interface (Lindberg and 

McLaughlin 1986). The gas-exchange processes for the common gaseous species 

such as 0 SO2, HNO and NO2 have been parameterized in dry deposition models 

(Bal docch?'and otherZ91987). 

Differences in the rates of pol lutant deposition could, in part, account 

for the variation in plant responses that are usually considered a function of 

genotype, environment, pollutant exposure dynamics (movement of pollutants 

into the leaves), or the presence of other pollutants. Because the physiological sites of action for most gaseous pollutants are cells in the leaf 

interior (table lo), any factor that influences pol 1 utant deposition (e. g. , rate, concentration, form) will change the relationship between the concentrations of the pollutant in the atmosphere and the corresponding concentration at 

the sensitive sites within leaf cells. An understanding of the role of 

deposition processes should, therefore, aid in explaining patterns in pollutant 

effects on the physiological status, growth, development, and productivity of 

plants (Taylor and others 1988). 

Table 10. Types of Interactions of Gases With Plant Surfaces 

Adsorb to or react with all surfaces 

Interact primarily inside leaf tissue 

Exchange slowly with plants; 

potentially important interactions 

H202, HF, HC1, HN03 

S02, NH3, Cl2, NO2, H2C0, 

03, PAN, H20*, metabol ical ly reactive C02*, 02* 

N20, NO, CO, organic 

molecules such as 

hydrocarbons, pheromones, 

ethylene, and terpenes 

"Normal large-scal e exchange in metabol ic processes. 

Source: Hosker and Li ndberg (1982). 

86 
A variety of physical, chemical and biological factors control deposition 

rates (Bennett and Hill 1975). These factors include the micrometeorological 

conditions above and within the plant canopy, leaf surface characteristics 

(shape, degree of pubescence, and surface chemistry) and the chemical and 

biological properties of individual pollutants, such as diffusivity, water 

solubility, and biological reactivity within cell systems (Lindberg and 

McLaughlin 1986). 

Pollutant deposition can be addressed from different levels of organization: 1) at the plant canopy level, focusing on atmospheric processes 

governi ng turbulence in terrestrial 1 andscapes and the di stri buti on of pol 1 utant deposition sites within plant canopies; 2) at the level of the individual 

leaf, focusing on foliar sites of deposition and the role of cuticular, 

stomata1 , and mesophyl 1 resistances in governing trace gas rates of exchange 

(Bennett and others 1973); and 3) at the molecular level, emphasizing the 

physiochemical processes at the intercellular space and mesophyll cell surface 

interface in the leaf interior (Tingey and Taylor 1982); 

Dry deposition of gaseous pol 1 utants to plant canopies and individual 

leaves occurs because a chemical gradient exists between the atmosphere and the 

deposition sites, either on the exterior leaf surfaces or on cells of the leaf 

interior. Dry deposition involves both gases and particles. Deposition of 

particles and cloud water to canopies and individual leaves, however, is 

governed primarily by processes other than diffusion. Vertical turbulence 

transports particles across the atmospheric boundary 1 ayer to the laminar 

boundary layer of individual leaves. Atmospheric turbulence, which is governed 

by mechanical mixing, buoyancy, and surface roughness, increases deposition. 

Particle cloud/water deposition is governed by gravitational settling, droplet 

impaction, interception efficiencies, and Brownian diffusion (Taylor and others 

1988). 

During wet deposition gases and particles are delivered to the canopy 

partly in solution, thus enhancing the possi bi 1 ity of sorption by fol iage. 

Other forms of wet deposition such as dew, mist, and/or fog interception can 

create conditions on the leaf surface that lead to extremely high concentrations of pollutants because of interactions with deposited particles and gases 

(Hosker and Lindberg 1982). Wet deposition is episodic and delivers the 

pollutants in highly irregular, intense concentrations. In addition to 

depositing pol 1 utants on plant surfaces, wetfall may simultaneously remove 

previously deposited substances, thus redistributing pollutants. Therefore, 

the net effects of precipitation may be either to increase or decrease the 

surface concentration of a particular pollutant (Lindberg and others 1979). 

Surface exchange processes occurring between the atmosphere and plant 

canopies present special problems in mass transport of air pollutants. Fluxes 

of air pollutants and materials such as CO and H 0 within the canopy vary 

because the vertical distribution of source$ and skks also varies with canopy 

height, the species composition, and with the physiological state of plants. 

Gases react with the various types of surfaces found within the vegetative 

canopies, especially leaves. Leaf surfaces of different species vary in degree 

of pubescence, number of trichomes, and in the presence of wax and exudates. 

The effectiveness of these competing sinks is not equivalent. The interaction 
of gases with these leaf components may alter the net effect of any individual 

gas on vegetation (Heagle and Johnston, 1979; Hosker and Lindberg 1982; Taylor 

and others 1988). 

When gases or particles approach plant leaves, they must eventually pass 

through the inner portion of the leaf boundary layer. Transfer within this 

area is no longer governed by turbulence effects. The leaf-air microenvironment within this sublayer may differ substantially from that outside the 

boundary layer. Gases moving into the leaf encounter concentrations of transpired water vapor and a variety of hydrocarbons such as terpenoids that are 

produced in secondary plant metabolism, volatilized into the gas phase of the 

leaf interior, and emitted by the leaf through the stomata. These substances 

are higher within the inner boundary layer and may significantly affect the 

surface uptake of gases and particles (Hosker and Lindberg 1982). 

Gas exchange at the atmosphere-leaf interface has certain unique aspects. 

These are: (1) a pronounced and viscous/stagnant layer in the gas phase 

adjacent to the leaf surface (boundary layer) and in the leaf interior 

(sub-stomatal chamber and intercellular space); (2) a variable diffusive 

resistance in the gas phase due to changes in stomatal porosity; (3) a 

distributed array of extensive deposition sites both on the leaf surface and 

within the leaf interior; (4) a biochemically variable capacity for trace gas 

deposition and assimilation in the mesophyll tissues; and (5) at the level of 

the plant canopy, a mix of functionally varying leaf surfaces comprised of 

different age classes and species, each having its own distinct local microclimate such that the canopy's effective deposition/sink potential varies in 

space and time. 

Stomata1 control of diffusion into the leaf is a major factor influencing 

the exchange of gaseous pollutants between the atmosphere and the internal 

tissues of leaves. All studies indicate that stomatal resistance (the extent 

to which stomata are open or closed) is the most dynamic and most influential 

resistance to 0 ?' so?' 

and NO2 transfer when the canopy is dry (Baldocchi and 

others 1987). unct oning of stomata in turn is controlled by both physiological and bioenvironmental processes (Hosker and Lindberg 1982; Wesely and 

Hicks 1977; Winner and others 1985). The rate of diffusion through the stomata 

depends on (1) the diffusivity of the gases, (2) the magnitude of the concentration gradient for a specific gas, and (3) the degree to which stomata are 

open (Winner and others 1985). 

The transfer of gases from the atmosphere to the leaf interior involves 

both a gas phase and a liquid phase. The leaf interior is a highly porous and 

physiological ly complex organ. Given the leaf' s morphology, the ease of 

diffusion and site deposition for gaseous pollutants are likely to vary appreciably among gases in accordance with their physical and chemical properties 

(Taylor and others 1988). 

The gas-phase flux depends upon the concentration gradient between ambient 

air and the wet cell surfaces within the leaf and the resistance to mass 

transfer along the diffusion pathway. Within the leaf, following deposition on 

wet cell surfaces, flux is governed by liquid-phase processes including diffusion, formation of pollutant derivatives, and productivity of plants (Tingey and 

Tay 1 or 1982). 
Uptake of some dry deposited gaseous pol 1 utants (sulfur dioxide, nitrogen 

oxides, ozone, peroxyacyl nitrates, nitric acid, ammonia, and hydrogen peroxide) occurs by absorption and adsorption of the gas on the moist walls of 

mesophyll cel 1s. After diffusing into the intercellular space of the leaf 

interior, gas molecules are partitioned across this gas-to-liquid interface on 

cell surfaces at a rate determined by the solubility of the gas in the external 

cell solution and the chemical reactivity of the gas in aqueous phase (Taylor 

and others 1983). Uptake of NO, particularly, is controlled by its extremely 

low solubility (Malhoutra and Kahn 1984). Tingey and Taylor (1982) have 

proposed that deposition sites for highly reactive water-soluble gases such as 

HMO, and SO, are the leaf surface and the substomatal cavity, whereas, the site 

forJ less soyuble gases (e. g. , 03) is the mesophyll tissue (Taylor and others 

1988). 

Uptake of sulfate, nitrate, ammonium, and hydrogen ions and nutrient 

cations and anions in fine and coarse particles, in coarse particulate matter, 

and in precipitation and cloud water after impaction or gravitational settling 

on vegetation surfaces, occurs more slowly than uptake of gases. Uptake of 

these ions after dissolution occurs mainly by diffusion through the cuticle on 

leaves or through the bark, and by diffusion through fine feeder roots and 

mycorrhizae after percolation and diffusion through the soil. 

Absorption rates depend on pollutant concentration, cuticle thickness and 

hydration, and the affinity of cuticular compounds for the solutes involved. 

Sol ubl-e compounds may enter leaves through cuticular breaks, trichomes, wounds, 

or stomata. Leaf exudates also may play a role in the entrllnce of deposited 

substances (Hosker and Lindberg 1982). 

Lichens and mosses, important nonvascular plants growing in forests, have 

low tolerances to air pollutants (Winner and Bewley 1978a, 1978b). Their 

sensitivity is greater, in part, because these plants accumulate ai rborne 

compounds at a faster rate than vascular plants do (Winner and others 1978). 

Lichens are composed of a fungus and an alga. The thallus (plant body) of 

lichens, unlike the leaves of vascular plants, have no waxy cuticle. Bare 

patches, holes, pores, and depressions on the surfaces of the lichen permit the 

interior of the thallus to be exposed to ambient air all times of the day or 

night. For this reason, airborne chemicals can diffuse rapidly into the 

thallus where they can accumulate and cause injury. The algal cells of lichens 

are generally more sensitive to injury by sulfur dioxide and ozone than the 

fungal cells (Anderson and Treshow 1984). 

Mosses have low tolerances to air pollutants. Their leaves are usually a 

single cell thick and have no cuticle, therefore, readily absorb airborne 

chemicals. Winner and others (19781, by using sulfur isotopes, showed that 

mosses absorb airborne sulfur. Because they can accumulate large quantities of 

sulfur rapidly, they are more vulnerable to effluents such as SO2 than vascular 

plants (Winner and others 1978). 

Exposure of plants to airborne gases fine or coarse aerosols, or coarse 

particulate matter through the soi 1 , is indirect. Four separate processes 

are involved: 
* Deposition; 

. Dissolution into the soil water; 

* Percolation or diffusion through the soil; and finally, 

Absorption through the surfaces of fine feeder roots or 

mycorrhi zae. 

Many biotic, physical, and chemical factors influence these processes. The 

emphasis of this section is on foliage-mediated effects; therefore, the soilmediated exposures are not discussed at this time. 

Microorganisms readily absorb sulfur- and nitrogen-containing compounds 

from soil in forest ecosystems. Microorganisms can immobilize nutrients and 

reduce the quantity in the soil available to plants via nutrient assimilation 

(A1 exander 1977). 

In summary, the physical, chemical, and biological processes involved in 

plant uptake of sulfur- and nitrogen-derived compounds from both air and soil 

are influenced by a variety of biotic and abiotic factors. The biotic factors 

i ncl lade 

* stage of plant development; 

genetic variation in plant species; 

competition among several species and individuals of the plant 

community for light, water, growing space; and 

interaction among the pol 1 utant, the plant, plant pathogens 

and pests, and root mycorrhizal fungi. 

The abiotic factors include 

photoperiod; 

* lightintensity; 

* temperature; 

relative humidity; 

amount of rainfall and its timing; 

b soi 1 moisture; 

soil fertility; 

interactions among co-occurring pollutant chemicals; and 

dewfall (Heck 1982). 
4.2.2 What is the Mode of Action of the Different Airborne Chemical 

Pol 1 utants? 

Gaseous pollutants must enter plants to cause an effect. Alteration of 

physiological processes begins at the biochemical level within the cells of 

leaves. Entrance is mainly through the stomata in the leaves, though some 

entrance may occur through the cuticle, especially if it is weathered or damaged. 

Pollutant uptake is controlled by the stomata and the environmental conditions 

that determine stomatal behavior. Any action of a pollutant on stomatal 

behavior affects gas exchange and water loss from plants (Black 1985). Leaf 

(i . e. , stomatal) conductance regulates the movement of the gas from the leaf 

boundary layer into the leaves (Black 1985; Tingey and Taylor 1982; Winner and 

Mooney 1980a, 1980b). The mode of action of sulfur- and nitrogen-containing 

gaseous pollutants, of ozone, and of mixtures of these pollutants is discussed 

in the sections that follow. 

Sulfur dioxide, nitrogen oxides, ozone and PAN (and possibly hydrogen 

peroxide, H 0 ), the major pollutants discussed in this chapter, have a number 

of features2i?i common: 

1) All five pollutants are gases that enter plants primarily 

through stomata; 

2) At least the first three alter the normal regulatory function of 

guard cells and thus can increase or (usually) decrease stomatal 

conductance; 

3) All five are highly reactive oxidizing agents, and at least the 

first three attack the membranes and organelles of cells and 

thus alter general metabolic processes; 

4) The first four are known to inhibit photosynthesis, increase 

respiration, and a1 ter carbon a1 1 ocation between shoots and 

roots. 

a) Sulfur Compounds 

Sulfur Dioxide. --Uptake of SO2 through stomata occurs mainly in the summer. 

Stomata1 activity in winter is low, although uptake in even subzero temperatures 

may occur in some evergreen trees. Deciduous trees a1 so can absorb SO2 in 

winter if their lenticels are active (Garsed 1985). 

There is no simple pattern of stomatal response to SO . Exposure to SO2 

may induce either increased stomatal opening or closure it? the same species. 

Low concentrations in some studies increased stomatal opening, while high 

concentrations induced closure (Black 1985). Responses to SO are so diverse 

and depend on so many factors, it is extremely difficult to Jssess the longterm imp1 ications of exposure. Within a forest canopy, stomatal responses vary 

spatially and with time in concert with variation in pollutant concentrations, 

environmental conditions, and the biological state of each particular plant 

part (Bl ac k 1985). 
Regardless of the ambient air concentration of SO to which trees are 

exposed, tree response is determined by the amount of 50 that enters the 

leaves of the trees. After diffusing through the s$_omata, %O dissolve_s in the 

water on the cell ul ar surfaces to form sul Pi te (SO ), bisulfite (HSO ) and, 

depending on the pH of the surrounding medium, 0th r ionic species (~ilhoutra 

and Kahn 1984). Both sulfite and bisulfite are toxic to many biochemical and 

physiological processes (Malhoutra and Hocking 1976; Ziegler 1975). 

Plants can detoxify sulfite and bisulfite by further oxidizing them to 

sulfate. Toxicity is determined by the site at which conversion occurs. 

Toxicity may occur if, during short episodes with high SO concentrations, the 

rate of SO conversion to sulfate is exceeded (Ziegler k75). At low SO 

concentratigns, plants can accumulate sulfur compounds; however, chroniz 

symptoms may appear if the capacity of a plant to metabolically incorporate 

sulfate is exceeded. Detrimental toxic effects are usually associated with 

concentrations, while growth stimulation or no effects may be 

d with exposure to 1 ow concentrations (U. S. Environmental Protection 

Agency 1982a). 

Sulfate Aerosols.--Sulfate aerosols are secondary pollutants fyyed in the 

atmosphere through the oxidation of SO, to sulfates. Sulfates (SO, ) are an 

L important component of acid deposition, but studies dealing with the4e-ffects of 

sulfates are 1 imited. The majority of the studies reporting the effects on 

vegetation of sul fate-containing acid deposition emphasize the hydrogen ion, 

the associated changes in pH, and the effects of the gaseous components 

(Altshuller and Linthurst 1984). 

Aerosols must be soluble to enter leaves through stomata, trichomes, or 

cuticular breaks and wounds before injury can occur at the cellular level 

within a plant (Hosker and Lindberg 1982). Chevone and others (1986) suggest 

that sulfuric acid aerosols can enter leaves through the stomata; howevy, no 

visible injury was noted in soybeans exposed to acid aerosols (500 pg/m for 

4 hours), a concentration much qreater than is usually observed in the ambient 

air. ~bi iar injury was observed; however, when pinto -beans (Phaseol us 

L. ) were exposed for a 3-week period to submicrometer ammonium sulfate aerosols 

using an aerosol concentration two orders of magnitude above the ambient (Gmur 

and others 1983). Leaf injury was probably due to the ammonium fraction of the 

aerosol as sulfate in leaves has not been shown to be injurious (Nihlgard 

3.985). No field studies dealing with the uptake of sulfate have been found in 

the literature (Krupa and Legge 1986). 

b) Ni trogen Compounds 

.--Gaseous nitrogen compounds, principally NO and NO must 

diffuse through stomata or cuticle to cause a plant effect. ~ertu&ation 

within mesophyll cells in the leaf results after gaseous NO or NO from the 

ambient air enters the stomata, dissolves in the extracel lular wa?er on the 

moist cell surfaces, and reacts to form nitrous oxide which then dissociates to 

nitrite (NO ) The solubility of NO and NO in the extracel lular water is a 

major influgnce in determining the rate of up?ake. While NO is poorly soluble, 

NO2 is highly soluble (Mansfield and Freer-Smith 1981). The solubility of NO 

in xylem sap is greater than in distilled water. This fact may play an 
important role in its uptake because xylem sap is continuous with the extracet l ul ar water i n a leaf (Anderson and Mansf iel d 1979). Both stomata1 and 

mesophyll resistance control uptake (Srivastava and others 1975a). 

Plants have the ability to metabolize th~ dissolved NO via the following 

general metabol i c pathway: NOx+NO +NO +NH4 +ami no aci ds+brotei ns. Nitrite 

is usual 1y considered to be a to c io%, though not all scientists agree. 

Available evidence suggests that fects of NO are a consequence of increased 

nitrite and nitrate concentrations within plan't cel ls, particularly nitrite. 

Normal nitrogen metabolism follows the pathway indicated above. If amino acids 

are not formed, toxic compounds wi 11 accumulate. Injury results when nitrite 

reduction in leaves is inhibited (Mansfield and Freer-Smi th 1981). Plants 

that have been exposed to NO2 exhibit increased concentrations of nitrate 

reductase (NiR) in the leaves. Plants exposed to both NO and SO2 failed to 

show an increase in nitrate reductase in the leaves (Mansf eld and Freer-Smith 

1981). Most of the effects of NOx on nitrate metabolism can be explained by 

assuming pol lutant-induced changes in the rates of reduction of NO3 and NO2 

ions by plants (Rowland and others 1985). 

.--The concentrations required to cause 

een well quantified. Foliaqe is the most 

susceptible plant part (~ational Research Council 1977a). ~mmonia dissolves in 

the waterfilm on the walls of the mesophyll cells after entrance through the 

stomata. This view is supported by experiments in which pinto bean plants were 

exposed for 12 to 21 days to submicrometer ammonium sulfate aerosols at concegtrations two orders of magnitude above the ambient concentration of 60-100 pg . Vi s i bl e i n jury was observed a1 ong the 1 eaf margi ns ; however, i ni ti a1 i n jury5 was 

to the spongy mesophyll cells (Gmur # others 1983). Experiments using NH3 indicate that once inside the cell, N is incorporated into amino acids and 

proteins and ultimately is transported to the roots (van Hove and others 1987). 

Nitrates. --Dry deposition of nitrates into forests can be an important 

source of inorganic nitrogen. Lovett and Lindberg (1986) reported that dry 

deposition of atmospheric nitrates supplied nearly 60 percent of the inorganic 

nitrogen deposited in the Walker Branch Watershed in eastern Tennessee. Canopy 

net uptake was estimated to have been 3.2 kg NO -N/ha/year. Canopy uptake of 

deposited nitrate is considered to be a general henomenon in forest ecosystems, 

Taylor and others (1988) point out that the deposition site for HN03 vapor is 

the leaf surface and the substomatal cavity. Nitrates, as well as sulfates 

deposited on the soil, can enhance growth. The fate of nitrate aerosols after 

deposition on plant foliage is an area needing study. 

Peroxyacetyl Nitrates. PAN is the most common member of a series of 

homologues that increase in phytotoxicity with increase in molecular weight. 

Only PAN is found in ambient air at concentrations that are injurious to 

vegetation, and then only in limited areas of the country (U. S. Environmental 

Protection Agency 1986). Because concentrations injurious to vegetation have 

not been reported from the eastern United States, and because woody plants are 

not very susceptible to injury by PAN, its mode of action will not be discussed 

here. 

Only ozone that enters the plant through the leaf stomata can impair plant 

processes. An effect will cxur only if sufficient ozone reaches the sensitive 

sites within a leaf. The e;~'ects range from subtle modifications of cellular 

biochemistry and whole plant ~hysiology to visible injury. 
c) Ozone and Hydrogen Peroxide 

Ozone. --Ozone enters the leaf through stomata; once within the leaf it 

quickly dissolves in the aqueous layer on the cells lining the air spaces. 

Ozone, or its decomposition products, then diffuses through the cell wall and 

membrane into the cell, where it may affect c e l l u l ar metabol ism. At any point 

along this pathway, ozone or its decomposition products may react with cellular 

components. Altered cell structure and function may result in changes in 

membrane permeabi 1 i ty , carbon di oxi de fixation, and many secondary metabol i c 

processes (Tingey and Taylor 1982). 

Ozone injury will not be detected if (1) the rate of ozone uptake is 

sufficiently small so that the plant is able to detoxify or metabolize ozone or 

its derivalives; or (2) the plant is able to repair or compensate for the ozone 

impacts (Tingey and Taylor 1982). 

The uptake and movement of ozone to sensitive cellular sites are subject 

to various physiological and biochemical controls. The magnitude of ozoneinduced effects will depend upon the physical environment of the plant, including macro- and microclimatic factors; the chemical environment of the plant, 

including other gaseous air pollutants; and biological factors, including 

genetic potential, developmental age of the plant, and interaction with plant 

diseases and insects, Cellular injury may subsequently manifest itself in a 

number of ways, including visible foliar injury; premature senescence; reduced 

growth or yield, or both; reduced plant vigor; and sometimes death. 

Some scientists consider membranes to be the primary site of action of 

ozone (Heath 1980; Tingey and Taylor 1982). Alteration in plasma membrane 

function is an earb event in the sequence of ozone-induced effects that 

eventually leads to leaf injury and subsequent growth reduction. Changes in 

the semipermeabi 1 ily of the membrane are evidenced by changes in fluxes of 

carbohydrates, ami no acids , i norgani c ions , and water (Heath 1975, 1980 ; 

Tingey and Taylor 1982). These functional changes lead to cell ular dysfunction and death similar to that induced by water stress and pathogenesis (Heath 

1975; Thomson 1975; Tingey and others 1976b). 

.--Study of plant response to H O2 is in the early 

(1986) reported injury to ~orw$~ spruce ( Picea 

exposures to an acidic mist containing H 

reported a significant decrease of cell area in the mesophyll in bot 

and secondary needles and also in the vascular bundles; however, in some cases 

changes were found only in the primary needles. Movement into the leaf was 

considered to be through the stomata. The authors state that H O2 must be 

considered an important factor in forest dieback in Europe. At $resent this 

view is subject to question, as scientists in the United States have not been 

able to duplicate the work of Mallant and others (1986). 

6) Mixtures of Gaseous Pollutants 

Little is known about how stomata1 regulation and associated biochemical 

and physiological processes are influenced by mixtures of pollutants (Malhoutra 

and Kahn 1984). The effects of pollutant mixtures are complex and differ with 

species, cultivar, and experimental conditions (Ormrod 1982). The effects of 

mixtures of gases on plants are possibly mediated by the following: 
1) reactions between the pollutant gases after entering the leaf; 2) an effect 

by one of the gases on the stomata1 aperture; 3) competition between the gases 

for the reaction sites; 4) a change at the reaction sites in sensitivity; or 

5) a combination of these effects (Heagle and Johnston 1979). 

Mixtures of Gases and Aerosols.--Fine particulate aerosols and ozone are 

formed concurrently in the atmosphere. The interaction of particulate sulfate 

aerosols and ozone is a possible reason for concern. Fine particulate aerosols 

accumulate at night in a reactive plume, while ozone decreases at night and 

then peaks during the daytime hours. Thus, because of the diurnal photochemical patterns of ozone, vegetation is first exposed to peak sulfate aerosols 

concentrations before exposure to ozone (Krupa and Legge 1986). Under these 

circumstances when conditions are acidic, the residence time of ozone is 

si gni f icantly i ncreased (Chevone and others 1986; Krupa and Legge 1986). 

During periods of wet deposition, ambient ozone concentrations are general ly below- detection 1 imi ts; however, the acidification of receptor surfaces 

cod d predi spose fol i ar surfaces to - subsequent ozone exposure. Laboratory 

studies suggest this possibility. Neither soybean (Glycine max L.) nor pinto 

bean (Phase01 us L. ) exvbited visible fol iar- injuryThen exposed to 

sulfuric acid ae 500 pg/m for 4 hours); however, when exposed to ozone 

at 190 ppb for 4 hours, visible injury was observed (Krupa and Legge 1986). 

Chevone and others (1986) reported similar results using similar concentrations 

of ozone and acid aerosols. Additional studies are needed to determine whether 

predisposition occurs when sulfuric acid aerosols at concentrations near 

ambient are used in the exposures. 

e) Acidic Deposition 

Acidic deposition, in wet form, is a dilute solution of hydrogen ions and 

associated anions. his solution, -depending on the surface ;t comes into 

contact with, is easily altered. If the surface is basic, the pH increases; if 

acidic, the pH may decrease (Bennett and others 1985). The soluble-ions usually 

encguntered in precipitation are hydrogen (H ), bigarbonate (HCO 3, calcium 

(Ca 1, magnesium (Mg-), sodium (Na ), potpssium (K ), sulfate (S~L), nitrate 

(NO, ), chloride (C1 ), and ammonium (NH, ). Sulfate and nitrate ions are the 

domfnant ones in precipitation in the as? (Stensland and others 1986). 

Much has been written about the possible effects of acidic deposition on 

forests and its relationship to forest decline. No evidence to date supports 

the concept that there is a direct relationship between acidic deposition and 

forest effects. A possible indirect effect of acidic deposition on forests via 

soil-mediated response mechanisms has been hypothesized (Amthor 1984; Bennett 

and others 1985; National Acid Precipitation Assessment Program 1987;). In a 

separate document, Binkley and others (1989) discuss soil-mediated responses. 

Summary.--Gaseous atmospheric pollutants must enter plants, chiefly 

through the stomata, to cause a response. Pollutant chemicals in the gas phase 

diffuse through the stomata into the substomatal cavity where they enter the 

liquid phase by dissolving in the water film on the walls of the cells lining 

the substomatal cavity. From this point along this pathway, the phenomena that 

influence the behavior of the pollutant chemicals are not well understood. The 

cell membrane has been suggested as the site of ozone-induced perturbation 
(Tingey and Taylor 1982). Other pollutants may also move through the membrane 

into the cell. Perturbation, changes in cell structive or function, may 

result. The physiological and biochemical changes that cause perturbation are 

not well understood for the majority of the pollutants. Sulfur dioxide, 

nitrogen oxides, and ozone have been studied the most. Response of plants to 

nitrogen oxides is based solely on experiments conducted in the laboratory, 

greenhouse, or in the field under simulated ambient conditions. At present no 

data exist indicating that vegetation in the field in the United States has 

been injured by nitrogen oxides alone. 

Acidic deposition has not been shown to have an adverse effect on forests 

via foliage-mediated response mechanisms. 

4.2.3 Is Photosynthesis in Forest Trees Altered by Exposure to Airborne 

Chemical Pol 1 utants? 

Movement of pollutant chemicals through stomata into cells in the leaf 

interior can a1 ter cellular function. Inhibition of photosynthesis is one of 

the physiological changes attributed to the entrance of airborne pollutant 

chemicals into plant leaves. 

Carbon is the basic building block of the large organic molecules necessary 

for life. Plants accumulate, store, and use carbon to build their structure 

and maintain their physiological processes. Plants obtain carbon from the 

atmosphere and during the process of photosynthesis use sunlight to convert 

carbon dioxide (CO ) into carbohydrates. Carbohydrates serve as the raw 

materials for furtzer biochemical synthesis. Photosynthesis takes place in 

chloroplasts that are located in the palisade parenchyma and spongy mesophyll 

cells of leaves (Waring and Schlesinger 1985). 

From the atmosphere, gaseous carbon dioxide diffuses through the stomata 

into the substomata cavity and the intercellular spaces of the leaf. Then, in 

the liquid phase, it diffuses through the cell membrane into the palisade and 

spongy mesophyll cells (Waring and Schlesinger, 1985). Alteration or disruption of the movement of carbon dioxide into tree leaves can affect the process 

of photosynthesis. 

a) Sulfur Compounds 

Sulfur Dioxide. Inhibition of photosynthesis often is regarded as the 

first sign of SO toxicity in plants (Darrall 1986; Hzllgren 1978). Pollutants, such as sd! have a direct effect on the photosynthetic capacity of 

leaves even at ve6 low concentrations. The rate of photosynthesis in a leaf 

at any time i s governed by certain physiological factors, i ncl udi ng stomatal 

conductance; the nutrient content of leaves; and the biochemical integrity of 

organelles, membranes, and enzymes in the leaf mesophyll cells (Winner and 

others 1985). By altering stomatal conductance or by changing the metabolic 

capacity of mesophyll cells, SO2 can affect photosynthesis (Winner and Mooney 

1980a; Ziegler 1975). 

Sulfur dioxide, regardless of whether it induces stomatal opening or 

cl osure , general ly has =i negative i nf l uence on photosynthes i s (Wi nner and 

others 1985). At least initially, decreased photosynthesis is related to 

decreased stomatal conductance. Leaves with the highest conductance, (the 

least stomatal resistance to the entry of atmospheric gases) will receive the 
ws a, m7575 a, CDO n x =corn 

"Jc +u 0 TTJ 

C3" d.0 rc-J.a, 0"00-5 

a23 t: am C= mi% -5lz:V)33 Zj,It,":m=) CtCL 

a, ct.* 

( -'*V,-=r-J-'T;I -5 

@Ir ZfCDpls =Z.rco fD 

rcx QJ QJ 2 Q,a 

m (*a, " n, 

QJ +a, '*O?j 2. 

$3 Z3"3"3 

g~ccL1-5 0 3-m 

m "m 3 "awu V,* $E c m o 0 4." 

=-5&V)- =- 2.3 +f= 

qzZmmwd0 "am 3'" " J 

rcJ ""ti +' rc 
V) 

' rV) 

CU 

L: 

c, 

c 

%' 

0 

c, 

0 

1=: 

P. 

rt: 

' F 

c 

0 

C 

c, 

0 

3 

a 

a, 

L 

3 

TIC> 

rt 

M 

r-i 

Ln 

0 

ln 

Y 

3 

M 

L 

0 

Y2 

a 

\ 

L 

1Z 

dH 

Y 

rd 

0 

A 

U 

rd 

Cm 

V) 

F 

V) 

a, 

IZ 

-4-' 

t 

2' 

0 

c, 

0 

I: 

a 

c 

rIZ 

0 

.C 

c, 

U 

3 

a 

a, 

L 

cV 

a 

rl 

M 

r-i 

Ln 

a 

V) 

3 

a 

CV 

I 

r-i 

0 

f: 

ln 

Q 

a, 

c, 

r 

II: 

3 
permanent. Recovery began immediately with the termination of exposure. The 

time required for complete recovery depended on the magnitude of inhibition 

(Hill and Bennett 1970). 

The concentrations of NO PAN, and PPM and the period of exposure required 

to inhibit photosynthesis di&r with the plant species or cultivar, leaf age, 

humidity, temperature and amount of l ight (Sri vastava and others 1975b; U. S. 

Environmental Protection Agency 1986). Nitric oxide (NO) was reported to 

be approximately half as effective in reducing apparent photosynthesis in 

alfalfa and oats as NO ; however, the decrease in the rate of photosynthesis as 

we1 l as the rate of rgcovery was more rapid (Hill and Bennett 1970). Photosynthesi s i n beans was si gni f i cantly depressed by nitrogen di oxi de concentpations of 1,000 ppb or higher for exposures of half an hour or longer, The 

degree of inhibition was increased by increasing NQ concentration and increasi ng exposure time (Srivastava and others 1975a). he percentage of i nhi bi ti on 

changed with leaf age and was highest for leaves that had the highest rates sf 

photosynthesis. The reduction of photosynthesis to a greater extent than 

transpiration or stomata1 conductance suggests that the inhibitory effect of 

NO2 occurred within the leaf and not at the point of CQ2 entry. 

The effects observed in the studies discussed above occurred at exposures 

that have never been reported (1,000 ppb for 1 hour or longer) in the ambient 

air with the possible exception of the highest concentrations observed around a 

very few strong emission sources. Capron and Mansfield (19761, however, 

observed a reduction in the ~hotosvnthetic rate of tomato (Lvco~ersicon 

exculentum Mi 11. ) plants expos'ed to 350 ppb of NU2 and 250 ppd of i0 or higher 

concentrations over a 20-hour period. The effect of the two gases on combination resulted in an additive inhibition of photosynthesis. ., 

c) Ozone 

The effects of 0 on leaf gas exchange are not as well known as for S8 

This is surprising, s?nce 0 is considered to be the more injurious pol luta 

of the two. However, few sy$tems have been built which permit the introduction 

of O3 into leaf cuvettes so that dynamic leaf responses to 0 can be measured. 

Lack of such systems also prevents studies that define the e&ects of O3 on the 

light-harvesting efficiency of leaves, the activity of RuBP carboxylase, and 

other diagnostic features necessary for assessing mechanisms of 0 injury. 

Consequently most research using 0 involves fumigating plants in oie chamber, 

then turning off ozone or moving &e plants to another site before measuring 

photosynthesis and other gas exchange parameters. 

Ozone has the potential for inhibiting photosynthesis. Species in which 

ozone has been shown to reduce photosynthesis are northern red oak ( 

rubra L.) (Reich and Amundson 1985), loblolly pine (Pinus -- taeda i. ), slash 

pine (P. elliottii Engelm. ex Vasey) (Barnes 1972a), ponderosa pine (P. 

ponderosa Dougl. x Laws) (Coyne and Bingham 1981; Miller and others 19691, 

eastern white pine (P. strobus L. ) (Barnes 1972a; Reich and Amundson, 1985; 

Botkin and others 1971, 1972), black oak (Quercus velutina Lam. ), sugar maple 

(Acer saccharum Marsh.) (Carlson 1979; Reich and Amundson 19851, and one poplar 

hybrid (Populus deltoides x trichocarpa) (Reich and Amundson 1985) (see 

table 12). 

Some of the studies cited above involve ozone fumigation experiments with 

concentrations at or below 120 ppb; these are of particular value because they 

occur in the ambient atmosphere. Barnes (1972a) examined the impact of ozone 

on seedlings of three species of pine at concentrations of 50 or 150 ppb 

continuously for 19 days to 18 weeks. In younger seedlings of eastern white 

pine, which bore only primary needles, ozone had 1 ittle influence on rate of 

photosynthesis. In older seedlings with secondary needles, photosynthesis was 

slightly depressed. With seedlings of slash, eastern white, and loblolly 

pines, exposure at 105-ppb ozone had a relatively consistent depressing influence on photosynthesis of a11 species. At 50 ppb, however, ozone appeared to 

stimulate photosynthesis in older secondary needles and to depress photosynthesis in younger secondary needles. Barnes (1972a, 1972b) used a Mast meter to 

measure ozone; the Mast meter can underestimate the ozone concentration unless 

it is calibrated against a reference standard. Also, the sample size used in 

these experiments was very small, four to nine seedlings. It is possible that 

variation among samples may have masked potential effects in some of the 

experiments (Barnes 1972a). 

Miller and others (1969) found that 3-year-old ponderosa pine seedlings 

sustained a 25-percent reduction in apparent photosynthesis after a 60-day 

exposure to an ozone concentration of 150 ppb for 9 hours per day. In another 

study, Coyne and Bingham (1981) measured changes in gross photosynthesis in 

needles of ponderosa pine trees of various sensitivities to ozone, Needles 

sustaining slight, moderate, and severe injury exhibited a 90 percent reduction 

in gross photosynthesis after exposure to a cumulative dose of 800,000, 700,000, 

and 450,000 ppb-hours ozone, respectively, in a 3-year time period (2 years for 

the most sensitive class of trees). The percentage inhibition in gross photosynthesis was based on photosynthesic rates of newly emerged needles; no true 

controls were used in the experiment. The authors emphasized that the decline 

in photosynthesis reflected the superimposition of ozone effects on normal 

aging. 

Reich and Amundson (1985) over a period of several years measured the 

photosynthetic response of four tree species [sugar maple, eastern white pine, 

hybrid poplar (Populus deltoides x trichocarpa) and northern red oak] to ozone 

concentrations representative of those found in the ambient air of the Eastern 

United States. Exposure concentrations ranged from 20 to 140 ppb. A 1 inear 

reduction in net photosynthesis was observed after long-term exposures (5 to 

7 hour per day, 3 to 7 days per week, for 3 to 12 weeks) in a11 species. 

Reduction ranged from 10 percent in white pine at 30,000 ppb-hours to 40 percent 

in hybrid poplar at 20,000 ppb-hours. The authors suggest that, based on the 

results of their studies, reductions in net photosynthesis may be occurring 

over much of the eastern United States and southeastern Canada. 

Botkin and others (1972) observed suppression of photosynthesis in eastern 

white pine when 5-year-old sap1 ings were fumigated in small chambers. Three 

categories of sensitivity to ozone were observed. In sensitive trees, 

exposures of 900 to 1,000 ppb of ozone for 10 hours reduced net photosynthesis 

to zero; in trees of intermediate sensitivity, the same exposures reduced 

photosynthesis by approximately 50 percent; and in resistant trees, similar 

exposures had no effect upon photosynthesis. Ozone-induced suppression of 

photosynthesis in the trees with intermediate sensitivity was reversible if an 

ozone-free period fol lowed exposure. Photosynthetic suppression and recovery 

occurred prior to or without the expression of visible injury symptoms. 
d) Mixtures of Gaseous Pollutants 

The effects of exposure to 03, SO , NO, or NO2 individually on photosynthesis in plants has been discusse8 above. The effects of mixtures of 

these substances on photosynthesis, however, have been poorly studied. 

Sulfur dioxide, nitrogen oxides, and ozone individually inhibit the 

process of photosynthesis. Interaction of these chemicals frequently elicits a 

greater plant response than any of them alone do. The rate of reduction of 

net photosynthesis during the first 2 days of exposure was greater than 

additive when sugar maple and white ash were fumigated simultaneously with 

500 ppb of SO and 03. For sugar maple, the reductions persisted for 1 week 

(Carl son 1979 . The concentrations (500 ppb) used in this experiment are much 

greater than those usually encountered in the ambient air, 

Fumigation of plants with high concentrations (1200 ppb) of NO2 stimulates 

the formation of nitrate reductase (NiR). Plants exposed to NO and SO2 

simultaneously failed to show an increase in NiR levels (Mansffeld and 

Freer-Smith 1981). Wellburn (1982, 1985) has proposed a mechanistic explanation for the NO /SO interaction. The presence of SO completely prevents the 

induction of in&eas$d activity by NO . This phenomena% was observed in several 

clones of Lolium as well as in other rasses fumigated with low levels (68 ppb) 

of the two pollutants, while fumigation with NO2 alone increased the levels of 

NiR (Wellburn 1982). 

In a shdy of the effect of SO and NO in combination on alfalfa 

( Medicago sativa L. ), White and othgs (19743 observed inhibition of the 

apparent rate of photosynthesis. Exposure to 2,500 ppb SO for 2 hours 

resulted in 2 to 3 percent inhibition. Exposure to 2,500 pp8 of NO produced 

no measurable inhibition, but exposure to both gases simultaneously it 2500 ppb 

for 2 hours caused a 9 to 15 percent inhibition, while exposure to 15 ppb 

reduced photosynthesis 7 percent. The degree of synergism decreased as the 

concentrations increased until exposure concentrations reached 5,000 ppb. At 

that point, no synergism was observed. Bull and Mansfield (1974) also reported 

inhibition of photosynthesis in pea (Pisum sativum L. ) by combinations of SO2 

and NO2. 

In the sunflower (He1 ianthus annuus L. cv Russian Mammoth), comb.inations 

of NO, SO, and 0, measurably reduced the rate of photosynthesis when compared 

with fhe effects 6roduced by each pollutant alone (~urukawa and Totsuka 1979). 

Inhibition of photosynthesis caused by NO /S02/03 combinations resembled those 

caused by S02/02 (Furukawa and Totsuka 197%). 

Summary. --Photosynthesis in forest trees is a1 tered by exposure to airborne chemical pollutants. Sulfur dioxide, nitrogen oxides, and ozone, as well 

as mixtures of the three chemicals, can inhibit the process of photosynthesis 

and decrease the amount of photosynthate produced. Regardless of the mechanism, a substantial reduction in photosynthesis will ultimately affect the 

growth, vigor, and reproduction of trees if repair does not occur. 
4.2.4 Is Respiration in Forest frees Altered by Exposure to Airborne Chemical 

Pol l utants? 

Respiration is the process through which the energy from the carbohydrates 

produced by plants during photosynthesis is released for growth, maintenance, 

and reproduction, In addition to its use in the formation of plant tissue 

and maintenance, some photosynthate is allocated for storage of compounds such 

as organic and amino acids, sugars, starches, fats, and proleins. Carbohydrates 

are also used to make secondary compounds such as alkaloids, tannins, pigments, 

and growth regulators (Waring and Schlesinger 1985). These processes occur at 

several sites within plant cells, and are likely to be vulnerable to disruption 

by gaseous air pollutants. Respiratory pathways and organelles may be affected 

(Black 1984). 

Increased respiration is a well-recognized response to gaseous pollutant 

exposure (McLaughlin and Shriner 1980). Exposure to SO2, 0 ,or NO can a1 ter 

such major physiological processes as photosynthesis in tre s (section 4.2.3) 

and affect the rate of respiration. Respiration is sensitive to a variety of 

chemicals (e.g. , SO ) that may inhibit various stages of the overall process 

(Kramer and ~ozl owsai 1979). However, the conversio" of carbohydrates, fats, 

and proteins into new plant tissue to repair pollutant-induced injuries requires 

large amounts of energy to be supplied through respiration (Kramer and 

Kozlowski 1979). Increased respiration, therefore, is caused by cellular 

injury rather than a cause of cellular injury (Kozlowski and Constantinidou 

1986a). 

Trees, when compared with plants in which foliage and seed production are 

predominant, require a disproportionate amount of carbohydrate for respiration 

because they have such a large amount of conducting tissue associated with 

their sapwood (Wari ng and Schlesinger 1985). Injury, therefore, requires 

changes in the normal allocation patterns and places stress on the starch 

reserves in trees. The maximum level of starch reserves in twigs and stems may 

serve as an indicator of the degree of injury a tree can tolerate (Waring and 

Schl esi nger 1985). 

Summary.--Increased respiration is a well-recognized response to gaseous 

pollutants. Energy normally used for maintenance is required to repair any 

injury caused by pollutants. 

4.2.5 Are Carbon Uptake and Allocation Altered by Exposure to Airborne 

Chemical Pol 1 utants? 

Carbohydrates produced in the process of photosynthesis are normally used 

for the growth and maintenance of trees. Deciduous trees grow new leaves each 

year; conifers produce new needles. Roots, too, require energy to produce new, 

fine roots and develop mycorrhizae. Patterns of carbon allocation directly 

influence growth rate. The strategy for a1 location of carbon may change during 

the life of the plant, as well as with different environmental conditions 

(fig. 27; Winner and Atkinson 1986). Factors such as air pollutants that 

limit carbon fixation shift allocation to new leaves, while factors that limit 

the availability of nitrogen or water wi 11 shift a1 location to the roots. The 

increase in the allocation of photosynthate to new leaves due to the reduced 

photosynthetic capacity of plants exposed to ozone and to SO2 has been reported 

by Oshima and others (1979) and by Reinert and Gray (1981). Therefore, air 

pol 1 utants that l imi t carbon gain and nutrient availabi 1 i ty may a1 so suppress 

the growth rate and total biomass production (Winner and Atkinson 1986). 
l N lTlAL STATE 

PHOTO ROOT NITROGEN 

X RAE S UPTAKE 

REDUCE CARBON SUPPLY 

LOWER UGW 

LOWER C02 

CLOSE rnMATA 

ESlS 

NEW ALLOCATION STATE 

LEiW PHO E~C NITROGEN 

BIOMASS x UFTME 

Figure 27. Model relating interaction between root and shoot allocation. 

Source: Winner and others (1985). 
Changes in normal allocation patterns are an indication of stress (McLaughlin 

and Shriner 3980). Maring (1987) points out that the accumulation of amina 

acids in foliage is associated with reduced allocation of carbohydrates to the 

roots. 

In the following section, changes in carbon allocation associated with 

experimental exposures to SO2 and O3 are discussed. 

a) Sulfur . --Gaseous pol lutants general ly, and sulfur dioxide in 

particular, decrease the photosynthetic capacity of plants and shift carbon 

allocation towards the leaves. They also alter carbon translocation from the 

needles or leaves of trees. Suppression of root growth may result (Winner and 

Atkinson 1986). Inhibition of translocation of sugars by SO without the 

accompanying inhibition of net photosynthesis a1 so has been rgported (Noyes 

1980). 

Inhibition of translocation was observed in bean3(Phaseol us vulgaris L. 

cv. Black Valentine) when exposed to 100 ppb (26 pg/m ) of SO for 2 hours. 

Trans1 ocation was reduced 39 percent. Autoradiograms of 4~- fabel ed primary 

leaves indicated that the phloem-loading processes and/or axial transport 

within sieve tubes were significantly inhibited or delayed by SO 

(Noyes 1980). Teh and Swanson (1982) also reported a reduction 

loading. Translocation rates out of SO -stressed leaves were lower than 

expected on the basis of decreases in ph%tosynthesi s. Though photosynthesis 

recovers from SO inhibition after exposures terminate, the effect of SO2 on 

translocation pe8sists (Darral 1 1986; Noyes 1980; Teh and Swanson 1982). 

Reinert and Gray (1981) observed changes in carbon a1 location in radish 

(Raphanus sativus L. ) when fumigated with 260 and 400 ppb SO2 for 3 or 6 hours. 

A significant change in the root/shoot dry weight ratio was observed. Winner 

and others (1985) also reported a shift in carbon allocation from root to shoot 

in radish exposed to SOT 

b) Nitrogen Compounds. --Reinert and Gray (1981) reported no change in root/ 

shoot ratio in plants exposed to NO, at 200 and 400 ppb for either 3 or 6 hours. 

Though many studies cite the inhibiting effects of NO and NO on photosynthesis, 

not many studies mention changes in carbon allocation resilting from the 

inhibition. However, the information regarding carbon allocation derived from 

studies using SO and 0 makes it possible to hypothesize that exposure to NO 

concentr&ions tiat inhibit photosynthesis does a1 ter carbon a1 1 ocation 

c) Ozone.--Depression or inhibition of photosynthesis in the foliage can 

al ter the a1 location and translocation of photosynthate (e. g. , sucrose) from 

the shoots to the roots and other organs (Tingey 1974). Tingey and others 

(1976a) observed that ozone exposure differentially affected the metabolite 

pools in the roots and tops of ponderosa pine seedlings grown in field chambers. 

The amounts of soluble sugars, starches, and phenols tended to increase in the 

tops and decrease in the plant roots exposed to 0.10 ppm O3 for 6 hours per day 

for 20 weeks. The sugars and starches stored in the tree roots were significantly less than those in the roots of the controls. 
McLaughlin and others (1982) observed that the decreased availability of 

carbohydrate in trees exposed to O3 reduced the vigor of root systems and 

possibly enhanced the susceptibility of the trees to root diseases. The loss 

in vigor of the trees was accompanied by reduced annual radial growth and a 

loss in the capacity to respond in years when conditions were favorable for 

growth, The primary cause of decrease in radial growth appeared to be exposure 

to high concentrations of ozone and a sequence of events and conditions that 

led to premature senescence and loss of older needles, lower growth, and 

reduced photosynthate availability for growth and maintenance of trees 

(McLaughlin and others 1982). Carbon-14 transport patterns indicated that 

older needles were sources of photosynthate for new needlelqrowth in spring and 

were storage sinks in the fall. The higher retention of C-photosynthate by 

foliage and branches of sensitive trees indicated that the export of photosynthate to trunks and roots was reduced. 

Data from herbaceous plants support the fact that 0 inhibits the 

a1 1 ocation and translocation of carbohydrates from shoots 80 the roots (U. S. 

Envi ronmental Protection Agency 1986). 

d) Mixtures of Gaseous Pollutants.--Reinert and Gray (1981) reported that O3 

exposure at 40T ppb reduced root dry weight more than exposure at 200 ppb 

during exposures of 3 or 6 hours. Sulfur dioxide also depressed the root 

weight at both concentrations (200 ppb and 400 ppb); however, when NO and SO2 

were present together, there was a synergistic depression of the roo%/shoot 

ratio. The average 03-induced reduction of root weight of radish was additive 

and SO2 were present. Though the foliage was the direct recipient of 

ss, the response appeared as a reduction in root weight. Exposures to 

pollutant mixtures do alter carbon allocation in herbaceous plants. If pollutant combinations inhibit photosynthesis and the production of carbohydrates, or 

a1 ter biochemical or physiological processes in trees, changes in carbon 

allocation should result. The responses of trees to pollutant combinations and 

the effect on carbon allocation have not been studied. 

Summary. Gaseous airborne pollutants, by decreasing the photosynthetic 

capacity of plants, shift the a1 location of carbon into new leaves and away 

from the roots and also decrease the carbohydrates available for plant maintenance and storage. The amount of starch stored in branches and twigs may be an 

indicator of how much injury a tree can endure. 

4.2.6 Is Increased Foliar Leaching in Forest Trees a Response to Airborne 

Chemical Pollutant Exposure? 

beaching of nutrients From the leaves of trees has been implicated as a 

result of ozone exposure. Leaching, a widespread natural phenomenon, is a 

biologically passive process that can remove a variety of substances from 

plants and temporarily reduce their concentrations in the tissues (Tukey 

1970). Metabolites are leached by the action of aqueous solutions such as 

rain, dew, fog, or mist from the wide variety of plants that have been studied. 

Inorganic nutrients, organic substances, all of the naturally occurring amino 

acids, and many organic acids have been detected in leachates (Tukey 1980). 

The amounts of solutes leached can be greater than the amounts present in 

foliage prior to leaching, suggesting a connection with the transpirational 

stream. The rates of root uptake and translocation may increase during leaching (Tukey 1970; Tukey and others 1958). 
a, -om 3 cf m-53mzr 

o rno -hn at sc 

0 m 3 mu3 

4. 

:%g 

<s 

4, X Q 

= mu' 

~2. 

m 

0 3 

ru m 

-5 m 

:!5 

3- m 

V V, 

Q 

-5 ct 0, 0 

ct 

m c+ 

3- 

!?am 

-V, 

0 --r* 

n cf 

ru m rt v, -1. 

0 0 3 -h 

!a" 

3 -5 0-ID 

Tu 3 cf C m 

CT v, 

3. 

m 3 3 m 

ctcf P, s u- 2" 

Cu -0 T-n 

m 

a, 

4 n 

m ct V, --1* 

s c d d. 

ctrt 

mu U 

'd. 

3 cf 

3- 

!2=' 

m -. u-=3 

Cu t-r. n c 

7; 

Q 

a 

-5 

m 

f" 

2 

ct 

m 

4 

m 

Q 
1) Atmospheric deposition may contribute basic materials such as 

soil dust, or acid-forming materials such as gaseous SO2 and NOx 

and sulfate aerosols; 

2) The canopy may acidify precipitation through the release of 

weak, organic acids; 

3) Canopy surfaces may increase pH by releasing basic cations in 

exshange for hydrogen ions in the incident precipitation. 

Gaseous, particulate, and dissolved matter are potentially released from, 

taken up by, and transformed at, a1 1 canopy surfaces (Parker 1987). Because 

all of the processes cited above contribute to the alteration of precipitation 

quality, forest canopies commonly change the pH of precipitation reaching the 

forest floor. In addition, most of the field studies suffer because neither 

dry deposition inputs of hydrogen ions nor the stemflow releases of hydrogen 

and other cations are accounted for in the data. Moreover, highly acidic 

rainfall events, when compared with less acidic events at a given location, do 

not appear to enhance foliar leaching. Clearly, on the basis of the above 

reasons, change in pH alone is not a valid indicator that acidic deposition 

increases 1 eachi ng. 

A wide range of acidities (pH 6.7 to 2.3) has been used in laboratory 

studies of the effects of leaching. The degree of enhanced leaching is relatively small considering the range sf acidities used. Moreover, the acidities 

used in the experiments are much greater than those normally encountered in 

bu1 k precipitation. No evidence exists for a pH level below which losses in 

solutes are dramatically increased (Parker 1987). 

Summary. Foliar leaching is a widespread, natural, biologically passive 

phenomenon that is influenced by many external and internal factors. Gaseous, 

particulate, and dissolved matter are potential ly released from, taken up by, 

and transformed at, all canopy surfaces. Release of materials from the forest 

canopy involves the action from two broad source categories: 1) from within 

the forest ecosystem, and 2) from without the ecosystem, chiefly from atmospheric deposition. At the present time ozone is the only gaseous airborne 

chemical that has been implicated as increasing foliar leaching, and then, when 

in association with acidic fog. The effects of acidic deposition are difficult 

to assess because forest canopies alter the pH of precipitation reaching the 

forest floor: 1) by contributing basic materials, or acid-forming materials 

(e-g., SO and NO ); 2) through the release of weak, organic acids; and 

finally, 3) by exc%anging basic cations for hydrogen ions in incident precipitation. These processes of throughfa1 l enhaficement make it. difficult to 

attribute leaching to a single phenomenon such as acidic precipitation. In 

addition, highly acidic rainfall incidents at a given location, when compared 

with those of low acidity, do not appear to increase foliar nutrient leaching. 

4.2.7 Does Exposure to Airborne Chemical Pollutants Disrupt the Processes of 

Regeneration and Reproduction in Trees? 

Reduction in the physbological efficiency of leaves often decreases crop 

yield and qua1 ity of seeds and fruits. Many studies indicate a reduction in 

yield in agronomic crops exposed to SO2 and O3 (see table 9 for critical 
concentrations). Sensitivity to pol 1 utants varies with the crop, cul tivar, 

envjronmental conditions and with the concentration and length of exposure. 

Regeneration, reproduction, and yield are likely to suffer if pollutant exposures inhibit photosynthesis and the production of carbohydrates necessary for 

growth and maintenance. A few studies have shown pollutant inhibition of 

pollen germination and tube elongation under experimental conditions (fig. 28; 

Kozfowski and Constantinidou 1986a; Smith 1981). 

Tree vigor is important in reproduction. In forest stands, dominant trees 

usual 1y produce the most seeds. Dominant trees produced almost a1 1 of the 

cones in stands of ponderosa (Pinus Dougl. ex Laws.) and sugar (P. 

lambertiana Dougl. ) pine trees. Codominant trees produced only 1 to 1.5 percent, and almost no cones were produced by the intermediate and suppressed 

trees. In the crowns of red pine trees (P. resinosa Ait.), the most vigorously 

growing branches produced 1 arger cones c&tai ni ng more and higher qua1 i ty seeds 

than did the cones on the less vigorously growing branches (Kozlowski and 

Constantinidou 1986a). Finally, the importance of tree vigor to reproduction 

is illustrated by the fact that individual trees of loblolly pine (P. -- taeda L.) 

when released from competition (i.e., thinned) produced 10 times as many cones 

as did trees not released from competition (Kozlowski and Constantinidou 

1986a). 

Both regeneration and reproduction in trees can be affected by changes in 

carbon aff ocation (fig. 271, AS indicated above, tree vigor i s important 

for reproductive growth. Injury causes trees to shift their resources from 

growth, storage, and reproduction to injury repair. Defoliation is a classic 

example of a circumstance under which trees must regenerate tissue (McLaughlin 

and Shriner 1980). The ability of a tree to withstand such stresses can be 

assessed by eval uati ng the solible and easi ly mobi 1 ized carbohydrate reserves 

near points of need. Webb (1981) reported that Douglas-fir [: 

menziesii (Mirb.) Franco] and grand fir [Abies (Doug1 . Lindl.] could be completely defoliated andm1 initiate new foliage, as long 

as starch concentrations in twigs remained above a certain minimim. The - 

maximum amount of carbohydrate reserves in twigs and stems may be an indication 

of the degree of injury a tree can endure and still recover. Starch reserves 

are particularly important in deciduous trees that must survive much of the 

year on energy derived from storage (Waring and Schlesinger 1985). 

a) Sulfur Compounds 

Sulfur Dioxide.--Reduced production of cones in western larch (Larix 

occidental i s Nutt. ) , l odgepol e pi ne (Pi nus conhrta Dougl . ex Loud), ponderosa 

pine ( P. ponderosa Laws. ) and Douglas-fir [ menziesi i (Mi rb. ) 

Franco] exposed to high concentrations of SO, was reporLed by Scheffer and 

Hedgcock (1955). Houston and Dochi nger (1977) a1 so reported significant 

differences in cone. seed, and pol len-characteristics when 15-to 25-year-old 

eastern white pine and red pine (Pi nus resinosa Ai t. ) growing in SO -bol luted 

areas were compared with trees growing in SO,-free areas. In whi?e pine, 

percent pol len germination was reduced, seeds 6er cone were fewer in number and 

lighter in weight. In red pine, cone length, seed weight, percent filled seed, 

percent seed germination, and pollen tube length were significantly less in 

SO,-exposed than in non-exposed trees. The reductions in cone, seed, and 

pof len- production occurred hespi te the absence of visible injury to the trees. 
SEEDS 

MATURE TREES 

- 

Figure 28. Potential points of interaction between air pollutants and 

sexual reproduction of forest trees. 

Source: Smith (1 981 ). 
The effects on reproduction discussed above were observed on trees growing in 

the vicinity of power plants where SO concentrations are usually high. A 

definitive statement concerning the ef&cts of SO at ambient concentrations on 

reproduction in forest trees cannot be made a t t h 

b) Nitrogen Compounds 

Oxides. --Nitrogen oxides have not been reported as a1 teri ng 

reproduction and regeneration, probably because ambient concentrations do not 

usbal ly reach injuribus levels. 

Oxidant exposure of ponderosa and Jeffrey (Pinus &rev. & Balf.) 

pine in the San Bernardino forest of California was observed to reduce seed 

production. Dominant trees of both ponderosa and Jeffrey pines produced the 

most cones, and cone production increased significantly with tree age. However, 

dominant trees of both species 130 years or older, when severely injured by 

oxidant air pollution, bore significantly fewer cone crops over a 6-year period 

than did uninjured trees (Luck 1980). Miller and others (1982) suggest ozone 

injury-related reductions in photosynthesis, carbohydrate formation and allocation are the reason for reduced cone production. 

In the eastern United States, reduction in cone or fruit production has 

not been reported even though ozone exposure has been reported to alter carbon 

uptake and allocation and to reduce vigor in sensitive trees (sec. 4.2.5). 

McLaughlin and others (1982) associated needle injury and senescence with a 

loss of vigor and reduced photosynthate availability in eastern white pine 

(Pinus strobus L. ) trees exposed to ozone. Loss in vigor was accompanied by 

reduced annual radial growth and loss in capacity to respond to stress in years 

when conditions were favorable. Further, they state that repair of foliage 

injured by ozone exposure can place a drain on carbohydrate reserves. Moreover, 

a reduction in photosynthesis reduces the production of carbohydrate. Though 

the authors did not mention reduction in cone production, reduction in vigor 

and amount of photosynthate, as mentioned above by Luck (1980), affect 

reproduction. 

Summary. Exposure to air pollution, specifically SO and 0 , by inhibiting 

photosynthesis and a1 teri ng the a1 location of carbohyd6ates can decrease 

reproduction and inf 1 uence regeneration in trees. Evidence from the ambient 

ozone exposure of ponderosa and Jeffrey pine in the San Bernardino Forest of 

California, as well as from experimental studies, indicates that dominant trees 

and trees with the most vigor produce the most cones. Trees in stands that 

were thinned produced more cones than trees not released from competition. The 

effects of gaseous pollutant mixtures on reproduction and regeneration have not 

been reported; however, if photosynthesis and carbohydrate production are 

reduced by exposure to gaseous pollutant mixtures, then reproduction is likely 

to be affected. 

4.2.8 Is Host Susceptibility to Pests, Pathogens, and Mycorrhizal Colonization Altered by Exposure to Airborne Chemical Pollutants? 

Plant tissue is generally not very nutritious to most insects and pathogens. Drought, air pol 1 uti on and other environmental stresses, however, can 
alter host susceptibility to attack from insect pests and pathogens by altering 

biochemical and physiological processes that influence the formation and 

concentrations of carbohydrates, nitrogen, sugars, and plant allelochemicals. 

Drought stress-related changes in nitrogen and carbohydrate-containing defense 

compounds can enhance the ability of insects to attack trees (Mattson and Haack 

1987; Sharpe and others 1986; Waring and Schlesinger 1985). Similar changes 

have been observed in trees injured by air pollution. Mattson and Haack (1987) 

suggest that plants have a more or less universal response to stress. Insects 

appear capable of detecting and capitalizing on the changes caused by stress. 

Tree vigor plays an important role in insect and pathogen attack (Hanover, 

1975; Kozlowski and Constantinidou 1986b; Larsson and others 1983). A common 

observation is that vigorous standing trees of many different species are more 

resistant to bark beetle attack than freshly cut or 1 ightning-struck trees 

(Hanover 1975). Larsson and others (1983) observed that ponderosa pine trees 

of low vigor were more prone to attack by beetles than trees of high vigor. 

Vigor was measured as stem growth per unit of leaf area. The attacks by 

beetles increased below a vigor threshold equivalent to production of approximately 100 grams of wood per square meter of leaf area. Susceptibility of 

lodgepole pine ( Pi nus contorta Doug1 . ex Loud. ) to mountain pine beetle 

(Dendroctonus k also was found to increase below the same 

vigor index (Waring and Pitman 1985). 

Christiansen and others (1987) hypothesize that the ability of trees to 

withstand attacks by bark beetles and their associated fungi is 1 inked to the 

amount of carbohydrates that can be utilized directly for defense wound reactions. Therefore, any environmental factor that restricts the size of the 

canopy or its photosynthetic efficiency can weaken a tree's resistance. 

To coexist, plants and parasites must be in a state of equilibrium in 

which neither total ly dominates the other. Plants have evolved a variety of 

defense mechanisms, and parasites and pathogens have in return evolved countermeasures for avoiding or overcoming plant defenses. Neither plant defenses nor 

parasite countermeasures are completely effective; hence, a balance exists in 

which neither dominates. Air pollution-induced biochemical changes within 

plants can upset the relationship that the host and parasites or pathogens have 

developed over time (Hughes and Laurence 1984). 

Different pine species may have evolved different defense strategies. 

Pines in the Southern United States have developed the resin ducts (constitutive defense) as contrasted with western conifers that have developed a hypersensitive wound reaction (induced defense). In the former, the defense system 

is present when the beetle atGacks occur; in the latter, the defense system, a 

wound reaction, develops at the time of the attack. It has been hypothesized 

that the constitutive defense system evolved because southern trees experience 

asynchronous attacks by several multivoltine (having multiple broods in a 

season) bark beetle species. The resin duct system is the most energy efficient 

for them (Christiansen and others 1987). Air pollutants weaken and predispose 

trees to certain diseases, particularly those caused by non-obligate pathogens. 

Large infestations of bark insects were observed in forest stands that were 

weakened by air pollution (Hughes and Laurence 1984; Scheffer and Hedgcock 

1955). 
Plant parasites, specifically herbivorous insects, depend upon their host 

for nutrition and for many chemical and physical cues that regulate their 

feeding and reproductive behavior. Imbalances in the insecl-tree relationship 

may be caused by variation in any of the four basic host characteristics: 

1) morphology and anatomy of the host; 2) chemical repellents produced by the 

host; 3) chemical attractants produced by the host; 4) nutritional status of 

the host (Hanover 1975). Biochemical changes induced by air pol 1 utants can 

alter these relationships. The ensuing section cites some instances when this 

has occurred. Again, probably because their concentrations in ambient air are 

not high enough to cause injury to trees, studies mentioning the effects of 

nitrogen oxides in altering pathogen or parasite relationships do not exist. 

a) Sulfur Compounds 

Sulfur Dioxide.--Increased production of both chemical insect attractants 

and repellents has been reported in trees exposed to SO . Renwick and Potter 

(1981) reported increased emissions of the volatile te6penes that serve as 

chemical attractants for many insect pests after balsam fir [Abies balsamea 

(L. ) Mil 1. ] trees were exposed to SO (200 ppb, 5 hours per dayrm. 

On the other hand, ethane, a strong2 repellent of the wood-boring beetle 

Monochamus a1 ternatus, ameared after red ~ine ( P. resinosa Ait. 1- and Daner 

birch (Betul a papyri <era '~arsh. ) were expo'sed SO at 300 ppb conti n;o;sl 

for 3 todays (Kimmerer and Kozlowski 1982). ~Joduction of oleoresin. 

anoth-er chemical repellent produced by trees injured by ozone, has not been 

reported for trees injured by SO,. Sulfur dioxide exposure, however, increases 

the amounts of solubie sugars ahd nitrogen in plant -foliage, inner bark, and 

sapwood of exposed trees (Mattson and Haack 1987). 

Scheffer and Hedgcock (1955) observed increased numbers of bark beetles 

(Dendroctonus spp.) and engraver beetles (+ spp.) on many of the larger 

ponderosa pines injured by SO exposure. The authors suggest the attacks 

followed weakening of the tree$ and were not the primary cause of the weakening. Most species of bark beetles can only breed in trees that exhibit severe 

decline, or are already dead; therefore, they are merely promoting decomposition and mineralization. However, a few species are capable of attacking and 

killing living, sometimes quite healthy, trees (Christiansen and others 1987). 

Air pollutants may increase or decrease the severity and incidence of 

plant disease depending upon whether it is the host or the pathogen that is 

adversely affected by pollutant exposure (Treshow 1980b). 

Exposure to SO according to reports, can suppress fungal growth. 

Scheffer and ~edgco& (1955) reported that fungi parasitic on needles and 

leaves appeared to be retarded or inhibited in areas exposed to the greatest 

concentrations of S o, two species of rust fungi, 

on quaking aspen ( tremuloides Michx,) and M. s 

cottonwood (Populus trichocarpa Torr. & Gray) were absent in areas with the 

highest SO, concentrations, sparsely present in areas where trees were 

moderately Li njured, and abundant where trees were uninjured. Li nzon (1978) 

reported the virtual absence of white pine bl i ster rust (Cronarti urn ri bicul a) 

from' forests within 25 miles of the Sudbury, Ontario, Canada, smelters in line 

with the prevailing wind. A reduced incidence of fungi that attack plant 

fol iage was also reported. In contrast, root-rotting and heart-rot fungi such 
as Armillaria mellea abietinum (syn Lenzites abietina), Trametes 

serialis and T more common in trees injured by SO2 than in 

uni njured trees. 

b) Ozone 

The majority of the studies relating the effects of ozone exposure on 

host-pathogen or insect-host relationship were made in the West where ozone 

injury of forests has been more prevalent. 

Bark beetles are the most damaging and economical ly sign! f icant insect 

pests of commercially important conifers in the United States (Stark and Cobb 

1969). Beetle outbreaks in western forests are associated with several predisposing factors. These include host weakening caused by photochemical oxidants; 

root disease initiated by the fungi ~eterobasidion annosum (syn Fomes annosus) 

or Verticicladiella wagenerii (Stark and Cobb 1969); insect defoliation, such 

as pine looper stripping of ponderosa pine (Dewey and others 1974); and various 

climatic stresses, such as drought and windthrow (uprooting and breakage by 

strong winds) (Rudi nsky 1962). 

Ponderosa pines exhibiting advanced ozone injury symptoms were more 

frequently attacked by bark beetles than those with less severe symptoms (Stark 

and others 1968). The mountain beetle was not found infesting healthy trees. 

Ozone injury reduced oleoresin formation, exudation pressure, resin yield, and 

the rate of resin flow. The crystallization rate, however, increased. Further 

effects associated with ozone injury included reductions in sapwood and phloem 

moisture, thickness, and carbohydrates. Cobb and Stark (1970) concluded that 

the changes in host chemistry brought about by ozone injury made the trees 

more susceptible to bark beetle infestations. Mattson and Haack (1987) observed 

similar changes in oleoresin composition in trees under drought stress. They 

associated increased host susceptibility to insect attack with the changes. 

Jones and Coleman (1988) and Coleman and Jones (1988a, 1988b) studied the 

i nteracti on of pl ant stress and i nsect behavior and performance on eastern 

cottonwood (~opulus deltoides Bartr. ex Marsh.) using ozone as the stress. 

Based on their studies, they conclude that an acute dose of ozone (200 ppb for 

5 hours) that does not directly affect plant growth, or produce signs of 

visible injury, does not significantly change feeding and oviposi tion preference 

of the beetle versicolora. Both larvae and adult beetles preferred 

feeding on, and adults consumed more foliage of, ozone-stressed trees than 

controls when presented with a choice. In contrast, females preferred to 

oviposit on unstressed foliage. The changes in preference were maintained for 

3 years and to approximately the same degree in two cottonwood clones. 

The consequences of ozone stress on cottonwood and insect feeding and 

oviposition behavior are complex and not immediately obvious. To attempt a 

prediction of the effects of ozone stress and tree resistance to insect predation requires a comprehensive evaluation of the oviposition preference, studies 

on growth, survivorship, fecundity of the insects, their relationship to other 

members of the insect and pathogen community present on the cottonwood trees 

and, finally, the interaction of all of these organisms with the ozone-stessed 

trees (Jones and Coleman 1988). The authors suggest that the relationships of 

the cottonwood aphid ( ) and the leaf spot fungus 
(Marssonina brunnea), which are unaffected by the same ozone exposures that 

affect the beetles and the cottonwood rust (Melampsora medusae) whose reproduction is reduced by those exposures, could be altered because of ozone stress 

having changed the leaf resource availability. 

Mycorrhizae assi st in protecti ng coni fer roots from pathogens such as 

Heterobasidion annosum (Krupa and Fries 1971). Injury to the mycorrhizae or 

reduction in the number of mycorrhizae, such as can be induced by ozone exposure, can remove this protection. Non-mycorrhizal and mycorrhizal root systems 

contain essentially the same major volatile compounds; however, studies using 

Scots pine (Pinus sylvestris L. ) indicate that the concentrations of monoterpenes and sesquiterpenes increase twofold to eightfold in roots infected 

by mycorrhizae. Many of the monoterpenes identified in mycorrhizal root 

systems are constituents of the oleoresins commonly found in conifers. 

Volatile oleoresin components from ponderosa pine have been shown to inhibit 

the growth of H. annosum and four ~eratoc~stis species (Cobb and others 1968), 

and are believgd to aid in defense of tree from bark beetles (Stark and Cobb 

1969). James and co-workers associated decreased oleoresi n exudation with 

increased susceptibility to infection by H. annosum in roots (James and others 

1980a) and cut stumps (James and others 19%0bmonderosa and Jeffrey pines. 

Christiansen and others (1987) also mention that production of resins, as 

we1 1 as phenolic compounds, in coniferous trees prevents fungal as we1 1 as 

beetle attack. When fungi, bacteria, or viruses infect plants, plants respond 

with a hypersensitive reaction. A necrotic area that deprives the invader of 

living tissues for food forms around the wound. If, due to the number of 

attacks, the resin concentration of the reaction zone decreases, the ability of 

the fungi to invade sapwood increases (Christiansen and others 1987). Skelly 

(1980) reported increased incidence of root disease caused by Verticicladiella 

procera in oxidant-injured eastern white pine in Virginia. 

Air pollution injury and reduced tree vigor have been implicated as the 

predisposing factors in the infestation by the balsam wooly adelgid (BWA) of 

Fraser fir [Abies fraseri (Pursh.) Poir.] growing in the Appalachian Mountains. 

The Fraser fir in the Southern Appalachians is particularly susceptible to the 

BWA that arrived in this country from Europe around 1900. - ~tmos~heric deposition is considered by Hain and Arthur (1985) to be the stress factor that 

alters the defense-response of Fraser fir and increases the host susceptibility 

to insect attack. Fraser fir decline on Mt. Mitchell, the highest peak in 

eastern North America, has in the past been attributed to the BWA. Recent 

observations suggest atmospheric deposition of SO2, ozone, and/or airborne 

microscopic particles of metals (e. g. , lead copper and nickel) probably play a 

role (Hain and Arthur 1985). 

Tree mortality patterns of Fraser fir vary at the six major locations where 

they grow. Trees on Mt. Mitchell are more susceptible than trees growing 

farther north in Virginia on Mt. Rogers. Genetic differences and the greater 

ability of the trees on Mt. Rogers to produce monoterpenes have been associated 

with their greater tolerance to BWA attack. In experimental studies, monoterpene production between pre- and post-wounding control samples and the 

hypersensitive zone was more variable in Mt. Rogers trees. Terpene production 

in Mt. Mitchell trees, on the other hand, was stable. In addition, necrotic 

ectomycorrhizal roots were observed on the trees on Mt. Mitchell; this condition 

was not reported from trees growing on Mt. Rogers (Hain and Arthur 1985). 
Mycorrhizae are sensitive to changes in the photosynthetic capacity of the 

host trees and their capacity to translocate carbon compounds to the roots 

(Hacskaylo 1973). For example, when seedlings of Virginia pine (Pinus * ** Mill,), inoculated with the mycorrhizal fungus terrestris 

and growing under a 16-hour photoperiod, were switched to 9-hour photoperiods, 

the seed1 ings became dormant within 4 weeks. No further invasion of rootlets 

by the fungus occurred even though root growth continued. Fungal sporophores 

were formed on the seed1 i ngs that remained under the 16-hour photoperiod. 

Studies have shown that simple sugars provided by plant roots are readily 

utilized by mycorrhizae and enhance fungal inoculation (Hacskaylo 1973; Krupa 

and Fries 1971). Berry (1961), in his studies to determine the cause of 

"emergence tipburn," dug up the roots of injured trees near Marlton, 

WV, where injury of eastern white pine trees had been prevalent for the past 

several years. The root systems of 10 healthy and 10 injured trees were 

examined. He observed that 72 percent of ozone-injured trees had dead feeder 

roots as compared with only 28 percent of the roots of healthy trees. Ozone 

exposure, therefore, not only inhibits the photosynthetic capacity of the 

eastern white pine, but also the a1 tered carbon a1 location is expressed as a 

change in the relationship between the roots and the mycorrhizal fungi. 

Wargo (1972) reported that a reduction in the starch concentration in the 

roots of repeatedly defoliated sugar maple trees was related to a series of 

complex physiological changes that resulted in small er amounts of sugars and 

changes in the number and concentrations of amino acids in the roots. These 

changes predisposed roots to attack by Armillaria mellea. 

Ozone has also been shown to influence bacterial symbiosis in herbaceous 

species. Whether it does so in trees and woody shrubs appears not to have been 

investigated. In herbaceous species, the rate of nitrogen fixation by symbiotic bacteria is dependent on the rate of photosynthesis by the plant. Symbiotic 

nitrogen fixation is the major biological source of fixed nitrogen (Tingey and 

Blum 1973). 

Ozone has inhibited infection of herbaceous plants by pathogens in controlled experiments; however, the inhibition occurred when ozone concentrations 

were several times greater than those usually encountered in the ambient air. 

Inhibition of forest pathogens by ozone has not been reported. 

Summary.--Tree vigor is very important in host tree-insect interactions. 

Most species of bark beetles are able to breed only in trees that exhibit 

severe decline or are already dead. Air pollutants such as SO and 03, by 

altering the physiological processes within trees, tend to ma8e them more 

susceptible to insect and pathogen attack, In addition, air pollutants can 

interfere with the natural defense processes within trees. Increased beet1 e 

and fungal attacks have been observed in forests subjected to SO or 0 

exposure. The abi 1 i ty of trees to withstand environmental , insect, %nd pa 

attacks has been related to the amount of stored carbohydrate. It has been 

hypothesized that the ability of trees to withstand attacks by bark beetles and 

their associated fungi is linked to the amount of carbohydrates that can be 

utilized directly for defensive wound reactions. Therefore, any environmental 

factor that restricts the size of the canopy or its photosynthetic efficiency 

can weaken a tree's resistance. Air pollutants, by inhibiting the process of 

photosynthesi s and the formation of photosynthate, can affect the tree's 

ability to withstand depredation by insects or pathogens. 
A tree's abi 1 i ty to form mycorrhi zal re1 ati onships has al so been re1 ated 

to its vigor. inhibiting photosynthesis and the amount of carbohydrate available for translocation to the roots can i n t e r f e r e w i t h mycorrhizal formation, 

and thus further weaken the abiliLy of trees to withstand environmental and 

biological stresses such as those posed by insecLs or pathogens. 

4.3 

Forest Trees? 

Tree growth is the culmination of many biochemical and physiological 

processes. The carbon necessary for growth, maintenance, and reproduction of 

trees is obtained during the process of photosynthesis. Carbon dioxide 

diffuses from the atmosphere into the leaves of green plants through the 

stomata and into the intercellular spaces, then into the mesophyll cells where 

photosynthesis takes place. Within the chloroplasts in the mesophyll cells, 

the energy of sunlight is used to convert carbon dioxide into carbohydrates 

that then can be used in the other biochemical processes necessary for plant 

survival. Virtually every perturbation of a plant community results in stress 

and affects the performance and survival of individual plants. The discussions 

in the foregoing sections point out how exposure to airborne chemical stresses 

by affecting physiological and biochemical processes inhibit photosynthesis, 

a1 ter carbon a1 location, influence vigor and reproduction, and can increase 

susceptibility of trees to pathogens and insect pests. The impact that ambient 

air pollutants stresses (e.g. , sulfur dioxide or ozone) have on growth 

is determined by the age of the tree, the complex interactions of pollutant 

and natural stresses with various environmental and genetic factors and their 

combined action on physiological and biochemical processes. 

Because forest trees and shrubs are perennial plants, they must cope with 

the cumulative effects of both short- and long-term stresses. Responses of 

trees to some stresses may appear rapidly as, for example, when sensitive 

eastern white pine needles express visible symptoms within days after exposure 

to episodes of high ozone concentrations. In the majority of instances, however, responses are more subtle and may not be observable for many years. 

Adaptation and response of trees to stresses are expressed by differential 

growth, a result of changes in carbon a1 location (Waring and Schlesinger 

1985; McLaughlin and Shriner 1980). It is, therefore, usually difficult to 

attribute responses observed today to a stress or stresses that occurred many 

years previous. The sections that follow discuss studies in which air pollutants have been implicated as the cause of a reduction in tree growth. 

a) Sulfur Compounds 

Many observations under ambient field conditions and experimental studies 

indicate that exposure to SO affects tree growth. The majority of the 

experimental studies used segdlings to study exposure-response because: 

1) they require little space, are readily transported and grown, and can be 

easily exposed under controlled conditions; 2) they are highly susceptible; 

3) they permit early detection of adverse effects; and 4) the experimental 

results are easily related to natural regeneration. Natural regeneration is 

important in forestry as an alternative to planting and in maintaining genetic 

diversity. A disadvantage, however, is that while the responses of older 

seedlings more closely resemble grown trees, the responses of young seedlings 

to air pollutants may not be applicable to large trees (Keller 1985). An 
additional shortcoming of many of the studies is that the concentrations used 

usually were much greater than those encountered in ambient air. The advantage of using high concentrations, on the other hand, is that they permit the 

early detection of adverse effects, particularly visible effects. 

In the studies that used concentrations approaching or equaling those 

encountered in the ambient air, foliar injury was usually the indicator of 

exposure effects. Fol iar injury, however, is not a re1 iable indicator 

t growth reductions have occurred or will occur. Growth reduction can 

occur without foliar injury (U.S. Environmental Protection Agency 1986). 

Carlson (1979) observed reduction in photosynthesis in sugar maple, black oak, 

and white ash without any foliar symptoms (table 12; sec. 4.3.2). Growth 

reductions were not mentioned, but the author suggests that plant productivity 

in pol 1 uted environments could be greatly reduced without vi si bl e symptoms 

appeari ng. 

Only a few major investigations have reported observations based on trees 

growing under natural conditions (Dreisinger 1965; Dreisinger and McGovern 

1970; Legge 1980; Legge and others; Linzon 1971, 1980; Scheffer and Hedgcock 

1955; 1981). All of these studies were associated with point sources of SO2 

and were made in Canada. Scheffer and Hedgcock (1955) made the classical 

study associated with emissions from the smelters near Trail, British Columbia; 

Linzon and his co-workers reported on the Sudbury, Ontario area, while Legge 

and his co-worker studied the effects of SO2 on a forest ecosystem near 

Whitecourt in west central Alberta. 

In areas of moderate sulfur dioxide concentrations, Douglas fir and ponderosa pine sap1 ings were reduced in vigor and many were ki 1 led. Near the 

smelters, the Douglas fir were most severely injured. The presence of SO in 

the atmosphere resulted in the presence of abnormal concentrations of sul%ur 

in the leaves (Scheffer and Hedgcock 1955). 

Table 13 indicates the degree of injury to eastern white pine observed 

over a 10-year period (1953-1963) in the sulfur-fume-effects area of smelters 

near Sudbury, (Linzon 1980). Linzon (1971, 1978) indicated that a pol 1 ution 

(SO ) gradient existed within the designated study areas. Effects correlated 

wel! with this gradient. Chronic effects on forest growth were prominent 

where SO air concentrations annually averaged 17- or 45-ppb (45 or 115 pg/m ) 

SO an{ such effects were not reported in areas receiving only 8 ppb 

(2fYpg/m ) SO annual ly. A1 though monitoring of SO was conducted in these 

studies, nei t6er concentrations of other air pol 1 utints nor their potential 

interactions were evaluated by the authors. It is also not clear to what 

extent high, short-term, peak SO exposures may have contributed to the 

effects that have been associated $i th annual average SO2 concentrations. 

In a study of the effects of low SO concentrations on a forest 

ecosystem, the main ecological process dir&ctly and indirectly affected was 

nutrient cycling (Legge 1980; Legge and others 1980). The deposition of 

sulfur onto the soil progressively altered the mineral nutrient balances of 

ecosystem components by modifying the physiological and biochemical functions 

and relationships among the components. Laboratory and field measurements 

revealed a decrease in the rate of photosynthesis in lodgepole pine x jack 

pine (Pinus contorta Dougl. ex Loud. x - P. banksiana Lamb. ) hybrids in the 

field. Reduction of adenosine triphosphate (ATP) concentration in pine tissue 

during SO fumigation in the field also was observed. Complete recovery 

fol lowed ?ermi nation of SO2 fumigation. Basal area increment measurements of 

200 lodgepole pine x jack pine trees from five ecologically comparable 

sampling sites revealed a decrease in wood production was directly related to 

the SO concentrationr. The maximum reduction in basal area increments 

occurrgd nearest the Whitecount Gas plant and fell to zero at 9.6 km. 

Recovery in the area began with the termination of SO emissions. Again, 

these studies indicate greatest tree injury was observgd near point sources 

and downwind from the SO2 source (Legge 1980). 

Reductions in tree growth observed after SO2 exposures undoubtedly result 

from a reduction in the rate of photosynthesis and alterations in carbon 

allocation and translocation. These processes-together with other processes 

such as mobilization of storage reserves, reallocation of nutrients between 

old and new leaves, and the factors that regulate the distribution of plant 

nutrient resources between roots, shoots, and reproductive plant partsinfluence the patterns of tree growth (Winner and others 1985). 

b) Nitrogen Compounds 

No published reports relating growth reduction of trees with exposures to 

NO or NO in ambient air exist. Ambient air concentrations do not reach 

levels c&able of causing injury to vegetation (National Research Counci 1 

1978b; U. S. Environmental Protection Agency 1982b). A study in Cal i fornia in 

the early 1970s compared the effects of NOx in smog with greenhouse fumigations (Thompson and others 1971). Leaf drop of navel orange trees was used to 

determine pollutant effect. Ambient concentrations of NO in smog varied 

considerably, depending on the level of emissions, time of gay, and the amount 

of sunlight. In early summer, peak concentrations of 120 ppb were observed; 

in late summer and early fa1 1, peaks reached 240 ppb. The concentrations of 

other chemicals in the smog are not mentioned. Greenhouse fumigations were 

with 100 and 500 ppb. Exposures lasted 35 days. The authors concluded that 

levels in the 10s Angeles Basin are not causing significant injury to 

us" (Thompson and others 1971). 

Skelly and his co-workers (Skelly and others 1972; Stone and Skelly 1974) 

related reductions in growth of eastern white pine and ye1 low-poplar to emissions of NO alone or in combination with SO2 from the U.S. Army Ammunition 

Plant at Raiford, VA. These studies are the only published instances where 

injury to trees in the field have been associated with NOx emissions. Sulfur 

dioxide, however, also was being emitted at the same time as NOx Since no 

air quality measurements were made during the period the growth reductions 

occurred, it is difficult to determine the concentrations of either pollutant 

at the time or determine which pollutant was causing the injury. 

Munitions production levels at the arsenal were used as an indicator of 

the amounts of NO2 and SO2 being emitted because actual measurements were not 

available for the periods (during World War II, 1941 through 1945, and during 

the Korean conf 1 ict, 1950 through 1956) when the growth reductions occurred. 

Nor were they avai lable for the period between World War 11 and the Korean 

conflict (1946-1949) when the tree growth increased. free growth was 

determined by using annual increment cores. Reductions in growth of eastern 
white pine and yellow poplar were inversely correlated with production levels. 

High production rates were associated with decreased tree growth, low production levels with increased growth. Emi ssion measurements were made during 

1967 and 1969 at the time of the Vietnam conflict when production levels were 

high, as a possible basis for estimating the emissions during the years no 

measurements were made. Maximum l-hour concentrations of NO and SO2 in 1967 

were 120 and 40 ppb, respectively. For 1969, the maximum 1-%cur NO concentrations of 580 ppb and 670 ppb of SO2 were substantially higher. 2~itrogen 

dioxide concentrations for either year are far below those shown experimentally 

to cause vegetational injury (table 11; sec. 4.3)- 

Kress and others (1982) studied the growth impact of 03, NO2 and/or SO 

on loblol ly pine seed1 ings. Significant growth suppression was observed it? 

the relatively sensitive full-sib loblolly pine families in all treatments 

except NO2 alone. The concentration of NO2 was 100 ppb. In another study, 

Kress and Skelly (1982) exposed willow oak, green ash, white ash, sugar maple, 

sweet gum, sycamore, yellow-poplar, pitch pine and loblolly pine seedlings to 

chronic doses of 0 and NO . Height growth of loblolly and Virginia pine was 

suppressed by expo?ure to $40 at 110 ppb; however, white ash and green ash 

exhibited greater height groith than the control. In their final assessment, 

the authors state that "the relationship that data such as these have to 

ambient Pi el d condi ti ons i s tenuous '' 

The foregoing discussions of data reporting the effects of NO exposure 

serve to reiterate the statement made earlier that concentration$ of NO 

sufficient to injure trees growing in the forests are seldom if ever encofintered except in the vicinity of point sources such as power plants or munitions 

factories. 

c) Ozone 

The effects of ozone on growth of vegetation, especially crop plants, are 

well documented and have been widely reviewed (Heck and others 1982; Krupa and 

Legge 1986; National Research Council 1977~; U.S. Environmental Protection 

Agency 1986). The effects of ozone on the growth of forest trees have not 

been studied as thoroughly. As in the case of SO , most of the studies cite 

the effects of ozone on seedlings. The majority o these experimental studies 

has exposed the seedlings in growth chambers or continuously stirred tank 

reactors; some have used open-top chambers in the field, Reich (1987) cites 

the results of many of these studies. The widespread distribution and 

regionality of ozone, however, have also made possible the observation of the 

effects of ozone exposure on trees growing in forests. The best documented 

studies are those in the San Bernardin~ Forest of California (Miller 1984; 

Miller and others 1982). There are, however, field studies and observations 

that indicate that ozone has been an important pollutant in the eastern United 

States for many years (Berry 1961; Berry 1964; Berry and Hepting 1964; Berry 

and Ripperton 1963; Hayes and Skelly 1977). 

Generally, exposure to ozone is considered not beneficial to tree growth; 

however, growth stimulation of both herbaceous and woody plants by low ozone 

concentrations (20 to 50 ppb) has been reported (Bennett and others 1974; 

Kress and Skelly 1982). Growth stimulation of herbaceous plants occurred when 

plants grown in low ozone concentrations were compared with those grown in 
carbon-filtered, ozone-free air. Bennett and others (1974) hypothesized that 

the apparent growth stimulation occurred because domestic herbaceous plants, 

such as the ones they used, were selected for low sensitivity to ambient ozone 

concentrations of 10 to 40 ppb. They were, therefore, not adapted to growing 

in an ozone-free atmosphere. Other studies also have observed that plants 

that are genetically adapted to growth in stressful environments frequently do 

not grow as well as when transferred to stress-free conditions (Roose and 

others 1982). 

Kress and Skelly (1982) reported apparent growth stimulation when 2-to4-week-old ye1 low-poplar ( Liriodendron tul ipifera L. ), sweetgum (Liquidambar 

styraciflua L.), sycamore (Platanus occidentalis L.) and sugar maple seedlings 

qrown in carbon-filtered, ozone-free air were compared with those exposed to 

50, 100, and 150 ppb of ozone 6 hours per day for 28 consecutive days. 

Significant stimulation of growth in height of sweet gum, yellow poplar, and 

sycamore was observed at 50 ppb, and of sugar maple at 100 ppb. Growth 

suppression, however, was observed for sweetgum and sycamore at 100 ppb (see 

below). Kress and Skelly, when analyzing their results, stated that the 

relationship between their data and ambient field conditions was tenuous. 

Growth suppression rather than stimulation is a much more common phenomenon resul ti ng from ozone exposure. Under control 1 ed conditions , Kress and 

Skelly (1982) observed significant suppression of height growth in 2- to 

4-week old tree seed1 ings exposed to ozone for 6 hours per day for 28 days. 

At 50 ppb: loblolly pine (Pinus -- taeda L.) showed 18 percent reduction, 

American sycamore, 9 percent; at 100 ppb: pitch pine (P. rigida Mil 1. ), 13 

percent; sweetgum, 29 percent; green ash (Fraxinus enns lvanica Marsh.), 24 

percent; and at 150 ppb: willow oak (Quercus phellos fr_ L. and sugar maple, 19 

and 25 percent, respectively. Other investigators have reported similar 

results for other tree species exposed to ozone under other regimes (e.g., 

Constantinidou and Kozlowski 1979; Dochinger and Townsend 1979; Kress and 

others 1982; Mooi 1980; Patton 1981; Pye 1988). 

Hogsett and others (1985), using exposures that simulated ambient conditions, noted a reduction in growth in height, diameter, and root length in two 

varieties of slash pine seedlings receiving chronic ozone exposure. The 

seed1 ings were exposed to one of three regimes: (1) charcoal-filtered air; 

(2) an exposure profile with a daily 1-hour maximum of 126 ppb at around 

2 p.m., a 7-hour (9 a.m. to 4 p.m.) seasonal mean of 104 ppb, and an integrated 

exposure of 155,000 ppb-hour (sum of hourly ppm >0); or (3) a similar ozone 

exposure profile but with a daily 1-hour maximum of 94 ppb, a 7-hour seasonal 

mean of 76 ppb, and an integrated exposure of 122,000 ppb-hours. Both 

varieties of slash pine exhibited an increasing reduction in growth with 

increasing ozone concentration. A signi f icant reduction (p <0.001) in stem 

diameter occurred by day 112 for both ozone treatments: 24 percent less than 

that of controls for "elliottii" and 30 percent less for "densa" at the lowest 

exposure; and 40 percent and 50 percent below control plants for "elliottii" 

d "densa," respectively, at the highest O exposure. Both 0 exposures also 

caused significant reductions in growth ii height (p <0.001?. The most 

pronounced change was observed in the growth of roots, which in "el 1 iotti it' 

was reduced 33 percent by day 21 at an integrated exposure of 29 ppm per 

hour, and 27 percent by day 56 with an exposure of 63 ppm per hour. 
Studies dealing with the effects of ozone on growth of trees in their 

natural habitats have focused mainly on the San Bernardino forest ecosystem in 

California, where the effects of air pollution on forests were first observed. 

Studies of air pollution effects on forests in the Eastern United States have 

developed more slowly and, until reports (Johnson and others 1984; Johnson and 

Siccama 1984; Siccama and others 1982) of forest decline in the Northeast 

began to appear, were concentrated i n the southern Appalachian Mountai ns, 

probably because air pollution injury of forests has not been a readily 

vi si ble phenomenon. 

Oxidant-induced injury of vegetation has been observed in the Appalachian 

Mountains for many years. The U.S. Forest Service in the 1950's studied the 

decline of eastern white pine in an area covering several hundred square miles 

on the Cumberland Plateau of East Tennessee, and concluded that unidentified 

atmospheric constituents were the cause (Berry and Hepting 1964). 

Needle blight (also called emergence tipburn) of eastern white pine was 

first reported in the early 1900's, but it was not until 1963 that it was shown 

by Berry and Ripperton (1963) to be the result of acute and chronic oxidant 

exposure. An unusually high incidence of emergence tipburn of eastern white 

pine occurred during the summer of 1961 in western North Carolina and 

surrounding States. During the same period, researchers at North Carol i na 

State University reported a very high incidence of weatherfleck of tobacco, a 

condition now known to result from ozone exposure (Berry 1961). 

Investigations to determine the cause of the tipburn included examination 

of roots of injured and uninjured trees. On injured trees, 72 percent of the 

primary roots had dead ends; on healthy trees, 28 percent (Berry 1961). 

Associated with the injury were oxidant concentrations of 65 ppb for the 

period from 1:30 to 3:00 p.m. on July 1, and from 1: 15 to 5:15 p.m. on July 2, 

1961; earlier, on June 27, 50 ppb had also been measured in Pocahantas County, 

WV (Berry 1964). Measurements were made the following year to compare 

concentrations between the valley and mountain top. The highest concentrations were recorded during the day in the valley and at night on the mountain 

top. On the night of June 28, oxidant peaks of 220 ppb were recorded at 

1: 15 a.m. , 120 ppb at 2: 30 a.m. , and 90 ppb at 8:45 a, m. For the valley, the 

highest concentrations (250 ppb) were observed between 12 noon and 5 p.m., 

during July 6 to 8, 1962. (Oxidant measurements were made with a Mast meter.) 

Despite these early reports of Berry (1961, 1964), Berry and Hepting 

(1964), and Berry and Ripperton (19631, no concerted efforts were made to 

determine the effects of ozone on the vegetation of the Appalachian Mountains 

until Hayes and Skelly (1977) began monitoring total oxidants and reported 

oxidant-associated needle injury of eastern white pine at three rural Virginia 

sites between April 1975 and March 1976. The injury was associated with 

oxidant peaks of 80 ppb or higher. Hayes and Skelly suggested that continued 

exposure of sensitive and intermediately sensitive white pine to acute and 

chronic oxidant concentrations could ultimately influence their vegetative 

vigor and reproductive ability. 

Needle i n jury symptoms and decreased growth of 25-year-ol d white pi ne 

growing on a plantation on the Cumberland Plateau near Oak Ridge, TN, were 

associated with ozone exposure (Mann and others 1980). To better understand 
the mechanisms I'nvolved in growth reduc$l"on of the trees, Mann and others 

(1980) and McLaughlin and others (1982) conducted studies of the seasonal 

physi ol ogi cal responses of white pi ne. Several ozone epi sodes with l- houraverage concentrations greater than 80 ppb were experienced between April 1 

and July 17, 1977. A peak of 150 ppb was reached on June 16. Concentrations 

during the previous year (19761, however, had not been as high. Concentrations 

exceeded 80 ppb for only 15 days, as compared with 40 in 1977. Sulfur dioxide 

(100 ppb) and NO (200 ppb) levels during the study period were below those 

concentrations k#own to cause vegetati anal in jury. 

In a subsequent study of the same trees, McLaughlin and others (1982) 

found reduced annual radial growth, which they also atbibuted to high concentrations of ozone. McLaughlin and others (1982) divided trees into three 

sensitivity classes-tolerant, intermediate, and sensitive- on the basis of 

needle color and length, and duration of retention. A steady decline in 

annual ring increment of sensitive whib pines was observed during the years 

1962 through 1979. Reductions of 70 percent in average annual growth of 

sensitive trees, when compared to the growth of tolerant and intermediate 

trees, were noted. Tolerant trees showed a consistently higher growth rate of 

5 to 15 percent (p = 0. 005) than intermediate bees for the 1960 to 1968 

interval , simi l ar growth from 1969 through 1975, and reduced growth of 5 to 

15 percent (but significant only at p = 0.10) for the period 1976 through 

1979, compared to trees of intermediate sensitivity. Needles of ozonesensitive trees were 15 to 45 percent shorter than those of either of the 

other classes. Decline was attributed primarily to chronic exposure to ozone, 

which frequently occurred at phytotoxic concentrations in the area. For the 

years 1975 through 1979, the incidence rates for hourly concentrations >80 ppb 

were: 1976, 190 hours; 1977, >339 hours; 1978, 190 hours; 1979, 125 Fours. 

Maximum 1-hour concentrations ranged from 120 to 300 ppb during this period. 

The pollutants sulfur dioxide and fluoride have been measured in the area, but 

the premature loss of needles and occasional tip necrosis of needles of the 

current year are manifestations usually associated with ozone, and no causeand-effect relaLionship with sulfur dioxide is indicated by the available 

data. 

In the studies discussed above (Mann and others 1980; McLaughlin and 

others 1982), decline in vigor and reduction in the growth of coniferous trees 

were usually associated with the following sequence of events and conditions: 

(1) premature senescence and 105s of older needles at the end of the growing 

season, probably leading to (2) reduced carbohydrate storage capacity in the 

fall and reduced resupply capacity in the spring to support new needle growth; 

and (3) increased reliance of new needles on self-support during growth. 

This resulted in (4) shorter new needles, lower gross phabsynthetic productivity and (5) a higher retention of current photosynthate by foliage, thus 

reducing the availability of photosynthate for external use (including repair 

sf chronically stressed tissues of older needles) (Mctaughljn and others 

1982). 

Benoit and others (1982) studied the annual ring increment of native 

eastern white pines sf reproducing age to evaluate the possible effects of 

oxidant air pollution on the long-term growth of forest species in a region of 

the Blue Ridge Mountains of Virginia, extending from the northern end of 

Sklyline Drive in Shenandoah National Park to the southernmod portion of the 
Blue Ridge Parkway lying in Virginia. The three white pines in each study 

plot were classified as sensitive, intermediate, or tolerant, based on a 

fo1 iar rating scale that incorporated needle length, needle retention by 

number of years, and the presence of typical ozone symptoms on needles. The 

mean ages of ozone-tot erant , intermediate, and ozone-sensi ti ve tree c1 asses 

were 53, 52, and 55 years, respectively. From ?955 to 1978, growth in mean 

annual radial increment for sensitive trees was 25 percent less, and for 

intermediate trees 15 percent less than that of tolerant trees, respectively, 

Only the 25 percent decrease for ozone-sensitive trees, however, was significant (p = 0.01; table 14). Smaller mean increments in the last 10 years 

compared to the previous 24 years indicated a trend toward decline in overall 

growth rates in all classes of trees. A comparison of growth from 1974 to 

1978 with that from 1955 to 1959 showed decreases of 26, 37, and 51 percent 

for tolerant, intermediate, and sensitive trees, respectively. 

The significant reduction in radial growth of ozone-sensitive white pines 

was assumed by the authors (Benoit and others 1982) to be the result of 

cumulative stress caused by the reduced photosynthetic capacity of oxidantinjured trees. Their assumption was based on the fact that no significant 

changes in seasonal precipitation had occurred between the 1955-1963 and 

1963-1978 periods. Increasing ozone concentrations, therefore, could account 

for growth reductions, particularly since for the latter period there was a 

negative correlation between precipitation and radial growth. Sulfur dioxide 

concentrations were too low (undetectable to 30 ppb) to have any vegetational 

effects. 

Duchelle and others (1982) reported that during a typical ozone season 

concentrations increased from January through June, stabilized during the 

summer months, and then decreased around October. During this time, ozone 

concentrations are low at night, begin to rise in the morning, peak in early 

afternoon, and then drop again at sunset. Episodes in the Blue Ridge Mountains 

lasting from 1 to 5 consecutive days have been reported by Skelly and others 

(1984) (table 15). Monitoring of ozone indicated the presence of monthly 

average concentrations of 50 to 60 ppb on a recurring basis with episodic 

peaks (1-hour) frequently in excess of 90 ppb (Duchelle and others 1982). 

Earlier, Hayes and Skelly (1977) suggested that the episodes in the area 

resulted from long-range transport of ozone from urban areas. This suggestion 

was supported by Fankhauser (1976), whose observations indicated that ozone 

was transported in a giant loop strekhing from New York City, Philadelphia, 

Baltimore and Washington, DC, though West Virginia and Ohio and then back to 

the Wheeling, WV-Pittsburgh, PA, area. Ozone was observed to follow this path 

for 4 to 5 days in September 1972. Earlier, in Nay 1972, another instance of 

ozone transport had been observed when a stagnant "high" and a slow-moving 

"l ow'' transported air from Chicago and Pittsburgh areas southward a1 ong the 

Appalachian Mountains do the Gulf of Mexico and then back Lo Miami, FL. 

The experiments of Duchelle and others (1982, 1983) exposing native tree 

seedlings and native vegetation in the Big Meadows area of the Shenandoah 

National Park in the Blue Ridge Mountains of Virginia to ambient ozone, indicate that ambient concentrations can reduce both the growth of native trees 

other than eastern white pine and the productivity of the native herbaceous 

vegetation found growing in the forested areas. 
Table 14. Annual Mean Radial Growth Increment, 1955 Through 1978,~ for 

Three Ozone Sensitivity Classess of Native Eastern White Pines (Pinus strobus 

L.) Growing in 10 Plots of Three Trees Each in the Blue Ridge Moutains in 

Vi rgi ni a 

(mm) 

Plot To1 erant ~rees~ Intermedi ate Trees Sensitive Trees 

a~hite pines rated tolerant, intermediate, or sensitive to O3 based on 

foliar symptoms. 

b~ensi tvity classes with the same letter are not significantly different 

at p = 0.01 based on Duncan's multiple range test. 

Source: Benoit and others (1982). 

Suppressed growth of seven indigenous forest tree species was observed 

when seed1 ings grown in open plots or ambient air open-top chambers were 

compared with those grown in charcoal-filtered air (Duchelle and others 1982). 

Growth reductions of wild-type seedlings of tulip poplar, green ash, sweet 

gum, black 1 ocust (Robinia pseudoacacia- L. ), eastern hemlock [(Tsuga canadensi s 

(L. ) Car. ) Table Mountain (Pinus pungens Lamb. ), Virginia (P. virginiana 

Mil 1. ) and pitch (P. Mi 11. ) pines usual ly occurred wi th&t expression 

of foliar symptoms. Open-top chambers were operated continuously in 1979 from 

May 9 until October 9, and in 1980 from April 24 until September 15. 

Injury to and growth reductions of native herbaceous vegetation growing 

in the Big Meadows area also were observed (Duchel le and others 1983). Few 

studies have reported the effects of ambient air pollutants on growth and 

productivity of native vegetation. Reduced growth and productivity of 

graminoid and forb vegetation were observed in experiments conducted at the 

same time and under the same exposure conditions as the tree seedling study 

cited above. Biomass production (weight of living tissue) was greatest in 

filtered open-top chambers when compared to those in open-air plots. Total 

3-year cumulative dry weight of plants in filtered chambers was significantly 

greater (p = <0.05) than that of plants in non-fi 1 tered and open-air plots. 

Reductions in biomass were attributed to a decrease in root growth, a result 
Table 15. Peak Hourly Ozone Concentration in Responses Recorded at Three 

Monitoring Sites in Western Virginia, Spring and Summer, 1979 Through 1982 

Date 

1- hr O3 concn. , ppb 

Big Meadows Rocky Knob Horton Center 

1979, April 11-12 

1979, June 5-6 

1979, September 12 

1980, May 29-30 

1980, June 13-15 

1980, July 14-15 

1980, July 31 

1981, May 24 

1981, June 23-23 

1981, June 29-30 

1981, July 8-11 

1982, May 11-13 

1982, May 16-17 

1982, July 24-26 

1982, July 27-28 

1982, August 3-4 

1982, August 19-20 

1982, October 1-3 

a~.~. = data not available. 

Source: Skelly and others (1984). 

of ozone exposure. Common mi 1 kweed ( Asclepias syriaca L. ) and common 

blackberry ( Rubus Porter) were the only two native species to 

develop visible injury. Milkweed previously had been reported to be sensitive 

to ozone and a good biological indicator (Duchelle and Skelly 1981). 

In the introduction to this section it was pointed out that perennial 

trees and shrubs must cope with the cumulative effects of both short- and 

long-term stresses, and that responses to these stresses may be subtle and not 

observable for many years. McLaughlin and others (1982) associated long-term 

(1962-1979) growth reductions of eastern white pine with ambient ozone concentrations by using data from increment cores. Benoit and others (1982) likewise 

used increment cores to relate growth reduction of eastern white pine in the 

Blue Ridge Mountains during the years 1955 to 1978 to ambient ozone concentrations. More recently, dendroecoloqical studies of dieback and decline of 

red spruce ( -- Picea rubens Sarg. ) in the Northeast (Johnson and Siccama 1983) 

and of reduced arowth rates of red swuce, balsam fir (Abies bal samea (L. ) 

Mill.), and ~raier fir [(A. fraseri (~ursh) Poir.)] in GKEiGl West ~irginia 

and western Virginia (Adams and others 1985) have provided further evidence 

that growth reductions measurable today probably began at least 20 years ago. 

Currently, there is no agreement as to the factor or factors that precipitated 

the dieback, mortality, and decline in growth, but multiple stresses including 

air pollution have been hypothesized (Adams and others 1985; Johnson and 

Si ccama 1983). 
The dendroecological studies of Phipps and Whiton (1988) also evidence 

the di f f i cul ty of determi ni ng the cause today of tree growth responses that 

began many years ago because of a major stress or stresses existing at that 

time. Growth decline, defined as actual growth less than that expected, was 

observed in 40 of 60 tree-ring collections judged to have been unaffected by 

local site histories, From 89 locations throughout much of the range (from 

Connecticut to North Carolina and west to ~owa) of white oak ( a1 ba 

L 1. Onset of decl ine began somewhere between the mid-1950's and the early 

1960'~~ and was essentially the same in the Northeast, Midwest, and the foutheast. The chronologies studied were of trees that had reached canopy dominance 

by 1850; some individuals were more than 200 years old at that time. A linear 

growth trend prior to 1950, as determined by examination of the basal area 

index (BAl), was observed in these dominant trees. A number of the individuals 

were cross-dated as having been in the canopy before 1600. Studies of the BAI 

of these individual s indicate no departure from the observed 1 i near trend 

prior to 1950. 

The onset of growth decline, though rather abrupt, did not appear in all 

trees at the same time, suggesting that some trees were more tolerant of the 

stress than others. Some trees showed no decline at all. Changes in climate 

did not appear to mimic growth trends except in North Carolina where a temperature maximum was observed around 1950. Estimations of 1 ong-term regional 

sulfur and nitrogen trends vary for the different regions (Phipps and Whiton 

1988). Trends in the Southeast generally show a marked increase in sulfur 

emissions since 1950, and those in the Midwest appear to have been increasing 

since the 1920's. In the Northeast, trends in sulfur emissions have stabilized or decreased somewhat since the 1970's while nitrogen emissions have 

increased. Only in the Southeast have emission increases correlated with the 

decline in white oak growth; however, trees that did not exhibit a decrease in 

growth in the 1950's still do not exhibit a decrease even though emissions 

have increased. The authors hypothesize that growth decl i ne was initiated 

during a period beginning in the mid-1950's to the early 1960's by conditions 

or events as yet unidentified, and that subsequent growth has not been in 

response to tree age, geographic location, site quality, climate trends, or 

regional emissions of sulfur and nitrogen oxides (Phipps and Whiton 1988). 

Phipps and Whiton (1988) rule out sulfur and nitrogen emissions as 

initiators of the growth decline observed in white oak since the 1950s. 

Could ozone have been an initiator of the decline? The authors do not speculate. It should be noted, however, that regular oxidant-monitoring stations 

were not established east of the Mississippi River until 1962 (U.S. Department 

of Health, Education, and Welfare 11970). Maximum oxidant concentrations 

recorded between 1964 and 11967 indicate that Cincinnati had 18 days above 

150 ppb and 55 above 100 ppb; Philadelphia 13 days above 150 ppb and 60 above 

100 pph; and WashingLan, DC, 7 days above 150 ppb and 65 above 100 ppb, Based 

on these data, ozone concentrations sufficient to injure vegetation were 

already present in the &masphere from Chicago to Phjladelphia and Washington, 

DC, by the middle 1960'5. The observations of Berry (1961) and Berry and 

Ripperton (1963) mentioning oxidant concentrations above 100 ppb during 1961 

and 1962 in West Virginia and North Carolina indicate that the area with high 

ozone concentrations also extended southward. Oxidant exposure at this time 

was just being established as a probable cause of the foliar injury that had 

been observed for many years on sensitive eastern white pine (Berry and 

Ripperton 1963). 
d) Mixtures of Gaseous Pollutants -- 

The possibility that exposure to mixtures of gases could have an effect 

on vegetation greater than a single gas has been a concern since the 1960's. 

A majority of the studies conducted since that tine has focused on the effects 

of pollutant mixtures on agricultural crops, and the S02/03 combination has 

received the greatest attention. 

Dochinger and others (1970) suggested that 0 /SO acting synergistically 

could cause needle mottling of white pine. ~arnossy fi976) observed more than 

additive effects when he exposed trembling aspen ( tremuloides Michx.) 

to combinations of 03/S0 He also observed clonal variation in the response 

of the different clones t% the gas combinations. Clones shown to be sensitive 

to SO or O3 alone were more sensitive to the combinations of the two. These 

assesgments were based on leaf injury rather than growth response. Both of 

the above studies were conducted in chambers. 

Miller an (1984) exposed Jeffrey (Pinus Grev. & Balf,), 

ponderosa ( P. Doug1.e~ Laws), and Digger pi abiniana Dougl.) 

to mixtures?f Shoot growth was not as sensitive as root growth. 

Root growth was m6re3affected by 03/S02 in combination than by either pol lutant a1 one. 

The co-occurrence of SO and NO is most frequently measured in the 

vicinity of power plants (Lehhn and fingey 1984). Ambient concentrations of 

NO rarely approach the injury threshold. Concern regarding its presence in 

th$ atmosphere is related to its potential for interaction with other pol lutants such as SO as demonstrated by Tingey and others (1971). The majority of 

the experiments gonducted to determine the effects of SO2 + NO2 in combination 

have used agri cul tural and horti cul tural crops. 

In England, Whitmore and Freer-Smith (1982) and Freer-Smith (1984) 

observed that combinations of SO + NO affected cuttings of a number of 

British tree species. Second yea? cuttqngs were exposed to SO 

from March through August. The SO combination (62 + 62 

interactive effects resulting in gr han additive responses. Significant 

reductions in stem height and diameter and dry weight resulted from the 

exposures (Whitmore and Freer-Smith 1982). Freer-Smith (1984) also reported 

that decreases in growth produced by SO2 + NO2 were greater than by either 

pol 1 utant a1 one. 

Stone and Skelly (1974) and Phillips and others (1977a,b) measured growth 

responses in the field to emissions from the Radford Army Ammunition Plant in 

Radford, VA. Both SO and NO2 were present in concentrations sufficient to 

i nf 1 uence the growth f white pine, yellow poplar, and 1~blo11y pines. No 

attempt was made at this time to see whether the pollutant combination potentiated the effect of either pollutant alone; however, synergism was suggested. 

The O and NO combination has been studied infrequently. Kress and 

Skelly (1932) expozed seedlings of 10 forest tree species to O and NO 

concentrations of 100 ppb. The only significant interaction efdects no&d 

between the two polluLants were less than additive. 
The growth impact of 0 , and/or SO2 on lob101 y pine was studied by 

Kress and others (1982). Aidd NO to the O3 and SO combination had a 

slightly stimulatory effect o owl%$ however, the "s nsitive" seedling of 

the full-sib family suffered a 30 percent height suppression when exposed to 

the three-pollutant combination as compared with a 14-percent suppression for 

the resistant family. 

Summary. Evidence that exposure to sulfur oxides and ozone can decrease 

the growth of trees is well established by the many experimental studies that 

have been conducted during the past 20 years. Evidence also is strong that 

the growth of trees in their natural habitats also can be inhibited by 

exposure to these two pollutants. Sulfur dioxide inhibits growth and causes 

decline and mortality of trees in the vicinity of point sources. Ozone also 

inhibits growth and causes decline and mortality in areas where episodic 

concentrations (frequently transported from metropolitan areas some miles 

away) last for several days and reach 80 ppb or higher. Data indicating 

nitrogen oxide injury of trees growing in their natural habitats are lacking. 

The concentrations used to cause injury experimentally have not been observed 

in the field. 

One of the biggest problems facing scientists at the present time is 

determining what stress or stressors precipitated tree injury or growth 

responses visible or measurable today when the exposure took place a number of 

years ago. Also, though sulfur dioxide and ozone do inhibit tree growth, the 

exact manner in which they alter tree biochemistry is as yet unknown, and as 

yet it is not possible to establish an exposure-response relationship. 

4.4 What is the Likelihood That Airborne Chemical Pollutants Affect Forest 

Ecosystems Through Soil-Mediated Response Mechanisms? 

Soils are one of the most stable components of ecosystems, and together 

with climate determine the limits of ecosystem productivity. Soils are a 

mixture of inorganic minerals, organic matter, soil animals, microorganisms, 

gases, and water. They function as the anchor for most terrestrial and many 

aquatic plants. They provide and store water, nutrients, and oxygen needed by 

plant roots for growth. Chemical transformations and animal and microorgani smal activity within soi 1 s i nf 1 uence the avai 1 abi 1 i ty of nutrients to 

plants. Should airborne chemical pol 1 utants inhibit or a1 ter chemical 

processes, or influence animal and/or microorganismal behavior in soils, the 

growth of the above-ground flora could be impacted. 

Exposure of plants to airborne gases, fine aerosols, or coarse particles 

through the soil is indirect. To have an effect, these chemical compounds 

must in some way influence the movement or uptake of plant nutrients. Four 

separate processes are involved in exposure: 1) deposition, 2) dissolution 

into soil water, 3) percolation or diffusion through the soil, and 4) absorption from the soil through the surfaces of fine feeder roots and mycorrhizae. 

Many biotic, physical and chemical factors influence these processes. 

The general consensus of most investigators at the present time is that 

there are no apparent drastic or potentially catastrophic short-term effects 

from the atmospheric deposition of ambient gaseous air pollutants on soils. 

There is a lack of information, however, concerning either temporary or 
permanent effects of long-term pollutant exposures on soils (Weidensaul and 

McClenahen 1986). Most of the available information concerning pollutantsoil interactions deals with soils as a sink. For this reason, the efhcts 

of deposition and fate of gaseous airborne chemical pollutants on soi 1s are 

discussed in a very general manner. 

The effects of acidic deposition on soil are of greatest concern. A 

number of publications have reviewed this topic (Altshuller and Linthurst 

1984; Bennett and others 1985; Binkley and others 1987; Brandt 1987). These 

publications discuss in some detail the possible impact on soils of acidic 

deposition and the resulting effects on vegetation changes. The fol lowing 

section deals with the relative likelihood of some of the major hypothetical 

mechanisms by which soil changes could affect forest plants. 

4.4.1 Airborne Gases, Fine Aerosols, and Coarse Particfes 

Soils generally sorb organic gases faster and in greater amounts as their 

molecular weights increase. Absorption rates al so increase when functional 

groups such as nitrogen, oxygen, and sulfur are attached to the organic gases. 

Uptake of organic gases usually depends on the development of an appropriate 

microorganismal population, while removal of inorganic gases is attributed to 

chemical and/or physical processes (Bohn 1972). 

Most gaseous SO is removed by soil within 15 minutes, but the same 

amount of NO requirzs approximately 24 hours. The sorption rate for each of 

the compound$ is similar in sterile and nonsteri le soi ls, suggesting that most 

and NO is removed by some chemical reaction (Abeles and others 1971). 

magnitzde of SO2 and NO sorption is influenced by soil type. Soils not 

only sorb NO gases, but a1 o release nitrogen gases such as N2, N20, O, NO, 

and some ~0~ 'Lia chemical reactions (Bohn 1972). 

Nitrates and sulfates diffuse rapidly into the soil. Length of nitrate 

retention depends on their assimilation by pl2nts and microorganisms, as they 

are readily leached. The ammonium ion (NH4 ) is frequently attached to 

nitrates or sulfates. Nitrates, sulfates, and ammonium when in ionic form in 

the soil are readily - taken up by tree roots (Donahue and others 1977). 

Retention of SO depends on soil pH and the presence of hydrous oxides, iron, 

and aluminum (~%rahamsen and Dol lard 1979). Acidic nitrates and sulfates as 

aerosols or in rain tend to acidify soil. The nutrient roles of sulfur and 

nitrogen are mentioned in section 4.3.1; the effects resulting from deposition 

in acid aerosols are discussed below. 

Soils have developed a nutrient capacity during their formation from 

parent materials over time. Trees and other vegetation growing on these soils 

are adapted to them. Forests, therefore, are not nutrient limited, except for 

a lumberman's point of view. Today, as a result of atmospheric pollution many 

coniferous forest soils are receiving nitrogen in excess of the growth requirements of the trees. Appearance of nitrate in the soil solution is an early 

symptom of excess nitrogen in these systems. A second symptom, because of the 

inability of roots of the densely growing conifers to metabolize large amounts 

of nitrate, is the presence of nitrate in the xylem sap (Waring 1987). 

Nitrogen uptake by trees may not necessarily enhance growth, particularly 

in boreal and subalpine conifers. The photosynthetic capacity of conifer 
foliage is low and not greatly enhanced by increasing the nitrogen content. 

In addition, excess nitrogen not only alters the way in which carbon is 

allocated between shoots and roots, but also i s i n a form difficult to metabo1 ize, and increases sensitivity to frost. The capacity of gymnosperms in 

general, and boreal and subalpine species in particular, to synthesize the 

enzymes required to reduce nitrate in foliage or roots appears to be limited. 

A reduced allocation of carbohydrates to the roots is associated with the 

accumulation of amino acids in foliage (Waring 1987). 

Excess nitrogen in subalpine or boreal ecosystems is 1 i kely to cause 

nutrient imbalances and make conifers more susceptible to drought, other 

pollutants, and pathogens. Competition under these circumstances would favor 

deciduous species (Waring 1987). Whether nitrate deposition in the Southeastern United States is capable of causing the phenomena presented above has 

not been assessed at this time. 

4.4.2 Acidic Deposition 

Acidification can strongly influence soil dynamics by altering the soil's 

biological, chemical, and physical behavior. Changes in soil structure, soil 

aeration, water status, and the environment of soil organisms can, in turn, 

influence the avai labil i ty of nutrients and toxic substances. Soi 1 acidification can result from natural sources of hydrogen ion in soils as well as 

from the atmospheric deposition of acidifying substances (Brandt 1987). 

Sulfate and nitrate deposition are expected to have a greater influence on 

most soils than hydrogen ion deposition (Bennett and others 1985). If the 

sulfate and nitrate in acidic deposition are not assimilated by plants and 

microorganisms, the addition of the hydrogen ion could increase soil acidity 

and lower soi 1 pH (Binkley and others 1989). 

Effects of soil acidification in general may be outlined as follows 

(Brandt 1987) : 

* Natural soil development in humid regions includes base leaching and acidification. Acidic precipitation can intensify 

these processes; 

* The extent to which soils can be acidified depends on the 

initial pH and buffering capacity of the soils. Soils 

naturally low in basic cations, with low pH and limited cation 

exchange capacity, may be particularly sensitive; 

. Soi 1 acidification is detrimental to many microbiolgical 

processes in soils, for example the mineralization and 

humi fication of organic matter and the mobi 1 ization and fixation of nutrients and toxic substances (e.g., toxic metallic ion 

concentrations) ; 

* Use of fertilizers, especially nitrogen fertilizers with 

physiological acidic reactions, as well as other cultivation 

and land-use practices can increase soil acidification. 
The soil factors mast likely to be affected by acidic deposition are 

(1) an increase in sol ubil i ty of toxic metals such as aluminum and manganese 

so that they affect roots or mycorrhizae; (2) an alteration of nutrient supply 

through leaching or by altering microbial populations or processes (toxicity 

of mycorrhizae due to increased nitrogen has been postula-ted); and (3) a 

reduction in the solubility of phosphorus and molybdenum. Acidic deposition 

can increase natural acidification processes; however, at current levels the 

contributions are cansidered to be minor; (A1"cshu'ller and Linthurst 1984; 

Binkley and others 1987). The likelihood of increased acidification of soils 

occurring is discussed in the paragraphs that follow. 

Soi l acidification is the primary condition control l ing metal mobi l ization. Acid soils often contain increased concentrations of aluminum in the 

soil solution. A substantial body of knowledge has accumulated over the years 

concerning the toxicity of aluminum to agricultural crops. Many plants are 

unable to to1 erate high al umi num concentrati ons because they impair root 

activity. Excess aluminum in highly acid soils causes root necrosis and 

reduced root growth; these in turn increase plant susceptibility to moisture 

and/or nutrient stress (Black 1968; Foy and others 1978; Schier 1985). Many 

tree species, however, grow on very acidic soils and appear to tolerate the 

high aluminum concentrations in soi 1 solutions. Forest soi l s are commonly 

acid with pH values of 3.5 to 5.5. Tree species may become intolerant if high 

aluminum to nutrient ratios limit root uptake of calcium and magnesium and 

create a nutrient deficiency (Bi nkl ey and others 1989). 

Shortl e and Smith (1988) attributed al umi num-i nduced cal ci um deficiency 

syndrome in declining red spruce to restricted calcium uptake by trees growing 

in highly acid soils. Though soluble aluminum in high concentrations is toxic 

to the roots of a wide range of agricultural and Parest plants, there are 

healthy stands of trees on soils of similar acidity and with similar aluminum 

concentrations both in the soil and fine roots. In addition to toxicity, 

aluminum interferes with calcium uptake by fine roots when the soluble ions 

are in equimolar concentrations. Shortle and Smith propose that inhibition of 

calcium uptake causes decline in cambial growth and a loss of crown in mature 

red spruce. 

The demand for calcium per unit area of cambium surface is essentially 

constant and the surface area expands exponentially as mature trees add 

secondavy xylem. Restricted uptake of calcium suppresses cambial growth, and 

a decrease in the widths of annual rings ~esults. This reduces the amount of 

functioning sapwood since the older sapwood is continuously transformed into 

heartwood. Both decl ining and heal thy trees have the same number of annual 

rings, but in healthy trees the rings are wider. Trees become susceptible to 

death from secondary pathogens and insects when the amount of sapwood forms 

less than 25 percent of the cross-sectional area. Crown density and associated 

leaf mass and area are positively correlated with the cross-sectional area of 

sapwood- As the ecress-sections1 area of sapwood decreases, crown densfty 

also would be expected to decrease (Shortle and Smith 1988). 

A comparison by Shortle and Smith (1988) of trees growing on Mt. Abraham, 

VT, where more than half the canopy spruce are dead, with trees growing near 

Beddi ngton, ME, where canopy spruce are uni formly heal thy, supports thei r 

theory of aluminum-to-calcium imbalance. The humus layer for a1 l stands 
investigated had a pH of 3.2 to 4.6. Equilibrium soil solution analysis 

indicated that the aluminum-to-calcium ratio was greater than 1 on Nt. Abraham 

and was significantly higher than in the soil solution from Beddington. 

Spruce trees at high elevations on Mt. Abraham (1000 m) have been most 

affected because the aluminum-to-calcium ratio is greater than I in both humus 

and subsoil, and limits calcium supply. Continued input of sulfur and 

nitrogen anions will increase the amounts of aluminum in solution, and the 

increased aluminum will reduce calcium uptake. Aluminum toxicity is not the 

problem under these circumstances (Shortle and Smith 1988). 

At present it is not known whether acidic deposition is inducing nutrient 

leaching in the forest soils of eastern North America because: a) the nutrient status of eastern forests is not characterized well enough to make such a 

determination, and b) in cases where nutrient leaching is occurring, it is not 

known whether the amounts being leached are greater than would be expected in 

the absence of ambient acidic deposition. Turner and others (1986) estimated 

that base-cation leaching would result in growth reductions in 51 percent of 

forested soils in the Southeast. The Piedmont and Coastal Plain were judged 

to have the largest proportion of areas sensitive to acidic deposition (Turner 

and others 1986). Some have questioned the data used to establish sensitivity 

by Turner and associates. Binkley and others (1989), however, state that the 

data base used by Turner and co-workers is probably the "most extensive and 

qua1 ity-assured" available. 

Turner and others (1986) estimated that it would take approximately 

30 years at current rates of acidic deposition to decrease soil base saturation of the most sensitive forest soils from 60 to 40 percent, and 80 years 

from 60 to 25 percent. Replenishment of base cations by deposition was not 

considered. 

Tabatabai (1985) predicts that at current rates nutrient leaching induced 

by atmospheric deposition is unlikely in most eastern soils because: (1) sulfate adsorption capacities are relatively large; (2) cation exchange capacities 

are large; (3) most forests are nitrogen 1 imited and most nitrate will be 

taken up by forest vegetation; and (4) soil base cations are being replenished 

by atmospheric deposition and mi neral weatheri ng. 

Field studies of base cation and anion leaching in eastern forests 

indicate that the amounts of these ions are insignificant in soil leachate of 

most forest watersheds studied (Walker Branch and Camp Branch, TN; Hubbard 

Brook, NH; and Huntington Forest, NY) (Bormann 1979; Johnson 1983; Johnson 

and others 1985a; Johnson and others 1985b; Mollitor and Raynal 1982). 

Binkley and others (1989) report that to date, forests in the South have 

responded most strongly to additions of phosphorus and nitrogen, probably 

because growth of most forest stands in the South is phosphorus and nitrogen 

limited. There is, however, some evidence that deficiencies in other elements 

have also 1 imited growth. Data on phosphate and molybdenum availability in 

acidified soils are limited; however, molybdenum deficiency in southern 

forests has not been reported. Potassium deficiencies may be common on sandy 

soils. 
Though soil acidification does affect many biological processes, at 

present it is difficult to distinguish between direct effects of acidification 

and secondary effects resulting from acid-induced changes in the soil solution, Important effects under field conditions have not been demonstrated 

(A1 tshul ler and Linthurst 1984). The possible effects of nitrogen toxicity 

to mycorrhizae have been reviewed (Cline and others 19887). The authors 

concluded that at the current rate of nitrogen deposition in the southeastern 

United States, nitrogen is not toxic to mycorrhizae. 

.--The consensus of most investigators at the present time is that 

no apparent drastic or potentially catastrophic short-term effects result from 

the atmospheric deposition of gaseous air pollutant chemicals on soils. In 

general, the long-term effects are unknown. Concern exists chiefly for the 

possible long-term effects of acidic deposition. Most studies have emphasized 

changes in soil pH resulting from deposition of acidic or acidifying 

substances. Deposition of sulfate and nitrate may pose a greater problem. 

Data suggest that nitrogen in excess of the amounts required for growth 

of trees is accumulating in coniferous forests. Appearance of nitrate in the 

soil solution and in the xylem sap of conifers are symptoms of this problem. 

Nitrate appears in the xylem sap because the roots of densely growing conifers 

are unable to metabolize large amounts. In addition, excess nitrogen in 

foliage of boreal and subalpine conifers does not enhance photosynthesis. It 

not only alters the manner in which carbon is allocated between shoots and 

roots but also provides nitrogen in a form difficult to metabolize and one 

that increases sensitivity to frost. Resulting nutrient imbalances are likely 

to make conifers more susceptible to drought, other pollutants, and pathogens. 

Competition under these circumstances favors deciduous species. 

Soil acidification is the primary condition controlling metal mobilization. An increase in the concentrations of metals such as aluminum and 

manganese toxic to roots and mycorrhizae of has been postulated as a possible 

result of acidic deposition. While many plants are sensitive to high aluminum 

concentrations in soil solutions, healthy stands of trees do grow on very 

acidic soils (pH 3.5 to 5.5) and appear to tolerate the high aluminum concentrations. Aluminum, on the other hand, also interferes with calcium uptake by 

fine roots when soluble ions are in equimolar concentrations. Calcium 

deficiency results in reduced cambial growth, reduced formation of new sapwood, 

and a decrease in functioning sapwood as the older sapwood is transformed into 

heartwood. Both declining and healthy trees have the same number of annual 

rings, but in healthy trees the rings are wider, When sapwood forms less than 

25 percent of the cross-sectional area, trees become susceptible to death from 

secondary pathogens and insects. 

At the present time the effects of acidic deposition on nutrient leaching 

are not known. Tabatabai (1985) predicts that at the current rates atmospheric 

deposition is unlikely to induce nutrient leaching in most eastern soils. 

Field studies of base cation and anion leaching in eastern forest suggest that 

the amounts of these ions in soil leachates for most forest watersheds are 

insignificant. 

Current data suggest that the forests in the South have responded most 

strongly to additions of phosphorus and nitrogen. There is, however, some 

evidence that other nutrients have also limited growth. Aqueous aluminum may 

which can be either severe episodic stresses (1 day to a week of acute high 

exposures) or long-term chronic stresses, are superimposed upon the naturally 

occurring stresses already being encountered by the trees. They may accelerate 

the processes of change already under way. The effects of the natural 

stresses upon ecosystems, unless they are catastrophic (e.g., fire, flood, or 

windstorm), are frequently difficult to determine. 

Causation for ecosystem breakdown in eastern North America has been 

definitely established only around point sources of pollution. Emissions of 

from smelters and an iron sintering plant resulted in trees and shrubs 

ng eliminated from the vicinity of Copper Hill Basin near Ducktown, TN 

(Hedgcock 19141, and from around Wawa and Sudbury, Ontario, Canada (Gordon 

and Gorham 1963; Linzon 1978). The pollutant dosage in the above situations 

was catastrophic, resulting in long-term injury to the ecosystems. 

Responses of forest communities or ecosystems to catastrophic or acute 

disturbances are, as a rule, readily observable and measurable. Catastrophic 

disturbances cause mature forests to be returned to an earlier successional 

stage. During normal succession in unpolluted atmospheres, the number of 

species in a community usually increases. Productivity, biomass, community 

height, and structural complexity also increase. Stress alters svccessicn and 

diverts energy from growth and reproduction to maintenance. Biomass accumulation and production decrease and there is a reduction in structural complexity, 

species diversity, environmental modification, and nutrient control. Succession reverts to an earlier stage. 

The responses of forest ecosystems to air pollution are integrated 

responses because the air pollution stress is superimposed upon and interacts 

with the multiple stresses the component species are already experiencing. 

The composition and successional stage of forests are determined by the kind, 

frequency, and magnitude of disturbance (table 16) (Kozlowski 1985). 

Studies of radiation effects (Woodwell 1970) and data from air pollution 

effects around strong point sources (see above), as we1 l as studies and 

observations of ozone effects on natural ecosystems such as the San Bernardino 

Forest, suggest that forest ecosystems respond to increasing air pollution 

stresses in a predictable pattern that may be thought of as a continuum of 

responses (tab1 e 16) (Bormann 1985; Guderi an and Kueppers 1980 ; Kozl ows ki 

1985; Smith 1981; Woodwell 1970). Circumstantial evidence for forest change 

may be obtained by comparing field data gleaned from studies of ambient 

pollution effects on individual trees, surveys to determine visible injury 

symptoms on forest trees and vegetation in the field, and studies of soil 

changes with the continuum of responses (table 16) of Bormann (1985). 

The responses of decline and dieback seen at the present time in eastern 

North America are the result of chronic stresses. Responses of forest 

ecosystems to chronic stresses are diffuse and, therefore, less readily observed 

and measured. Subtle and indirect effects of pollutant dosages on individual 

trees and other perennial species are cumulative. Sensitive annual species 

respond during the year they are exposed. Cumulative pollutant stress provides 

a selective force that eliminates those species that lack the genetic diversity 

to survive, suppresses some genotypes, and favors others. The distribution of 

a species and the productivity of an ecosystem both depend upon the success or 
failure of individual plants (Billings 1978). When removal of an individual 

species alters species interactions and changes population dynamics, then 

energy flow and chemical nutrient cycling of a community or ecosystem also are 

altered, thus setting the stage for changes in community structure that possibly could be irreversible (table 16) (Guderian and Kueppers 1980). The 

problem, once the changes have occurred or are occurring, is to discover what 

caused the changes. Because of lack of data (monitoring data in particular), 

we are left with hypotheses (table 17). 

Surveys made in 1982 give quantitative evidence of spruce decline in 

stands throughout the Appalachian Mountains from New Hampshire and Vermont to 

North Carolina (Johnson and Siccama 1983). Red spruce (Picea -- rubens Sarg. ) 

is the most characteristic species of the subalpine coniferous forests that 

occupy the higher peaks and ridges of the Appalachian Mountains from Maine to 

North Carolina and Tennessee. In the North, balsam fir [(Abies balsamea (L.) 

Mil 1. )] is commonly present as a codominant species, whereas Fraser fir [(A. 

fraseri (Pursh) Poir.)], a closely related species, is the codominant in the 

Adams and others 1985). 

In the high elevation forests of New York, Vermont, and New Hampshire, 

red spruce has shown marked dieback and a large reduction in basal area and 

density. Spruce mortality has occurred rather evenly regardless of tree size, 

variety of stand composition, and growth on a variety of soils. Red spruce 

dieback was observed in substantial areas of Vermont and New Hampshire in the 

1950's and 1960's. The areas were studied to determine the cause. No primary 

pathogens or insect infestation were observed on the trees. Fungal pathogens 

present (Polyporus boreal is, Fomes * ' , and Armil laria me1 lea) were those 

whose invasion is usually triggered by predisposing agents. Data from longterm (1780 to 1965) tree ri ng records indicated 1 arge- f 1 uctuati ons i n annual 

growth, but no consistent pattern of declining growth (Johnson and Siccama 

1983). 

The red spruce dieback is characteristic of a stress-related disease. 

These characteristics include the absence of an obvious cause, dieback (loss 

of foliage beginning in the crown and progressing over time downward and from 

branch tips inward), reduced vigor, and subsequent invasion by secondary 

organisms that normally do not cause substantial injury to vigorous trees (see 

table 16, 11-3) (Johnson and Siccama 1983; Siccama and others 1982). Spruce 

regeneration is reported to be severely depressed. 

Hardwood forests also show quantitative decline. There is a reduction in 

mosses and mycorrhizae, and the natural litter decomposition process has been 

affected. Seedlings have malformed roots, and in the spring when plant growth 

is most vigorous, the highly acidic soi 1 nutrient solution contains toxic 

concentrations of metals (Klein 1985). Many hypotheses have been suggested 

to explain these effects (table 17). The multiple stress hypothesis is most 

current (Johnson and Siccama 1983; Klein and Perkins 1987). 

In the Southeast, the decline and mortality of the Fraser fir in the 

Great Smoky Mountains National Park, Plott Balsam Mountains, NC, and the 

Black Mountains, which include Mt. Mitchell, have been attributed to infestation by the balsam woolly adelgid (Hain and Arthur 1985). Results from a 

20-year experimental study of Fraser fir growing in the Smoky Mountains 
Table 16. Ecosystem Response to Pollutant Stress 

Stage of Response of Vegetation Response of Ecosystem 

Conti nuum 

O Anthropogenic pollutants 

insignificant 

I Pollutant concentrations low; 

no measurable physiological 

response 

I1 Pollutant concentrations injurious 

to sensitive species 

(1) Reduced photosynthesis, 

altered carbon allocation, 

reduced growth and vigor 

(2) Reduced reproduction 

(3) Predisposition to entomological or microbiological 

stress 

I11 Severe pollution stress. 

Large plants of sensitive 

species die. Forest layers are 

peeled off: first trees, tall 

shrubs and, under most severe 

conditions, short shrubs and 

herbs. 

Unaffected; systems pristine 

Ecosystem functions unaffected; 

pollutants transferred from 

atmosphere to organic or 

available nutrient compartments 

Altered species composition; 

populations of sensitive species 

decline; same individuals are lost. 

Their effectiveness as functional 

ecosystem members diminishes; they 

could be lost from the system. 

Ecosystem reverts to an earlier 

stage. 

(1) Simplification, basic ecosystem structure changes, 

becomes dominated by weedy 

species not previously 

present. 

(2) Reduced stability and productivity; loss of capability for 

repairing itself. Runoff increases, nutrient loss, and 

erosion accelerates; a barren 

zone results. Ecosystem 

col 1 apses. 

Source: Adapted from Bormann (1985) ; Kozl ows ki (1985) ; Smith (1974) ; 

National Park indicate that the balsam woolly adelgid had killed almost all of 

the canopy trees and had reduced basal area in two plots established in the 

1960's. Spruce basal area remained about the same for the same period. The 

report does not mention monitoring of atmospheric pollutants or possible 

pollutant-insect interaction or predisposition (Busing and others 1988). 

Recent studies on Mt. Mitchell, however, do not attribute the death of Fraser 

fir solely to the balsam woolly adelgid, but suggest that atmospheric deposition and multiple pollutant stresses also have a role in tree mortality (Bruck 

and others 1988; Hain and Arthur 1985). The studies also cited exposure to 

gaseous air pollutants, particularly ozone, and cloud water deposition of 

acidic substances as among the possible stresses having increased host 

susceptibility to attack by the balsam woolly adelgid (Bruck and others 1988; 

Hain and Arthur 1985). 
fable 17. Some Hypotheses to Explain Forest Decline in Eastern North America 

Catego~y Wypo"clres-i"s 

MaLural process Synchronous canopy dieback 

Biotic pathogens 

Assorted abiotic stress agents 

Forest stand history Landscape conversion 

Forest management practices (e.g., genetic stock) 

Fcrest fire frequency and intensity 

Antkropogenic stress Nitrogen subsidy (e.g. , nitric acid vapor, 

nitrate deposition) 

Direct ozone effects (e. g. , inhibition of 

photosynthesis) 

Indirect ozone effects (e. g. , leaching) 

Direct effects of hydrogen ion deposition 

Indirect effect of hydrogen ion deposition 

(e. g. , rhizosphere a1 urninurn toxicity) 

Deposition sf sulphur 

Deposition of organics 

Deposition of muleipk pollutants 

(wed and dry deposition) 

Deposition of trace elements 

Acid cloud moisture 

Multiple factors Site-specific combination of assorted natural 

processes, forest stand history, and anthropogenic stress 

Adapted from Taylor and Narby (1985)- 

Sol" 1 conditions resul tjng From the depositjon of airborne sul fur- and 

nitrogen-derived chemicals have been hypothesized as having been involved in 

the dieback and decline of red spruce and responsible for the slowdown in 

growth of other trees (sec. 4.4). The deposition of nitrogen in excess of 

amounts required for coniferous tree growth as well as the imbalance of 

aluminum and calcium in the fine roots of trees have recently been proposed 

as the cause. Nitrogen in excess of growth requirements can be detrimental to 

conifers. In eight regions where forest decline has been observed, the 

nitrate concentration in the soil was an order of magnitude higher than in 

areas where forest decline has not yet been observed (Waring 1987). Enrichment of the soil with atmospheric nitrogen in excess of growth requirements of 

boreal and subalpine conifers will not necessarily enhance growth, but is 

l i kely to cause nutrient imbalances and make conifers more susceptible to 

drought, other pol 1 utants, and pathogens (Wari ng 1987). An imbalance of 

calcium and aluminram in the Pine roots of "trees reduces the rate of wood 

formation, decreases the amount of functional sapwood and live crown, and 

results in large trees being more vulnerable to secondary diseases and insect 

pests (Shsrtle and Smith 1988)- 
In the preceding sections (4.2.1 to 4.2.7), the authors have discussed 

both field observations and experimental data dealing with the effects of 

acidic deposition, sulfur dioxide, nitrogen oxides, ozone, and gaseous pollutant mixtures on individual trees and seedlings via foliage-mediated response 

mechanisms. Exposure of trees to sulfur dioxide and ozone in their natural 

habitats has been shown to inhibit photosynthesis; alter carbon allocation; 

decrease vigor; disrupt reproduction; change host susceptibility to insects, 

pathogens, and mycorrhizal colonization; and, in the final analysis, affect 

tree growth. Evidence that nitrogen oxides or gaseous pollutant mixtures 

cause growth reductions under ambient conditions is lacking. 

Sulfur dioxide is emitted from point sources. In the eastern United 

States, 70 percent of the emissions are from power plants with tall slacks, In 

the summer, residence time in the Midwestern States is approximately 2 days. 

During that time, the plumes from the power plants are likely to have been 

transported up to 1000 km, broadened, and the SO to have been transformed 

into sulfate or to have been removed by dry or 6et deposition (Husar and 

Holloway 1983; sec. 3.4). The highest average SO concentrations at rural 

sites in eastern North America are in the Ohio ~i%er Valley with values of 

10 ppb. Average concentrations decrease with increasing distance from this 

area, reaching approximately 2 ppb in New England and 1 ppb in the South 

(Mueller and Hidy 1983). 

Ozone, on the other hand, is a regional gaseous air pollutant (see 

chapter 3. ). It affects vegetation throughout the United States, impairing 

craps, native vegetation, and ecosystems more than any other air pollutant 

(Dai nes and others 1960; Heck and others 1980; U. 5. Envi ranmental Protection 

Agency 1986). Ozone is episodic and may be transferred long distances. 

During an episode, ozone trajectories may cover very large areas. For example, 

an ozone episode covering most of a 20-state area occurred April 12-23, 1976 

(Wolff and others 1977). During this episode, ozone concentrations exceeding 

80 ppb occurred simultaneously across the Midwest and the Northeastern United 

States. On the afternoon of April 12, ozone concentrations in excess of 

80 ppb were measured in Indiana and Ohio ahead of the high pressure system 

that moved out of Canada and was centered over Lake Michigan. By the evening 

of April 13, the high ozone concentrations (60 to 80 ppb) observed over 

Indiana and Ohio were advected to the East Coast of the Mid-Atlantic States, 

with trajectories extending from Ohio to New Jersey. The center of the high 

pressure system continued to move southeastward, and by the mcrning of 

April 16 was centered off the coast of North Carolina. At this time, concentrations ranging from 60 to 80 ppb extended from Maine to Virginia and from 

Massachusetts to Virginia in the Boston to Washington corridor (Wol ff and 

others $977)- 

Additional studies indicate that two other amone episodes that took place 

during August 1-6 and August 8-14, 1976, exhibited trajectories similar to the 

one described above (Wolff and others 1980). In both cases, the high pressure 

originated in Canada and moved southeastward into the Midwest. Concentrations 

ranged from 30 to 50 ppb in the high as it moved into the U. S. During the 

first or second day that the center of the high was over the Midwest, two 

areas (the Midwest and the Northeast Corridor) developed ozone concentrations 

greater than 80 ppb. One area extended from the western shore of Lake Michigan 

to St. Louis, across eastern Indiana, Ohio, southeastern Michigan, and 
northwestern Pennsylvania to southern Ontario. The second area was the 

Washington, D. C. -Boston, MA, corridor. As the high pressure area persisted, 

the two areas merged and the ozone concentrations continued to rise until a 

cold front moved down from Canada. Maximum ozone concentrations observed 

during the two August episodes exceeded 200 ppb (Wolff and others 1980). The 

second episode varied from the first in that the trajectory extended south 

over Lexington, KY. In both studies, the episodes included the area from West 

Virginia and Virginia in the South to Maine in the Northeast. Ozone concentrations in trajectories increase when the air parcels pass over urban areas 

where additional ozone is present. For example, in the above episode, an air 

parcel bearing ozone concentrations exceeding 100 ppb added another 100 ppb 

when passing over Philadelphia (Wolff and others 1980). The above studies 

point out that ozone episodes are not an abnormal phenomenon, but can occur 

several times during a single year. 

All ozone episodes are not associated with high pressure systems originating in Canada. Studies of ozone episodes made during July 12-21, July 21-24 

and July 26-30, 1977, indicate that ozone from the western coast of the Gulf 

of Mexico also may be transported into the Northeast (Wolff and Lioy 1980). 

During the episode of July 12-21, 1977, ozone transport was associated with a 

high pressure area over the Gulf of Mexico. Air parcels originating in the 

vicinity of Texas and Louisiana traveled northeastward through the Midwest to 

the Atlantic Coast. The episode that began on July 12 lasted until July 21, 

with the area extending from the Texas-Louisiana Gulf Coast to the northeastern 

Atlantic Coast being exposed to ozone concentrations averaging 120 to 130 ppb. 

The highest concentration of 328 ppb was observed in Connecticut (Wolff and 

Lioy 1980). Unl i ke the July 12-21 episode, the second and third episodes 

were associated with high pressure systems that moved down from Canada and 

suppressed the tropical gulf air. The ozone "river" reaching from the Gulf 

Coast to New England that had come into existence during the July 12-21 

episode was prolonged when the cold fronts of July 21 and July 23, 1977, from 

Canada moved eastward and off the Atlantic Coast. These episodes ~imyjta-~ 

neously exposed nearly two-thirds of the United States, an area 2 x 10 mi 

(Wol f f and Li oy 1980). 

The foregoing discussion not only depicts the episodic nature of ozone 

exposures, but also points out the fact that a major portion of the United 

States east of the Mississippi River has for many years been frequently 

exposed to phytotoxic ozone concentrations. Even though the understanding 

that ozone was not just a local problem but could be transported long 

distances did not develop unti 1 during the 1970'~~ ozone episodes had been 

occurring for some time. Linzon (1960) points out that by 1957, foliar injury 

of white pine had been reported (14 references are cited by Linzon) throughout 

its range under a variety of names, though the cause of the injury was still 

unknown. Foliar injury of white pine under the name of "white pine needle 

bl ight" was reported at the Petawawa Forest Experiment Station, Chalk River, 

Ontario, during the month of July in 1957 and in July and September 1958. 

During 6 of the 8 years from 1957-1964, fol iar injury was observed on white 

pine. Ozone was suspected as a cause of the injury, but attempts to establish 

a definite relationship were not successful though ozone concentrations 

sufficient to cause "weatherfleck" of tobacco were observed (Linzon 1967). 
In the southeastern United States, fol iar injury of eastern white pine 

resulting from ozone exposure also has been observed for some time. Berry and 

Ripperton (1963) reported that, though intensive studies to determine the 

cause of "emergence tipburn" of eastern white pine were not begun in Pocahontas 

County, WV, until 1957, many people in the same area had been reporting casual 

observations regarding the phenomenon for over 20 years. We now know 

"emergence tipburn" and "white pine needle b1 ight" to be visible symptoms of 

ozone injury. Berry (1961) reported that "emergence tipburn seems to be more 

prevalent and more damaging in the northerly latitudes, being very common in 

the Northeastern United States and Canada, and rare on white pine in Georgia." 

Further, Berry cites the occurrence of an ozone episode during July of 1961 

when he states that an unusually high incidence of "emergence tipburn" 

occurred in western North Carolina and "surrounding States." In other words, 

ozone episodes were occurring long before scientists became aware, not only of 

their existence, but of the fact that ozone episodes were causing the injury 

observed on white pine. Hence, beginning no later than the early 19501s, both 

natural and agricultural ecosystems from the Gulf Coasts of Texas and Louisiana 

to the Midwest and Canada and throughout the East from North Carolina and 

Tennessee to Maine have been exposed for many years. Hayes and Skelly (1977) 

associated injury to eastern white pine at three rural sites in Virginia 

between July 1-5, 1975, with a high pressure system over the Great Lakes and a 

low, Hurricane Amy, off the Atlantic Coast. Air parcels bearing ozone moved 

from the Northeast and Midwest into Virginia. The episode was dissipated when 

the cold front from the Midwest moved across Virginia into the Atlantic Ocean. 

Since this time, when ozone episodes have occurred in the same region, they 

have been associated with meteorological phenomena such as cited above (Skelly 

and others 1984). 

Vegetation injury due to ozone was not unknown in the eastern United 

States in the 19401s, though it was not identified as such until 1960. An 

unusual pattern of fol i ar in jury of many cultivated crops, ornamental s , and 

indigenous vegetation was observed in the New Jersey area beginning in 1944. 

The injury was frequently followed by partial or complete defoliation of many 

deciduous and coniferous plants (Daines and others 1960). Ozone was first 

recognized as the causal factor of the foliar injury after Heggestad and 

Middleton (1959) recognized that weatherfleck of tobacco was due to ozone 

injury. Their observations indicate that during an ozone episode in October 

1958, concentrations at Beltsville, MD as measured with a Mast meter reached 

380 ppb on October 9, 500 ppb on October 10, and 430 ppb on October 16, 1958, 

respectively. Ozone injury to crops in New Jersey during the 1959 growing 

season, reported by Daines and others (1960), indicates that another episode 

occurred that year during the growing season. 

It is possible, therefore, that the growth reductions in forest trees 

observable today and attributed to many different factors (table 17) are the 

cumul ati ve responses from exposure to mu1 tip1 e O3 epi sodes that began many 

years ago. 

Many reports point out that trees at high altitudes frequently are 

observed to more severely injured than those lower on the mountain. This 

could be due to O3 exposure because the diurnal behavior of ozone at higher 

a1 titudes results in high a1 titude forest ecosystems suffering greater ozone 

injury than those at 1 ower a1 titudes (Lefohn and Jones 1986; Wolff and others 
1987). Though daily maximum and midday ozone concentrations are similar at 

different altitudes, the dosage increases with height. Ozone is rapidly 

depleted near the surface below the nocturnal inversion layer. However, 

mountainous sites above the nocturnal inversion layer do not experience this 

depletion. Therefore, the total exposure to ozone in mountainous areas can be 

much higher than in nearby valleys (see table 18). Because of this, several 

consideradions need to be taken into account when assessjng exp~stdre-response 

rela&ionships of forest ecosystems at high altitudes: 

I. Si-&es at elevations abwe the raocturaal inversion layer can be 

exposed to higher peak and higher Gob1 concentrations of ozone 

than 5-8" tes at 1 owe? elevati sns (Mi l 1 er and others 1977; Ma" f 1 er 

and others 1982; Walff and others 1987). 

2, The maximum ozone concentrations observed at elevated nountainaus sites, as well as at many nonmounla7"nous rural and remo'i;e 

sites, often occur at night (Lefohn and Jones 1980; WdalFP and 

others 1987). For species such as white pine, in which the 

stomata remain fully or partially open after dark, this is 

particularly important. 

3. Sites at higher elevations are often exposed to sustained or 

mu1 tiple peak concentrations of ozone within a given 24-hour 

period as the result of conditions such as (a) trapping inversions; (b) the successive transport of ozone from multiple 

urban sources upwind, either aloft or across terrain devoid of 

sufficient arone scavengers; and (c) the occurrence of 

mountain-valley and upslope-downslope Flaws, such that the 

trajectory of an air parcel passes back over the same forest or 

stand of trees (Wolff and others 1987). 

Table 18. Comparison of Ozone Data at Different Altitudes and at an Urban Area 

7/22/75 3123175 

No. of 

1-hr max 24-hr mean 1-kr max 24-hr mean hr 

PP~ PP~ PP~ PP~ - >80 ppb 

High Point, 500 m 66 49 136 81 13 

High Point, 300 rn 61 38 118 61 9 

High Point, 140 m 59 26 120 52 9 

Bayonne, N9, sea level 69 33 119 48 7 

High Point Mountain is in New Jersey, 

After Wslff and others (1987)- 

Most studies of forest ecosystems have monitored only the daylight ozone 

concentrations. Thus, the reported exposures may not represent either the 

number length or magnitude to which the forests have been exposed (U.S. 

Envi ronmental Protection Agency 1986). 
The chronic nature of the episodic ozone stresses to which the emsysdems 

of eastern North America Rave Seen subjected has made it difficult to determine 

what changes, if any, have resulted from them. Some, but not a7 1, eastern 

white pine trees are sensitive to ozone. McLaughlin and others (1982) and 

Benoit and oGhers (1982) divided the species inta three categories: sensiliiie, 

intermediate, and tolerant. The sensitive members of the white pine species 

are indicators that a forest stand, community, or ecssystem has been exposed 

to ssone, but do not necessarily inform scientists whether the ecosystem as a 

whole is being affected, Many tree species are sensitive to ozone, Davis and 

LdS l hour (11976) used fa1 iar injury to determine ozone sensl"tivity, Twentythree species of trees grown in North America were listed as sensitive and 18 

as intermediate. More recently, Pye (19881, using experimental data on seedlings, stated that ozone can reduce photosynthesis and growth of 43 tree 

species and hybrids. 

Given the fact that ozone has been a regional stress fop decades, the 

most sensitive plant species have undoubtedly been removed and supplanLed by 

tolerant species. Hayes and Skelly (1977) reported that in 1971, 15 sensitive 

white pine trees were tagged. After 2 years of exposure to ambient ozone 

concentrations, 13 had declined in health, and after 3 years of consecutive 

ozone exposure, 6 of those 13 were dead. The psssfhi 1 ity that some plant 

species may have evolved genetic resistance has been prrsposed (Taylor and 

Nsrby 19851, 

Selection for genetic resistance has undoubtedly occurred -in urban 

habitats. As with many agricultural crops and some forest tree species, trees 

and othgr woody plants that are grown in a variety of urban habitats have been 

selected over time for viability in sLressed environments. Umbaeh and Davis 

(1984) point out that trees obtained from commercial nurseries are not likely 

to represent the full range of genotypes present in the wild population of the 

species. The work of Rhoads and others ($980), along with that of Karnosky 

(1981, 19831, suggests that, based on foliar injury, the majority of the 

plants growing dong streets and urban parks, arboreta, remnant woodlots, and 

suburban communities are relatively insensitive to O3 exposure. Their relative 

insensitivity may be the combined result of genetic selection and physical 

factors affecting stornatal processes. 

Un7ike the ponderosa or Jeffrey pines in the San Bernardins Forest, the 

eastern white pine is not what has recently been termed a "keystone species," 

one so important to its ecosystem that its removal or demise will cause the 

col lapse of that ecosystem (Sunquist 1988). Therefore, ttae demise of the 

white pine, as in the situation cited above, did not make a readily observable 

impact an the ecosystem. Except for the observations of McClenahen (1978) on 

the subtle changes occurring in fogaests wi-ren exposed to pallulilnts Par 

40 years, and Lk experiments sf Dtrchetie and others (1982) to determine 

the effect of ozone exposure on native vegetation in the Blue Ridge, no 

studies have been made in the eastern United States to determine how exposure 

to ozone has affecLed ecosystem components other than trees, 

Had the deaths sf sensitive easGern white pine resulted in dieback and 

mortaljly as severe as have the deaths of the codominant spru@e and fir trees 

in the high elevation forests of North Carolina, New Hampshire, and Vermont, 

our knowledge concernjng the jmpachf ozone on the forest ecosystems sf 
eastern North America would be much greater today. As it is, because we lack 

sufficient data, we can only speculate about the changes that undoubtedly have 

resulted from long-term ozone exposure. 

--The ferest ecosystems of eastern North America are in the 

process of change. Decreases in growth, dieback, and mortality of certain 

tree species have been reported in the high-elevation forests ecosystems from 

Maine, New Hamsphire, and Vermont south to North Carolina and Tennessee. 

Dendroecological studies indicate that the decrease in growth of forest trees 

began during the 1950's or early 1960's. The extent of forest declines and 

the factors that precipitated them, however, are questions of controversy. 

The problem is establishing causation. Many hypotheses have been advanced as 

to the possible cause or causes. 

Attributing the observed growth reductions to air pol 1 ution is difficult 

because data indicating the presence many years ago of injurious concentrations 

of phytotoxic air pollutants is lacking. Of the gaseous air pollutants, ozone 

is the most phytotoxic. Sulfur dioxide injury to vegetation is found most 

frequently in the vicinity of strong point sources. Vegetation exposure to 

phytotoxic concentrations in these areas is usually of short duration. 

(Exceptions are Wawa and Sudbury, Ontario, Trail British Columbia, and Copper 

Hill Basin near Ducktown, TN, where catastrophic emissions resulted in longterm injury to the ecosystems). Nitrogen oxides have not been measured in the 

atmosphere in concentrations sufficient to cause vegetational injury. 

Ozone is a regional pollutant and is, therefore, the only one capable of 

having exposed the entire area where vegetation injury and growth reductions 

have been observed over the past several years without leaving a trace. 

Ozone molecules are ephemeral. By-products of ozone, unl i ke residual s of 

sulfur or nitrogen oxides, fluorides, heavy metals, or radionuclides, are not 

selectively retained or accumulated in any particular components of the 

ecosys tem. 

Limited monitoring of SO and NO concentrations in the Eastern United 

States did not begin until 1&9 or 1%0, while valid oxidant data did not 

become available until 1964 and then only for certain large cities. However, 

emission trends data indicate a sharp increase for SOx and NO beginning in 

the late 1940's. Increase in ozone concentrations, fueled by farge-scale 

emissions that began during World War 11, probably started in the late 1940 s 

even though data do not exist to indicate this. Circumstantial evidence 

suggests that the ecosystems of eastern North American have been subjected to 

episodic phytotoxic concentrations of ozone similar to those cited below for 

at least 30, possibly 40 years. 

Typical ozone episodes in the Eastern United States are associated with 

high-pressure systems which originate out of Canada and move southeastward 

into the Midwest enroute to the Atlantic Coast. Ozone trajectories during an 

episode may cover very large areas. For example, an episode covering most of 

a 20-state area occurred April 12-23, 1976. During this episode, ozone 

concentrations exceeding 80 ppb occurred simultaneously from the Midwest 

across to the Atlantic Coast and into the northeast. Additional studies 

indicate that two other ozone episodes, each lasting one week and exhibiting 

trajectories similar to the one mentioned above, took place during August, 

1976. Maximum concentrations measured during the two August episodes were 

200 ppb. 
The following year there were three episodes, July 12-21; July 21-24 and 

July 26-30, 1977. The last two, like the 1976 episodes, were associated with 

high-pressure systems originating in Canada, but the first episode originated 

in the Texas-Louisiana area and formed "a river of ozone" "from the Texas Gulf 

Coast through the Midwest to the Atlantic Coast. The highest concentration 

of 328 ppb was observed in Connecticut. Ozone concentrations in trajectories 

increase as air parcels pass over urban areas where additional ozone is 

present. There also have been reports for many years of ozone episodes 

originating in the Washington-Philadelphia area and moving westward and south 

across West Virginia and Virginia down into the Carolinas. 

Published reports of field observations support the view that ozone 

episodes have been occurring for at least 40 years. Vegetation injury in the 

eastern United States resulting from ozone was first observed in the 19401s, 

though it was not recognized as such until 1960. Beginning in 1944, an 

unusual pattern of fol iar injury of many cultivated crops, ornamentals, and 

indigenous vegetation was observed in the New Jersey area. Injury was 

frequently followed by partial or complete defoliation of many deciduous and 

coniferous species. Ozone was first recognized as the causal factor after 

Heggestad and Middleton in 1960 reported that weatherfleck of tobacco observed 

during the 1958 growing season in Be1 tsvi 1 le, MD was due to ozone. Ozone 

injury of crops in New Jersey also was observed during the 1959 growing 

season, 

Field observations also suggest that oxidant-induced injury of vegetation 

has been occurring in the Appalachian Mountains for many years. Studies of 

forest decline of eastern white pine in an area covering several hundred 

square mi 1 es of the Cumber1 and Pl ateau of East Tennessee i n the 1950' s 

concluded that unidentified atmospheric constituents were the cause. In 

addition, people living in Pocahontas County, WV in 1957 had been reporting 

casual observations of visible ozone injury of eastern white pine for more 

than 20 years. 

Forest ecosystems at high altitudes can suffer greater ozone injury than 

those at lower altitudes due to the diurnal behavior of ozone. Though daily 

maximum and mid-day ozone concentrations are similar at different altitudes, 

the dosage increases with height. Ozone is rapidly depleted near the surface 

below the nocturnal inversion layer; however, mountainous sites above the 

nocturnal inversion layer do not experience this depletion. Therefore, the 

total exposure to ozone in mountainous areas can be much higher than in nearby 

valleys. Maximum ozone concentrations observed at elevated mountainous sites, 

as well as nonmountainous rural and remote sites often occur at night. In 

addition, sites at higher elevations are often exposed to sustained or 

multiple peak concentrations of ozone within a given 24-hour period. When 

assessing exposure-response relationships of forest ecosystems at high 

altitudes these considerations need to be taken into account. Most studies 

of forest ecosystems have monitored only the daylight ozone concentrations. 

An additional factor that makes determining the effects of pollutant 

stress on ecosystems difficult, is that mature ecosystems are not completely 

stable, but maintain themselves in an oscillating steady state. Severe 

competitional among trees for light, water, and mineral nutrients results 

in the elimination of old and suppressed trees and the addition of new ones. 
The elimination of the trees most susceptible to competitional stresses is 

accelerated by additional and severe periodic stresses such as drought, 

f l oodi ng , f i re and attacks by i nsects or di sease-causi ng organisms. Succession results, causing changes within an ecosystem. The effects of natural 

stresses upon ecosystems, unless they are catastrophic (e. g. , fire, flood, or 

wi ndstorm) are frequently di f Fi cul t to determi ne, 

Pollutant stresses, regardless of whether they are foliage or soil 

mediated, are superimposed upon naturally occurring stresses and may 

a~celerate the processes of change a7 ready underway in ecosystems. Gaseous 

pollutant stresses, whether severe episodic stresses (one or more days of 

acute high exposure) or long-term chronic stresses, occur chiefly through the 

fol iage. 

Ozone is known to impair the growth of agricultural crops, and of native 

vegetation in ecosystems throughout the United States, more than any other air 

pol 1 utant. Exposure of vegetation to ozone inhibits photosynthesis, al ters 

carban allocation and interferes with mycorrhizal formation in tree roots. 

Disruption of these important physiological processes can suppress the growth 

of trees, shrubs, and herbaceous vegetation by decreasing their capacity to 

form carbon compounds needed for growth and their ability to absorb the water 

and nutrients they require for life from the soil. In addition, lass of vigor 

increases susceptibility of trees to insect and pathogens and reduces their 

capabi l i ty to reproduce. 

It is not possible to state unequivocally what the effects of airborne 

pollutant chemicals on forest ecosystems have been. Stresses on trees from 

these pollutants have been both above ground through the foliage and through 

the soil. The responses of decreased growth, dieback and mortality observed 

at the present time in the forests of eastern North America are the result of 

long-term pollutant exposures and other multiple stresses upsn individual 

trees and other perennial species. The effects of continuing pollutant 

stresses are cumulative and appear as growth responses. Growth responses take 

time. Such responses in perennial woody plants usually are not observable 

during the year they occur but often require a generation. This is 

particularly true of injury resulting from repeated acute episodic ozone 

exposures because there is time for recovery between peaks. It is possible, 

therefore, that the growth reductions in forest trees observable today in the 

eastern United States and attributed to many different factors are the result 

of physiological changes to which episodic ozone exposures that began many 

years ago were a major contributor. 
5. REFERENCES 

Abeles, F. B. ; Craker, L. E. ; Forrence, L. E. [and others]. 1971. Fate of air 

pol 1 utants: removal of ethylene, sulfur dioxide, and nitrogen dioxide by 

soil. Science 173914-916. 

Abrahamsen, G. ; Do1 lard, G. J. 1979. Effects of acid precipitation on forest 

vegetation and soil. pp. 1-17. In: Ecological effects of acid precipitation: report of a workshop; 1978 September; Gatehouse-of-Fleet, Galloway, 

United Kingdom. EPRI report no. EPRI S0A77-043. Palo Alto, CA: Electric 

Power Research Institute. 

Adams, H. S. ; Stephenson, S. L. ; Blasing, T. J. [and others]. 1985. Growth-trend 

decl ines of spruce and fir in mid-Appalachian subalpine forests. Environmental and Experimental Botany 25:315-325. 

Alexander, M. 1977. Introduction to soil microbiology. 2nd ed. pp. 225-230; 

239-246; 272-286; 355-356. New York: John Wiley & Sons. 

Altshuller, A.P. 1986. The role of nitrogen oxides in nonurban ozone formation 

in the planetary boundary layer over N. America, W. Europe and adjacent 

areas of ocean. Atmospheric Environment 20:245-268. 

Altshuller, A. P. ; Linthurst, R.A. , eds. 1984. Acidic deposition phenomenon and 

its effects: critical assessment review papers. vol. 1. Atmospheric 

sciences. vol. 2. Effects sciences. EPA report nos. EPA/600/8-83/016aF-bF. 

Washi ngton: U. S. Envi ronmental Protecti on Agency, Off i ce of Research and 

Development. [Available from NTIS, 5285 Port Royal Road, Springfield, VA 

22161; PB85-100030 and PB85-1000481 

Amthor, J.S. 1984. Does acid rain directly influence plant growth? Some 

comments and observations. Environmental Pollution, Series A 36:l-6. 

Anderson, F. K. ; Treshow, M. 1984. Responses of l ichens to atmospheric pol 1 ution. pp. 259-289. In: Treshow, M., ed. Air pollution and plant life. 

(Environmental monographs and symposia). New York: John Wiley & Sons. 

Anderson, L.S. ; Mansfield, T.A. 1979. The effects of nitric oxide pollution on 

the growth of tomato. Environmental Pollution 20:113-121. 

Andreae, M.O. ; Raemdonck, H. 1983. Dimethyl sulfide in the surface ocean and 

the marine atmosphere: a global view. Science 221:744-747. 

Aneja, V.P. ; Aneja, A.P.; Adams, D.F. 1982. Biogenic sulfur compounds and the 

global sulfur cycle. Journal of the Air Pol lution Control Association 

32: 803-807. 
Baldocchi, D.D.; Camara, P.; Hicks, B.B. 1987. A canopy stomata1 resistance 

model for gaseous deposition to vegetated surfaces. Atmospheric Environment 21:91-101. 

Barchet, R. 1987. Acid deposition and its gaseous precursors. In: Interim 

assessment: the causes and effects of acidic deposition. vol . 3. 

Atmospheric processes. Washington: Office of the DirectQr of Research; 

Chapter 5. 

Barnes, R.L. 1972a. Effects of chronic exposure to ozone on photosynthesis and 

respiration of pi nes. Envi ronmental Pol 1 uti on 3: 133-138. 

Barnes, R.L. 1972b. Effects of chronic exposure to ozone on soluble sugar and 

ascorbic acid contents of pine seedlings. Canadian Journal of Botany 

50: 215-219. 

Bennett, D.A.; Goble, R.L.; Linthurst, R.A. 1985. The acidic deposition 

phenomenon and its effects: critical assessment document. EPA report 

no. EPA/600/8-85/001. Washi ngton: U. S. Environmental Protection Agency; 

Office of Acid Deposition, Envi ronmental Moni tori ng , and Qua1 i ty 

Assurance. [Available from NTIS, 5285 Port Royal Road, Springfield, 

VA 22161; PB85-2333931 

Bennett, J. H. ; Hi 11, A. C. 1975. Interactions of air pollutants with canopies of 

vegetation. pp. 273-306. In: Mudd, J. B. ; Kozlowski , T. T. , eds. Responses 

of plants to air pollution. New York: Academic Press. 

Bennett, J. H. ; Hi 11, A. C. ; Gates, D. M. 1973. A model for gaseous pol lutant 

sorption by leaves. Journal of the Air Pollution Control Association 

23: 957-962. 

Bennett, J. P. ; Resh, H.M. ; Runeckles, V. C. 1974. Apparent stimulations of plant 

growth by ai r pol l utants. Canadian Journal of Botany 52: 35-41. 

Benoit, L. F. ; Skelly, J.M. ; Moore, L.D. [and others]. 1982. Radial growth 

reductions of Pinus strobus L. correlated with foliar ozone sensitivity as 

an indicator of ozone-induced losses in eastern forests. Canadian Journal 

of Forest Research 12673-678. 

Berry, C. R. 1961. White pine emergence tipburn, a physiogenic di sturbance. 

Station Paper 130. Asheville, NC: U.S. Department of Agriculture, Forest 

Service, Southeastern Forest Experiment Station. 8pp. 

Berry, C.R. ; Hepting, G.H. 1964. Injury to eastern white pine by unidentified 

atmospheric constituents. Forest Science 10:2-13. 

Berry, C.R. ; Ripperton, L.A. 1963. Ozone, a possible cause of white pine 

emergence tipburn. Phytopathology 53:552-557. 

Berry, C. R. 1964. Differences in concentrations of surface oxidant between 

valley and mountaintop conditions in the southern Appalachians. Journal 

of the Air Pollution Control Association 14938-239. 
Billings, W.D. 1978. Plants and the ecosystem. 3rd ed. pp. 1-62; 83-108. 

Belmont, CA: Wadsworth Pub1 ishing Co. , Inc. 

Binkley, D. ; Driscoll , C.T. ; A1 len, H. L. [and others]. 1989. Acidic deposition 

and forest soils: Gontex and case studies in the southern United States 

pp. 149. (Ecological Studies: 52). Burl in, Federal Republ ic of Germany: 

Spri nger-Verlag. 

Black, C.A. 1968. Soil-plant relationships. 2nd ed. New York: John Wiley & 

Sons. 

Black, V.J. 1984. The effect of air pollutants on apparent respiration. 

pp. 231-248. In: Koziol , M. J. ; Whatley, F. R. , eds. Gaseous air pol 1 utants 

and plant metabolism. London, United Kingdom: Butterworths. 

Black, V.J. 1985. SO2 effects on stomata1 behavior. pp. 96-117. In: Winner, 

W. E. ; Mooney, H.A. ; Goldstein, R.A. , eds. Sulfur dioxide and vegetation: 

physiology, ecology, and policy issues. Stanford, CA: Stanford University 

Press. 

Black, V. J. ; Unsworth, M.H. 1979. Effects of low concentrations of sulphur 

dioxide on net photosynthesis and dark respiration of -- Vicia faba. Journal 

of Experimental Botany 30:473-483. 

Bohn, H. L. 1972. Soi 1 absorption of air pol 1 utants. Journal of Environmental 

Qua1 i ty 1: 372-377. 

Bolin, B. ; Granat, L. ; Ingelstam, L. [and others]. 1971. Sweden's case study 

for the United Nations conference on the human environment. Air pollution 

across national boundaries: the impact on the environment of sulfur in air 

and precipitation. Stockholm, Sweden: P. A. Norstedt and Sons. 

Bormann, F.H. 1979. Catastrophic disturbance and the steady state in northern 

hardwood forests. American Scientist 67:660-669. 

Bormann, F. H. 1985. Air pol 1 ution and forests: an ecosystem perspective. 

Bi oSci ence 35: 434-441. 

Botkin, D.B. ; Smith, W.H. ; Carlson, R.W. 1971. Ozone suppression of white pine 

net photosynthesis. Journal of the Air Pollution Control Association 

21: 778-780. 

Botkin, D.B. ; Smith, W.H. ; Carlson, R.W. [and others]. 1972. Effects of ozone 

on white pine saplings: variation in inhibition and recovery of net 

photosynthesis. Environmental Pollution 3:273-289. 

Brandt, C. J. 1987. Acidic precipitation: formation and impact on terrestrial 

ecosystems. pp. 91-113; 123-146. Duesseldorf, Federal Republic of Germany: 

Verein Deutscher Ingenieure-Kommission Reinhaltung der Luft. 

Brown, V. K. 1984. Secondary succession: insect-plant relationships. BioScience 

34: 710-716. 
Bruck, R. I. ; Robarge, W. ; Khorram, S. [and others). 1988. Observations of 

forest decline in the boreal montane ecosystem of Mt. Mitchel 1, North 

Carolina - an integrated forest response approach. In: Proceedings of the 

1I.S.-F.R.G. symposium on forest decline; 1987 October; Burlington, VT. 

Broomall , PA: U. S. Forest Service. 

Bull, J. N. ; Mansfield, ?.A. 1374. Photosynthesis in leaves exposed to SO2 and 

NO2. Nature 250:443-444. 

Busing, R.T. ; Clebsch, E.E.C. ; Eagar, C.C. [and others]. 1988. Two decades of 

change in a Great Smokey Mountains spruce-fir forest. Bulletin of the 

Torrey Botanical Cl ub 115: 25-31. 

Cadle, S.H. ; Countess, R. J. ; Kelly, N.A. 1982. Nitric acid and ammonia in urban 

and rural locations. Atmospheric Environment 16:2501-2506. 

Capron, T. M. ; Mansfield, T. A. 1976. Inhibition of net photosynthesis in tomato 

in air polluted with NO and NOz. Journal of Experimental Botany 

27: 1181-1186. 

Carlson, R.W. 1979. Reduction in the photosynthetic rate of Acer, Quercus and 

Fraxinus species caused by sulphur dioxide and ozone. Environmental 

Pol 1 uti on 18: 159-170. 

Chevone, B. I. ; Herzfeld, D. E. ; Krupa, S. V. [and others]. 1986. Direct effects 

of atmospheric sulfate deposition on vegetation. Journal of the Air 

Pollution Control Association 36:813-815. 

Christiansen, E. ; Waring, R. H. ; Berryman, A. A. 1987. Resistance of conifers to 

bark beetle attack: searching for general relationships. Forest Ecology 

and Management 22:89-106. 

Cline, M. L. ; Stephans, R. J. ; Marx, D.H. 1987. Influence of atmospherically 

deposited nitrogen on mycorrhizae: a critical literature review. Raleigh, 

NC: U. S. Department of Agriculture, Forest Service, Southern Commercial 

Forest Research Cooperative. 

Cobb, F.W. , Jr. ; Stark, R.W. 1970. Decline and mortality of smog-injured 

ponderosa pine. Journal of Forestry 68: 147-149. 

Cobb, F.W. , Jr. ; Wood, D. L. ; Stark, R.W. [and others). 1968. IV. Theory on the 

relationship between oxidant injury and bark beetle infestation. Hilgardia 

39: 141-152. 

Coleman, J. S. ; Jones, C. G. 1988a. Acute ozone stress on eastern cottonwood 

deltoides Bartr.) and the pest potential of the Aphid, 

Thomas (Homoptera: Aphidi dae). Envi ronmental 

Entomol ogy 17: 207-212. 

Coleman, J.S.; Jones, C.G. 1988b. Plant stress and insect performance: 

cottonwood, ozone and a leaf beetle. Oecologia 76:in press. 
Constantinidou, H. A. ; Kozlowski , T. T. 1979. Effects of sulfur dioxide and ozone 

on Ulmus americana seedlings: 1. visible injury and growth. Canadian 

Journal of Botany 57: 170-175. 

Cowling, E.B. 1985. Prepared discussion. Journal of the Air Pcllution Gontrd 

Association 35: 916-919. 

Coyne, P. I. ; Bingham, G. E. 1981. Comparative ozone dose response of gas 

exchange in a ponderosa pine stand exposed to long-term fumigations. 

Journal of the Air Pollution Control Association 31:38-41. 

Daines, R.H. ; Leone, I.A. ; Brennan, E. 1960. Air pollution as it affects 

agriculture in New Jersey. Bull. 794-A. New Brunswick, NJ: New Jersey 

Agricultural Experiment Station, Rutgers. 

Darrall, N.M. 1986. The sensitivity of net photosynthesis in several plant 

species to short-term fumigation with sulphur dioxide. Journal of 

Experimental Botany 37: 1313-1322. 

Davis, D. D. ; Wi 1 hour, R. G. 1976. Susceptibi 1 ity of woody plants to sulfur 

dioxide and photochemical oxidants. EPA report no. EPA-600/3-76-102. 

Corvall is, OR: U. S. Environmental Protection Agency, Corvall is Envi ronmental Research Laboratory. [Avai 1 able from NTIS, 5285 Port Royal Road, 

Springfield, VA 22161; PB-2608811 

Dewey, J. E. ; Ciesla, W.M. ; Meyer, H. E. 1974. Insect defoliation as a predisposing agent to bark beet1 e outbreak in eastern Montana. Environmental 

Entomology 3: 722. 

Dochinger, L.S. ; Bender, F.W.; Fox, F.L. [and others]. 1970. Chlorotic dwarf of 

eastern white pine caused by an ozone and sulfur dioxide interaction. 

Nature 225: 476. 

Dochinger, L.S. ; Seliskar, C.E. 1970. Air pollution and the chlorotic dwarf 

disease of eastern white pine. Forest Science 16:46-55. 

Dochinger, L. S. ; Townsend, A.M. 1979. Effects of roadside deicer salts and 

ozone on red maple progenies. Environmental Pollution 19:229-237. 

Donahue, R.L.; Miller, R.W.; Shickluna, J.C. 1977. Soils: an introduction to 

soils and plant growth. 4th ed. pp. 123-141. Englewood Cliffs, NJ: 

Prenti ce-Hal 1 , Inc. 

Dreisinger, B. R. 1965. Sulphur dioxide levels and the effects of the gas on 

vegetation near Sudbury, Ontario. Presented at: 58th annual meeting of the 

Air Pollution Control Association; 1965 June; Toronto, ON, Canada. Paper 

no. 65-121. Pittsburgh, PA: Air Pol lution Control Association. 

Dreisinger, B. R. ; McGovern, P. C. 1970. Monitoring atmospheric sulphur dioxide 

and correlating its effects on crops and forests in the Sudbury area. In: 

Linzon, S. N. , ed. Impact of air pol lution on vegetation: proceedings of a 

speci a1 ty conference; 1970 Apri 1 ; Toronto, ON, Canada. Toronto, ON, 

Canada: Ontario Department of Energy and Resource Management. 
Duchelle, S. F. ; Skelly, J.M. 1981. Response of common milkweed to oxidant air 

pollution in the Shenandoah National Park in Virginia. Plant Disease 

65: 661-663. 

Duchelle, S.F. ; Skelly, J.M.; Chevone, B.I. 1982. Oxidant effects on forest 

tree seedling growth in the Appalachian Mountains. Water, Air, and Soil 

Pollution 18:363-373. 

Duchelle, S.F. ; Skelly, J.M. ; Sharick, T. L. [and others]. 1983. Effects of 

ozone on the productivity of natural vegetation in a high meadow of the 

Shenandoah National Park of Virginia. Journal of Environmental Management 

17: 299-308. 

Ehrlich, P. R. ; Mooney, H.A. 1983. Extinction, substitution, and ecosystem 

services. BioScience 33:248-254. 

Evans, G. ; Finkelstein, P. ; Martin, B. [and others]. 1983. Ozone measurements 

from a network of remote sites. Journal of the Air Pollution Control 

Association 33: 291-296. 

Fankhauser, R. K. 1976. Ozone levels in the vicinity of 33 cities. Journal of 

the Air Pol 1 ution Control Association 26: 771-777. 

Fowler, D. 1980. Removal of sulphur and nitrogen compounds from the atmosphere 

in rain and by dry deposition. pp. 22-32. In: Drablos, D. ; Tollan, A., 

eds. Ecological impact of acid precipitation: proceedings of an international conference; 1980 March; Sandefjord, Norway. Oslo-As, Norway: SNSF 

Project no. NOR-SNSF/001. 

Foy, C. D. ; Chaney, R.C. ; White, M.C. 1978. The physiology of metal toxicity to 

plants. Annual Review of Plant Physiology 29: 511-566. 

Freer-Smith, P.H. 1984. The responses of six broadleaved trees during long-term 

exposure to SO, and NO,. New Phytologist 97:49-61. 

Furukawa, T. ; Totsuka, T. 1979. Effects of NO, SO2 and 0, alone and in 

combinations on net photosynthesis in sunflower. Seibutsu Kankyo Chosetsu 

17: 161-166. 

Gal loway, J. N. ; Whelpdale, D. M. 1980. An atmospheric sulfur budget for eastern 

North America. Atmospheric Environment 14: 409-417. 

Garsed, S.G. 1985. SO2 uptake and transport. pp. 75-95. In: Winner, W. E. ; 

Mooney, H. A. ; Goldstein, R. A. , eds. Sulphur dioxide and vegetation: 

physiology, ecology, and pol icy issues. Stanford, CA: Stanford University 

Press. 

Gmur, N.F. ; Evans, L.S. ; Cunningham, E.A. 1983. Effects of ammonium sulfate 

aerosols on vegetation 2. mode of entry and responses of vegetation. 

Atmospheric Environment 17:715-721. 

Gordon, A.G. ; Gorham, E. 1963. Ecological aspects of air pollution from an 

i ron-sintering plant at Wawa, Ontario. Canadian Journal of Botany 

41: 1063-1077. 
Gschwandtner, G. ; Gschwandtner, K. ; Eldridge, K. [and others]. 1986. Historic 

emissions of sulfur and nitrogen oxides in the United States from 1900 to 

1980. Journal of the Air Pollution Control Association 36:139-149. 

Guderian, R. , ed. 1985. Air pol 1 ution by photochemical oxidants. Formation, 

transport, control, and effects on plants. (Ecological studies:52). 

Berl i n, Federal Republ ic of Germany: Spri nger-Verl ag. 

Guderian, R. ; Kueppers, K. 1980. Response of plant communities to air pol lution. pp. 187-199. In: Miller, P.R., ed. Proceedings of the symposium on 

effects of air pollutants on Mediterranean and temperate forest 

ecosystems; 1980 June; Riverside, CA. Gen. Tech. Rep. PSW-43. Berkeley, 

CA: U.S. Department of Agriculture, Forest Service, Pacific Southwest 

Forest and Range Experiment Station. [Available from NTIS, 5285 Port Royal 

Road, Springfield, VA 22161; PB81-1337201 

Hacskaylo, E. 1973. The Torrey symposium on current aspects of fungal development: IV. dependence of mycorrhizal fungi on hosts. Bulletin of the Torrey 

Botanical Cl ub 100: 217-223. 

Hain, F.P. ; Arthur, F.H. 1985. The role of atmospheric deposition in the 

latitudinal variation of Fraser fir mortality caused by the Balsam Woolly 

Adelgid, Adelges piceae(Ratz. ) (Hemipt. , Adel gidae): a hypothesis. 

Zei tschri ft fuer Angewandte Entomol ogie 99: 145-152. 

Hales, J.M. 1984. Precipitation scavaging processes. pp. 6-1 to 6-80. In: 

A1 tshul ler, A. P. ; Linthurst, R.A. , eds. Acidic deposition phenomenon and 

its effects: critical assessment review papers, vol. 1. Atmospheric 

sciences. EPA report no. EPA/600/8-83/016AF. Washi ngton: U. S. Envi ronmental Protection Agency, Office of Research and Development. [Available 

from NTIS, 5285 Port Royal Road, Springfield, VA 22161; PB85-1000301 

Hallgren, J.-E. 1978. Physiological and biochemical effects of sulfur dioxide 

on plants. pp. 163-209. In: Nriagu, J.O., ed. Sulfur in the environment. 

Part 2: Ecological impacts. New York: John Wiley & Sons. 

Hanover, J.W. 1975. Physiology of tree resistance to insects. Annual Review of 

Entomology 20: 75-95, 

Hayes, E. M. ; Skel ly , J. M. 1977. Transport of ozone from the northeast U. S. into 

Virginia and its effect on eastern white pines. Plant Disease Reporter 

61: 778-782. 

Heagle, A. S. ; Johnston, J. W. 1979. Variable responses of soybeans to mixtures 

of ozone and sulfur dioxide. Journal of the Air Pollution Control Association 29: 729-732. 

Heath, R. L. 1975. Ozone. pp. 23-55. In: Mudd, J. B. ; Kozlowski , T. T. , eds. 

Response of plants to air pollution. New York: Academic Press, Inc. 

Heath, R. L. 1980. Initial events in injury to plants by air pollutants. Annual 

Review of Plant Physiology 31: 395-431. 
Heck, W.W. 1982, Future directions in air pollution research. pp. 411-435. In: 

Unsworth, M.H. ; Ormrod, D. P., eds. Effects of gaseous air pollution in 

agriculture and horticulture. Boston, MA: Butterworth Scientific. 

Heck, M. M. ; Larsen, R. I, ; Heagl e, A. S. N80. Measuring the acute dose-response 

of plants to ozone. pp. 32-49. In: Teng, P. S. ; Krupa, S. V. , eds. Crop l oss 

assessment: prcceedi ngs of E. C. Stakman commemorative symposi um. Mi scellaneous Publ. 7. St, Paul, MN: University of Minnesota, Agricultural 

Experiment Station. 

Heck, W.W. ; Taylor, O.C. ; Adams, R. [and others]. 1982. Assessment of crop loss 

from ozone. Journal of the Air Pollution Control Association 32:353-361. 

Hedgcock, G. G. 1914. In jury by smelter smoke in southeastern Tennessee. Journal 

of the Washington Academy of Sciences 4:70-71. 

Heggestad, H. E. ; Middleton, J. T. 1959. Ozone in high concentrations as cause of 

tobacco 1 eaf i n jury. Science 129: 208-210. 

Heikes, B.G.; Kok, G.L. ; Walega, J.G. [and others]. 1987. H202, O3 and SO2 

measurements in the lower troposphere over the eastern United States 

during fall. Journal of Geophysical Research 92:915-931. 

Hicks, B.B. 1984. Dry deposition processes. pp. 7-1 to 7-70. In: Altshuller, 

A. P. ; Linthurst, R.A. , eds. Acidic deposition phenomenon and its effects: 

critical assessment review papers, vol. 1. Atmospheric sciences. EPA 

report no. EPA/600/8-83/016AF. Washington: U. S. Environmental Protection 

Agency, Office of Research and Devel opment. [Avai l abl e from NTIS, 5285 

Port Royal Road, Springfield, VA 22161; PB85-1000301 

Hi1 1, A. C. ; Bennett, J. H. 1970. Inhibition of apparent photosynthesis by 

nitrogen oxides. Atmospheric Envi ronment 4: 341-348. 

Hogsett, W. E. ; Plocher, M. ; Wildman, V. [and others]. 1985. Growth response of 

two varieties of slash pine seedlings to chronic ozone exposures. Canadian 

Jouranl of Botany 63: 2369-2376. 

Hosker, R.P., Jr.; Lindberg, S.E. 1982. Review: atmospheric deposition and 

plant assimilation of gases and particles. Atmospheric Environment 

16: 889-910. 

Houston, D.B. ; Dochinger, 1.5. 1977. Effects of ambient air pollution on cone, 

seed, and pollen characteristics in eastern white and red pines. Environmental Pol 1 ution 12: 1-5. 

Hughes, P. R. ; Laurence, J.A. 1984. Relationship of biochemical effects of air 

pollutants on plants to environmental problems: insect and microbial 

interactions. pp. 361-377, In: Koziol , PI. J. ; Whatley, F. R. , eds, Gaseous 

air pollutants and plant metabolism. London, United Kingdom: Butterworths. 

Husar, R.B. 1986. Emissions of sulfur dioxide and nitrogen oxides and trends 

for Eastern North America. pp. 48-92. In: Acid deposition: long-term 

trends. Washington: National Academy Press. 

Karnosky, D. F. 1976. Threshold levels for fol iar injury to Popul us tremuloides 

by sulfur dioxide and ozone. Canadian Journal of Forest Reseach 6366-169. 

Karnosky, D. F. 1981. Chamber and field evaluations of air pol 1 ution tolerances 

of urban trees. Journal of Arboricul ture 7: 99-105. 

Karnosky, D. F. 1983. Variation in air pollution responses of hardwood trees. 

Aqui lo Ser. Botanica 19: 183-188. 

Keller, T. 1985. SO2 effects on tree growth. pp. 250-263. In: Winner, W.E. ; 

Mooney, H.A. ; Goldstein, R. A. , eds. Sulfur dioxide and vegetation: physiology, ecology, and policy issues. Stanford, CA: Stanford University 

Press. 

Keller, Th. 1977. The effect of long duration, low SO, concentrations upon 

photosynthesis of conifers. pp. 81-83. In: Kasuga, 5.; Suzuki, N.; Yamada, 

T. [and others], eds. Proceedings of the fourth international clean air 

congress; 1977 May; Tokyo, Japan. Tokyo, Japan: Japanese Union of Air 

Pollution Prevention Associations. 

Kelly, J.; Lambert, M.J. 1972. The relationship between sulphur and nitrogen in 

the foliage of Pinus radiata. Plant and Soil 37: 395-408. 

Kelly, N.A. ; Wolff, G.T. ; Ferman, M.A. 1984. Sources and sinks of ozone in 

rural areas. Atmospheric Envi ronment 18: 1251-1266. 

Kimmerer, T. W. ; Kozlowski , T. T. 1982. Ethylene, ethane, acetaldehyde, and 

ethanol production by plants under stress. Plant Physiology 69:840-847. 

Klein, R. 1985. At Camels Hump and other forests: 'something very serious' is 

happeni ng. Envi ronmental Forum 4(7): 22, 24-25. 

Klein, R.M. ; ~erki-ns, T.D. 1987. Cascades of causes and effects of forest 

decline. Ambio 16:86-93. 

Kozlowski, T.T. 1980. Impacts of air pollution on forest ecosystems. BioScience 

30: 88-93. 

Kozlowski, T.T. 1985. SO, effects on plant community structure. pp. 431-453. 

In: Winner, W. E. ; Mooney, H.A. ; Goldstein, R.A. , eds. Sulfur dioxide and 

vegetation: physiology, ecology, and policy issues. Stanford, CA: Stanford 

University Press. 

Kozlowski , T. T. ; Constantinidou, H. A. 1986a. Responses of woody plants to 

environmental pollution: part 1. sources and types of pollutants and plant 

responses. Forestry Abstracts 47:5-51. 

Kozlowski , T. T. ; Constantinidou, H. A. 1986b. Environmental pol lution and tree 

growth: part 2. factors affecting responses to pollution and alleviation 

of pollution effects. Forestry Abstracts 47:105-132. 

Kramer, P. J. 1983. Water relations of plants. New York: Academic Press. 
Kramer, P.3.; Kozlowski, T.T. 1979. Physiology of woody plants. pp. 1-12. New 

York: Academic Press. 

Krause, G.H.M. ; Prinz, B. ; Jung, K.D. 1984. Forest effects in West Germany. 

pp. 297-332. In: Davis, D. 0. ; Mi i ]en, A. A. ; Dochinger, L. , eds. Air 

pollution and the productivity of the forest: proceedings of the 

symposium; 1983 October; Washington, DC. Arlington, VA: Izaak Malton 

League of America. 

Kress, L.W.; Skelly, J.M. 1982. Response of several eastern forest tree species 

to chronic doses of ozone and nitrogen dioxide. Plant Disease 

66: 1149- 1152. 

Kress, L.W. ; Skelly, J.M. ; Hinkelmann, K.H. 1982. Growth impact of 03, NOa 

and/or SO2 on -- Pi nus taeda. Envi ronmental Monitoring and Assessment 

1: 229-239. 

Krupa, 5.; Fries, N. 1971. Studies on ectomycorrhizae of pine. 1. Production of 

volatile organic compounds. Canadian Journal of Botany 493425-1431. 

Krupa, S.V. ; Legge, A.H. 1986. Single or joint effects of coarse and fine 

particulate sulfur aerosols and ozone on vegetation. pp. 879-887. In: Lee, 

S. D. ; Schneider, T. ; Grant, L. D. [and others], eds. Aerosol s: research, 

risk assessment and control strategies: proceedings of the second 

U. S. -Dutch international symposium; 1985 May; Mi 11 iamsburg, VA. Chelsea, 

MI: Lewis Publishers, Inc. 

Larsson, S. ; Oren, R. ; Waring, R.H. [and others]. 1983. Attacks of mountain 

pine beetle as related to tree vigor of ponderosa pine. Forest Science 

29: 395-402. 

Lefohn, A. S. ; Davis, C. E. ; Jones, C. K. [and others]. 1987. Co-occurrence 

patterns of gaseous ai r pol 1 utant pairs at different minimum concentrations in the United States. Atmospheric Environment 21:2435-2444. 

Lefohn, A. S. ; Jones, C. K. 1986. The characterization of ozone and sulfur 

dioxide air quality data for assessing possible vegetation effects. 

Journal of the Air Pollution Control Association 36:1123-1129. 

Lefohn, A.S.; Mohnen, V.A. 1986. The characterization of ozone, sulfur dioxide, 

and nitrogen dioxide for selected monitoring sites in the Federal Republic 

of Germany. Journal of the Air Pollution Control Association 36:1329-1337. 

Lefohn, A. S. ; Tingey, D. T. 1984. The co-occurrence of potential ly phytotoxic 

concentrations of various gaseous air pollutants. Atmospheric Environment 

18: 2521-2526. 

Legge, A. H. 1980. Primary productivity, sulfur dioxide and the forest 

ecosystem: an overview of a case study. pp. 51-62. In: Mi 1 ler, P. R. , ed. 

Proceedings of the symposium on effects of air pollutants on Mediterranean 

and temperate forest ecosystems; 1980 June; Riverside, CA. Gen. Tech. Rep. 

PSW-43. Berkeley, CA: U. S. Department of Agriculture, Pacific Southwest 

Forest and Range Experiment Station. [Available from NTIS, 5285 Port Royal 

Road, Springfield, VA 22161; P881-1337201 
Legge, A.H.; Jaques, D.R.; Harvey, G.W. [and others]. 1981. Sulphur gas emissions in the Boreal Forest: the West Whitecourt case study. Water, Air, 

and Soi 1 Pol 1 ution 15: 77-85, 

Lindberg, S.E.; Harriss, R.C.; Turner, R.R. [and others]. 1979. Mechanisms and 

rates of atmospheric deposition of selected trace elements and sulfate to 

a deciduous forest watershed. Report no ORNL/TM-6674. Oak Ridge, TN: U. S. 

Department of Energy, Oak Ridge National Laboratory. [Available from NTIS, 

5285 Port Royal Road, Springfield, VA 22161; ORNLITM-66741 

Lindberg, S. E. ; McLaughl in, S. B. 1986. Air pollutant interactions with vegetation: research needs in data acquisition and interpretation. pp. 449-503. 

In: Legge, A.H. ; Krupa, S.V. , eds. Air pollutants and their effects on the 

terrestri a1 ecosystem. (Nri agu, J. 0. , ed. Advances i n environmental 

science and technology: vol. 18). New York: John Wiley & Sons, Inc. 

Linzon, S.N. 1960. The development of foliar symptoms and the possible cause 

and origin of white pine needle blight. Canadian Journal of Botany 

38: 153-161. 

Linzon, S.N. 1967. Ozone damage and semimature-tissue needle blight of eastern 

white pine. Canadian Journal of Botany 45:2047-2061. 

Linzon, S.N. 1971. Economic effects of sulfur dioxide on forest growth. Journal 

of the Air Pollution Control Association 21:81-86. 

Linzon, S.N. 1978. Effects of airborne sulfur pollutants on plants. pp. 109-162. 

In: Nriagu, J.O., ed. Sulfur in the environment: part 2: ecological 

impacts. New York: John Wi ley & Sons. 

Linzon, S.N. 1980. Acute and chronic effects of sulfur dioxide on natural 

vegetation. pp. 50-61. In: The proposed SOX and particulate standard: a 

specialty conference sponsored by APCA, TE Effects Division; 1980 

September; Atlanta, GA. Pittsburgh, PA: Air Pollution Control Association. 

Lovett, G.M. ; Lindberg, S. E. 1986. Dry deposition of nitrate to a deciduous 

forest. Biogeochemi stry 2: 137-148. 

Lovett, G.M.; Reiners, W.A. ; Olson, R.K. 1982. Cloud droplet deposition in 

subalpine balsam fir forests: hydrological and chemical inputs. Science 

218: 1303-1304. 

Luck, R.F. 1980. Impact of oxidant air pollution on ponderosa and Jeffrey pine 

cone production. p. 240. In: Miller, P.R., ed. Proceedings of the 

symposium on effects of air pollutants on Mediterranean and temperate 

forest ecosystems; 1980 June; Riverside, CA. Gen. Tech. Rep. PSW-43. 

Berkeley, CA: U.S. Department of Agriculture, Pacific Southwest Forest and 

Range Experiment Station. [Available from NTIS, 5285 Port Royal Road, 

Springfield, VA 22161; P881-1337201 

Ma1 hotra, S. S. ; Hocking, D. 1976. Biochemical and cytological effects of 

sulphur dioxide on plant metabolism. New Phytologist 76:227-337. 
Malhoutra, S.S.; Kahn, A.A. 1984. Biochemical and physiological impact of major 

pollutants. pp. 113-157. In: Treshow, M., ed. Air pollution and plant 

life. (Polunin, N., ed. Environmental monographs and symposia). New York: 

John Wiley & Sons. 

Ma1 lant, R. K. A. M. ; Slanina, J. ; Masuch, G. [and others]. 1986. Experiments on 

H20,-containing fog exposures of young trees. pp. 901-908. In: Lee, S.D. ; 

Schneider, T. ; Grant, L. D. [and others], eds. Aerosols: research, risk 

assessment and control strategies: proceedings of the second U.S.-Dutch 

international symposium; 1985 May; Wi ll iamsburg, VA. Chelsea, MI: Lewis 

Publishers, Inc. 

Mann, L. K. ; McLaughl in, S. B. ; Shriner, D. S. 1380. Seasonal physiological 

responses of white pine under chronic air pollution stress. Environmental 

and Experimental Botany 20: 99-105. 

Mansfield, T.A. ; Freer-Smith, P.H. 1981. Effects of urban air pollution on 

plant growth. Biological Review 56:343-368. 

Marx, D.H. 1988. Southern Forest Altas Project. Presented at: 81st. annual 

meeting, Air Polution Control Association, June; Dallas, TX. Pittsburgh, 

PA. Air Pol 1 ution Control Association; paper No. 88-70.7. 

Mattson, W. J. ; Haack, R.A. 1987. The role of drought in outbreaks of 

plant-eating insects. BioScience 37:llO-118. 

McClenahen, J.R. 1978. Community changes in a deciduous forest exposed to air 

pollution. Canadian Journal of Forest Research 8:432-438. 

McClenahen, 3.R. 1984. Air pollutant effects on forest communities. pp. 83-94. 

In: Davis, D.D. ; Mil len, A.A. ; Dochinger, L. , eds. Air pollution and the 

productivity of the forest: proceedings of the symposi um; 1983 October; 

Washington, DC. Arlington, VA: Izaak Walton League of America. 

McLaughl in, S. B. ; McConathy, R. K. ; Duvick, D. [and others]. 1982. Effects of 

chronic air pol lution stress on photosynthesis, carbon a1 location, and 

growth of white pine trees. Forest Science 28:60-70. 

McLaughl in, S. B. ; Shriner, D. S. 1980. A1 location of resources to defense and 

repair. pp. 407-431. In: Horsfall, J.B. ; Cowling, E.B. , eds. Plant disease: vol. 5. New York: Academic Press, Inc. 

Miller, P.L. 1973. Oxidant-induced community change in a mixed conifer forest. 

pp. 101-113. In: Naegele, J.A., ed. Air pollution damage to vegetation. 

(Advances in chemistry series: no. 122). Washington: American Chemical 

Society. 

Miller, P.R. 1984. Ozone effects in the San Bernardino National Forest. 

pp. 161-197. In: Davis, D. D. ; Mi 1 len, A. A. ; Dochinger, L. , eds. Air 

pollution and the productivity of the forest: proceedings of the 

symposium; 1983 October; Washington, DC. Arlington, VA: Izaak Walton 

League of America. 
Miller, P.R,; Kickert, R.N.; Taylor, O.C. [and others]. 1977. Photochemical 

oxidant air pollutant effects on a mixed conifer forest ecosystem: a 

progress report, 1976. EPA report no. EPA-6OO/3-77-104. Corvallis, OR: 

U.S. Environmental Protection Agency, Gorvallis Environmental Research 

Laboratory. [Avai lable from NTIS, 5285 Port Royal Road, Springfield, VA 

22161; PB-2745311 

Miller, P.R. ; McBride, J.R. 1975. Effects of air pollutants on forests. 

pp. 195-235. In: Mudd, J. B. ; Kozlowski , T. T. , eds. Responses of plants to 

air pollution. (Kozlowski, T. T. , ed. Physiological ecology: a series of 

monographs, texts, and treatises). New York: Academic Press. 

Miller, P.R.; Parmeter, J.R., Jr.; Flick, B.H. [and others]. 1969. Ozone dosage 

response of ponderosa pine seedlings. Journal of the Air Pollution Control 

Association 19:435-438. 

Miller, P.R.; Stolte, K.W. 1984. Response of forest species to 03, SO2, and NO2 

mixtures. Paper no. 84-30.5. Presented at: 77th annual meeting of the Air 

Pollution Control Association; 1984 June; San Francisco, CA. Pittsburgh, 

PA: Air Pol 1 ution Control Association. 

Miller, P.R. ; Taylor, O.C. ; Wilhour, R.G. 1982. Oxidant air pollution effects 

on a western con7 ferous forest ecosystem. EPA report no. EPA-600/D-82-276. 

Corval 1 is, OR: U. S. Environmental Protection Agency, Environmental 

Research Laboratory. [Avai 1 able from NTIS, 5285 Port Royal Road, 

Springfield, VA 22161; PB83-1893081. 

Mohnen, V.A. 1987. The mountain cloud chemistry and forest exposure monitoring 

project: program review report. A1 bany, NY: New York State University. 

Moll itor, A. V. ; Raynal , D. J. 1982. Atmospheric deposition and ionic input in 

Adirondack forests. Journal of the Air Pol 1 ution Control Association 

33: 1032-1036. 

Mooi, J. 1980. Influence of ozone on growth of two poplar cultivars. Plant 

Di sease 64: 772-773. 

Mudd, J. B. ; Kozlowski , T.T. , eds. 1975. Responses of plants to air pollution. 

Mew York: Academic Press. 

Muel 1 er, P. K. ; Hidy, G. M. 1983. The sul fate regional experiment (SURE): report 

of findings. Report no. EA-1901. Palo Alto, CA: Electric Power Research 

Institute. 3 vols. 

Muel ler, P. K. ; Hidy, G. M. ; Warren, K. [and others]. 1980. The occurrence of 

atmospheric aerosols in the northeastern United States. Annals of the New 

York Academy of Sciences 338:463-482. 

National Acid Precipitation Assessment Program. 1987. Interim assessment: the 

causes and effects of acidic deposition, vol. 1. executive summary. 

Washington: Office of the Director of Research. 

National Atmospheric Deposition Program. 1985. Annual data summary: precipi tation chemistry in the United States - 1983. Fort Collins, CO: Colorado 

State University, Natural Resources Ecology Laboratory. 

162 
m0, cfi 

nra zu -.'- 1 V, 

m -..z u 3no* 

4u- 

-Pa, 

I-5T 

-P* '-4 4. mJ 

a*' i-' lucu 

3 

0 H V, 
Patton, R. L. 1981. Effects of ozone and sulfur dioxide on height and stem 

specific gravity of Popul us hybrids. Forest Service Res. Pap. NE-471, 

Broomall, PA: Northeastern Forest Experiment Station. [Available from 

NTIS, 5285 Port Royal Road, Springfield, VA 22161; P881-1712741 

Phil 1 ips, S.O. ; Skel ly, J.M. ; Burkhart, H. E. 1977a. Growth fluctuation of 

loblolly pine due to periodic air pollution levels: interaction of rainfall and age. Phytopathology 67:716-720. 

Phil1 ips, 5.0. ; Skelly, J.M. ; Burkhart, H. E. 1977b. Eastern white pine exhibits 

growth retardation by f 1 uctuati ng air pol 1 utant level s: interaction of 

rainfall, age, and symptom expression. Phytopathology 67:721-725. 

Phipps, R. L. ; Whiton, J. C. 1988. Decline in long-term growth trends of white 

oak. Canadian Journal of Forest Research 18: 24-32. 

Pinkerton, J. E. ; Lefohn, A. S. 1986. Characterization of ambient ozone concentrations in commercial timberlands using available monitoring data. Tappi 

69: 58-62. 

Pye, J.M. 1988. Impact of ozone on the growth and yield of trees: a review. 

Journal of Environmental Quality 17347-360. 

Reich, P. B. 1987. Quantifying plant response to ozone: a unifying theory. Tree 

Physiology 3: 63-91. 

Reich, P. B. ; Amundson, R. G. 1985. Ambient levels of ozone reduce net photosynthesis in tree and crop species. Science 230:566-570. 

Reinert, R.A. ; Gray, T. N. 1981. The response of radish to nitrogen dioxide, 

sulfur dioxide, and ozone, alone and in combination. Journal of Environmental Qua1 i ty 10: 240-243. 

Renwick, J.A.A. ; Potter, J. 1981. Effects of sulfur dioxide on volatile terpene 

emission from balsam fir. Journal of the Air Pollution Control Association 

31: 65-66. 

Rhoads, A. ; Harkov, R. ; Brennan, E. 1980. Trees and shrubs relatively insensitive to oxidant pol 1 ution in New Jersey and southeastern Pennsylvania. 

Plant Disease 64:1106-1108. 

Roose, M. L. ; Bradshaw, A. D. ; Roberts, T. M. 1982. Evolution of resistance to 

gaseous air pol lutants. pp. 379-409. In: Unsworth, M. H. ; Ormrod, D. P. , eds. Effects of gaseous air pol 1 ution in agriculture and horticulture. 

Boston, MA: Butterworth Scientific. 

Rowland, A. ; Murray, A. J.S. ; Wellburn, A. R. 1985. Oxides of nitrogen and their 

impact upon vegetation. Reviews on Environmental Health 5295-342. 

Rudinsky, J. A. 1962. Ecology of scolytidae. Annual Review of Entomology 

7: 327-348. 
SAROAD, Storage and Retrieval of Aerometric Data [data base]. 1985. [Data file 

for 19791. Research Triangle Park, NC: U. S. Environmental Protection 

Agency, Office of Air Quality Planning and Standards. Disc; ASCII. 

Scheffer, T.C.; Hedgecock, G.G. 1955. Injury to northwestern forest trees by 

sulfur dioxide from smelters. USDA Tech. Bull. 1117. Washington: U.S. 

Department of Agriculture, Forest Service. 

Schier, G.A. 1985. Response of red spruce and balsam fir seedlings to aluminum 

toxicity i n nutrient sol uti ons. Canadian Journal of Forest Research 

15: 29-33. 

Schuett, P. ; Cowl i ng, E. B. 1985. Waldsterben--a general decl i ne of forests in 

central Europe: symptoms, development, and possible causes of a beginning 

breakdown of forest ecosystems. Plant Disease 69548-558. 

Sharpe, P.J.H.; Newton, R.J.; Spence, R.D. 1986. Forest pests: the role of 

phloem osmotic adjustment in the defensive response of conifers to bark 

beetle attack. pp. 113-131. In: Hennessey, T.C.; Dougherty, P.M.; Kossuth, 

S. V. [and others], eds. Stress physiology and forest productivity: 

proceedings of the physiology working group technical session, Society of 

American Foresters national convention; 1985 July; Fort Collins, CO. 

Dordrecht, The Netherlands: Martinus Mijhoff Publishers, 

Shortle, W. C. ; Smith, K.T. 1988. Aluminum-induced calcium deficiency syndrome 

i n decl i ni ng red spruce. Science 240: 1017-1018. 

Shriner, D. S. ; Cowling, E. B. 1980. Effects of rainfall acidification on plant 

pathogens. pp. 435-442. In: Hutchinson, T. C. ; Havas, M. , eds. Effects of 

acid precipitation on terrestrial ecosystems. New York: Plenum Press. 

Shugart, H. H. 1987, Dynamic ecosystem consequences of tree birth and death 

patterns: a set of computer models predicts long-term behavior of forests. 

BioScience 37: 596-602. 

Siccama, T.G.; Bliss, M.; Vogelmann, H.W. 1982. Decline of red spruce in the 

Green Mountains of Vermont. Bulletin of the Torrey Botanical Club 

109: 163-168. 

Sij, J.W. ; Swanson, C.A. 1974. Short-term kinetic studies on the inhibition of 

photosynthesis by sul fur dioxide. Journal of Environmental Qua1 i ty 

3: 103-107. 

Skiffington, R.A. ; Wilson, E. J. 1988. Excess nitrsgen deposition: Issues for 

consideration. Environmental Pollution 54: 159-184. 

Skelly, J.M. 1980. Photochemical oxidant impact on Mediterranean and temperate 

forest ecosystems: real and potential effects. pp. 38-50. In: Mi 11 er, 

P. R. , ed. Proceedings of symposium on effects of air pol lutants on Mediterranean and temperate forest ecosystems; 1980 June; Riverside, CA. Gen. 

Tech. Rep. PSW-43. Berkeley, CA: U. S. Department of Agricui ture, Forest 

Service, Pacific Southwest Forest and Range Experiment Station. [Available 

from NTIS, 5285 Port Royal Road, Springfield, VA 22161; PB81-1337201 
Skel ly, J. M. ; Moore, L. D. ; Stone, 1. L. 1972. Symptom expression of eastern 

white pine located near a source of oxides of nitrogen and sulfur dioxide. 

Plant Disease Reporter 56:3-6, 

Skel ly, J.M. ; Yang, Y. -5. ; Chevone, B. I. [and others]. 1984. Ozone concentrations and their influence on forest species in the Blue Ridge Mountains of 

\li rgi nia. pp. 143-159, In: Davis, D, D. ; Mi 1 len, A, A. ; Dochi nger, L. eds, 

Air pollution and the productivity of the forest: proceedings of the 

symposi um; 1983 October; Washi ngton, DC. Arl i ngton, VA: Izaak Wa1 ton 

league of America. 

Smith, W.H. 1974. Air pollution--effects on the structure and function of the 

temperate forest ecosystem. Environmental Pollution 6:111-129. 

Smith, W. W. 1980. Air p o l l u t i o n - - a 20th century a1 logenic influence on forest 

ecosystems. pp. 79-87. In: Miller, P. R., ed. Proceedings of the symposium 

on effects of air pollutants on Mediterranean and temperate forest 

ecosystems; 1980 June; Riverside, CA. Gen. Tech. Rep. PSW-43. Berkeley, 

CA: U.S. Department of Agriculture, Forest Service, Pacific Southwest 

Forest and Range Experiment Station. [Available from NTIS, 5285 Port Royal 

Road, Springfield, VA 22161; PB81-1337201 

Smith, W. ti. 1981. Air pol 'iution and forests: interactions between air contaminants and forest ecosystems. New York: Spri nger-Verl ag. 

Soderlund, R.; Svensson, B.H. 1976. The global nitrogen cycle. In: Svensson, 

B. H. ; Soderlund, R. , eds. Nitrogen, phosphorus and sulphur-global cycles: 

SCOPE report 7. Ecological Bull eti ns 22: 23-73. 

Srivastava, H. S. ; Jolliffe, P.A. ; Runeckles, V.C. 1975a. Inhibition of gas 

exchange in bean leaves by NO2. Canadian Journal of Botany 53466-474. 

Srivastava, H. S. ; Jolliffe, P.A. ; Runeckf es, KC. 1975b. The effects of 

environmental conditions on the inhibition of leaf gas exchange by NO2. 

Canadian Journal of Botany 53:475-482. 

Stark, R.W.; Cobb, F.W. 1969. Smog injury, root diseases and bark beetle damage 

in ponderosa pine. California Agriculture 23:13-15. 

Stark, R. W. ; Miller, P. R. ; Cob, F. W. , Jr. [and others J. 1968. Photochemical 

oxidant injury and bark beetle (Coleoptera: Scolytidae) infestation of 

ponderosa pine. Hi lgardia 39: 121-126. 

Stensland, 6.9. ; Whelpdale, B. M. ; Oehlert, 6. 1986. Precipitation chemistry. 

pp. 1213-.a99. In: kid deposition: long-term trends. Washi ngton: National 

Academy Press. 

Stone, L. L. ; Skel l y, J.M. 3.974. The growth of two forest tree species adjacent 

to a periodic source of air polluLion. Phytopathology 641773-778, 

Summers, P. W. ; Bowersox, V. C. ; Stensland, G. J. 1986. The geographical 

ds'stribution (and temporal variations of acidic deposition in eastern 

North America, Water, Air, and Soi 1 Pol 1 ution 31: 523-535. 
Sunquist, F. 1988. Zeroing in on keystone species. International Wildlife 

18: 18-23. 

Sze, N. D. ; Ko, M. K. W. 1980. Photochemistry of COS, CS,, CH3SCH3 and H2S: 

implications for the atmospheric sulfur cycle. Atmospheric Environment 

14: 1223-1239, 

Tabatabai, M-A. 1985. Effect of acid rain on soils. CRC Crjt-ical Reviews in 

Envi ronment Control 15: 65-110. 

Taylor, G.E., Jr.; Hanson, P.J.; Baldocchi, D.D. 1988. Pollutant deposition to 

individual leaves and piant canopies: sites of regulation and relationship 

to injury. pp. 227-257. In: Heck, W.W.; Taylor, OX.; Tingey, D.T., eds. 

Assessment of crop loss from air pollutants, Mew York: El sevier 

Pub1 ishers. 

Taylor, G. E. , Jr. ; McLaughl in, S. B. , Jr. ; Shriner, D. S. [and others]. 1983. The 

f 1 ux of sul fur-contai ni ng gases to vegetation. Atmospheric Envi ronment 

17: 789-796. 

Taylor, G.E., Jr.; Norby, R.J. 1985. The significance of elevated levels of 

ozone on natural ecosystems of North America. pp. 152-175. In: Lee, S. D. , ed. Eva1 uation of the scientific basis for ozone/oxidants standards: 

proceedings of an APCA international specialty conference; 1984 November; 

Houston, TX. (APCA transactions: TR-4). Pittsburgh, PA: Ai r Pol 1 ution 

Control Association. 

Teh, K.H. ; Swanson, LA. 1982. Sulfur dioxide inhibit-r'on of translocation in 

bean plants. Plant Physiology 69: 88-92. 

Thompson, C.R.; Kats, G.; Hensel, E.G. 1971. Effects of ambient levels of NO, 

on navel oranges. Environmental Science and Technology 5:1017-1019. 

Thomson, W. W. 1975. Effects of air pol 1 utants on plant ul trastructure. 

pp. 179-194. In: Mudd, J. B. ; Kozl owski, T. T. , eds. Responses of plants to 

air pollution. (Kozlowski , T. T. , ed. Physiological ecology: a series of 

monographs, texts, and treatises). New York: Academic Press. 

Tingey, D.T. 1974. Ozone induced alterations in the metabolite pools and enzyme 

activities of plants. pp. 40-57. In: Dugger, M., ed. Air pollution effects 

on plant growth. (American Chemical Society symposium series 3). 

Washington: American Chemical Society. 

Tingey, D.T. ; Blum, U. 1973. Effects of ozone on soybean nodules. Journal of 

Environmental Qua1 i ty 2: 341-342, 

Tingey, D.T.; Fites, R.C.; Llickliff, C. 1976b. Differential foliar sensitivity 

of soybean cultivars to ozone associated with differential enzyme activities. Physiological Plant Pathology 37: 69-72. 

Tingey, D.T. ; Taylor, G. E. , Jr. 1982. Variation in plant response to ozone: a 

conceptual model of physiological events, pp. 113-X38, In: Unsworth, M.H. ; 

Ormrod, D. P, , 4s. Effects of gaseous air pol lution -r"n agriculture and 

horticul lure. Boston, MA: Butterworth Seienti f i c, 
Tingey, D.T. ; Wil hour, R.G. ; Standley, C. 1976a. The effect of chronic ozone 

exposures on the metabol i te content of ponderosa pi ne seedl i ngs. Forest 

Science 22: 234-241. 

Tingey, D.T.J. ; Reinert, R. A. ; Dunning, J.A. [and others]. 1971. Vegetation 

injury from the interaction of nitrogen dioxide and sulfur dioxide. 

Phytopatholagy 61: 1506-151 1. 

Tomlinson, G.H., II. 1983. Air pollutants and forest decline. Environmental 

Science and Technology 17:246A-256A. 

Treshow, M. 1970. Environment and plant response. New York: McGraw-Hi 11 Book 

Co. 

Treshow, M. 1980a. Pollution effects on plant distribution. Environmental 

Conservation 7:279-286. 

Treshow, M. 1980b. Interactions of air pollutants and plant disease. 

pp. 103-109. In: Miller, P.R., ed. Proceedings of the symposium on effects 

of air pollutants on Mediterranean and temperate forest ecosystems; 1980 

June; Riverside, CA. Gen. Tech. Rep. PSW-43. Berkeley, CA: U. S. Department 

of Agriculture, Forest Service, Pacific Southwest Forest and Range Experiment Station. [Avas"l;able from NTIS, 5285 Port Royal Road, Springfield, VA 

22161; PB81-1337201 

Treshow, M. ; Anderson, F. K. ; Harner, F. 1967. Responses of Douglas-fir to 

elevated atmospheric fluorides. Forest Science 1334-120. 

Tukey, H. B. , Jr. 1970. The leaching of substances from plants. Annual Review of 

Plant Physiology 21:305-324. 

Tukey, H.B., Jr. 1980. Some effects of rain and mist on plants, with implications for acid precipitation. pp. 141-150. In: Hutchinson, T. C. ; Havas, 

M. , eds. Effects of acid precipitation on terrestrial ecosystems. Mew 

York: Plenum Press. 

Tukey, H. B. , Jr. ; Tukey, H. B. ; Wittwer, S. H. 1958. Loss of nutrients by fol iar 

1 eachi ng as determined by radioisotopes. Proceedings of the American 

Society for Horticultural Science 71:496-506. 

Turner, J. ; Lambert, M.J. 1980. Sulfur nutrition of forests. pp. 321-333. In: 

Shriner, D. S. ; Richmond, C. R. ; Lindberg, S. E. , eds. Atmospheric sulfur 

deposition: environmental impact and health effects. Ann Arbor, MI: Ann 

Arbor Science Publishers, Inc. 

Turner, R. S. ; 01 son, R. J. ; Brandt, C. C. 1986. Areas having soi l characteristics 

that may indicate sensitivity to acidic deposition under alternative 

forest damage hypotheses. Report no. ORNL/TM-9917. Oak Ridge, TN: Oak 

Ridge National Laboratory, Environmental Sciences Division publ. no. 2720. 

U.S. Congress. 1984. Acid rain and the transported air pollutants: implications 

for public policy. Report no. OTA-0-0204. Washington: Office of Technology 

Assessment. 
U. S. Department of Health, Education, and Welfare. 1968. Air quality data from 

the national ai r survei 1 lance networks and contri buti ng state and l ocal 

networks: 1966 edition* MAPCA Publication no. APTD 68-9. Durham, NC: 

National Air Pollution Control Administration. Pvailable from NTIS, 5285 

Port Royal Road, Springfield, VA 22161; PB-1951551 

U. S. Department of Health, Education, and Wel fare. l970. Air qual ity criteria 

for photochemical oxidants. Report no. NAPCA-PUB-AP-63. Washington: 

National Air Pollution Control Administration. [Available from NTIS, 5285 

Port Royal Road, Spri ngf ield, VA 22161; PB-190262/BA] 

U. S. Environmental Protection Agency. 1973. Effects of sulfur oxides in the 

atmosphere on vegetation; revised chapter 5 for air quality criteria for 

sulfur oxides. EPA report no. EPA/R3-73-030. Research Triangle Park, NC: 

National Environmental Research Center, Office of Research and Development. [Available from NTIS, 5285 Port Royal Road, Springfield, VA 22161; 

PB-2263141 

U. S. Environmental Protection Agency. 1982a Air qual i ty criteria for particulate matter and sulfur oxides. EPA report no. EPA-600/8-82-029aF-cF. 

Research Triangle Park, NC: Office of Health and Environmental Assessment, 

Envi rsnmental Criteria and Assessment Office. [Avai lable from NTIS, 5285 

Port Royal Road, Springfield, VA 22161; PB84-1567771 

U. S. Environmental Protection Agency. 1982b. Air qual i ty criteria for oxides of 

nitrogen. EPA report no. EPA-600/8-82-026F. Research Triangle Park, NC: 

Office of Health and Environmental Assessment, Environmental Criteria 

and Assessment Office. [Available from NTIS, 5285 Port Royal Road, 

Springfield, VA 22161; PB83-1633373 

U. S. Environmental Protection Agency. 1986. Air qual i ty criteria for ozone and 

other photochemical oxidants: vol. 2 and vol. 3. [draft final]. EPA report 

nos. EPA-600/8-84-020bF and EPA-600/8-84/020bF-cF. Research Triangle Park, 

NC: Office of Health and Environmental Assessment, Environmental Criteria 

and Assessment Office. [Available from NTIS, 5285 Port Royal Road, 

Springfield, VA 22161; PB87-1429491 

Umbach, D.M. ; Davis, D. D. 1984. Severity and frequency of S02-induced leaf 

necrosis on seedlings of 57 tree species. Forest Science 30:587-596. 

Unsworth, M. H. ; Biscoe, P.V. ; Black, V. 1976. Analysis of gas exchange between 

plants and polluted atmospheres. pp. 5-16. In: Mansfield, T.A., ed. 

Effects of air pol 1 utants on plants. (Society for Experimental Biology 

semi nar series 1). Cambridge, Uni ted Ki ngdom: Cambridge University Press. 

van Hove, L. W. A. ; Koops, A. 3. ; Adema, E. H. [and others]. 1987, Analys-i s of 

the uptake of atmospheric ammonia by leaves of Phaseolus L, 

Atmospheric Environment 21: 1759-1763. 

Vogl, R.3. 1980. The ecological factors that produce perturbation-dependent 

ecosystems. pp. 63-94. In: Cairns, J. , Jr. , ed. The recovery process of 

damaged ecosystems. Ann Arbor, MI: Ann Arbor Science Publishers, Inc. 
Wargo, P. M. 1972. Defol iation-indud chemical chaj-rges in sugar maple roots 

stimulate growth of Armillaria mellea. Phytopathology 62:L278-1283. 

Maring, R.H. 1987. Nitrate pollution: a particular danger to boreal and 

subalpine coniferous forests. pp. 93-IO5. In: Fujimori , T. ; Kimura, tul. , eds. Human impacts and management of mountain forests: [proceedings of a 

symposi um]. Ibaraki , Japan: Forestry and Forest Products Research 

Institute. 

Waring, R.H.; Pitman, G.B. 1985. Modifying fodgepole pine stands to ckange 

susceptibility to mountain pine beetle attack. Ecology 66:883-897. 

Waring, R.H.; Schlesinger, W.H. 1985. Forest ecosystems: concepts and management. pp. 7-77. Orlando, FL: Academic Press, Inc, 

Weathers, K, C. ; Likens, G, E. ; Bormann, F. H. [and others]. 1986. A regional 

acidic cloud/fog water event in the eastern United States. Nature 

319: 657-658. 

Webb, W. L. 1981. Relation of starch content to conifer mortality and growth 

1 oss after defol iation by the Douglas-f ir Tussock moth. Forest Science 

27: 224-232, 

Wel'densaul , T. C. ; McClenahen, J. R. 1986. Soil-air pollutant interactions. 

pp. 397-414. In: Legge, A. W. ; Krupa, S. V. , eds, Air pol lutants and their 

effects on the terrestrial ecosystem. (Nriagu, J. O., ed. Advances in 

environmental science and technology: v. 18). New York: John Wiley & Sons. 

Wellburn, A. R. 1982. Effects of SO, and NO, on metabolic function. pp. 169-187. 

In: Unsworth, M. H. ; Ormrod, D. P. , eds. Effects of gaseous air pollution in 

agriculture and horticul ture. London, United Kingdom: Butterworth 

Scientific. 

Wellburn, A.R. 1985. SO2 effects on stromal and thylakoid function. 

pp. 133-147. In: Winner, W. E. ; Mooney, H.A. ; Goldstein, R.A. , eds. Sulfur 

dioxide and vegetation: physiology, ecology, and policy issues. Stanford, 

CA: Stanford University Press. 

Wesely, M. L. ; Hicks, B. B. 1977. Some factors that affect the deposition rates 

of sulfur dioxide and similar gases on vegetation. Journal of the Air 

Pol 1 ution Control Association 27: 1110-1116. 

White, K. I. ; Will, A. 6. ; Bennett, J. H. 1974. Synergistic inhibition of apparent 

photosynthesis rate of a1 fa1 fa by combinations of sulfur dioxide and 

nitrogen dioxide. Environmental Science and Technology 8:574-576* 

Whitmore, M. E. ; Freer-Smith, P.H. 1982. Growth effects of SO, and/or NO2 on 

woody plants and grasses during spring and summer. Nature 300:55-57. 

Winner, W. E. ; Atkinson, C. J. 1986. Absorption of air pollution by plants, and 

consequences for growth. U.S. Department of Energy, Idaho Operations 

Office [Report] TREE 1:15-18. 
Winner, W. E. ; Bewley, J. D. (19783) Terrestrial mosses as bioindicators of SO2 

pollution stress: synecological analysis and the index of atmospheric 

purity. Oecologia 35:221-230. 

Winner, W. E. ; Bewley, J. D. 1978b. Contrasts between hryophyte and vascular 

plant synecological responses in an SO2-stressed white spruce association 

in central A1 berta. Oecol ogi a 33: 311-325. 

Winner, W. E. ; Bewley, J.D. ; Krouse, H. R. [and others]. 1978, Stable sulfur 

isotope analysis of SO, pol 1 ution impact on vegetation. Oecologia 

36: 351-361. 

Winner, W.E.; Koch, G.W.; Mooney, H.A. 1982. Ecology of SO2 resistance: 4. 

predicting metabolic responses of fumigated shrubs and trees. Oecologia 

52: 16-22. 

Winner, W.E. ; Mooney, H.A. 198Oa. Ecology of SO, resistance: 1. effects of 

fumigations on gas exchange of deciduous and evergreen shrubs. Oecologia 

44: 290-295. 

Winner, W. E. ; Mooney, H. A. 1980b. Ecology of SO2 resistance: 2. photosynthetic 

changes of shrubs in relation to SO2 absorption and stomata1 behavior. 

Oecol ogi a 44: 296-302. 

Winner, W. E. ; Mooney, H.A. ; Williams, K. [and others]. 1985. Measuring and 

assessing SO, effects on photosynthesis and plant growth. p. 118-132. In: 

Winner, W.E.; Mooney, H.A.; Goldstein, R.A., eds. Sulfur dioxide and 

vegetation: physiology, ecology, and policy issues. Stanford, CA: Stanford 

University Press. 

Wol ff, G. T. ; Lioy, P. J. 1980. Development of an ozone river associated with 

synoptic scale episodes i n the eastern Uni Led States. Environmental 

Science and Technology 14:1257-1260. 

Wolff, G.T. ; Lioy, P. J. ; Taylor, R. S. 1987. The diurnal variations of ozone at 

different altitudes on a rural mountain in the eastern United States. 

Journal of the Air Pollution Control Association 37:45-48. 

Wolff, G.T. ; Lioy, P. J. ; Wight, G.D. [and others]. 1977. An investigation of 

long-range transport of ozone across the midwestern and eastern United 

States. Atmospheric Environment 11:797-802. 

Wolff, G.T. ; Lioy, P. J. ; Wight, G.D. 1980. Transport of ozone associated with 

an air mass. Journal of Environmental Science and Health Part A 

A15 : 183- 199, 

Wood, T.; Bormann, F.H. 1974. The effects of an artificial acid mist upon the 

growth of Betula Bri tt, Environmental Pol 1 ution 7: 259-268. 

Wood, T. ; Bormann, F.H. 1975. Increases in foliar leaching caused by acidification of an artificial mist, Ambio 4:169-171. 

Woodman, J.N.; Cowling, E.B. 1987. Airborne chemicals and forest health. 

Envi ronmental Science and Techno1 ogy 21: 120-126. 
Woodwell, G.M. 1970. Effects of pollution on the structure and physiology of 

ecosystems: changes in natural ecosystems caused by many different types 

of disturbances are similar and predictable. Science 168:429-433. 

Yang, Y.-S. ; Skelly, J.M. ; Chevone, B. I. [and others]. 1983. Effects of longterm ozone exposure on photasynthesis and dark respiration of easLern 

white pine. Envi ronmental Science and Technology 17: 371-373. 

Ziegler, I. 1975. The effect of air-polluting gases on plant metabolism. 

Environmental Qua1 i ty and Safety 2: 182-208. 

The Forest Sewice, US. Department sf 

Agriculture, is dedicated to the principle of 

multiple use management of the Nation" forest resources 

For sustained yields of wood, water, forage, wildlife, and 

recreation. Through forestry research, cooperation with the 

States and private forest owners, and management sf the 

National Forests and National Grasslands, it strives-as 

directed by Congress-to provide increasingly greater 

service to a growing Nation. 

USDA policy prohibits discrimination because of race, 

color, national origin, sex, age, religion, or handicapping 

condition. Any person who believes he or she has been 

discriminated against in any USDA-related activity should 

immediabely contact the Secretary of Agriculture, 

Washington, C)@ 20250.