A multi-layer model to describe the atmospheric transport and deposition of ammonia in Great Britain

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Atmospheric Environment Vol. 32, No. 3, pp. 393—399, 1998 ( 1998 Elsevier Science Ltd All rights reserved. Printed in Great Britain S1352–2310(97)00238–0 1352—2310/98 $19.00#0.00

A MULTI-LAYER MODEL TO DESCRIBE THE ATMOSPHERIC TRANSPORT AND DEPOSITION OF AMMONIA IN GREAT BRITAIN R. SINGLES,*,†,‡ M. A. SUTTON‡ and K. J. WESTON† † Department of Meteorology, Kings Buildings, University of Edinburgh, West Mains Road, Edinburgh, EH9 3JZ, Scotland; and ‡ Institute of Terrestrial Ecology, Edinburgh Research station, Bush Estate, Penicuik, Midlothian, EH26 0QB, Scotland (First received 1 November 1995 and in final form 1 October 1996. Published February 1998) Abstract—The large spatial variability of ammonia (NH ) emissions and deposition makes it difficult to 3 estimate the input of reduced nitrogen to different ecosystems from measurements, since to quantify accurately the spatial variation in NH concentration would require a very large number of monitoring 3 stations. Such quantification is an important requirement for assessing the impacts of acidifying and nitrogen entrophicating deposition. These problems have been addressed in the current study by applying an existing multi-layer model of ammonia transport (TERN) (ApSimon et al., Atmospheric Environment 1994, 28, 665—678), together with a detailed NH emission field, to develop a long-term statistical trajectory 3 model (FRAME) over Great Britain at a 5 km]5 km grid resolution. The model treats the vertical concentration gradient explicitly with 33 layers, as well as provides a description of the land-use dependence of dry deposition using a resistance formulation dependent on major surface landcover classes. By treating vertical diffusion explicitly, the model provides the capability for examining the spatial variability at a finer scale than models assuming instantaneous mixing of emissions, and also avoids the need for air concentration and deposition correction factors. Model results are presented in the form of maps of annual mean NH concentration and net dry deposition, and a comparison is made with NH concentrations measured by a3 national monitoring network. A total annual budget for reduced nitrogen3 is given which shows the directional dependence of both total deposition and export of reduced nitrogen. ( 1998 Elsevier Science Ltd. All rights reserved. Key word index: Ammonia, long-range transport, NH deposition, annual budgets. 3

1. INTRODUCTION

Atmospheric ammonia is the main alkaline gas present in the atmosphere. It is emitted in gaseous form as NH , but rapid reactions with other atmospheric 3 species (such as the oxidation products of SO and 2 NO ) lead to the formation of ammonium (NH` ) x 4 containing aerosols, e.g. ammonium sulphates ((NH ) SO ) and nitrates (NH NO ). Because of the 42 4 4 3 importance of nitrogen as a plant nutrient and of the soil nitrification of NH to NO~ , deposition of NH x 3 3 and NH` can lead to changes in plant health and 4 competition, and can cause acidification of the soil (e.g. Hornung et al., 1995). Emissions of NH occur mainly from low-level 3 sources (farm animals and agricultural land) and, depending on prevailing meteorological conditions, rapidly disperse in the atmosphere as plumes becomes

* Author to whom correspondence should be addressed.

better mixed with the surrounding air. Because large concentrations can occur close to sources, and because NH frequently has a large deposition velocity 3 (on semi-natural land and forests), this can lead to major gradients in local dry deposition, depending on the land surface (Sutton et al., 1993). In contrast, NH` 4 aerosol generally has a small deposition velocity and is transported over much larger distances. The large spatial variability in ground level concentrations and dry deposition of NH means that it is very difficult to 3 produce accurate country-wide maps of concentrations from measurements without setting up a very large number of stations. An alternative method is to use an atmospheric transport model to estimate air concentrations and deposition, and use measurements where available to test the accuracy of the model results. Several models have already been applied to Great Britain, using a variety of spatial resolutions ranging from 150]150 km grid squares for the EMEP model (e.g. Sandnes and Styves, 1992) to the 20]20 km grid squares for HARM (Metcalfe et al., 1995). All these models describe the process of vertical diffusion by

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assuming instantaneous mixing of emissions which results in a uniform vertical concentration profile. Used on its own, the assumption of instantaneous mixing leads to an underestimation of ground level concentrations, and without the use of appropriate correction factors, will also lead to the underestimation of dry deposition. These correction factors have been applied in the EMEP model (u factor) to account for the local dry deposition. Since the main purpose of the EMEP model is to describe long-range transport, these values represent a method of determining atmospheric removal by dry deposition over a large area and do not account for the very large spatial variation in dry deposition of NH that can 3 occur within the vicinity of the emissions source. Most of the NH emissioins are rural, and are often situated 3 close to areas which are susceptible to ecological effects. There exists the need for a model which is able to both describe this short-range deposition of NH 3 explicitly, and also be applicable on a national scale to describe the behaviour of ammonium aerosols. This implies that such a model requires a fine resolution of emissions, in order to distinguish clearly between source and sink areas, and be able to parameterise the exchange of NH over different land-types. It should 3 also have an acceptable description of vertical dispersion to account for the high concentrations and large dry-deposition fluxes which can occur in the close vicinity of high-emission areas. The application of the TREND model to the Netherlands using 5 km resolution emissions data (Asman and van Jaarsveld, 1992) clearly shows the spatial variability of NH dry de3 position that may be produced where fine-resolution emissions data are available. This study describes the mean annual spatial distribution of NH ground-level concentrations and dry 3 deposition in Great Britain using a new statistical atmospheric transport model (FRAME). This model uses a multi-layer scheme to describe vertical diffusion explicitly (curtailing the need for correction factors), and uses a recently developed emissions database of NH for Great Britain on a resolution of 5 km]5 km. 3 To account for ecosystem dependence of ammonia dry deposition, the model incorporates a land-use database applying different canopy resistances and deposition velocities dependent on a major ecosystem class.

FRAME has been created to be used as a statistical atmospheric transport model. The main purpose of it is to describe the average long-term behaviour of atmospheric pollutants, particularly NH , on a national scale, and to 3 describe the spatial distribution of annually averaged NH 3 surface concentrations and annual deposition fluxes of NH . x The model utilises straight-line trajectories, in relation to specified wind directions. Twenty four wind directions are considered, and the results are combined statistically, suitably weighted by the frequency of the winds from each direction. The wind data are taken from a 1981 study of wind trajectories (Jones, 1981). In the present study the model is run at a wind speed of 7.5 m s~1 to represent the mean value from the wind rose of Jones. Trajectories start at four different times of the day (00 h, 06 : 00 h, 12 : 00 h, 18 : 00 h) and the results are combined accordingly to accommodate the effect of the diurnal cycles of various modelled processes such as the vertical diffusivity and chemical reactions. Vertical mixing is described using K-theory eddy diffusivity, with the exchange of material between layers determined by the equation

A B

Ls L Ls " K z Lz Lt Lz

(1)

where s is the concentration of the species under consideration and t is time. The vertical diffusivity (K ) is defined as a function of z height (z), atmospheric stability and time of day. K is z modelled to have a linearly increasing value up a specified height and then remains up to the top of the mixing layer. This multi-layer treatment of the diffusion process means that local dry deposition of material emitted from ground level sources can be modelled explicitly and no correction factors need be applied (e.g. EMEP, u factor). A detailed description of the parameterisation of the vertical diffusion process used in FRAME and TERN can be found in ApSimon et al. (1994). The model has coupled chemistry of SO , 2 NO and NH , which is based on the same scheme as used in x 3 the EMEP model (Sandnes et al., 1992). A more detailed description of the chemical processes can be found in ApSimon et al. (1994). Emissions of NH are introduced into the lowest layer of 3 the model at a height of 1 m, whereas SO and NO emis2 x sions are evenly distributed throughout the lowest 300 m of the mixing layer to reflect the more variable emission heights of these species. In this study all emissions are assumed to be temporally constant and no account is made of diurnal and seasonal variations. Material is removed from the air column by way of dry deposition, wet deposition and chemical conversion. Dry deposition is treated using deposition velocities (» ) and $ these are listed in Table 1 together with the relevant chemical species. Although the original version of the TERN model used a single value of » for NH , this process has now been $ 3 expanded and is dealt with in Section 2.2.

Table 1. Chemical species used in the model showing the assigned values of the deposition velocity (» ) ! 2. MODEL DESCRIPTION

The model applied in this study, referred to as Fine Resolution AMmonia Exchange (FRAME), has been developed from the TERN Model, of which a more detailed description can be found in ApSimon et al. (1994). It is a Lagrangian model which describes the main atmospheric processes taking place in a column of air extending from ground to a maximum height of 2500 m. The model consists of 33 vertical layers, of variable depth, with the tops at 2, 4, 6, 10, 25, 50, 75, 100, 150, 200, and thereafter in 100 m steps up through the mixing layer.

Species SO 2 SO all forms 4 NH 3 NH` aerosol 4 PAN NO NO 2 NO~ aerosol 3 HNO 3

» (m s~1) $ 0.008 0.001 see Section 2.2 0.001 0.002 0.000 0.001 0.001 0.04

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Table 2. Scavenging coefficients (j) used by the model to calculate wet deposition, based on an annual rainfall of 1000 mm. After ApSimon et al. (1994) Species

j (s~1)

SO 2 SO2~ all forms 4 NH 3 NH` aerosol 4 PAN NO NO 2 NO~ aerosol 3 HNO 3

1.1]10~5 4]10~5 5.7]10~5 4]10~5 0.0 0.0 0.0 4]10~5 5.7]10~5

Wet deposition is performed by the use of washout coefficients and material is removed from all layers in the column at a rate dependent on concentrations and the specified rainfall rate. The coefficients are consistent with the washout ratios used in the EMEP model (Sandnes et al., 1992) when an air column of 1000 m depth with constant concentrations is assumed and are listed in Table 2. Annual rainfall data on a 5 km]5 km resolution are used to calculate wet deposition of ammonium and other components using a time constant average scavenging coefficient. As with the emission flux, rainfall is assumed to have a constant rate of precipitation. The daytime depth of the mixing layer is determined using the ‘‘slab’’ model of Carson (1973), which calculates the daytime growth of the layer due to heat flux from the surface and entrainment of more stable air from above. The nighttime depth of the mixing layer is determined in a more statistical manner using Pasquill stability classes. Values of this parameter can range from 100 m at night with very stable conditions, to a maximum daytime value of 1500 m.

2.1. Model input and parameters Spatially disaggregated ammonia emissions for the model have been taken from the results of Sutton et al. (1995b), and to allow comparison with the results of the model later in the paper, the emissions used are shown in Fig. 1. These data were also used in earlier development work of the model (Singles et al., 1995), but in that instance were aggregated up to 20 km]20 km grid resolution to allow computationally less intensive model testing model. Emissions data of sulphur dioxide and nitrogen oxides were taken from the National Atmospheric Emissions Inventory (Eggeleston, 1992a). The total emission flux of NH —N, calculated by Sutton et al. 3 (1995b) was 368 Gg year~1, and is similar to the estimate of Asman (1992) of 385 Gg year~1. However, it is significantly different from the values of 443 Gg yr~1, from Eggleston (1992b), and 210 Gg yr~1 calculated by Jarvis and Pain (1990), and there is significant uncertainty in this term. The transboundary import of foreign material has been modelled by using a set of concentration profiles for the edge of the Great Britain domain, which are used to initialise trajectories. These profiles have been created by running the original TERN model along a series of 96 h trajectories across Europe which terminate at the edge of the model domain. The much lower resolution (150 km grid squares) EMEP emissions of NH , SO and NO were used as input 3 2 x data to create these boundary profiles. Models such as HARM (Metcalfe et al., 1995) run trajectories across a large model domain, including the UK and much of Europe, but the intensive computational time in the FRAME model, caused by the multi-layer diffusion scheme, means that standard model runs are restricted to Great Britain.

Fig. 1. Total annual ammonia emissions for Great Britain from agricultural and other miscellaneous sources on a 5 km]5 km grid resolution as used in the model.

2.2. Dry deposition For most of the chemical species in the model, dry deposition is described by the use of an average » (Table 1). $ However, in the case of NH , a land-use database is used to 3 create a set of land-dependent values of » . The land-use $ database contains information on the classification of land into different categories for a specified grid square, and also the percentage land cover. There are five land categories, arable, forest, moorland, grassland and urban. Values of » are calculated for each land category using a resistance $ model. This resistance model assumes that the transport of material between a specified height in the atmosphere and the surface takes place between three resistances in series, the aerodynamic resistance (R ), the laminar boundary layer ! resistance (R ) and a surface resistance (R ). " # » "(R #R #R )~1. (2) $ ! " # To describe land-dependent dry deposition, a specific value of R is assigned to each land category, and these are com# bined with calculated values of R and R to create a set of ! " land-dependent values of » . The formulation of both $ R and R are taken from Garland (1977). The values of ! " R used are listed in Table 3. Due to the large surface # roughness of forests, R will dominate the calculation of # » during the day, but the increased stability at night will $ lessen this effect. All other surface classes, apart from moorland, are dominated by the very large values of R that are # used. These values are chosen to provide a simple description consistent with the literature on surface atmosphere exchange of NH with different land-types, e.g. Sutton et al. 3 (1993, 1994). Since emissions from fertilisers and crop decomposition are already treated from arable land in the emissions inventory, and the long-term exchange over this vegetation (not including the above mentioned emissions) suggests an extremely limited net deposition, a value of R is # set to 1000 s m~1. Rapid deposition is generally seen to forest/moorland vegetation, and for both these classes

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Table 3. Land-dependent values of the canopy resistance (R ) used in the resistance model to calculate # diurnal values of » for NH $ 3 Parameter

Arable

Forest

Grass

Moorland

Urban

Sea

R (sm~1) #

1000

20

600

20

240

240

a value of 20 s m~1 is applied. Grassland represents a more problematic category. In the NH emissions inventory used 3 for this study (Fig. 1), the data for grassland is defined as the net emission from this system. Sutton et al. (1993) recommended that surface atmosphere exchange should be set to zero over this surface. This has been implemented in the present version of the FRAME model by restricting the calculated » for grassland to very small values, by using a large value $ of R . The sensitivity to this parameter will be investigated in # future activities. Very little is known about the deposition of ammonia in urban areas, for example, to hard surfaces and buildings, and the large canopy resistance here reflects limited uptake by vegetation in urban areas. Sea may act both as a minor source as well as a sink, but that is not the focus of this study, and so an approximate value has been used for this surface.

3. RESULTS AND DISCUSSION

The field of NH3 ground-level concentrations (1—2 m layer), created by using the model conditions described here, is presented in Fig. 2. The modelled air concentration field shows the very large predicted spatial variability in NH3 ground level air concentrations. This results from more limited vertical diffusion than would be represented by instantaneous boundary layer mixing, as well as from rapid dry deposition. The distribution of air concentrations is consistent with the emissions database used in this study (Fig. 1) showing the highest air concentrations in a broad band in the borders of England and Wales. This corresponds with mainly livestock, cattle and sheep farming in these areas, as well as more local high emission areas in North West England (west Lancashire, North Cumbria). A further high emission area in eastern England (East Anglia) is associated with large poultry and pig farming. One of the most significant features of this exercise is the model estimation of extremely small air concentrations over the whole of the Scottish Highlands reflecting an extremely low emission density in this area. To test the results of the model, comparisons have been made with measured data. There are relatively few data on vertical concentration profiles, but data from the Cabauw mast in the Netherlands were compared to modelled concentrations from the TERN model (ApSimon et al., 1994). Measured data were recorded up to a height of 200 m, and the modelled data were in reasonable agreement with the observed concentrations. Surface concentration data from FRAME were also compared with measured NH3 concentrations from a network of passive diffusion tubes, operated over 1987—1990, which provide the most extensive set of measurements available in the United Kingdom at the present time (Atkins and Lee,

Fig. 2. Annual average NH surface (1—2 m) concentration 3 field produced by the model in the present study.

1992). The greatest number of monitoring sites were operational in the period of 1987—1988, and these are the data that were used to compare with the model results. One note of caution should be sounded when using results from diffusion tubes. Anderson (1991) reported that the samplers were thought to give a value which is greater than the actual air concentrations present, and a scaling factor of 0.45 has been applied to the diffusion tube results. This factor, though very uncertain, has been supported by further studies in the UK (Sutton et al., unpublished data), and was also used in a previous comparison of model results with the U.K. measurements by Lee and Johnson (1993). The results of the comparison with the present modelling study are shown in Fig. 3. Surface concentrations of NH3 can vary greatly over the short range of a few hundred meters, and thus over a range of 5 km, there may be a great deal of variations. Any comparison with measurement should be treated with some caution and be viewed as giving an indication of the overall effectiveness of the model performance, as opposed to accurately modelling concentrations at specific locations. Although there are significant divergences between the model predictions and the results of correcting the measurements, Fig. 3 shows that there is a significant

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Fig. 3. Correlation plot of modelled surface concentration values against measured values ]0.45. The central dotted line is a one to one agreement and the solid line is the best fit line produced by a regression analysis.

correlation between the measurements, with a value of R2 of 0.49 and P of 8.7]10~9. The gradient of the regression line is close to the 1 : 1 line. However, whilst other disagreements may be due to deficiencies in the formulation of the model, to some extent it may also be caused by limitations in the measurement data. For example, the correction factor for the diffusion tubes may include an intercept, as there is evidence that the correction factor becomes larger at smaller air concentrations. This point is borne out by the prediction of the model air concentrations in northwest Scottland in the region of 0.1 kg m~3, whereas the diffusion tube measurements suggest air concentrations in the range 0.7—2 kg m~3, for these sites. Other independent measurements of NH3 at sites in the Scottish Highlands have been made using continuous annular denuders (Sutton et al., 1995a) of the type described by Wyers et al. (1993), and these confirm the expected model results of air concentrations in the range 0.02—0.2 kg m~3. An important conclusion from this finding is that it is expected that dry-deposition maps, based on the diffusion-tube air concentration data, are likely to overestimate dry deposition significantly in low concentration areas (cf. INDITE, 1994). Since remote areas are extremely sensitive to nitrogen deposition, for example, moorland and bog areas of North west Scotland, this finding has very important consequences for estimation of critical load exceedances in these areas. Figure 4 shows a map of modelled annual dry deposition in Great Britain, and as with the plots of air concentrations, the spatial distribution of dry deposition is comparable to the emissions with large gradients occurring between areas of high and low emission values. Large deposition occurs in the vicinity of major emissions, such as East Anglia and the England/Wales border, and there is little dry deposition in remote areas such as northern Scotland. The largest

Fig. 4. Net annual modelled dry deposition of the N—NH 3 predicted by the FRAME model for different 5 km]5 km grid squares integrating the different rates of dry deposition into the ranges of ecosystem types.

deposition occurs in areas where there is both large emission and substantial areas of semi-natural ecosystems, such as forest and moorland, generally in the west. It should be noted that this map shows the integrated flux budget of component dry deposition to each different land-surface type, weighted according to fraction of surface area in each grid square. Actual inputs to specific land-use types will be different, with, for example, much larger inputs to forest and moorland areas. At present there is a need for further model evaluation regarding the magnitude of these fluxes, in particular, the description of the rate of dry deposition to grassland which is very uncertain. However, this map already shows the significant spatial variability of NH that may be expected where emissions are 3 applied with a fine resolution atmospheric transport transfer model able to describe sufficiently the vertical gradient and mixing of NH3 in the atmosphere. The increased spatial resolution allows the distribution of land types across the country to be more clearly defined. Dry deposition to sensitive ecosystems in the regions of sources may receive a much larger modelled dry-deposition flux than would occur with a low-resolution model or one based on interpretation of measured air concentrations. In contrast, remote areas or hill ranges of northern England receive less dry deposition than might have been imagined on the basis of an instantaneous vertical dispersion scheme. In the present case, it may be seen from Fig. 2 how the air concentrations over the Pennine Ridge of north

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Table 4. Estimated annual budget of reduced nitrogen in relation to trajectory origin and annual mean budget using the present formation of the model; values in brackets are frequency weighted totals for each wind sector; the units are Gg N yr~1 Wind sector origin of trajectories

Emissions of NH —N x

Import of NH —N x

Dry deposition of NH —N x

Wet deposition of NH —N x

Export of NH —N x

0—45

310.7

45— 90

310.7

90—135

310.7

135—180

310.7

180—225

310.7

225—270

310.7

270—315

310.7

315—360

310.7

Averaged annual modelled budget of NH —N x

310.7

1.3 (0.1) 29.0 (2.3) 94.4 (6.2) 76.2 (6.6) 49.1 (5.7) 33.6 (6.5) 28.5 (5.1) 11.3 (2.0) 35.6

105.0 (11.4) 107.0 (8.5) 105.8 (7.0) 106.3 (9.1) 108.2 (12.5) 108.7 (21.0) 107.9 (19.2) 105.9 (18.3) 107.0

109.1 (11.9) 126.5 (10.0) 166.4 (11.0) 170.8 (14.7) 135.3 (15.7) 109.7 (21.2) 105.8 (18.8) 103.0 (17.8) 122.1

97.9 (10.7) 106.3 (8.4) 132.9 (8.8) 109.8 (9.4) 116.4 (13.5) 126.0 (24.3) 125.4 (22.3) 113.1 (19.6) 117.2

central England are depleted in the sink region, thereby limiting the dry deposition of NH3 . The annual budget of reduced nitrogen for Great Britain predicted by the present implementation of the model is shown in Table 4. The greatest net export of reduced nitrogen occurs when the wind comes from the southwest. Ammonia-rich air, resulting from passage over source areas in the west and southwest of England, is subsequently transported over regions of large NOx and SO2 emissions. Thus, a sizeable fraction of the NH3 is converted to NH` 4 aerosol. Since these species have a longer atmospheric residence time than NH , they are the main form of reduced nitrogen 3 involved in long-range transport. By contrast, winds from the south to southeast show the least net export, but this is accompanied by the largest total of wet deposition, approximately 55% of the emissions. Trajectories from this wind sector pass through areas of northwest Scotland which experience the largest annual rainfall in the country, and thus much of the material in the air column is removed by rainfall scavenging and deposited. Dry deposition shows little dependence on wind direction.

4. CONCLUSIONS

To conclude, results are presented here of the development of a multi-layer atmospheric transport model of the emission, transport and deposition of NH3 over Great Britain. The availability of a detailed 5 km]5 km emission inventory, together with the explicit treatment of multi-layer diffusion in the model, has allowed the model to be run on this fine

spatial resolution providing a more realistic impression of the spatial variability of NH3 concentrations at ground level. In contrast, models assuming instantaneous mixing of NH3 emissions are likely to smooth out the air concentration field artificially, providing higher concentrations in remote areas, and underestimated concentrations in source areas if correction factors are not properly determined. Inclusion of a dry deposition dependent on land-use in the model has allowed the high removal at source to be directly incorporated into the transport scheme, without the use of correction factors. Compared to previous studies, the map shows larger rates of dry deposition in source areas and lower concentrations in sink areas, particularly northwest Scotland, though overall air concentrations predicted by the model show a good agreement with available measurement data. An NH3 budget for the country is presented showing that inputs by dry and wet deposition are of similar magnitude, 110 and 133 Gg of N, respectively. A comparable budget from INDITE (1994) gives the annual dry deposition flux of NH3—N to be 100 Gg and the wet deposition flux as 131 Gg. The modelled export—import (net export) of 81.6 Gg of N yr~1 is somewhat smaller than the estimate from INDITE of 119.0 Gg of N yr~1. This is largely as a result of the explicit description of diffusion in the present model which allows greater re-deposition in Great Britain. Acknowledgements—Financial support for this work and its presentation is gratefully acknowledged from the UK Natural Environmental Research Council (RJS) and the UK Department of the Environment (MAS). The authors wish to thank H. ApSimon and B. Barker for supplying the TERN

Atmospheric transport and deposition of ammonia code which formed the basis for the model, and also for their help and useful suggestions. Also, many thanks to R. Smith for his contributions to this work.

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Anderson H. (1991). Ammonia monitoring—passive diffusion tube sampling. In ‘Nitrogen and Phosphorus in Soil and Air’ Project Abstracts of the Danish NPo Research Programme. Ministry of the Environment Protection, Kobenhavn, Denmark. ApSimon H. M., Goddard A. J. H., Wrigley J. and Crompton S. (1984). Atmospheric transport of radioisotopes and the assessment of population doses on a European scale. CEC report EºR 9128, EN. ApSimon H. M., Barker B. M. and Kayin S. (1994). Modelling studies of the atmospheric release and transport of ammonia—applications of the TERN model to an EMEP site in eastern England in anticyclonic episodes. Atmospheric Environment 28, 665—678. Asman W. A. H. (1992). Ammonia Emission in Europe: ºpdated Emission and Emission »ariations. RIVM report. 228471008. RIVM, Bilthoven, The Netherlands. Asman W. A. H. and van Jaarsveld J. A. (1992). A variableresolution transport model applied for NH for Europe. x Atmospheric Environment 26A, 445—464. Atkins D. H. F. and Lee D. S. (1992). ¹he Distribution of Ammonia in the ºnited Kingdom. AEA Technology AEAEE-0469, Harwell Laboratory, Oxfordshire. Carson D. J. (1973). The development of a dry inversion-capped convectively unstable boundary layer. Quarterly Journal of the Royal Meteorological Society 99, 450. Eggleston H. S. (1992a). Pollution in the Atmosphere: future Emissions from the º.K. Report for the UK Department of Environment. Warren Spring Laboratory, Stevenage, Hertfordshire, SG1 2BX. Eggleston H. S. (1992b). An improved U.K. ammonia emission inventory. In Ammonia emissions in Europe: emission coefficients and abatement costs. Proceedings of a workshop 4—6 February 1991 (edited by Klassen G.), pp. 95—107. IIASA, Laxemburg, Austria. Garland J. A. (1977). The dry deposition of sulphur dioxide to land and water surfaces. Proceedings of the Royal Society of ¸ondon A 345, 245—268. Hornung M., Sutton M. A., and Wilson R. B. (eds.) (1995). Mapping and modelling of critical loads for nitrogen—a workshop report. (Report of an UN-ECE workshop, Grange over Sands, 23—26 October 1994.) Institute of Terrestrial Ecology, Edinburgh. INDITE (1994). Inputs of nitrogen deposition in terrestrial ecosystems. DoE, London.

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