Ammonia: emission, atmospheric transport and deposition

New Phytol. (1998), 139, 27–48

Ammonia : emission, atmospheric transport and deposition B  W I L L E M A. H. A S M A N"*, M A R K A. S U T T O N#    J A N K. S C H J Ø R R I N G$ " National Environmental Research Institute, Frederiksborgvej 399, DK-4000 Roskilde, Denmark # Institute of Terrestrial Ecology, Bush Estate, Penicuik, Midlothian, EH26 0QB, Scotland, UK $ Plant Nutrition Laboratory and Centre for Ecology and Environment, Royal Veterinary and Agricultural University, Thorvaldsensvej 40, DK-1871 Frederiksberg C, Copenhagen, Denmark (Received 5 September 1997)  The global emission of ammonia (NH ) is about 54 Mt N. The major global sources are excreta from domestic $ animals and fertilizers, but oceans, biomass burning and crops are also important. About 60 % of the global NH $ emission is estimated to come from anthropogenic sources. NH -N emissions are of the same order as the NOx-N $ emissions on both global and European scales. Emitted NH returns to the surface mainly in the form of dry $ deposition of NH and wet deposition of ammonium (NH +). In countries with high NH emission densities, dry $ % $ deposition of NH from local sources and wet deposition of NH + from remote sources dominate the deposition. $ % In countries with low NH emission densities only wet deposition of NH + from remote sources dominates the $ % deposition. Surface exchange of NH is essentially bi-directional, depending on the NH compensation point $ $ concentration of the vegetation and the airborne concentration. In general, the compensation point is larger for agricultural than semi-natural plants, and varies with plant growth stage. According to basic thermodynamics the leaf tissue or stomatal compensation point of NH doubles for each increase of 5 °C. However, exchange of NH $ $ does not only occur through the stomata, but it can also be deposited to leaf surfaces, as well as emitted back to the atmosphere from drying leaf surfaces. Atmospheric transport and deposition models can be used to interpolate NH concentrations and depositions in space and time, to calculate import}export balances and to estimate past $ or future situations. Adverse effects on sensitive ecosystems caused by high N deposition can be reduced by lowering the emissions and, to a limited extent, also by removing sources close to the ecosystem to be protected. Key words : Ammonia, dry deposition, compensation point, wet deposition, deposition models, ecosystem protection, pollution emissions.

 Ammonia (NH ), and its reaction product $ ammonium (NH +), are important atmospheric com% ponents, NH being the most abundant alkaline $ component in the atmosphere. A substantial part of the acid generated in the atmosphere by the oxidation of sulphur dioxide (SO ), and nitrogen oxides (NOx), # is neutralized by NH . As a result NH + is a major $ % component of atmospheric aerosols and precipitation. NH and NH + (collectively termed NHx) are $ % also nutrients. Excessive deposition of these components to oligotrophic ecosystems might lead to a * To whom correspondence should be addressed. E-mail : wa!dma.dk

shift of plant species to more nitrophilic ones (Bobbink et al., 1992). When NHx is deposited, nitrification can occur leading to the formation of nitric acid in the soil. This acidification might lead to K+ and Mg#+ deficiencies in vegetation followed by severe stress and, as a consequence, premature shedding of needles of coniferous trees (Roelofs et al., 1985). The species composition in forest undergrowth could also be changed (van der Eerden, de Vries & van Dobben, 1998). Nitrogen deposition is not only caused by NH $ and NH +, but also by NOx and its reaction products : % gaseous nitric acid (HNO ), and particulate NO −. $ $ Expressed as emissions of N, both groups of N components are similarly important in Europe,

28

W. A. H. Asman and others

although, as NH is preferentially deposited in semi$ natural ecosystems, NHx will often exert the larger ecological impact (Sutton, Asman & Schjørring, 1994). This paper discusses processes that influence the atmospheric behaviour of NH and NH +. Two $ % forms of NHx deposition are important : dry deposition of NH which is most important close to a $ source, and wet deposition of NH +, which is most % important some distance downwind from the source. Wet deposition of NH + to large areas, such as % countries, can be estimated relatively easily from measurements at a limited number of sites. Dry deposition of NH shows a high spatial variability, $ because it depends on the many scattered local sources of NH . Therefore measurements from very $ many sites would be needed to get a reliable estimate of the deposition of NH to a country. Moreover, the $ continuous measurement of dry deposition requires many resources. For all these reasons, estimates of dry deposition of NH to a country are often based $ mainly on the results of atmospheric transport and deposition models. This is why, in this paper, the atmospheric processes of emission, reaction, deposition and atmospheric transport are discussed with the emphasis on modelling the processes in such a way that they can be incorporated in atmospheric transport models.

 Sources of ammonia and emission processes Virtually all NHx emission occurs in the form of NH , with NH + in the atmosphere originating from $ % reactions of NH . A key feature of the exchange of $

Table 1. Global NH emissions from Bouwman et al. $ (1997) Source Domestic animals Human excrements Synthetic fertilizers Agricultural crops Biomass burning in agriculture and biofuel use Fossil fuel combustion Industry Subtotal anthropogenic emissions Wild animals Undisturbed ecosystems Biomass burning in natural ecosystems Sea Subtotal natural emissions Total

Emission (Mt N yr−") 21±7 2±6 9±0 3±6 2±7 0±1 0±2 39±9 0±1 2±4 3±2 8±2 13±9 53±8

NH between the biosphere and the atmosphere is its $ bi-directional nature. For NH this means that the $ flux will depend on the difference between the surface concentration and the concentration in the air overlying that surface (χa). Emission occurs if the first concentration exceeds the latter (see also the section on dry deposition}exchange). This will usually be the case if the surface contains high concentrations of NH + at alkaline pH, such as animal manure or % fertilizer. In the case of seawater, or for plants, air concentrations might be of the same order as surface concentrations, so that the magnitude of emissions is affected by the prevailing air concentrations. The total global NH emission has been estimated $ to be of the order of 50 Mt N yr−" (Schlesinger & Hartley, 1992 ; Bouwman et al., 1997) and is presented in Table 1. The map in Figure 1 shows that the emission density is highest in India, China and western Europe. The European NH emission (land only, excluding $ the former USSR countries) has been estimated at c. 4±5 Mt N yr−", arising mainly from animal husbandry (80 %) and synthetic fertilizers (Asman, 1992 ; ECETOC, 1994). It has been estimated that European NH emissions have doubled since 1950 $ (Asman, Drukker & Janssen, 1988), caused by the increase in animal numbers and the increasing application of synthetic fertilizer, and they might continue to rise owing to increasing intensification of livestock production (Sutton, Pitcairn & Fowler, 1993 c). By comparison, the total global NOx emission is estimated to be of the order of 40 Mt N yr−" (Lee et al., 1997). This is about the same order as the NH $ emission. The NOx emission in Europe (land only, excluding the former USSR countries) is estimated as 5±1 Mt N yr−", which is also comparable to the NH emission for the same area (Pacyna, Larssen & $ Semb, 1991). The uncertainty in the global and European NH and NOx emissions is at least of the $ order of ³40 %. However, the uncertainty in some natural NH and NOx emissions can be much larger, $ in many cases by more than a factor of 4 (³). Figure 2 shows that there are large variations in the annually averaged NH emissions across Europe. $ The highest values occur in the Netherlands, Belgium and Denmark. The average emission density in the Netherlands is c. 50 kg N ha−" yr−", compared with 1 kg N ha−" yr−" for Sweden with its lower population density of both humans and livestock. Ammonia emissions from livestock occur during animal housing, manure storage and spreading (the emission occurs primarily after and not during spreading) and when animals are grazing. The amount of NH emitted depends on : the N content $ and amino acid content of the animal food, the conversion efficiency of the N in animal food to that in milk, meat or eggs (which determines the amount

Emission, transport and deposition of ammonia g N m–2 yr –1

29

90

90

60

60

30

30

0

0

0–0·05 0·05–0·1

0·5–1 1–2 2–3 >3

– 30

– 30

– 60

– 60

– 90

– 90 – 180 – 150 – 120

Degrees latitude

0·1–0·5

– 90

– 60

– 30 0 30 Degrees longitude

60

90

120

150

180

Figure 1. Global NH emission on a 1¬1 degree grid (g N m−# yr−" ; (1 g N m−# yr−" ¯ 10 kg N ha−" yr−")) $ (Bouwman et al. 1997).

of N left in the manure). Emissions also depend on the type of livestock and their age}weight, and on density (number ha−"). Ammonia emissions from manure occur when it is exposed to the air. The emission rate depends on the airborne NH concentration (χmf) just above the $ manure that is in equilibrium with the NHx concentration in the manure, the airborne NH $ concentration at a certain reference height over the manure (χa) and on air turbulence. The value of χmf depends on the temperature and pH of the manure and the underlying surface. In the field the air turbulence is determined by meteorological factors and is well known. Over manure in housings and in storage facilities, however, the environment is poorly characterized with respect to turbulence and the airborne concentration, and consequently the emission rate cannot be calculated from the turbulence and χmf. In the case of buildings with mechanical ventilation it is possible to measure the emission rate from the vents, but measurements become much more complex for buildings with natural ventilation and from storage facilities (Demmers et al., 1998). The fact that the turbulence and NH concentration over manure in housings and $ storage facilities is not known also makes it difficult to model the emission rate for a specific situation over a specific period. The emission over a longer period (c. 6 months) can, however, also be estimated from the difference in the NHx concentration in manure after production and before spreading. Emission after spreading is also influenced by the viscosity and dry matter content of the manure and the amount of manure applied per unit land area. The application method determines the extent to

which the manure is exposed to the air and hence also influences the emission. Ammonia emissions from synthetic fertilizers likewise depend on χmf and the turbulence. χmf is dependent on the type of fertilizer, varying by up to a factor of 10 (Asman, 1992). Until now, the annually averaged emission has been estimated using livestock numbers and an emission factor per animal, or the application rate of fertilizers with an emission factor per tonne. Figures 1–3 are constructed in such a way. However, these calculations have ignored the variation caused by meteorology. Diurnal variations in the emission rate are also likely, with an afternoon peak related to warmer temperatures and maximum turbulence (Asman, 1992). In addition to natural seasonal effects, farming operations vary over the season, giving rise to peak emissions in spring and autumn coinciding with manure spreading (Asman, 1992). In Denmark and the UK, NH concentrations $ are also high during the summer, possibly reflecting emissions from plants. Distribution of ammonia emissions on different spatial scales A key aspect in estimating the magnitude of NH $ emissions is establishing the spatial distribution of the sources. This is necessary to provide maps of NH emission as inputs to atmospheric transport $ and deposition models. Such models can then be used for risk assessment by comparing modelled air concentrations and deposition with the locations of sensitive areas. The first mapped estimates of NH $ emissions were provided for the Netherlands (Buijsman, Maas & Asman, 1984), England and

30


> 60 40–60 20–40 10–20 5–10 2·5–5 0–2·5

Figure 2. European NH emissions (kg N ha−" yr−") (Asman, 1992). $

Wales (ApSimon, Kruse & Bell, 1987 ; Kruse, ApSimon & Bell, 1989) and for Europe (Buijsman, Maas & Asman, 1987 ; Asman 1992). The resolution of NH emissions has been limited $ by both the availability of data on source distributions as well as the capabilities of atmospheric models. On a global scale inventories have been presented on 10¬10 degree scale (Dentener & Crutzen, 1994) and on a 1¬1 degree scale (Bouwman et al., 1997). On a European scale inventories have been provided with a grid resolution of 75 km (Asman, 1992), whereas on national scales maps have been constructed at grid resolutions down to 5 km for the Netherlands (Buijsman et al., 1984 ; Erisman, 1989), Belgium (Asman, 1992), Denmark (Asman, 1990), Switzerland (Rihm, Kunz & Engel, 1992) and Great Britain (Sutton et al., 1995 c ; Dragosits, Sutton & Place, 1996). The importance of scale in modelling NH $ emission inventories is illustrated in Figures 1–3, which show the global, European and the UK

national maps, at 1 degree 75 km, and 5-km grid resolutions respectively. On a national scale, where source distributions are based on parish or municipality data, this often provides the finest reasonable balance between spatial uncertainty and resolution. Nevertheless, environmental effects can still dominate at scales much finer than 5 km. As part of the work to improve 5-km resolution emissions in Britain, the models developed (referenced above) reallocate variable scale parish level emissions to a 1km level in relation to landcover and the statistical distribution of sources, then re-aggregate to the 5km level. Although the 1-km emission maps show a great uncertainty where livestock is farmed intensively, such as where pigs and poultry are housed, the approach has proved useful in providing improved values where livestock numbers are determined by grazing area and manure is spread over the land (Dragosits et al., unpublished ; Rihm, Kunz & Engel, 1992). For an even finer level of resolution spatial

Emission, transport and deposition of ammonia

31

AMMONIA EMISSIONS FROM AGRICULTURE AND NON-AGRICULTURAL SOURCES IN THE UNITED KINGDOM 1988

Ammonia emissions in kg ha–1 NH3-N 0·0–2·5 > 2·5–5·0 > 5·0–10·0 > 10·0–20·0 > 20·0–30·0 > 30·0–40·0 > 40·0–50·0 > 50·0

0

100

200 km

Figure 3. Total estimated ammonia emissions for Great Britain mapped with a 5-km grid resolution, including agricultural and non-agricultural sources (Dragosits, Sutton & Place, 1996). The map is a re-aggregation of a statistical 1 km distribution of ammonia emissions combining agricultural source and land-use datasets.

inventories of NH based on information provided at $ a field by field and farm level can be used. Such inventories have been developed for sample areas in the Netherlands (Boermans & Erisman, 1993) and are being developed in Denmark and the UK (Bouwman & Asman, 1997 ; Dragosits et al., 1997). Such local scale inventories allow emissions to be estimated at the spatial scale of environmental effects, and these might be used in local scale atmospheric transport models to address questions of spatial variability of deposition and impacts.   Gaseous NH reacts with acidic aerosols that contain $ sulphuric acid (H SO ), and also with gaseous nitric # % acid (HNO ), or gaseous hydrochloric acid (HCl). $

The reaction rates depend on the acid concentration, humidity and temperature, and NH +-containing % aerosols are formed (Stelson, Friedlander & Seinfeld, 1979 ; Huntzicker, Cary & Ling, 1980). Reverse reactions also occur, but are only important with HNO and HCl. The reaction rates for NH show $ $ temporal and spatial variations and are also dependent on the height above ground level. In principle it is possible to estimate these reaction rates from field experiments using the changes in the ratio NH : NH + in air as a function of the vertical or $ % horizontal distance (Erisman et al., 1988 ; Harrison & Kitto, 1992). However, many assumptions have to be made and results vary. Alternatively the reaction rate can be estimated by changing the rate function in atmospheric transport models until the best agreement is obtained with measured concentrations


32

va

va

Ra

va

va

Ft Ra

Ra

Ra

Rs Rb

Rb

Rb

Rb

vs > 0

vc

vc Rw Rc

Rs

Rw

vs > 0 (a)

(b)

Rg

Rs

vs > 0 (c)

vg > 0 (d)

Figure 4. Resistance models applied to describe the exchange of ammonia with vegetation. The net flux (Ft) is defined by Ohm’s law in relation to air concentrations (χa) and component resistances, including the atmospheric resistance (Ra) and the laminar boundary layer resistance (Rb). (a) The classical ‘ canopy or surface resistance ’ (Rc) model, defines deposition only on the assumption that the concentration at the surface is zero. Rc can be distinguished into several parallel component resistances (b–d ). (b) The classical ‘ stomatal compensation point ’ (χs) model allows bi-directional exchange via stomata (Rs). (c) A simple ‘ canopy compensation point model ’ calculating the canopy concentration (χc), as a result of competing stomatal exchange and deposition to leaf cuticles (Rw). (d ) A more complex canopy-exchange model accounting for exchange with a ground concentration (χg) and diffusion through the canopy air space (Rg).

of NH and NH + in air and NH + in precipitation. $ % % Using this technique Asman & Janssen (1987) found a pseudo-first-order reaction rate of 8¬10−& s−" (c. 30 % h−"), which also gives reasonable results on a global scale (Dentener & Crutzen, 1994). This rate is appropriate as a first estimate to calculate an annually averaged conversion, but cannot be used to describe situations over shorter time periods or for one particular site. The reaction rate has a large influence on the deposition of NHx as a function of distance downwind from the source because NH is usually dry $ deposited at a high rate, whereas NH + is deposited % much more slowly. The uncertainty in the reaction rate exerts a relatively small effect on the total NHx deposition at one site, because some effects compensate each other. At faster rates the contribution from nearby sources becomes less important relative to the contribution from more remote sources, favouring deposition in the form of wet deposition instead of dry deposition. A sensitivity analysis, where the reaction rate was varied uniformly over Europe, showed that the difference in total NHx deposition to Dutch and Swedish sites varied by only approx. 10 % even if the average reaction rate was varied by a factor of 4 (Asman & van Jaarsveld, 1992). The modelled total NHx deposition would show larger variations, however, if the reaction rate in the model showed spatial and temporal variations or if the deposition caused by only one source were calculated.

            }         Models Many pollutants, such as SO , NO and O , are # # $ normally absorbed by land and water surfaces. By contrast, for NH the presence of significant liquid$ phase concentrations in the environment results in a variable surface potential for emission, allowing both upward or downward fluxes depending on the relative magnitude of the surface NH concentration $ and NH concentrations in the atmosphere (χa). The $ gas phase concentration at equilibrium with liquid NH concentrations in plants is often termed the $ ammonia compensation point (χcp), since when χa ¯ χcp there is no net flux, and the biological processes liberating NHx at the surface balance consuming processes. The idea of a compensation point for NH $ was first applied to plant canopies to parallel to the concept of CO exchange (Farquhar et al., 1980 ; # Lemon & van Houtte, 1980) and has since been utilized by many others (Dabney & Bouldin, 1990 ; Schjørring, 1991 ; Sutton et al., 1993 c). The compensation point concept has also recently been applied to bi-directional exchange processes over seawater where χcp can be controlled largely by phytoplankton activity (Asman, Harrison & Ottley, 1994 ; Barrett, 1998). The first model used to consider the NH $ compensation point over vegetation assumed that the exchange process takes place through the stomata

— 1 to 50 — 1 to 11 1 to 30 — —

®1300 to 300 ®950 to 650 ®1000 to 700 ®60 to ®0±2 ®400 to 550 ®33 to ®0±2 ®40 to 20 ®150 to 300 ®20 to 20*

0±8–27 0±2–25 5 (mean) 0±1–1±2 0±5–40 0±2–3±7 0±2–9 1–20 0±05–16

Spindler, Rolle & Gru$ ner (1996) Sutton et al. (1997 a)

Sutton et al. (1995 a)

Sutton, Fowler & Moncrieff (1993 b) Erisman & Wyers (1993) Sutton, Fowler & Moncrieff (1993 b)

— 2–14 (for 5–20 °C)

—

—

—

—

Short grass : 8–50 Long grass : 2–12±5 (for 5–20 °C) 0±4 to 2±5 (for 5–20 °C)

0±2–4±5

0 to 1000 — 40 to 2000 (mean : 125) —

Short grass : 12620 Long grass : 3150 630

—

— —

—

8500

— 3470

—

Estimates of Apoplastic stomatal compensation [NH +]}[H+] % (M}M) point, χs (µg m−$)

0 to 200 (medians for 5–34 (for 5–20 °C) humidity classes) ®20 to 2000 (mean : 6) —

®45 to 2450 —

—

Canopy resistances, Rc (s m−")

Negative Ft indicates deposition ; negative Vd indicates emission. Rc values given for deposition fluxes only ; negative values indicate deposition faster than is possible by turbulence, often due to scatter in the measurements. * Measurements immediately before anthesis.

Wheat

Unfertilized calcareous grassland Low N agricultural grassland High N agricultural grassland

Moorland

Duyzer et al. (1994) Sutton, Schjørring & Wyers (1995 d ) Wyers & Erisman (1998)

®125 to 200 (mean : 26) ®180 to 100— —

Andersen et al. (1993)

Forest

®300 to 50

0±05–4

Authors

Canopy

Deposition velocities, Vd (mm s−")

NH concs. $ (µg m−$)

Range of fluxes, Ft (ng m−# s−")

Table 2. Example micrometeorological measurements illustrating the variability in ammonia exchange fluxes recorded for different vegetation types

Emission, transport and deposition of ammonia 33

34

W. A. H. Asman and others Bi-directional diffusive transport of ammonia Above canopy chemical interactions with HNO3 HCl? Flower/seed layer apoplast

(fomation/destruction of ammonium aerosois)

Flower/seed layer surfaces

Upper canopy apoplast

Upper canopy surfaces

Lower canopy apoplast

Lower canopy surfaces

Parallel fluxes of other pollutants SO2, HNO3, HCl, O3 and interactions on canopy surfaces

Within canopy chemical interactions?

Litter decomposition Soil

Figure 5. Schematic of ammonia exchange processes within and above a plant canopy. Although in many situations the canopy can be considered in terms of bulk apoplastic}stomatal exchange and bulk leaf cuticle exchange, in complex canopies significant differences occur between soil, litter, lower canopy, upper canopy and reproductive layers. To this might be added the need for models to consider interactions on leaf cuticles with other pollutants as well as gas–aerosol interactions above and within the canopy.

in relation to the NH + concentration in the leaf % apoplast (Farquhar et al., 1980). On the basis of the standard thermodynamics of NH solubility and $ dissociation in solution, the gaseous NH concen$ tration is known to be a temperature dependent function of the ratio [NH +]}[H+] in the apoplast. % The flux (Ft) was therefore assumed as : Ft ¯ (χs®χa)}(RaRbRs),

(1)

where the three component resistances act in series, these being the turbulent atmospheric resistance (Ra), the quasi-laminar boundary layer resistance (Rb) and the resistance for diffusion through stomata (Rs) (Farquhar et al., 1980 ; Schjørring, Husted & Mattsson, 1998) and χs is the stomatal compensation point (see Fig. 4 b). This model works well in a number of situations, and particularly in the laboratory. However, in the field the frequent occurrence of high humidity or wet conditions leads to significant fluxes to plant surfaces, requiring the application of more complex models. A commonly applied model is based on the deposition velocity (Vd) (Fig. 4 a). This permits uptake onto different canopy surfaces, but it does not permit emission from the canopy to be treated satisfactorily. The model assumes that the concentration at the absorbing canopy surface is zero (χs ¯ 0), and any excess resistance or concentration limitation at the surface is expressed by a surface or canopy resistance (Rc) : Ft ¯®χa}(RaRbRc) ¯®χa Vd.

(2)

This model usefully describes exchange in situations where emission is rare, but its use is limited. Many measurements of NH fluxes have $ been reported in terms of Vd, but because of the bidirectional nature of the exchange, as well as the variability of atmospheric resistances, it is generally more helpful to summarize results in terms of canopy resistances (Rc), compensation point estimates (χs) and the apoplastic ratio of [NH +] : [H+], as well as % the range of fluxes measured (Table 2). Uncertainties In addition to foliar emissions, in some situations soil surfaces, leaf litter or reproductive tissues within canopies must be considered as significant sources of NH emission. Figure 5 provides a schematic view of $ a plant canopy and potential sites for NH exchange. $ In complex canopies, where each layer acts very differently, this would necessitate the development of multi-layer models to describe net fluxes (Sutton et al., 1997 b ; Nemitz et al., 1997). Further uncertainties include the potential for interactions with other atmospheric pollutants such as SO , O , # $ HNO and HCl on plant surfaces (Erisman & Wyers, $ 1993 ; Sutton et al., 1994 ; Sutton, Schjørring & Wyers, 1995 d ; Cape et al., 1998), or with HNO and $ HCl in the air above or within canopies (Brost, Delany & Huebert, 1988 ; Kramm & Dlugi, 1994 ; Nemitz et al., 1996). The major uncertainties linked with each of these issues are considered in the following sections.


35

Exchange of ammonia with leaf tissues

4

Given the relationship χs ¯ f(T)[NH +]}[H+], both % the apoplastic NH + concentration and pH are % critical in defining the NH compensation point. $ The apoplastic NH + concentration is very sensitive % to leaf N status and external N supply. Hence much larger values are found for agricultural species receiving high N inputs than for those receiving little N, and semi-natural species. For example, in leaves of Brassica napus ssp. napus for plants receiving lowN treatment the apoplastic NH + concentration was % c. 0±2 m, whereas in leaves from plants receiving high-N the corresponding concentration was tenfold higher (Husted & Schjørring, 1996). In Hordeum vulgare, leaf apoplastic NH + concentrations ranged % from 0±04³0±003 at an external NH + concentration % of 0 m to 2±28³0±42 at 10 m (Mattsson, Husted & Schjørring, 1998). Apoplastic NH + concentrations % in the two grass species Arrhenatherum elatius and Bromus erectus growing with a low N supply in the rooting medium and exposed to atmospheric NH $ mole fractions of 0 or 70 nmol mol−" for 25 d were in all cases ! 0±15 m (Hanstein, Mattsson & Schjørring, unpublished). Apoplastic pH has been estimated to lie mostly between 5±5 and 6±5 (Husted & Schjørring, 1995 ; Mattsson & Schjørring, 1996). Hoffman, Planker & Mengel (1992) measured apoplastic pH values by means of fluorescence and observed lower pH values for plants given NH + (pH 5±8) than for those given % NO − (pH 6±3). These results contradict those of $ Dannel, Pfeffer & Marschner (1995) and Mu$ hling & Sattelmacher (1995) who found no difference in apoplastic pH with NO − or NH + supply. The $ % extent of apoplastic pH buffering, which affects both the regulation of χs and the interpretation of measurements, needs to be resolved. In B. napus, apoplastic air and water volume increased with leaf age, and apoplastic NH + con% centration declined from 1±3³0±08 to 1±1³0±24 m with ageing (Husted & Schjørring, 1995). By comparison, apoplastic pH (5±8³0±2) was not affected by leaf age. Senescing leaves have a much higher bulk tissue NH + concentration than young % leaves (Husted, Mattson & Schjørring, 1996 ; Schjørring et al., 1998), but extracting apoplastic solution from senescing leaves is difficult and so far apoplastic NH + concentrations and pH have not % been measured directly. The apoplastic NH + concentration is also sen% sitive to the activity of glutamine synthetase (GS) the key enzyme in NH + assimilation : Inhibition of GS % by addition of methionine sulphoxime (MSO) increased the apoplastic NH + concentration in B. % napus even up to 25 m (Husted & Schjørring, 1995). Similarly, barley mutants with only 66 % of normal chloroplastic GS activity had higher apoplastic NH + concentrations than did the wild type %

3 NH3 exchange (nmol m–2 s–1)

Emission

2 1 0 –1 –2 –3 Absorption –4 10

15

20 25 30 Leaf temperature (°C)

35

40

Figure 6. Flux of atmospheric ammonia in oilseed rape (Brassica napus) using a cuvette gas exchange system under laboratory conditions. The substantial temperature response of the compensation point (χs) led to a change from deposition to emission at approx. 30 °C. Ammonia concentrations in the chamber were E 12 µg m−$ (0±7 µmol m−$), with an irradiance of 550 µmol m−# s−", and 20 % r.h. The experiment was performed in duplicate over 2 consecutive days. – – –, theoretical NH flux in response to $ changing leaf temperature at an NH concentration of 12 $ µg m−$, referenced to apoplastic [NH +]}[H+] calculated % from the measurements at 32 °C and assuming a temperature-independent leaf NH conductance (0±12 mol $ m−" s−").

(1±08³0±26 vs. 0±34³0±08  ; Mattsson et al., 1997). The apoplastic pH in the mutant was also somewhat higher than that in the wild type plants (5±86³0±14 against 5±32³0±06), causing its NH compensation $ point to be 7±72 nmol mol−" compared with only 0±75 nmol mol−" in the wild type. The temperature response of the stomatal compensation point cs is substantial, leading to an approximate doubling of the concentration of dissolved NH for each 5 °C increase (Farquhar et al., $ 1980 ; Sutton et al., 1993 c ; Schjørring et al., 1998). This provides a major control on the direction of NH fluxes, as demonstrated in Figure 6, which $ shows how an increase in temperature can lead to a switch from deposition to emission in laboratory measurements of NH exchange (Schjørring et al., $ 1998). In the field the effect of temperature on net fluxes becomes more complex, as increased emissions during warm conditions from all sources lead to increased NH concentrations in air. Thus tem$ perature, air concentrations and plant exchange fluxes are moderately coupled. Diurnal variations in apoplastic NH + and H+ % concentrations were fairly small in field-grown B. napus (Husted et al., 1997). However, when stomatal NH compensation points were derived from the $ apoplastic measurements and adjusted for diurnal variations in air temperature, the compensation point

36


peaked around noon and attained a minimum at night. A similar time-course was observed for the NH exchange flux. Thus, diurnal variations in NH $ $ exchange seem mainly to relate to changes in leaf temperature and stomatal conductance, rather than the concentrations of NH + and H+ in the leaf % apoplast. At present it can be concluded that larger values of [NH +] : [H+] occur in agricultural plants compared % with semi-natural plants, and that this is closely related to the available N supply from fertilizers. The result is a tendency for NH emissions to be $ greater from agricultural crops than from seminatural systems where deposition generally dominates. Little information is available on compensation points and apoplastic concentrations for semi-natural species and this area merits much greater attention. For example, feedbacks between long term atmospheric N deposition and NH $ compensation points for semi-natural plants have been suggested (Sutton et al., 1995 b). The consequence of such feedbacks is the potential for a longterm limitation of NH deposition in areas which $ have received a high N load. Controlled experiments are required to quantify the extent of these links, which might, for example, explain the large values of χs estimated for a polluted forest in the Netherlands (Table 2 ; Sutton et al., 1995 d ; Wyers & Erisman, 1998). Exchange of ammonia with leaf surfaces Dry deposition rates of NH measured in the field $ indicate that NH has a high affinity for leaf cuticles $ and other plant surfaces (Erisman & Wyers, 1993 ; Sutton, Fowler & Moncrieff, 1993 b ; Duyzer et al., 1994). Even in the absence of free water on leaves, a significant amount of wetness appears to be associated with leaf cuticles, increasing at high humidities, which provides a sink for NH uptake. $ This has been illustrated by the hydration of dead leaves at different humidities (van Hove & Adema, 1996), and is also supported by observations of water adsorption onto glass in relation to r.h. (Benner, Ogorevc & Novakov, 1992). The amounts of NH adsorbed onto leaf cuticles $ increases both in relation to increasing r.h. (van Hove, Adema & Vredenberg, 1988 ; Sutton et al., 1995 d) and in relation to decreasing vapour pressure deficit (van Hove et al., 1989, Nemitz et al., 1997, and unpublished). Clear relationships have been demonstrated between estimates of the leaf cuticle resistance for NH deposition (Rw) and r.h. over a $ Douglas-fir forest (Sutton et al., 1995 d ; Wyers & Erisman, 1998) as well as upland moorland (Fowler et al., 1998). Typical estimates of Rw range from 0 to 30 s m−" for r.h. above 85 %, increasing to typically 100–500 s m−" at ! 50 % r.h. (Sutton et al., 1995 d). In laboratory cuvette studies of NH gas exchange, $

generally performed at r.h. ! 60–70 %, it is generally assumed that stomatal exchange dominates, with cuticular uptake becoming more important in cool conditions (van Hove et al., 1989). The discrepancy between the measured and modelled fluxes in Figure 6 at ! 15 °C might be a consequence of this effect. There is ample evidence that the chemistry on leaf surfaces influences the rate of NH dry deposition, $ especially at high humidities or where leaves are wet. The presence of alkaline soil particles on leaves might reduce the rates of NH dry deposition (Sutton $ et al., 1993 b), whereas, conversely, the deposition of gaseous SO to leaf surfaces could enhance the rate of # NH deposition, sometimes referred to as ‘ co$ deposition ’ (van Breemen et al., 1982 ; van Hove et al., 1989 ; Erisman & Wyers, 1993 ; Sutton et al., 1995 d ; Cape et al., 1998). Such effects are closely linked to regulation of leaf surface acidity and the capacity for absorbing NH , which is greatly $ increased in acidic conditions. This effect provides an important link between the atmospheric S and NH cycles, since reducing SO emissions and $ # concentrations might be expected to reduce the rates of NH dry deposition (Sutton, Asman & Schjørring, $ 1994). Coupled with reduced formation of (NH ) SO aerosols and consequent wet deposition %# % (see sections on reaction and on wet deposition), this might increase the atmospheric lifetime of NHx leading to increased export and long range transport distances (Metcalfe, Whyatt & Derwent, 1998). Such effects have been indicated from long-term measurements of throughfall chemistry (van Breemen et al., 1982 ; Draaijers, Ivens & Bleuten, 1987 ; Cape et al., 1998). Despite this, the interpretation of real time deposition estimates from micrometeorological flux measurements has often had difficulty demonstrating these effects, and this might be linked to dynamic processes controlling these effects. The dynamic chemical behaviour of NH ad$ sorption with leaf surfaces poses a particular challenge for parameterizing NH exchange. Although $ most work has treated cuticular uptake as a steadystate absorption, governed by a resistance, the evidence is that in many situations it is more realistically described by a capacitance for NH $ adsorption. Based on modelled thickness of water films in relation to humidity, Sutton et al. (1995 a) provided an initial cuticular capacitance model, which was able to reproduce the occurrence of a morning pulse of NH emission following leaf $ drying. Nonetheless, major simplifications were made regarding leaf surface chemistry and there is a need to develop multi-species chemical models to consider these interactions. Although the instantaneous fluxes of NH and SO (e.g. ! 30 min time $ # resolution) must contribute to the co-deposition effect, it is likely that the interaction is mediated over longer timescales (e.g. days to weeks) as leaf chemistry effects accumulate. This might explain the

Emission, transport and deposition of ammonia difficulty in observing short-term interactions in micrometeorological measurements (Erisman & Wyers, 1993 ; Sutton et al., 1993 a). Higher average rates of SO deposition have been measured in NH # $ polluted regions (Erisman et al., 1993) compared with cleaner SO areas (Fowler et al., 1992) pro# viding evidence of these co-deposition interactions.

Exchange of ammonia with plant litter and soil surfaces The upper part of a plant canopy is that most aerodynamically connected to the atmosphere, with much larger resistances operating through the canopy to the soil surface. As a result, upper canopy cuticles and stomata play a key role in the exchange of NH with the atmosphere. In situations where $ NH dry deposition occurs to the canopy as a whole, $ the ground surface is only a minor sink. The situation is different, however, when the soil surface, or a covering layer of leaf litter, is a large source of NH $ emission. Soil-surface NH emissions are most $ closely associated with the application of mineral fertilizers (Whitehead & Raistrick, 1990 ; Sutton et al., 1995 a ; van der Weerden & Jarvis, 1997), especially following urea application. Decomposition of leaf litter is also a ground source of NH emission $ and its importance needs further investigation. Emissions from B. napus canopies following leaf fall appear to be significant (Sutton et al., 1997 b ; Nemitz et al., 1997). However, the potential for NH $ emissions from litter in semi-natural and forest ecosystems is unquantified, although this is likely to depend on the N content of the litter.

Integrating canopy exchange of ammonia As the understanding of the different canopy components contributing to NH exchange has improved $ it has become necessary to build quantitative models to evaluate the integration of these effects in generating net fluxes. Noting the components of Figure 5, a significant amount of internal cycling of gaseous NH might be expected, with emissions $ occurring that do not escape to the atmosphere. This was first indicated over 20 yr ago by Denmead, Freney & Simpson (1976) who used within-canopy concentration profiles and gradient diffusion theory in a grass-clover canopy to estimate fluxes at different levels within the canopy. Their conclusion was that most of the NH emitted by the clover understorey $ was recaptured by the overlaying grass canopy. A concern of this early approach, however, is the inadequacy of gradient diffusion theory within canopies, owing to the contribution of near-field or foliage-wake effects (Raupach, 1989 ; Denmead, 1995). A relatively new approach to address these questions has been developed by Raupach (1989),

37 which considers transfer using a matrix of different source}sink and concentration layers together with a Lagrangian model operated in reverse, the so called ‘ inverse Lagrangian approach ’ (ILA). This approach has been recently applied for NH (Nemitz $ et al., 1997) ; the analysis, for B. napus, indicated that litter fluxes were generally recaptured by the overlaying canopy. However, under windy conditions at night a significant fraction would be expected to escape from the canopy. Since the measurement equipment and mathematical resources required to implement the ILA are much larger than for the classical gradient approach, a key aspect of this work must be a quantitative assessment of the uncertainties involved when applying the simple gradient theory within different canopy types. Although the ILA provides measured estimates of source}sink strengths at different layers within the canopy, there is a need to develop predictive models that can simulate the net exchange fluxes. Two other more sophisticated models of the NH exchange $ process to that outlined in eqns (1) and (2) (Figures 4 b, a) are shown in Figure 4 c, d. Figure 4 c is essentially a single-layer model of the canopy, which integrates between cuticular deposition and bidirectional stomatal exchange in relation to a compensation point (Sutton et al., 1995 d). The key to this model is the prediction of surface concentration or compensation point for the canopy as a whole (χc). This model has been implemented in Figure 7 for an intensive agricultural grassland, where fluxes were measured by the micrometeorological aerodynamic gradient method (Sutton et al., 1997 a). The model is applied with a simple r.h. response to Rw, whereas the value of χs is fitted using two discrete values of the ratio [NH +] : [H+], of 12 600 and 3150 for short % and long grass, respectively. The encouraging fit of the model to the measurements using only approximate values in the parameterization provides an indication of the robustness of the model simplification. The measurements imply that the compensation point of the short grass was four times larger than that of the long grass, which appeared to be a consequence of cutting about a week previously. At 20 °C the values of χs were 49±9 and 12±5 µg m−$, for short and long grass, respectively. In many circumstances, the model used in Figure 7 (Fig. 4 c), would be sufficient to provide a broad description of the exchange process. Nevertheless, it is clear that for complex canopies such as those of B. napus, there is a need for the further development of multi-layer models to integrate ground surface fluxes with those of the upper canopy. Dynamic models are also needed, for the consideration of temporal chemical adsorption}desorption processes on leaf cuticles, as well as models addressing the potential for gas-particle reaction within and above canopies. Such vegetation exchange models might also begin to be built into atmospheric transport models, and a

38

W. A. H. Asman and others 300 Measured F t

250

Modelled F t (long grass) Ammonia flux (ng m–2 s–1)

200

Modelled F t (short grass)

150 100 Emission

50 0

0

0200

0400

0600 0800

1000

– 50

1200

1400 1600 Time (GMT)

1800

2000

2200

0

– 100 Deposition – 150 Measurements over long grass

Measurements over short grass

– 200

Figure 7. Ammonia fluxes measured over grassland compared with an implementation of the canopy compensation point bi-directional flux model (Fig. 4 c). The model was fitted with a r.h. response in Rw and apoplastic [NH +] : [H+] of 3200 and 13 200, respectively. The two grass canopies were different parts of the % same field, with the short grass cut approx. 1 wk before the measurements were made.

preliminary implementation of the model in Figure 4 c has already been applied by Sorteborg & Hov (1996) on a European scale. Dry deposition of particulate ammonium The dry deposition process for aerosol NH + is % somewhat different from that for gaseous NH $ (Davidson & Wu, 1990). They share the fact that atmospheric turbulence dominates the transport from the atmosphere to the quasi-laminar boundary layer over the surface to which deposition occurs. Ammonium-containing particles are so small that gravitation is not an important transport process. Transport through the quasi-laminar boundary layer occurs by molecular diffusion (gases) or Brownian diffusion (particles). Particles containing NH + (mass % median diameter c. 0±5 µm, Davidson & Wu (1990)) are much larger than NH molecules and conse$ quently their diffusivity is much smaller than that of NH . Once transported through the laminar bound$ ary layer there seems to be no important resistance to surface uptake (Rc). Particles of this size will not reemit easily, which means that the NH + flux at the % surface itself is, contrary to the NH flux, uni$ directional. Observations of apparent emission of NH + aerosol are therefore most likely to be a % consequence of the near surface gas-particle conversion effects noted above (Nemitz et al., 1996). The dry deposition velocity of particles is determined by the physical properties (which are a function of their size), their chemical properties, the properties of the surface (roughness, presence of obstacles like hairs) and the meteorological conditions. The dry deposition velocity to forests will be

higher than to grass due to the larger canopy roughness which generates more turbulence. It is very difficult to measure reliably the dry deposition velocity of particles, owing to the small vertical concentration gradients encountered and the changes in the size of the particles, which is a function of the height-dependent r.h., and few measurements exist (Ruijgrok et al., 1995). Duyzer (1994) measured the dry-deposition velocity of particulate NH + to heathland in the % Netherlands and found an average value of 1±7 mm s−". During the same period the drydeposition velocity of NH was estimated to be $ 14 mm s−", i.e. about a factor of 10 higher. This illustrates that the dry-deposition velocity of NH + % is about an order of magnitude lower than that of NH over land, if the compensation point of NH is $ $ very low. More measurements of the aerosol NH + % dry-deposition velocity to different types of surfaces are needed, in particular clarifying the interactions with near-surface formation}destruction processes.   Wet removal of NHx from the atmosphere occurs by two processes with different efficiencies. Close to sources below cloud, NHx is scavenged out of its plume rather inefficiently by raindrops and snow. (Asman, 1994 ; Asman, 1995). This means that the wet deposition close to sources is not so important compared with wet deposition further afield and that elevated concentrations in precipitation reported close to sources are presumably caused by dry deposition to the surface of the precipitation samplers (Jensen & Asman, 1995). When the plume is

Emission, transport and deposition of ammonia mixed into the clouds NHx is transferred into cloud droplets. These aggregate by various microphysical processes to snowflakes and raindrops, which are then deposited from the atmosphere. This more efficient in-cloud scavenging process for NH differs $ from that for NH + aerosols. Aerosols act as % condensation nuclei, such that condensation of water vapour on the aerosol surface converts them into cloud droplets. Once formed, they might then readily absorb gaseous NH . $ Cloud-water and precipitation are usually acidic, so that most of the NH taken up by drops reacts $ with H+ to form NH +. It is therefore only possible % to distinguish between the contribution from NH $ and NH + to precipitation if models are used which % describe the uptake processes of these components separately. Calculations with a wet-deposition model applied to Denmark showed that the contribution of different processes to the wet deposition of NHx was 15 % from in-cloud scavenging of NH , 77 % from $ in-cloud scavenging of NH +, 6 % from below-cloud % scavenging of NH and 2 % from below-cloud $ scavenging of NH + (Asman & Jensen, 1993). % Maps of European wet deposition of NHx have been provided by Buijsman & Erisman (1988) and Hjellbrekke, Schaug & Skjelmoen (1996). Measurements of the wet deposition close to sources are needed to validate the conclusion that nearby sources do not significantly contribute to wet deposition.        x It is important to develop numerical models of atmospheric transport and deposition for several reasons : (1) Models allow the interpolation in space and time between measurements of air concentrations and deposition. This is especially useful in the case of NH , where the high spatial and temporal $ variability, makes it very difficult to estimate national distributions of NH concentrations and deposition $ from measurements alone. (2) Models can be used to obtain a better strategy of where to look and which processes to look for during field measurements. For example, they can be used to indicate where the largest concentrations gradients would be expected. With this knowledge, measuring sites can be chosen in such a way that the maximum amount of information can be gathered with the available resources. (3) Models provide a framework for integrating present knowledge, and they allow the calculation of the contribution of different processes (e.g. in-cloud, below-cloud scavenging) to deposition, which cannot otherwise be separated (Asman & Jensen, 1993). This makes it possible to identify the most important processes for deposition. (4) Continental-scale models allow the calculation

39 of the contribution of different countries to the deposition in a country (Asman & Janssen, 1987 ; Barrett & Berge, 1996 ; Hjellbrekke et al., 1996). Such information is normally very difficult to obtain from measurements. (5) Models can be used to describe situations in the past. For example, an important question is how high the total NHx deposition was before ecological impacts of N deposition were observed. As there were no measurements of the total deposition at the beginning of this century, and only a few of wet deposition, this is difficult to estimate. It is, however, possible to make model calculations based on estimates of historical NHx emissions (Asman et al., 1988). (6) Models can also be used to predict future concentrations and deposition for different scenarios. This can, for example, be important if it is aimed at reducing the deposition in particular areas, investigating the effectiveness of different emission abatement options. Such calculations are made in connection with the international negotiations on the NOx protocol of the UN-ECE (Alcamo, Shaw & Hordijk, 1990). Importance of scale Atmospheric transport models necessarily incorporate each of the key processes of the atmospheric NH budget : emission, transport and diffusion, dry $ and wet deposition. The information on processes is available from measurements, such as described in the previous sections. Moreover, model results are as far as possible verified with other measurements, such as monitored spatial and temporal distributions of concentrations and deposition. The scales of models range from local (up to a few km from a source) to continental or even global. In theory, irrespective of scale models should accommodate all processes of the atmospheric budget. In practice, however, limitations due to computational resources, poor understanding of the system and data availability mean it is generally impossible to take all processes into account or treat them all in detail. Also it must be recognized that processes that are of major importance on one scale might be less important on another scale. On a local scale it is essential to have a detailed description of the diffusion process, otherwise the concentration and deposition as a function of downwind distance from the source is not accurately known. For global-scale models it is still important to take the deposition close to sources into account, but it does not matter whether the deposition occurs at exactly at 100 m or 10 km from the source, because the spatial resolution of this type of models exceeds 200 km. Modelling the atmospheric transport of NHx is not easy. This is because the horizontal gradient in dry deposition of NH close to the source (! 1 km) $

40

W. A. H. Asman and others 1 dry NH3 wet NH3 +

dry NH4

+

Fraction deposited

wet NH4

total NHx

0·1

0·01

0·001 101

102

103 104 Distance downwind (m)

105

106

Figure 8. Fate of atmospheric NH emissions : cumulative deposition of different forms as a function of $ downwind distance from a 1-m-high point source. The deposition is integrated over all wind directions and is expressed as a fraction of the NH emission. The calculations are based on Dutch climatology and surrounding $ land being rough grassland with an estimated Rc of 30 s m−". (Reprinted from Asman & van Jaarsveld (1992). Copyright 1992, with kind permission from Elsevier Scientific, UK).

should be described in detail as well as the longrange transport and deposition of particulate NH +. % Most atmospheric transport models describe either processes close to a source in detail, or only longrange transport, but it is difficult to include processes on both scales in one model. Asman (1994) gives an overview of atmospheric transport models for NH . $ More information on these model types and their limitations can be found in Asman & Janssen (1987), Janssen & Asman, (1988), Asman & van Jaarsveld (1992), Dentener & Crutzen (1994), Barrett & Berge, (1996). As an example to illustrate the scale of dispersion and deposition of NH , the modelled fate of NH $ $ emitted to the atmosphere from a 1-m-high point source in a Dutch climate is shown in Figure 8 (Asman & van Jaarsveld, 1992). The accumulated deposition downwind is presented in this figure as fraction of the emission. At an infinite distance from the source all NHx should be deposited and consequently the fraction should equal one. These model results show that 44 % of the emitted NH is $ estimated to be dry-deposited as NH , 6 % is wet$ deposited as the contribution of NH to the wet $ deposition of NHx, 14 % is dry-deposited as particulate NH + and 36 % is wet-deposited as the % contribution of particulate NH + to the wet de% position of NHx. Figure 8 shows that 10 % of the emission is estimated to be dry-deposited as NH $ within 100 m of the source and that 20 % is deposited within 1000 m of the source. This is because the emissions occur close to the ground for most NH $ sources, which leads to large concentrations at ground level close to the source. Coupled with potentially high NH dry-deposition velocities (Rc $

and Rw are small), this leads to very large dry deposition close to the source (Fig. 8). For rough surfaces like forests the percentage of the emission that is dry deposited close to the source is even higher owing to the higher turbulence. Figure 8 also shows that the deposition close to the source occurs mainly in the form of dry deposition of NH and $ further away from the source mainly as wet deposition of NH +. This means that NH is mainly % $ exported to other countries in the form of NH +. % Figure 9 shows the sum of modelled total deposition of NHx in Europe (Asman & van Jaarsveld, 1992). It can be seen from this figure and Figure 2 that the deposition is highest in areas with a high NH emission density. In these areas the $ contribution of dry deposition NH originating from $ local sources is important, as well as wet deposition of NH + from more distant sources. In more remote % areas like northern Sweden, the wet deposition of NH + dominates. % Regional models can be used to calculate the contribution of domestic and foreign sources to the deposition in a country (Asman & Janssen, 1987 ; Barrett & Berge, 1996). However, it is difficult to generalize the fractions imported and exported, since this depends not only on the emission densities within the country and the adjacent areas, but also on the size of the country and the meteorological} climatological conditions. In a small country with a high emission density like the Netherlands c. 80 % of the NHx deposition is estimated to originate from domestic sources. However, in a country like Sweden, with a low emission density, only c. 35 % comes from domestic sources (Barrett & Berge, 1996). For European countries (except Russia,


< 100

< 200

< 400

< 800

< 1600

41

< 3200

> 3200

Figure 9. Total (dry plus wet) modelled NHx deposition in Europe (mol ha−" yr−" ; 1000 mol ha−" yr−" ¯ 14 kgN ha−" yr−"). (Reprinted from Asman & van Jaarsveld (1992). Copyright 1992, with kind permission from Elsevier Scientific, UK.)

because it is so large) typically 40–60 % of the NH $ emission emitted by a country is estimated to be exported to other countries, which illustrates that NHx is not only a local problem (Barrett & Berge, 1996). Comparison with NOx and reaction products It is helpful to contrast the atmospheric behaviour of NOx with that of NHx. Although both oxidized and reduced N contribute to eutrophication and acidification, they differ in several ways : (1) The dry deposition velocity of NO (and its rapidly formed reaction product NO ) is an order of #

magnitude (NO) or a factor of 3 lower (NO ) than # that of NH . Although an important part of the NOx $ is emitted from low-level sources (traffic, domestic heating) this does not lead to as much local dry deposition as for NH , because of the much lower $ dry-deposition velocities. (2) Nitric oxide and NO are not very soluble in # water. By contrast, their reaction product HNO is $ very soluble, and the resulting NO − particles act $ efficiently as condensation nuclei. As a consequence, NO and NO are inefficiently removed by pre# cipitation, whereas HNO and NO − are scavenged $ $ quickly. For this reason wet deposition of oxidized N near a source is small and only becomes important

42


after NOx has been converted to HNO and NO −. $ $ Ammonium is removed much more efficiently than Nox by precipitation, at least if it has reached cloud level. (3) The reaction rate of NO to HNO , NO − # $ $ aerosol and other products is on average 3 % h−", which is a factor of 10 slower than for NH . $ As a result, N deposition caused by NO and NO # is not so important close to sources and relatively more NOx and reaction products are exported than is the case for NHx. In European countries typically 80–95 % of the NOx and its reaction products are exported, whereas the corresponding number is typically only 40–60 % for NHx. New developments in atmospheric transport modelling of ammonia It was noted earlier that the emission of NH from $ manure and fertilizer is a function of the atmospheric turbulence. Atmospheric diffusion and surface exchange are also functions of atmospheric turbulence. In most atmospheric transport models the diffusion and often, also, the dry deposition are made functions of the turbulence, but emission is not, which is in fact incorrect. Hutchings, Asman & Sommer (1997) tried to integrate an emission model (Hutchings, Sommer & Jarvis, 1996) with an atmospheric transport and deposition model (Asman, 1998) at the farm level. They made the emissions from manure and fertilizer functions of the turbulence and temperature, and the emissions from housing and storage facilities functions only of temperature. The windspeed was taken to be the same for all model runs (3±5 m s−" at 10 m height), but different atmospheric stabilities were chosen. Their results show that not only its emission rate higher in the case of an unstable atmosphere, but also that the emitted NH is transported further $ afield. This means that meteorological conditions favouring emission are associated with transport over longer distances. This illustrates how important it is to integrate all related processes in one model, and further work is required in this area. Can NHx deposition to sensitive ecosystems be reduced ? Given the known impacts of high N deposition to semi-natural ecosystems, a sustainable environmental policy must include ways of reducing deposition in affected areas. Reductions in NHx deposition can be achieved by either direct abatement of emissions, primarily from livestock agriculture, or by indirect abatement by modifying local scale atmospheric transport. The latter might include modifying near source deposition or providing buffer zones where emissions are selectively removed close to seminatural areas of concern.

Direct abatement of agricultural NH emissions is $ possible by several means : (1) Reducing the N content of animal feeds to obtain optimal nutrition for each stage of development of the animal. In current practice the same food is frequently used for animals of almost all ages. (2) Adding some specific amino acids to feed for non-ruminants (pigs etc.), so that supplying surpluses of other amino acids can be avoided. (3) Employing housing and storage systems which minimize losses of manure NH . $ (4) Using biofilters which reduce the NH concen$ trations in the emissions from housing and storage facilities. (5) Switching to low-NH -emission mineral $ fertilizers (Asman 1992), and, on bare arable fields, incorporating fertilizer directly into the soil. (6) Using a low-NH -emission application tech$ nique : incorporating farmyard manure rapidly into the soil, injection of slurry, use of trail hoses in a growing crop. (7) Reducing livestock numbers. It should be noted that measures to reduce flows into the system are to be preferred (N in diets, livestock numbers), since savings of N by minimizing losses as NH only lead to increased N $ contents of the soil, which in turn might have other adverse effects including increased NO − leaching or $ N O emissions. However, efficient use of manure # does reduce the requirement for mineral fertilizers. Little work has been done in developing indirect abatement measures for NH deposition, which is $ important in areas with high NH emissions. The $ fact that a relatively large percentage of the emission is dry-deposited close to the source (Asman et al., 1989) suggests that this deserves further investigation. Benefits might be achieved, for example, by planting managed farm woodlands around known sources to increase local deposition and reduce deposition to more critical areas, or by selectively removing emission sources surrounding sensitive semi-natural systems. A simple analysis of the effects of such approaches is shown in Figure 10. Calculations were made for a theoretical nature reserve of an area of 1¬1 km, surrounded by rings of 0±5 km width with an average homogeneously distributed emission density of 40 kg N ha−" yr−", which is representative of Danish conditions. It was assumed that the whole area, including the nature reserve, had a roughness length of 0±3 m, that Rc was 30 s m−" for NH and 200 s m−" for NH +, the reaction rate was $ % 8¬10−& s−" and that the height of the sources was 4 m. The minimal effect (! 10 %) of precipitation was ignored. Figure 10 shows the contribution of the first 40 emission rings and the sum of the contributions of all other emission rings to deposition in the nature reserve. The total estimated dry deposition caused by all emission rings was 14±1 kg N ha−" yr−", of which 11±5 kg N ha−" yr−" was


all other rings

Contribution of emission ring (kg N ha–1 y –1)

3

2

1

0 0

5 10 15 20 Distance from nature reserve (km)

25

Figure 10. Contribution of different 500-m-wide emission rings around a model nature reserve of 1¬1 km# to the dry deposition in the nature reserve. The emission rings have a homogeneous emission density of 40 kg N ha−" yr−". The distance is the approximate distance of the centre of the emission ring to the border of the nature reserve.

caused by the first 40 emission rings (up to c. 20 km from the border of the nature reserve). The model estimates that the average deposition to the reserve will be reduced by c. 2±7 kg N ha−" yr−" when all emissions up to 0±5 km from the reserve are removed, and by 6±1 kg N ha−" yr−" when all emissions up to 2±5 km are removed. Almost all the N acquired from these nearby sources is dry deposition of NH . There are, however, other $ contributions to the N deposition. Under Danish conditions wet deposition of NHx caused by remote sources contributes c. 5 kg N ha−" yr−", dry deposition of NOx and reaction products contributes c. 4 kg N ha−" yr−" and wet deposition of NO − c. $ 5 kg N ha−" yr−". This means that the total N deposition would only decrease from 28 to 22 kg N ha−" yr−" if all NH emissions were removed $ up to 2±5 km from the reserve. If only the emissions up to 0±5 km from the reserve were removed the total N deposition would decrease from 28 to 25 kg N ha−" yr−". These reductions are not very

Vertical flux (lg NH3 m–2 s–1)

0·10

43 large because of the size of contributions from other N species relative to the dry deposition of NH in $ this example. Larger percentage reductions would be achieved if the nature reserve were a forest, as NH dry deposition would contribute a larger $ fraction of total N deposition. Similarly, if the emission density around the nature reserve were larger than the average emission density used here, the potential for reduction of the N deposition by removing emissions close to the reserve would also be larger. It should be noted that the deposition in the nature reserve is highest near the edges (Asman et al., 1989) and emission reductions close to the reserve will cause the largest percentage reduction in deposition near the edges. Influence of the compensation point on deposition close to sources The existence of a compensation point in plants has a major influence on NH fluxes and budgets on a $ landscape scale, particularly for fertilized agricultural land. Immediately downwind of a major source, concentrations of NH would be expected to be $ larger than χc allowing deposition to the canopy. As concentrations decrease away from the source, χc might be larger than χa, allowing net emission from the canopy. It is useful to establish the distance over which net deposition is possible in relation to the magnitude of χc for a given emission (Sutton et al., 1997 a). The local-scale transport-model of Asman (1998) was applied to investigate this by incorporating a range of compensation point values between 0 and 10 µg m−$ for a tall grass}crop canopy downwind of a 3-m-high source representing a housing}storage complex for 500 pigs. The model run incorporating an application of eqn (1) assumed a wind speed of 3±46 m s−" at 10-m height, neutral atmospheric conditions, a roughness length of 0±1 m, (RaRb) of 25 s m−" and Rs of 100 s m−". The vertical flux as a

Compensation point (lm NH3 m–3) 0 1 2 5 10

0·05

0·00

– 0·05

– 0·10

0

400

800 1200 Distance from source (m)

1600

2000

Figure 11. Modelled net vertical NH flux (emission is positive) as a function of distance from a 3-m-high point $ source and as a function of the compensation point of the surrounding area. (Reprinted from Asman (1998). Copyright 1998, with kind permission from Elsevier Scientific UK.)

44


function of distance in the plume centre is presented in Figure 11. It should be noted that the ordinate is chosen so that the large negative fluxes very close to the source are not shown. The distance at which the net flux is zero in this example varies from 450 m (χs ¯ 10 µg NH m−$) to 1450 m (χs ¯ 1 µg $ NH m−$). This can be compared with the field $ measurements of Sutton et al. (1997 a), who estimated a zero flux distance of around 50 m in an integrated case study investigating recapture of NH $ emissions from slurry spreading onto downwind intensive grassland, where χs was 50 µg m−$. Currently, most atmospheric transport and deposition models do not take into account the existence of a compensation point. In this case : (i) The total NH emission used in these models $ might be an underestimate because NH $ emissions from plants were not taken into account (see section on emissions). (ii) The modelled net deposition of NH to agri$ cultural crops might be somewhat overestimated.  Global NH emissions are estimated to be c. $ 54 Mt N, with c. 60 % coming from anthropogenic sources. These emissions are of the same order of magnitude as the NOx-N emissions on both global and European scales. NH emissions are estimated to $ have doubled since 1950. Emitted NH returns to the surface mainly in the $ form of dry deposition of NH and wet deposition $ caused by scavenging of particulate NH +. Dry % deposition dominates inputs near sources (within a few km), whereas wet deposition dominates inputs at remote locations several 100 km from sources. This difference is the result of the low source height of NH sources, the potentially high dry removal rates $ of NH , the small dry-deposition velocity of aerosol $ NH +, and the relatively fast conversion rate of NH % $ to particulate NH +. % In parts of Europe with high NH emissions, like $ the Netherlands, Belgium and Denmark, dry deposition of NH represents the largest contribution $ to total NHx deposition, but wet deposition of NH + % is also important. In these countries c. 75 % of the deposition is caused by domestic sources. In more remote areas, like Norway and parts of Sweden, wet deposition of NHx caused by scavenging of particulate NH + dominates total NHx deposition. % About 35 % of the total NHx deposition in these areas is estimated to come from domestic sources. Concentrations of NH decrease rapidly from $ sources (within the first 1–2 km), causing large spatial variation in dry deposition of NH in the rural $ landscape where there are many sources. For this reason it is important to use emission inventories and models with a high resolution in areas with a high

NH emission density, where dry deposition of NH $ $ dominates the total NHx deposition. For areas with a lower emissions like Norway and parts of Sweden, where wet deposition of NH + dominates the total % NHx deposition, spatial variations are much less because deposition is usually derived from remote sources, except where sources are nearby or when orographic enhancement of wet deposition plays a role in hill areas. Surface exchange of NH is essentially bi-di$ rectional, with both dry deposition and emission being possible , depending on the compensation point of vegetation canopy (χc) and air concentrations (χa). The same holds true in principle for NH $ concentrations over manure or fertilizer (χmf), although normally χmf ( χa. The value of χc depends on the stomatal compensation point (χs) of plant tissues and the extent of recapture onto leaf cuticles, which depends on air humidity and the presence of other acidic or basic species on leaf surfaces. In general χs is larger for agricultural plants than seminatural, and varies with plant growth stage, although detailed information is only presently available for a few species. It is also highly dependent on temperature, doubling for each increase of 5 °C. In complex canopies NH surface exchange is affected $ by the interaction of emission from ground surface litter and bi-directional exchange in various layers of the canopy. In most cases net dry deposition will occur to semi-natural areas. However, particular environmental conditions, such as high temperatures and response to long-term N deposition which raises the N status of the vegetation might also cause periods of NH emission from plant communities $ growing in semi-natural terrestrial ecosystems. In general, plant communities on arable land represent a net source of NH to the atmosphere, and NH $ $ emission might lead to a loss of up to 5 % of the shoot N content. Ammonia emissions from crops contribute significantly to atmospheric NH pollution $ (15–20 % of the total NH emission). This will affect $ atmospheric budgets and cause larger monitored air concentrations during summer. Atmospheric transport and deposition models are useful tools allowing interpolation in space and time between measurements of air concentrations and deposition, calculation of import}export balances for countries and estimation of past and future situations. Owing to the large spatial and temporal variability in the dry deposition of NH , many $ monitoring stations would be needed to infer the dry deposition of NH to a country from measurements, $ which would be very expensive. In this case models can be applied to calculate the deposition, and a limited number of monitoring stations can be used to collect measurements to validate the model results. High N deposition to sensitive ecosystems leads to adverse effects. Deposition to particular receptors can be controlled either by reducing NH emissions $

Emission, transport and deposition of ammonia or by modifying the spatial patterns of deposition. Emissions of NH can be reduced by controlling the $ N content in animal food, using housing and storage systems which minimize losses of NH from manure, $ switching to mineral fertilizers, which produce lower NH emissions, better application techniques, as $ well as by reducing the number of livestock. In areas with a high emission density it is possible to reduce the N deposition to specific ecosystems by management practices affecting local dispersion and deposition, such as selectively removing NH sources $ close to a nature reserve. This possibility for ‘ indirect abatement ’ would be most useful in those cases where there are rather strong NH sources close to a $ nature reserve. The approach would be most effective in that situation, because background N deposition caused by more remote NH sources and $ by NOx sources would contribute a smaller fraction of the total deposition. Research in the following fields is needed to improve our knowledge about atmospheric NH . $ (1) Natural emissions, including emissions from plants. (2) Emissions from animal excreta and mineral fertilizer outside Europe and North America. (3) Diurnal and seasonal variations in NH $ emission rates. (4) Emissions and deposition on a local scale. These should be measured and validated models should be developed that can describe deposition on local scales. Such models should ideally include meteorology dependent NH emissions, the exist$ ence of compensation points and different dry deposition velocities to different surfaces. (5) The magnitude of stomatal compensation points of plants for more species and as a function of time and N supply. This should enable modelling of the compensation point. (6) Future application of the stomatal NH com$ pensation point in models for NH exchange over $ various types of terrestrial ecosystems at different N input levels, and under different climatic conditions, requires quantification of the functional dependence of the NH compensation point on central driving $ processes in plant growth, N uptake (N availability) and N metabolism. (7) The chemical interactions affecting NH fluxes $ with vegetation, both on leaf surfaces, within the canopy space and above canopies. (8) Temporal variations of the reaction rate of NH at different sites, so that it is possible to $ quantify dependence on (pollution) climate using chemical models. (9) Dry deposition velocity of NH + to different % types of vegetation, and its interactions with near canopy gas-particle interconversion as affected by particle size. (10) Influence of nearby sources on the dry and wet deposition of NHx and identification of sec-

45 ondary abatement strategies to reduce NH de$ position to sensitive areas.                WAHA acknowledges partial financial support from the Danish Environmental Research Programme, the Danish Environmental Protection Agency (Project ‘ Spredning og effekter af ammoniak ’) and the Netherlands Foundation for Atmospheric Chemistry. MAS gratefully acknowledges support from the UK Department of Environment, Transport and the Regions (EPG 1}3}94) and the UK Ministry of Agriculture, Fisheries and Food (ADEPT, WAO613}CSA2644). Financial support is gratefully acknowledged from the Danish Agricultural and Veterinary Research Council, the Danish Research Academy (DANVIS programme) and the Hofmangave Foundation (JKS). The work of MAS and JKS was also supported by the European Commission DG XII}D under the EXAMINE and GRAMINAE projects (EV5VCT94-0426). Lex Bouwman, National Environmental Research Institute (RIVM), Bilthoven, the Netherlands is thanked for preparing Figure 1, and Ulli Dragosits of the University of Edinburgh, for preparing Figure 3. We are grateful to Lucy Sheppard, Institute of Terrestrial Ecology, for her comments on this paper.

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