Hatfield J.L., Gandini A. (eds.) Nitrogen in the Environment

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364 Nitrogen in the Environment

at this scale since the reaction rate of NH

is such that little of the atmospheric NH

will react before it is deposited or leaves the modeling domain. Local-scale models

provide the link between investigations at the site level (e.g., environmental impact

assessments) and the predictions by larger-scale atmospheric dispersion models.

4 . EFFECTS OF AMMONIA

4.1 . Vegetation Effects

4.1.1 . Background

Effects of enhanced N deposition, both positive and negative, are widely

acknowledged ( Nihlgård, 1985 ; Bobbink and Heil, 1993 ; Lee and Caporn, 1998 ).

Figure 14 . ( Continued )

0.1–1

1– 5

5–10

0 1 2 km

Nature reserves

Farms

10–25

25–50

 50

Deposition (kg N / ha / year)

(b)

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Sources, Dispersion and Fate of Atmospheric Ammonia 365

But the significance of the form of the N deposition has only recently been ques-

tioned, driven by the need to show whether one N form is more toxic than another,

for mitigation purposes. Few experiments to date have addressed the effects of

different N forms, especially under field conditions. The reviews of impacts of

reduced N and specifically those of atmospheric NH

( Pearson and Stewart, 1993 ;

Fangmeier et al., 1994 ; Krupa, 2003 ) include few references to field or long-term

effects. Most studies on effects of gaseous NH

have been conducted over a few

weeks and have used continuous constant exposure scenarios that, arguably, under-

mine their application to predicting effects at the ecosystem level. Probably the

most informative data concerning field effects have come from transect studies

down wind of significant NH

sources ( Van Herk, 1999 ), particularly where these

have coincided with NH

concentration monitoring ( Pitcairn et al., 1995, 1998 ;

Wolseley et al., 2006 ).

Here we highlight some properties of NH

that make it a potential threat to

plant communities and ecosystems. Normally terrestrial plants acquire nutri-

ent N from the soil via their roots. In areas with elevated levels of atmospheric

there is also the potential for nutrient N acquisition in gaseous form via the

leaves. The consequences of foliar uptake and processing of an alkaline gas for cel-

lular functions appear to drive the deleterious effects of NH

on terrestrial plants.

Alkalinity is also thought to be a key driver for NH

effects on epiphytic lichens

( Van Herk, 2001 ). It should be remembered, however, that NH

is a natural prod-

uct of metabolic pathways responsible for photorespiration ( Keys, 1999 , and ref-

erences therein) and is generated during senescence, seed, and/or fruit production

( Marschner, 1995 ).

4.1.2 . Ecosystems at risk

In much of Europe, NH

effects on vegetation can be detected within 100 m o r

up to 1 km from farm sources, depending on both the source strength, the lifetime/

presence of the source and the sensitivity of the receptors. However, as far as the

receptors are concerned, annual atmospheric NH

concentrations or monthly

mean concentrations, based on passive chemical sampling, are time averaged

and may conceal larger short-term peaks in concentrations which contribute to

adverse effects. Exposure to significantly elevated NH

concentrations, for example,

when the wind is blowing from a nearby NH

source, will be intermittent and the

concentrations variable. Atmospheric NH

also impacts local ecosystems as NH



when the NH

deposits to plant surfaces, dissolves, and is washed through the

canopy into the soil where it can increase soil acidity and interfere with base cation

uptake ( Pearson and Stewart, 1993 ; Fangmeier et al., 1994 ; Krupa, 2003 ). Many

of the effects reported for NH

represent the combined effects of uptake through

shoots as NH

/NH



and roots as NH



. Transport of NH

as NH



aerosols

enables NH

to impact ecosystems remotely, when NH



is deposited in rain or

cloud water.

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366 Nitrogen in the Environment

4.1.3 . Mediation of effects at the ecosystem level

There are many reports of detrimental effects of atmospheric NH

on ecosys-

tems in rural areas dominated by livestock rearing, comprehensively summarized

by Krupa (2003) . Ecosystems on soils with low levels of mineral N and organisms

that rely on atmospheric deposition for nutrients, that is, mosses and lichens are

mostly at risk. The latter have limited detoxification capacity relative to their uptake

potential ( Pearson and Stewart, 1993 ). Negative effects of NH

can occur via direct

toxicity, when uptake exceeds detoxification capacity and, via N accumulation,

which increases the likelihood of detrimental interactions with other abiotic and

biotic stressors. Ammonia can affect slow growing species, especially under storey

species through eutrophication. This is when nitrophytes (N loving plants) use the

additional N to overgrow and shade out under storey species. Eutrophication affects

competition for resources, favoring fast growing species with fast N assimilation

rates. The net result has been large-scale changes in the flora of central Europe,

where cattle, pig, and poultry production are widespread and intensive. Species-rich

bogs and heathlands have been transformed into grasslands dominated by a few

nitrophytic grasses ( Heil and Diemont, 1983 ; Bobbink and Heil, 1993 ). Acidophytic

lichen epiphytes have disappeared from the bark of Oak trees and other previously

acid surfaces in the Netherlands and 95% of Denmark, though some of this effect

may be explained by the fall in SO

concentrations over the last few decades ( Van

Dobben and Ter Braak, 1998 ; Van Herk, 2001 ).

4.1.4 . Ammonia uptake and mediation of effects at the cellular level

In terrestrial plants, gas exchange occurs via stomata, embedded within the

cuticle (water-impermeable protective layer), which also facilitate water loss (i.e.,

transpiration). Stomata are lined with a water film linked to the apoplast of the mes-

ophyll cells (inner leaf tissue where photosynthesis occurs). Their main function

is to facilitate the uptake of CO

while minimizing water loss. Stomatal opening

and closing responds to temperature, light and moisture and has mostly evolved to

maximize CO

uptake ( Raven and Yin, 1998 ). At low photon flux and humidity

stomata will close, excluding NH

, while stomata open at high light and humid-

ity levels. The potential for absorption of NH

into the apoplast is thus driven by

physical and chemical equilibria, not in response to any biological demand for N.

Ammonia can also positively affect stomatal opening, increasing the potential for

uptake ( Hanstein and Felle, 1999 ). For these reasons, there is the potential for

to access plant cells in concentrations that could be toxic.

The main factor that provides some biological control to apoplastic ammonia

uptake is the stomatal ammonia compensation point. As already outlined (Section

3.2), this is the air concentration of NH

that is in equilibrium with the pH and

[NH



] of the apoplast. Temperature is one of the strongest drivers, with 

dou-

bling every 5°C increase for a given value of  ( [NH



]/[H

]). ( Sutton et al.,

2001c ). However,  is under clear biological control representing the product of

competing NH

supply production and consumption processes (e.g., Riedo et al.,

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Sources, Dispersion and Fate of Atmospheric Ammonia 367

2002 ). Thus the apoplastic solution constitutes a highly dynamic NH



pool, to

which NH



is constantly added via efflux from the mesophyll cells, uptake from

the xylem and/or from the atmosphere, and there are a range of transport systems

that operate to maintain its homeostasis. Thus  substantially increases following

nitrogen fertilization, as well as for senescent arable crops. Linked to such adapta-

tion to high internal NH

supply, agricultural plants may be relatively less sensitive

to atmospheric ammonia. By contrast, native species in semi-natural contexts may

not be adapted to deal with a high NH

supply to the apoplast. At the same time,

since 

(and hence 

) is smallest for such species, the input by NH

dry deposi-

tion will be largest.

The level of trans-cuticular uptake of gaseous NH

was earlier presumed to

be minimal ( Van Hove et al., 1987 ). Deposition to the cuticle occurs either to the

surface water film, which may be present even at relative humidities below 50%,

due to the presence of deliquescent salts ( Burkhardt and Eiden, 1994 ), or through

adsorption to surface waxes ( Sutton et al., 1995c ). However, plants can absorb sig-

nificant amounts of the NH

gas dissolved in solution as NH



through their aerial

shoots (e.g., Calluna vulgaris ), (  50% of the NH



in throughfall, Skiba et al.,

1986 ; Bobbink et al., 1992 ) and Sitka spruce ( Chiwa et al., 2003 ), and this has been

supported by the application of such experimental data in resistance models ( Sutton

et al., 1995c ). Uptake of NH



ions often leads to a loss of base cations as NH



uptake is achieved via cation exchange ( Bobbink et al., 1992 ).

To summarize, at the cellular level Van der Eerden (1982) and Fangmeier et al.

(1994) suggest that effects of NH

include direct toxicity (acutely high NH

con-

centrations), changes in metabolism and changes in carbon (C) sinks (assimilation

products). The effects of diverting C into CN rich products has repercussions for

stress resistance, while the wide ranging differential changes in growth of different

species have implications for species homeostasis, favoring species able to use N at

the expense of oligotrophic species, through changes in competitive advantage. The

consequences of these effects will vary with the species and habitat.

4.1.5 . Influences on plant sensitivity to ammonia and damage symptoms

Higher plant sensitivity to NH

is dependent on ecosystem, species, concentra-

tion, and climate ( WHO, 1997 ). Uptake of gaseous NH

occurs more readily than

that of other N forms (Scots pine seedlings, Pe¯rez-Soba and Van der Eerden, 1993 ).

Crop plants, which are fast growing with high N requirements, are generally the

least sensitive, while semi-natural plants, growing on nutrient poor soils in low N

environments, will be most sensitive. Krupa (2003) lists the sensitivities of  150

species. Forest trees with their large rough canopies provide a large potential area

for NH

deposition, although uptake depends on the physical and chemical proper-

ties of the leaf surface. Ammonia damage to trees is described in Krupa (2003) .

The tendency for N to accumulate in the foliage, which happens in most plants

subjected to NH

pollution, once the demands for growth have been satisfied, can

significantly increase susceptibility to stress. Because NH

is the most common

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368 Nitrogen in the Environment

N pollutant for vegetation growing near animal units, there is a high likelihood of

stress injury. However, all the causes and effects are not specific to NH

, but many

can be induced by other N forms too. Ammonia exposure even at low concentra-

tions can weaken plants through enhanced C turnover ( Leith et al., 2001 ), increas-

ing susceptibility to stress and accelerating senescence. Incidences of interactions

between NH

and abiotic stress are widely reported in Europe. In evergreen spe-

cies such as pines and Calluna , NH

exposure increases susceptibility to freezing

temperatures ( De Temmerman and Coosemans, 1989 ; Clement, 1996 ) and win-

ter desiccation ( Sheppard and Leith, 2002 ). The incidence of spring frost dam-

age in Calluna growing on an ombrotrophic peat bog and damage from pathogens

such as Phytophora and Botrytis was much higher when Calluna was exposed to

, compared with Calluna that received wet deposited N either in reduced or

oxidized forms ( Sheppard et al., 2007a ). De Temmerman and Coosemans (1989)

reported widespread fungal infection in pines damaged by frost, which appeared to

be linked to locally high NH

emissions (see also Roelofs et al., 1985 ). Incidence

of insect damage is far higher in plants affected by NH

and other N pollutants.

The widespread effects of the heather beetle in The Netherlands, which appeared

on an 8-year rotation, as opposed to 20 years, when NH

pollution reached its peak,

represents the worst described interaction ( Van der Eerden et al., 1991 ). Of all the

N pollutants, the interactions with abiotic stresses are worse with NH

, because

detoxification of NH

into amino acids changes C partitioning. High concentrations

of amino acids attract pests and pathogens and are formed at the expense of soluble

sugars which protect plants from intracellular freezing ( Van der Eerden, 1982 ; Van

der Eerden et al., 1990, 1998 ).

Many lichen species appear to be very sensitive to even small increases in NH

concentrations above cf. 1  g/m

( Wolseley et al., 2006 ). Current evidence sug-

gests that the absence of acidophytic lichens from twigs and trunks of acid-barked

trees, growing in NH

rich environments, is due to NH

neutralizing the bark pH

( Van Herk, 2001 ). The most obvious impact of NH

on epiphytic lichens is the loss

of acidophytic species and the proliferation of a less diverse group of nitrophytic

lichens ( Wolseley and James, 2002, 2004 ; Wolseley et al., 2006 ). However, given

that another survey in Western Oregon and Washington ( Geiser and Neitlich, 2007 )

related the disappearance of sensitive species to slightly elevated NH



concentra-

tions (10  m), the effects of NH

per se , can often be synonymous with those of



. Sheppard et al. (2004) compared the effects of field exposures to NH

with

those of NH

Cl ([4 mM] equivalent to 56 kg N/ha/year). They found that monthly

concentrations  2 0  g/m

decimated Cladonia portentosa ( Figure 15 ) popu-

lations in  1 year and that after 3 years the critical concentration had fallen to

 3  g/m

. Wet deposited NH



caused only restricted damage.

In mosses, NH

exposure can increase both the N and amino acid content of

ectohydric pleurocarpous mosses. Elevations in N and amino acid content have

been proposed as a well-coupled indicator of NH

-N deposition ( Pitcairn et al.,

2006 ). Moss species differ with respect to their N uptake, and presumably their

CH11-P374347.indd 368CH11-P374347.indd 368 5/31/2008 6:11:23 PM5/31/2008 6:11:23 PM

Sources, Dispersion and Fate of Atmospheric Ammonia 369

Figure 15 . Ammonia damage to Cladonia portentosa and Sphagnum capillifolium,

photographed in situ, growing along an ammonia release transect on an ombrotrophic

bog (Whim Bog in the Scottish Borders, UK). (a) healthy C. portentosa, (b) damaged

C. portentosa, (c) healthy S. capillifolium, (d) damaged S. capillifolium. Bleaching is a

common symptom of ammonia damage. Photographs by Ian D. Leith, CEH Edinburgh.

(a) (b)

tolerance ( Pitcairn et al., 2006 ). Some Sphagnum (bog mosses) appear to be very

sensitive, especially those that lack the red-orange pigments (carotenoids). The fact

that Sphagnum mosses ( Figure 15 ) are often wet makes them an ideal sink for NH

( Jones, 2006 ).

4.2 . Effects on Humans and Animals

High atmospheric concentrations of NH

are not just toxic to plants, but can

also lead to toxic effects on humans and animals through the formation of sec-

ondary particulate matter including a significant contribution from NH



. Large

amounts of atmospheric N in the form of fine particles are linked to respiratory and

cardiovascular diseases, asthma, reduced lung function, and overall mortality ( Pope

et al., 2002 ; Townsend et al., 2003 ). Malm et al. (2000) , quoted in Townsend et al.

2003 , suggest that particles containing NH



and NO



contribute up to 65% of the

total atmospheric particle load in southern California.

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370 Nitrogen in the Environment

Ammonia in combination with large concentrations of dust, as found in poultry

houses, is known to increase susceptibility to dust ( Donham and Cumro, 1999 ) in

poultry workers. The combined presence of NH

and poultry dust can also cause

acute or chronic respiratory disease including bronchitis, asthma, etc. ( Whyte,

2002 ) and changes in lung function ( Golbabaei and Islami, 2000 ). Golbabaei and

Islami (2000) also report symptoms such as burning eyes, light sensitivity, sore

throat, headache, and nausea due to exposure to very high levels of NH

. In the

UK, an occupational exposure standard (OES) has been set at 25 ppm (equivalent

cf. 14 mg/m

) for NH

, based on a time-weighted average of 7 h exposure every 24 h

( Health and Safety Executive, 2001 ; quoted in Whyte, 2002 ). A short-term exposure

limit of 35 ppm (c. 20 mg/m

) has been set for a 15-min period. It should be noted,

however, that these concentrations are more than three orders of magnitude larger

than typical ambient concentrations, and are generally relevant for indoor human

exposure at indoor source locations only.

Effects on animals are parallel to those on humans. For example, Kristensen

and Wathes (2000) found that exposure of poultry to high concentrations of NH

led to irritation of mucous membranes, eye infections, impaired vision, damage to

the respiratory tract, and increases susceptibility to diseases in intensive farming.

4.3 . Protection of Ecosystems from Ammonia – Critical Loads/Levels

During the 1980s and 1990s the initial focus on controlling transboundary air

pollution was through reducing national air pollution emissions by simple percent-

ages. Thus the first Sulfur Protocol ( UNECE, 1985 ) was sometimes known infor-

mally as the “ 30% club. ” Subsequently, it was recognized that the link between

these “ flat rate ” emission reductions and a reduction in the impacts of atmospheric

pollutants was not sufficiently strong, and that different ecosystems showed differ-

ent sensitivities to air concentrations and the deposition of pollutants. To account

for these differential sensitivities, and provide a tool to link air pollution abatement

with environmental condition, the approach of “ critical loads ” and “ critical levels ”

was developed. Critical loads are defined as the atmospheric deposition below

which effects on specific elements of the environment do not occur, according to

current knowledge (e.g., Nilsson and Grennfelt, 1988 ; Bull, 1991 ; Bull and Sutton,

1998 ). Critical levels are similar, but relate effects to estimated atmospheric con-

centrations. They are 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 ” ( Van der Eerden, 1982 ; Van der Eerden

et al., 1991, 1998 ). The critical loads/levels approach to environmental protec-

tion assumes that it is possible to define threshold deposition levels for pollutants,

below which specific ecosystems will not experience adverse effects. Maps of criti-

cal loads for ecosystems, developed to reflect the sensitivity of different ecosystems

to particular pollutants, may be compared with deposition maps for these pollutants.

This allows the identification of areas with critical loads exceedance which need

greater attention, and makes it possible to devise effective abatement strategies

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Sources, Dispersion and Fate of Atmospheric Ammonia 371

that reduce deposition to these areas ( Hornung et al., 1995 ). The UK, as part of the

wider approach in the UNECE CLRTAP, has adopted the critical loads concept as

a key element of its strategy to control atmospheric deposition ( DoE, 1990, 1991 ;

NEGTAP, 2001 ). The critical loads approach was initially developed for total acid-

ity (i.e., to combat “ acid rain ” ) and also for nutrient nitrogen (which includes both

reduced and oxidized nitrogen). Hence there is no critical load for NH

as such

(although critical levels apply). Empirical critical loads for N have been estimated

for a wide range of ecosystems ( Acherman and Bobbink, 2003 ) including, for

example, 10–20 kg N/ha/year for temperate forests or dry heaths, or 5–10 kg N/ha/

year for ombrotrophic bogs.

Until recently, the critical level for NH

did not include any differentiation

according to plants/ecosystems, with different critical levels being set according

to exposure period. Hence the annual critical level was 8  g NH

, the monthly

level was 23 mg/m

, the daily critical level was 270 ug/m

and the hourly level

3,300 ug/m

( Van der Eerden et al., 1991 ). These values have many shortcomings,

being based on short-term exposure experiments at uniformly high concentrations

under artificial conditions, ignoring the influence of environmental conditions that

drive NH

uptake and detoxification. Also, it was not apparent that for long-term

exposures (i.e.,  1 year) the previous exposure history can reduce the critical level

for effects ( Sheppard et al., 2007b , see www.ammonia-ws.ceh.ac.uk ). With these

values, and typical deposition velocities, it was noted by Burkhardt et al. (1998)

that the critical load for nitrogen would invariably be exceeded at concentrations

much less than the critical load. Hence, until very recently there was very little

interest in the relevance of the ammonia critical level. Under the Convention on

LRTAP, an Expert Workshop held in Edinburgh ( www.ammonia-ws.ceh.ac.uk ;

UNECE, 2007 ) concluded that the critical level was set too high, and new values

were adopted by the Convention. The revised critical levels represent long term

mean NH

air concentrations, suitable for protection over several years, although

the Expert Workshop concluded that they could not be assumed to provide pro-

tection for longer than 20–30 years. The new critical level values are 1  g/m

for

lichens, bryophytes, and ecosystems where these plant groups are critical to eco-

system integrity and 3 (2–4)  g/m

for other grasslands, shrublands, and woodland

vegetation.

Due to the varying affinity and compensation points of ammonia for different

habitats, expressed in differences in mean deposition velocities, the rates of ammonia

deposition vary greatly between habitat types. This means that maps of NH

dry dep-

osition need to be interpreted with care, noting whether they refer to inputs to specific

habitat types (e.g., woodland, shrublands, croplands) or net dry deposition averaged

over entire grid squares. For the purpose of assessing critical loads exceedance, the

deposition figures for the relevant habitats need to be used, rather than the grid aver-

ages. Such maps (e.g., NEGTAP, 2001 ) show that for the most sensitive habitats in

the UK, there is substantial exceedance of the critical loads for nitrogen, with the total

deposition to these habitats being dominated by reduced nitrogen.

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372 Nitrogen in the Environment

In the national UK models, deposition is estimated at 5-km grid resolution and

critical loads at 1-km resolution. Although this shows a substantial variation across

the country, the deposition patterns are nevertheless artificially smoothed compared

with the real variability. It is possible to model this fine-scale spatial variability in

critical loads exceedance for specific examples at a local level. This is illustrated

for the landscape study area investigated by Dragosits et al. (2002, 2006) . In Figure

16 , critical loads exceedance is modeled at a 25-m grid resolution, showing the

major sub-5 km grid variability that occurs as a result of local NH

dispersion and

differences in rates of NH

dry deposition. With national modeling, the area in

Figure 16 would generally not exceed the critical load. This indicates potential large

No exceedance

0–1

1– 5

0 1 2 km

Farms

5–10

 10

Exceedance of critical loads for Nitrogen (kg N / ha / yr)

Figure 16 . Critical loads exceedance of nutrient N for woodland and heathland in a

study area in central England (kg N/ha/year).

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Sources, Dispersion and Fate of Atmospheric Ammonia 373

underestimates of sensitive receptors at risk from NH

pollution, both at a national

level as well as at coarser levels such as the EMEP maps for Europe at a 50 k m 

50 km grid resolution.

4.4 . Biomonitors for Ammonia

Biomonitoring makes use of living biological organisms to monitor the impacts

of N deposition. A range of biomonitors for N have been comprehensively evalu-

ated for distinguishing between different N forms for UK application ( Sutton et al.,

2004a, 2004c ; Leith et al., 2005 ). Biomonitors can provide a measure of the bio-

logically available NH

, mediated as a change in biomass (indicator plants) and/or

chemical N concentration (%N dry weight and water soluble NH



) and biochemi-

cal composition (amino acids). Unlike chemical monitoring, not all biomonitors

effectively discriminate between different N forms, despite showing form prefer-

ences ( Marschner, 1995 ). Plants may already be growing at a site (e.g., pleurocar-

pous mosses) or may be transplanted into a site (e.g., moss bags or standardized

indicator plants) ( Leith et al., 2005 ). The majority of biomonitors evaluated by

Leith et al. (2005) gave the best correlations with dry deposition (i.e., NH

concen-

trations) rather than with N deposition per se. The degree of N enhancement is spe-

cies specific, reflecting differences in species uptake capacities.

Lichen diversity indices, which do not rely on the presence of specific species,

based on the groupings of Van Herk (1999) , offer the most sensitive and robust

indicator for NH

, and this has been extended for lichens on twigs by Wolseley et al.

(2006) . The composition of some macrolichen assemblages on tree trunks and

twigs is a function of bark pH. Nitrophytic lichens increase in the presence of NH

Acidophytes, with a requirement for acid bark, disappear in response to increas-

ing NH

concentrations, especially species growing on twigs, which are unable to

tolerate even small increases in pH. Nitrophytic lichens start to dominate epiphytic

communities once the NH

concentration exceeds 3  g/m

( Leith et al., 2005 ). The

Ellenberg index for higher plants ( Ellenberg, 1979 ), which was devised in Europe

in the 1970s to describe individual species responses to a range of ecological condi-

tions, including N availability, is the least sensitive indicator of anthropogenic N

from point sources (both for reduced and oxidized N). The Ellenberg N index does

indicate eutrophication, but was developed for differences in soil available N rather

than anthropogenically manipulated N, which is available for N uptake above and

below ground.

5 . METHODS TO REDUCE EMISSIONS AND IMPACTS OF AMMONIA

5.1 . Introduction

Emissions of NH

can lead to impacts on semi-natural ecosystems and human

health; it is therefore desirable to reduce both the amount that is emitted into

the atmosphere and/or the amount that reaches sensitive receptors (ecosystems

and humans). This has led to policies aimed at emission reduction at national and

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