Nitrogen, oxides of (EHC 188, 1997, 2nd edition)

UNITED NATIONS ENVIRONMENT PROGRAMME
INTERNATIONAL LABOUR ORGANISATION
WORLD HEALTH ORGANIZATION

INTERNATIONAL PROGRAMME ON CHEMICAL SAFETY

ENVIRONMENTAL HEALTH CRITERIA 188

Nitrogen Oxides

(Second Edition)

This report contains the collective views of an international group of
experts and does not necessarily represent the decisions or the stated
policy of the United Nations Environment Programme, the International
Labour Organisation, or the World Health Organization.

First draft prepared by Drs J.A. Graham, L.D. Grant, L.J. Folinsbee,
D.J. Kotchmar and J.H.B. Garner, US Environmental Protection Agency

Published under the joint sponsorship of the United Nations
Environment Programme, the International Labour Organisation, and the
World Health Organization, and produced within the framework of the
Inter-Organization Programme for the Sound Management of Chemicals.

World Health Organization
Geneva, 1997

The International Programme on Chemical Safety (IPCS) is a joint
venture of the United Nations Environment Programme, the International
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is to promote coordination of the policies and activities pursued by
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WHO Library Cataloguing in Publication Data

Nitrogen oxides - 2nd ed.

(Environmental health criteria ; 188)

1.Nitrogen dioxide 2.Nitrogen oxides
I.Series

ISBN 92 4 157188 8 (NLM Classification: WA 754)
ISSN 0250-863X

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CONTENTS

ENVIRONMENTAL HEALTH CRITERIA FOR NITROGEN OXIDES

Preamble

1. SUMMARY

1.1. Nitrogen oxides and related compounds
1.1.1. Atmospheric transport
1.1.2. Measurement
1.1.3. Exposure
1.2. Effects of atmospheric nitrogen species, particularly
nitrogen oxides, on vegetation
1.3. Health effects of exposures to nitrogen dioxide
1.3.1. Studies of the effects of nitrogen compounds on
experimental animals
1.3.1.1 Biochemical and cellular mechanisms of
action of nitrogen oxides
1.3.1.2 Effects on host defence
1.3.1.3 Effects of chronic exposure on the
development of chronic lung disease
1.3.1.4 Potential carcinogenic or co-carcinogenic
effects
1.3.1.5 Age susceptibility
1.3.1.6 Influence of exposure patterns
1.3.2. Controlled human exposure studies on nitrogen
oxides
1.3.3. Epidemiology studies on nitrogen dioxide
1.3.4. Health-based guidance values for nitrogen dioxide

2. PHYSICAL AND CHEMICAL PROPERTIES, AIR SAMPLING AND ANALYSIS,
TRANSFORMATIONS AND TRANSPORT IN THE ATMOSPHERE

2.1. Introduction
2.1.1. The nomenclature and measurement of atmospheric
nitrogen species
2.2. Nitrogen species and their physical and chemical properties
2.2.1. Nitrogen oxides
2.2.1.1 Nitric oxide
2.2.1.2 Nitrogen dioxide
2.2.1.3 Nitrous oxide
2.2.1.4 Other nitrogen oxides
2.2.2. Nitrogen acids
2.2.2.1 Nitric acid
2.2.2.2 Nitrous acid
2.2.3. Ammonia
2.2.4. Ammonium nitrate
2.2.5. Peroxyacetyl nitrate
2.2.6. Organic nitrites and nitrates

2.3. Sampling and analysis methods
2.3.1. Nitric oxide
2.3.1.1 Nitric oxide continuous methods
2.3.1.2 Passive samplers for NO
2.3.1.3 Calibration of NO analysis methods
2.3.1.4 Sampling considerations for NO
2.3.2. Nitrogen dioxide
2.3.2.1 Chemiluminescence (NO + O₃)
2.3.2.2 Chemiluminescence (luminol)
2.3.2.3 Laser-induced fluorescence and tuneable
diode laser absorption spectrometry
2.3.2.4 Wet chemical methods
2.3.2.5 Other methods
2.3.2.6 Passive samplers
2.3.2.7 Calibration
2.3.3. Total reactive odd nitrogen
2.3.4. Peroxyacetyl nitrate
2.3.5. Other organic nitrates
2.3.6. Nitric acid
2.3.7. Nitrous acid
2.3.8. Dinitrogen pentoxide and nitrate radicals
2.3.9. Particulate nitrate
2.3.10. Nitrous oxide
2.3.11. Summary
2.4. Transport and transformation of nitrogen oxides in the air
2.4.1. Introduction
2.4.2. Chemical transformations of oxides of nitrogen
2.4.2.1 Nitric oxide, nitrogen dioxide and ozone
2.4.2.2 Transformations in indoor air
2.4.2.3 Formation of other oxidized nitrogen
species
2.4.3. Advection and dispersion of atmospheric nitrogen
species
2.4.3.1 Transport of reactive nitrogen species
in urban plumes
2.4.3.2 Air quality models
2.4.3.3 Regional transport
2.5. Conversion factor for nitrogen dioxide
2.6. Summary

3. SOURCES, EMISSIONS AND AIR CONCENTRATIONS

3.1. Introduction
3.2. Sources of nitrogen oxides
3.2.1. Sources of NO_x emission
3.2.1.1 Fuel combustion
3.2.1.2 Biomass burning
3.2.1.3 Lightning
3.2.1.4 Soils
3.2.1.5 Oceans

3.2.2. Removal from the ambient environment
3.2.3. Summary of global budgets for nitrogen oxides
3.3. Ambient concentrations of nitrogen oxides
3.3.1. International comparison studies of NO_x
concentrations
3.3.2. Example case studies of NO_x and NO₂
concentrations
3.4. Occurrence of nitrogen oxides indoors
3.4.1. Indoor sources
3.4.1.1 Gas-fuelled cooking stoves
3.4.1.2 Unvented gas space heaters and water
heaters
3.4.1.3 Kerosene space heaters
3.4.1.4 Wood stoves
3.4.1.5 Tobacco products
3.4.2. Removal of nitrogen oxides from indoor environments
3.5. Indoor concentrations of nitrogen oxides
3.5.1. Homes without indoor combustion sources
3.5.2. Homes with combustion appliances
3.5.3. Homes with combustion space heaters
3.5.4. Indoor nitrous acid concentrations
3.5.5. Predictive models for indoor NO₂ concentration
3.6. Human exposure
3.7. Exposure of plants and ecosystems

4. EFFECTS OF ATMOSPHERIC NITROGEN COMPOUNDS (PARTICULARLY NITROGEN
OXIDES) ON PLANTS

4.1. Properties of NO_x and NH_y
4.1.1. Adsorption and uptake
4.1.2. Toxicity, detoxification and assimilation
4.1.3. Physiology and growth aspects
4.1.4. Interactions with climatic conditions
4.1.5. Interactions with the habitat
4.1.6. Increasing pest incidence
4.1.7. Conclusions for various atmospheric nitrogen
species and mixtures
4.1.7.1 NO₂
4.1.7.2 NO
4.1.7.3 NH₃
4.1.7.4 NH₄⁺ and NO₃^- in wet and occult
deposition
4.1.7.5 Mixtures
4.1.8. Appraisal
4.1.8.1 Representativity of the data
4.1.9. General conclusions
4.2. Effects on natural and semi-natural ecosystems
4.2.1. Effects on freshwater and intertidal ecosystems
4.2.1.1 Effects of nitrogen deposition on
shallow softwater lakes

4.2.1.2 Effects of nitrogen deposition on lakes
and streams
4.2.2. Effects on ombrotrophic bogs and wetlands
4.2.2.1 Effects on ombrotrophic (raised) bogs
4.2.2.2 Effects on mesotrophic fens
4.2.2.3 Effects on fresh- and saltwater marshes
4.2.3. Effects on species-rich grasslands
4.2.3.1 Effects of nitrogen on calcareous
grasslands
4.2.3.2 Critical loads for nitrogen in
calcareous grasslands
4.2.3.3 Comparison with other semi-natural
grasslands
4.2.4. Effects on heathlands
4.2.4.1 Effects on inland dry heathlands
4.2.4.2 Effects of nitrogen on inland wet
heathlands
4.2.4.3 Effects of nitrogen on arctic and alpine
healthlands
4.2.4.4 Effects on herbs of matgrass swards
4.2.5. Effects of nitrogen deposition on forests
4.2.5.1 Effects on forest tree species
4.2.5.2 Effects on tree epiphytes, ground
vegetation and ground fauna of forests
4.2.6. Effects on estuarine and marine ecosystems
4.2.7. Appraisal and conclusions

5. STUDIES OF THE EFFECTS OF NITROGEN OXIDES ON EXPERIMENTAL ANIMALS

5.1. Introduction
5.2. Nitrogen dioxide
5.2.1. Dosimetry
5.2.1.1 Respiratory tract dosimetry
5.2.1.2 Systemic dosimetry
5.2.2. Respiratory tract effects
5.2.2.1 Host defence mechanisms
5.2.2.2 Lung biochemistry
5.2.2.3 Pulmonary function
5.2.2.4 Morphological studies
5.2.3. Genotoxicity, potential carcinogenic or
co-carcinogenic effects
5.2.4. Extrapulmonary effects
5.3. Effects of mixtures containing nitrogen dioxide
5.4. Effects of other nitrogen oxide compounds
5.4.1. Nitric oxide
5.4.1.1 Endogenous formation of NO
5.4.1.2 Absorption of NO
5.4.1.3 Effects of NO on pulmonary function,
morphology and host lung defence
function

5.4.1.4 Metabolic effects
5.4.1.5 Haematological changes
5.4.1.6 Biochemical mechanisms for nitric oxide
effects: reaction with iron and effects
on enzymes and nucleic acids
5.4.2. Nitric acid
5.4.3. Nitrates
5.5. Summary of studies of the effects of nitrogen compounds on
experimental animals

6. CONTROLLED HUMAN EXPOSURE STUDIES OF NITROGEN OXIDES

6.1. Introduction
6.2. Effects of nitrogen dioxide
6.2.1. Nitrogen dioxide effects on pulmonary function and
airway responsiveness to bronchoconstrictive agents
6.2.1.1 Nitrogen dioxide effects in healthy
subjects
6.2.1.2 Nitrogen dioxide effects on asthmatics
6.2.1.3 Nitrogen dioxide effects on patients
with chronic obstructive pulmonary
disease
6.2.1.4 Age-related differential susceptibility
6.2.2. Nitrogen dioxide effects on pulmonary host defences
and bronchoalveolar lavage fluid biomarkers
6.2.3. Other classes of nitrogen dioxide effects
6.3. Effects of other nitrogen oxide compounds
6.4. Effects of nitrogen dioxide/gas or gas/aerosol mixtures on
lung function
6.5. Summary of controlled human exposure studies of oxides of
nitrogen

7. EPIDEMIOLOGICAL STUDIES OF NITROGEN OXIDES

7.1. Introduction
7.2. Methodological considerations
7.2.1. Measurement error
7.2.2. Misclassification of the health outcome
7.2.3. Adjustment for covariates
7.2.4. Selection bias
7.2.5. Internal consistency
7.2.6. Plausibility of the effect
7.3. Studies of respiratory illness
7.3.1. Indoor air studies
7.3.1.1 St Thomas' Hospital Medical School
Studies (United Kingdom)
7.3.1.2 Harvard University - Six Cities Studies
(USA)
7.3.1.3 University of Iowa Study (USA)

7.3.1.4 Agricultural University of Wageningen
(The Netherlands)
7.3.1.5 Ohio State University Study (USA)
7.3.1.6 University of Dundee (United Kingdom)
7.3.1.7 Harvard University - Chestnut Ridge
Study (USA)
7.3.1.8 University of New Mexico Study (USA)
7.3.1.9 University of Basel Study (Switzerland)
7.3.1.10 Yale University Study (USA)
7.3.1.11 Freiburg University Study (Germany)
7.3.1.12 McGill University Study (Canada)
7.3.1.13 Health and Welfare Canada Study (Canada)
7.3.1.14 University of North Carolina Study (USA)
7.3.1.15 University of Tucson Study (USA)
7.3.1.16 Hong Kong Anti-Cancer Society Study
(Hong Kong)
7.3.1.17 Recent studies
7.3.2. Outdoor studies
7.3.2.1 Harvard University - Six City Studies
(USA)
7.3.2.2 University of Basel Study (Switzerland)
7.3.2.3 University of Wuppertal Studies
(Germany)
7.3.2.4 University of Tubigen (Germany)
7.3.2.5 Harvard University - Chestnut Ridge
Study (USA)
7.3.2.6 University of Helsinki Studies (Finland)
7.3.2.7 Helsinki City Health Department Study
(Finland)
7.3.2.8 Oulu University Study (Finland)
7.3.2.9 Seth GS Medical College Study (India)
7.4. Pulmonary function studies
7.4.1. Harvard University - Six City Studies (USA)
7.4.2. National Health and Nutrition Examination Survey
Study (USA)
7.4.3. Harvard University - Chestnut Ridge Study (USA)
7.4.4. Other pulmonary function studies
7.5. Other exposure settings
7.5.1. Skating rink exposures
7.6. Occupational exposures
7.7. Synthesis of the evidence for school-age children
7.7.1. Health outcome measures
7.7.2. Biologically plausible hypothesis
7.7.3. Publication bias
7.7.4. Selection of studies
7.7.4.1 Brief description of selected studies
7.7.4.2 Studies not selected for quantitative
analysis
7.7.5. Quantitative analysis

7.8. Synthesis of the evidence for young children
7.9. Summary

8. EVALUATION OF HEALTH AND ENVIRONMENT RISKS ASSOCIATED WITH
NITROGEN OXIDES

8.1. Sources and exposure
8.2. Evaluation of the effects of atmospheric nitrogen species
on the environment
8.2.1. Guidance values - critical levels for air
concentrations of nitrogen oxides
8.2.2. Environment-based guidance values - critical loads
for total nitrogen deposition
8.3. Evaluation of health risks associated with nitrogen oxides
8.3.1. Concentration-response relationships
8.3.2. Subpopulations potentially at risk
8.3.3. Derivation of health-based guidance values

9. CONCLUSIONS AND RECOMMENDATIONS FOR PROTECTION OF HUMAN HEALTH
AND THE ENVIRONMENT

10. FURTHER RESEARCH

REFERENCES

RESUME

RESUMEN

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WHO TASK GROUP ON ENVIRONMENTAL HEALTH CRITERIA FOR NITROGEN OXIDES

Members

Dr K. Bentley^*, Health and Environment Policy Section, Department
of Community Services and Health, Canberra ACT, Australia

Dr S. Dobson, Institute of Terrestrial Ecology, Monks Wood
Experimental Station, Abbots Ripton, Huntingdon, Cambridgeshire,
United Kingdom

Dr L. van der Eerden, Centre "De Bom" Wageningen, The Netherlands

Dr L. Folinsbee, Health Effects Research Laboratory, US Environmental
Protection Agency, Research Triangle Park, North Carolina, USA
(Rapporteur)

Dr L. Grant^*, National Center for Environmental Assessment, US
Environmental Protection Agency, Research Triangle Park, North
Carolina, USA

Mr L. Heiskanen, Health and Environment Policy Section, Department of
Community Services and Health, Canberra ACT, Australia

Mr G.M. Johnson, CSIRO, Division of Coal and Energy Technology, Centre
for Pollution Assessment and Control, North Ryde, NSW, Australia

Dr J. Kagawa, Professor of Hygiene and Public Health, Tokyo Women's
Medical College, Shinjuku-ku, Tokyo, Japan

Dr R.R. Khan, Ministry of Environment and Forests, Paryavaran Bhawan,
New Delhi, India

Dr D.B. Menzel, University of California, Department of Community &
Environment and Medicine, California, USA

Dr L. Neas, Department of Environmental Health, Environmental
Epidemiology Program, Harvard School of Public Health, Boston,
Massachusetts, USA

Dr S.E. Paulson, Department of Atmospheric Sciences, University of
California, Los Angeles, California, USA

Dr P.J.A. Rombout, Department for Inhalation Toxicology, National
Institute of Public Health and Environmental Hygiene, Bilthoven,
The Netherlands (Chairman)

^* Invited, but unable to attend

Dr W. Tyler, Veterinary Anatomy and Cell Biology, University of
California, California, USA

Dr K. Victorin, Karolinska Institute, Institute of Environmental
Medicine, Stockholm, Sweden

Dr A. Woodward, Department of Community Medicine, University of
Adelaide, Adelaide, Australia

Dr R. Ye, Deputy Director, National Environmental Protection Agency,
Xizhimennei Nanziaojie, Beijing, People's Republic of China

Observers

Professor M. Moore, National Research Centre for Environmental
Toxicology, Nathan, Australia

Dr M. Pain, Department of Thoracic Medicine, Royal Melbourne Hospital,
Melbourne VIC, Australia

Dr P. Psaila-Savona, WA Department of Health, Perth WA, Australia

Mr B. Taylor, Policy and Planning Group, Public and Planning Group,
Public Health Commission, Wellington, New Zealand

Mr B. Saxby, AGL Gas Companies, North Sydney NSW, New Zealand

Secretariat

Dr B.H. Chen, International Programme on Chemical Safety, World Health
Organization, Geneva, Switzerland (Secretary)

Dr M. Younes, WHO European Centre for Environment & Health, Bilthoven,
The Netherlands

ENVIRONMENTAL HEALTH CRITERIA FOR NITROGEN OXIDES

A WHO Task Group on Environmental Health Criteria for Nitrogen
Oxides met in Melbourne, Australia from 14 to 18 November 1994. The
meeting was hosted by the Clean Air Society of Australia and New
Zealand and the Victorian Departments of Health and Environment,
Australia. Dr B.H. Chen, IPCS, opened the meeting and welcomed the
participants on behalf of the Director, IPCS, and the three IPCS
cooperating organizations (UNEP/ILO/WHO). The Task Group reviewed and
revised the draft criteria monograph and made an evaluation of the
risks for human health and the environment from exposure to nitrogen
oxides.

The first draft of this monograph was prepared by Drs J.A.
Graham, L.D. Grant, L.J. Folinsbee, D.J. Kotchmar and J.H.B. Garner,
US EPA. Drs W.G. Ewald, T.B. McMullen and B.E. Tilton, US EPA,
contributed to the preparation of the first draft. The second draft
was prepared by Dr L.D. Grant incorporating comments received
following the circulation of the first draft to the IPCS Contact
Points for Environmental Health Criteria. Drs R. Bobbink, L. Van der
Eerden and S. Dobson prepared the final text of the environmental
section. Mr G.M. Johnson contributed to the final text of the
chemistry section.

Dr B.H. Chen and Dr P.G. Jenkins, both members of the IPCS
Central Unit, were responsible for the overall scientific content and
technical editing, respectively.

The efforts of all who helped in the preparation and finalization
of the document are gratefully acknowledged.

Financial support for this Task Group meeting was provided by the
Department of Community Services and Health, Australia, Victorian
Departments of Health and Environment, Australia, and the Clean Air
Society of Australia and New Zealand.

ABBREVIATIONS

ADP adenosine diphosphate
AM alveolar macrophages
AQG Air Quality Guidelines
BAL bronchoalveolar lavage
BHPN N-bis (2-hydroxypropyl) nitrosamine
CI confidence interval
CLM chemiluminescence method
COPD chronic obstructive pulmonary disease
ECD electron capture detection
FEF forced expiratory flow
FEV forced expiratory volume
FTIR Fourier transformed infrared
FVC forced vital capacity
GC gas chromatography
GDH glutamate dehydrogenase
(c)GMP (cyclic) guanosine monophosphate
GS glutamine synthetase
HNO₂ nitrous acid
HNO₃ nitric acid
LIF laser-induced fluorescence
MS mass spectrometry
N₂ nitrogen (elemental)
NH₃ ammonia
NH₄⁺ ammonium ion
NH_y the sum of NH₃ and NH₄⁺
NiR nitrate reductase
NK natural killer
NO nitric oxide
NO₂ nitrogen dioxide
NO₂^- nitrite ion
NO₃^- nitrate ion
N₂O nitrous oxide
N₂O₅ nitrogen pentoxide
NO_x nitric oxide plus nitrogen dioxide
NO_y gas-phase oxidized nitrogen species (except nitrous oxide)
NPSH non-protein sulfhydryl
NR nitrate reductase
O₃ ozone
PAN peroxyacetyl nitrate
PBzN peroxybenzoyl nitrate
PEF peak expiratory flow
PFC plaque-forming cell
PMN polymorphonuclear leukocyte
ppb parts per billion (10^-9)
ppm parts per million (10^-6)
ppt parts per trillion (10^-12)
pptv parts per trillion (by volume)
PSD passive sampling device

R_aw airway resistance
ROC reactive organic carbon
RUBISCO ribulose 1,5-biphosphate carboxylase
SD standard deviation
SES socioeconomic status
SG_aw specific airway conductance
SO₂ sulfur dioxide
SO_y sulfur oxides
SPM suspended particulate matter
SR_aw specific airway resistance
TDLAS tuneable diode laser absorption spectrometry
TSP total suspended particulate
VOC volatile organic carbon

1. SUMMARY

1.1 Nitrogen oxides and related compounds

Nitrogen oxides can be present at significant concentrations in
ambient air and in indoor air. The types and concentrations of
nitrogenous compounds present can vary greatly from location to
location, with time of day, and with season. The main sources of
nitrogen oxide emissions are combustion processes. Fossil fuel power
stations, motor vehicles and domestic combustion appliances emit
nitrogen oxides, mostly in the form of nitric oxide (NO) and some
(usually less than about 10%) in the form of nitrogen dioxide (NO₂).
In the air, chemical reactions occur that oxidize NO to NO₂ and other
products. There are also biological processes that liberate nitrogen
species from soils, including nitrous oxide (N₂O). Emissions of N₂O
can cause perturbation of the stratospheric ozone layer.

Human health may be affected when significant concentrations of
NO₂ or other nitrogenous species, such as peroxyacetyl nitrate (PAN),
nitric acid (HNO₃), nitrous acid (HNO₂), and nitrated organic
compounds, are present. In addition, nitrates and HNO₃ may cause
health effects and significant effects on ecosystems when deposited on
the ground.

The sum of NO and NO₂ is generally referred to as NO_x. Once
released into the air, NO is oxidized to NO₂ by available oxidants
(particularly ozone, O₃). This happens rapidly under some conditions
in outdoor air; in indoor air, it is generally a much slower process.
Nitrogen oxides are a controlling precursor of photochemical oxidant
air pollution resulting in ozone and smog formation; interactions of
nitrogen oxides (except N₂O) with reactive organic compounds and
sunlight form ozone in the troposphere and smog in urban areas.

NO and NO₂ may also undergo reactions to form a range of other
oxides of nitrogen, both in indoor and outdoor air, including HNO₂,
HNO₃, nitrogen trioxide (NO₃), dinitrogen pentoxide (N₂O₅), PAN
and other organic nitrates. The complex range of gas-phase nitrogen
oxides is referred to as NO_y. The partitioning of oxides of nitrogen
among these compounds is strongly dependent on the concentrations of
other oxidants and on the meteorological history of the air.

HNO₃ is formed from the reaction of OH^- and NO₂. It is a
major sink for active nitrogen and also a contributor to acidic
deposition. Potential physical and chemical sinks for HNO₃ include
wet and dry deposition, photolysis, reaction with OH radicals, and
reaction with gaseous ammonia to form ammonium nitrate aerosol.

PANs are formed from the combination of organic peroxy radicals
with NO₂. PAN is the most abundant organic nitrate in the
troposphere and can serve as a temporary reservoir for reactive
nitrogen, which may be regionally transported.

The NO₃ radical, a short-lived NO_y species that is formed in
the troposphere primarily by the reaction of NO₂ with O₃, undergoes
rapid photolysis in daylight or reaction with NO. Appreciable
concentrations are observed during the night.

N₂O₅ is primarily a night-time constituent of ambient air as it
is formed from the reaction of NO₃ and NO₂. In ambient air, N₂O₅
reacts heterogeneously with water to form HNO₃, which in turn is
deposited.

N₂O is ubiquitous because it is a product of natural biological
processes in soil. It is not known, however, to be involved in any
reactions in the troposphere. N₂O participates in upper atmospheric
reactions contributing to stratospheric ozone (O₃) depletion and is
also a relatively potent greenhouse gas that contributes to global
warming.

1.1.1 Atmospheric transport

The transport and dispersion of the various nitrogenous
species in the lower troposphere is dependent on both meteorological
and chemical parameters. Advection, diffusion and chemical
transformations combine to dictate the atmospheric residence times.
In turn, atmospheric residence times help determine the geographic
extent of transport of given species. Surface emissions are dispersed
vertically and horizontally through the atmosphere by turbulent mixing
processes that are dependent to a large extent on the vertical
temperature structure and wind speed.

As the result of meteorological processes, NO_x emitted in the
early morning hours in an urban area typically disperses vertically
and moves downwind as the day progresses. On sunny summer days, most
of the NO_x will have been converted to HNO₃ and PAN by sunset, with
concomitant formation of ozone. Much of the HNO₃ is removed by
deposition as the air mass is transported, but HNO₃ and PAN carried
in layers aloft (above the nighttime inversion layer but below a
higher subsidence inversion) can potentially be transported long
distances in oxidant-laden air masses.

1.1.2 Measurement

There are a number of methods available to measure airborne
nitrogen-containing species. This document briefly covers
methodologies currently available or in general use for in situ

monitoring of airborne concentrations in both ambient and indoor
environments. The species considered are NO, NO₂, NO_x, total
reactive odd nitrogen (NO_y), PAN and other organic nitrates, HNO₃,
HNO₂, N₂O₅, the nitrate radical, NO₃^-, and N₂O.

Measuring concentrations of nitrogen oxides is not trivial.
While a straightforward, widely available method exists for measuring
NO (the chemiluminescent reaction with ozone), this is an exception
for nitrogen oxides. Chemiluminescence is also the most common
technique used for NO₂; NO₂ is first reduced to NO. Unfortunately,
the catalyst typically used for the reduction is not specific, and has
various conversion efficiencies for other oxidized nitrogen compounds.
For this reason, great care must be taken in interpreting the results
of the common chemiluminescence analyser in terms of NO₂, as the
signal may include many other compounds. Additional difficulties
arise from nitrogen oxides that may partition between the gaseous and
particulate phases both in the atmosphere and in the sampling
procedure.

1.1.3 Exposure

Human and environmental exposure to nitrogen oxides varies
greatly from indoors to outdoors, from cities to the countryside, and
with time of day and season. The concentrations of NO and NO₂
typically present outdoors in a range of urban situations are
relatively well established. The concentrations encountered indoors
depend on the specific details of the nature of combustion appliances,
chimneys and ventilation. When unvented combustion appliances are
used for cooking or heating, indoor concentrations of nitrogen oxides
typically greatly exceed those existing outside. Recent research has
shown in these circumstances that HNO₂ can reach significant
concentrations. One report showed that HNO₂ can represent over 10% of
the concentrations usually reported as NO₂.

1.2 Effects of atmospheric nitrogen species, particularly nitrogen
oxides, on vegetation

Most of earth's biodiversity is found in (semi-)natural
ecosystems, both in aquatic and terrestrial habitats. Nitrogen is the
limiting nutrient for plant growth in many (semi-)natural ecosystems.
Most of the plant species from these habitats are adapted to nutrient-
poor conditions, and can only compete successfully on soils with low
nitrogen levels.

Human activities, both industrial and agricultural, have greatly
increased the amount of biologically available nitrogen compounds,
thereby disturbing the natural nitrogen cycle. Various forms of
nitrogen pollute the air: mainly NO, NO₂ and ammonia (NH₃) as dry
deposition; and nitrate (NO₃^-) and ammonium (NH₄⁺) as wet
deposition. NH_y refers to the sum of NH₃ and NH₄⁺. Another

contribution is from occult deposition (fog and clouds). There are
many more nitrogen-containing air pollutants (e.g., N₂O₅, PAN, N₂O,
amines), but these are neglected here, either because their
contribution to the total nitrogen deposition is believed to be small,
or because their concentrations are probably far below effect
thresholds.

Nitrogen-containing air pollutants can affect vegetation
indirectly, via photochemical reaction products, or directly after
being deposited on vegetation, soil or water surface. The indirect
pathway is largely neglected here although it includes very relevant
processes, and should be taken into account when evaluating the entire
impact of nitrogen-containing air pollutants: NO₂ is a precursor for
tropospheric O₃, which acts both as a phytotoxin and a greenhouse
gas.

The impacts of increased nitrogen deposition upon biological
systems can be the result of direct uptake by foliage or uptake via
the soil. At the level of individual plants, the most relevant
effects are injury to the tissue, changes in biomass production and
increased susceptibility to secondary stress factors. At the
vegetation level, deposited nitrogen acts as a nutrient; this results
in changes in competitive relationships between species and loss of
biodiversity. The critical loads for nitrogen depend on (i) the type
of ecosystem; (ii) the land use and management in the past and
present; and (iii) the abiotic conditions (especially those that
influence the nitrification potential and immobilization rate in the
soil).

Adsorption on the outer surface of the leaves takes place and may
damage wax layers of the cuticle, but the quantitative relevance for
the field situation has not yet been proved. Uptake of NO_x and NH₃
is driven by the concentration gradient between atmosphere and
mesophyll. It generally, but not always, is directly determined by
stomatal conductance and thus depends on factors influencing stomatal
aperture. There is increasing evidence that foliar uptake of nitrogen
reduces the uptake of nitrogen by the roots. Uptake and exchange of
ions through the leaf surface is a relatively slow process, and thus
is only relevant if the surface remains wet for longer periods.

NO is only slightly soluble in water, but the presence of other
substances can alter the solubility. NO₂ has a higher solubility,
while that of NH₃ is much higher. NO₂^- (the primary reaction
product of NO_x), NH₃ and NH₄⁺ are all highly phytotoxic, and could
well be the cause of adverse effects of nitrogen-containing air
pollutants. The free radical *N=O may play a role in the phytotoxicity
of NO.

More-than-additive effects (synergism) have been found in nearly
all studies concerning SO₂ plus NO₂. With other NO₂ mixtures (NO,
O₃ and CO₂), interactive effects are the exception rather than the
rule.

When climatic conditions and supply of other nutrients allow
biomass production, both NO_x and NH_y result in growth stimulation at
low concentrations and growth reduction at higher concentrations.
However, the exposure level at which growth stimulation turns into
growth inhibition is much lower for NO_x than for NH_y.

Evidence exists that plants are more sensitive at low light
intensity (e.g., at night and in winter) and at low temperatures (just
above 0°C). NO_x and NH_y can increase the sensitivity of plants to
frost, drought, wind and insect damage.

An interaction exists between soil chemistry and sensitivity of
vegetation to nitrogen deposition; this is related to pH and nitrogen
availability.

The relative contribution of NO and NO₂ to the NO_x effect on
plants is unclear. The vast majority of information is on effects of
NO₂ but available information on NO suggests that NO and NO₂ have
comparable phytotoxic effects.

Air quality guidelines refer to thresholds for adverse effects.
Two different types of effect thresholds exist: critical levels (CLEs)
and critical loads (CLOs). The critical level is defined as the
concentration in the atmosphere above which direct adverse effects on
receptors, such as plants, ecosystems or materials, may occur
according to present knowledge. The critical load is defined as a
quantitative estimate of an exposure (deposition) to one or more
pollutants below which significant harmful effects on specified
sensitive elements of the environment do not occur according to
present knowledge.

According to current practice, critical levels have been derived
from assessment of the lowest exposure concentrations causing adverse
effects on physiology or growth of plants (biochemical effects were
excluded), using a graphical method.

To include the impact of NO, a critical level for NO_x is
proposed instead of one for NO₂; for this purpose it has been assumed
that NO and NO₂ act in an additive manner. A strong case can be made
for the provision of critical levels for short-term exposure. However,
currently there are insufficient data to provide these with sufficient
confidence. Current evidence suggests a critical level of about
75 µg/m³ for NO_x as a 24-h mean.

The critical level for NO_x (NO and NO₂ added in ppb and
expressed as NO₂ in µg/m³) is considered to be 30 µg/m³ as an
annual mean.

Information on organisms in the environment is almost exclusively
restricted to plants, with minimum data on soil fauna. This
evaluation and guidance values are, therefore, expressed in terms of
nitrogen species effects on vegetation. However, it is expected that
plants will form the most sensitive component of natural systems and
that the effect on biodiversity of plant communities is a sensitive
indicator of effects on the whole ecosystem.

Critical loads are derived from empirical data and steady-state
soil models. Estimated critical loads for total nitrogen deposition
in a variety of natural aquatic and terrestrial ecosystems are given.
Possible differential effects of deposited nitrogen species (NO_x and
NH_y) are insufficiently known to differentiate between nitrogen
species for critical load estimation.

The great majority of ecosystems for which there is sufficient
information to estimate critical loads are from temperate climates.
The few arctic and montane ecosystems included, which might be
expected to be representative of higher latitudes, have the least
reliable basis. There is no information on tropical ecosystems and
little on estuarine or marine ecosystems in any climatic zone.
Nutrient-poor tropical ecosystems such as rain forests and mangrove
swamps are likely to be adversely affected by nitrogen deposition.
The lack of both deposition data and effect thresholds make it
impossible to make risk assessments for these climatic regions.

The most sensitive ecosystems (ombrotrophic bogs, shallow soft-
water lakes and arctic and alpine heaths) for which effects thresholds
can be estimated show critical loads of 5-10 kg N.ha^-1.year^-1 based
on decreased biological diversity in plant communities. A more
average value for the limited range of ecosystems studied is 15-20 kg
N.ha^-1.year^-1, which applies to forest trees.

The atmospheric chemistry of nitrogen oxides includes the
capacity for ozone generation in the troposphere, ozone depletion in
the stratosphere, and contribution to global warming as greenhouse
gases. Nitrogen oxides and ammonia contribute to soil acidification
(along with sulfur oxides) and thereby to increased bioavailability of
aluminium.

The phytotoxic effects of nitrogen oxides on plants have little
direct relevance to crop plants when concentrations marginally exceed
the critical level. However, the role of NO_x in the generation of
ozone and other phytotoxic substances, e.g., organic nitrates leads to

crop loss. Nitrogen deposited on growing crops will represent a very
small increase in total available nitrogen compared to that added as
fertilizer.

1.3 Health effects of exposures to nitrogen dioxide

A large number of studies designed to evaluate the health effects
of NO_x have been conducted. Of the NO_x compounds, NO₂ has been
most studied. The discussion in this section focuses on NO₂, NO,
HNO₂ and HNO₃, while nitrates are mentioned briefly.

1.3.1 Studies of the effects of nitrogen compounds on experimental
animals

Extrapolating animal data to humans has both qualitative and
quantitative components. As summarized below, NO₂ causes a
constellation of effects in several animal species; most notably,
effects on host defence against infectious pulmonary disease, lung
metabolism/biochemistry, lung function and lung structure. Because of
basic physiological, metabolic and structural similarities in all
mammals (laboratory animals and humans), the commonality of the
observations in several animal species leads to a reasonable
conclusion that NO₂ could cause similar types of effects in humans.
However, because of the differences between mammalian species, exactly
what exposures would actually cause these effects in humans is not yet
known. That is the topic of quantitative extrapolation. Limited
modelling research on the dosimetric aspect (i.e., the dose to the
target tissue/cell that actually causes toxicity) of quantitative
extrapolation suggests that the distribution of the deposition of NO₂
within the respiratory tract of animals and humans is similar,
without yet providing adequate values to use for animal-to-human
extrapolation. Unfortunately, very little information is available on
the other key aspect of extrapolation, species sensitivity (i.e., the
response of the tissues of different species to a given dose). Thus,
from currently available animal studies, we know which human health
effects NO₂ may cause. We are unable to assert with great confidence
the effects that are actually caused by a given inhaled dose of
NO₂.

With the above issues in mind, the animal toxicology database
for NO₂ is summarized below according to major classes of effects
and topics of special interest. Although it is clear that the
effects of NO₂ exposure extend beyond the confines of the lung, the
interpretation of these systemic effects relative to potential human
risk is not clear. Therefore they are not summarized further here,
but are discussed in later chapters. Although interactions of NO₂
and other co-occurring pollutants, such as O₃ and sulfuric acid
(H₂SO₄), can be quite important, especially if synergism occurs, the
database does not yet allow conclusions that enable assessment of
real-world potential interactions.

1.3.1.1 Biochemical and cellular mechanisms of action of nitrogen
oxides

NO₂ acts as a strong oxidant. Unsaturated lipids are readily
oxidized with peroxides as the dominant product. Both ascorbic acid
(vitamin C) and alpha-tocopherol (vitamin E) inhibit the peroxidation
of unsaturated lipids. When ascorbic acid is sealed within bilayer
liposomes, NO₂ rapidly oxidizes the sealed ascorbic acid. The
protective effects of alpha-tocopherol and ascorbic acid in animals
and humans are due to the inhibition of NO₂ oxidation. NO₂ also
oxidizes membrane proteins. The oxidation of either membrane lipids
or proteins results in the loss of cell permeability control. The
lungs of NO₂-exposed humans and experimental animals have larger
amounts of protein within the lumen. The recruitment of inflammatory
cells and the changes in the lung are due to these events.

The oxidant properties of NO₂ also induce the peroxide
detoxification pathway of glutathione peroxidase, glutathione
reductase and glucose-6-phosphate dehydrogenase. Following NO₂
exposure the increase in the peroxide detoxification pathway in
animals follows an exposure-response relationship.

The mechanism of action of NO is less clear. NO is readily
oxidized to NO₂ and peroxidation then occurs. Because of the
concurrent exposure to some NO₂ in NO exposures, it is difficult to
discriminate NO effects from NO₂. NO functions as an intracellular
second messenger modulating a wide variety of essential enzymes, and
it inhibits its own production (e.g., negative feedback). NO
activates guanylate cyclase which in turn increases intracellular cGMP
levels. A possible mechanism of action of nitrates may be through the
release of histamine from mast cell granules. Acidic nitrogenous air
pollutants, particularly HNO₃, may act by alteration of intracellular
pH.

PAN decomposes in water, generating hydrogen peroxide. Little is
known of the mechanism of action, but oxidative stress is likely for
PAN and its congeners.

Inorganic nitrates may act through alterations in intracellular
pH. Nitrate ion is transported into alveolar type 2 cells acidifying
the cell. Nitrate also mobilizes histamine from mast cells. HNO₂
could also act to alter intracellular pH, but this mechanism is
unclear.

The mechanisms of action of the other nitrogen oxides are
unknown.

Acute exposure to NO₂ at a concentration of 750 µg/m³ (0.4 ppm)
can result in lipid peroxidation. NO₂ can oxidize polyunsaturated

fatty acids in cell membranes as well as functional groups of proteins
(either soluble proteins in the cell, such as enzymes, or structural
proteins, such as components of cell membranes). Such oxidation
reactions (mediated by free radicals) are a mechanism by which NO₂
exerts direct toxicity on lung cells. This mechanism of action is
supported by animal studies showing the importance of lung antioxidant
defences, both endogenous (e.g., maintenance of lung glutathione
levels) and exogenous (e.g., dietary vitamins C and E), in protecting
against the effects of NO₂. Many studies have suggested that various
enzymes in the lung, including glutathione peroxidase, superoxide
dismutase and catalase, may also serve to defend the lung against
oxidant attack.

1.3.1.2 Effects on host defence

Although the primary function of the respiratory tract is to
ensure an efficient exchange of gases, this organ system also provides
the body with a first line of defence against inhaled viable and non-
viable airborne agents. An extensive database clearly shows that
exposure to NO₂ can result in the dysfunction of these host defences,
increasing susceptibility to infectious respiratory disease. The
host-defence parameters affected by NO₂ include the functional and
biochemical activity of cells in lungs, alveolar macrophages (AMs),
immunological competence, susceptibility to experimentally induced
respiratory infections, and the rate of mucociliary clearance.

Alveolar macrophages are affected by NO₂. These cells
are responsible for maintaining the sterility of the pulmonary
region, clearing particles from this region, and participating in
immunological functions. Functional changes that have been reported
include the following: the suppression of phagocytic ability and
stimulation of lung clearance at 560 µg/m³ (0.3 ppm) 2 h/day for
13 days; a decrease in bactericidal activity at 4320 µg/m³ (2.3 ppm)
for 17 h; and a decreased response to migration inhibition factor at
3760 µg/m³ (2.0 ppm) 8 h/day, 5 days/week for 6 months. The
morphological appearance of these defence cells changes after chronic
exposure to NO₂.

The importance of host defences becomes evident when animals have
to cope with laboratory-induced pulmonary infections. Animals exposed
to NO₂ succumb to bacterial or viral infection in a concentration-
dependent manner. Mortality also increases with increased NO₂
concentration or duration of exposure. After acute exposure,
effects are observed at concentrations as low as 3760 µg/m³ (2 ppm).
Exposure to concentrations as low as 940 µg/m³ (0.5 ppm) will cause
effects in the infectivity model after 6 months.

Both humoral and cell-mediated defence systems are changed by
NO₂ exposure. In the cases in which the immune system has been
investigated, effects have been observed after short-term exposure to

concentrations > 9400 µg/m³ (5 ppm). The effects are complex
since the direction of the change (i.e., increase or decrease) is
dependent upon NO₂ concentration and the length of exposure.

1.3.1.3 Effects of chronic exposure on the development of chronic
lung disease

Humans are chronically exposed to NO₂. Therefore, such
exposures in animals have been studied rather extensively, typically
using morphological and/or morphometric methods. This research has
generally shown that a variety of pulmonary structural and correlated
functional alterations occur. Some of these changes may be reversible
when exposure ceases.

Pulmonary function may be altered following chronic NO₂ exposure
of experimental animals. Impaired gas exchange occurred following
exposure to 7520 µg/m³ (4.0 ppm) NO₂ for four months and this was
reflected in decreased arterial O₂ tension, impaired physical
performance and increased anaerobic metabolism.

Although NO₂ produces morphological changes in the respiratory
tract, the database is sometimes confusing due to quantitative and
qualitative variability in responsiveness between, and even within,
species. The rat, the most commonly used experimental animal in
morphological assessments of exposure, appears to be relatively
resistant to NO₂. Short-term exposures to concentrations of
9400 µg/m³ (5.0 ppm) or less generally have little effect in the
rat, where similar exposures in the guinea-pig may result in some
centriacinar epithelial damage.

Longer-term exposures result in lesions in some species with
concentrations as low as 560 to 940 µg/m³ (0.3 to 0.5 ppm). These
are characterized by epithelial remodelling similar to that described
above, but with the involvement of more proximal airways and
thickening of the interstitium. Many of these changes, however, will
resolve even with continued exposure, and long-term exposures to
levels above about 3760 µg/m³ (2.0 ppm) are required for more
extensive and permanent changes in the lungs. Some effects are
relatively persistent (e.g., bronchiolitis), whereas others tend to be
reversible and limited even with continued exposure. In any case, it
seems that for either short- or long-term exposure, the response is
more dependent upon concentration than duration of exposure.

There is substantial evidence that long-term exposure of several
species of laboratory animals to high concentrations of NO₂ results
in morphological lung lesions. Destruction of alveolar walls, an
essential additional criterion for human emphysema, has been reliably
reported in lungs from animals in a limited number of studies. The

lowest NO₂ concentration for the shortest exposure duration that will
result in emphysematous lung lesions cannot be determined from these
published studies.

1.3.1.4 Potential carcinogenic or co-carcinogenic effects

NO₂ has been shown to be mutagenic in Salmonella bacteria, but
was not mutagenic in one study with a mammalian cell culture. Other
studies using cell cultures have demonstrated sister chromatid
exchanges (SCE) and DNA single strand breaks. No genotoxic effects
have been demonstrated in vivo concerning lymphocytes, spermatocytes
or bone marrow cells, but two inhalation studies with high
concentrations (50 760 and 56 400 µg/m³, 27 and 30 ppm) for 3 h and
16 h, respectively, have demonstrated such effects in lung cells.

Literature searches revealed no published reports of NO₂ studies
using classical whole-animal chronic bioassays for carcinogenesis.
Research with mice having spontaneously high tumour rates was
equivocal. In one study, NO₂ at 18 800 µg/m³ (10 ppm) slightly
enhanced the incidence of lung adenomas in a sensitive strain of mice
(A/J). Although several co-carcinogenesis investigations have been
undertaken, conclusions are precluded because of problems with
methodology and interpretation. Reports on whether NO₂ facilitates
the metastasis of tumours to the lung are also inadequate to form
conclusions. Other investigations have centred on whether NO₂ could
produce nitrates and nitrites that, by reacting with amines in the
body, could produce nitrosamines. A few studies suggest that
nitrosamines are formed in animals treated with high doses of amines
and exposed to NO₂, but other studies have indicated that nitrosamine
formation is unlikely.

1.3.1.5 Age susceptibility

Investigations into age dependency are inadequate and results so
far are equivocal.

1.3.1.6 Influence of exposure patterns

Several animal toxicological studies have elucidated the
relationships between concentration (C) and duration (T) of exposure,
indicating that the relationship is complex. Most of this research
has used the infectivity model. Early C × T studies demonstrated that
concentration had more impact on mortality than did duration of
exposure. An evaluation of the toxicity of NO₂ exposures cannot be
delineated by C × T relationships.

1.3.2 Controlled human exposure studies on nitrogen oxides

Human responses to a variety of oxidized nitrogen compounds have
been evaluated. By far, the largest database and the one most

suitable for risk assessment is that available for controlled
exposures to NO₂. The database on human responses to NO, HNO₃
vapour, HNO₂ vapour and inorganic nitrate aerosols is not as
extensive. A number of sensitive or potentially sensitive subgroups
have been examined, including adolescent and adult asthmatics, older
adults, and patients with chronic obstructive pulmonary disease (COPD)
and pulmonary hypertension. Exercise during exposure increases the
total uptake and alters the distribution of the deposited inhaled
material within the lung. The relative proportion of NO₂ deposited in
the lower respiratory tract is also increased by exercise. This may
increase the effects of the above compounds in people who exercise
during exposure.

As is typical with human biological response to inhaled particles
and gases, there is variability in the biological response to NO₂.
Healthy individuals tend to be less responsive to the effects of NO₂
than individuals with lung disease. Asthmatics are clearly the most
responsive group to NO₂ that has been studied to date. Individuals
with COPD may be more responsive than healthy individuals, but they
have limited capacity to respond to NO₂ and thus quantitative
differences between COPD patients and others are difficult to assess.
Sufficient information is not available at present to evaluate whether
age and sex play a role in the response to NO₂.

Healthy subjects can detect the odour of NO₂, in some cases at
concentrations below 188 µg/m³ (0.1 ppm). Generally, NO₂ exposure
did not increase respiratory symptoms in any of the subject groups
tested.

NO₂ causes decrements in lung function, particularly increased
airway resistance in resting healthy subjects at 2-h concentrations as
low as 4700 µg/m³ (approx.2.5 ppm). Available data are insufficient
to determine the nature of the concentration-response relationship.

Exposure to NO₂ results in increased airway responsiveness to
bronchoconstrictive agents in exercising healthy, non-smoking subjects
exposed to concentrations as low as 2800 µg/m³ (approx.1.5 ppm) for
1 h or longer.

Exposure of asthmatics to NO₂ causes, in some subjects,
increased airway responsiveness to a variety of provocative mediators,
including cholinergic and histaminergic chemicals, SO₂ and cold air.
The presence of these responses appears to be influenced by the
exposure protocol, particularly whether or not the exposure includes
exercise. These responses may begin at concentrations as low as
380 µg/m³ (0.2 ppm). A meta-analysis suggests that effects may occur
at even lower concentrations. However, an unambiguous concentration-
response relationship is observed between 350 to 1150 µg/m³
(approx.0.2 to 0.6 ppm).

The implications of this overall trend are unclear, but increased
airway responsiveness could potentially lead to increased response to
aeroallergens or temporary exacerbation of asthma, possibly leading to
increased medication usage or even increased hospital admissions.

Modest increases in airway resistance may occur in COPD patients
from brief exposure (15-60 min) to concentrations of NO₂ as low as
2800 µg/m³ (approx.1.5 ppm), and decrements in spirometric measures
of lung function (3 to 8% change in FEV₁ (forced expiratory volume
in 1 second)) may also be observed with longer exposures (3 h) to
concentrations as low as 600 µg/m³ (approx.0.3 ppm).

Exposure to NO₂ at levels above 2800 µg/m³ (approx.1.5 ppm) may
alter the numbers and types of inflammatory cells in the distal
airways or alveoli. NO₂ may alter the functioning of cells within
the lungs and production of mediators that may be important in lung
host defences. The constellation of changes in host defences,
alterations in lung cells and their activities, and changes in
biochemical mediators is consistent with the epidemiological findings
of increased host susceptibility associated with NO₂ exposure.

In studies on mixtures of NO₂ with other pollutants, NO₂ has
not been observed to increase responses to other co-occurring
pollutant(s) beyond that which would be observed for the other
pollutant(s) alone. A notable exception is the observation that
pre-exposure to NO₂ enhanced the ozone-induced change in airway
responsiveness in healthy exercising subjects during a subsequent
ozone exposure. This observation suggests the possibility of delayed
or persistent responses to NO₂.

Within an NO₂ concentration range that may be of interest with
regard to risk evaluation (i.e., 100-600 µg/m³), the characteristics
of the concentration-response relationship for acute changes in lung
function, airway responsiveness to bronchoconstricting agents or
symptoms cannot be determined from the available data.

On the basis of an effect at 400 µg/m³ and the possibility of
effects at lower levels, based on a meta analysis, a one-hour average
daily maximum NO₂ concentration of 200 µg/m³ (approx.0.11 ppm) is
recommended as a short-term guideline.

NO is acknowledged as an important endogenous second messenger
within several organ systems. Inhaled NO concentrations above
6000 µg/m³ (approx.5 ppm) can cause vasodilation in the pulmonary
circulation without affecting the systemic circulation. The lowest
effective concentration has not been established. Information on
pulmonary function and lung host defences consequent to NO exposure
are too limited for any conclusions to be drawn at this time.

Relatively high concentrations (> 40 000 µg/m³) have been used in
clinical applications for brief periods (< 1 h) without reported
adverse reactions.

Nitric acid levels in the range of 250-500 µg/m³ (97-194 ppb)
may cause some pulmonary function responses in adolescent asthmatics,
but not in healthy adults.

Limited information on HNO₂ suggests that it may cause eye
inflammation at 760 µg/m³ (0.40 ppm). There are currently no
published data on human pulmonary responses to HNO₂.

Limited data on inorganic nitrates suggest that there are no lung
function effects of nitrate aerosols at concentrations of 7000 µg/m³
or less.

1.3.3 Epidemiology studies on nitrogen dioxide

Epidemiological studies on the health effects of nitrogen oxides
have mainly focused on NO₂. Many indoor and outdoor epidemiological
studies designed to evaluate the health effects of NO₂ have been
conducted. Two health outcome measurements of NO₂ exposure are
generally considered: lung function measurements and respiratory
symptoms and diseases.

The evidence from individual studies of the effect of NO₂ on
lower respiratory symptoms and disease in school-aged children is
somewhat mixed. The consistency of these studies was examined and
the evidence synthesized in a combined quantitative analysis
(meta-analysis) of the subject studies. Most of the indoor studies
showed increased lower respiratory morbidity in children associated
with long-term exposure to NO₂. Mean weekly NO₂ concentrations
in bedrooms in studies reporting NO₂ levels were predominantly
between 15 and 122 µg/m³ (0.008 and 0.065 ppm). Combining the
indoor studies as if the end-points were similar gives an estimated
odds ratio of 1.2 (95% confidence limits of 1.1 and 1.3) for the effect
per 28.3 µg/m³ (0.015 ppm) increase of NO₂ on lower respiratory
morbidity. This suggests that, subject to assumptions made for the
combined analysis, an increase of about 20% in the odds of lower
respiratory symptoms and disease corresponds to each increase of
28.3 µg/m³ (0.015 ppm) in estimated 2-week average NO₂
exposure. Thus, the combined evidence is supportive for the effects
of estimated exposure to NO₂ on lower respiratory symptoms and
disease in children aged 5 to 12 years.

In individual indoor studies of infants 2 years of age or younger,
no consistent relationship was found between estimates of NO₂
exposure and the prevalence of respiratory symptoms and disease. Based

on a meta-analysis of these indoor infant studies, subject to the
assumptions made for the meta-analysis, the combined odds ratio for the
increase in respiratory disease per increase of 28.2 µg/m³ (0.015 ppm)
NO₂ was 1.09 with a 95% confidence interval of 0.95 to 1.26, where
mean weekly NO₂ concentrations in bedrooms were predominantly between
9.4 and 94 µg/m³ (0.005 and 0.050 ppm) in studies reporting levels.
The increase in risk was very small and was not reported consistently
by all studies. We cannot conclude that the evidence suggests an effect
in infants comparable to that seen in older children. The reasons for
these age-related differences are not clear.

The measured NO₂ studies gave a higher estimated odds ratio than
the surrogate estimates, which is consistent with a measurement error
effect. The effect of having adjusted for covariates such as
socioeconomic status, smoking and sex was that those studies that
adjusted for a particular covariate found larger odds ratios than
those that did not.

Although many of the epidemiological studies that involved
measured NO₂ levels used measurements over only 1 or 2 weeks, these
levels were used to characterize children's exposures over a much
longer period. The standard respiratory symptom questionnaire used by
most of these studies summarizes information on health status over an
entire year. The 28.2 µg/m³ (0.015 ppm) difference in NO₂ levels
used in the meta-analyses relates to a difference in the household
annual average exposure between gas and electric cooking stoves.
Some studies measured NO₂ levels only in the winter and may have
overestimated annual average exposures. This would tend to have
underestimated the health effect of a 28.2 µg/m³ (0.015 ppm)
difference in the annual NO₂ exposure. A study based on a household
annual average exposure measured in both the winter and summer found a
stronger health effect than many of the other studies. The true
biologically relevant exposure period is unknown, but these exposures
extended over a lengthy period up to the entire lifetime of the child.

The association between outdoor NO₂ and respiratory health is
not clear from current research. There is some evidence that the
duration of respiratory illness may be increased at higher ambient
NO₂ levels. A major difficulty in the analysis of outdoor studies is
distinguishing possible effects of NO₂ from those of other associated
pollutants.

Several uncertainties need to be considered in interpreting the
above studies and meta-analysis. Error in measuring exposure is
potentially one of the most important methodological problems in
epidemiological studies of NO₂. Although there is evidence that
symptoms are associated with indicators of NO₂ exposure, the quality
of these exposure estimates may be inadequate to determine a
quantitative relationship between exposure and symptoms. Most of the
studies that measured NO₂ exposure did so only for periods of 1 to

2 weeks and reported the values as averages. Few of the studies
attempted to relate the observed effects to the pattern of exposure
(e.g., transient NO₂ peaks). Furthermore, measured NO₂ concentration
may not be the biologically relevant dose; estimating actual exposure
requires knowledge of pollutant species, levels and related human
activity patterns. However, only very limited activity and aerometric
data are available that examine such factors. The extrapolation to
possible patterns of ambient exposure is difficult. In addition,
although the level of similarity and common elements between the
outcome measures in the NO₂ studies provide some confidence in their
use in the quantitative analysis, the symptoms and illnesses combined
are to some extent different and could indeed reflect different
underlying processes. Thus, caution is necessary in interpreting the
meta-analysis results.

Other epidemiological studies have attempted to relate some
measure of indoor and/or outdoor NO₂ exposure to changes in pulmonary
function. These changes were marginally significant. Most studies
did not find any effects, which is consistent with controlled human
exposure study data. However, there is insufficient epidemiological
evidence to draw any conclusions about the long- or short-term effects
of NO₂ on pulmonary function.

On the basis of a background level of 15 µg/m³ (0.008 ppm) and
the fact that significant adverse health effects occur with an
additional level of 28.2 µg/m³ (0.015 ppm) or more, an annual
guideline value of 40 µg/m³ (0.023 ppm) is proposed. This value will
avoid the most severe exposures. The fact that a no-effect level for
subchronic or chronic NO₂ exposure concentrations has not yet been
determined should be emphasized.

1.3.4 Health-based guidance values for nitrogen dioxide

On the basis of human controlled exposure studies, the
recommended short-term guidance value is for a one-hour average NO₂
daily maximum concentration of 200 µg/m³ (0.11 ppm). The recommended
long-term guidance value, based on epidemiological studies of
increased risk of respiratory illness in children, is 40 µg/m³
(0.023 ppm) annual average.

2. PHYSICAL AND CHEMICAL PROPERTIES, AIR SAMPLING AND ANALYSIS,
TRANSFORMATIONS AND TRANSPORT IN THE ATMOSPHERE

2.1 Introduction

Nitrogen oxides are produced by combustion processes and are
emitted to the air mainly as NO together with some NO₂. Natural
biological processes and lightning also emit NO and N₂O. In the
atmosphere nitrogen oxides undergo complex chemical and photochemical
reactions; NO is oxidized to NO₂ and other products and eventually to
HNO₃ and nitrates. Nitrogenous species are removed from the air to
the ground by wet and dry deposition processes. Oxidized nitrogen
compounds can have impacts on human health and the environment, and
are important to the formation of photochemical smog and tropospheric
ozone.

In this chapter the properties of nitrogen compounds are briefly
described and techniques for their sampling and analysis outlined.
Atmospheric chemical reactions that cause the oxidation of NO to NO₂
and the production of ozone, organic nitrates and HNO₃ are described.
The differences between night-time and day-time chemistry and the
composition of the atmosphere are discussed. The nature of the
nitrogen species and their chemical reactions in urban regions, in
chimney plumes such as those from power stations, in air advected away
from urban regions and in rural and remote areas are described. The
role of nitrogen oxides in photochemical smog production and the
effects of nitrous oxide on stratospheric ozone are briefly discussed.

2.1.1 The nomenclature and measurement of atmospheric nitrogen
species

There are several methods available for determining nitrogen
species, but many of these techniques are nonspecific.

To denote various mixtures of nitrogen species, the terms NO_x,
NO_y and NO_z are often employed. It is customary to refer to the sum
of NO and NO₂ emitted from a source as NO_x, the unit of measure for
NO_x being the NO₂ mass equivalent of the NO plus NO₂.

The term NO_y is frequently used to denote the sum of the gas
phase oxidized nitrogen species (except N₂O) and NO_z to denote the
sum of NO_y plus the oxidized nitrogen present as particulate matter.
Measurement of NO_z requires a combination of particulate and gas
phase sampling and analysis.

A confusion arises because one of the most commonly used methods
for determining NO₂ in ambient air (thermal conversion of NO₂ to NO
and measurement of the resultant NO by chemiluminescent reaction with
O₃) is nonspecific and responds to several gaseous species in
addition to NO₂. These include organic nitrogen compounds and,

depending on the converter, HNO₃, although HNO₃ can be readily lost
to the sampling system. Therefore, depending on the composition of
the air being sampled, the results from this type of instrument can be
representative of NO_y rather than NO_x (or NO₂) concentrations.
This technique is used in most routine determinations of ambient NO_x
and NO₂ concentrations but the discrepancy between these values and
true NO_x and NO₂ can be considerable for air in which the pollutant
emissions have undergone substantial exposure to sunlight.

Nitrous oxide is ubiquitous in the atmosphere because it is a
product of biological processes in soil as well as anthropogenic
activities. It is not involved to any appreciable extent in chemical
reactions in the lower atmosphere, but it is an active "greenhouse"
gas. In the stratosphere N₂O forms NO by reaction with excited
oxygen atoms, and this NO then acts to deplete the stratospheric O₃
concentration.

Although NO₃, dinitrogen trioxide (N₂O₃), dinitrogen tetroxide
(N₂O₄), and N₂O₅ may play a role in atmospheric chemical reactions
leading to the transformation, transport, and ultimate removal of
nitrogen compounds from ambient air, they are present in very low
concentrations, even in polluted environments.

NH₃ is generated during decomposition of nitrogenous matter in
natural ecosystems and may be locally produced in high concentrations
by human activities such as intensive animal husbandry and feedlots.
Under suitable conditions NH₃ can react with oxidized nitrogen
species to form ammonium nitrate aerosol.

2.2 Nitrogen species and their physical and chemical properties

There are seven oxides of nitrogen that may be present in ambient
air, namely: NO, NO₂, N₂O, NO₃, N₂O₃, N₂O₄ and N₂O₅. In
addition these can be present as HNO₂, HNO₃ and various organic
nitrogen species, such as PAN, other organic nitrates and particles
containing oxidized nitrogen compounds (particularly adsorbed nitric
acid). Of these species, NO and NO₂ are the ones most often measured
and are present in the greatest concentrations in urban and industrial
air.

The chemical and physical properties of individual nitrogen
species are given below and are summarized in Table 1.

Table 1. Some physical and thermodynamic properties of oxides of nitrogen and other nitrogen compounds^a

Oxide Relative Melting point Boiling point Solubility in water Thermodynamic functions
molecular (°C)^b,c,d (°C)^b,c at 0°C (cm³ per 100 g)^b (Ideal gas, 1 atm, 25°C)
mass (g/mol)
Enthalpy of Entropy
formation (cal/mol-deg)
(kcal/mol)

NO 30.01 -163.6 -151.8 7.34 21.58 50.35

NO₂ 46.01 -11.2 21.2 Reacts with H₂O forming 7.91 57.34
HNO₂ and HNO₃

N₂O 44.01 -90.8 -88.5 130.52 19.61 52.55

N₂O₃ 76.01 -102 47 Reacts with H₂O forming 19.80 73.91
(decomposes) HNO₂

N₂O₄ 92.02 -11.3 21.2 Reacts with H₂O forming 2.17 72.72
HNO₂ and HNO₃

N₂O₅ 108.01 30 3.24 Reacts with H₂O forming 2.7 82.8
(decomposes) HNO₂

HNO₂ 47.01 - - - - -

HNO₃ 63.01 -42 83 -32.1 63.7

Table 1. (Con't)

PAN 121.06 - - - - -
(CH₃COOONO₂)

NH₄NO₃ 80.04 169.6 210 at 118.3 g/100 cm³ -87.37 36.11
11 torr H₂O at 0°C

^a Adopted from: US EPA (1993)
^b Matheson Gas Data Book (Matheson Company, 1966)
^c Handbook of Chemistry and Physics (Weast et al., 1986)
^d At 0°C and 1 atm pressure
2.2.1 Nitrogen oxides

2.2.1.1 Nitric oxide

NO is a colourless, odourless gas that is only slightly soluble
in water. It is a by-product of combustion processes, arising from
(i) high temperature oxidation of molecular nitrogen from the
combustion air, and (ii) from oxidation of nitrogen present in certain
fuels such as coal and heavy oil.

2.2.1.2 Nitrogen dioxide

NO₂ is a reddish-orange-brown gas with a characteristic pungent
odour. The boiling point is 21.1°C, but the low partial pressure of
NO₂ in the atmosphere prevents condensation. NO₂ is corrosive and
highly oxidizing. About 5 to 10% by volume of the total emissions of
NO_x from combustion sources is usually in the form of NO₂, although
substantial variations from one source type to another have been
observed.

In the atmosphere, photochemical reactions involving ozone
and organic compounds convert NO to NO₂. NO₂ is an efficient
absorber of light over a broad range of ultraviolet (UV) and visible
wavelengths. Because of its brown colour, NO₂ can contribute to
discoloration and reduced visibility of polluted air. Photolysis of
NO₂ by sunlight produces NO and an oxygen atom, which usually adds to
an oxygen molecule to produce ozone.

2.2.1.3 Nitrous oxide

N₂O is a colourless gas with a slight odour at high
concentrations. It is emitted to the atmosphere as a trace component
from some combustion sources and from the consumption of nitrate by
an ubiquitous group of denitrification bacteria that use nitrate as
their terminal electron acceptor in the absence of oxygen (Delwiche,
1970; Brezonik, 1972; Keeney, 1973; Focht & Verstraete, 1977). At
atmospheric concentrations N₂O has no significant physiological
effects in humans, although at higher concentrations it is employed as
an anaesthetic.

N₂O does not play a significant role in atmospheric reactions in
the lower troposphere. In the stratosphere it reacts with singlet
oxygen to produce NO, which participates in O₃ decomposition in
the stratosphere. These reactions are of concern because of the
possibility that increasing N₂O concentrations resulting from fossil
fuel use, and also from denitrification of excess fertilizer, may
contribute to a decrease in stratospheric O₃ (Council for
Agricultural Science and Technology, 1976; Crutzen, 1976) with
consequent potential for adverse impacts on ecosystems and human

health. Also of concern is the fact that N₂O absorbs long-wave
radiation, and therefore serves as a radiatively important greenhouse
gas that may contribute to global warming.

2.2.1.4 Other nitrogen oxides

Other nitrogen oxides can be present in trace quantities in the
air. NO₃ has been identified in laboratory systems containing
NO₂/O₃, NO₂/O and N₂O₅ as an important reactive transient
(Johnston, 1966). It is likely to be present in photochemical smog.
In the presence of sunlight, NO₃ is rapidly converted to either NO or
NO₂ (Wayne et al., 1991). Nitrogen trioxide is highly reactive
towards both NO and NO₂. Its expected concentration in polluted air
is very low (about 10^-6 µg/m³). However, traces of NO₃ may play an
important role in atmospheric chemistry, especially at night when it
may serve as a reservoir for NO_x (Wayne et al., 1991). In the
atmosphere N₂O₃ is in equilibrium with NO and NO₂. It reacts with
water to form HNO₂. N₂O₄ is the dimer of NO₂, formed in
equilibrium with NO₂ molecules, and it readily dissociates to NO₂.
N₂O₅ can be a trace night-time component of the air because it is
formed by a reaction between NO₂ and NO₃. Since NO₃ can exist in
appreciable quantities only in the absence of sunlight, N₂O₅ is only
important at night, when its reaction with water can be a significant
source of nitric acid.

2.2.2 Nitrogen acids

2.2.2.1 Nitric acid

HNO₃ is the most oxidized form of nitrogen. In the gaseous
state it is colourless. It is photochemically stable in the
troposphere. HNO₃ is volatile, so that at typical concentrations and
temperatures in the atmosphere the vapour does not coalesce into
aerosol and is not retained on particles unless the aerosol contains
reactants such as sodium chloride or ammonium salts to react with the
acid, when it produces particulate nitrates (Wolff, 1984).

In the aqueous phase (e.g., rain drops), HNO₃ dissociates to
form the nitrate ion (NO₃^-). Because nitrate is chemically
unreactive in dilute aqueous solution, nearly all of the
transformations involving nitrate in natural waters result from
biochemical pathways. The nitrate salts of all common metals are
quite soluble.

2.2.2.2 Nitrous acid

HNO₂ is formed when NO and NO₂ are present in the atmosphere,
as a result of their reaction with water. In sunlight, the dominant

pathway for HNO₂ formation is the reaction of NO with hydroxyl
radicals. During the daytime, atmospheric concentrations of HNO₂ are
limited by the photolysis of HNO₂ to produce NO and hydroxyl radical.

Nitrous acid is a weak reducing agent and is oxidized to nitrate
only by strong chemical oxidants and by nitrifying bacteria.

2.2.3 Ammonia

NH₃ is the completely reduced form of nitrogen. It is a
colourless gas with a pungent odour. It is extremely soluble in
water, forming ammonium (NH_y⁺) and hydroxyl (OH^-) ions. In the
atmosphere, NH₃ has been reported to be converted into NO_x by
reaction with hydroxyl radicals (Soederlund & Svensson, 1976). In the
stratosphere, NH₃ can be dissociated by irradiation with sunlight at
wavelengths below 230 nm (McConnell, 1973).

2.2.4 Ammonium nitrate

Gas-phase ammonia reacts with nitric acid to form ammonium
nitrate (NH₄NO₃). Ammonium nitrate is a solid at room temperature.
Like ammonia, it is very soluble in water and hence will be absorbed
by any water droplets present. Thus it readily forms an aerosol in
the atmosphere. Pathways to aerosol formation include nucleation and
condensation on existing particles. The presence of NH₄NO₃
particles can result in a visible haze.

2.2.5 Peroxyacetyl nitrate

Of the various peroxy nitrates found in ambient air, peroxyacetyl
nitrate (CH₃COOONO₂), or PAN, is found at the highest concentrations.
PAN undergoes a temperature-dependent decomposition to its precursors,
NO₂ and acetyl peroxy radicals. At low ambient temperatures PAN
can have a substantial lifetime in the atmosphere (Cox & Roffey, 1977).
In polluted air PAN concentrations can reach several parts per billion.

2.2.6 Organic nitrites and nitrates

A wide variety of organic nitrites (RNO₂) and nitrates (RNO₃),
where R denotes CH₃, CH₂CH₃, benzyl, etc., may be found in ambient
air. Some of these are emitted directly while others are formed by
photochemical reactions in the atmosphere.

2.3 Sampling and analysis methods

This section outlines methods for measuring nitrogen-containing
species in the atmosphere. The main focus is on methodologies
currently available and in general use for monitoring concentrations
in both ambient and indoor air.

Table 2 summarizes sampling and analytical methods for selected
species and addresses relevant characteristics, including the type of
method (i.e., in situ, remote, active, passive, continuous or
integrative), the stage of development of the method, sampling
duration, precision, accuracy and detection limits.

2.3.1 Nitric oxide

2.3.1.1 Nitric oxide continuous methods

Nitric oxide reacts rapidly with O₃ to give NO₂ in an excited
electronic stage. The transition of excited NO to the grand state can
be accompanied by the emission of light in the red-infrared spectral
range. When this chemiluminescent reaction occurs under controlled
conditions, the intensity of the emitted light is proportional to the
concentration of the NO reactant. This provides the basis of the
chemiluminescence method (CLM) for analysis of NO. This method is a
continuous technique and is the most commonly used method for
measuring NO in ambient air. Commercial instruments for measuring NO
and NO₂ are available with detection limits of approximately 5 ppb
and response times of the order of minutes. CLM measurement of NO₂
can also be accomplished by firstly converting the NO₂ of the sample
to NO. This is discussed in section 2.3.2.1.

Other NO analytical methods include laser-induced fluorescence
(LIF) (Bradshaw et al., 1985), absorption spectroscopy (e.g., tuneable
diode laser absorption spectroscopy, TDLAS) and passive samplers.

2.3.1.2 Passive samplers for NO

Passive samplers are used for air with higher-than-typical
ambient concentrations, which may be found indoors or in the
workplace. They are often used to obtain data at a large number of
sites. Sampling typically lasts a few hours.

The Palmes tube is a passive sampler that relies on diffusion of
an analyte molecule through a quiescent diffusion path of known length
and cross-sectional area to a reactive surface where the molecule is
captured by chemical reaction (Palmes et al., 1976). The Palmes tube
does not measure NO directly. Two tubes are required; the first one
has reactive grids coated with triethanolamine (TEA) to collect NO₂,
the second tube is similar but has an additional reactive surface
coated with chromic acid to convert NO to NO₂, which is in turn
collected by the TEA-coated grids. The NO concentration of the air is
determined from the difference in the results from the two tubes. The
data is corrected for the effects of the different diffusivities of NO
and NO₂ molecules. To ensure reliable results, contact between the
chromic-acid-coated surface and the TEA-coated grids for longer than
24 h must be avoided. Analysis of the material contained in the TEA

Table 2. Selected instruments and methods for determining oxides of nitrogen in ambient air (from: Sickles, 1992)

Species Methods^a Type^b Development Sample Performance Comments References
stage^c duration
Precision Accuracy MDL^d

NO CLM I, A, C C 5 min < 10% < 20% < 9 ppb - Finlayson-Pitts &
(NO + O₃) Pitts (1986)

TP-LIF I, A, C R 30 sec - 16% 10 ppt - Bradshaw et al. (1985);
Davis et al. (1987)

TDLAS I, A, C R, C 60 sec - - 0.5 ppb 40-m path length NASA (1983)

PSD I, P, IN C 24 h - - 70 ppb-h^e

NO₂ CLM I, A, C C 5 min 10% 20% 9 ppb Commonly used Finlayson-Pitts &
(NO + O₃) method; many Pitts (1986)
interferences

CLM I, A, C R < 100 sec 20 ppt 30% 10-25 ppt Uses thermal or Helas et al. (1987);
(NO + O₃) photolytic Fehsenfeld et al.
converters (1987)

CLM I, A, C C 100 sec 0.6 ppb - 10 ppt Interferences:
(Luminol) PAN, HNO₂, O₃

TP-LIF I, A, C R 2 min 20 ppt 16% 12 ppt - Davis (1988)

TDLAS I, A, C R, C 60 sec - 15% 100 ppt 150-m path length NASA (1983)

DOAS R, A, C R, C 12 min - 10% 4 ppb 800-m path length Platt & Perner (1983)

Bubbler I, A, IN RM 24 h 6 ppb 10% 8 ppb^e Purdue & Hauser (1980)

Table 2. (Con't)

Species Methods^a Type^b Development Sample Performance Comments References
stage^c duration
Precision Accuracy MDL^d

TEA I, A, IN L 24 h 15% 10% 0.2 ppb^e Interferences: Sickles et al. (1990)
filter PAN and HNO_2^f

Guaiacol I, A, IN L 1 h 4% - 0.1 ppb^e Stability of Buttini et al. (1987)
Denuder extract uncertain

DPA I, A, IN L 8 h 8% - 0.1 ppb^e DPA may volatilize; Lipari (1984)
Cartridge interferences:
HNO₂ and PAN

TEA PSD I, P, IN L 24 h 30% - 30 ppb-h^e Similar to Palmes
Tube; interferences
as above^f

NO_y CLM I, A, C R 10 sec - 15% 10 ppt CO with Au Fahey et al. (1986)
(NO + O₃) reducing catalyst

PAN GC-ECD I, A, IN R, RM 15 min - 30% 10 ppt^e Sensitivity can be Vierkorn-Rudolph
enhanced by using et al. (1985)
cryogenic sampling
and capillary
columns

GC-CLM I, A, IN L - - - - CLM (NO + O₃) and
(Luminol) reported

Other organic GC-ECD/MS I, A, C R 24 h - - 1 ppt^e Sample collected Atlas (1988)
Nitrates on charcoal

Table 2. (Con't)

Species Methods^a Type^b Development Sample Performance Comments References
stage^c duration
Precision Accuracy MDL^d

NHO₃ Filter I, A, IN R, RM 24 h 10% 20% 8 ppt^e May be nylon or Finlayson-Pitts &
calcium chloride Pitts (1986)
impregnated filter;
subject to
artifacts^f

Denuder I, A, IN R, RM 24 h 8% - 8 ppt^e Not subject to Sickles (1987);
above artifacts^f Sickles et al. (1989)

TDLAS I, A, C R, C 5 min - 20% 100 ppt 150-m path length NASA (1983)

HNO₂ Denuder I, A, IN R, RM 24 h 15% - 10 ppt^e Annular denuder Sickles et al. (1989);

preferred^f Vossler et al. (1988)

LIF I, A, C R 15 min - - 20 ppt OH detected
following photo-
fragmentation

DOAS R, A, C R, C 12 min - 30% 600 ppt 800-m path length Biermann et al. (1988)

Table 2. (Con't)

Species Methods^a Type^b Development Sample Performance Comments References
stage^c duration
Precision Accuracy MDL^d

NO₃ DOAS R, A, C R, C 12 min - 15% 20 ppt 800-m path length Platt & Perner (1983)

Particulate Denuder/ I, A, IN R, RM 24 h 10% - 40 ng/m^3e Use of denuders Vossler et al. (1988)
NO₃ Filter(s) avoids artifacts;
denuders collect
HNO₃ and NH₃;
teflon and nylon
filters used

N₂O GC-ECD I, A, IN R, RM 15 min 3% - 20 ppb^e -

^a CLM (NO + O₃) = Chemiluminescent using NO + O₃ reaction ^b I = In situ
TP-LIF = Two-photon laser-induced A = Active
TDLAS = Tuneable diode laser absorption spectroscopy C = Continuous
TTFMS = Two-tone frequency modulated spectroscopy P = Passive
PSD = Passive sampling device IN = Integrative
CLM (Luminol) = Chemiluminescent using reaction with Luminol R = Remote
DOAS = Differential optical absorption spectroscopy
DIAL = Differential absorption lidar ^c C = Commercially available
TEA = Triethanolamine R = Research tool
DPA = Diphenylamine L = Laboratory prototype
GC-ECD = Gas chromatography with electron capture detector RM = Routine method
CG-CLM = Gas chromatography with CLM detector
LIF = Laser-induced fluorescence ^d MDL = Minimum detection limit
GC-MS = gas chromatography with mass spectrometer ^e Depends on the sampled air volume (i.e., flow rate and sampling
duration)
^f Uses ion chromatographic or colorimetric analytical finish
is accomplished by extracting the grids into solution and analysing
the extract for NO₂^- by the use of the spectrophotometric or ion
chromatographic method (Miller, 1984). The colorimetric analysis is
calibrated by dilution of gravimetrically prepared nitrite solutions.
The Palmes Tube method was proposed for sampling occupational
exposures where the dosage does not exceed 25 ppm for 8 h (i.e.,
200 ppm-h). The reliability of this method for measuring NO in the
field at the parts-per-billion or parts-per-million level remains to
be demonstrated.

A badge-type sampler similar to the Palmes tube has been devised
by Yanagisawa & Nishimura (1982). This device uses a series of
12 layers of chromium-trioxide-impregnated glass fibre to oxidize NO
to NO₂. This technique is claimed to be more sensitive by
approximately a factor of 10 than the Palmes tube and to have a lower
limit dosage of 0.07 ppm-h.

2.3.1.3 Calibration of NO analysis methods

Calibration of CLM, TP-LIF and TDLAS measurement systems for NO
all rely on compressed gas mixtures of known concentration being
available. Typically compressed gas mixtures are supplied in
passivated aluminium/stainless steel gas bottles certified by the
manufacturer and with NO diluted with N₂ concentration in the rage of
1 to 50 ppm (Schiff et al., 1983; Carroll et al., 1985; Bradshaw et
al., 1985). Calibrations are performed by dynamic dilution of the
reference NO/N₂ mixture with air to give NO concentrations within the
range of 0.1 to 5 ppm.

For passive NO samplers, only the analysis portion of the
procedure is routinely calibrated (using gravimetrically prepared
nitrite solution).

2.3.1.4 Sampling considerations for NO

Oxides of nitrogen are reactive species and exhibit various
solubilities (Table 1). The most inert materials (i.e. glass and
Teflon^TM) are recommended for use in sampling trains. Since ambient
air contains water vapour that may be sorbed on sampling lines,
surface effects may influence the integrity of air samples containing
the more reactive and more soluble NO_y species. In hot, humid
conditions condensation in the sample lines of liquid water from the
air can cause difficulties when analysis equipment is installed in an
air-conditioned environment. To minimize contamination of the system
by dust and foreign matter, it is common practice to sample through an
inert (teflon) sample inlet filter. Of the NO_y species, NO is
probably the least susceptible to surface effects, whereas surface
effects are very important in the sampling of HNO₃.

Nitric oxide reacts rapidly with O₃ to form NO₂. In the
presence of sunlight NO₂ in air photolyses to yield NO and O₃. Thus
in daylight NO, O₃ and NO₂ can exist simultaneously in ambient air
in a condition known as a "photostationary state". The relative
amounts of the three species at any time are influenced by the
intensity of the sunlight present at that moment. Photolysis ceases
when a sample is drawn into a dark sampling line, but NO and O₃ can
continue to react to form NO₂. Therefore residence times in sampling
lines must be minimized to maintain the intensity of the NO/NO₂ ratio
of the sample.

2.3.2 Nitrogen dioxide

Airborne concentrations of NO₂ can be determined by several
methods including CLM, LIF, absorption spectroscopy, including
differential optical absorption spectroscopy (DOAS) and TDLAS, bubbler
and passive collection with subsequent wet chemical analysis. The
most common techniques are chemiluminescence and passive sampling.

2.3.2.1 Chemiluminescence (NO + O₃)

Instruments discussed in this section do not detect NO₂
directly. They sample continuously and rely on the conversion of some
or all of the NO₂ in the air sample to NO, followed by the CLM
reaction of NO and O₃. The NO₂ concentration is calculated from the
difference in the signal given by the sample after passing through the
converter compared to that when the converter is by-passed.

Several methods have been employed to reduce NO₂ to NO (Kelly,
1986). They include catalytic reduction using heated molybdenum or
stainless steel, reaction with carbon monoxide over a gold catalyst
surface, reaction with iron sulfate at room temperature, reaction with
carbon at 200°C, and photolysis of NO₂ to NO by light in the
wavelength range of 320 to 400 nm.

CLM instruments for the determination of NO₂ are readily
available commercially. Field evaluation of nine instruments showed
that the minimum detection limits (MDLs) ranged from 5 to 13 ppb
(Michie et al., 1983; Holland & McElroy, 1986).

Converters may be non-specific for NO₂ and may convert
several other nitrogen-containing compounds to NO, giving rise to
overestimates for NO₂ concentrations. Using commercial instruments,
Winer et al. (1974) found over 90% conversion of PAN, ethyl nitrate
and ethyl nitrite to NO with a molybdenum converter, and similar
responses to PAN and n-propyl nitrate with a carbon converter. With
a stainless steel converter at 650°C, Matthews et al. (1977) reported
100% conversion for NO₂, 86% for NH₃, 82% for CH₃NH₂, 68% for HCN,
1% for N₂O and 0% for N₂. Using a commercial instrument, Joseph &
Spicer (1978) found quantitative conversion of HNO₃ to NO with a

molybdenum converter at 350°C. Similar responses to PAN, methyl
nitrate, n-propyl nitrate, n-butyl nitrate and HNO₃, substantial
response to nitrocresol, and no response to peroxybenzoyl nitrate
(PBzN) were reported with a commercial instrument using a molybdenum
converter at 450°C (Grosjean & Harrison, 1985). These results were
confirmed for PAN and HNO₃ by Rickman & Wright (1986) using
commercial instruments with a molybdenum converter at 375°C and a
carbon converter at 285°C.

Interference from species that do not contain nitrogen have also
been reported. Joshi & Bufalini (1978), using a commercial instrument
with a carbon converter, found significant apparent NO₂ responses
to phosgene, trichloroacetyl chloride, chloroform, chlorine (Cl₂),
hydrogen chloride, and photochemical reaction products of a
perchloroethylene-NO_x mixture. Grosjean & Harrison (1985) reported
substantial responses to photochemical reaction products of Cl₂-NO_x
and Cl₂-methanethiol mixtures and small negative responses to
methanethiol, methyl sulfide, and ethyl sulfide. Sickles & Wright
(1979), using a commercial instrument with a molybdenum converter at
450°C, found small negative responses to 3-methylthiophene,
methanethiol, ethanethiol, ethyl sulfide, ethyl disulfide, methyl
disulfide, hydrogen sulfide, 2,5-dimethylthiophene, methyl sulfide
and methyl ethyl sulfide, and negligible responses to thiophene,
2-methylthiophene, carbonyl sulfide and carbon disulfide.

Methods of sample trapping followed by batch measurement of NO
and NO₂ in the desorbed sample using a chemiluminescence instrument
have been reported. Gallagher et al. (1985) used cryosampling of
stratospheric whole-air samples, and Braman et al. (1986) used
copper(I) iodide coated denuder tubes to sample NO₂ in ambient air.

2.3.2.2 Chemiluminescence (luminol)

A method for the direct chemiluminescence determination of NO₂
was reported by Maeda et al. (1980) and is based on the CLM reaction
of gaseous NO₂ with a surface wetted with an alkaline solution of
luminol (5-amino-2,3-dihydro-1,4-phthalazinedione). The light
emission is strong at wavelengths between 380 and 520 nm. The
intensity of the light can be proportional to the NO₂ concentration
in the sampled air, and the NO₂ concentration can be determined by
calibration of the instrument with air of known NO₂ concentration.

Since the introduction of the luminol method by Maeda et al.
(1980), improvements have been made to develop an instrument
suitable for use in the field (Wendel et al., 1983), and additional
modifications have been made recently to produce a continuous
commercial instrument (Schiff et al., 1986). Detection limits of 5 to
30 ppt and a response time of seconds have been claimed, based on
laboratory tests (Wendel et al., 1983; Schiff et al., 1986). Recent
laboratory evaluation of two instruments has revealed a detection

limit (i.e., twice the standard deviation of the clean air response)
of 5 ppt, and 95% rise and fall times of 110 and 15 seconds (Rickman
et al., 1988). Field tests of the same instruments have shown an
operating precision of ± 0.6 ppb.

2.3.2.3 Laser-induced fluorescence and tuneable diode laser
absorption spectrometry

Two newer techniques that show considerable promise for measuring
NO₂ specifically are photofragmentation/2-photon LIF and TDLAS. The
LIF and TDLAS techniques provide specific spectroscopic methods to
measure NO₂ directly and compare favourably to the sample photolysis-
chemiluminescence technique (Fehsenfeld et al., 1990; Gregory et al,
1990b). For NO₂ concentrations above 0.2 ppb, no interferences were
found for TDLAS (Fehsenfeld et al., 1990).

2.3.2.4 Wet chemical methods

Most wet chemical methods for measuring NO₂ involve the
collection of NO₂ in solution, followed by a colorimetric finish
using an azo dye. Many variations of this method exist, including
both manual and automated versions. These include the Griess-Saltzman
method, the continuous Saltzman method, the alkaline guiacol
method, the sodium arsenite method (manual or continuous), the
triethanolamine-guaiacol-sulfite (TGS) method and the TEA method.
These methods have been reviewed by Purdue & Hauser (1980).

2.3.2.5 Other methods

Several other methods for the determination of NO₂ have been
reported. Atmospheric pressure ionization mass spectrometry has been
investigated for the continuous measurement of NO₂ and SO₂ in
ambient air (Benoit, 1983). Methods employing photothermal detection
of NO₂ have been reported (Poizat & Atkinson, 1982; Higashi et al.,
1983; Adams et al., 1986).

A portable, battery-powered analyser specific to NO₂, which uses
an electrochemical cell as the detector, is commercially available.
By careful selection and design of the cell, levels down to
approximately 0.1 ppm (v/v) can be detected, although with
uncertainties of approximately 20-50%. The detection cell has a
finite life, dependent on the time integral of the NO₂ concentrations
measured. When the cell deteriorates, the instrument typically
develops a gradual drift.

2.3.2.6 Passive samplers

Passive samplers are frequently used in industrial hygiene,
indoor air and personal exposure studies and are less frequently used
for ambient air analysis. Namiesnik et al. (1984) have provided an
overview of passive samplers.

One type of passive NO₂ sampler for ambient application is the
nitration plate. It is essentially an open petri dish containing
TEA-impregnated filter paper. Mulik & Williams (1986) have adapted
the nitration plate concept by adding diffusion barriers in their
design of a passive sampling device (PSD) for NO₂ in ambient and
personal exposure applications. The device employs a TEA-coated
cellulose filter paper, two 200-mesh stainless steel diffusion screens
and two stainless steel perforated plates on each side of the coated
filter to act as diffusion barriers and permit NO₂ collection on both
faces of the filter paper. After sampling, the paper is removed
from the PSD, extracted in water, and analysed for NO_2^- by
ion chromatography. A sensitivity of 0.03 ppm-h and a rate of
2.6 cm³/second were claimed. Comparison of PSD results with
chemiluminescence determinations of NO₂ in laboratory tests at
concentrations between 10 and 250 ppb showed a linear relation and
high correlation (i.e., r = 0.996) (Mulik & Williams, 1987).
Interference from PAN and HNO₂ would be expected (Sickles, 1987).
Results of TDLAS and triplicate daily PSD NO₂ measurements in a
13-day field study showed good agreement between the study average
values but a correlation coefficient for daily results of only 0.47
(Mulik & Williams, 1987; Sickles et al., 1990). The Palmes tube
described in section 2.3.1.2 has been used to sample air in the
workplace and indoor environments to assess personal exposure to NO₂
(Palmes et al., 1976; Wallace & Ott, 1982).

2.3.2.7 Calibration

Calibration methods for NO₂ use permeation tubes or gas-phase
titration (GPT) to generate known concentrations of NO₂.
Calibrations are performed dynamically using dilution with purified
air.

GPT employs the rapid, quantitative gas-phase reaction between
NO, usually supplied as a known concentration from a gas cylinder, and
O₃ supplied from a stable O₃ generator, to produce one NO₂ molecule
for each NO molecule consumed by reaction. When O₃ is added to
excess NO in a titration system, the decrease in NO concentration
(and O₃) is equivalent to the increase in NO₂ produced (US EPA,
1987b).

Use of cylinders of compressed gas containing NO₂ for
calibration purposes (Fehsenfeld et al., 1987; Davis, 1988) is unwise
because of the uncertain stability of the NO₂ concentrations
delivered; this is a consequence of its relatively high boiling point.

2.3.3 Total reactive odd nitrogen

In this monograph, gas-phase total reactive odd nitrogen is
represented by NO_y. Individual components comprising NO_y are gas

phase NO, NO₂, NO₃, N₂O₅, HNO₂, HNO₃, peroxynitric acid
(HO₂NO₂), PAN, and other organic nitrates. NH₃ and N₂O are not
components of NO_y.

Researchers have successfully combined highly sensitive research-
grade CLM NO detectors with catalytic converters that are sufficiently
active to reduce most of the important gas phase NO_y species to NO
for subsequent detection (Helas et al., 1981; Dickerson, 1984; Fahey
et al., 1986; Fehsenfeld et al., 1987).

2.3.4 Peroxyacetyl nitrate

Several methods have been used to measure the concentration of
PAN in ambient air. Roberts (1990) has provided an overview of many
of these methods. A well-developed method is gas chromatography using
electron capture detection (GC-ECD) (Darley et al., 1963; Smith et
al., 1972; Stephens & Price, 1973; Singh & Salas, 1983).

2.3.5 Other organic nitrates

Other organic nitrates (e.g., alkyl nitrates, peroxypropionyl
nitrate and PBzN) can also be present in the atmosphere, but usually
at lower concentrations than PAN (Fahey et al., 1986). In general,
similar methods for sampling, analysis and calibration may be used for
other organic nitrates as are used for PAN (Stephens, 1969). FTIR,
GC-ECD and GC-MS may be used to measure these compounds.

2.3.6 Nitric acid

Several methods are available for the determination of HNO₃
concentrations in the atmosphere. These include filtration (Okita et
al., 1976; Spicer et al., 1978a), denuder tubes (Forrest et al., 1982;
De Santis et al., 1985; Ferm, 1986), CLM (Joseph and Spicer, 1978) and
absorption spectroscopy (Tuazon et al., 1978; Schiff et al., 1983;
Biermann et al., 1988). Many of these techniques carry significant
uncertainties, which have been compared by Hering et al. (1988).

2.3.7 Nitrous acid

Available techniques for the measurement of HNO₂ in ambient
atmospheres employ denuders (Ferm & Sjodin, 1985), annular denuders
(De Santis et al., 1985), CLM (Braman et al., 1986), PF/LIF (Rodgers &
Davis, 1989), absorption spectroscopy (Tuazon et al., 1978; Biermann
et al., 1988) and FTIR (Finlayson-Pitts & Pitts, 1986).

2.3.8 Dinitrogen pentoxide and nitrate radicals

N₂O₅ is readily reduced to NO at temperatures above 200°C and
may be measured nonspecifically as NO₂ with CLM NO₂ analysers
(Bollinger et al., 1983; Fahey et al., 1986).

Ambient concentrations of the NO₃ radical have been measured
using DOAS; concentrations between 1 and 430 ppt have been observed
(Atkinson et al., 1986).

2.3.9 Particulate nitrate

Many methods are available for sampling ambient aerosols,
including impactors, filtration, and filtration coupled with devices
to remove particles larger than a specified size (e.g., elutriators,
impactors and cyclones).

Particulate nitrate samples are generally collected by
filtration, extracted, and analysed directly or indirectly for nitrate
by ion chromatography or colorimetry.

2.3.10 Nitrous oxide

The most commonly used analytical method for N₂O employs GC-ECD.
It has a detection limit of 20 ppb (Thijsse, 1978) and a precision of
± 3% at the background level of 330 ppb (Cicerone et al., 1978).

2.3.11 Summary

Gas-phase CLM instruments have replaced manual (wet) methods
to a large extent in air quality monitoring network applications.
Gas-phase CLM measurement technology permits the determination of NO,
NO₂ and NO_y in the low ppt range. Although CLM NO detectors coupled
with catalytic NO₂ to NO converters are still not specific for NO₂,
they have proved to be useful for measuring NO_y. CLM NO detectors
coupled with photolytic NO₂ to NO converters have shown improved
specificity for NO₂. Most ambient NO₂ monitoring data reported are
from the nonspecific thermal conversing technique.

Passive samplers for NO₂ have been used primarily for workplace
and indoor applications, but hold promise for averaged ambient
measurements as well. GC-ECD is useful in the determination of PAN,
other organic nitrates and N₂O.

2.4 Transport and transformation of nitrogen oxides in the air

2.4.1 Introduction

Oxides of nitrogen are transformed by and removed from the
atmosphere by a complex web of reactions that are fundamental to the
formation and destruction of ozone and other oxidants. The
predominant form of oxidized nitrogen (NO, NO₂, HNO₃, etc.)
in the lower atmosphere varies, depending upon sunlight intensity,
temperature, pollutant emissions, period of time since these emissions
occurred and the meteorological history of an airmass.

2.4.2 Chemical transformations of oxides of nitrogen

2.4.2.1 Nitric oxide, nitrogen dioxide and ozone

The dominant source of nitrogen oxides in the air is combustion
processes (see chapter 3); 90-95% of these nitrogen oxides are usually
emitted as NO and 5-10% as NO₂. NO may be oxidized to NO₂ by
atmospheric oxygen according to reaction 2-1:

NO + NO + O₂ -> 2 NO₂ (2-1)

However at low NO concentrations this reaction is slow and is
important only when NO > 1 ppm (Boström C, 1993). NO concentrations
greater than 1 ppm are not frequently found in ambient air, but they
may possibly occur in indoor air and in plumes from industrial sources
(see Chapter 3). When concentrations are below 1 ppm, NO is oxidized
to NO₂ by two types of reaction. The first type of reaction is given
in equations 2-2 to 2-4. NO can react with O₃:

NO + O₃ -> NO₂ + O₂ (2-2)

Also O₃ is formed when NO₂ is photolysed, forming NO plus an O atom

NO₂ + hnu -> O + NO (2-3)

and O atoms react rapidly with O₂ to form ozone:

M
O + O₂ -> O₃ (2-4)

Thus reactions 2-2, 2-3 and 2-4 recycle O₃ rather than producing a
net increase in O₃ concentrations, where the "M" represents a third
molecule such as N₂, O₂, etc., that absorbs excess vibrational
energy from the newly formed O₃ molecules. However, a second
oxidation path involving the reaction of organic species can lead to
increases in O₃ concentrations and in the conversion rate of NO to
NO₂ (2-9 and 2-10). Organic compounds in the air are commonly
referred to as VOC (volatile organic carbon), ROC (reactive organic
carbon) and non-methane hydrocarbons (NmHC). Urban areas are usually
characterized by significant sources of both nitrogen oxides and ROC
emissions. With suitable atmospheric conditions this can lead to the
formation of photochemical smog. The smog-forming reactions are
initiated by photolytic reactions which produce free radicals, for
example:

(i) the photolysis of O₃

O₃ + hnu -> O₂ + O* (2-5)

O* is an excited form of atomic oxygen, which can react with water to
produce the hydroxyl radical (OH):

O* + H₂O -> 2OH (2-6)

(ii) the photolysis of aldehydes, which also results in the production
of OH. Aldehydes are emitted in motor vehicle exhaust and are
produced in the air by reaction of ROC species with OH. OH is the
most important oxidizing agent in the lower atmosphere; it can react
with all organic compounds, usually forming water and producing an
organic radical.

For a generalized organic compound, R-H (R = CH₃, CHO, CH₂CH₃,
etc.), the principal elements of the reaction sequence are:

R-H + OH -> H₂O + R (2-7)

M
R + O₂ -> RO₂ (fast) (2-8)

RO₂ provides a pathway to oxidize NO to NO₂ without destroying O₃
(unlike reaction 2-2):

RO₂ + NO -> NO₂ + RO (2-9)

RO can undergo reactions that form additional HO₂ or RO₂. HO₂
reacts with NO to form NO₂ and regenerate OH:

HO₂ + NO -> NO₂ + OH (2-10)

In the case of photochemical smog episodes, the quantity of NO_x
emitted into the air determines the ultimate quantity of O₃ that may
be produced. The ROC concentration and sunlight intensity are the
major determinates of the rates at which NO will be oxidized to
produce net increases in NO₂ and O₃ concentrations. Ozone
production is terminated when NO and NO₂ are consumed by reaction to
form products such as HNO₃ (see below), resulting in insufficient NO
concentration for reactions 2-9 and 2-10 to proceed at significant
rates.

In large cities with sunny climates and poor dispersion of
emissions (e.g., Los Angeles and Mexico City), O₃ concentrations in
excess of 200 ppb are not uncommon.

2.4.2.2 Transformations in indoor air

Oxides of nitrogen in indoor air arise from two sources: a)
outdoor air; and b) indoor sources, such as combustion appliances and
heaters. Photochemical reactions do not take place under artificial
lighting, so chemical transformations are limited by the amounts of
oxidizing species (HO₂, O₃, etc.) that arrive in outdoor air, or are
generated by combustion sources.

2.4.2.3 Formation of other oxidized nitrogen species

Oxidation products of NO_x arising from tropospheric
photochemical reactions include HNO₃, HO₂NO₂, HNO₂,
peroxyacylnitrates (RC(O)O₂NO₂), N₂O₅, nitrate radical (NO₃) and
organic nitrates (RNO₃).

Fig. 1 shows a summary for the interconversion pathways for
oxides of nitrogen. These pathways govern urban and indoor air, as
well as "clean" air, but the partitioning between the nitrogen oxide
species varies according to the specific conditions characteristic of
each type of airmass.

a) Nitric acid

Nitric acid is a strong mineral acid that contributes to acidic
deposition from the air. In terms of atmospheric chemistry, HNO₃ is
a major sink for active nitrogen. In daylight, HNO₃ is formed by the
reaction of NO₂ with the OH radical:

M
NO₂ + OH -> HNO₃ (2-11)

This reaction is a chain-terminating step in the free radical
chemistry that produces urban photochemical smog and it removes
reactive nitrogen as well as the hydroxyl radical. Reaction 2-11 is a
relatively fast reaction that can produce significant amounts of HNO₃
over a period of a few hours. At night, in polluted air containing
significant ozone concentrations, the heterogeneous reaction between
gaseous N₂O₅ and liquid water is thought to be a source of HNO₃
(N₂O₅ is produced from NO₃ (see section 2.4.3.5) and NO₂). This
pathway to HNO₃ production is negligible during daytime, because the
NO₃ radical photolyses rapidly and is not present in sufficient
quantities to react with NO₂. The NO₃ radical can also abstract a
hydrogen atom from certain organic compounds (such as aldehydes,
dicarbonyls and cresols) to provide another night-time source of
HNO₃.

Logan (1983) has estimated a lifetime of 1 to 10 days for
HNO₃ in the lower troposphere. The primary removal mechanism is
deposition. The loss of HNO₃ by rain-out is subject to precipitation
frequency while the loss rate by dry deposition varies with the nature

of the ground and vegetation and atmospheric mixing characteristics of
the boundary layer. Chemical destruction mechanisms for HNO₃ also
exist, but their importance is not well understood and is suspected to
be minor for the lower troposphere.

In the presence of NH₃, HNO₃ may form the salt, ammonium
nitrate:

HNO₃(g) + NH₃(g) -> NH₄NO₃ (2-12)

Ammonium nitrate gas readily condenses to the particulate phase.
Ammonium nitrate aerosol can be responsible for significant visibility
reduction and particulate pollution, e.g., where nitric acid is
produced in air from urban areas and this interacts with NH₃ emitted
from agricultural operations.

b) Nitrous acid

HNO₂ is produced from the reaction of NO and OH:

M
NO + OH -> HNO₂ (2-13)

In indoor air other reactions (particularly surface reactions)
may be important sources of nitrous acid.

There have been a few measurements of nitrous acid in urban
environments (Harris et al., 1982; Winer et al., l987). Daytime
levels of nitrous acid are expected to be low because it photolyses
rapidly, yielding NO and ·OH. This reaction probably serves as a
source of OH radicals during the morning in urban regions, where
nitrous acid may form (from NO, NO₂ and H₂O) and accumulate during
the night-time hours. Reaction 2-13 may lead to a build up of nitrous
acid in urban air only during the late afternoon and evening hours
when sunlight intensities are low but some OH radicals are still
present.

c) Peroxynitric acid

While peroxynitric acid (HO₂NO₂) has never been measured in the
atmosphere, it is expected to be present in the upper troposphere.
Models suggest concentrations in the 10 to 100 ppt range at altitudes
above 6 kilometres (Logan, 1983; Singh, 1987). HO₂NO₂ is thermally
unstable, so that boundary layer concentrations are expected to be
extremely low (< 1 ppt). Peroxynitric acid is formed through the
combination of a hydroperoxy (HO₂) radical with NO₂. In the upper
troposphere, HO₂NO₂ is destroyed by photolysis or by reaction with
OH radicals.

d) Peroxyacyl nitrates

Peroxyacetyl nitrate (PAN) is the most abundant of this family of
organic nitrates. The second most abundant homologue, peroxypropionyl
nitrate (PPN), is generally less than 10% of the PAN concentration,
and species with higher relative molecular mass, such as PBzN, are
expected to have even lower concentrations. PAN is a strong oxidant
and is known to be phytotoxic; it is formed from the reaction of
acetylperoxy radical with NO:

CH₃C(O)OO + NO₂ +M -> CH₃C(O)O₂NO₂ +M (2-14)

PAN is thermally unstable and so its lifetime is very dependent
on ambient temperature. For example, PAN lifetimes of about 5 and
20 h have been calculated for 20°C and 10°C, respectively.

In cold conditions PAN can serve as a reservoir for reactive
nitrogen, which is liberated when the temperature of the air is
increased. PAN can be lost from the atmosphere by dry deposition over
land, but it is very likely that a significant fraction of PAN
produced in urban plumes can be transported into the regional
environment.

e) Nitrate radical

The nitrate (NO₃) radical is a short-lived species formed mainly
by the reaction of NO₂ with O₃, although other sources of NO₃
radicals exist (Wayne et al., 1991).

NO₂ + O₃ -> NO₃ + O₂ (2-15)

NO₃ also reacts with NO₂ to form N₂O₅

M
NO₂ + NO₃ -> N₂O₅ (2-16)

Nitrate radicals rapidly photolyse, resulting in a lifetime of
about 5 seconds at midday. They also react rapidly with NO, which
limits their lifetime both during the day- and night-time hours. At
night if atmospheric NO concentrations are approximately 320 pptv,
then the lifetime of NO₃ radicals is similar to that at midday (about
5 seconds).

At night, NO₃ concentrations range from about 0.3 ppt in clean
tropospheric air to 70 ppt in urban areas (Biermann et al., 1988). In
clean background environments, it has been reported that measured NO₃
radical levels are significantly less than those predicted by the
above reactions. Several loss mechanisms have been suggested (Noxon
et al., 1980; Platt et al., 1981): (i) NO₃ radical reaction with

organic compounds; (ii) heterogeneous losses of NO₃ radicals and/or
N₂O₅ on particle surfaces; (iii) reactions of NO₃ radicals with
H₂O vapour; and (iv) reaction of NO₃ radicals with NO.

f) Dinitrogen pentoxide

N₂O₅ is formed from NO₃ and NO₂ (reaction 2-15). Since NO₃
is present only at night, N₂O₅ is also primarily a night-time
species. N₂O₅ is thermally unstable, decomposing to NO₃ and NO₂
(reaction 2-15). At high altitudes in the troposphere, where
temperatures are low, N₂O₅ can act as a temporary reservoir for
NO₃. Dinitrogen pentoxide photolyses at wavelengths less than 330 nm
to give NO₃ and NO₂.

Dinitrogen pentoxide reacts heterogeneously with water to form
HNO₃. This serves as the main night-time production mechanism for
HNO₃ and it provides an important route for removal of oxidized
nitrogen from the atmosphere, since HNO₃ is readily removed by dry
and wet deposition. Other atmospheric reactions of N₂O₅ include its
reaction with gas-phase water to form HNO₃ and possible reactions
with aromatic VOCs such as naphthalene and pyrene (Pitts et al., 1985;
Atkinson et al., 1986). Nitroarenes appear to be the product of
N₂O₅-aromatic reactions.

2.4.3 Advection and dispersion of atmospheric nitrogen species

The transport and dispersion of the various nitrogen species is
dependent on both meteorological and chemical parameters. Advection,
diffusion and chemical transformations dictate the atmospheric
residence time of a particular trace gas. Nitrogenous species that
undergo slow chemical changes in the troposphere and are not readily
removed by depositional processes can have atmospheric lifetimes of
several months. Gases with lifetimes of the order of months can be
dispersed over continental scales and possibly even over an entire
hemisphere. At the other extreme are gases that undergo rapid
chemical transformation and/or depositional losses limiting their
atmospheric residence times to a few hours or less. Dispersion of
these short-lived species may be limited to only a few kilometres from
their point of emission.

Surface emissions are dispersed vertically and horizontally
through the atmosphere by turbulent mixing processes. These processes
are dependent to a large extent on the vertical temperature structure
of the boundary layers and on wind speed. In the vertical dimension,
transport occurs as follows (see also Fig. 2.):

a) the daytime and/or night-time mixed layer; this layer can extend
from the surface up to a few hundred metres at night or to
several thousand metres during the daytime;

b) a layer that can exist during the night-time above a low level
surface inversion and below the daytime mixing height; this layer
generally is situated between 200 and 2000 m altitude;

c) the free troposphere; this transport zone is above the boundary
layer mixing region.

During the warm, summertime period, vertical mixing follows a
fairly predictable diurnal cycle. A surface inversion normally
develops during the evening hours and persists throughout the night-
time and morning period until broken by sunlight heating the surface
of the earth. While the inversion is in place, surface NO_x emissions
can lead to relatively high local concentrations because of restricted
vertical dispersion. Following the break-up of the night-time surface
inversion, vertical mixing will increase and surface-based emissions
will disperse to higher altitudes. The depth of the vertical mixing
during the daytime is often controlled by synoptic weather features.
Temperature inversions aloft, associated with high pressure systems,
are common in many parts of the world.

The dispersion processes described above, coupled with the
chemical transformations of reactive nitrogen compounds, determine the
distances oxidized nitrogen will be transported in the troposphere. A
reasonable understanding exists concerning the short-term (daylight
hours) fate of NO_x emitted in urban areas during the morning hours.
As described above, NO_x emitted in the early morning hours in an
urban area will disperse vertically and move downwind as the day
progresses. On sunny summer days, most of the NO_x will have been
converted to HNO₃ and PAN by sunset. Much of the HNO₃ will be
removed by depositional processes as the air mass moves along. After
dusk, an upper portion of the daytime mixed layer will be decoupled
from the surface because of formation of a low-level radiation
inversion. Transport will continue in this upper level during the
night-time hours and, although photochemical processes will cease,
dark-phase chemical reactions can proceed. Peroxyacetyl nitrate and
HNO₃, if carried along in this layer, can be transported long
distances.

2.4.3.1 Transport of reactive nitrogen species in urban plumes

Overall removal rates for reactive nitrogen species during
daytime at mid-latitudes have been measured or calculated for a few
areas. For example, in the plume from Boston, USA, after correction
for dilution, removal rates ranged from 0.14 to 0.24 h^-1 on 4 days
(Spicer, 1982, Altshuller, 1986). In Los Angeles and Detroit, the
removal rate has been estimated to be 0.04-0.1 h^-1 (Calvert, 1976;
Chang et al., 1979; Kelly, 1987). Formation and removal of HNO₃ is
thought to be the rate-controlling step for removal of reactive
nitrogen.

2.4.3.2 Air quality models

Air quality models are mathematical descriptions of pollutant
emissions, atmospheric transport, diffusion and chemical reactions
of pollutants. However, air quality models are very complex and
difficult to test for validity. Inputs include emissions, topography
and meteorology of a region. Air quality models represent an
integration of knowledge for the chemistry and physics of the
atmospheric system; they offer some predictive capability for the
effectiveness of pollution control strategies. Models have also been
developed for indoor air.

2.4.3.3 Regional transport

Transport of reactive nitrogen species in regional air masses can
involve several mechanisms. Mesoscale phenomena, such as land-sea
breeze circulations or mountain-valley wind flows, will transport
pollutants over distances of ten to hundreds of kilometres. On a
larger scale, synoptic weather systems such as the migratory highs
that cross the eastern USA and other areas of the world in the
summertime influence air quality over many hundreds of kilometres.
The accumulation and fate of nitrogen compounds will differ somewhat
between the mesoscale and synoptic systems. Mountain-valley and land-
water transport mechanisms have dual temporal scales because of their
dependence on solar heating. However, in the larger-scale synoptic
systems, reactive nitrogen species can build up over multiday periods.
The residence time of air parcels within a slow-moving high pressure
system can be as long as 6 days (Vukovich et al., 1977).

In many cases, the transport mechanisms mentioned above are
interrelated. Mountain-valley or land-water breezes can dictate
pollutant transport in the immediate vicinity of sources, but the
eventual fate of reactive nitrogen species will be distribution into
the synoptic system.

2.5 Conversion factor for nitrogen dioxide

1 ppm = 1.88 mg/m³
1 mg/m³ = 0.53 ppm

2.6 Summary

Combustion provides the major source of oxides of nitrogen in
both indoor and outdoor air, producing mostly NO with some NO₂. The
sum of NO and NO₂ is generally referred to as NO_x. Once released
into the air, NO is oxidized to NO₂ by available oxidants,
particularly O₃, and by photochemical reactions involving reactive
organic compounds. This happens rapidly under some conditions in
outdoor air; for indoor air, it is generally a much slower process.
Nitrogen oxides are a controlling precursor of ozone and smog

formation; interactions of nitrogen oxides (except N₂O) with reactive
organic compounds and sunlight form ozone in the troposphere and smog
in urban areas.

In both indoor and outdoor air, NO and NO₂ may undergo reactions
to form a suite of other nitrogenous species including HNO₂, HNO₃,
NO₃, N₂O₅, PAN and other organic nitrates. The complete suite of
gas-phase nitrogen oxides is referred to as NO_y. The partitioning
of nitrogen among these compounds is strongly dependent on the
concentrations of other oxidants, sunlight exposure, the presence of
reactive organic compounds and the meteorological history of the air.

A sensitive, specific and reliable analytical method exists for
measuring NO (by the chemiluminescent reaction with ozone), but this
is an exception for NO_y species. Chemiluminescence is also the
most common technique used for NO₂, which is first reduced to NO.
Unfortunately, the method of reduction usually used is not specific
for NO₂, and it has various conversion efficiencies for other
oxidized nitrogen compounds that may also be present in the air
sample. For this reason, care must be taken in interpreting the NO₂
values given by the common chemiluminescence analyser, as the signal
may include responses from interfering compounds. Additional
difficulties arise from nitrogen species such as HNO₃ that may
partition between the gas and particulate phases both in the
atmosphere and in the sampling procedure.

3. SOURCES, EMISSIONS AND AIR CONCENTRATIONS

3.1 Introduction

Oxides of nitrogen can have significant concentrations in ambient
air and in indoor air. The types and concentrations of nitrogenous
compounds present can vary greatly from location to location, with
time of day, and with the season. The main sources of nitrogen oxides
emissions are combustion processes. Fossil fuel power stations, motor
vehicles and domestic combustion appliances emit nitrogen oxides,
mostly in the form of NO but with some (usually less than about 10%)
in the form of NO₂. In the air chemical reactions occur which
oxidize NO to NO₂ and other products (chapter 2). Also, there are
biological processes in soils which liberate nitrogen species,
including N₂O. Emissions of N₂O can cause perturbation of the
stratospheric ozone layer.

Human health may be affected when significant concentrations of
NO₂ or other nitrogenous species, such as PAN, HNO₃, HNO₂ and
nitrated organic compounds, are present. In addition, nitrates and
nitric acid can cause significant effects on ecosystems when deposited
on the ground.

Indoors, the use of combustion appliances for cooking and heating
can give rise to greater NO and NO₂ concentrations than are present
outdoors, especially when the appliance is not vented to the outside.
Recent research has shown that in these circumstances nitrous acid can
reach significant concentrations (Brauer et al., 1993).

This chapter discusses both ambient and indoor sources
of nitrogenous compounds, their emissions, and the resulting
concentrations that may directly affect human health or participate
in atmospheric chemical pathways leading to effects on human health
and welfare. Nitrogen-containing compounds are also of particular
interest because of their secondary impacts. For example, production
of photochemical smog and ozone pollution depends on emissions to the
air of nitrogen oxides together with volatile organic compounds.
Nitric acid, which is produced in the air by the reaction of hydroxyl
radicals (OH*) with NO₂, is one of the major components of acidic
precipitation. As well as being present in the gas phase, oxidized
nitrogen can, by reaction and adsorption, become incorporated into
aerosol particles. Graedel et al. (1986) identified 20 inorganic
nitrogen-containing species detectable in the atmosphere. Near
cities and urban regions the species usually present in greatest
concentrations are NO and NO₂, and these are the most reliably
measured and frequently monitored nitrogen oxide species.

Knowledge of emission patterns and concentrations of nitrogenous
compounds is critically important for air quality planning and human
health and environment risk assessments. Because nitrogen oxides and

their reaction products have lifetimes of several days in the
atmosphere, they can be transported long distances by the wind and
give rise to environmental impacts far from their source of emission.

3.2 Sources of nitrogen oxides

Combustion systems emit NO and NO₂ and together these species
are usually denoted as NO_x.

When NO_x emissions are expressed in mass units, the mass is
expressed as if all the NO had been converted to NO₂. Another
convention adopted in some of the following sections is to report the
emissions on a mass basis in terms of the nitrogen content.

3.2.1 Sources of NO_x emission

3.2.1.1 Fuel combustion

Annual production of NO_x from combustion of fossil fuels is
typically estimated from emission factors for various combustion
processes, combined with worldwide consumption data for coal, oil and
natural gas. Logan (1983) provided a tabular summary of emission
factors, which has been updated by the US National Acid Precipitation
Assessment Program (Placet et al., 1991). Owing to variations in
process operating conditions, the emission factors must be considered
to be uncertain by about ± 30%. Table 3 provides a summary of global
emission estimates for NO_x according to fuel type. The estimates of
Logan (1983) are slightly higher than those of Ehhalt & Drummond
(1982), the largest discrepancies being in emission estimates for the
transportation sector. The differences arise because Logan (1983)
based estimates of emissions on fuel usage, while Ehhalt & Drummond
(1982) scaled the totals somewhat indirectly by using world automobile
population numbers.

Dignon (1992) has assembled a database for mapping (with a
resolution of one degree in latitude and longitude) and estimated
global NO_x and sulfur oxides emissions from their common principal
anthropogenic source, i.e. fossil fuel combustion. For 1980, the
global total was estimated to be 22 million tonnes, as nitrogen.
Countries heading the list (in millions of tonnes of nitrogen per
year) were: USA, 6.4; USSR, 4.4; China, 1.7; Japan, 0.80; and Federal
Republic of Germany, 0.66. An estimated 95% of NO_x emissions from
fossil fuel combustion originates in the northern hemisphere.

For oceanic regions, shipping is a source of NO_x emissions.
Aircraft also emit nitrogen oxides and this may be significant for the
upper troposphere and stratosphere.

Table 3. Estimates of global emissions of nitrogen oxides (NO_x) from combustion of fossil fuels and biomass (from: US EPA, 1993)^a

Source type Annual consumption Emission factors^b Global source strength
(10⁶ tonnes, unless indicated otherwise) (10⁶ tonnes nitrogen/year)

(E & D) (L) (C et al.) (E & D) (L) (E & D) (L) (C et al.)

Fossil fuels^c

Hard coal 2150 2696 - 1.0-2.8 2.7 3.9 (1.9-5.8) 6.4 -
Lignite 810 - 0.9-2.7 1.6 (0.8-2.3) -
Light fuel oil 300 1.39 - 1.5-3.0 2.2^d 0.7 (0.5-0.9) 3.1 -
Heavy fuel oil 470 1.5-3.1 1.1 (0.7-1.5) -
Natural gas 1.04 1.2 × 10⁹ m³ - 0.6-3.0 1.9^d 1.9 (0.6-3.1) 2.3 -
Industrial sources - - - 1.2 -
Automobiles (4.1-5.4) 1.0 × 10⁹ m³ - 0.9-1.2^e 8.0^d 4.3 (3.7-6.4) 8.0
× 10¹² km

Total 13.5 (8.2-18.5) 19.9

Biomass burning^f

Savanna (6-14) × 10³ 2000 1200 1.0 1.7 3.1 (1.8-4.3) 3.4 2.1
Forest clearings (2.7-6.7) × 10³ 4100 2700 1.0-1.6 2.0 2.1 (0.8-3.4) 8.2 4.7
Fuel wood - 850 1100 - 0.5 2.0 (1-3) 0.4 0.5
Agricultural waste - 15 1900 - 1.6 4.0 (2-6) 0.02 3.3

Total 11.2 (5.6-16.4) 12.0 10.6

^a Estimates according to Ehhalt & Drummond (1982) (E & D) and Logan (1983) (L). Ranges are given in parentheses.
^b Emission factors refer to grams of nitrogen per kg of fuel consumed, unless indicated otherwise
^c Petroleum refining and manufacture of nitric acid and cement; global emissions were obtained by scaling USA emissions for each
industrial process
^d Grams of nitrogen per m³ of fuel consumed
^e Grams of nitrogen per km
^f For biomass-burning, Crutzen et al. (1979) (C et al.) have given annual consumption rates differing somewhat from those of the other
authors. The data of Crutzen et al. (1979) and the resulting nitrogen oxides production rates are included for comparison
3.2.1.2 Biomass burning

Table 3 includes a breakdown of estimates for release of NO_x
from burning of biomass. In natural fires and the burning of wood,
temperatures are rarely high enough to cause oxidation of nitrogen
molecules of the air. The emissions are thereby more closely related
to the fixed nitrogen content of the fuel. Logan (1983) reviewed a
number of experimental determinations of nitrogen emission factors
that indicate yields are highest for grass and agricultural refuse
fires (1.3 g nitrogen/kg fuel), less for prescribed forest fires
(0.6 g nitrogen/kg fuel), and still lower for burning of fuel wood in
stoves and fireplaces (0.4 g nitrogen/kg fuel). The values roughly
reflect differences in nitrogen content of the materials burned.
Biomass burning is mainly associated with agricultural practices in
the tropics, which include plant, slash, and shift practices as well
as natural or intentional burning of savanna vegetation at the end of
the dry season. Forest wildfires and use of wood as fuel make a
lesser contribution.

3.2.1.3 Lightning

Thunderstorm activity has been viewed as a major NO_x source
since 1827, when Von Liebig proposed it as a natural mechanism for
fixation of atmospheric nitrogen. Electrical discharges in air
generate NO_x by thermal dissociation of nitrogen molecules due to
ohmic heating inside the discharge channel and shockwave heating of
the surroundings. Laboratory studies by Chameides et al. (1977) and
Levine et al. (1981) indicate an NO_x yield of 6 × 10¹⁶ molecules per
joule of spent energy. Great uncertainties exist, however, about the
total energy generated by lightning in the atmosphere. Noxon (1976,
1978) first studied the increase of NO_x in the air during a
thunderstorm. His results provide the basis for many of the estimates
shown in Table 4. Reviews by Kowalczyk & Bauer (1981) Borucki &
Chameides (1984) and Albritton et al. (1984) provide a best estimate
of annual generation by lightning: 1 million tonnes of NO_x in North
America and 13 million tonnes globally (Placet et al., 1991).

3.2.1.4 Soils

The biochemical release of NO_x from soils is poorly understood,
and the flux estimates must be viewed with caution. Both rely on the
observations by Galbally & Roy (1978), who used the flux box method in
conjunction with chemiluminescence detection of NO_x. They found
average fluxes of 5.7 and 12.6 µg nitrogen/m²*h on ungrazed and
grazed pastures, respectively, where NO was the main product. More
recent measurements of Slemr & Seiler (1984) indicate that the release
of NO_x from soils depends critically on the temperature and moisture

content of the soil, which in turn complicates the estimate of the
global emissions. Slemr & Seiler (1984) also found an average release
rate of 20 µg nitrogen/m² per h for uncovered natural soils, evenly
divided between NO and NO₂. Grass coverage reduced the escape flux,
whereas fertilization enhanced it. Ammonium fertilizers were about
five times more effective than nitrate fertilizers. This suggests
that nitrification as a source of NO_x is more important than
denitrification. According to Slemr & Seiler (1984), an annual global
flux of 10 million tonnes of nitrogen represents an upper limit to the
release of NO_x from soils. Galbally et al. (1985) presented more
detailed estimates for arid lands, and Table 4 provides a compilation
of current literature used to develop the global budgets. Soil is
also a source of N₂O and NH₃ emissions.

In the presence of low concentrations, plants can emit NH₃,
rather than absorb it. This is especially true with scenescing and
with highly fertilized plants (Grünhage et al., 1992; Holtan-Hartwig &
Bockman 1994; Fangmeijer et al., 1994). Release to the atmosphere of
N₂ and NO by plants has also been reported. In some cases this was
part of the response following exposure to nitrogen-containing
pollutants, but other mechanisms are involved (Wellburn, 1990). NO and
N₂O are emitted in significant quantities by the soil. The reason
why the deposition velocity of NO is relatively low see (see Table 5)
is partly due to the fact that the downward flux (and uptake by the
canopy) is "mathematically" compensated by soil emissions. In other
words: a low deposition velocity does not always mean that the uptake
by the vegetation is low. In the case of N₂O, soil emissions are
mostly larger than deposition; this emission is the result of
denitrification and is positively related to the nitrogen and water
content and the temperature of the soil. This is why the release of
nitrogen from the ecosystem in the form of N₂O is dependent on the
ecosystem type, climate and land use (fertilization and water table
height). Skiba et al. (1992) estimated for the United Kingdom the NO
and N₂O emissions from agricultural land to be 2-6% of the nationwide
NO_x emissions and 16-64% of the N₂O emissions, respectively.

Table 4. Global and North America natural emissions (average and range) of nitrogen oxides (NO_x)
from lightning, soils and oceans

Global North America Reference
(10⁶ tonnes/year) (10⁶ tonnes/year)

Lightning 8.6 (2.6-26) Borucki & Chameides (1984)
18 1.7 Albritton et al. (1984)
13 (7-26) 1 (0.3-2) Kowalczyk & Bauer (1981); Placet et al. (1991)

Soils 50 (as NO₂) Lipschultz et al. (1981)
30 (as NO) Levine et al. (1984); Galbally & Roy (1978)
36 Slemr & Seiler (1984)
2 Placet et al. (1991)

Oceans 0.35 Zafiriou & McFarland (1981); Logan (1983)

Table 5. Deposition velocity of nitrogen-containing gases and
aerosols

Deposition velocity Reference
(mm/second)

NO₂ 0.1-10 Grennfelt et al. (1983);
Anonymous (1991)

NO 0.2-1 Prinz (1982)

NH₃ 12 (-5 - +40) Grünhage et al. (1992);
Sutton et al. (1993);
Fangmeijer et al. (1994);
Holtan-Hartwig & Bockman (1994)

NH₄⁺ 1.4 (0.03-15) Fangmeijer et al. (1994)

Estimates of global emissions of N₂O and ammonia are summarized
in Table 6.

Table 6. Annual global estimates (average and range) of N₂O and
NH₃ emissions to the troposphere (10⁶ tonnes
of nitrogen)

Source N₂O NH₃ Reference

Soils 10 (2-20) 15 Dawson (1977);
Boettger et al. (1978)

Ocean 26 (12-38) Hahn (1981)

Biomass burning 2 2-8 Crutzen et al. (1979);
Crutzen (1983)

Fossil fuels 1.6 0.2 Weiss & Craig (1976);
Boettger et al. (1978)

Fertilizer 0.1 3 Boettger et al. (1978);
Crutzen et al. (1979);
Crutzen (1983);
Stedman & Shetter (1983)

Domestic animals 22 Soederlund & Svensson (1976)
Boettger et al. (1978);
Crutzen et al. (1979);
Crutzen (1983);
Stedman & Shetter (1983)

3.2.1.5 Oceans

There have been few measurements of NO_x, N₂O or NH₃ fluxes
over the ocean, and current literature suggests that the sea is a
negligible source of NO. Zafiriou & McFarland (1981) observed a
supersaturation of seawater with regard to NO in regions of relatively
high concentrations of nitrite, owing to upwelling conditions. The

excess NO must, in this case, arise from photochemical decomposition
of nitrite by sunlight. Logan (1983) estimated a local source
strength of 1.3 × 10¹² molecules/m² per second under these
conditions. Linear extrapolation results in an annual global flux
estimate of 350 000 tonnes of nitrogen.

3.2.2 Removal from the ambient environment

Wet precipitation and dry deposition provide two of the major
mechanisms for removal of NO_x from the atmosphere. The addition to
the plant soil ecosystem of nitrate (and ammonium) by rainwater
constitutes an important source of fixed nitrogen to the terrestrial
biosphere, and until 1930 practically all studies of nitrate in
rainwater were concerned with the input of fixed nitrogen into
agricultural soils. Eriksson (1952) and Boettger et al. (1978) have
compiled many of the available data. Despite the wealth of
information, it remains difficult to derive a global average for the
deposition of nitrate, because of an uneven global coverage of the
data, unfavourably short measurement periods at many locations, and
inadequate collection and handling techniques for rainwater samples.
In addition, the concentration of nitrate in rainwater has increased
in those parts of the world where the utilization of fossil fuels has
led to a rise in the emissions of NO_x, i.e. primarily western Europe
and the USA.

Dry deposition is important as a sink for those gases that are
readily absorbed by materials covering the earth surface. In the
budget of NO_x, the gases affected most by dry deposition are NO₂ and
HNO₃. The deposition velocity of NO is too small and the
concentration of peroxyacetyl nitrates is not high enough for a
significant contribution.

According to Grennfelt et al. (1983) and Wellburn (1990), NO_3^-
and HNO₃ have a higher deposition velocity then NH₃, but this was
not quantified. HNO₂ is assumed to have a deposition velocity equal
to SO₂: 1-30 mm/second (Table 5).

There are several other nitrogen-containing air pollutants with
relatively high deposition velocities. These generally add only small
amounts to the total nitrogen deposition, because most of the time
their ambient concentrations are relatively low.

Atmospheric nitrogen deposition can significantly change the
chemical composition of the soil. In the rooting zone these changes
have an impact on vegetation. The changes in deeper soil layers
are particularly relevant if groundwater is used as a source of
drinking-water. Groundwater under fertilized agricultural land can be
heavily polluted with nitrate (and aluminium), but this is beyond the
scope of this chapter. Due to atmospheric nitrogen deposition, the

groundwater under forests and other non-fertilized vegetation can
become polluted with nitrate. For instance, in 20% of the forested
area of the Netherlands, the nitrate concentration in phreatic
groundwater is higher than 50 mg/litre (the EC drinking-water
standard); in 37% it is higher than 25 mg/litre (Boumans & Beltman,
1991). The average annual nitrogen deposition in the Netherlands is
45 kg/ha; approximately 10 kg/ha is from dry deposition of NO_x. The
nitrate concentration in groundwater is strongly related to the soil
type. With the same atmospheric deposition, the nitrate concentration
increases as follows: peaty soils < moderately drained sandy soils
< well-drained rich sandy soils (Boumans, 1994). A distinct relation
also exists concerning the age of the trees: tree stands in Wales
showed nitrate leaching (measured in the stream water draining the
catchments), but only with stands older than 30 years. Younger trees
used the nitrogen as nutrient, but the nitrogen demand of the older
trees was lower. The annual nitrogen deposition in that region was
estimated to be 20 kg/ha (Emmett et al., 1993).

3.2.3 Summary of global budgets for nitrogen oxides

The principal routes to the production of NO_x are combustion
processes, nitrification and denitrification in soils, and lightning
discharges. The major removal mechanism is oxidation to HNO₃,
followed by wet and dry deposition. In developing Table 7, the dry
deposition velocities for NO₂ over bare soil, grass and agricultural
crops were assumed to fall in the range of 3 to 8 mm/second. However,
over water the velocities are significantly smaller, so that losses of
NO₂ by deposition onto the ocean surface can be ignored. The
absorption of nitric acid by soil, grass and water is rapid, and dry
deposition correspondingly important, but the global flux is difficult
to estimate because information on HNO₃ mixing ratios is still
sparse. Logan (1983) adopted NO mixing ratios of 50 pptv over the
oceans and 100 pptv over the continents. The mixing ratios assumed
for NO₂ were 100 and 400 pptv, respectively. Allowance was made for
higher mixing ratios in industrialized areas affected by pollution.
Logan (1983) included the deposition of particulate nitrate over the
oceans, using a settling velocity of 3 mm/second. This process
contributes 2 million tonnes nitrogen/year to a total dry deposition
rate of 12 to 22 million tonnes nitrogen/year.

Efforts by Boettger et al. (1978), Ehhalt & Drummond (1982),
Galbally et al. (1985) and Warneck (1988) to quantify the sources and
sinks have led to an improved understanding of the global budget of
NO_x, in which the flux of NO_x into the troposphere and the rate of
nitrate deposition are approximately balanced. Ehhalt & Drummond
(1982) relied on the detailed evaluation of data by Boettger et al.
(1978). Their analysis emphasized measurements from the period 1950
to 1977, and they prepared a world map for nitrate deposition rates,

which were then integrated along 5° latitude belts. Logan (1983)
considered recent network data from North America and Europe; Galloway
et al. (1982) reported measurements of nitrate in precipitation at
remote locations in Alaska, South America, Australia and the Indian
Ocean. Both estimates gave wet nitrate deposition rates in the range
of 2 to 14 million tonnes nitrogen/year for the marine environment and
8 to 30 million tonnes nitrogen/year on the continents. An earlier
appraisal by Soederlund & Svensson (1976) led to rather similar
values, i.e. 5 to 16 and 13 to 30 million tonnes nitrogen/year,
respectively, although it was primarily based on Eriksson's (1952)
compilation of data from the period 1880 to 1930.

Table 7. Global budget (average and range) of nitrogen oxides
in the troposphere (from US EPA, 1993)^a

Type of source or sink Global flux
(10⁶ tonnes nitrogen/year)

Ehhalt & Logan (1983)
Drummond (1982)

Production

Fossil-fuel combustion 13.5 (8.2-18.5) 21 (14-28)
Biomass burning 11.2 (5.6-16.4) 12 (4-24)
Release from soils 5.5 (1-10) 8 (4-16)
Lightning discharges 5.0 (2-8) 8 (2-20)
NH₃ oxidation 3.1 (1.2-4.9) uncertain (1-10)
Ocean surface (biologic) - < 1
High-flying aircraft 0.3 (0.2-0.4) -
Stratosphere 0.6 (0.3-0.9) approx. 0.5

Total production 39 (19-59) 50 (25-99)

Losses

Wet deposition of NO₃^-, land 17 (10-24) 19 (8-30)
Wet deposition of NO₃^-, oceans 8 (2-14) 8 (4-12)
Wet deposition, combined 24 (15-33) 27 (12-42)
Dry deposition of NO_x - 16 (12-22)

Total loss 24 (15-40) 43 (24-64)

^a Derived from estimates according to Ehhalt & Drummond (1982)
and Logan (1983)

On continents, one should also consider the interception of
aerosol particulates by high growing vegetation. The interception of
nitrate is expected to be particularly effective. Hoefken &
Gravenhorst (1982) studied the enrichment of nitrate in rainwater
collected underneath forest canopies compared to that collected in
open areas outside forests. The effect is caused by the wash-off of
dry-deposited material from foliage. Hoefken & Gravenhorst (1982)
found that, in a beech forest, nitrate was enhanced by a factor of
1.4, whereas in a spruce forest enhancement by a factor of 4.1
occurred. Unfortunately, they were unable to differentiate between
contributions of particulate nitrate versus gaseous nitrate to the
total dry deposition.

If losses of NO₂ and HNO₃ by dry deposition are included in the
total budget of NO_x, one obtains a reasonable balance between the
sources and sinks, as Table 7 shows. Ehhalt & Drummond (1982) noted
that an appreciable part of their dry deposition is already included
in their wet deposition rates, because rain gauges frequently are left
open continuously, so that the collection of nitrate occurs during
both wet and dry periods. For NO₂, they estimated a dry deposition
rate of 7 million tonnes nitrogen/year. Because of the uncertainty,
they chose to include it in the error bounds and not in the mean value
of total NO_x-derived nitrogen deposition. Clearly, the total budget
of NO_x is far from being well defined. Moreover, in view of the
relatively short residence times of chemical species involved in the
NO_x cycle, it is questionable whether a global budget gives an
adequate description of the tropospheric behaviour of NO_x and its
reaction products. Supplemental regional budgets could be more
appropriate.

3.3 Ambient concentrations of nitrogen oxides

Because cities usually have an aggregation of emissions sources
ambient concentrations of NO and NO₂ tend to be greatest in cities.
High concentrations of NO are common in street canyons, owing to motor
vehicle emissions. In rural areas the emissions may have spent
considerable time in the atmosphere and have undergone reactions to
produce significant concentrations of other species, such as HNO₃ and
PAN.

3.3.1 International comparison studies of NO_x concentrations

Data for monthly average concentrations of NO_x collected by the
World Meteorological Organization at five background locations in
Europe for the period 1983 to 1985 are summarized in Fig. 3 (WMO,
1988, 1989). Fig. 4 presents published monthly averages of NO₂ in
1987 for 12 stations in a cooperative network under the Organisation
for Economic Co-operation and Development (OECD) (Grennfelt et al.,
1989). These two figures show that concentrations of both NO_x and
NO₂ tend to be higher during winter months.

Measurements of NO₂ in several countries during the late 1970s
and early 1980s are summarized in "Assessment of Urban Air Quality"
(WHO, 1988). The trends in composite annual averages for urban NO₂
monitoring stations in five countries are portrayed in Fig. 5 for the
period 1975 to 1985. The trend in the Canadian data appears to have
been downward, but essentially stable trends were evident for data
from the other countries. Annual averages in the 1980-1984 period for
42 cities around the world are summarized in the same report (WHO,
1988). During that period, only one city, Sao Paulo, reported an
annual average greater than 0.053 ppm (100 µg/m³).

Short-term peak values (1-h or 30-min maxima, or 98th or 95th
percentile values) have been reported for 18 cities during the
1980-1984 period (WHO, 1988). Ten of these cities (Amsterdam,
Brussels, Hamilton, Hong Kong, Jerusalem, Montreal, Munich, Rotterdam,
Tel Aviv and Toronto) reported values above the WHO 1-h guideline
level of 400 µg/m³ (0.21 ppm) for at least one year during that
5-year period. For eleven cities in the WHO report, both the annual
average and a "1-hour" peak statistic were reported for the 1980-1984
period. Fig. 6 compares these two statistics. It shows that three
cities, Amsterdam, Jerusalem and Tel Aviv, reported an average peak
value above the WHO 1-hour guideline value of 400 µg/m³ (0.21 ppm).
It should be kept in mind that the peak-value statistic is more
susceptible to undetected spurious measurements than is the annual
average. Data from the remaining eight cities place them in the
quadrant below the target levels for both the annual average and the
1-hour peak. A similar situation is seen in the majority of cities in
the USA and is discussed in the next section.

More recent data on NO₂ trends in the world's largest cities
have been reported by WHO/UNEP (1992) in the monograph "Urban Air
Pollution in Megacities of the World". Such trends for six selected
cities from various regions of the world are illustrated in Fig. 7, a
composite of figures extracted directly from the WHO/UNEP (1992)
report. In general, the overall trends appeared to be relatively
stable for most of the cities (and/or specific neighbourhoods).
However, there were a few exceptions, e.g., an apparent decrease in
the late 1980s for Bombay and an apparent increase during the same
period for some areas of Moscow. There are substantial differences in
the concentrations reported for different cities.

Table 8 summarizes emissions of nitrogen oxides and ambient
monitoring data from the WHO/UNEP (1992) report for the years
indicated. Included are estimates for total emissions and percentages
attributed to mobile sources, primarily private motor vehicles and
public land transport systems. However, the quality and type of
information contained in the report is mixed, reflecting a variety of
monitoring methods and reporting policies in different countries.
Ambient data in some cities was reported as NO_x, and in others as
NO₂; reporting periods varied from one hour to one year.

Table 8. Estimated mobile and stationary source emissions of nitrogen oxides in
megacities (from: WHO/UNEP, 1992)^a

City Total emissions of Mobile source Ambient concentration
nitrogen oxides contribution (µg/m³)
(tonnes/year) (%)

Bangkok 60 000 (1990) 30 max 1 h NO_x (as NO₂)
270 at one site; < 320 at
three stations (1987)

Beijing na

Bombay 56 000 (1990) 52 NO₂ 70-85 (annual 98th
percentile, 1990)

Buenos Aires 27 000 (1989) 48 na

Cairo 24 700 (1989) 23 NO_x 380-1400 (1979,
monthly means; single
study)

Calcutta 36 550 (1990) 29

Delhi 73 000 (1990) 20 NO₂ 500 (1990, 8 h)
(mostly diesel)

Jakarta 20 500 (1989) 75 NO_x 28 (1990, annual mean)

Karachi 50 000 (1989) 38 38-544 (12-13 June 1988;
single study)

Table 8. (Con't)

City Total emissions of Mobile source Ambient concentration
nitrogen oxides contribution (µg/m³)
(tonnes/year) (%)

London 79 000 (1983) 75 (1984) NO₂ max 1 h 867; > 600
for 8 h; > 205 for 72 h
(episode 12-15 Dec. 1991);
98th percentile > 135;
50th percentile > 50 (1989);
NO recorded but not
reported

Los Angeles 440 000 (1987) 76 NO₂ max 1 h 526; > 400
at 8 out of 24 stations (1990)

Manila 119 000 (1990 - 90 na
dubious accuracy)

Mexico City 177 300 (1991) 75 NO₂ hourly maxima
301-714 (1986-91)

Moscow 210 000 (1990) 19 NO₂ max daily means
100-150

New York 120 000 New York na NO₂ 1 h max 402; daily
City; 513 000 New max 160; annual mean 87
York metropolitan (1990)
area (1985)

Rio de Janeiro 63 000 (1978) 92 na

Sao Paulo 245 000 (1988) 82 NO₂ max 1 h
600-1500 (1988)

Table 8. (Con't)

City Total emissions of Mobile source Ambient concentration
nitrogen oxides contribution (µg/m³)
(tonnes/year) (%)

Seoul 270 000 (1990) 78 NO₂ annual means only

Shanghai 127 000 (1983); na NO_x annual mean 50;
1991 emissions indoor level 90
assumed 50%
higher, i.e.
approx. 190 000

Tokyo 52 700 (1985) 67% from motor daily mean 98th percentile
vehicles; 5% from > 115 tolerable level at
ship and aircraft 25% of stations

^a na = not available

As shown in Table 8, of importance for air quality management is
the large contribution of NO_x from motor vehicles reported for some
cities and the continuing growth in this contribution. For example,
emissions from vehicles in Bombay (about 29 000 tonnes per year in
1990) are expected to increase by an additional 14 600 tonnes/year by
the year 2000 (WHO/UNEP, 1992).

Estimates for Jakarta attribute some three-quarters of NO_x
emissions to motor vehicles, which is comparable with London, Los
Angeles and Mexico City. Data from Manila indicate that some 90% of
NO_x originates from motor vehicles.

3.3.2 Example case studies of NO_x and NO₂ concentrations

Data from a range of countries and locations are given in Table 9
(Agra, India) and Tables 10 and 11 (various cities in China).

Table 9. Concentrations of NO₂ measured in the vicinity of the
Taj Mahal, Agra India^a

Year Mean monthly concentration range (µg/m³)

1987 5.5 to 41.9
1988 6.3 to 33.1
1989 4.2 to 15.2

^a Highest concentrations tend to occur in winter
Personal communication from R.R. Khan, Ministry of Environment and
Forests, New Delhi, India (1994)

In urban areas in the USA, hourly patterns at fixed-site ambient
air monitors often follow a bimodal pattern of morning and evening
peaks, related to motor vehicular traffic patterns. Sites affected by
large stationary sources of NO₂ (or NO that reacts to produce NO₂)
are often characterized by short episodes at relatively high
concentrations, as the plume moves to downwind areas.

Since 1980, the annual average level among NO₂-reporting
stations in the USA has been below 0.03 ppm, with no significant
trend evident. This is exemplified in Fig. 8 (US EPA, 1991) by
annual averages for the period 1980 to 1989 for 60 metropolitan
areas subdivided into three population categories: 16 areas with a
population of 250 000 to 500 000, 14 with 500 000 to one million, and

Table 10. Annual average NO_X concentration (µg/m³) in China from 1981 to 1990^a

Year Cities all over China Southern cities Northern cities

Concentration Annual Concentration Annual Concentration Annual
range average range average range average

1981 10-90 50 10-80 40 20-90 60
1982 10-110 45 10-90 40 30-110 50
1983 9-94 46 9-79 36 29-94 55
1984 10-95 42 13-75 37 10-95 46
1985 13-49 50 13-84 41 22-49 59
1986 14-108 48 14-98 41 18-108 55
1987 17-199 56 17-60 43 30-199 69
1988 9-110 45 9-110 42 8-120 48
1989 10-140 47 10-133 43 12-140 51
1990 7-130 43 12-71 38 7-130 47

^a General Environmental Monitoring Station of China (1991)

Table 11. Statistical data for the percentiles of ambient annual average NO_x concentrations (µg/m³) for Chinese cities (1986-1990)^a

Year Number Minimum Percentile Maximum Arithmetic Geometric
of cities value value
5 10 25 50 75 90 95 Average Standard Average Standard
deviation deviation

1986 71 14 17 20 30 43 60 81 88 108 48 22 43 488

1987 71 13 16 21 33 46 60 74 80 105 48 20 44 478

1988 73 8 11 18 30 43 58 67 84 120 45 22 40 547

1989 63 10 14 19 30 44 58 64 87 140 47 26 41 546

1990 59 7 13 17 27 38 51 71 86 130 43 23 37 554

^a General Environmental Monitoring Station of China (1991)

30 with over one million. No group exhibited a time trend, but the
areas with more than one million people clearly reported levels higher
than the smaller metropolitan areas. For 103 Metropolitan Statistical
Areas (MSA) reporting a valid year's data for at least one station in
1988 and/or 1989, peak annual averages ranged from 0.007 to 0.061 ppm
(Fig. 9). The only recently measured concentrations exceeding the USA
annual average standard (0.053 ppm) have occurred at stations in
southern California.

The seasonal patterns at stations in California are usually quite
marked and reach their highest levels through the autumn and winter
months. Stations elsewhere in the USA usually have less prominent
seasonal patterns and may peak in the winter or summer, or may contain
little discernable variation (Fig. 10) (US EPA, 1991).

One-hour NO₂ values at stations in the USA can exceed 0.2 ppm,
but in 1988 only 16 stations (12 of which are in California) reported
an apparently credible second high 1-h value above 0.2 ppm (Fig. 11).
Because at least 98% of 1-h values at most stations are below 0.1 ppm,
these values above 0.2 ppm are quite rare excursions whose validity
should be verified (US EPA, 1991).

3.4 Occurrence of nitrogen oxides indoors

This section summarizes emissions of NO_x from sources that
affect indoor air quality and are commonly found in residential
environments. There are several reasons for considering these
emissions. Firstly, examining emissions from several types of sources
and source categories can help identify the relative impact of each
source on indoor air quality and thus its influence on human exposure.
Secondly, such information is needed to understand the fundamental
physical and chemical processes influencing emissions. This
understanding can be used to help develop strategies for reducing
emissions. Finally, studying emissions from indoor sources can
provide source strength input data needed for indoor air quality
modelling. Knowledge of indoor concentrations is an important
component in estimating the total exposure of individuals to nitrogen
oxides.

An important factor for indoor air quality is how (or if) the
combustion products from appliances are vented outside the building.
It should be noted that several common types of vented appliances
usually emit NO_x to the outdoors; examples include gas-fired
furnaces, water heaters and clothes dryers, as well as stoves and
furnaces using wood, coal and other fuels. Under some circumstances
even these vented emissions may filter back inside and contribute to
elevated NO_x levels indoors. For example, Hollowell et al. (1977)
reported high NO and NO₂ concentrations in a house where a vented
forced-air gas-fired heating system was used. Elevated concentrations
may also be a problem with malfunctioning vented appliances. Other

data (e.g., Fortmann et al., 1984), however, suggest that fugitive
emissions of NO_x from vented appliances are small. The importance of
unvented appliances to indoor NO_x levels is well documented; this
section focuses on emissions from such appliances.

3.4.1 Indoor sources

3.4.1.1 Gas-fuelled cooking stoves

Several research programmes have investigated NO_x emissions
from stoves fuelled with natural and liquid petroleum gas (Himmel &
DeWerth, 1974; Cote et al., 1974; Massachusetts Institute of
Technology, 1976; Yamanaka et al., 1979; Traynor et al., 1982b; Cole
et al., 1983; Caceres et al., 1983; Fortmann et al., 1984;
Moschandreas et al., 1985; Cole & Zawacki, 1985; Tikalsky et al.,
1987; Borrazzo et al., 1987a). Most of these studies have included
investigations of several other pollutants, including CO, aldehydes
and unburned hydrocarbons. Table 12 lists average emission factors
for range-top burners and for oven and broiler burners operated at
maximum heat input rate. Data are shown for both well-adjusted blue
flames and for poorly adjusted yellow flames. Each of the averages is
based on the total number of stoves tested for that category, using
data from the above studies. For top burners with blue flames, a
total of 27 values are represented; for yellow flames, there are 23
values (24 for NO_x). Averages for the oven and broiler burners
represent 20 blue flame and 16 yellow flame values. Values are
generally very similar for emissions from these two types of burners
on the same stove. Overall, the results show that well-adjusted blue
flames emit more NO but less NO₂ than poorly adjusted yellow flames.
Emission factors from range-top burners are comparable to those from
oven and broiler burners.

Table 12. Average emission factors for nitric oxide (NO),
nitrogen dioxide (NO₂) and nitrogen oxides (NO_x)
from burners on gas stoves

Flame Factor for Factor for Factor for
type NO (µg/kJ) NO₂ (µg/kJ) NO_x (µg/kJ)

Top burners blue 20.0 ± 4.5 10.2 ± 3.1 41.0 ± 8.2
Top burners yellow 16.9 ± 4.5 15.0 ± 4.8 42.0 ± 9.1
Ovens and broilers blue 21.9 ± 6.3 7.23 ± 3.01 40.9 ± 8.6
Ovens and broilers yellow 19.8 ± 9.6 11.4 ± 5.7 39.0 ± 10.8

3.4.1.2 Unvented gas space heaters and water heaters

The findings of several investigators (Thrasher & DeWerth, 1979;
Traynor et al., 1983a, 1984b; Zawacki et al., 1986) are summarized in
Table 13. The most significant result is the markedly lower emissions
from heaters equipped with catalytic burners, radiant ceramic tile
burners and improved-design steel burners (radiant and Bunsen),
compared to emissions from simpler convection designs using
conventional cast-iron Bunsen burners. Equipping convective heaters
with radiant tiles does not make much difference to emission levels,
nor does the choice of natural gas or liquid petroleum gas fuel.
Other studies by Billick et al. (1984), Zawacki et al. (1984) and
Moschandreas et al. (1985) produced similar results.

3.4.1.3 Kerosene space heaters

The data presented in Table 14 show that emission factors of NO
and NO₂ for radiant kerosene heaters are generally much smaller than
those for convective kerosene heaters. Emissions of NO from two-stage
heaters are only slightly greater than those from radiant heaters,
whereas emissions of NO₂ are the lowest of the three heater types.
Most of the emissions from radiant heaters are in the form of NO₂;
for convective heaters that are two-stage heaters, the emissions of NO
and NO₂ are of comparable magnitude. There are insufficient data
to evaluate changes in emissions as kerosene heaters age. Other
products, including particles, present in these emissions may also be
of concern for their possible health effects.

3.4.1.4 Wood stoves

A number of studies have examined pollutant emissions from wood
stoves. Some of these studies have developed emission factors based
on concentrations in the flue gases; such information would be useful
for assessing the contribution of wood stove emissions to ambient air
quality. Very little information is available, however, on fugitive
emissions from wood stoves into the indoor living space.

In a detailed literature survey, Smith (1987) reported that
emissions of pollutants from wood stoves are highly variable,
depending on the type of wood used, stove design, the way the stove is
used and other factors. He reported emission factors for NO_x and
other pollutants for wood stoves used in developing countries. Many
of these stoves are unvented, which results in excessive indoor
concentrations as the combustion products are exhausted into the room.
The major health concerns for wood fires without chimneys arise from
pollutants other than NO₂, such as particulate matter.

Table 13. Summary of studies with unvented convective (C) and infrared (I) space heaters

Type of Number Heat input NO emission NO₂ emission NO_x emission Reference
heater (kJ/min) (µg/kJ) (µg/kJ) (µg/kJ)

Convective 5 86-661 24-47 2.2-7.3 39-77 Thrasher & DeWerth (1979)

Convective 8 188-830 9.5-22 9.5-20 34-47 Traynor et al. (1983a)

Infrared 5 245-352 0.1-1 4.1-6.2 4.9-6.2 Traynor et al. (1984b)
Convective 4 335-626 17.8-28.7 10-18.3 40.1-57.5

Infrared 5 264-334 0.005-1.7 1.6-4.8 2.7-5.7 Zawacki et al. (1986)
Convective 5 176-703 5.3-44.4 7.6-23.3 27.1-76.4

Table 14. Average emission factors for nitric oxide (NO), nitrogen dioxide (NO₂) and nitrogen oxides (NO_x) from kerosene heaters

Type of heater Heat input rate Emission factor Emission factor Emission factor Reference
(kJ/min) for NO (µg/kJ) for NO₂ (µg/kJ) for NO_x (µg/kJ)

Radiant, new 144 0.45 ± 0.05 4.4 ± 0.2 5.1 ± 0.2 Leaderer (1982)
Radiant, new 113 0.08 ± 0.05 5.0 ± 0.2 5.1 ± 0.2
Radiant, new 84.4 0 5.9 ± 0.3 5.9 ± 0.3

Convective, new 158 17 ± 0.3 7.0 ± 0.4 33 ± 0.6
Convective, new 97.9 12 ± 0.6 15 ± 0.3 33 ± 1.0
Convective, new 37.3 11 ± 0.9 17 ± 1.0 34 ± 1.7

Radiant, new 137 1.3 ± 0.7 4.6 ± 0.8 6.6 ± 1.3 Traynor et al. (1983b)

Radiant, 1 year old 111 2.1 5.1 8.3

Convective, new 131 25 ± 0.7 13 ± 0.8 51 ± 1.3

Convective, 5 years old 94.8 11 ± 0.1 32 ± 2.8 49 ± 2.8

Radiant 110-200 - - 13 ± 1.8 Yamanaka et al. (1979)

Convective 110-200 - - 70 ± 6.8

Traynor et al. (1984a) have studied wood stoves (three airtight
and one non-airtight) used in a house. For each experiment, airborne
concentrations of several pollutants were measured inside and outside
the house during operation of one of the stoves. The results showed
that all indoor and outdoor concentrations of NO and NO₂ were
below 0.02 ppm. Moreover, indoor air concentrations of some other
pollutants were high during use of the non-airtight stove. The
airtight stoves had little influence on indoor concentrations of any
pollutants. In another study, Traynor et al. (1982a) found elevated
airborne concentrations of NO and NO₂ in three occupied houses during
operation of wood stoves and a wood furnace. The concentrations were
highly variable.

Because of the limited data, it is difficult to reach
quantitative conclusions regarding the importance of wood stoves.
However, the limited information available suggests that wood stoves
are not a major contributor to indoor nitrogen oxide exposures. This
is consistent with the small NO emission rates expected from the low
temperature combustion processes characteristic of wood stoves.

3.4.1.5 Tobacco products

A number of studies have compared concentrations of NO_x and
other pollutants in houses with smokers and houses without smokers.
In general, these studies have shown that concentrations are somewhat
greater in the homes of smokers.

A few studies have reported emissions of NO_x from cigarettes
while sampling both sidestream and mainstream smoke together.
Woods (1983) reported 0.079 mg NO_x/cigarette, while Moschandreas
et al. (1985) listed emissions of 2.78 mg/cigarette for NO and
0.73 mg/cigarette for NO₂. The National Research Council (1986)
reported total NO_x emissions of 100 to 600 µg/cigarette for
mainstream smoke, with values 4 to 10 times greater for sidestream
smoke. According to the report, virtually all of the emitted NO_x is
in the form of NO; once emitted, the NO is gradually oxidized to NO₂.
Thus environments containing cigarette smoke may have higher
concentrations of both NO and NO₂ than environments without such
smoke. The NO₂ concentration on trains travelling between Changchun
and Harbin, China, was found to be related to the amount of cigarette
smoking, which was greater on daytime trains than on night-time ones.
On a one-way daytime train the average NO₂ concentration was 54 ppb
(range, 37-84 ppb), whereas on a two-way night-time train it was
40.6 ppb (range, 30-59 ppb) (Du et al., 1992).

3.4.2 Removal of nitrogen oxides from indoor environments

A number of field studies of NO₂ levels in residences have
reported that NO₂ is removed more rapidly than can be accounted for
by infiltration alone (Wade et al., 1975; Macriss & Elkins, 1977;

Oezkaynak et al., 1982; Traynor et al., 1982a; Ryan et al.,
1983; Leaderer et al., 1986). Indoors, NO₂ is removed by
infiltration/ventilation and by interior surfaces and furnishings.
The removal of NO₂ by interior surfaces and furnishings and reactions
occurring in air is often referred to as the reactive decay rate of
NO₂, and it can be a significant factor in the actual NO₂ levels
measured in residences. Failure to account for the reactive decay
rate can lead to a serious underestimation of emission rate
measurements in chamber and test house studies and a serious
overestimation of indoor concentrations when using emission rates to
model indoor levels. The NO₂ reactive decay rate is typically
determined by subtracting the decay of NO₂, after a source is shut
off, from that of a relatively non-reactive gas (e.g., CO, CO₂, SF₆,
NO), which can be related to ventilation rates, expressed in room
air changes per hour. The measured reactive decay rates in the
above-mentioned field studies ranged from 0.1 to 1.6 air change
times/hour. All studies noted that the reactive decay of NO₂ is as
important and in some cases more important than infiltration in
removing NO₂ indoors. Leaderer et al. (1986) monitored NO₂, NO, CO
and CO₂ continuously in seven houses over periods ranging from 2 to
8 days. They reported that the NO₂ decay rate was always greater
than that due to infiltration alone and was highly variable among
houses and among time periods within a house.

In an effort to identify the factors that control the NO₂
reactive decay rate, a number of small chamber (Miyazaki, 1984; Spicer
et al., 1986), large chamber (Moschandreas et al., 1985; Leaderer et
al., 1986) and test house studies (Yamanaka, 1984; Borrazzo et al.,
1987b; Fortmann et al., 1987) have been conducted. The most extensive
small chamber work was reported by Spicer et al. (1986), where 35
residential materials were screened for NO₂ reactivity in a 1.64-m³
chamber and a limited number of the materials were tested for the
impact of relative humidity on the reactivity rate. Fig. 12 shows the
relative rates of NO₂ removal for the materials screened. The figure
indicates that many of the materials used for building construction
and furnishings are significant sinks for NO₂ and that their removal
rate is highly variable. Many of the materials were found to reduce a
significant proportion of the removed NO₂ to NO. In no cases was
NO₂ re-emitted, although some materials emitted NO. The authors
noted that the materials that removed NO₂ most rapidly fall in two
categories: (1) porous mineral materials of high surface area; and (2)
cellulosic material derived from plant matter. Higher relative
humidities were found to enhance the removal rate for some materials
(e.g., wool carpet), reduce the removal rate for some (e.g., cement
block), and have little effect on others (e.g., wallboard). In a
series of small (0.69 m³) chamber studies (Miyazaki, 1984) reactive
decay rates for NO₂ were found to vary as a function of material type
and to increase with increasing surface area of the material, degree
of stirring in the chamber, temperature and relative humidity. A
saturation effect was noted on some of the carpets tested.

In a series of large chamber studies (34-m³ chamber), Leaderer
et al. (1986) evaluated the reactive decay rate of NO₂ as a function
of material type, surface area of material, relative humidity and air
mixing. The reactive decay rate was found to vary as a function of
material surface roughness and surface area. Carpeting was found to
be most effective in removing NO₂, whereas painted wallboard was
least effective. Increases in relative humidity were associated with
increases in removal rates for all materials tested, but the slope was
a shallow one. Of particular interest is the finding in this study
that the degree of air mixing and turbulence was a dominant variable
in determining the reactive decay rate for NO₂. Moschandreas et al.
(1985) evaluated six materials in a 14.5-m³ chamber and found
variations in decay rates according to material types and a positive
impact of relative humidity on NO₂ decay rates in an empty chamber.

Yamanaka (1984), in assessing NO₂ reactive decay rates in a
Japanese living room, found the decay to consist of both homogeneous
and heterogeneous processes. The rates were found to vary as a
function of surface property and sharply as a function of relative
humidity. NO production during the decay was noted. In a test house
study, Fortmann et al. (1987) noted that the NO₂ decay rate tends to
decrease as the concentration increases. It is not clear whether this
is due to surface saturation or second-order kinetics. This study
also noted a sharp increase in NO levels during the NO₂ decay,
indicating NO production as a result of the NO₂ decay. In a test
house study conducted over a 7-month period, Borrazzo et al. (1987b)
found that reaction rates for NO₂ in the test house were sensitive to
the location in the house where they were measured. This indicates
that reaction losses during transport of NO₂ from room to room in a
house may be important.

Reactive decay of NO₂ associated with interior surface materials
and furnishings is an important mechanism for removing NO₂ from the
air within homes. Reactive decay rates for NO₂ vary as a function of
the type and surface area of the material. The impact of relative
humidity on the decay rate is unclear, with some studies showing a
pronounced impact (Yamanaka, 1984), while others show only moderate or
little impact (e.g., Spicer et al., 1986; Leaderer et al., 1986). The
degree of air mixing or turbulence can have an important effect on the
reactive decay rate. A by-product of NO₂ removal by materials may be
NO production, and a saturation effect may occur for some materials.
Reactive decay of NO₂ in residences is highly variable between
residences, within rooms in a residence, and on a temporal basis
within a residence. The large number of variables controlling the
reactive decay rate make it very difficult to assess in large field
studies through questionnaire or integrated air sampling.

3.5 Indoor concentrations of nitrogen oxides

Indoor concentrations of NO₂ are a function of outdoor
concentrations, indoor sources (source type, condition of source,
source use, etc.), infiltration/ventilation, air mixing within and
between rooms, reactive decay by interior surfaces, and air cleaning
or source venting.

3.5.1 Homes without indoor combustion sources

Typical studies in homes without indoor sources of NO₂,
summarized in Table 15, have reported concentrations lower than
outdoor levels due to removal from the air of NO_x by the building
envelope and interior surfaces. Thus indoor/outdoor concentration
ratios are consistently less than unity. These homes provide some
degree of protection from outdoor concentrations. Indoor/outdoor
ratios vary considerably according to the season of the year, the
lowest ratios occurring in the winter and highest occurring during the
summer. Although urban concentrations are often highest in winter,
this pattern in the indoor/outdoor ratio, attributed to seasonal
differences in infiltration rates, NO₂ reactivity rates, the
penetration factor and outdoor concentrations, can result in higher
indoor concentrations in summer than in winter. The indoor-to-outdoor
ratio for these homes does not appear to depend on geographical area,
housing type or outdoor concentration. Results of monitoring in
Portage, Wisconsin, USA, show that the presence of a gas stove
contributes dramatically to the indoor NO₂ levels. Table 16, taken
from the report of Quackenboss et al. (1986) and based on data
collected in 1981 and 1982, clearly shows that gas stoves increase not
only indoor concentrations but also the personal exposure of children.

3.5.2 Homes with combustion appliances

It is estimated that gas (natural gas and liquid propane) is used
for cooking, heating water or drying clothes in about 45% of all homes
in the USA (US Bureau of the Census, 1982) and in nearly 100% of homes
in some other countries (e.g., the Netherlands). Gas appliances
(gas cooker/oven, water heater, etc.) are the major indoor source
category for indoor residential NO₂ by virtue of the number of homes
with such sources. NO₂ concentrations in homes with gas appliances
are higher than those without such appliances. Within this category,
the gas cooker/oven and unvented heaters are by far the major
contributors. Cookers and ovens are especially important sources when
used inappropriately as a supplementary room heater. Average indoor
concentrations (based on a 1- to 2-week measurement period) in excess
of 100 µg/m³ have been measured in some homes with gas cookers
(Table 17). Homes where gas cookers with pilot lights are used have

higher NO₂ levels than homes that have gas cookers without pilot
lights. Average NO₂ concentrations in homes with gas cookers/ovens
exhibit a spatial gradient within and between rooms. Kitchen
concentrations of NO₂ are higher than other rooms and a steep
vertical concentration gradient in the kitchen has been observed in
some homes, concentrations being highest nearest the ceiling. Average
NO₂ concentrations are highest during the winter months and lowest
during the summer months. This seasonal temporal gradient is
attributed to differences in infiltration, appliance use, NO₂
reactivity rates and indoors and outdoor concentrations. The impact
of gas appliance use on indoor NO₂ levels may be superimposed upon
the background level resulting from outdoor concentrations. Only very
limited data exist on short-term average (3 h or less) indoor
concentrations of NO₂ associated with gas appliance use. These data
suggest that short-term average concentrations of NO₂ are several
times the longer-term average concentrations measured.

A wide variety of fuel types can be used for cooking and heating
in different localities. These can produce various effects on indoor
air quality. As an example, Table 18 gives data for indoor NO_x
concentrations measured at Lanzhou City, China, where coal and
liquified gas were used in apartments and houses (Duan et al., 1992).

Table 15. Average outdoor concentrations of nitrogen dioxide (NO₂) and average indoor/outdoor ratios in homes without gas appliances or
unvented space heaters^a

Location Housing Averaging Seasons Number Average NO₂ Indoor/outdoor ratios Reference
type^b time of outdoor
homes concentration
(µg/m³) Kitchen Bedroom

Southern California Mixed 7 days Summer 70 71.9 0.80 0.75 Southern California
Spring 100 43.5 0.72 0.60 Gas Company (1986)
Winter 69 91.2 0.56 0.47

New Haven, CT Single family 14 days Winter 60 13.2 0.56 0.55 Leaderer et al. (1986)
unattached

Albuquerque, NM Mixed 14 days Winter 1 60 14.1 - 0.50 Marbury et al. (1988)
Winter 2 56 19.6 - 0.32

California Mobile homes 7 days Summer 46 25.9 0.61 0.54 Petreas et al. (1988)
Winter 23 44.6 0.27 0.26

Portage, WI Mixed 7 days Summer 47 15.2 0.91 0.72 Quackenboss et al. (1986)
Winter 47 17.2 0.65 0.45

Tucson, AZ Mixed 14 days Summer 56 19.9 0.86 0.76 Quackenboss et al. (1986)
Spring/Autumn 41 25.6 0.71 0.55
Winter 23 36.8 0.64 0.52

Boston, MA Mixed 14 days Summer 117 31.7 0.76 0.75 Ryan et al. (1988)
Autumn 117 37.8 0.43 0.40
Winter/Spring 124 33.5 0.53 0.47

Table 15. (Con't)

Location Housing Averaging Seasons Number Average NO₂ Indoor/outdoor ratios Reference
type^b time of outdoor
homes concentration
(µg/m³) Kitchen Bedroom

Northern Central Single family 5 days Winter 9 53.8 Koontz et al. (1986)
Texas unattached

Suffolk County, Single family 7 days Winter 49 35.5 0.47 - Research Triangle
NY unattached Institute (1990)

Onondago County, Single family 7 days Winter 66 21.7 0.70 -
NY unattached

Portage, WI Single family 7 days Average over 25 12.8 0.65 0.51 Spengler et al. (1983)
unattached all seasons

Watertown, MA Not given 3-4 days November 18 37.0 0.65 0.51 Clausing et al. (1984)
December 10 46.0 0.39 0.30

Middlesbrough, UK Not given 7 days Winter 87 35.0 0.97 0.75 Goldstein et al. (1979)

Middlesbrough, UK Not given 7 days Winter 15 34.7 - 0.75 Melia et al. (1982a,b)

^a Data from field studies of private residences in the USA and United Kingdom
^b "Mixed" indicates a single family in an attached or unattached dwelling, condominium or apartment
Table 16. Nitrogen dioxide concentrations (ppm) according to season and
stove type in Portage, Wisconsin, USA^a

Season Stove Indoor Outdoor Personal

Mean SD Mean SD Mean SD

Summer Gas 0.016 0.006 0.006 0.003 0.014 0.004
Electric 0.007 0.003 0.008 0.003 0.009 0.003

Winter Gas 0.027 0.013 0.008 0.003 0.023 0.009
Electric 0.005 0.003 0.009 0.003 0.008 0.003

^a From: Quackenboss et al. (1986); SD = standard deviation
Table 17. Indoor and outdoor concentrations of nitrogen dioxide (NO₂) in homes with gas appliances, and the calculated average
contribution of those appliances to indoor residential NO₂ levels

Location Housing Averaging Type of Season No. of Average measured NO₂ Indoor NO₂ due to source Reference
type^a time appliance homes (µg/m³) (µg/m³)
(days)
Outdoors Kitchen Bedroom Other Kitchen Bedroom Other Comment^b

USA

Southern Mixed 7 Oven/range, Summer 147 75.3 91.6 68.4 - 31 12 - 1,2 Southern
California ± pilot Spring 202 49.2 79.2 51.3 - 35 22 - 1,2 California
Winter 141 104 101.5 69 - 48 20 - 1,2 Gas Company
(1986)
Oven/range, Winter 98 107 113 76 - 53 26 - 1,2
pilot

Oven/range, Winter 38 97 74 53 - 20 7 - 1,2
no pilot

Water heater Winter 21 92 59 50 - 11 11 - 1,2,3
in home

Wall furnace Winter 90 121 161 113 - 49 36 - 1,4

Floor Summer 42 119 177 126 - 66 52 - 1,4
furnace

Table 17. (Con't)

New Haven, Single 14 Oven/range, Winter 42 14.8 44.7 27.6 30.4 36 20 22 1,5 Leaderer
CT family, ± pilot et al.
unattached (1986)

Albuquerque, Mixed 14 Oven/range, Winter 82 19.1 - 33.1 41.9 - 24 31 1,5,6 Marbury et
NM ± pilot Winter 75 20.3 - 30.9 39.3 - 24 32 al. (1988)

California Mobile 7 Oven/range, Summer 265 21.1 43.1 30.2 - 30 19 - Petreas et
homes ± pilot Winter 231 42.1 53.7 37.5 - 42 27 - 1,7 al. (1988)

Portage, Mixed 7 Oven/range, Summer 36 11.5 38.9 21.1 29.6 29 13 20 Quackenboss
WI ± pilot Winter 34 15.4 69.6 31.2 50.7 60 15 42 1,8 et al.
(1986)

Tucson, Mixed 14 Oven/range, Summer 13 23.1 39.1 26.3 30.7 19 8 11 Quackenboss
AZ ± pilot Spring/ 11 36.3 45.8 31.9 42.4 20 12 17 et al.
Autumn (1986)
Winter 10 45.2 60.6 43.4 50.7 32 20 25 1,9

Boston, Mixed 14 Oven/range, Summer 301 41.6 65.9 45.6 50.9 33 15 19 Ryan et al.
MA ± pilot Autumn 277 43.7 74.3 47.5 52.8 56 30 34 (1988)
Winter/ 298 40.5 73.5 48.6 55.1 52 30 34 1,9
Spring

Table 17. (Con't)

Central Single 5 Oven/range, Winter 22 34.6 - - 54.1 - - 37 1,10 Koontz et
Texas family, ± pilot al. (1986)
unattached

Suffolk Co., Single 7 Oven/range, Winter 42 37.6 77.5 - 52.4 60 - 37 Research
NY family, ± pilot Triangle
unattached Institute
(1990)

Onondago Single 7 Oven/range, Winter 56 30.6 62.6 0 50 41 - 27 1,9
Co., NY family, ± pilot
unattached

New York, Apartments 2 Oven/range Summer 14 109 122 98 106 30 6 13 Goldstein
NY Autumn 1 15 61 96 65 71 53 22 18 et al.
Autumn 2 9 73 108 66 76 45 15 25 (1985)
± pilot Winter 1 8 100 121 76 95 61 16 35
Winter 2 18 75 126 63 82 81 18 37 9,11,12
Spring 13 95 121 82 99 55 16 33

Portage, WI Single 7 Natural gas All 36 15.8 65.5 36.7 - 55 29 - Spengler et
family, Oven/range, seasons al. (1983)
unattached no pilot

Liquified All 76 11.6 65.6 37.6 - 58 31 - 1,13
petroleum seasons
gas
Oven/range,
no pilot

Table 17. (Con't)

Watertown, Not given 3-4 Gas cooking Novemb. 60 37 74 45 51 50 26 33 1,9,14 Clausing et
MA Decemb. 51 46 86 46 60 68 32 44 al. (1984)

Netherlands

Arnet Not given 7 Gas cooking Autumn/ 294 35 118 - 97 97 - 37 Noy et al.
Enschede no pilot Winter (1984)
Water
heaters

Ede Not given 7 Gas cooking Autumn/ 173 44 113 43 54 89 17 28 Noy et al.
no pilot Winter (1984)
Water
heaters

Vlagttwedde Rural area 7 Gas cooking Autumn/ 162 28 107 24 51 90 7 34
no pilot Winter Water
heaters

Rotterdam I, Inner city 7 Gas cooking Autumn/ 228 45 144 51 80 117 24 53
no pilot Winter
Water
heaters

Rotterdam II, Inner city 7 Gas cooking Autumn/ 102 45 143 64 73 117 37 46 9,17
no pilot Winter Water
heaters

Table 17. (Con't)

United Kingdom

Middlesbrough Not given 7 Gas cooking Winter 428 35 213 58 - 179 24 - 1,15 Goldstein
no pilot et al.
(1979)

Middlesbrough Not given 7 Gas cooking Winter 183 34.7 - 60 82.7 - 39 61 1,16 Melia et
al.
(1982a,b)

^a "Mixed" indicates a single family in an attached or unattached dwelling, condominium or apartment
^b 1. Background correction determined by multiplying: (a) the indoor/outdoor ratio for homes in the study with no indoor NO₂ sources
for a given season; by (b) the outdoor NO₂ concentration measured for the home with sources; and subtracting the product from
the indoor level measured in the house.
2. Homes containing forced air gas furnace. These homes are thought not to contribute significantly to indoor levels for this
sample.
3. Homes with electric cooker/oven, forced air gas furnace, and gas water heater in home. Comparison is made with electric
cooker/oven, forced air gas furnace, and gas water heater located outside home.
4. Homes have gas cooker/oven with source contribution calculated after correction of a gas cooker/oven. Values are background
corrected with gas stove.
5. Living room or activity room.
6. Sampling was done over two different periods for the same houses within the same winter period.
7. Outdoor values were obtained from five locations, housing type, mobile home.
8. Other location in home; bedroom refers to average of levels in one or more bedrooms in house.
9. Other location in the main living room.
10. Other location is point nearest centre of home.

Table 17 (Con't)

11. 48-h samples over 30 consecutive days.
12. Indoor/outdoor (I/O) ratio is assessed to be 0.6, 0.7, and 0.85 for the Winter, Spring/Autumn and Summer periods,
respectively, for all locations, because no control home (no gas appliances) mean measurements were available. Using these
I/O ratios, the impact of sources was calculated as footnote 1.
13. Each home was sampled six times over a 1-year period.
14. Outdoor levels are average for homes with or without gas appliances.
15. Outdoor levels were recorded at 75 locations in the general sampling area and were not home-specific. Bedroom levels were
obtained for 107 of the 428 homes.
16. Outdoor levels were recorded at 82 locations in the general sampling areas and were not home-specific. Outdoor levels were
recorded at the beginning and end of the study.
17. Indoor/outdoor (I/O) ratio is assumed to be 0.6 for all locations, because no control home (no gas appliances) measurements
were available. Using I/O ratio of 0.6, the impact of sources was calculated as in footnote 1.
Table 18. Indoor concentration of NO_x in Lanzhou city, China^a

Type of residence Average NO_x
concentration (mg/m³)

Winter Summer

Apartment building with central 0.141 0.059
heating, liquified gas for cooking

Apartment building without central 0.136 0.059
heating, coal for cooking and heating

One-storey house, coal for cooking 0.106 0.046
and heating

^a From: Duan et al. (1992)

3.5.3 Homes with combustion space heaters

Unvented kerosene and gas space heaters are important sources of
NO and NO₂ in homes, both because of the NO and NO₂ production rates
of the heaters and the long periods of time that they are in use. The
concentrations of NO emitted are usually several times higher than
those of NO₂. However, in the literature, indoor air concentrations
of NO are frequently not reported.

Field studies indicate that average residential concentrations
(1- or 2-week average levels) exhibit a wide variation, depending
primarily on the amount of heater use and the type of heater
(Table 19). Under similar operating conditions, unvented gas space
heaters appear to be associated with higher indoor NO₂ concentrations
than kerosene heaters. Average concentrations in homes using unvented
kerosene heaters have been found to be well in excess of 100 µg/m³.
In one study (Leaderer et al., 1986), calculations of NO₂
concentrations in residences during kerosene heater use (in homes
without gas appliances) indicate that approximately 50% of the homes
have NO₂ concentrations above 100 µg/m³ and 8% above 480 µg/m³. A
peak NO₂ concentration of 847 µg/m³ was measured over a 1-h period
in a home during use of a kerosene heater.

Table 19. Two-week average nitrogen dioxide (NO₂) levels for homes
in New Haven, Connecticut, USA, during winter, 1983^a

Source category; NO₂ (µg/m³)
location

n Mean SD^b % above
100 µg/m³

No kerosene heater
or gas stove
Outdoors 144 13.2 5.3 0
House average 145 7.4 4.2 0
Kitchen 147 7.6 3.7 0
Living room 146 7.3 3.4 0
Bedroom 145 7.3 8.6 0

One kerosene heater,
no gas stove
Outdoors 95 12.9 4.6 0
House average 95 36.8 32.8 2.1
Kitchen 96 39.1 35.5 4.2
Living room 96 38.5 36.6 5.2
Bedroom 95 31.9 30.8 5.3

No kerosene heater,
one gas stove
Outdoors 42 14.8 4.2 0
House average 42 34.3 26.2 4.8
Kitchen 42 44.7 31.4 4.8
Living room 42 30.4 24.8 4.8
Bedroom 42 27.8 25.1 4.8

One kerosene heater,
one gas stove
Outdoors 18 14.5 5.2 0
House average 18 66.8 43.9 16.7
Kitchen 18 74.5 52.0 22.2
Living room 18 57.4 38.6 11.1
Bedroom 18 68.5 56.5 16.7

Two kerosene heaters,
no gas stove
Outdoors 13 16.5 9.4 0
House average 13 69.5 38.0 23.0
Kitchen 13 73.0 31.7 23.0
Living room 13 73.6 44.3 38.5
Bedroom 13 67.8 44.9 23.1

Table 19. (Con't)

Source category; NO₂ (µg/m³)
location

n Mean SD^b % above
100 µg/m³

Two kerosene heaters,
one gas stove
Outdoors 3 22.1 6.2 0
House average 3 85.8 24.4 33.3
Kitchen 3 94.0 22.7 66.6
Living room 3 77.6 38.4 33.3
Bedroom 3 85.8 19.5 33.3

^a From: Leaderer et al. (1986); repeat monitoring data (n = 19)
are included
^b SD = standard deviation

A large field study (Koontz et al., 1986) of indoor NO₂
concentrations in Texas homes using unvented gas space heaters (most
also had gas cookers) found that approximately 70% of the homes had
average concentrations in excess of 100 µg/m³ and 20% had average
concentrations in excess of 480 µg/m³. This study found that the
indoor/outdoor temperature difference was the best indicator of
average indoor NO₂ levels during the colder winter periods when
heating demands are greatest.

Only limited data have so far been published on short-term peak
indoor concentrations for homes with unvented gas space heaters, and
no data are available on spatial variations or concentrations solely
during the hours of heater operation.

No spatial gradient of NO₂ was found in homes with unvented
kerosene space heaters, contrary to the strong spatial gradient noted
for homes with gas appliances. This is probably due to the strong
convective heat output and the long operating hours of the heaters,
which result in rapid mixing within the homes.

Ferrari et al. (1988) conducted a study of air quality in
homes with unvented space heaters in Sydney, Australia, over
two winters. NO₂ concentrations were measured by both continuous
(chemiluminescence with O₃ method) and passive monitors (badges and

Palmes tubes). Concentrations of NO₂ exceeded 0.10 ppm (average
concentration) in 85% of homes tested, and 0.16 ppm in 44% of homes.
More than 10% of homes had average NO₂ concentrations exceeding
0.32 ppm, and the maximum recorded was greater than 0.5 ppm. Unvented
gas space heaters are common in Sydney, and average use is about 3 h
per night during the winter. As a result, an estimated 0.5 million
residents are exposed to NO₂ concentrations exceeding 0.16 ppm for
several hours per night during the colder months of the year.

Improper use of gas appliances (e.g., using a gas oven or
stove to heat a living space) and improperly operating gas appliances
or vented heating systems (e.g., out-of-repair gas cooker or
improper operation of a gas wall or floor furnace) can be important
contributors to indoor NO₂ concentrations, but few data are available
to assess the magnitude of that contribution. Little or no field data
exist that would allow for an assessment of the contributions of wood-
or coal-burning stoves or fireplaces to indoor NO₂ concentrations,
but such a contribution would be expected to be small. Cigarette
smoking is expected to add relatively small amounts of NO₂ to homes
(see also Tables 15 and 18).

In developing countries, biomass fuels (e.g., wood, biogas,
animal dung, etc.) are much more widely used for home heating and/or
cooking than in developed countries, these fuels often being burnt in
open hearth fires or poorly vented appliances within indoor spaces of
residential dwellings (WHO, 1992). As noted by Sims & Kjellström
(1991), a very conservatively estimated 400 million people are
affected by biomass smoke problems worldwide, mostly in rural areas of
developing countries. A disproportionate number of women and young
children are exposed, owing to the greater numbers of hours typically
spent by them indoors and their involvement in cooking and other
household chores. Increased NO_x concentrations, as well as greater
concentrations of carbon monoxide, suspended particulate matter (SPM)
and volatile organic compounds (VOCs) are normally found in biomass
smoke (Chen et al., 1990). Reviews of field studies in rural areas of
developing countries indicate that exposure levels to biomass smoke
components are usually rather high. Indoor concentrations for NO₂,
for example, were found to fall within the range of 0.1 to 0.3 mg/m³
in India, Nepal, Nigeria, Kenya, Guatemala and Papua New Guinea, as
reported in reviews by WHO (1984) and Smith (1986, 1987). Similarly,
Hong (1991) reported NO₂ concentrations in the range of 0.01 to
0.22 mg/m³ resulting from indoor combustion of biogas in homes in
Chengdu, Szechuan Province, China. Hong (1991) also reported NO_x
concentrations in the range of 0.02 to 0.16 mg/m³ in other houses in
Gansu Province, China, where dried cow dung was used as a fuel. The
above NO₂ indoor air concentrations from biomass smoke should be
compared with the WHO Air Quality Guidelines recommendation of
0.15 mg/m³ for daily exposures to NO₂ (WHO, 1987).

3.5.4 Indoor nitrous acid concentrations

Recent studies have demonstrated that substantial concentrations
of HNO₂ can be present inside residential buildings, especially when
unvented combustion sources are used. HNO₂ is formed by the reaction
of NO₂ with water on surfaces and the reaction is promoted by high
humidity. HNO₂ may also be produced by other mechanisms, and this is
the subject of active research. Brauer et al. (1993) found that HNO₂
can represent over 10% of the concentrations usually reported as NO₂.

3.5.5 Predictive models for indoor NO₂ concentration

Efforts to model indoor NO₂ levels have employed two distinct
approaches: physical/chemical and empirical/statistical models.

The physical/chemical modelling approach has been used by
numerous investigators in chamber, test house and small field studies
(involving a small number of homes) to estimate emission rates of NO₂
from combustion sources (e.g., Traynor et al., 1982a; Leaderer, 1982;
Moschandreas et al., 1984), to estimate reactive decay rates (e.g.,
Oezkaynak et al., 1982; Yamanaka, 1984; Leaderer et al., 1986; Spicer
et al., 1986; Borrazzo et al., 1987a), to estimate the impact of
ventilation and mixing on the spatial and temporal distribution of
NO₂ (e.g., Oezkaynak et al., 1982; Traynor et al., 1982b; Borrazzo
et al., 1987a), and to evaluate the applicability of emission
rates determined under controlled conditions in estimating indoor
concentrations of NO₂ (e.g., Traynor et al., 1982b; Borrazzo et al.,
1987a). More recently, studies have reported the use of distributions
of the input variables to the mass balance equation (emission rates,
source use, decay rates, ventilation rates, etc.), determined from the
published literature, to estimate distributions of indoor NO₂ levels
for specific sources and combinations of sources (Traynor et al.,
1987; Hemphill et al., 1987).

Prediction of indoor concentrations or concentration
distributions of NO₂ in homes with combustion sources using
physical/chemical (mass-balance) models requires accurate information
for input parameters (e.g., emission rates). Although data are
available for some of the input parameters under controlled
experimental conditions, there are very limited data available
concerning either the variability of such input parameters in actual
homes or the factors that control variability (e.g., variability of
emission or decay rates). Obtaining field measurements or estimates
of the inputs in large numbers of homes would be expensive and
time-consuming. Such modelling efforts do, however, help to identify
the potential range of indoor NO₂ concentrations, factors that may
result in high levels, and the potential effectiveness of mitigation
efforts.

Empirical/statistical models have been developed from large field
studies that have measured NO₂ concentrations in residences and
associated outdoor levels for time periods of a week or more. These
have typically used questionnaires to obtain information on sources in
the residences, source use, building characteristics (house volume,
number of rooms, etc.), building use, and meteorological conditions.
In some cases, additional measurements, including temperature have
been recorded. Several investigators have attempted to fit simple
regression models to their field study databases in an effort to
determine whether the variations in NO₂ levels seen among houses can
be explained by variations in questionnaire responses. The goal has
been to see how well questionnaire information or easily available
information (meteorological data) can predict indoor NO₂ levels. In
most cases a linear model has been used, but several investigators
have used log transformations of variables. These employ
questionnaire responses and measured physical data (house volume,
etc.) as independent variables and have met with moderate success.
Linear regression models, with the exception of the Petreas et al.
(1988) model, explain from 40 to 70% of the variations in residential
NO₂ levels and typically have large standard errors associated with
their estimates. Although log transformations of variables have
always produced a higher percentage of explained variation due to the
skewed distribution of the original variables, interpretation of the
coefficients in a nonlinear model can require special attention.

Regression models developed from field studies employing
questionnaires to explain variations in indoor levels of NO₂ have met
with only moderate success.

Better information, through additional measurements and better
questionnaire design, is needed on a range of factors if the
statistical/empirical models are to be used to estimate indoor
concentrations of NO₂ in homes without measurements. Factors include
source type and condition, source use, contaminant removal
(infiltration and reactive decay) and between and within room mixing.

3.6 Human exposure

To assess the health impact of exposure to nitrogen oxides, it is
essential to conduct an accurate exposure assessment. Such data are
of paramount importance for the definition of dose-effect and
dose-response relationships. In fact, the risk to human health is not
simply determined by indoor and outdoor concentrations of nitrogen
oxides, but rather by the personal exposure of every individual. The
integrated exposure is the sum of the individual exposures to oxides
of nitrogen over all possible time intervals for all settings or
environments. It requires, thus, the consideration of long-term

average concentration level, variations and short-term exposures, as
well as the activity patterns and personal and home characteristics of
individuals (Berglund et al., 1994).

Exposure is a function of concentration and time. People spend
various periods in different types of micro-environments with various
concentration levels. On average, people spend about 90% of their
time indoors (at home, work, school, etc.), about 5% in transit
(Chapin, 1974), and 7% (range 3-12%) near smokers (Quackenboss et al.,
1982). These values vary with the season, day of the week, age,
occupation and other factors but it is decidedly important to predict
indoor pollutant levels when total exposure is being estimated.

Adequate exposure assessment for NO₂ is particularly critical in
conducting and evaluating epidemiological studies. Failure to measure
or estimate exposures adequately and address the uncertainty in the
measured exposures can lead to misclassification errors (Shy et al.,
1978; Gladen & Rogan, 1979; Oezkaynak et al., 1986; Willett, 1989;
Dosemeci et al., 1990; Lebret, 1990). Early studies comparing the
incidence of respiratory illness in children living in homes with and
without gas stoves used a simple categorical variable of NO₂
exposure; the presence or absence of a gas cooker. Such a dichotomous
grouping can result in a severe non-differential misclassification
error in assigning exposure categories. This type of error is likely
to underestimate the true relationship and could possibly result in a
null finding.

In assessing human exposure to NO₂ (and other oxides of
nitrogen), averaging times chosen should account for the type of
effect to be expected. With regard to NO₂, the principal biological
responses include (a) relatively transient effects on respiratory
function associated with acute, short-term (< 1 h) exposures, and (b)
the likelihood of increased risk for respiratory disease in children
associated with frequently repeated short-term peak exposures and/or
lower level long-term exposures.

Indirect and direct methods for personal exposure assessment are
available. Indirect methods combine measures of concentrations at
fixed sites in various types of micro-environments with information
on where people have spent their time (time-activity patterns).
Time-weighted average (TWA) exposure models have been developed to
estimate total personal exposure (Fugas, 1975; Duan, 1982; Duan,
1991). The NO₂ exposure levels predicted from TWA exposure models
have been shown to correlate closely with the exposure levels obtained
by direct measurements of personal exposure (Nitta & Maeda, 1982;
Quackenboss et al., 1986; Sega & Fugas, 1991). However, the large
variation in NO₂ concentrations (distribution) within each type of

micro-environment (because of variability in, for example, stove use,
emission rates, ventilation frequencies, and the day-to-day and
person-to-person variations in the use of time) decreases the accuracy
of the predicted exposure and increases the risk for misclassification
of the exposure.

Direct measurements of concentrations in the breathing zone
of a person using personal passive exposure monitors provide
time-integrated measurements of exposure for a certain period across
the various micro-environments where a person spends time. It is
important to collect exposure data over time intervals consistent with
the expected effects. Effects from long-term, low-level exposure may
be different from effects from short periods of high concentration
(intermittent peak exposure). Intermittent peak exposure, which
occurs during cooking on a gas stove, may be significant to total
exposure and adverse health effects. If effects from peak exposure
are to be considered in the exposure assessment, the sampling time
must be short enough to detect these peak exposures. Such a short
sampling time is possible with the more sensitive passive samplers and
with conventional air monitors, such as chemiluminescence NO_x
monitors. However, direct methods of measuring personal exposure
are relatively costly and time-consuming. Within-person and
between-person variability, both in personal exposure and personal
use of time, can be large.

Hence a sufficient number of personal exposure measurements must
be collected for each person (repeated measurements), and a sufficient
number of individuals must be sampled before the measurements can be
considered to be representative. Personal daily exposures have been
shown to vary between individuals on the same day by a factor of up to
about 15 in the urban area of Stockholm and between days for the same
individual by a factor of up to 10 (Berglund et al., 1993).

A comparison of personal NO₂ exposures, as measured by Palmes
diffusion tubes, and NO₂ exposures measured in residences had a
correlation of 0.94 for a subsample of 23 individuals (Leaderer et
al., 1986). Results of this comparison are depicted in Fig. 13 and
show an excellent correlation between average household exposure and
measured personal exposure.

It is important to note that indoor concentrations are strong
predictors of personal exposure. In the case of homes with gas or
electric stoves, personal exposure has been shown to be closely
related to the household indoor average concentrations (Quackenboss
et al., 1986; Harlos et al., 1987a).

Results of monitoring in Portage, Wisconsin, verify that the
presence of a gas stove contributes dramatically to personal NO₂
exposure levels. Table 16, derived from the reports of Quackenboss et
al. (1986) and based on data collected in 1981 and 1982, clearly shows
that gas stoves increase not only indoor concentrations but also the
personal exposure of children.

On the other hand, outdoor concentrations, even if measured
outside each residence, have been found to be relatively poor
predictors of personal exposure (Quackenboss et al., 1986; Leaderer et
al., 1986). The association between personal exposure and outdoor
levels of NO₂ is weakest during the winter for both gas and electric
stove groups.

The only route of NO₂ exposure is inhalation. The dose is
dependent on the inhalation volume and thus on physical activity, age,
etc. Lung absorption of NO₂ is about 80-90% during rest and over 90%
during physical activity (WHO, 1987).

Efforts have been made to find a sufficient biological marker for
NO₂ exposure and dose. Increased urinary excretion of collagen and
elastin (pulmonary connective tissue) breakdown products (including
hydroxyproline, hydroxylysine and desmosine) has been suggested as a
marker of diffuse pulmonary injury related to inhaled NO₂. A
significant relationship between urinary hydroxyproline excretion and
daily NO₂ exposure was found among housewives in Japan, but the
hydroxyproline excretion fell within the normal distribution for
healthy people (Yanagisawa et al., 1986). The majority of the
housewives were exposed to active or passive cigarette smoke, and this
exposure was independently related to the excretion of hydroxyproline.
Other investigators have not been able to substantiate the
relationship between urinary hydroxyproline excretion and NO₂
exposure (Muelenaer et al., 1987; Adgate et al., 1992).

Measurements of the NO-haem protein complex in bronchoalveolar
lavage (Maples et al., 1991) and of 3-nitrotyrosine in urine (Oshima
et al., 1990) have been suggested as biological markers for NO₂
exposure. The work in progress to find a suitable biological marker
for NO₂ exposure at levels found in the general environment is
promising; however, no metabolite has yet proved to be sensitive or
specific enough.

Personal exposure to air pollutants can be assessed by direct or
indirect measures. Direct measures include biomarkers and use of
personal monitors. No validated biomarkers for exposure are presently
available for NO₂.

Studies using passive monitors to measure NO₂ exposures lasting
one day to one week have been conducted in the USA (Dockery et al.,
1981; Clausing et al., 1986; Leaderer et al., 1986; Quackenboss et
al., 1986; Harlos et al., 1987; Schwab et al., 1990), in the
Netherlands (Hoek et al., 1984), in Japan (Nitta & Maeda, 1982;
Yanagisawa et al., 1984), and in Hong Kong (Koo et al., 1990). These
studies generally indicate that outdoor levels of NO₂, although
related to both personal levels and indoor concentrations, are poor
predictors of personal exposures for most populations. Average indoor
air residential concentrations (e.g., whole-house average or bedroom
level) tend to be the best predictor of personal exposure, typically
explaining 50 to 80% of the variation in personal exposures.

Indirect measures of personal exposure to NO₂ employ various
degrees of micro-environmental monitoring and questionnaires to
estimate an individual's or population's total exposure. One such
model (Billick et al., 1991), developed from an extensive monitoring
and questionnaire database, can estimate population exposure
distributions from easily obtained data on outdoor NO₂ concentrations,
housing characteristics and time-activity patterns. This model is
proposed for use in evaluating the impact of various NO₂ mitigation
measures. The model is promising, but has not yet been validated nor
has associated uncertainty been characterized.

3.7 Exposure of plants and ecosystems

The sensitivity of plants to nitrogen oxides is determined both
by their genetic characteristics and by environmental conditions.

The relation between exposure and uptake by plants depends on
aerodynamic and stomatal resistance and thus increases with increasing
light intensity, wind velocity and air humidity. After uptake, the
response of a plant depends on its metabolic activity, and thus
increases with poorer nutritional supply and lower temperature.

Moreover, the sensitivity of plants depends on the general air
pollution situation. Emission of SO₂ is often combined with NO_x,
and these compounds act synergistically. Therefore, the impact of
NO_x may be higher in regions with elevated SO₂ concentrations. NO_x
forms part of photochemical smog. Although ozone is the most
phytotoxic, the contribution of NO_x to adverse effects on plants is
not negligible.

For vegetation and ecosystems the impact of NO_x is through its
contribution to total nitrogen disposition rather than its direct
toxicity. Thus, other nitrogen-containing pollutants have to be taken
into consideration.

The dependencies of sensitivity, as summarized above, mean that
wide variation exists in the vulnerability of different regions of the
world.

4. EFFECTS OF ATMOSPHERIC NITROGEN COMPOUNDS (PARTICULARLY NITROGEN
OXIDES) ON PLANTS

Effects of nitrogen on ecosystems are caused through deposition
onto soil and foliar uptake of nitrogen in various forms. Total
effects are often difficult to separate into component effects. This
section, therefore, covers nitrogen inputs in all forms to ecosystems.
Much of the research focuses on European ecosystems, where the
majority of the research has been conducted. Here NH_y deposition
either dominates or is a major constituent of total nitrogen input.
However, this is not true for other parts of the world. All effects
of atmospheric nitrogen input, in whatever form, are included as
indicators of more globally relevant effects on ecosystems but the
reader should bear in mind local conditions of nitrogen input when
assessing likely local consequences.

NO_x, as used in this chapter, refers to the total nitrogen
measured by chemiluminescence detectors; this is NO₂ following
conversion to NO, and NO itself. Other nitrogen oxides may interfere
to some extent in this method.

Elemental nitrogen (N₂) forms 80% of the atmosphere of the
earth. This is equivalent to about 75 × 10⁶ kg above each hectare of
the earth's surface. In unpolluted conditions a small fraction
(1-15 kg nitrogen per ha per year) is converted by nitrogen-fixing
microorganisms to biologically more active forms of nitrogen: NH₄⁺
and NO₃^-. The natural deposition of nitrogen-containing atmospheric
compounds other than N₂ is much less. The soil contains 5 times more
nitrogen than the atmosphere, but weathering of rock is a negligible
source of biologically active nitrogen. By denitrification (reduction
of NO₃^- to N₂ and to a lesser extent N₂O, NO and NH₃), 1-30 kg
nitrogen per ha per year is recycled from the earth to the atmosphere.

Human activities, both industrial and agricultural, have greatly
increased the amount of biologically active nitrogen compounds,
thereby disturbing the natural nitrogen cycle. Various forms of
nitrogen pollute the air, mainly NO, NO₂ and NH₃ as dry deposition
and NO₃^- and NH₄⁺ as wet deposition. Another contribution
is from occult deposition (fog and clouds). There are many more
nitrogen-containing air pollutants (e.g., N₂O₅, PAN, N₂O, amines)
but these have not been considered in this chapter, either because
their contribution to the total nitrogen deposition is considered to
be small or because their concentrations are probably far below the
effect thresholds.

Transformations of nitrogen, as it moves from the atmosphere,
through ecosystems and back to the atmosphere, form the nitrogen
cycle. This is illustrated, together with anthropogenic sources of
nitrogen, in Fig. 14. The component processes affected by chronic
nitrogen deposition are indicated in Fig. 15.

Nitrogen-containing air pollutants can affect vegetation
indirectly, via chemical reactions in the atmosphere, or directly
after being deposited on vegetation, soil or water surfaces. The
indirect pathway is largely neglected in this chapter, although it
includes very relevant processes, and should be taken into account
when evaluating the entire impact of nitrogen-containing air
pollutants: NO and NO₂ are precursors for tropospheric ozone (O₃),
which acts both as a phytotoxin and a greenhouse gas. The effects of
O₃ will be discussed in a forthcoming Environmental Health Criteria
monograph. N₂O contributes to the depletion of stratospheric O₃,
resulting in increasing ultraviolet radiation. This and other aspects
of global climate change will be evaluated in a WHO/WMO/UNEP document
entitled "Climate and Health: potential impacts of climate change".
The direct impact of airborne nitrogen is due to toxic effects,
eutrophication and soil acidification. One effect of soil
acidification is that aluminum enters into solution, hence increasing
its bioavailability. This result in root damage. Aluminum toxicity
will be discussed in a further Environmental Health Criteria
monograph.

Most biodiversity is found in (semi-)natural ecosystems, both
aquatic and terrestrial. Nitrogen is the limiting nutrient for plant
growth in many (semi-)natural ecosystems. Most of the plant species
from these (semi-)natural habitats are adapted to nutrient-poor
conditions, and can only compete successfully in soils with low
nitrogen levels (Chapin, 1980; Tamm, 1991). Ellenberg (1988b) surveyed
the nitrogen requirements of 1805 plant species from Germany and
concluded that 50% can compete successfully only in habitats that are
deficient in nitrogen. Furthermore, of the plants threatened by
increased nitrogen deposition, 75-80% are indicator species for
low-nitrogen habitats. When stratified by ecosystem type, it is also
clear that the trend of rare species occurring with greater frequency
in nitrogen-poor habitats is a common phenomenon across many
ecosystems (Fig. 16 and Fig. 17). Plant species threatened by high
nitrogen deposition are common across many ecosystem types (Ellenberg,
1988b). The critical loads for nitrogen depend on (i) the type of
ecosystem; (ii) the land use and management in the past and present;
and (iii) the abiotic conditions (especially those which influence the
nitrification potential and immobilization rate in the soil). The
impact of increased nitrogen deposition upon biological systems can be
the result of direct uptake by the foliage or uptake via the soil.
The most relevant effects at the level of individual plants are injury
to the tissue, changes in biomass production and increased
susceptibility to secondary stress factors. At the vegetation level,
this results in changes in competitive relationships between species
and loss of biodiversity.

Effects on individual plants are discussed in section 4.1. The
following natural ecosystems are treated in detail in section 4.2:
freshwater ecosystems (shallow soft-water bodies, lakes and streams)
and terrestrial ecosystems (wetlands and bogs, species-rich
grasslands, heathlands and forests). Estuarine and marine systems
are also considered.

Air quality guidelines refer to thresholds for adverse effects.
Two different types of effect thresholds exist: critical levels and
critical loads.

The critical level is defined as:

the concentration in the atmosphere above which direct adverse
effects on receptors, such as plants, ecosystems or materials,
may occur according to present knowledge.

The critical load is defined as:

a quantitative estimate of an exposure (deposition) to one or
more pollutants below which significant harmful effects on
specified sensitive elements of the environment do not occur
according to present knowledge.

Generally, critical levels for nitrogen-containing air pollutants
are expressed in terms of exposure (µg/m³ and exposure duration),
while critical loads are expressed in terms of deposition (kg nitrogen
per ha per year). Both critical level and load are intended to be
set so as to protect vegetation, and can be converted into each
other knowing the deposition velocity. Thus, it might seem to be
superfluous to assess both critical levels and loads. However, with
the currently accepted approach, critical levels and loads are more or
less complementary: critical levels focus on effect thresholds for
short-term exposure (1 year or less), while critical loads focus on
safe deposition quantities for long-term exposure (10-100 years):
critical levels are not aimed to protect plants completely against
adverse effects. No-observed-effect concentrations (NOECs) are
usually lower than critical levels. For instance, a critical level
can be set at 5% yield reduction. Thus, owing simply to differences in
definition, the critical level is generally higher than the critical
load (Fig. 18b).

In current practice there are other differences between critical
levels and loads: critical levels give details on individual compounds
and focus on responses on plant level, while critical loads cover all
nitrogen-containing compounds and focus on the vegetation or ecosystem
level. In other words: critical loads focus on functioning of the
ecosystem, while critical levels focus on protection of the relatively
sensitive plant species.

In the critical level concept, the different nitrogen-containing
compounds are evaluated separately, because of their differences in
phytotoxic properties, even when their load in terms of kg nitrogen
per ha per year is the same (Ashenden et al., 1993). Another
difference between critical level and critical load is that critical
level considers the possibility of more- or less-than-additive effects
(Wellburn, 1990), while in the critical load concept additivity of
nitrogen-containing or acidifying compounds is presumed. Moreover,
nitrogen-containing air pollutants have their impact not only because
of their contribution to the nitrogen supply. Sometimes other effects
seem to dominate. For instance, although occult deposition is
generally small in terms of nitrogen deposition, it may be of great
significance because of its ability to affect plant surfaces.

It was concluded for these reasons that both critical levels and
loads are necessary within the scope of air quality guidelines for
nitrogen-containing compounds.

Assessing effect thresholds is relatively simple in the case of
toxic compounds with an exposure/response relationship which follows
the usual sigmoid curve: the lowest exposure level that results in a
response that is significantly different from the control treatment is
the effect threshold. However, this approach is essentially invalid
for exposure of nitrogen-limited vegetation to nitrogen-containing air
pollutants. Nitrogen is a macro-nutrient and so each addition of
nitrogen can result in a physiological response: growth stimulation
gradually increases with higher exposure levels and changes in growth
inhibition at higher levels (Fig. 18a). Moreover, depending on the
definition of adverse effects, the status of the vegetation may not be
optimal at background levels (Fig. 18b). These features complicate
the assessment of effect thresholds for nitrogen-containing compounds.
Nevertheless, in this chapter effect thresholds are presented,
according to current practice.

4.1 Properties of NO_x and NH_y

In this section general information is initially presented on
uptake, detoxification, metabolism and growth aspects. In the
following subsections the data determining the critical levels for
individual compounds and mixtures are discussed. The relevance of this
information and possibilities for generalization are discussed in
sections 4.1.8 and 4.1.9, where the critical levels are estimated.
Deposition on and emission from soils and vegetation is discussed in
chapter 3.

4.1.1 Adsorption and uptake

The impact of a pollutant on plants is determined by its
adsorption, rate of uptake (flux) and the reaction of the plants.
Foliar uptake is probably dominant for NO, NO₂ (Wellburn, 1990) and
NH₃ (Pérez-Soba & Van der Eerden, 1993), while the pathway via soil
and roots is the major route for nitrogen-containing pollutants in wet
deposition.

The flux of the compounds from the atmosphere into the plant is a
complicated process, which is highly dependent on the properties of
both plant and compound and on environmental conditions. This is why
deposition velocities proved to be highly variable (chapter 3).

The flux from the atmosphere to the leaf surface (and soil)
depends on the aerodynamic and boundary layer resistances, which
are determined by meteorological conditions and plant and leaf
architecture. The flux from the leaf surface to the final site of
reaction in the cell is determined by stomatal, cuticular and
mesophyll resistance. The reaction of the plant to the nitrogen that
arrives at the target side is dependent on the intrinsic properties of
the plant and on its nutritional status, and again on environmental
conditions.

The flux of atmospheric nitrogen through the soil is conditioned
by properties of soil and vegetation and by meteorological conditions.
The chemical composition of soil water, the rate of nitrification
(NH₄⁺ -> NO₃^-), the preference of the plant for either NH₄⁺ or
NO₃^-, the root architecture and the metabolic activity of the plants
play major roles in this uptake (Schulze et al., 1989).

Adsorption on the outer surface of leaves certainly takes
place. Exposure to relatively high concentrations of gaseous NH₃
(180 µg/m³) or NH₄⁺ in rainwater (5 mmol/litre) damages the
crystalline structure of the epicuticular wax layer of the needles of
Pseudotsuga menziesii (Van der Eerden & Pérez-Soba, 1992). NO₂
(Fowler et al., 1980) and NH₄⁺ and NO₃^- in wet and occult
deposition can disturb leaf surfaces in several ways (Jacobson, 1991).
The quantitative relevance of this effect for the field situations has
not yet been shown in detail.

Uptake of NH₃ and NO_x is driven by the concentration gradient
between atmosphere and mesophyll. It is generally directly determined
by stomatal conductance and thus depends on factors influencing
stomatal aperture. Although in higher plants uptake through the
stomata strongly dominates, there are indications that penetration
through the cuticle is not completely negligible. This has been
demonstrated for NO and NO₂ (Wellburn, 1990). While stomata greatly
influence the foliar uptake of aerial nitrogen compounds, many of
these compounds subsequently alter stomatal aperture and the extent of

further uptake. The nitrogen status of plants is also known to affect
stomatal behaviour towards other environmental conditions such as
drought (Ghashghaie & Saugier, 1989).

The flux of NH₃ into a plant appeared to be linearly related to
the atmospheric concentration (Van der Eerden et al., 1991), there
being no mesophyll resistance (Van Hove et al., 1989). This relation
can become less then linear with high concentrations, e.g., some
hundreds of µg/m³ (Wollenheber & Raven, 1993). Mesophyll resistance
is, however, probably more significant for NO and NO₂ (Capron et al.,
1994).

There is increasing evidence that foliar uptake of nitrogen
reduces the uptake of nitrogen by the roots (Srivastava & Ormrod,
1986; Pérez-Soba & van der Eerden, 1993), although the driving
mechanism is not yet clear.

In the presence of low concentrations plants can emit NH₃,
rather than absorb it (chapter 3). NO and N₂O are emitted in
significant quantities by the soil (chapter 3).

Rain, clouds, fog and aerosols always contain significant amounts
of ions including NH₄⁺ and NO₃^-. In the past, foliar uptake of
nitrogen from wet deposition was considered to be negligible, but
recent research using ¹⁵N and throughfall analysis shows that this
path can contribute a high proportion of the total plant uptake (see
Pearson & Stewart, 1993, and section 2.4). In general, cations (e.g.,
NH₄⁺) are more easily taken up through the cuticle than anions
(e.g., NO₃^-). A substantial foliar uptake of NH₄⁺ from rainwater
has been measured in several tree species (Garten & Hanson, 1989).
Lower plants, such as bryophytes and lichens do not have stomata and a
waxy waterproof cuticle, and thus the potential for direct uptake of
pollutants in the form of wet or dry deposition is greater. Epiphytic
lichens are active absorbers of both NH₄⁺ and NO₃^- (Reiners &
Olson, 1984). Uptake and exchange of ions through the leaf surface is
a relatively slow process, and thus is only relevant if the surface
remains wet for long periods.

4.1.2 Toxicity, detoxification and assimilation

One would expect a positive relationship between the solubility
of a compound and its biological impact. NO is only slightly soluble
in water, but the presence of other substances can alter its
solubility. NO₂ has a higher solubility, while that of NH₃ is much
higher.

Much information exists on mechanisms of toxicity, although it is
sometimes confusing. NO₂, NO, HNO₂ and HNO₃ can be incorporated
into nitrogen metabolism using the pathway: NO₃^- -> NO₂^- ->

(NH₃ <--> NH₄⁺) <--> glutamate -> glutamine -> other amino
acids, amides, proteins, polyamines, etc. The enzymes involved
include nitrate reductase (NR), nitrite reductase (NiR) and glutamine
synthetase (GS). Glutamate dehydrogenase (GDH) plays a role in the
internal cycling of NH₄⁺.

After exposure to NO₂, nitrate can accumulate for some weeks;
accumulation of nitrite is rarely reported, although it is certainly
an intermediate. Nitrite levels can be elevated for some hours due to
the fact that NR activity is induced faster than that of NiR. In many
cases storage of excess nitrogen has been found to be in the form of
arginine (Van Dijk & Roelofs, 1988), which could last months or
longer.

NO₂^-, NH₃ and NH₄⁺ are highly phytotoxic, and could well be
the cause of adverse effects of nitrogen-containing air pollutants.
Wellburn (1990) suggested that the free radical *N=O plays a role in
the phytotoxicity of NO_x.

High concentrations can cause visible injury via lipid breakdown
and cellular plasmolysis. At lower concentrations inhibition of lipid
biosynthesis may dominate, rather than damage of existing lipids
(Wellburn, 1990).

Raven (1988) assumed that the adverse effects of nitrogen-
containing compounds are due to their interference with cellular
acid/base regulation. They can influence cellular pH both before
and after assimilation. Assimilation of most air pollutants,
including NH₃, has been shown to result in production of protons
(Wollenheber & Raven, 1993). Assimilation of nitrate and a high
buffer capacity can prevent the plant from being damaged by this
acidification (Pearson & Stewart, 1993). If these adverse effects can
effectively be counteracted, assimilation of nitrogen-containing
compounds will result in growth stimulation.

Synergistic effects have been found in nearly all studies
concerning SO₂ and NO₂ (Wellburn et al., 1981). Inhibition of NiR
by SO₂, resulting in the inability of the plant to detoxify nitrite,
might be the cause of this interaction.

4.1.3 Physiology and growth aspects

When climatic conditions and nutrient supply allow biomass
production, both NO_x and NH_y result in growth stimulation at low
concentrations and growth reduction at higher concentrations.
However, the exposure level at which growth stimulation turns into
growth inhibition is much lower for NO_x than for NH_y (see Fig. 18a).

Foliar uptake of NH₃ generally results in an increase in
photosynthesis and dark respiration, and in the amount of RUBISCO
(ribulose 1,5-biphosphate carboxylase oxygenase) and chlorophyll.
Some authors have shown that stomatal conductance increases and the
stomata remain open in the dark, resulting in enhanced transpiration
and drought sensitivity (Van der Eerden & Pérez-Soba, 1992). Most
experiments with NO and NO₂ have been conducted with relatively high
concentration levels (> 200 µg/m³). These experiments show
inhibition of photosynthesis by both NO and NO₂, possibly additively
(Capron & Mansfield, 1976). Inhibition by NO may be stronger than
that of NO₂ (Saxe, 1986). The threshold for this response is well
below the threshold for visible injury (Wellburn, 1990) and
transpiration (Saxe, 1986). With lower (nearer to ambient) NO_x
concentrations, stimulation of photosynthesis may well occur. Both
NO_x and NH_y generally cause an increase in shoot/root ratio. The
specific root length and the amount of mycorrhizal infection can be
reduced by both compounds. However, these alterations in root
properties resemble general responses to increased nitrogen nutrient
supply.

4.1.4 Interactions with climatic conditions

Evidence suggests that exposure of vegetation to NH₃ and to
mixtures of NO₂ and SO₂ can influence the subsequent response to
drought and frost stress. There is also evidence that environmental
conditions can affect the response to NO_x and to NH₃.

The foliar uptake of nitrogenous compounds in the form of wet and
occult deposition is largely via the cuticle. Uptake and exchange of
ions through the leaf surface is a relatively slow process, and thus
is especially relevant if the surface remains wet for longer periods,
e.g., in regions where exposure to mist and clouds is frequent.

The solubility of most gases, including NO, NO₂ and NH₃, rises
as temperature falls, while the metabolic activity of plants and thus
their detoxification capacity is lower. On the other hand, stomatal
conductivity and thus the influx of gases generally falls as
temperature falls.

Guderian (1988) proposed a lower critical level in winter than
for the whole year, in acknowledgement of several results that
indicate greater toxicity of NO₂ during winter conditions. For
example, exposure of Poa pratensis in outdoor chambers to 120 µg/m³
inhibited growth during winter but had little effect or even
stimulated growth in summer and autumn (Whitmore & Freer-Smith, 1982).
Mortensen (1986) found that low light and non-injurious low
temperature conditions can enhance the toxicity of NO_x. Capron et
al. (1991) found that the depression relative to the control of net
photosynthesis by 1250 µg NO/m³ plus 575 µg NO₂/m³ at 10°C was
three times, and at 5°C was almost five times, that recorded at 20°C.

An interaction between NO_x and temperature may also occur at lower
realistic concentrations. This is suggested by the observation of
nitrite accumulation at low temperatures during fumigation of
Picea rubra with 38 µg NO₂/m³ plus 54 µg SO₂/m³ (Wolfenden et
al., 1991). This temperature effect may play a role in combination
with elevated concentrations of CO₂ as well: the stimulating effect
of CO₂ on net photosynthesis was inhibited by NO_x to a larger extent
when applied at lower temperature (Capron et al., 1994). Observation
of NH₃ injury to plants also indicates that this is greatest in
winter (Van der Eerden, 1982).

In contrast with the view that NO_x (and NH₃) injury is greater
at low temperatures, Srivastava et al. (1975) found that inhibition by
NO_x of photosynthesis was greatest under optimal temperature and high
light conditions, when stomatal conductance to the gas would be
highest.

The exposure of plants to NO_x and NH₃ may reduce their ability
to withstand drought stress, owing to loss of control of transpiration
by stomata and to an increase in the shoot/root ratio (Lucas, 1990;
Atkinson et al., 1991; Fangmeijer et al., 1994).

4.1.5 Interactions with the habitat

Whether the atmospheric input of nitrogen has a positive or
negative impact depends on the plant species and habitat. Based on
experimental evidence, Pearson & Stewart (1993) hypothesized that
species which are part of a climax vegetation on nutrient-poor acidic
soils are often relatively sensitive to NO_x and NH_y. Morgan et al.
(1992) found that NO_x disrupted the NR activity to a greater extent
in calcifuge than calcicole moss species. Ombrotrophic mires and
other strongly nitrogen-limited systems may be especially prone to
detrimental effects from input of nitrogen-containing air pollutants.

The assimilation of low concentrations of NO₂ and the
incorporation into amino acids by NR (Morgan et al., 1992; Table 20)
are indicators that this pollutant can contribute to the nitrogen
budget of plants (Pérez-Soba et al., 1994). The contribution of NO_x
to the nitrogen supply increases as root-available nitrogen is lowered
(Okano & Totsuka, 1986; Rowland et al., 1987). Srivastava & Ormrod
(1986) observed reduced ability to respond to a supply of nitrate to
the roots when Hordeum vulgare was fumigated with NO₂. Similarly,
Pérez-Soba & Van der Eerden (1993) found reduced uptake of NH_4⁺ from
the soil when Pinus sylvestris was fumigated with NH₃. Although
there is much evidence that nitrogen-containing air pollutants play a
role in the nitrogen demand and nitrogen metabolism of the plant,
Ashenden et al. (1993) found no obvious relationship between
sensitivity to NO₂ and nitrogen preference, as indicated by Ellenberg
(1985).

4.1.6 Increasing pest incidence

Any change in chemical composition of plants due to the uptake of
nitrogenous air pollutants could alter the behaviour of pests and
pathogens. Evidence indicates that, in general, NO_x and NH_y
increase the growth rate of herbivorous insects (Dohmen et al., 1984;
Flückiger & Braun, 1986; Houlden et al., 1990; Van der Eerden et al.,
1991). This may also apply to fungal pathogens (van Dijk et al.,
1992).

4.1.7 Conclusions for various atmospheric nitrogen species and
mixtures

4.1.7.1 NO₂

In Table 20 the lowest effective exposure levels for NO₂ are
given. Three different types of effects are considered:

* (bio)chemical: e.g., enzyme activity, consumption quality
* physiological: e.g., CO₂ assimilation, stomatal conductivity
* growth aspects: e.g., biomass, reproduction, stress sensitivity

Four exposure durations are used in this table. These are
(including an indication of the exposure durations and the margins):

* short term (hours): < 8 h
* air pollution episodes (days): 8 h to 2 weeks
* growing season or winter season (months): 2 weeks to 6 months
* long term (years): > 6 months

To avoid the information being too selective, in each cell in
this table a species is used only once. For each cell the three
lowest effective concentrations and exposure durations are given;
species and references are mentioned in footnotes. Exposure levels
far higher than current levels measured in the field situation have
not been considered.

The fact that not all cells in Table 20 are filled with
information is because many of the experiments have been conducted
with unrealistically high concentrations. The majority of observations
mentioned in the table are on crops; several of these show growth
stimulation. Most of the responses on a biochemical level deal with
enhanced NR activity, which shows that the plants are capable of
assimilating the NO₂. A general effect threshold as derived from
Table 20 would be substantially higher if enhanced NR and biomass
production of crops were not assumed to be an adverse effect.
However, growth stimulation is often considered an adverse effect in
most types of natural vegetation. Moreover, Pearson & Stewart (1993)

assumed detoxification of NH_y and NO_x to be a potentially adverse
effect, because it contributes to cellular acidification, which can
not always be counteracted.

4.1.7.2 NO

In Table 21 the lowest effective exposure levels for NO are
given.

Most research into the effects of nitric oxide has been based on
glasshouse crops, particularly the tomato (Lycopersicon esculentum).
Virtually all experiments deal with photosynthesis or enzymatic
reactions and a few growth aspects were measured. The existing data
are rather difficult to interpret since controlled fumigation with NO
inevitably results in some oxidation to NO₂. Thus atmospheres will
contain a mixture of the oxides. There is growing interest in the
distinct properties and effects of NO and NO₂, and the mechanisms of
their cellular action probably differ (Wellburn, 1990). The exchange
properties of NO and NO₂ over vegetation (personal communication by
D. Fowler to the IPCS) and single plants (Saxe, 1986) appear quite
different. Their effects are also contrasting, and there is clearly
some dispute over which oxide is the most toxic. Earlier studies of
the inhibition of photosynthesis found NO to act more rapidly than
NO₂ (at several ppm) but to cause less overall depression of the
photosynthetic rate (Hill & Bennet, 1970). More recent photosynthetic
studies by Saxe (1986), using similar concentrations, found NO to be
considerably more toxic than NO₂. There is very little information
on contrasting effects of the two oxides at low concentrations, but
this also adds weight to the suggestion that NO is biologically more
toxic. In her studies of NR in bryophytes, Morgan et al. (1992)
discovered that exposure to NO initially inhibited NR while NO₂
induced activity. At present, however, there is insufficient
knowledge across a range of species to establish separate critical
levels for NO and NO₂, and studies using a wider variety of
vegetation are urgently required.

4.1.7.3 NH₃

The lowest effective exposure levels for NH₃ are given in Table
22.

The toxicity of NH₃ during very short exposure periods has been
tested for the purpose of evaluating accidental releases during
transport or industrial processes. The estimated critical level for
10 min is (100 ppm) (personal communication by Lee & Davison to the
IPCS). This type of exposure is out of the context of this monograph.

Table 20. Lowest exposure levels (in µg/m³) and durations at which NO₂
caused significant effects^a

(Bio)chemical Physiological Growth aspects

Long term 200 (130); 104 h/week;
7 months^r
120-500; 9.5 months^s
122; 37 weeks^t

Growing season 50; 39 days^b 120; 22 days^j 10-43; 130 days^u
or winter 125; 140 days^c 190 (65); 105 h 55-75; 62 days^v
940; 19 days^d in 15 days^k 150-190 (28-33);
120 h in 40 days^w

Air pollution 140; 1 day^e 375 (165); 35 h in 375; 2 weeks^x
episodes 160; 7 days^f 5 days^l 190; 3 days^m 100 (25);
65; 1 day^g 375 (165); 35 h 20 h in 5 days^y
in 5 daysⁿ

Short term 7500, 6 h^h 940; 1 h^o 2000-3000; 3.5 h^z
7500; 4 hⁱ 850; 7 h^p
1100; 1.5 h^q

^a If the fumigation was not continuous an average value has been estimated
and quoted in parentheses (calculated assuming 10 µg/m³ during the periods
in which the fumigation was switched off).
^b Pinus sylvestris; changes in amino acid composition, with no physiological
changes (Näsholm et al., 1991)
^c Lolium perenne; increase in GDH activity (Wellburn et al., 1981)
^d Lycopersicum esculentum; decrease in nitrate content of the leaves (Taylor
& Eaton, 1966)
^e Picea rubens, increase in NR activity (Norby et al., 1989)

Table 20 (Con't)

^f Pinus sylvestris, increase in NR activity (Wingsle et al., 1987)
^g Several bryophyte species; increase in NR activity (Morgan et al., 1992)
^h Zea mais; increase in NiR activity (Yoneyama et al., 1979)
ⁱ Vicia faba; change in amino acid composition (Ito et al., 1984)
^j Betula sp; increased water loss (Neighbour et al., 1988)
^k Phaseolus vulgaris; reversible increase in dark respiration (Sandhu & Gupta, 1989)
^l Glycine max; increase in photosynthesis (Sabarathnam et al., 1988a,b)
^m Phaseolus vulgaris; increase in transpiration (Ashenden, 1979)
ⁿ Glycine max; enhanced dark respiration (Sabarathnam et al., 1988b)
^o Vicia faba; reversible structural damage on cellular level (Wellburn et al., 1972)
^p Pisum sativum; emission of stress ethylene (Mehlhorn & Wellburn, 1987)
^q Medicago sativa, Avena sativa; inhibition of photosynthesis (Hill & Bennet, 1970)
^r Several grass species; reduction in shoot growth (Whitmore & Mansfield, 1983)
^s Citrus sinensis; increased fruit drop (Thompson et al., 1970)
^t Polytrichum formosum and 3 fern species; injury and changes in growth (Ashenden
et al., 1990; Bell et al., 1992)
^u Brassica napus and Hordeum vulgare; growth stimulation (resp.: Adaros et al.,
1991a,b)
^v Phaseolus vulgaris; increase in total dry matter, not in yield (Bender et al.,
1991)
^w Raphanus sativus; growth stimulation (Runeckles & Palmer, 1987)
^x Helianthus annuus; reduction in net assimilation rate (Okano et al., 1985b)
^y Pinus strobus; slight needle necrosis in 2 of 8 clones (Yang et al., 1983)
^z Nicotiana tabacum; leaf necrosis (Bush et al., 1962)

Table 21. Lowest exposure levels (in µg/m³) at which NO caused
significant effects^a

(Bio)chemical Physiological Growth aspects

Growing season 44; 21 days^b 625; 16 daysⁿ
500; 28 days^c 500;^o

Air pollution 375; 8 days^d 1250; 4 daysⁱ 1250; 5 days^p
episodes 44; 8-24 h^e 125; 20 h^j
1875; 18 h^f

Short term 188; 7 h^g 750; 1 h^k
500; 3 h^h 2500; 10 min^l
1875; 20 min^m

^a If the fumigation was not continuous an average value has been
estimated and quoted in parentheses (calculated assuming 10 µg/m³
during the periods in which the fumigation was switched off).
^b Four bryophyte species; inhibition of nitrate-induction of NR
(Morgan et al., 1992)
^c Lycopersicon esculentum; induction of NiR (Wellburn et al., 1980)
^d Lactuca sativa; induction of NiR (Besford & Hand, 1989)
^e Ctenidium molluscum (bryophyte); inhibition of NR (Morgan et al., 1992)
^f Capsicum annum; reduction in NiR activity (Murray & Wellburn, 1980)
^g Pisum sativum; increase in ethylene release (Mehlhorn & Wellburn, 1987)
^h Lycopersicon esculentum; induction of NiR (Wellburn et al., 1980)
ⁱ Eight indoor ornamental species; inhibition of photosynthesis
(Saxe, 1986)
^j Lycopersicon esculentum; inhibition of photosynthesis (Capron &
Mansfield, 1989)
^k Avena sativa & Medicago sativa; inhibition of photosynthesis (Hill &
Bennet, 1970)
^l Lactuca sativa; inhibition of photosynthesis (Capron, 1989)
^m Lycopersicon esculentum; inhibition of photosynthesis (Mortensen, 1986)
ⁿ Lactuca sativa; reduction in plant mass (Capron et al., 1991)
^o Lycopersicon esculentum; reduction in plant mass (Anderson &
Mansfield, 1979)
^p Lycopersicon esculentum; reduction in plant mass (Bruggink et al., 1988)

Table 22. Lowest exposure levels (in µg/m³) at which NH₃ caused
significant effects^a

(Bio)chemical Physiological Growth aspects

Long term 50; 8 months^b 53; 9 months^h 25; 1 year^k
53; 8 months^l
35; 16 months^m

Growing season 100; 6 weeks^c 50; 6 weeksⁱ 60; 2 monthsⁿ
or winter 60; 14 weeks^d 20; 90 days^o
180; 13 weeks^e 30; 23 days^p

Air pollution 2000; 24 h^f 213; 5 days^j 120; 11 days^q
episodes 213; 5 days^g 1000; 2 weeks^r
300; 3 days^s

Short term 30 000; 1 h^t
2000 2 h^u
2000 6 h^v

^a If the fumigation was not continuous an average value has been
estimated and quoted in parentheses (calculated assuming
10 µg/m³ during the periods in which the fumigation was
switched off).
^b Species of Violion caninea alliance; imbalanced nutrient
status (Dueck & Elderson, 1992)
^c Deschampsia flexuosa; change in amino acid composition (Van
der Eerden et al., 1990)
^d Pinus sylvestris; increased GS activity (Pérez-Soba et al.,
1990)
^e Pseudotsuga menziesii; imbalanced nutrient status (Van der
Eerden et al., 1992)
^f Lycopersicum esculentum; increase of NH_4⁺ in the dark
(Van der Eerden, 1982)
^g Lolium perenne; 30% of N in the plant is derived from
foliar uptake (Wollenheber & Raven, 1993)
^h Pinus sylvestris; increased loss of water after two weeks
of desiccation (Dueck et al., 1990)
ⁱ Populus sp.; increase in stomatal conductance in leaves;
increase in mesophyll conductance and maximum photosynthetic
rate at a slightly higher exposure level (Van Hove et al., 1989)
^j Lolium perenne; significant impact acid/base regulation and
nutrients status

Table 22 (Con't)

^k Pseudotsuga menziesii; erosion of wax layer (Thijse & Baas,
1990; the authors have some doubts about the causality of this
effect (personal communication)
^l Calluna vulgaris; reduction in survival rate after winter
(Dueck, 1990)
^m Arnica montana; reduced survival after winter and flowering
(Van der Eerden et al., 1991)
ⁿ Field exposure during winter; median concentration; severe
injury of several conifer species (Van der Eerden, 1982)
^o Viola canina, Agrostis capillaris; 50% growth stimulation
of the shoot (not of the roots) (Van der Eerden et al., 1991)
^p Racomitrium lanuginosum; chlorosis (Van der Eerden et al.,
1991)
^q Hypnum jutlandicum; chlorosis (Van der Eerden et al., 1991)
^r Lepidium sativum; reduction in dry weight (Van Haut &
Prinz, 1979)
^s Several horticultural crops; leaf injury
^t Various deciduous trees; leaf injury (Ewert, 1979)
^u Brassica sp., Helianthus sp.; leaf injury (Benedict & Breen,
1955)
^v Rosa sp.; leaf injury rose (Garber, 1935)

Several cells in Table 22 could not be filled; the majority of quoted
effects are on biomass production and tissue injury. It is clear that
the data in this table are not random; nearly all of the information
originating from one Dutch research group. Only a few pollution
climates were considered. The results may be representative for
mild oceanic climates, but probably not for cold climates with dark
winters: toxicity of NH₃ increases with lower temperature and lower
light intensity. The effects of NH₃ need to be studied with more
plant species and under more climatic conditions in order to obtain
critical levels with sufficient potential for generalization.

4.1.7.4 NH₄⁺ and NO₃^- in wet and occult deposition

NH₄⁺, NO₃^- and H⁺ make up about half of the ionic
composition of rain, clouds, fog and aerosols. The impact of the
acidity of rain and clouds has received much attention in recent years
(Jacobson, 1991). This is not the case with other compounds in wet
deposition, although their relevance is recognized. At the same pH,
Cape et al. (1991) found a much greater effect of sulfuric acid than
of nitric acid, indicating that the impact of acid rain is not only
through protons, but also through anions.

There is an abundance of information on the effects of NH₄⁺ in
soil solution. However, threshold concentrations for NH₄⁺ in the
soil (e.g. Schenk & Wehrman, 1979) can not be considered a critical
level for rain water quality, because the type of exposure and
response is completely different.

Wet deposition containing NH₄⁺ can reduce frost tolerance (Cape
et al., 1990) and induce leaching of K⁺ and other cations (Roelofs
et al., 1985). It is not yet clear whether this type of ion exchange
can have deleterious effects on its own in the field situation.

Currently, too few quantitative data on the effects of nitrogen-
containing wet and occult deposition are available for critical levels
for this group of compounds to be derived.

4.1.7.5 Mixtures

A polluted atmosphere generally consists of a cocktail of
compounds, but certain combinations are more frequent. Because of
their role in the formation of tropospheric O₃, simultaneous
co-occurrence of relatively high levels of O₃ and NO are rarely
observed, while sequential co-occurrences are much more frequent
(Kosta-Rick & Manning, 1993). If burning of fossil fuels results in
emission of SO₂, this is often combined with emission of NO_x.

a) SO₂ plus NO₂

Synergism has been found in nearly all studies concerning this
combination, with only few exceptions (Kuppers & Klump 1988; Murray et
al., 1992). Based on data presented by Whitmore (1985), for Poa
pratensis the effect threshold for combinations of SO₂ and NO₂, in
equal concentrations when expressed in ppm, is in the range of 1.2-2.0
ppm.days (Fig. 19). This threshold applies to effects by combinations
of SO₂ and NO₂; the effects of single exposures were not assessed in
this study. However, it is reasonable from other references to expect
synergism, and thus to include this threshold in Table 23, in which
combined effects are summarized. Another threshold for combinations
of SO₂ and NO₂ was defined by Van der Eerden & Duym (1988) (Fig.
20; Table 23).

b) SO₂ plus NH₃

Adsorption of either NH₃ or SO₂ on leaf surfaces is enhanced by
the presence of the other compound (Van Hove et al., 1989).
Interactive physiological effects have been found as well (Dueck,
1990; Dueck et al., 1990; Dueck & Elderson, 1992). Currently, there
is far too little information on this combination to quantify this
interaction.

Table 23. Lowest exposure levels at which NO₂ increases the
effect of SO₂, O₃, or SO₂ plus O₃

(Bio)chemical Physiological Growth aspects

Long term 150-190; 9 months^f
220; 60 weeks^g
19; 10-41 weeks^h

Growing season 55-75; 34 days^b 135; 28 days^d 30; 38 daysⁱ
or winter 135; 28 days^c 10-43; 130 days^j
30; 43 days^k

Air pollution 80; 2 weeks^l
episodes 75; 1 day^m

Short term 153; 1 h^e 325; 1 h^m
400; 1 hⁿ

^a If the fumigation was not continuous an average value has
been estimated and quoted in parentheses (calculated
assuming 10 µg/m³ during the periods in which the
fumigation was switched off).
^b Phaseolus vulgaris; inhibition of parts of nitrogen
metabolism, when exposed sequentially with O₃
(100-120 µg/m³; 8 h/day)
^c Lolium perenne; decrease in proline content during winter
hardening when applied in combination with SO₂ at
188 µg/m³ (Davison et al., 1987)
^d Lolium perenne; less negative osmotic potential during
winter hardening when applied in combination with SO₂ at
188 µg/m³ (Davison et al., 1987)
^e Phaseolus vulgaris; Inhibition of photosynthesis when in
combination with SO₂ (215 µg/m³); without SO₂
inhibition at 760 µg/m³ (Bennet et al., 1990)
^f Several crops; growth stimulation by NO₂ turns into a
reduction in synergism with sequential treatment with O₃
(160-200 µg/m³; 6 h/day) (Runeckles & Palmer, 1987)
^g Six tree species; reduced plant growth in combination with
SO₂ (280 µg/m³), both antagonism and synergism
(Freer-Smith, 1984)
^h 10 grass species were tested in combination with SO₂
(27 µg/m³). Three species showed growth stimulation.
Reduced growth was found at higher concentrations.
Interactions with acidic mist and with O₃ were found
(Ashenden et al., 1993).

Table 23 (Con't)

ⁱ Poa pratensis; inhibition of biomass production; in
combination with SO₂ (42 µg/m³) for 38 days; the longest
exposure period used in the experiments. Calculated from
data from Whitmore (1985), assuming synergism and a critical
level for SO₂ plus NO₂ of 1.2 ppm.days (Whitmore,
1985).
^j Brassica napus and Hordeum vulgare; antagonism (and
rarely synergism) with O₃ (6-44 µg/m³; 8 h/day) and
SO₂ (9-33 µg/m³, continuously): enhanced yield turns into
reduction (Adaros et al., 1991a,b)
^k Plantago mayor; reduced shoot dry weight synergism with
SO₂ (60 µg/m³) and O₃ (60 µg/m³, 8 h/day)
(Mooi, 1984)
^l Poa pratensis; inhibition of biomass production; in
combination with SO₂ (110 µg/m³) for 2 weeks (the upper
margin of the exposure period of this cell in the table; the
shortest fumigation in this survey was 20 days. Calculated
from data from Whitmore (1985), assuming synergism and a
critical level for SO₂ plus NO₂ of 1.2 ppm.days
(Whitmore, 1985).
^m Critical level for NO₂ assuming SO₂ to be present at
70 µg/m³; considered to be a critical level for a 24-h mean
(UNECE, 1994) (Van der Eerden & Duym, 1988)
ⁿ Lycopersicon esculentum; reduction in plant mass if in
combination or preceded by O₃ (160 µg/m³; 1 h)
(Goodyear & Ormrod, 1988).

c) NO plus NO₂

When activated charcoal has been used as filter material in NO₂
fumigation experiments, NO must have been present as well, because
activated charcoal has virtually no capacity to absorb NO. In those
studies, responses must have been due to NO₂ plus NO. Although the
toxicity of NO was often considered to be much less than that of NO₂,
currently the two compounds are assumed to be equally toxic and to
act additively. However, Wellburn (1990) and others have stated
that NO is more toxic, and Saxe (1994) showed that the variation in
sensitivity amongst species is different for the two compounds. This
supports the suggestion of Wellburn that the mechanism of toxicity is
different.

For the purpose of deriving critical levels, the assumption of
additivity may result in an underestimation. However, there are not
enough data to quantify this.

d) Mixtures with O₃

The combination NH₃ plus O₃ has rarely been studied. No
statistically significant interactions have been found as yet, but in
one study the threshold for leaf injury was higher in the presence
of NH₃ (Van der Eerden et al., 1994). The combination NO₂ plus O₃
has been studied more frequently, but the responses differed
considerably between experiments and species. Additivity or
antagonism was found by Runeckles & Palmer (1987), Adaros et al.
(1991a,b), and Bender et al. (1991). Synergism was reported by Ito et
al. (1984), Runeckles & Palmer (1987) and Kosta-Rick & Manning (1993).

The combination of SO₂ plus O₃ plus NO₂ has also been studied.
Again the responses varied between plant species and experiment.
Antagonism, additivity and synergism have all been found (Kosta-Rick &
Manning, 1993).

e) Mixtures with elevated CO₂

Generally, an increased supply of CO₂ to crops results in an
enhanced biomass production. The responses of native species are more
variable but are also frequently positive. This growth stimulation is
limited by a deficiency of other nutrients. Nitrogen can be one such
limiting factor, and for this reason a nitrogen fertilizer such as
NH_y and possibly low NO_x concentrations could be hypothesized to
have a more-than-additive relationship with CO₂. However, as yet
there is no experimental evidence for this. Van der Eerden
et al. (1994) and Pérez-Soba et al. (1994) found stimulation of
photosynthesis and growth by both NH₃ and CO₂, but not by a
combination of these two compounds.

Effects of the combination of NO_x and CO₂ have not yet been
studied within the scope of global climate change. But some relevant
information could be gained from the literature dealing with CO₂
enrichment in glasshouses. When the flue gases of the heating system
are used as a CO₂ source, NO_x (in which NO is dominant) becomes a
major contaminant. The fertilizing effect of elevated CO₂ can
largely disappear in the presence of NO_x (Anderson & Mansfield, 1979;
Saxe & Voight Christensen, 1984; Mortensen, 1985; Bruggink et al.,
1988; Capron et al., 1994).

The CO₂, NH₃ and NO_x concentrations used in combination in
these experiments were relatively high and therefore cannot be used
in the critical level assessment. More experiments with lower
concentrations are required.

Table 23 indicates, surprisingly, that the effective long-term
exposures are generally higher than those of shorter duration.
However, long-term responses (more than half a year) have rarely been
studied. Therefore, the information on effects over growing season
periods may be much more representative of long-term effects.

A study included in a report by UNECE (1994) used 21 µg SO₂/m³
and 11 µg NO₂/m³, over the entire growing season and found synergism
in reducing biomass production of Pisum sativum and Spinacea
oleracea. Similar results were found for Hordeum vulgare and
Brassica oleracea, when fumigation was conducted for 120-190 days
with 30-40 µg SO₂/m³ and 30-50 µg NO₂/m³. This study cannot be
used for the assessment of critical levels because it has not yet been
published, but it indicates that lower levels of the two pollutants
than those quoted in Table 23 can influence plant responses.

4.1.8 Appraisal

Table 24 shows the former air quality guidelines for NO₂ and
some other critical levels assessed in the same period. Fig. 21
summarizes the results given in Tables 20 to 23. In this figure
curves are drawn to estimate critical levels according to current
practice, known as the "envelope" approach. After having plotted all
effective exposure levels in a graph of concentration and exposure
time, a curve is drawn just below the lowest effective exposures.
Critical levels can be derived from this curve. Fig. 21 shows that
more experiments with exposure periods of 0.5 to 5 days are required
to give a more solid basis for the estimation of critical levels of
24 h.

Table 24. Critical levels for NO₂

Concentration Exposure time Reference
(µg/m³)

95 4 h WHO (1987)
30^a annual mean WHO (1987)
800 1 h Guderian (1988)
60 growing season Guderian (1988)
40 winter Guderian (1988)

^a SO₂ and O₃ not higher than 30 µg/m³ and 60 µg/m³, respectively

A second approach to arrive at critical levels is the statistical
model of Kooijman (1987). Based on the variation in sensitivity
between species, critical levels are calculated taking into account
the number of tested species and the level of uncertainty (Van der
Eerden et al., 1991). The second approach is better, but only part of
the available data is suitable for this approach.

Tables 20 to 23 show that some new relevant information has
appeared. Comparing the data of Table 20 with those of Table 21
(Fig. 21a and 21b), it appears that NO₂ has slightly higher effect
thresholds than NO. However, this probably reflects the separate
attention paid to these compounds, rather than any difference in
toxicity. It is now obvious that the toxicity of NO cannot be
ignored, and it should be included in the guidance values. The
consideration of NO and NO₂ together (leading to a guidance value for
NO_x) seems the best way of evaluating the impact of NO. However,
future research should evaluate the specific phytotoxic properties of
the individual compounds and their combinations.

It is not yet possible to discriminate in the critical level
assessment between separate types of vegetation, such as crops,
plantation forests, natural forests and other natural vegetation. A
1-h average for NO₂ of 800 µg/m³ to prevent acute damage
(Table 24) is probably too high. A critical level for NO_x of around
300 µg/m³ would be better. A critical level of 95 µg/m³ as a 4-h
mean, as proposed in the previous WHO guidelines (WHO, 1987), is still
realistic, but not very practical. If critical levels for short
periods (e.g., 1 or 8 h) should be defined, it is probably necessary
to separate day- and night-time exposures. A critical level for a
24-h mean is more practical, as this is relevant for both peak
concentrations of several hours and air pollution episodes of several
days.

For the growing season and winter half year, Guderian (1988)
suggested critical levels of 60 and 40 µg/m³, respectively. From
Table 20 it can be seen that the critical level of 60 µg/m³ can cause
substantial growth stimulation rather than reduction. Within the
context of air quality guidelines, this type of response must be
regarded as potentially adverse because, for instance, of its
influence on competition within the natural vegetation. From current
knowledge it is evident that 60 µg/m³ is too high to prevent growth
stimulation. In addition, the critical level of 30 µg/m³ for an
annual mean, given in the 1987 WHO guidelines, will almost certainly
not protect all plant species. However, for crops, where growth
stimulation is rarely an adverse effect, this could be acceptable if
secondary effects are negligible. The experimental basis for the WHO
air quality guidelines of 1987 was relatively poor, but evidence is
increasing that these are certainly not unrealistically low. Not even

all direct adverse effects are eliminated by these levels (Adaros et
al., 1991a,b; Bender et al., 1991; Ashenden et al., 1993). Thus, the
updated guidelines/guidance values should be lower than the ones of
1987.

A long-term critical level for NO₂ of 10 µg/m³, especially to
avoid eutrophication of nutrient-poor vegetation, was proposed by
Guderian (1988) and Zierock et al. (1986). The basis for this
proposal was the work of Lee et al. (1985) and Press et al. (1986),
who found reduced growth of Sphagnum cuspidatum in regions with an
annual mean concentration of 38 and 11 µg/m³, respectively, compared
to the growth in another region with 4 µg/m³ after 140 days of
exposure. However, Lee et al. (1985) also showed that the poor growth
of Sphagnum was more closely related to the excessively high
concentrations of nitrate and ammonium ions in bog water rather than
to the concentration of NO₂ alone. Thus, this information could well
be used to assess water quality guidelines, but is not very useful as
a basis for air quality guidelines.

4.1.8.1 Representativity of the data

Critical levels for adverse effects of NH₃ on plants were
estimated using the model of Kooijman (Van der Eerden et al., 1991).
To protect 95% of the species at P < 0.05, a 24-h critical level of
270 and an annual mean critical level of 8 µg/m³ were estimated.
With the graphical approach the 24-h average was a little lower and
the annual mean somewhat higher (13 and 200 µg/m³, respectively;
Fig. 21).

On the basis of a review by Cape (1994), critical levels for H⁺
and NH₄⁺ were adopted for locations where ground-level cloud is
present for more than 10% of the time and where calcium and magnesium
concentrations in rain or cloud do not exceed H⁺ and NH₄⁺
concentrations (mainly high elevation areas in cold climate zones):
300 µmol NH₄⁺/litre as an annual mean (UNECE, 1994).

There remains considerable deficiency in the amount and scope of
experimentally derived information on which to base air quality
guidelines. This results from the fact that most experiments have
been performed under conditions that cannot directly be compared to
outdoor circumstances. In most experiments, only primary effects such
as photosynthesis and biomass production were evaluated, and rarely
secondary effects such as alteration of stress tolerance or
competitive ability. The plant species chosen in most experiments
were crops, although evidence suggests that some native species are
relatively more sensitive. For instance, lower plants such as
bryophytes and lichens are not protected by a waxy waterproof cuticle
and do not have the potential to close stomata. Furthermore, Pearson
& Stewart (1993) suggested that plants species from nutrient-poor
acidic soils are more sensitive.

Further work, employing low concentrations of NH_y and NO_x
(especially NO) in different climates, is urgently required. It is
not realistic to screen for all likely growth and physico-chemical
effects in the majority of species in order to arrive at general
effect thresholds. Selections must be made on the basis of an
understanding of differences in sensitivity between species. However,
an obvious mechanistic explanation for sensitivity differences is not
yet available. For instance, there appears to be no relationship
between the sensitivity to NO₂ and the nitrogen preference
(Ellenberg, 1985; Ashenden et al., 1993). Sensitivity classifications
for some tens of species have been made for NO₂ and NH₃ (e.g. US
EPA, 1978; Taylor et al., 1987), but it appears difficult to extend
predicitions very far beyond those examined. The hypotheses of Raven
(1988) and Pearson & Stewart (1993) should be studied in more detail
in laboratory experiments and field studies, as they could result in
an efficient selection criterium for further screening.

An attempt to determine the global situation regarding
nitrogen-containing compounds is being made. The assumption that all
deposited nitrogen-containing compounds (which is part of the critical
load concept) act additionally in their impact on vegetation is poorly
based on experimental results and is probably not valid for the short
term.

Generalizations and simplifications have to be made to arrive at
conclusions that are applicable in environmental policy making, but
this should be done with great care. Mechanistic simulation models
can become a powerful tool for making general predictions on the basis
of various air pollution experiments (Van de Geijn et al., 1993).
However, sufficient knowledge of biochemical and physiological
mechanisms to incorporate the impact of air pollution on vegetation
into these models is still lacking. This applies especially to
natural vegetation where stress sensitivity and competition are key
factors.

Many gaps in understanding the impact of nitrogen-containing air
pollution on vegetation still exist, and this is a good reason to use
a safety factor in determining critical levels and loads. However,
currently there is still no broadly accepted approach to quantify such
a safety factor.

4.1.9 General conclusions

The sum of information on gaseous NH₃ and on NH₄⁺ in wet and
occult deposition is still too limited to arrive at air quality
guidelines, as they should have a broad applicability. The critical
levels for NH₃ and NH₄⁺ are probably only valid for temperate
oceanic climatic zones (see sections 4.1.7.3, 4.1.7.4 and 4.1.8).

In most studies with NO and NO₂, no significant effects were
found at levels below 100 µg/m³, but several relevant exceptions
exist. NO₂ altered the response to O₃ generally with a
less-than-additive interaction. In combination with SO₂, NO₂ acted
more-than-additively in most cases. With CO₂ and with NO, no
interaction and thus additivity were generally found. The lowest
effective concentration levels of NO₂ are about equal for NO₂ alone
and in combination with O₃ or SO₂, although, generally, at
concentrations near to its effect threshold NO₂ causes growth
stimulation if it is the only pollutant, while in combination with
SO₂ and/or O₃ it results in growth inhibition.

To include the impact of NO, a critical level for NO_x instead of
one for NO₂ is proposed, assuming that NO and NO₂ act in an additive
manner. A strong case can be made for the provision of critical
levels for short-term exposures, but currently there are insufficient
data to provide these with sufficient confidence. Current evidence
exists for a critical level of around 75 µg/m³ for NO_x as a 24-h
mean.

The critical level for NO_x (NO and NO₂, added in ppb and
expressed as NO₂ in µg/m³) is 30 µg/m³ as an annual mean. At
concentrations slightly above this critical level, growth stimulation
will be the dominant effect if the ambient conditions allow growth and
NO_x is the only pollutant. This stimulation may be combined with a
minor increase in sensitivity to biotic and abiotic stresses. In
cases where biomass production is a positive effect, e.g., in
agriculture and plantation forests, the critical level can be higher.
Current knowledge is insufficient to arrive at critical levels for
these systems.

The critical level can be converted into deposition quantities.
With deposition velocities of 3 and 0.3 mm/second for NO₂ and NO,
respectively (see section 3.2.2 and Table 5), the annual deposition
corresponding to a NO_x concentration of 30 µg/m³ is 4.8 kg/ha when
half of the NO_x is NO₂ and 8.3 kg/ha when all is NO₂. This
indicates that at a concentration near to its critical level the
contribution of NO_x to the nitrogen demand is negligible for
fertilized crops but not for natural vegetation (see section 4.2).

4.2 Effects on natural and semi-natural ecosystems

4.2.1 Effects on freshwater and intertidal ecosystems

In this section the effects of atmospheric nitrogen deposition
on freshwater and intertidal ecosystems are evaluated. The
effects of increased emissions of nitrogen compounds with respect to
eutrophication are examined in order to establish ecosystem guidelines

for nitrogen deposition. The ecological effects of nitrogen
deposition are reviewed for (i) shallow softwater lakes and (ii) lakes
and streams.

4.2.1.1 Effects of nitrogen deposition on shallow softwater lakes

In the lowlands of western Europe, soft water is often found on
sandy soil which is poor in calcium carbonate or almost devoid of it.
The water is poorly buffered and the concentrations of calcium in the
water layer are very low. The water bodies are shallow and fully
mixed, with periodically fluctuating water levels. They are mainly
fed by rain water and thus are oligotrophic. These softwater
ecosystems are characterized by plant communities from the
phytosociological alliance Littorellion (Schoof-van Pelt, 1973;
Wittig, 1982; Roelofs, 1986; Vöge, 1988; Arts, 1990). The stands of
these communities are characterized by the presence of rare and
endangered isoetids, such as Littorella uniflora, Lobelia dortmanna,
Isoetes lacustris, I. echinospora, Echinodorus species, Luronium
natans and many other softwater macrophytes. These softwater bodies
are now almost all within nature reserves and have become very rare in
western Europe. A strong decline in amphibians has also been observed
in these water bodies (Leuven et al., 1986).

The effects of nitrogen pollutants on these softwater bodies have
been intensively studied in the Netherlands both in field surveys and
experimental studies. Field observations on about 70 softwater bodies
(with well-developed isoetid vegetation in the 1950s) showed that the
water, in which these macrophytes were still abundant in the early
1980s, was poorly buffered (alkalinity of 50-500 µeq/litre), slightly
acidic (pH=5-6) and very poor in nitrogen (Roelofs, 1983; Arts et al.,
1990). The softwater sites where these plant species had disappeared
could be divided into two groups. In 12 of the 53 softwater sites,
eutrophication, resulting from nutrient-enriched water, seemed to be
the cause of the decline. In this group of non-acidified water
bodies, plant species, such as Myriophyllum alterniflorum, Lemna
minor or Riccia fluitans had become dominant. High concentrations
of phosphate and ammonium ions were measured in the sediment. In some
of the larger water bodies no macrophytes were found, as a result of
dense plankton bloom. In the second group of lakes and pools (41 out
of 53) another development had taken place: the isoetid species were
replaced by dense stands of Juncus bulbosus or aquatic mosses such
as Sphagnum cuspidatum or Drepanocladus fluitans. This clearly
indicates acidification of the water in recent decades, probably
caused by enhanced atmospheric deposition. In the same field study it
was shown that the nitrogen levels in the water were higher in
ecosystems where the natural vegetation had disappeared, compared with
ecosystems where the Littorellion stands were still present (Roelofs,
1983). This strongly suggests the detrimental effects of atmospheric
nitrogen deposition in these softwater lakes.

Several ecophysiological studies have revealed the importance of
(i) inorganic carbon status of the water as a result of intermediate
levels of alkalinity, and (ii) low nitrogen concentrations for the
growth of the endangered isoetid macrophytes. Furthermore, almost all
of the typical softwater plants had a relatively low potential growth
rate. Increased acidity and higher concentrations of ammonium ion in
the water clearly stimulated the development of Juncus bulbosus and
submerged mosses such as Sphagnum and Drepanocladus species
(Roelofs et al., 1984; Den Hartog, 1986). Cultivation experiments
confirmed that the nitrogen species involved (ammonium or nitrate
ions) differentially influenced the growth of the studied species of
water plants. Almost all of the characteristic softwater isoetids
developed better when nitrate was added instead of ammonium, whereas
Juncus bulbosus and aquatic mosses (Sphagnum & Drepanocladus) were
clearly stimulated by ammonium (Schuurkes et al., 1986). The
importance of ammonium for the growth of these aquatic mosses was also
reported by Glime (1992).

The effects of atmospheric deposition have been studied in
small-scale softwater systems during a 2-year treatment with different
artificial rainwaters. Acidification, without airborne nitrogen input
(using sulfuric acid), did not result in a mass growth of Juncus
bulbosus, and a diverse isoetid vegetation remained present.
However, after increasing the nitrogen concentration in the
precipitation (as ammonium sulfate), similar changes to those seen in
field conditions were observed, i.e. a dramatic increase in the
dominance of Juncus bulbosus, of submerged aquatic mosses and of
Agrostic canina (Schuurkes et al., 1987). It became obvious that
the observed changes occurred because of the effects of ammonium
sulfate deposition, leading to both eutrophication and acidification.
The increased levels of ammonium in the system directly stimulated the
growth of plants such as Juncus bulbosus, whereas the surplus
ammonium would be nitrified in this water (pH > 4.0). During this
nitrification process, H⁺ ions are produced, which increases the
acidity of the system. The results of this study clearly demonstrated
that the changes in composition of the vegetation had already occurred
after a 2-year treatment with > 19 kg nitrogen per ha per year. A
reliable critical load for nitrogen deposition in these shallow
softwater lakes is thus most likely to be below 19 kg nitrogen per ha
per year and probably between 5 to 10 kg nitrogen per ha per year.
This value is supported by the observation that the greatest decline
in the species composition of the Dutch Litorellion communities has
coincided with nitrogen loads of around 10-13 kg nitrogen per ha per
year (Arts, 1990).

4.2.1.2 Effects of nitrogen deposition on lakes and streams

There is ample evidence that an increase of acidic and
acidifying compounds in atmospheric deposition had resulted in recent
acidification of lakes and streams in geologically sensitive regions

of Scandinavia, western Europe, Canada and the USA (Hultberg, 1988;
Muniz, 1991). This acidification is characterized by a decrease in pH
and acid neutralizing capacity and by increases in concentrations of
sulfate, aluminium, and sometimes nitrate and ammonium. It has been
shown since the 1970s, using various approaches (field surveys,
laboratory studies, whole-lake experiments), that these changes have
had major consequences for plant and animal species (macrofauna,
fishes) and for the functioning of these aquatic ecosystems.

The critical loads of acidity (from SO_y and NO_y) for aquatic
ecosystems, based on steady-state water chemistry models, were
published by the UN Economic Commission for Europe (UNECE) in 1988 and
1992. These models incorporate both sulfur and nitrogen acidity, and
critical loads are calculated on the basis of: (i) base cation
deposition; (ii) internal alkalinity production or base cation
concentrations; and (iii) nitrate leaching from the water system. The
calculated critical loads are thus site-specific (sensitive areas or
not) and also depend on the local hydrology and precipitation (for
full details, see Hultberg (1988), Henriksen (1988) and Kämäri et al.
(1992)). The critical loads of nitrogen acidity (kg nitrogen per ha
per year) for the most sensitive lakes and streams are:

Scandinavian 1.4-4.2 (Hultberg, 1988; Henriksen,
waters 1988; Kämäri et al., 1992)

Alpine lakes 3.5-6.1 (Marchetto et al., 1994)

Humic moorland 3.5-4.5 (Schuurkes et al., 1987;
pools van Dam & Buskens, 1993)

In many areas with high water alkalinity and/or high base cation
deposition, the values of the critical load for nitrogen acidity are
much higher than those for sensitive freshwaters. At present, the
possible effects of nitrogen eutrophication by ammonia/ammonium or
nitrate deposition are not incorporated in the establishment of
critical loads for nitrogen, except for shallow softwater lakes (see
section 4.2.1.1). This is because primary production in almost all
aquatic ecosystems is limited by phosphorus availability, and thus
nitrogen enrichment has been considered unimportant in this respect
(Moss, 1988). This certainly holds for those aquatic ecosystems
considered above, where the critical load with respect to acidifying
effects are certainly more relevant than the effects of nitrogen
eutrophication. It is, however, to be expected that the following
aquatic ecosystems are sensitive to nitrogen enrichment: (i) alpine
lakes; (ii) water with low background nitrogen; and (iii) humic lakes
(Kämäri et al., 1992). The effects of nitrogen eutrophication
(including ammonia/ammonium) in these water bodies need further
research and should be taken into account in future critical loads
determinations for nitrogen. At present it is not possible to present

reliable critical loads for nitrogen eutrophication in these aquatic
ecosystems. An overview of critical loads for nitrogen in aquatic
ecosystems is given in section 8.2.2.

4.2.2 Effects on ombrotrophic bogs and wetlands

In this section the effects of atmospheric nitrogen deposition in
(semi-)natural wetlands are evaluated. The effects of enhanced
nitrogen inputs are considered for: (i) ombrotrophic (raised) bogs;
(ii) fens; and (iii) intertidal fresh- and saltwater marshes. A
common feature of all these systems is the anaerobic nature of their
waterlogged soils, characterized by low redox potential, high
concentrations of toxic reduced substances and high rates of
denitrification (Gambrell & Patrick, 1978; Schlesinger, 1991).

4.2.2.1 Effects on ombrotrophic (raised) bogs

Ombrotrophic ("rain-nourished") bogs, which receive all their
nutrients from the atmosphere, are particularly sensitive to airborne
nitrogen loads. These bogs are systems of acidic wet areas and are
very common in the boreal and temperate parts of Europe. Because of
the anaerobic conditions, decomposition rates are slow, favouring the
development of peat. In western Europe and high northern latitudes,
typical plant species include bog-mosses ( Sphagnum species), sedges
(Carex; Eriophorum) and heathers ( Andromeda, Calluna and Erica).
Insectivorous plant species (e.g., Drosera) are especially
characteristic of these bogs; they compensate for low nitrogen
concentrations by trapping and digesting insects (Ellenberg, 1988b).

Clear effects of nitrogen eutrophication have been observed in
Dutch ombrotrophic bogs. The composition of the moss layer in the
small remnants of the formerly large bog areas has markedly changed in
recent decades as nitrogen loads have increased to 20-40 kg nitrogen
per ha per year (especially as NH_4⁺/NH₃). The most characteristic
species (Sphagnum) are replaced by the more nitrophilous mosses
(Greven, 1992). A national survey of Danish ombrotrophic bogs has
shown a decline of the original bog vegetation together with an
increase of more nitrogen-dependent species in areas with high ammonia
deposition (> 25 kg ammonium nitrogen per ha per year (Aaby, 1990).

The effects of atmospheric nitrogen deposition on ombrotrophic
bogs have also been intensively studied in the United Kingdom (Lee et
al., 1989; Lee & Studholme, 1992). Many characteristic Sphagnum
species are now largely absent from affected ombrotrophic bog areas
in the United Kingdom, such as the southern Pennines in England.
Atmospheric nitrogen deposition has more than doubled in these areas
to around 30 kg nitrogen per ha per year, compared with areas of
healthy Sphagnum growth. In contrast to the situation in
continental western Europe, most of the nitrogen deposition in the
United Kingdom is of nitrogen oxides, although the importance of

ammonia/ammonium deposition is also increasing in the United Kingdom
(Fowler et al., 1980; Sutton et al., 1993). Several studies on bogs
in the United Kingdom have shown that increased supplies of nitrogen
are rapidly absorbed and utilized by bog-mosses (Sphagnum),
reflecting the importance of nitrogen as a nutrient and its scarcity
in unpolluted regions (Woodin et al., 1985; Woodin & Lee, 1987). The
high nitrogen loadings are, however, supraoptimal for the growth of
many characteristic Sphagnum species, as demonstrated by restricted
development in growth experiments and transplantation studies between
clean and polluted locations. In areas with high nitrogen loads, such
as the Pennines, the growth of Sphagnum is in general less than in
unpolluted areas (Lee & Studholme, 1992). After transplantation of
Sphagnum from an unpolluted site to a bog in the southern Pennines,
a rapid increase in plant nitrogen content from around 12 to 20 mg/g
dry weight was observed (Press et al., 1986). A large increase in
arginine in the shoots of these bog-mosses was also found after
application of nitrogen. In field experiments in northern and
southern parts of Sweden, nitrogen was found to be the limiting factor
for the growth of Sphagnum. However, other nutrients, especially
phosphorus, may become secondarily limiting to plant growth when
nitrogen inputs reach a threshold (Aerts et al., 1992).

A further possible effect of the increased nitrogen content of
Sphagnum is an increased decay rate of the peat, as nitrogen content
strongly influences decomposition rates (Swift et al., 1979). The
decay rate of Sphagnum peat in Swedish ombrotrophic bogs has been
studied along a gradient of nitrogen deposition (Hogg et al., 1994).
The results of this short-term decay experiment indicated that the
decomposition rate of Sphagnum peat is more influenced by the
phosphorus content of the material than by its nitrogen content,
although some relation with nitrogen supply has been observed.
Further evidence is necessary to evaluate the long-term effects of
enhanced nitrogen supply on the decay of peat.

All these studies strongly indicate the detrimental effects of
atmospheric nitrogen on the development of the bog-forming Sphagnum
species. However, enhanced nitrogen deposition can influence the
competitive relationships in nutrient-deficient vegetation such as
bogs. The effects of the supply of extra nitrogen on the population
ecology of Drosera rotundifolia has been recently studied in a
4-year experiment in Swedish ombrotrophic bogs (Redbo-Torstensson,
1994). It was demonstrated that experimental applications of more
than 10 kg nitrogen (as NH₄NO₃) per ha per year clearly affected the
population of this insectivorous species: the establishment of new
individuals and the survival of the plants significantly decreased in
the vegetation treated with extra nitrogen. This decrease in the
total population density of the characteristic bog species Drosera
was not caused by toxic effects of nitrogen, but by enhanced

competition for light with tall species such as Eriophorum and
Andromeda, which responded positively to the increased nitrogen
inputs.

On the basis of the United Kingdom and Scandinavian studies, it
has become clear that increased nitrogen loads strongly affect
ombrotrophic bog ecosystems, especially because of the high nitrogen
retention capacity and closed nitrogen cycling. The growth of
bog-mosses is reduced, directly by nitrogen and indirectly by a
changed competitive relationship between the prostrate dominants
(e.g. Eriophorum) and the subordinate plant species. A reliable
critical load for nitrogen in these ombrotrophic bogs is 5-10 kg
nitrogen per ha per year, although additional long-term studies with
enhanced nitrogen (both nitrogen oxides and ammonia/ammonium) are
necessary to validate this figure.

4.2.2.2 Effects on mesotrophic fens

Fens are wetland ecosystems that are typical of alkaline to
slightly acidic habitats in many countries. The alkalinity is due to
groundwater draining from surrounding rocks or sediments that are
relatively rich in calcium carbonate. Most of these fen ecosystems
are characterized by rare and endangered plants species. The effects
of nitrogen enrichment upon mesotrophic fens have been intensively
studied in the Netherlands (Verhoeven & Schmitz 1991; Koerselman &
Verhoeven, 1992). These fen ecosystems are characterised by many
Carex species and are rich in forbs (e.g., Pedicularis palustris;
orchids). Most of these Dutch fen ecosystems are managed as hay
meadows, with removal of the plant material further restricting
nutrient levels, and are now nature reserves.

A considerable increase of tall graminoids (grass or Carex
species), with a somewhat higher potential growth rate, was observed
after experimentally adding nitrogen to three Dutch fen ecosystems
(Vermeer, 1986; Verhoeven & Schmitz, 1991). This increase caused a
significant decrease in the diversity of subordinate plant species.
In one of the Dutch fen sites investigated, which had a long history
of hay making, it has been shown that phosphorus deficiency was also a
major factor in the productivity of the system, since there was a high
output of phosphorus from the ecosystem with the hay (Verhoeven &
Schmitz, 1991; Koerselman & Verhoeven, 1992). Using the results of
fertilization trials and nutrient budget studies in these fen
ecosystems (Koerselman et al., 1990; Koerselman & Verhoeven, 1992),
with their relatively closed nitrogen cycle, it seems reasonable to
establish a critical load of 20-35 kg nitrogen per ha per year, based
upon the output of the nitrogen from the fen system via normal
management. In some fen ecosystems, the critical nitrogen load based
on the change in diversity may be substantially higher, because of the
limitation of productivity by phosphorus (Egloff, 1987; Verhoeven &
Schmitz, 1991). In this situation, however, the risks of nitrogen

losses to surface water or groundwater will increase because of
phosphorus limitation, and this effect should be taken into account in
critical load evaluation. High rates of denitrification could also
influence the establishment of critical loads for these fen
ecosystems, and this aspect needs further investigation.

4.2.2.3 Effects on fresh- and saltwater marshes

In the wetland ecosystems discussed above, the nitrogen cycle is
more closed than that of intertidal marshes. The data on atmospheric
nitrogen inputs to the nitrogen cycling in intertidal fresh- and
saltwater marshes (with large prostrate graminoids as species of
Spartina, Typha and Carex) have been reviewed by Morris (1991).
It has become evident that nitrogen inputs to these marsh ecosystems
via surface water (well above 100 kg nitrogen per ha per year) are
much higher than the atmospheric loading. In non-tidal freshwater
marshes, it has been demonstrated in many studies that denitrification
is very high with a very large output of nitrogen from the ecosystem
(Morris, 1991). Because of the combined effect of these processes,
atmospheric nitrogen deposition is of only minor importance for these
marshes, and it is not useful to establish a critical load for
airborne nitrogen to these systems. In his review Morris (1991)
formulated a critical load for atmospheric nitrogen in wetland
ecosystems of around 20 kg nitrogen per ha per year. It is more
appropriate to make a distinction for different types of wetlands, as
shown above. An overview of the critical loads for wetlands is given
in chapter 8.

4.2.3 Effects on species-rich grasslands

Almost all of the research on the effects of atmospheric
deposition on terrestrial vegetation has focused on ecosystems
(e.g., forest, heathland and bogs) involving poorly buffered acidic
soils. Semi-natural grasslands with traditional agricultural use have
also been an important part of the landscape in western and central
Europe, and contain, or used to contain, many rare and endangered
plant and animal species. A number of these grasslands have been set
aside as nature reserves in several European countries (Ellenberg,
1988b; Woodin & Farmer, 1993). These semi-natural grasslands, which
are of conservation interest, are generally nutrient-poor because of
long agricultural use with low levels of manure and the removal
of plant growth by grazing or hay making. The vegetation is
characterized by many low growing species and is of nutrient-poor soil
status (Ellenberg, 1988b). Although these grasslands are nowadays
rare, the proportion of endangered plant and animal species in these
communities is very high (Van Dijk, 1992). Many experiments have
shown that application of artificial fertilizer (nitrogen, phosphorus
and potassium) changes these grasslands into tall, species-poor
stands, dominated by a few highly productive crop grasses (Van Den
Bergh, 1979; Willems, 1980; Van Hecke et al., 1981). To maintain a

large diversity of species, addition of fertilizer has to be avoided.
It is thus to be expected that these species-rich grasslands will be
affected by increased atmospheric nitrogen input (Wellburn, 1988;
Liljelund & Torstensson, 1988; Ellenberg, 1988b).

Many semi-natural grassland types are present in western and
central Europe. Most of these grasslands belong to the so-called
neutral grasslands (Molinio-Arrhenateretea; moist to moderately dry),
to the dry calcareous grasslands (Festuca-Brometea) or to the acid
grasslands on very poor soils (Nardetalia). Detailed descriptions
have been given by Ellenberg (1988b) and Van Dijk (1992). To obtain
critical loads for nitrogen for all these grasslands, it would be
essential to study the effects of nitrogen eutrophication in a
representative range within these communities. Detailed data are,
however, scarce. Therefore, the results of an integrated research
programme on nitrogen eutrophication in Dutch calcareous grasslands
are used as a target study in this chapter to obtain (i) information
on observed changes in this type of grassland caused by enhanced
nitrogen input, and (ii) a reliable estimation of the critical load
for nitrogen in these well-buffered non-acidic grasslands. The
results of this calcareous grassland study will be discussed in
respect to other semi-natural grasslands.

4.2.3.1 Effects of nitrogen on calcareous grasslands

Calcareous grasslands are communities on limestone. The subsoils
consist of various kinds of limestone with high contents of calcium
carbonate (> 90%), covered by shallow well-buffered rendzina soils
(A/C-profiles; pH of the top soil is 7-8 with a calcium carbonate
content of around 10%). The depth of the soil varies between 10 and
50 cm and the availability of nitrogen and phosphorus is low. The
grasslands are generally found on slopes with limestone in the subsoil
and a deep groundwater table. Plant productivity is low, and the peak
standing crop is in general between 150 and 400 g/m². The canopy of
the vegetation is open and low (10-20 cm). Calcareous grasslands are
among the most species-rich plant communities in Europe and contain a
large number of rare and endangered species. The area of these
semi-natural grasslands has decreased substantially in Europe during
the second half of this century (Wolkinger & Plank, 1981; Ratcliffe,
1984). Some remnants have become nature reserves in several European
countries. To maintain the characteristic calcareous vegetation a
specific management is needed to prevent their natural succession
towards woodland (Wells, 1974; Dierschke, 1985). The calcareous
grasslands in the Netherlands are mown in autumn with removal of the
hay (Bobbink & Willems, 1987).

a) Nitrogen enrichment and vegetation composition

The effects of nitrogen enrichment in Dutch calcareous grasslands
on vegetation composition have been investigated in two field

experiments (Bobbink et al., 1988; Bobbink, 1991). Either potassium
(100 kg per ha per year), phosphorus (30 kg per ha per year) or
nitrogen (100 kg per ha per year), as well as a complete fertilization
(nitrogen, phosphorus and potassium), were applied for 3 years to
study the long-term effects on vegetation composition. Nitrogen was
given as ammonium nitrate and was applied to two nature reserves with
opposite aspects (north and south). Total above-ground biomass
increased considerably, as expected, after three years of nitrogen,
phosphorus and potassium fertilization. In the experiments where
the nutrients were applied individually, a moderate increase in
above-ground dry weight was only seen with nitrogen addition: (about
330 g/m² compared with about 210 g/m² in the untreated plots). The
dry weight distribution of the species was significantly affected by
nutrient treatments. In the nitrogen-treated vegetation, the dry
weight of the grass species Brachypodium pinnatum was about 3 times
higher than in the control plots. Nitrogen application also resulted
in a drastic reduction of the biomass of forb species (including
several Dutch Red List species) and of the total number of species.
The observed decrease in species diversity in the nitrogen-treated
vegetation was not caused by nitrogen toxicity, but by the change in
vertical structure of the grassland vegetation. The growth of
Brachypodium was strongly stimulated and its overtopping leaves
reduced the light within the vegetation. It overshadowed the other
characteristic species and growth of these species declined rapidly
(Bobbink et al., 1988; Bobbink, 1991). Stimulation of Brachypodium
growth and a substantial reduction in species diversity were observed
following application of nitrogen to a 5-year permanent plot study
using a factorial design (Willems et al., 1993).

Many characteristic lichens and mosses have also disappeared in
recent years from European calcareous grasslands (During & Willen,
1986). This has been caused partly by the indirect effects of extra
nitrogen inputs, as shown experimentally by Van Tooren et al. (1990).
Data on the effects of nitrogen eutrophication on the species-rich
fauna of calcareous grassland are not available. However, it is very
likely that the diversity of animals, especially of insects, will also
be reduced when tall grasses are strongly dominating the vegetation,
because of the decreasing abundance of many herbaceous flowering
species which act as host or forage plants.

b) Nitrogen enrichment and nutrient storage in calcareous grasslands

The seasonal distribution of nutrients after nitrogen
fertilization in spring (120 kg nitrogen as ammonium nitrate) has been
studied with the repeated harvest approach (Bobbink et al., 1989).
It resulted in a significantly increased peak standing crop of
Brachypodium . This grass proves to have very efficient nitrogen
uptake and very efficient withdrawal from its senescent shoots into
its well-developed rhizome system. Brachypodium can benefit from the
extra nitrogen redistributed to the below-ground rhizomes by enhanced

growth in the next spring. The distribution of nitrogen has also been
quantified in 3-year fertilization experiments. Brachypodium
greatly monopolized (> 75%) the nitrogen storage in both the
above-ground and below-ground compartments of the vegetation with
increasing nitrogen availability (Bobbink et al., 1988; Bobbink,
1991).

Nitrogen cycling and accumulation in calcareous grassland can be
significantly influenced by two major outputs from the system:
(i) leaching from the soil; and (ii) removal with management regimes.
Nitrogen losses by denitrification in dry calcareous grasslands are
low (< 1 kg nitrogen per ha per year), owing to the well-drained soil
conditions (Mosier et al., 1981). Ammonium and nitrate leaching has
been studied in Dutch calcareous grasslands by Van Dam et al. (1992).
Both the water fluxes and the solute fluxes at different soil depths
have been measured over 2 years in untreated vegetation and in
calcareous grassland vegetation sprayed at 2-weekly intervals with
ammonium sulfate (50 kg nitrogen per ha per year). The nitrogen
leaching from the untreated vegetation was very low (0.7 kg nitrogen
per ha per year) and amounted to only 2% of the total atmospheric
deposition of nitrogen. After the spraying with ammonium sulfate,
nitrogen leaching increased significantly to 3.5 kg nitrogen per ha
per year, although this figure was also a very small proportion (4%)
of the nitrogen input in this vegetation (Van Dam et al., 1992). It
is thus evident that calcareous grassland ecosystems retain nitrogen
almost completely in the system. This is caused by a combination of
enhanced plant uptake (Bobbink et al., 1988; Bobbink, 1991) and
increased immobilization in the soil organic matter (Van Dam et al.,
1992).

4.2.3.2 Critical loads for nitrogen in calcareous grasslands

The most important output of nitrogen from calcareous grassland
is via exploitation or management. The annual nitrogen removal in the
hay varies slightly between years and sites, but in general between
17 and 22 kg nitrogen per ha is removed from the system under normal
management conditions in the Netherlands (Bobbink, 1991; Bobbink &
Willems, 1991). The ambient nitrogen deposition in Dutch calcareous
grasslands, as determined by Van Dam (1990), is high (35-40 kg
nitrogen per ha per year) and is nowadays the major nitrogen input to
the system. Legume species (Leguminosae) also occur in calcareous
vegetation, and form an additional nitrogen input owing to the
nitrogen-fixing microorganisms in their root nodules (about 5 kg
nitrogen per ha per year).

The nitrogen mass balance of Dutch calcareous grasslands is
summarized in Table 25. It is obvious that calcareous grasslands now
significantly accumulate nitrogen (16-26 kg per ha per year) in the
Netherlands. A critical nitrogen load has been determined with a mass
balance model, because of the lack of long-term addition experiments

with low nitrogen loads. Assuming a critical long-term immobilization
rate for nitrogen of 0-6 kg nitrogen per ha per year, the critical
nitrogen load can be derived by adding the nitrogen fluxes due to net
uptake, denitrification and leaching, corrected for the nitrogen input
by fixation. In this way, 15-25 kg nitrogen per ha per year is
considered as nitrogen critical load for this ecosystem. Nitrogen
cycling within the system (between plants and soil) is not used for
this calculation.

Table 25. Nitrogen mass balance (kg nitrogen per ha per year)
for dry calcareous grassland in the Netherlands

Input Output

Atmospheric deposition 35-40 Harvest 17-22
Nitrogen fixation 5 Denitrification 1
Soil leaching 1

Total 40-45 Total 19-24

In calcareous grassland in England, addition of nitrogen
stimulated the dominance of grasses in most cases (Smith et al., 1971;
Jeffrey & Pigott, 1973). In these studies, the application of
50-100 kg nitrogen per ha per year resulted in a strong dominance of
the grasses Festuca rubra, F. ovina or Agrostis stolonifera.
However, Brachypodium and Bromus erectus, the most frequent
species in calcareous grassland in continental Europe, were absent
from these sites. Following a survey of data from a number of
conservation sites in southern England, Pitcairn et al. (1991)
concluded that Brachypodium had expanded in the United Kingdom
during the last 100 years. They considered that much of the early
spread could be attributed to a decline in grazing pressure but that
the more recent spread had, in some cases, taken place despite grazing
or mowing, and could be related to nitrogen inputs. However, a study
of chalk grassland at Parsonage Downs (United Kingdom) showed no
substantial change in species composition over the twenty years
between 1970 and 1990, a period when nitrogen deposition is thought to
have increased significantly (Wells et al., 1993). Brachypodium was
present in the sward but had not expanded as in the Dutch grasslands.
In a linked experimental study, applications of nitrogen to eight
forbs and one grass (Brachypodium) at levels of 20, 40 and 80 kg
nitrogen per ha per year for two years did not result in
Brachypodium becoming dominant.

Apart from the Dutch studies, the effects of enhanced nitrogen
inputs have been little investigated in continental European
calcareous grasslands. Some data from a recent fertilization

experiment at the alvar grasslands, a thin-soiled vegetation over flat
limestone, on the Swedish island Öland, suggest that the vegetation
hardly responds to applications of nitrogen or phosphorus (Sykes & Van
der Maarel, 1991; personal communication by Van der Maarel). Only
irrigation in combination with nutrients has caused an increase in
grasses. This is probably due to the low annual precipitation in this
area (400-500 mm).

Increased nitrogen availability is probably of major importance
in many European calcareous grasslands. An increased availability of
nitrogen is indicated by enhanced growth of some tall grasses,
especially stress-tolerant species, which have a slightly higher
potential growth rate and efficient nitrogen utilization. It clearly
depends on the original species composition, as to which of the
grass species will increase following enhanced nitrogen inputs.
Furthermore, the difference between the Dutch and United Kingdom
results may reflect differences in management; the impacts of grazing
in the United Kingdom grasslands could offset any competitive
advantage the grasses may have obtained from additional nitrogen
inputs. The critical load for nitrogen in these calcareous grasslands
could be influenced by management; long-term studies involving
additional nitrogen input with various management schemes are needed
to quantify these aspects.

4.2.3.3 Comparison with other semi-natural grasslands

Productivity in grasslands is strongly influenced by nutrients,
as shown in many agricultural studies (e.g. Chapin, 1980). It is also
well-known that large amounts of fertilizer (nitrogen, phosphorus and
potassium) alter almost all grassland types into tall, species-poor
swards dominated by a few highly productive crop grasses (e.g. Bakelaar
& Odum, 1978; Van Den Bergh, 1979; Willems, 1980; Van Hecke et al.,
1981). Most of these species-rich grasslands are deficient in
nitrogen or phosphorous, and thus characterized by plant species of
nutrient-poor habitats. It is thus likely that these grasslands are
sensitive to nitrogen eutrophication from the atmosphere (Wellburn,
1988; Ellenberg, 1988b). Thus, it is also important to establish
critical loads for nitrogen in the species-rich grasslands, although
data from experiments with nitrogen application in these semi-natural
grasslands are scarce.

Increased nitrogen availability can also affect other
semi-natural grasslands, although experimental evidence is quite
scarce. A classical study into the effects of nutrients on neutral
grasslands is the Park Grass experiment at Rothamsted, England, which
has been running since 1856 (Williams, 1978). After application of
nitrogen as ammonium sulfate or sodium nitrate (48 kg nitrogen per ha
per year), the vegetation became very poor in species and dominated by
grasses such as Holcus lanatus or Agrostis sp. The effects of
nutrients in dry and wet dune grasslands (1% calcium carbonate) on

sandy soils have been studied at Braunton Burrows (Devon, England) by
Willis (1963). Nutrients were applied over 2 years (6 × 40 kg
nitrogen per ha per year) using a factorial design for nitrogen and
phosphorus. Nitrogen proved to be the most important nutrient in
stimulating the growth of some grass species ( Festuca rubra and Poa
pratensis). This enhanced growth reduced significantly the abundance
of many small plants such as prostrate phanerogamic species, mosses
and lichens (Willis, 1963). In this coastal area with low nitrogen
deposition (currently about 10 kg nitrogen per ha per year) the
vegetation of dune grasslands is at present still species-rich,
whereas in many Dutch dune grasslands with higher nitrogen loading
(20-30 kg nitrogen per ha per year) certain grasses have increased and
it has become a problem to maintain diversity. Recent studies of the
response of mesothrophic grasslands in the United Kingdom have shown
that additions as small as 25 kg per ha per year can lead to changes
in species diversity after several years of fertilizer additions and
that changes take place more rapidly at higher rates of addition
(Mountford et al., 1994). This indicates that many of these
semi-natural grasslands are also sensitive to nitrogen eutrophication
and that the critical loads are likely to be of the same magnitude or
slightly higher (20-30 kg nitrogen per ha per year) than in calcareous
grasslands.

Many other semi-natural grassland types occur, especially in the
montane-subalpine regions, containing a large proportion of the
biodiversity of the area. However, data are too scarce to establish
reliable load for these grasslands, although it may be expected that:
(i) most of these grassland are sensitive to nitrogen; and (ii) the
critical load for nitrogen is probably lower than for lowland
(calcareous) grasslands. The presented critical loads for
species-rich grasslands are summarized in section 8.2.2.

4.2.4 Effects on heathlands

Various types of plant communities have been described as heath,
but the term is applied here to plant communities where the dominant
vegetation is small-leaved dwarf-shrubs forming a canopy of 1 m or
less above soil surface. Grasses and forbs may form discontinuous
strata, and there is frequently a ground layer of mosses or lichens
(Gimingham et al., 1979; De Smidt, 1979). Dwarf-shrub heathlands
occur in various parts of the world, especially in montane habitats,
but are widespread in the atlantic and sub-atlantic parts of Europe.
In these parts of the European continent, natural heathland is limited
to a narrow coastal zone. Inland lowland heathlands are man-made
(semi-natural), although they have existed for several centuries.
Lowland healths are widely dominated by the Ericaceae, especially
Calluna vulgaris in the dry heathlands and Erica tetralix in the
wet heathlands (Gimingham et al., 1979). In these heaths, development
towards woodland has been prevented by mowing, burning, sheep grazing
and sod removal.

Until the beginning of this century, the balance of nutrient
input and output was in equilibrium in the lowland heathlands of
western Europe (De Smidt, 1979; Gimingham & De Smidt, 1983). The
original land use implied a regular, periodic removal of nutrients
from the ecosystems via grazing and sod removal (Heil & Aerts, 1993).
Sod removal was practised less systematically in many Scandinavian and
Scottish heathlands (Gimingham & De Smidt, 1983). Here Calluna has
been renewed by burning at regular intervals, varying from 4-6 years
in southern Sweden to 15-20 years in western Norway (Nilsson, 1978;
Skogen, 1979). The original land use of the lowland heathland ceased
in the early 1900s and the area occupied by this community decreased
markedly all over its distribution area (Gimingham, 1972; De Smidt,
1979; Ellenberg, 1988b). Dwarf-shrub heathlands may be divided into
four categories according to broad differences in habitat: (1) dry
heathlands; (2) wet heathlands; (3) montane and (4) arctic-alpine
heathlands.

4.2.4.1 Effects on inland dry heathlands

During recent decades many lowland heathlands in western Europe
have become dominated by grass species. An evaluation, using aerial
photographs, has shown that more than 35% of Dutch heathland has been
altered into grassland (Van Kootwijk & Van der Voet, 1989). In recent
years, similar changes have been observed in SW Norway, which has the
largest local emission of ammonia as well as the heaviest nitrogen
input through long-range deposition in Norway (Anonymous, 1991). It
has been suggested that nitrogen eutrophication might be a significant
factor in this transition to grasslands. Field and laboratory
experiments confirm the importance of nutrients, especially in the
early phase of heathland development (Heil & Diemont, 1983; Roelofs
1986; Heil & Bruggink, 1987; Aerts et al., 1990). However, Calluna
can compete successfully with the grasses, even at high nitrogen
loading, if its canopy remains closed (Aerts et al., 1990). Apart
from the changes in competitive interactions between Calluna and the
grasses, heather beetle plagues and nitrogen accumulation in the soil
are important factors in the changing lowland heaths. Furthermore,
evidence is growing that frost sensitivity of the dominant
dwarf-shrubs may also be affected by increasing nitrogen inputs.

Heathland canopies have a strong filtering effect on air
pollutants, a varying canopy structure being an important factor. For
sulfur and nitrogen it has been shown that bulk deposition accounts
for only about 35-40% of total atmospheric input (Heil et al., 1987;
Bobbink et al., 1992b). Total atmospheric deposition of nitrogen is
30-45 kg nitrogen per ha per year in the heathland sites in the
eastern part of the Netherlands. More than 70% of the total nitrogen
input is deposited as ammonium or ammonia (Bobbink et al., 1992b;
Bobbink & Heil, 1993). Although data for nitrogen inputs in other

European lowland heathlands are missing, it is very likely that in
many European agricultural regions nitrogen deposition has increased
in recent years (Asman, 1987; Buijsman et al., 1987).

In Calluna heathland, outbreaks of the chrysomelid heather
beetle (Lochmaea suturalis) occur frequently. These beetles feed
exclusively on the green parts of Calluna. The closed Calluna
canopy is opened over large areas and the interception of light by
Calluna decreases strongly (Berdowski, 1987, 1993). Thus the growth
of the under-storey grasses ( Deschampsia or Molinia) is enhanced
significantly. In general insect grazing is affected by the nutritive
value of the plant material, and the nitrogen content is especially
important in this respect (Crawley, 1983). Experimental applications
of nitrogen to heathland vegetation cause the concentrations of this
element in the green parts of Calluna to increase (Heil & Bruggink,
1987; Bobbink & Heil, 1993). It is, therefore, very likely that the
frequency and intensity of heather beetle outbreaks are stimulated by
increased atmospheric nitrogen input in Dutch heathland. This
hypothesis is supported by the observations of Blankwaardt (1977), who
reported that from 1915 onwards heather beetle outbreaks were observed
in the Netherlands with an interval of about 20 years, whereas in the
last 15 years the outbreaks have occurred with a periodicity of less
than 8 years. It has also been observed that during a heather beetle
outbreak Calluna plants are more severely damaged in nitrogen-
fertilized vegetation (Heil & Diemont, 1983). In a rearing experiment
with larvae of the heather beetle, Brunsting & Heil (1985)
demonstrated that the growth of the larvae was increased by higher
nitrogen concentrations in the leaves of Calluna. Van der Eerden
et al. (1990) studied the effects of ammonium sulfate on the growth of
heather beetle after a outbreak of the beetle in vegetation
artificially sprayed under a cover. They found no significant effect
of the treatments on total number or on biomass of the first stage
larvae. However, the development of subsequent larval stages was
accelerated by the application of ammonium sulfate in the artificial
rain: the percentage of third stage larvae increased by 20%, compared
with larvae in the control treatment. Furthermore, heather beetle
larvae were put on Calluna shoots taken from plants which had been
fumigated with ammonia in open-top chambers (12 months; 4 to
105 µg/m³) (Van der Eerden et al., 1991). After 7 days the larvae
were counted and weighed. Both the mass and the development rate of
the larvae clearly increased with increasing concentrations of
ammonia. The heather beetle has recently been found in SW Norway and
it is expanding its territory. It is probably an important cause of
Calluna death in this region (Hansen, 1991). It can be concluded
that nitrogen inputs influence outbreaks of heather beetle, although
the exact relationship between both processes needs further research.

In the past Dutch inland heathlands were grazed by flocks of
sheep and sods were generally removed at intervals of 25-50 years
(De Smidt, 1979). Nowadays these heathlands are mostly managed by

mechanical sod removal, which can restore the heathland vegetation if
a seed bank of the original heathland species is still present
(Bruggink, 1993). The increase in organic matter and in the amounts
of nitrogen in the system during secondary succession is well known
(Begon et al., 1990). It was shown in the 1970s that during secondary
heathland succession the above-ground and below-ground biomass and the
amount of litter increase (Chapman et al., 1975; Gimingham et al.,
1979). It is likely that changes in nitrogen accumulation will have
occurred due to the increase in atmospheric deposition.

Berendse (1990) performed a detailed study on the accumulation of
organic matter and of nitrogen during the secondary succession after
sod removal in the Netherlands. He found a large increase in plant
biomass, soil organic matter and total nitrogen storage in the first
20 to 30 years after sod removal. Furthermore, it was demonstrated
that nitrogen mineralization was low during the first 10 years (about
10 kg nitrogen per ha per year), but increased considerably over the
next 20 years to 50-110 kg nitrogen per ha per year. Regression
analysis of the total nitrogen storage versus the years after sod
removal revealed an annual nitrogen increase in the system of about
33 kg nitrogen per ha per year (probably somewhat lower in the early
years and higher in later years) (Berendse, 1990). These values are
in good agreement with measured nitrogen deposition in Dutch
heathlands in the late 1980s (Bobbink et al., 1992b).

Thus, the organic matter in the soil increases rapidly after sod
removal, which removes almost all of the soil organic matter.
However, this process is accelerated by the enhanced dry matter
production and litter production of the dwarf shrubs caused by the
extra nitrogen inputs. Nitrogen accumulation in the system also
increases. Hardly any nitrogen disappears from the system because
nitrate leaching to deeper layers is only of minor importance in Dutch
heathlands, as shown by De Boer (1989) and Van Der Maas (1990).
Nitrogen availability from atmospheric inputs, in addition to
mineralization, is within a relatively short period (about 10 years)
high enough to stimulate the transition of heathland to grassland,
especially after the opening of the heather canopy by secondary
causes.

It has been demonstrated that frost sensitivity of some tree
species increases with increasing concentrations of air pollutants
(e.g. Aronsson, 1980; Dueck et al., 1991). This increase in frost
sensitivity is sometimes correlated with enhanced nitrogen
concentrations in the foliage of the trees. Long-term effects of air
pollutants on the frost sensitivity of Calluna and Erica are to be
expected because of (i) the evergreen growth form of these species and
(ii) the increasing content of nitrogen in the leaves of Calluna,
associated with increased nitrogen deposition in the Netherlands and
Norway (Heil & Bruggink, 1987; Hansen, 1991). It has been suggested
that damage to Calluna shoots in the successive severe winters of
the mid-1980s was at least partly caused by the increased frost

sensitivity. Investigations into the effects of air pollutants on the
frost sensitivity of heathland species outside the Netherlands started
in the early 1990s (Hansen, 1991; Uren, 1992).

The effects of sulfur dioxide, ammonium sulfate and ammonia upon
frost sensitivity in Calluna have been studied by Van der Eerden
et al. (1990). After fumigation with sulfur dioxide (90 µg/m³ for
3 months), increased frost injury in Calluna was only found at
temperatures that seldom occur in the Netherlands (lower than -20°C).
Fumigation with ammonia of Calluna plants in open-top chambers over
a 4-7 month period (100 µg/m³) revealed that frost sensitivity was
not affected in autumn (September or November), whereas in February,
just before growth started, frost injury increased significantly at
-12°C (Van der Eerden et al., 1991). These authors also studied the
frost sensitivity of Calluna vegetation sprayed with six different
levels of ammonium sulfate (3-91 kg nitrogen per ha per year). The
frost sensitivity increased slightly, although significantly, after
5 months in vegetation treated with the highest level of ammonium
sulfate (400 µmol/litre; 91 kg nitrogen per ha per year), compared
with the control treatments. However, frost sensitivity of Calluna
decreased again two months later and no significant effects of the
ammonium sulfate application upon frost hardiness were seen at that
time. Thus, high levels of ammonia or ammonium sulfate seem to
increase the frost sensitivity of Calluna plants, although the
significance of this phenomenon is still uncertain at ambient nitrogen
inputs. The relation between frost sensitivity and nitrogen input has
not yet been sufficiently quantified to use it for a precise
assessment of critical loads in this respect.

It has been shown above that atmospheric nitrogen is the trigger
for changes of lowland dry heathlands into grass swards in the
Netherlands. Lowland dry heathlands in the United Kingdom do not show
consistent patterns over the past 10 to 40 years. Pitcairn et al.
(1991) assessed changes in abundance of Calluna in three heaths in
East Anglia (eastern England) over recent decades. All three heaths
showed a decline in Calluna and an increase in grasses. The authors
concluded that increases in nitrogen deposition was at least partly
responsible for the changes, but also noted that the management had
changed. A wider assessment of heathlands in SE England showed that
in some cases Calluna had declined and subsequently been invaded by
grasses while other areas were still dominated by dwarf shrubs (Marrs,
1993). This clearly stresses the importance of management for the
maintenance of dwarf shrubs in heathlands. A simulation model, which
integrates processes such as atmospheric nitrogen input, heather
beetle outbreak, soil nitrogen accumulation, sod removal and
competition between species, has been used to establish the critical
loads of nitrogen deposition in lowland dry heathlands (Heil &
Bobbink, 1993a,b). The model has been calibrated with data from field
and laboratory experiments in the Netherlands. As an indicator of the
effects of atmospheric nitrogen, the proportion and increase of

grasses in the heathland system are used. Atmospheric nitrogen
deposition has varied between 5 and 75 kg nitrogen per ha per year in
steps of 5-10 kg nitrogen during different simulations. From these
simulations, the value for the critical load of nitrogen for the
changes from dwarf shrubs to grasses was 15-20 kg nitrogen per ha per
year.

4.2.4.2 Effects of nitrogen on inland wet heathlands

The western European lowland heathlands of wet habitats are
dominated by the dwarf shrub Erica tetralix (Ellenberg, 1988b) and
are generally richer in plant species than the dry heathlands. In
recent decades a drastic change in species composition of Dutch wet
heathlands has been observed. Nowadays, many wet heathlands that were
originally dominated by Erica have become monospecific stands of the
grass Molinia. Together with Erica almost all of the rare plant
species have disappeared from the system. It has been hypothesized
that this change has been caused by atmospheric nitrogen
eutrophication.

Competition experiments using micro-ecosystems have clearly shown
that Molinia is a better competitor than Erica at high nitrogen
availability. After 2 years of application of nitrogen (150 kg per ha
per year), the relative competitive strength of Molinia compared
with Erica doubled (Berendse & Aerts, 1984). A 3-year field
experiment with nitrogen application in Dutch lowland wet heathland
(around 160 kg nitrogen per ha per year) also indicated that Molinia
is able to outdo Erica at high nitrogen availability (Aerts &
Berendse, 1988). In contrast to the competitive relations between
Calluna and the grasses, Molinia can outdo Erica without opening
of the dwarf shrub canopy. This difference is caused by the lower
canopy of Erica (25-35 cm), compared with Calluna, and the tall
growth form of Molinia, which can overgrow and shade Erica if
enough nitrogen is available. It is in this respect also important
that heather beetle plagues do not occur in wet heathlands and that no
frost damage has been observed in this community.

It has been demonstrated that in many Dutch wet heathlands the
accumulation of litter and humus has led to increased nitrogen
mineralization (100-130 kg nitrogen per ha per year) (Berendse et al.,
1987). In the first 10 years after sod removal the annual nitrogen
mineralization is very low, but afterwards it increases rapidly. The
leaching of accumulated nitrogen from wet heathlands is extremely low
(Berendse, 1990). The observed nitrogen availabilities are high
enough to change Erica -dominated wet heathlands into monostands of
Molinia.

Berendse (1988) developed a wet heathland model to simulate
carbon and nitrogen dynamics during secondary succession. He
incorporated in this model the competitive relationships between

Erica and Molinia, the litter production from both species, soil
nitrogen accumulation and mineralization, leaching, atmospheric
nitrogen deposition and sheep grazing. He simulated the development
of lowland wet heathland after sod removal, because almost all of the
Dutch communities are already strongly dominated by Molinia and it
is impossible to expect changes in this situation without drastic
management. Using the biomass of Molinia with respect to Erica as
an indicator, his results suggested 17-22 kg nitrogen per ha per year
as the critical load for the transition of lowland wet heathland into
a grass-dominated sward (Berendse, 1988). The decrease in endangered
wet heathland forbs is partly caused by the overshading by Molinia,
but some species had already disappeared from wet heathlands before
the increase of Molinia started. The critical load for this decline
is probably lower than the given values and is discussed in section
4.2.4.4.

4.2.4.3 Effects of nitrogen on arctic and alpine heathlands

Semi-natural Calluna heathlands are found in the lowlands along
the Norwegian coast to 68°N and show distinct plant gradients in the
south-north direction, from coast to inland and from lowland to upland
areas (Fremstad et al., 1991). In central parts of western Norway the
plant composition changes at an altitude of about 400 m, above which
alpine species occur regularly in the heaths. At this altitude
oceanic upland Calluna and Erica heaths merge into alpine heaths,
which are naturally occurring, non-anthropogenic communities. Some
oligotrophic alpine heaths also contain Calluna, but most heaths in
Fennoscandia and in European parts of Russia are dominated by other
ericoid species ( Vaccinium spp., Empetrum nigrum s. lat.,
Arctostaphylos spp., Loiseleuria procumbens, Phyllodoce caerulea,
Betula nana, Juniperus communis and Salix spp.). Many heath types
have a more or less continuous layer of mosses and lichens. Related
heaths are found in alpine regions in the British Isles, in Iceland,
in southernmost Greenland, in northern Russia, and on siliceous rocks
in the Alps (Grabherr, 1979; Elvebakk, 1985; Ellenberg, 1988b).

Alpine and arctic habitats have many ecological characteristics
in common, although the climatic conditions are more severe in the
arctic regions than in most alpine regions. The growing season is
short (3-3.5 months in the low arctic zone), air and soil temperatures
are low, winds are frequent and strong, and the distribution of plant
communities depends on the distribution of snow during winter and
spring. Most alpine and all arctic zones are influenced by frost
activity or solifluction, except for soils in the low alpine and
hemiarctic zones, where podzolic soils are found. Decomposition of
organic matter and nutrient cycling are slow, and a large amount of
the nitrogen input is stored in the soil in forms which can not be
used by plants (Chapin, 1980). The low nutrient availability limits

primary production. Most species are adapted to a strict nitrogen
economy and their nitrogen indicator values are generally low
(Ellenberg, 1979).

Barsdate & Alexander (1975) investigated the nitrogen balance of
an arctic area in Alaska. The most important sources of nitrogen were
nitrogen fixation (75%) and ammonia in precipitation (22%). Most of
the nitrogen input is retained in living biomass, and very little is
leached from the soil. Denitrification is also low, partly due to
nutrient deficiency. Nitrogen metabolism as such does not seem to be
inhibited by low soil temperatures (Haag, 1974). Nitrogen fixation in
arctic habitats has been studied in bacteria, soil algae, lichens and
legume species (Leguminosae) (Novichkova-Ivanova, 1971). Blue-green
algae (cyanobacteria) are especially important in this respect, either
as free-living species, species associated with mosses or phycobionts
in lichens (e.g. Peltigera, Nephroma and Stereocaulon). The rate
of nitrogen fixation depends on temperature and moisture, and thus
varies through the year (Alexander & Schnell, 1973).

It is to be expected that arctic and alpine communities are
sensitive to increased atmospheric nitrogen input, because nitrogen
retention is very efficient, although primary production is also
strongly regulated by factors other than nitrogen (temperature,
moisture) (Tamm, 1991). The effects of increased nitrogen
availability on alpine/tundra vegetation have been studied in several
fertilizer experiments. In most experiments full nitrogen, phosphorus
and potassium fertilizer was used, although sometimes nitrogen was
applied separately. The following effects of nitrogen addition have
been observed:

* In alpine and arctic vegetation, nitrogen is quickly absorbed by
phanerogamic species and incorporated into their tissues. The
increase in nitrogen contents differs for graminoids, deciduous
and evergreen species (Summers, 1978; Shaver & Chapin, 1980;
Lechowicz & Shaver, 1982; Karlsson, 1987).

* Phanerogamic plant species respond to nitrogen application in
different ways: increased growth and biomass, enhanced number of
tillers, more flowers and changes in phenology (Henry et al.,
1986).

* Some phanerogamic plant species are damaged or even killed at
high doses of nitrogen fertilizer (Henry et al., 1986).

* Changes in species cover and composition are likely when nitrogen
is applied for a longer period of time (5-10 years).

All these studies concentrated on effects on phanerogamic plant
species; little information is available on the effects of nitrogen on
cryptogams. Many authors, however, stress that nitrogen fixation
probably will decrease when atmospheric deposition increases in arctic

and alpine ecosystems. In forest studies it has already been shown
that Cladonia spp. and some mosses are very sensitive to nitrogen
addition. The suggested critical load for arctic and alpine heaths
(5-15 kg nitrogen per ha per year) is lower than that for lowland
heathland (15-20 kg nitrogen per ha per year).

4.2.4.4 Effects on herbs of matgrass swards

In recent decades, in addition to the transition from
dwarf-shrub-dominated to grass-dominated heathlands, a reduced species
diversity in these ecosystems has been observed. Species of the
acidic Nardetalia grasslands and related dry and wet heathlands seem
to be especially sensitive. Many of these herbaceous species (e.g.,
Arnica montana, Antennaria dioica, Dactylorhiza maculata, Gentiana
pneumonanthe, Genista pilosa, Genista tinctoria, Lycopodium inundatum,
Narthecium ossifragum, Pedicularis sylvatica, Polygala serpyllifolia
and Thymus serpyllum) are declining or have even become locally extinct
in the Netherlands. The distribution of these species is related to
small-scale, spatial variability of the heathland soils. It has been
suggested that atmospheric deposition has caused such changes (Van Dam
et al., 1986). Dwarf shrubs as well as grass species are nowadays
dominant in the former habitats of these endangered species.

Enhanced nitrogen fluxes into nutrient-poor heathland soil
leads to an increased nitrogen availability in the soil. However,
most of the deposited nitrogen in western Europe originates from
ammonia/ammonium deposition and may also cause acidification as a
result of nitrification. Whether eutrophication or acidification or a
combination of both processes is important depends on pH, buffer
capacity and nitrification rates of the soil. Roelofs et al. (1985)
found that, in dwarf-shrub-dominated heathland soils, nitrification is
inhibited at pH 4.0-4.2 and that ammonium accumulates while nitrate
decreases to almost zero at these or lower pH values. Furthermore,
nitrification has been observed in the soils from the habitats of the
endangered species, owing to its somewhat higher pH and higher buffer
capacity. In soils within the pH rage of 4.1-5.9, the acidity
produced is buffered by cation exchange processes (Ulrich, 1983). The
pH will drop when calcium is depleted, and this may cause the decline
of those species that are generally found on soils with somewhat
higher pH. To study the pH effects on root growth and survival rate,
hydroculture experiments have been conducted over 4-week periods with
several of the endangered species ( Arnica, Antennaria, Viola,
Hieracium pilosella and Gentiana) and with the dominant species
( Molinia and Deschampsia) (Van Dobben, 1991). The dominant
species indeed have a lower pH optimum (3.5 and 4.0, respectively)
than the endangered species (4.2-6.0). However, the endangered
species could survive very low pH without visible injuries during this
short experimental period.

The pH decrease may indirectly result in an increased leaching of
base cations, increased aluminium mobilization and thus enhanced
aluminium/calcium (Al/Ca) ratios of the soil (Van Breemen et al.,
1982). Furthermore, the reduction of the soil pH may inhibit
nitrification and result in ammonium accumulation and consequently
increased NH₄/NO₃ ratios. In a recent field study the
characteristics of the soil of several of these threatened heathland
species have been compared with the soil characteristics of the
dominant species ( Calluna vulgaris, Erica tetralix and Molinia
caerulea) (Houdijk et al., 1993). Generally the endangered species
grow on soil with higher pH, lower nitrogen content, and lower Al/Ca
ratios than the dominant species. The NH₄⁺/NO₃ ratios were higher
in the dwarf-shrub-dominated soils than in the soil of the endangered
species. Fennema (1990, 1992) has demonstrated that soil from
locations where Arnica is still present had a higher pH and lower
Al/Ca ratio than soil of former Arnica stands. However, he found no
differences in total soil nitrogen or NH₄/NO₃ ratios. Both these
studies indicate that high Al/Ca ratios or even increased NH₄/NO₃
ratios are associated with the decline of these species. However, no
significant effects of Al and Al/Ca on growth rates have been observed
in hydroculture experiments in which the effects of Al and Al/Ca
ratios on root growth and survival rate were studied (Van Dobben,
1991). Comparable experiments of Pegtel (1987) with Arnica and
Deschampsia and Kroeze et al. (1989) with Antennaria, Viola,
Filago minima, and Deschampsia showed similar results. However,
results of a hydroculture experiment with Arnica showed that this
species is very sensitive to enhanced Al/Ca ratios at intermediate or
low nutrient levels (De Graaf, 1994). Pot experiments have indicated
that increased NH₄/NO₃ ratios lead to decreased health of Thymus.
Hydroculture experiments with this plant species confirmed that
increased NH₄/NO₃ ratios affected the cation uptake (Houdijk, 1993).
In a pot experiment Thymus, planted on acid heathland soil and on
artificially buffered heathland soil, was sprayed with 0, 15 and
150 kg nitrogen (as ammonium) per ha per year during 6 months (Houdijk
et al., 1993). In this relatively short period, a deposition of 15 kg
nitrogen (as ammonium) per ha per year on the acid soil did not lead
to ammonium accumulation in the soil. As a result of nitrification,
soil pH decreased faster than in the absence of ammonium deposition.
At the highest deposition (150 kg nitrogen (as ammonium) per ha per
year), nitrification rates in the acid heathland soils were too low to
prevent ammonium accumulation, and increased NH₄/NO₃ ratios probably
caused the decline of Thymus. Only in the artificially buffered
soils with higher pH were nitrification rates high enough to balance
ammonium and nitrate. Thymus plants on these soils were healthy
despite very high total nitrogen contents.

At present, however, there is too little information available on
these rare heathland and acidic grassland species to formulate a
critical load for nitrogen. The observation that these heathland

species generally disappear before dwarf shrubs are replaced by
grasses leads to the assumption that their critical load is lower than
the critical load for the transition to grasses (< 15-20 kg nitrogen
per ha per year) and probably between 10 and 15 kg nitrogen per ha per
year. An overview of the critical loads in heathlands is given in
section 8.2.2.

4.2.5 Effects of nitrogen deposition on forests

4.2.5.1 Effects on forest tree species

The growth of the vast majority of the forest tree species in the
Northern hemisphere was until recently limited by nitrogen. In
forestry, nitrogen fertilizers were used to increase wood production
(Tamm, 1991). An increase in the supply of an essential nutrient,
including nitrogen, will stimulate tree growth; the initial impact of
enhanced nitrogen deposition will, therefore, be a fertilizer effect.
However, continued high inputs of nitrogen produces negative effects
on tree growth (Chapin, 1980). Until the mid-1980s, almost all of the
research on forest decline focused on acidification, but it has now
become evident that enhanced nitrogen deposition may also be important
in recent forest decline.

The effects of high atmospheric nitrogen input are very complex
(Wellburn, 1988; Pitelka & Raynal, 1989; Heij et la., 1991; Pearson &
Stewart, 1993). Chronic nitrogen deposition may result in nitrogen
saturation, when enhanced nitrogen inputs no longer stimulate tree
growth, but start to disrupt ecosystem structure and function, and
increased amounts of nitrogen are lost from the ecosystem in leachate
(Agren, 1983; Aber et al., 1989; Tamm, 1991). The nitrogen input at
which saturation occurs depends on a number of factors including the
amount of deposition, vegetation type and age (see chapter 3), soil
type and management history. The following indirect processes,
besides the direct effect of gaseous pollutants on the shoots, are
important:

* Soil acidification, due to nitrification of ammonium. This
process leads to accelerating leaching of base cations and, in
poorly buffered soils, to increased dissolution of aluminium,
which can damage fine roots development and mycorrhizas, and thus
reduce nutrient uptake (Ulrich, 1983; Ritter, 1990).

* Eutrophication. Whether ammonium will accumulate in soil or
not is strongly dependent upon the nitrification rate and the
deposition levels (Boxman et al., 1988). In addition to an
initial growth stimulation and changes in root/shoot ratio,
ammonium accumulation will lead to an imbalance of the
nutritional state of the soil and concomitantly of the trees
(Roelofs et al., 1985; Van Dijk & Roelofs, 1988; Schulze et al.,
1989; Boxman et al., 1991). Accumulation of nitrates in the

ecosystem may also lead to eutrophication. As a consequence of
all these processes, the health of the trees declines and their
sensitivity to drought, frost, insect pests and to pathogens can
increase markedly (Wellburn, 1988). These phenomena may also
play a secondary, but certainly not unimportant, role in the
dieback of forest trees and have also been reviewed.

Although many tree species occur in natural forest ecosystems,
almost all studies on air pollution have concentrated on a few
forestry tree species from acidic, nutrient-poor soils. Most of these
species are conifers ( Picea, Pinus and Pseudotsuga spp.) and the
following section concentrates on the long-term soil-mediated effects
on these trees. Available data on broad-leaved species ( Fagus,
Quercus) are also considered. Long-term effects of nitrogen
eutrophication on the composition of the tree layer in natural forests
may be expected but have not yet been quantified. Soil acidification
per se has only been briefly reviewed, because the critical load for
acidity and tree growth is well established (Nilsson & Grennfelt,
1988; Downing et al., 1993).

a) Soil-mediated changes in nutritional status of forest tree species

It has been shown that in areas with high ammonia/ammonium
deposition, ammonium accumulates in acid forest soils with little or
no nitrification. Van Dijk & Roelofs (1988) found ammonium ion
accumulation in damaged Pinus and Pseudotsuga stands receiving
60-100 kg nitrogen per ha per year, although the pH of the soil was
the same as that in healthy stands. This build-up of ammonium ion
leads to increased ratios of ammonium to base cations (Roelofs et al.,
1985; Boxman et al., 1988), a reduction of base cation uptake and,
eventually, nutritional problems. Using soil columns with different
ammonium sulfate spraying treatments, critical ratios of excess
ammonium to base cations have been determined (Boxman et al., 1988).
The nutritional problems of the coniferous species studied have been
found above values of 5, 10 and 1, respectively, for the NH₄/K,
NH₄/Mg and Al/Ca ratios in soil solution. In soil with zero or a low
nitrification rate, 10-15 kg nitrogen per ha per year is a reliable
critical load to prevent critical ammonium to cation ratios, whereas
in base-cation-rich soil with moderate to high nitrification rates the
critical loads obtained are higher (20-30 kg nitrogen per ha per
year).

The nutritional status of the coniferous trees studied, after
enhanced nitrogen inputs, is affected by both ammonium accumulation
and soil acidification. Base cation concentrations in the soil are
reduced by leaching, whereas base cation uptake by plants is reduced
by excess of ammonium and of aluminium. Furthermore, root growth is
decreased (see later). Laboratory, greenhouse and field measurements
in the Netherlands, Germany and southern Sweden (Van Dijk & Roelofs,
1988; Van Dijk et al., 1989, 1990, 1992a; Hofmann et al., 1990;

Schulze & Freer-Smith, 1991; Boxman et al., 1991, 1994; Ericsson et
al., 1993) have shown that the complex of factors just noted produce
severe deficiencies of magnesium and potassium in coniferous trees.
Most of these studies were in areas, or involved experiments, with
large inputs (> 40-100 kg nitrogen per ha per year).

The magnesium and phosphorus concentrations in leaves of oak
trees (Fagus sylvatica), a common deciduous tree in Europe,
decreased significantly from 1984 to 1992 in permanent plots in NW
Switzerland. Furthermore, the magnesium concentrations in the leaves
of young Fagus sylvatica decreased significantly within a 4-year
period of fertilizer application at > 25 kg nitrogen per ha per
year (Flückiger & Braun, 1994). In Sweden, suboptimal concentrations
of magnesium and potassium in Fagus leaves were found in areas with
the highest nitrogen deposition (Balsberg-Pählsson, 1989) and addition
of nitrogen enhanced nutritional imbalance in a 120-year-old Fagus
stand (Balsberg-Pählsson, 1992). It is thus clear that this deciduous
tree species is also sensitive to nutritional imbalance induced by
enhanced nitrogen supply.

Base cations are also lost from the canopy by increased leaching,
linked to high amounts of atmospheric deposition (Wood & Bormann,
1975; Roelofs et al., 1985; Bobbink et al., 1992b). As a result of
high nitrogen inputs, the organic nitrogen concentration in the
needles of conifers has increased significantly to supra-optimal
levels (Van Dijk & Roelofs, 1988; De Kam et al., 1991). Concentrations
of nitrogen-rich free amino acids, especially arginine, have
significantly increased in the needles with high nitrogen concentration
(> 1.5% nitrogen in Picea abies) (Hällgren & Näsholm, 1988;
Pietila et al., 1991; Van Dijk et al., 1992) and in Fagus leaves
(Balsberg-Pählsson, 1992).

Although there is clear evidence that high NH₃/NH₄ loads
produce adverse changes in the nutritional status and the growth of
the investigated coniferous and broad-leaved trees, it is difficult to
obtain a critical load for nitrogen from these studies, because of the
complexity of the ecosystem. A quite reliable critical load for
nitrogen deposition on beech tree health is around 15-20 kg nitrogen
per ha per year, as demonstrated in the Swiss studies (Flückiger &
Braun, 1994).

The results of the EC nitrogen saturation study (NITREX), which
incorporates long-term experiments in both clean and nitrogen-polluted
areas and whole ecosystem manipulation of nitrogen inputs, are
providing important evidence on the effects of nitrogen deposition
on tree health and ecosystem health. Atmospheric deposition of
nitrogen was reduced from 40 to 2 kg nitrogen per ha per year in a
nitrogen-saturated Pinus sylvestris stand in the Netherlands (Boxman
et al., 1994, 1995). Throughfall water was intercepted with a roof
and replaced by clean throughfall water from 1989 onwards. In the

clean plot a quick response of the soil solution chemistry was
observed. The nitrogen concentrations in the upper soil and the
fluxes of this element through the soil profile decreased. As a
result, base cation leaching and the ratios of ammonium to various
cations also decreased; potassium and magnesium concentrations in the
needles increased significantly. The needle nitrogen concentrations
were only slightly reduced in the "clean" situation, but they were
significantly lower than in the needles of the control plots. The
concentration of arginine decreased significantly in the needles of
the trees from the clean throughfall plot. Furthermore, tree growth
became higher after 4 years of clean throughfall than in control plots
with high nitrogen deposition. No changes in the mycorrhizal status or
in the undergrowth have so far been observed (Boxman et al., 1994,
1995). This study clearly demonstrates the detrimental effects of
enhanced atmospheric nitrogen deposition on the nutritional balance of
coniferous trees.

b) Nitrogen deposition and tree susceptibility to frost, drought and
pathogens

It has been suggested by several authors that sensitivity of
trees to secondary stress factors is increased by high nitrogen
loading (Wellburn, 1988; Pitelka & Raynal, 1989). In field fertilizer
applications it is often observed that tree growth starts earlier in
the season, which may increase damage by late frost. Furthermore, it
has been shown, after nutrient applications, that frost damage to
Pinus sylvestris increases considerably at needle nitrogen
concentrations above 1.8% (Aronsson, 1980), although other fertilizer
studies have demonstrated reverse effects, i.e. improved nitrogen
status of the plants diminishes frost damage (De Hayes et al., 1989;
Klein et al., 1989; Cape et al., 1991).

Only few data are available with respect to frost damage in
direct relation to airborne nitrogen deposition. After exposure to
NH₃ and SO₂, Pinus sylvestris saplings became more frost sensitive
(< -10°C) than control plants (Dueck et al., 1990). Dueck et al.
(1990) also determined the frost sensitivity of Pinus sylvestris
growing in areas with low ammonia/ammonium pollution (approximately
4 µg NH₃/m³) and in highly polluted areas (40 µg NH₃/m³).
Surprisingly, the frost sensitivity was not higher in the polluted
area than in the other investigated sites, and was sometimes even
lower. After experimental treatment with ammonia (53 µg NH₃/m³) the
growth of the trees had increased, indicating that the observed change
in frost sensitivity might have occurred as a result of changes in
physiology and nutrient imbalance.

The effects of simulated acid mist containing sulfate, ammonium,
nitrate and H⁺ on the frost sensitivity of Picea rubens has been
studied (Sheppard et al., 1993; Sheppard, 1994). There was a strong

correlation between the application of sulfate-containing mist and an
increase in frost sensitivity, but no such correlation was seen after
treatment with ammonium or nitrate ions. Sulfur compounds clearly
affect the frost sensitivity of coniferous trees, but this effect may
be a consequence of the nutritional status (nitrogen, base cations) of
the trees (Sheppard, 1994). It is concluded that the effects of
increased nitrogen inputs on frost sensitivity remain uncertain.
Insufficient research has been carried out to use the results for
assessment of a critical load.

The water uptake of coniferous trees species may be affected by
increasing nitrogen deposition, owing to an increase in shoot-to-root
ratio and a reduction in fine-root length. Indeed, the health of many
tree species in the regions of the Netherlands with high nitrogen
deposition was particularly poor in the dry years in the mid-1980s,
but improved again during the subsequent normal years (Heij et al.,
1991). Many authors have mentioned a negative impact of high nitrogen
supply on the development of fine roots and mycorrhiza, although
positive effects have also been described (Persson & Ahlstrom, 1991).

Van Dijk et al. (1990) applied 0, 48, 480 kg nitrogen (as
ammonium sulfate) per ha per year to young Pinus sylvestris, Pinus
nigra and Pseudotsuga menziesii in a pot experiment. After seven
months the coarse root biomass had not changed, but the fine root
biomass decreased by 36% at the highest nitrogen application. In
parallel, a 63% decrease in mycorrhizal infection at the highest
nitrogen application was found. In the Dutch EC nitrogen saturation
study, the fine root biomass and the number of root tips of Pinus
sylvestris increased after reduction of the current nitrogen
deposition to pre-industrial levels, indicating restricted root growth
and nutrient uptake capacity at the ambient nitrogen load of about
40 kg nitrogen per ha per year (Boxman et al., 1994, 1995).

In a hydroculture experiment with Pinus nigra at pH=4.0, Boxman
et al. (1991) found an increase in coarse/fine root ratio after
increasing the ammonium concentration to 5000 µM. Furthermore, a
clear relation was found between the nitrogen content of the fine
roots and mycorrhizal infection (as measured as the number of
dichotomously branched roots). In a hydroculture experiment Jentschke
et al. (1991) found, however, that 2700 µM nitrate had hardly any
effect on the mycorrhizal development of Picea abies seedlings
inoculated with Lactarius rufus. Ammonium at 2700 µM only had a
slight negative effect on mycorrhizal development, whereas a reduction
in root growth was recorded. In a pot experiment with Picea abies,
Meyer (1988) found optimal mycorrhizal development when the mineral
nitrogen content of the soil was 40 mg nitrogen/kg dry soil, while at
350 mg nitrogen/kg dry soil a 95% reduction in mycorrhizal development
was found. In this study no correlation was found with the soil pH.
Alexander & Fairly (1983) found, after fertilizer application to a
35-year-old Picea sitchensis stand with 300 kg nitrogen (as ammonium

sulfate) per ha, a 15% reduction in mycorrhizal development in the
second year after application. Termorshuizen (1990) applied 0 to
400 kg nitrogen ha per year either as ammonium or nitrate to young
Pinus sylvestris inoculated with Paxillus involutus in a pot
experiment. Above application rates of 10 kg nitrogen per ha per year
there was a decrease in the amount of mycorrhizal root tips and the
number of sclerotia.

In addition to the above-mentioned data for coniferous trees, it
had been shown that the shoot-to-root ratios of young Fagus
sylvatica trees, grown in containers with acid forest soil,
increased significantly from about 1 to between 2 and 3 after a 4-year
experimental application of nitrogen (25 kg nitrogen per ha per year
or more) (Flückiger & Braun, 1994).

It is thus likely that enhanced nitrogen inputs affect drought
sensitivity through changes in shoot to root ratios, number of fine
roots and the ectomycorrhizal infection of the roots. However, the
data are too few to use for the assessment of a critical load of
nitrogen, based upon this aspect of reduced tree health.

There may also be significant effects of fungal pathogens or
insect pests associated with increasing nitrogen deposition. The
foliar concentrations of nitrogen increased markedly in tree needles
or leaves in experiments with nitrogen additions, and also in forest
sites with high atmospheric nitrogen loading (Roelofs et al., 1985;
Van Dijk & Roelofs, 1988; Balsberg-Pählsson, 1992). Animal grazing
generally increases with increasing palatability of the leaves or
shoots. Nitrogen is of major importance for the palatability of plant
material, and this certainly holds for insect grazing (Crawley, 1983).
Secondary plant chemicals, e.g., phenolics, are important for
increased resistance of plants. The total amount of phenolics in
Fagus leaves in a 120-year stand decreased by more than 30% after
fertilizer application of about 45 kg nitrogen per ha per year,
compared with the control treatment (Balsberg-Pählsson, 1992). An
ecologically important relation between nitrogen enrichment and insect
pests has been quantified for lowland heathland (Brunsting & Heil,
1985; Berdowski, 1993, see section 4.1) but not, so far, for forest
ecosystems.

From 1982 to 1985 an epidemic outbreak of the pathogenic fungus
Sphaeropsis sapinea was observed in coniferous forest (mainly
Pinus nigra) in the Netherlands. This greatly affected whole
stands, and was especially severe in the south-east part of the
Netherlands, where there was high airborne nitrogen deposition
(Roelofs et al., 1985). Van Dijk et al. (1992) showed that there was
a significantly higher foliar nitrogen concentration in the infected
stands, together with higher soil ammonium levels, than in the
uninfected stands. Most of the additional nitrogen in the needles of
the affected stands was stored as nitrogen-rich free amino acids,

especially arginine. Proline concentrations were also higher in the
infected trees, indicting a relation with water stress (Van Dijk
et al., 1992).

The effects of Sphaeropsis have also been studied by De Kam et
al. (1991). Two-year-old plants of Pinus nigra were grown for
3 years in pots and given five treatments of ammonium sulfate (very
low to about 300 kg nitrogen per ha per year), in combination with two
levels of potassium sulfate. The 5-year-old plants were then
inoculated with Sphaeropsis. The bark necroses were much more
frequent in the plants treated with ammonium sulfate than in the
controls. Effects of ammonium sulfate upon fungal damage were even
observed at an addition of 75 kg nitrogen per ha per year, but were
very significant in the plants treated with 150 kg nitrogen per ha per
year. After potassium addition the number of necroses caused by the
fungus was greatly reduced (De Kam et al., 1991).

In beech forests in NW Switzerland, a significant positive
correlation has been found between the nitrogen/potassium ratios in
the leaves and necroses caused by the beech cancer Nectria ditissima
(Flückiger & Braun, 1994). These authors also experimentally
inoculated Fagus sylvatica trees at different applications of
nitrogen with this beech cancer and observed increased dieback of new
leaves and shoots. Furthermore, the infestation of Fagus sylvatica
with beech aphids (Phyllaphis fagi) was also affected by the
nitrogen availabilities. The degree of infestation with the aphid
increased significantly with enhanced leaf nitrogen/potassium ratios
(Flückiger & Braun, 1994). Although evidence for nitrogen-mediated
changes in susceptibility to fungal pests and insect attacks has until
now been based upon observations of only few species, it is obvious
that trees became more susceptible to these attacks with increasing
nitrogen enrichment and this may play a crucial role in the dieback of
some forest stands.

A critical load for nitrogen had been established at 10-15 kg
nitrogen (at no or low nitrification) to 20-30 kg nitrogen per ha per
year in highly nitrifying soils, based upon nutritional imbalance of
coniferous species (Boxman et al., 1988). Recent evidence of Fagus
sylvatica tree health in acidic forests indicated a critical load
of 15-20 kg nitrogen per ha per year, based upon both field and
experimental observations. Elevated nitrogen deposition can
seriously affect tree healthy via a complex web of interactions (e.g.
susceptibility to frost and drought). Pathogens may play an important
role in tree decline, but at this moment it is not possible to combine
the observed processes and effects to an overall value for a critical
load of nitrogen for tree health.

4.2.5.2 Effects on tree epiphytes, ground vegetation and ground fauna
of forests

a) Effects on ground-living and epiphytic lichens and algae

The effects of SO_y as an acidifier on epiphytic lichens have
been extensively studied (Insarova et al., 1992; Van Dobben, 1993).
SO_y was previously the dominant airborne pollutant, and it has been
shown that most (epiphytic) lichens are more negatively affected by
acidity than by nitrogen compounds (except NO_y). Most lichens have
green algae as photobionts and are affected by acidity but not by
nitrogen. Some of them even react positively to nitrogen (Insarova et
al., 1992). However, 10% of all lichen species in the world have
cyanobacteria (blue-green algae) as the photobiont. These
cyanobacterial lichens are negatively affected by acidity, and also by
nitrogen. Most of the NW European lichens with cyanobacteria live on
the soil surface or are tree epiphytes. The most pollution-sensitive
lichens are among them and they are threatened by extinction in NW
Europe. This is probably the result of increased nitrogen deposition,
which inhibits the functioning of the cyanobacteria. In the
Netherlands, for example, all cyanobacterial lichens that were present
at the end of the 19th century are now absent. In Denmark, 96% of the
lichens with cyanobacteria are extinct or threatened. Furthermore,
the cyanobacterial lichens appear frequently on the Red List of the
European Union countries (Hallingbäck, 1991).

Very few data exist to establish a critical load for nitrogen for
these lichens with blue-green algae. Nohrstedt et al. (1988)
investigated the effects of nitrogen application (as ammonium nitrate
or calcium nitrate) on ground-living lichens ( Peltigera aphtosa and
Nephroma arcticum) with blue-green algae as photobionts. The plots
were treated once or three or four times with 120, 240 or 360 kg
nitrogen per ha. After a short period all Peltigera and Nephroma
lichens were eliminated and even 19 years later no recolonization had
occurred. However, it is impossible to transform these very high
doses to critical loads. The effects of air pollutants on lichens are
usually related to concentrations in the air or in the precipitation.
It is probably more relevant to relate the effects of nitrogen on
cyanobacterial lichens to deposition than to concentrations. For tree
epiphytes stemflow is most relevant, whereas for ground-living lichens
throughfall will be more important. Although much research is still
needed, it has been suggested that a load of 5-15 kg nitrogen per ha
per year is already critical for the growth of these cyanobacterial
lichens (Hallingbäck, 1991). These lichens may be the most sensitive
components of some forest ecosystems and thus determine the critical
load for these systems.

Free-living green algae, especially of the genus Pleurococcus
( Protococcus and Demococcus are synonyms), are strongly stimulated
by enhanced nitrogen deposition. They cover practically all outdoor

surfaces which are not subject to frequent desiccation in regions with
high nitrogen deposition, such as in the Netherlands and in Denmark.
The thickness and the colonization rate of spruce needles by green
algae has been investigated in the Swedish Environmental Monitoring
Programme (Brakenhielm, 1991). The Swedish data show that these algae
do not colonize spruce needles in regions with a total deposition
(throughfall) lower than about 5 kg nitrogen per ha per year. In
areas with deposition above 20 kg nitrogen per ha per year, the green
algal cover of the needles is so thick and the algae colonize so early
that they may impede the photosynthesis of the spruce trees.

b) Effects on forest ground vegetation

In the Netherlands the forest vegetation of a site in the central
part of the country was investigated in 1958 (with about 20 kg
nitrogen per ha per year) and in 1981 (with about 40 kg nitrogen per
ha per year). All lichens had disappeared during this period and a
considerable increase in Deschampsia flexuosa and Corydalis
claviculata was found. A large representative sample test (n=2000),
covering about 90% of the Dutch forests, revealed in the mid-1980s
that among the 40 most common forest plants were: Galeopsis
tetrahit, Rubus species, Deschampsia flexuosa, Dryoptesis
cathusiana, Molinia caerulea, Poa trivialis, and Urtica dioica
(Dirkse & Van Dobben, 1989; Dirkse, 1993). In Sweden, Quercus robur
stands in two geographical areas with different nitrogen deposition
were compared with special emphasis on nitrogen indicator species
(Tyler, 1987). The stands were quite comparable except for the
nitrogen inputs: 6-8 kg nitrogen per ha per year and 12-15 kg nitrogen
per ha per year, respectively. In the stand with the highest
deposition, the soil solution was more acidic, probably due to acidic
deposition as well (± 10 kg sulfur per ha per year), and it was
estimated that acidification of the soil has accelerated during the
last 30 to 50 years. The following species were more common in the
most polluted site: Urtica dioca, Epilobium augustifolium, Rubus
idaeus, Stellaria media, Galium aparine, Aegopodium podagraria and
Sambucus spp. Thus, both in Sweden and the Netherlands, species
indicative of nitrogen enrichment became common (Ellenberg, 1988b).

Comparable observations were reported by Falkengren-Grerup (1986)
and by Falkengren-Grerup & Eriksson (1990), who examined the changes
in soil and vegetation in Quercus and Fagus stands in southern
Sweden. They concluded that the exchangeable base cations were reduced
and that aluminium had doubled over the past 35 years. They also
found a decrease in soil pH, with a disappearance of several species
when pH dropped below a threshold. In spite of soil acidification
some species had increased in cover, and the most plausible
explanation seemed to be increased nitrogen deposition, which was
about 15-20 kg nitrogen per ha per year in southern Sweden and which
had doubled since 1955. A marked increase in cover was found for

Lactuca muralis, Dryopteris filix-max, Epilobium augustifolium,
Rubus idaeus, Melica uniflora, Aegopodium podagraria, Stellaria
holostea and S. nemorum, some of these species being nitrogen
indicators. Despite soil acidification, acid-tolerant species
( Deschampsia flexuosa, Maianthemum bifolium and Luzula pilosa) did
not increase. A distinct decrease was observed for Dentaria
bulbifera, Pulmonaria officinalis and Polygonatum multiflorum.
Furthermore, Rosen et al. (1992) found a significant positive
correlation between the increase of Deschampsia flexuosa cover in
the last 20 years in the Swedish forests and the pattern of nitrogen
deposition.

In a large semi-natural Fagus-Quercus forest in NE France,
about 50 permanent vegetation plots were investigated in 1972 and
1991. The changes in species composition on calcareous soils and in
moderately acidic habitats were followed. During the study period a
significant increase in nitrophilous ground flora was observed in the
high-pH (6.9) stands. This indicated that at this location (with
ambient deposition of 15-20 kg nitrogen per ha per year) there was a
distinct effect of increasing nitrogen availability (Thimonier et al.,
1994).

From 1968 to 1985, three sites in a 30-year-old Pinus
sylvestris forest in Lisselbo (central Sweden) were annually
fertilized with 0, 20, 40 and 60 kg nitrogen per ha per year (as
NH₄NO₃ plus ambient deposition of 10 kg nitrogen per ha per year).
The original ground vegetation consisted of Calluna vulgaris,
Vaccinium vitis-idea, V. myrtillus, Cladonia spp., Cladina spp.,
and the mosses Dicranum spp., Pleurozium spp. and Hylocomium
spp. The first changes were observed within 8 to 15 years and after
about 20 years the experimental plots were compared and statistically
analysed. The original species disappeared at nitrogen applications
above 20 kg (plus ambient deposition) nitrogen per ha per year and
were replaced by Epilobium augustifolium, Rubus idaeus, Deschampsia
flexuosa, Dryopteris carthusiana and the moss Brachythecium
oedipodium (Dirkse et al., 1991; Van Dobben, 1993). In another
experiment at Lisselbo the combined effects of acidification (addition
of H₂SO₄, pH=2.0) and nitrogen addition (0 and 40 kg nitrogen per ha
per year) were investigated. The increased nitrogen level seemed to
be the more important factor. Acidification was the next most
discriminating factor: all species disappeared, except for the moss
Pohlia nutans at high additions of acidity (Dirkse & Van Dobben,
1989; Dirkse et al., 1991).

In southern Sweden, Tyler et al. (1992) studied the effects of
the application of ammonium nitrate (60-180 kg nitrogen per ha per
year) over a 5-year period on stands of Fagus sylvatica. They
observed a large reduction in biomass of the ground vegetation with
the application of nitrogen, and the frequency of most herb layer
species declined significantly. Soil measurements revealed that, in

addition to eutrophication effects, the acidification of the soil
solution was also important for the decline of the original ground
vegetation. In an experiment on the effects of nitrogen fertilizer
application on bryophytes, it appeared that Brachythecium
oedipodium, B. reflexum and B. starkei increased significantly at
levels up to 60 kg nitrogen per ha per year. At higher doses these
species tended to decline, however. Hylocomium splendens and
Pleurozium schreberi declined considerably at doses of 30 to 60 kg
nitrogen per ha per year (Dirkse & Martaki, 1992).

c) Effects on macrofungi and mycorrhizas

During the last two decades many reports have described a
decrease in species diversity and abundance of macrofungi. These
changes can probably be attributed to indirect effects of air
pollution, in particular to increases in the amount of available
nitrogen (possibly in combination with acidification), and/or to
decreased health of trees with concomitant reduction of transport to
the roots (Arnolds, 1991).

When comparing sites over time, the number of fruiting bodies of
macrofungi showed marked differences. Most studies in western Europe,
however, have revealed that the number of ectomycorrhizal fungi
species has declined (Arnolds, 1991). In the Netherlands the average
number of ectomycorrhizal species per foray declined significantly
from 71 in 1912-1954 to 38 in 1973-1982. Similar changes have been
observed in Germany: 94 ectomycorrhizal species found in 1950-1979 in
the Völklinger area (Saarland) have not been recorded recently. From
the 236 species found in 1918-1942 in the Darmstadt area (Germany),
only 137 were recorded in the early 1970s, a loss of 99 species,
including many mycorrhizal fungi (Arnolds, 1991). In contrast to the
decline in mycorrhizal fungi, the number of saprotrophic species
remained practically unchanged, while the number of lignocolous
species increased. This may be related to soil acidification with a
increase in aluminium, since the proportion of forest areas in western
Europe with a soil pH below 4.2 increased from less than 1% in 1960 to
15% in 1988 (Schneider & Bresser, 1988).

Arnolds (1988, 1991) concluded that acidification has very little
effect on the diversity of ectomycorrhizal fungi, but rather triggers
changes in species composition. He regarded the increased nitrogen
flux to the forest floor as the most important factor in the decline
of mycorrhizal fungi. Termorshuizen & Schaffers (1987) found a
negative correlation between the total nitrogen input in mature
Pinus sylvestris stands and the abundance of fruit bodies of
ectomycorrhizal fungi. Similar results were obtained by Schlechte
(1986) who compared two sites with Picea abies in the Göttingen area
of Germany. An obvious negative relation was found between nitrogen
input (23 versus 42 kg nitrogen per ha per year) and ectomycorrhizal
species: 85 basidiomycetes including 21 ectomycorrhizas (25%) at the

less polluted site compared with 55 basidiomycetes including
3 ectomycorrhizas (5%) at the most polluted site. Environmental
factors other than nitrogen did not differ significantly. The
negative impact of nitrogen seems only to hold true for mature forests
(Termorshuizen & Schaffers, 1987). Jansen & de Vries (1988) found a
maximum in fruit-body production in > 20-year-old Pseudotsuga
menziesii stands at about 25 kg nitrogen per ha per year. Meyer
(1988) found a similar optimum when Picea abies was planted in soil
mixed with different amounts of sawdust having a high carbon/nitrogen
ratio.

Experiments with nitrogen fertilizer have produced similar
results. In a fertilizer trial with simulated nitrogen deposition in
a Fagus forest in southern Sweden (ambient deposition 15-20 kg
nitrogen per ha per year), Ruhling & Tyler (1991) found, after
applying NH₄NO₃ (60 and 180 kg nitrogen per ha per year), that
within 3 to 4 years almost all mycorrhizal species ceased fruit-body
production. In contrast, several decomposer species increased
fruit-body production. Wood decomposers showed no obvious reaction to
the treatment. No fruit-bodies were recovered when 300 kg nitrogen
per ha was applied to Pinus sylvestris stands as liquid manure
(Ritter & Tölle, 1978). The mycorrhizal frequency of the roots,
however, was still 55% as compared to 87% in the controls.
Application of 112 kg nitrogen (as NH₄NO₃) per ha to 11-year-old
Pinus taeda stands revealed an 88% reduction in the number of
fruit-bodies and a 14% decrease in the number of mycorrhizas per unit
of soil volume (Menge & Grand, 1978). In the Lisselbo study the
number of fruit-bodies decreased considerably at each nitrogen
fertilizer dose (Wasterlund, 1982). Termorshuizen (1990) applied
0, 30 and 60 kg nitrogen (as ammonium sulfate or nitrate) per ha per
year to young Pinus sylvestris stands. In general fruit-body
production was more negatively influenced by the higher ammonium
levels than nitrate levels. The mycorrhizal frequency and the number
of mycorrhizas per unit of soil volume were not influenced. It was
concluded by Termorshuizen (1990) that fruit-body production is much
more sensitive to nitrogen enrichment that mycorrhizal formation.
Branderud (1995) found after only 1.5 year a decrease in fruit-body
production of mycorrhizal species at a nitrogen application of 35 kg
nitrogen (as NH₄NO₃) per ha in a Picea abies stand at the Swedish
Nitrex stand.

In contrast, some studies have shown an increase in the number of
fruit-bodies of insensitive mycorrhizal fungi after nitrogen
fertilizer application, e.g., Paxillus involutes (Hora, 1959),
Laccaria bicolor (Ohenoja, 1988) and Lactarius rufus (Hora, 1959).

d) Effects on soil fauna of forests

Almost all studies of changes in faunal species composition due
to nitrogen enrichment have been conducted in arable fields or

agricultural grasslands using complete fertilization and thus cannot
be used to substantiate critical loads for semi-natural forest
ecosystems (Marshall, 1977). The relationship between acidity and
soil fauna has also been studied in northern coniferous forests, but
only very few studies have incorporated the effects of nitrogenous
compounds (Gärdenfors, 1987). The abundance of Nematoda,
Oligochaeta and microarthropods (especially Collembola) had
increased in some studies, but decreased in others, after application
of high doses of nitrogen fertilizers (> 150 kg nitrogen per ha per
year) (Abrahamsen & Thompson, 1979; Huhta et al., 1983; Vilkamaa &
Huhta, 1986). A reduction in the nitrogen deposition in a Pinus
sylvestris stand (Nitrex site Ysselstein) to pre-industrial levels
increased the species diversity of microarthropods due to a decreased
dominance of some species (Boxman et al., 1995). However, it is not
possible to use these few data to formulate a critical load for
changes in forest soil fauna due to increased nitrogen deposition.

On the basis of the results presented in this overview, the
critical load for changes in the ground vegetation of both coniferous
and deciduous acidic forest may be 15 to 20 kg nitrogen per ha per
year. The critical load for changes in the fruit-body production of
ectomycorrhizal fungi is probably about 30 kg nitrogen per ha per
year, while the critical load for changes in mycorrhizal frequency of
tree roots is hard to estimate, but certainly considerably higher.
There is insufficient data on the effects of enhanced nitrogen
deposition on faunal components of forest ecosystems to allow critical
loads to be set. Epiphytic or ground-living lichens with
cyanobacteria as the photobiont probably form a sensitive part of
forest ecosystems and have an estimated critical load of 10-15 kg
nitrogen per ha per year. A summary of the critical loads for forests
is given in chapter 8.

4.2.6 Effects on estuarine and marine ecosystems

Few topics in aquatic biology have received as much attention
over the past decade as the debate over whether estuarine and coastal
ecosystems are limited by nitrogen, phosphorus or some other factor
(Hecky & Kilham, 1988). Numerous geochemical and experimental studies
have suggested that nitrogen limitation is much more common in
estuarine and coastal waters than in freshwater systems. Taken as a
whole, the productivity of estuarine waters in the USA correlates more
closely with supply rates of nitrogen than with those of other
nutrients (Nixon & Pilson, 1983).

Estimation of the contribution of nitrogen deposition to the
eutrophication of estuarine and coastal waters is made difficult by
the multiple direct anthropogenic sources (e.g., from agriculture and
sewage) of nitrogen against which the importance of atmospheric
sources must be weighed. Estuaries and coastal areas are common
locations for cities and ports. The crux of any assessment of the

importance of nitrogen deposition to estuarine eutrophication lies in
establishing the relative importance of direct anthropogenic exposure
(e.g., sewage and agricultural run-off) and indirect effects
(e.g., atmospheric deposition).

The effects of nitrogen deposition in certain estuarine systems
have been investigated. Complete nitrogen budgets, as well as
information on nutrient limitation and seasonal nutrient dynamics,
have been compiled for two large "estuaries", the Baltic Sea
(Scandinavia) and the Chesapeake Bay (USA), and for the Mediterranean
Sea. In the case of the Mediterranean, Loye-Pilot et al. (1990)
suggest that 50% of the nitrogen load originates as deposition falling
directly on the water surface. In the case of the Baltic and
Chesapeake, deposition of atmospheric nitrogen has been suggested as a
major contributor to eutrophication. Data for other coastal and
estuarine systems are less complete, but similarities between these
two systems and other estuarine systems suggest that their results may
be more widely applicable. Discussion in this monograph is limited to
these two case studies, with some speculation about how other
estuaries may be related.

The Baltic Sea is perhaps the best-documented case study of the
effects of nitrogen additions in causing estuarine eutrophication.
Like many other coastal waters, the Baltic Sea has experienced a
rapidly increasing anthropogenic nutrient load. It has been estimated
that the supply of nitrogen has increased by a factor of 4, and
phosphorus by a factor of 8, since the beginning of the 20th century
(Larsson et al., 1985). The first observable changes attributable to
eutrophication of the Baltic were declines in the concentration of
dissolved oxygen in the 1960s (Rosenberg et al., 1990). Decreased
dissolved oxygen concentrations result when decomposition in deeper
waters is enhanced by the increased supply of sedimenting algal cells
from the surface water layers to the sediment. In the case of the
Baltic, the spring algal blooms that now result from nutrient
enrichment consist of large, rapidly sedimenting algal cells, which
supply large amounts of organic matter to the sediment for
decomposition (Enoksson et al., 1990). Since the 1960s, researchers
in the Baltic have documented increases in algal productivity,
increased incidence of nuisance algal blooms, and periodic failures
and unpredictability in fish and Norway Lobster catches (Fleischer &
Stibe, 1989; Rosenberg et al., 1990). It has now been shown by a
number of methods that algal productivity in nearly all areas of the
Baltic Sea is limited by nitrogen. Nitrogen-to-phosphorus ratios
range from 6:1 to 60:1 (Rosenberg et al., 1990), but the higher ratios
are only found in the remote and relatively unaffected area of the
Bothnian Bay (between Sweden and Finland). Productivity in the spring
(the season of highest algal biomass) is fuelled by nutrients supplied
from deeper waters during spring overturn (Graneli et al., 1990); deep
waters are low in nitrogen and high in phosphorus, resulting in
nitrogen-to-phosphorus ratios near 5 (Rosenberg et al., 1990),

suggesting potential nitrogen limitation when deep waters are mixed
with surface waters. Low nitrogen-to-phosphorus ratios in deep water
result from denitrification in the deep sediments (Shaffer & Rönner,
1984). Primary productivity measurements in the Kattegat (the portion
of the Baltic between Denmark and Sweden) correlate closely with
uptake of NO₃^-, but not of PO₄^3- (Rydberg et al., 1990). Level II
and III nutrient enrichment experiments conducted in coastal areas of
the Baltic, as well as in the Kattegat, indicate nitrogen limitation
at most seasons of the year (Graneli et al., 1990). Growth
stimulation of algae has also been produced by addition of rain water
to experimental enclosures, in amounts as small as 10% of the total
volume (Graneli et al., 1990); rain water in the Baltic is rich in
nitrogen but poor in phosphorus. In portions of the Baltic where
freshwater inputs keep the salinity low, blooms of the nitrogen-fixing
cyanobacterium Aphanizomenon flos-aquae are common (Graneli et al.,
1990); cyanobacterial blooms are common features of nitrogen-limited
freshwater lakes but are usually absent from marine waters.

Nitrogen budget estimates indicate that the Baltic Sea as a whole
receives 7.6 × 10¹⁰ eq of nitrogen per year, of which 2.8 × 10¹⁰ eq
per year (37%) comes directly from atmospheric deposition (Rosenberg
et al., 1990). Fleischer & Stibe (1989) reported that the nitrogen
flux from agricultural watersheds feeding the Baltic has been
decreasing since about 1980 but that the nitrogen contribution from
forested watersheds is increasing. They cite both increases in
nitrogen deposition and the spread of modern forestry practices as
causes for the increase. It should be noted, however, that the Baltic
also experiences a substantial phosphorus load from agricultural and
urban lands, and that phosphorus inputs may help to maintain
nitrogen-limited conditions (Graneli et al., 1990). If the Baltic had
received consistent nitrogen additions (e.g., from the atmosphere or
from agricultural run-off) in the absence of phosphorus additions, it
might well have evolved into a phosphorus-limited system some time
ago.

The physical structure of the Baltic Sea, with a shallow sill
limiting exchange of water with the North Sea contributes to the
eutrophication of the basin, by trapping nutrients in the basin once
they reach the deeper waters. Because the larger algal cells that
result from nutrient enrichment in the basin provide more nutrients to
the deep water through sedimentation, and because only shallow waters
have the ability to exchange with the North Sea, it is estimated that
less than 10% of nutrients added to the Baltic are exported over the
sill to the North Sea (Wulff et al., 1990). Throughout much of the
year (i.e., especially during the dry months) productivity in the
Baltic is maintained by nutrients recycled within the water column
(Enoksson et al., 1990). The trapping of nutrients within the basin
and recycling of nutrients from deeper water by circulation patterns

suggest that eutrophication of the Baltic is a self-accelerating
process (Enoksson et al., 1990) and has a long time-lag between
reductions of inputs and improvements in water quality.

In the USA, a large effort has been made to establish the
relative importance of sources of nitrogen to Chesapeake Bay (D'Elia
et al., 1982; Smullen et al., 1982; Fisher et al., 1988; Tyler, 1988).
Estimates of the contribution of nitrogen to Chesapeake Bay from each
individual source are very uncertain; estimating the proportion of
nitrogen deposition exported from forested watersheds is especially
problematic but critical to the analysis, because about 80% of the
Chesapeake Bay basin is forested. Nonetheless, three attempts at
determining the proportion of the total nitrate load to the Bay
attributable to nitrogen deposition all produce estimates in the range
of 18 to 31%. Supplies of nitrogen from deposition exceed supplies
from all other non-point sources to the Bay (e.g., agricultural
run-off, pastureland run-off, urban run-off), and only point source
inputs represent a greater input than deposition.

It is considered that the data from these studies are indicators
of the impact of anthropogenic nitrogen. Nevertheless, they are
insufficient to estimate critical loads for estuarine/marine systems.
It may well by that critical loads for these systems differ for
different climatic regions.

4.2.7 Appraisal and conclusions

Atmospheric deposition of nitrogen-containing and acidifying
compounds have an impact on soil and groundwater quality and on the
health and species composition of vegetation. Critical loads for
these effects are given in Table 26. Critical loads have been derived
using empirical data that relate loads directly to effects and
steady-state soil models that calculate critical loads from critical
chemical values for ion concentrations or ratios in foliage, soil
solution and groundwater (De Vries, 1993). Information on the effects
which occur when critical loads are exceeded is given in Table 27.
The values given in Tables 26 and 27 apply to forest vegetation in a
temperate climate. Whether they are representative of other climates
is uncertain. An overview of the critical loads for atmospheric
nitrogen deposition in a range of natural and semi-natural ecosystems
is given in chapter 8.

Effects of nitrogen and acidifying deposition on soil and
groundwater chemistry are most evident. Field studies showed that
deposited nitrogen is partly retained in the forest soil. Even at
high nitrogen deposition rates, as in the Netherlands, soil
acidification (which is mainly manifested by leaching of aluminium and
nitrate) is mainly caused by sulfur deposition. A relatively small
contribution of nitrogen to acidification does not imply that sulfur
has a larger impact on the health of forests, since the relationship

between soil acidification and forest health is not very clear. The
eutrophying impact of nitrogen is probably more important than the
acidifying impact at present.

There is substantial evidence from field surveys in several
countries of Europe that exceeding critical loads does not imply
dieback of the forest trees in the short term (one or two decades).
However, it does increase the risk of damage due to secondary stress
factors and it affects the long-term sustainability of forests. These
risks increase with the extent to which present loads exceed critical
loads and with the duration.

Table 26. Critical loads for acidity and nitrogen for forest ecosystems in temperate climates
(From: De Vries, 1993)

Effects Criteria^a Critical loads
(kg per ha per year)
(H for acidity;
N for eutrophication)

Coniferous Deciduous
forests forests

Acidity root damage; Al < 0.2 mol/m³ 1.1^b 1.4^b
inhibition of uptake; Al/Ca < 1.0 mol/mol 1.4^b 1.1^b
Al depletion; delta Al(OH)₃ = 0 mmol/m³ 1.2^b 1.3^b
Al pollution Al < 0.02 mol/m³ 0.5^b 0.3^b

Eutrophication inhibition of uptake of K; NH₄/K < 5 mol/mol 17-70^c
increased susceptibility; N < 1.8% 21-42^d
vegetation changes; NO₃ < 0.1 mol/m³ 7-20^e 11-20^e
nitrate pollution NO₃ < 0.4-0.8 mol/m³ 13-21^f 24-41^f

^a Background information on the various criteria is given in De Vries (1993). Critical Al and NO₃^-
concentrations and critical Al/Ca and NH₄/K ratios related to root damage, inhibition of nutrient
uptake and vegetation changes refer to the soil solution. Critical Al and NO₃^- concentrations
related to pollution refer to phreatic groundwater. Critical nitrogen contents related to an
increased risk for frost damage and diseases refer to the foliage.
^b Derived by a steady-state model. Al pollution refers to phreatic groundwater. For groundwater
used for the preparation of drinking-water, a critical acid load of 1600 mol/ha per year was derived
(De Vries, 1993).
^c Derived by a steady-state model assuming 50% nitrification in the mineral topsoil (second value).
^d Empirical data on the relation between nitrogen deposition and foliar nitrogen contents.

Table 26 (Con't)

^e The first value is derived by a steady-state model (worst case) and the second value is based
on empirical data.
^f Derived by a steady-state model using critical NO₃^- concentrations of 0.4 and 0.8 mol/m³,
respectively. NO₃^- pollution refers to phreatric groundwater. For deep groundwater, the
critical load will be higher because of denitrification.

Table 27. Possible and observed effects when critical loads are exceeded

Possible effects Average critical load Observed effects in the field
(kg per ha per year)^a

Root damage 1.1-1.4 H critical Al concentrations
exceeded greatly

Inhibition of 1.1-1.4 H critical Al/Ca ratios
uptake exceeded greatly

17-70 N critical NH₄/K ratios
exceeded slightly

Aluminium depletion 1.2-1.3 H depletion of secondary Al compounds

Groundwater 0.3-0.5 H critical Al concentrations
pollution exceeded greatly

13-21 N critical NO₃ concentrations
exceeded substantially

Increased 21-42 N critical N contents exceeded
susceptibility substantially; nutrient imbalances;
increased shoot/root ratios

Vegetation changes 7-20 N strong increase in nitrophilous species

^a H = acidity; N = total nitrogen

5. STUDIES OF THE EFFECTS OF NITROGEN OXIDES ON EXPERIMENTAL ANIMALS

5.1 Introduction

Most of the data reviewed in this chapter concerns the effects of
NO₂, since the bulk of the NO_x literature is on NO₂. The results
of the few comparative NO_x studies suggest that NO₂ is the most
toxic species studied so far. Most of the reports describe the
effects of NO₂ on the respiratory tract, but extrapulmonary effects
are also briefly discussed. A broad range of NO₂ concentrations has
been evaluated, but emphasis has been placed primarily on those
studies with exposure concentrations of 9400 µg/m³ (5.0 ppm) or less,
with the exception of studies on dosimetry and emphysema. Discussions
of available literature on the effects of other nitrogen compounds,
e.g., NO, HNO₃, and mixtures containing NO₂, also are included. WHO
(1987), Berglund et al. (1993) and US EPA (1993) comprise other
reviews of the animal toxicological literature concerning NO_x
effects.

5.2 Nitrogen dioxide

5.2.1 Dosimetry

It is generally agreed that effects of NO₂ observed in several
laboratory animal species can be qualitatively extrapolated to humans.
However, to extrapolate animal data quantitatively to humans,
knowledge of both dosimetry and species sensitivity must be
considered. Dosimetry refers to estimating the quantity of NO₂
absorbed by target sites within the respiratory tract. Even when two
species receive an identical local tissue/cellular dose, cellular
sensitivity to that dose is likely to show interspecies variability
due to differences in defence and repair mechanisms and other
physiological/metabolic parameters. Current knowledge of dosimetry is
more advanced than that of species sensitivity, impeding quantitative
animal-to-human extrapolation of effective NO₂ concentrations.
Nevertheless, information on dosimetry alone can be crucial to
interpretation of the data base. Both theoretical (modelling) and
experimental dosimetry studies are discussed below.

5.2.1.1 Respiratory tract dosimetry

The uptake of NO₂ in the upper respiratory tract (above the
larynx) has been experimentally studied in dogs, rats and rabbits.
The upper airways of dogs and rabbits exposed to 7520 to 77 080 µg/m³
(4.0 to 41.0 ppm) NO₂ removed 42.1% of the NO₂ drawn through
the nose (Yokoyama, 1968). The uptake of NO₂ by isolated
upper respiratory tracts of naive and previously exposed rats
(76 000 µg/m³, 40.4 ppm NO₂) was 28% and 25%, respectively (Cavanagh
& Morris, 1987). Kleinman & Mautz (1987) exposed dogs to 1880 or

9400 µg/m³ (1.0 or 5.0 ppm) NO₂ and found that more NO₂ was
absorbed in the upper respiratory tract with nasal breathing than with
oral breathing. In addition, the percentage uptake of NO₂ by the
upper respiratory tract decreased with increasing ventilation rates.
As ventilation increased up to four times resting values, NO₂ uptake
during nasal breathing decreased from approximately 85% to less than
80% and during oral breathing decreased from about 60% to approximately
45%. At rest, about 85% of the inhaled NO₂ entering the lungs was
absorbed by the lower respiratory tract; this increased to 100% with
high ventilation rates.

Miller et al. (1982) and Overton (1984) modelled NO₂ uptake in
the lower respiratory tract using the same dosimetry model described
by Miller et al. (1978) for ozone (O₃), but with the diffusion
coefficient and Henry's law constant appropriate to NO₂; however,
values of the latter constant and reaction chemistry were considered
uncertain. For all species modelled (i.e., rat, guinea-pig, rabbit