What is it?

Ammonia (NH3) is a highly reactive and soluble alkaline gas. It originates from both natural and anthropogenic sources, with the main source being agriculture, e.g. manures, slurries and fertiliser application.

Excess nitrogen can cause eutrophication and acidification effects on semi-natural ecosystems, which in turn can lead to species composition changes and other deleterious effects (Bobbink et al., 2010; Krupa, 2003; Pitcairn et al., 1998; Sheppard et al., 2008; Van den Berg et al., 2008; Wiedermann et al., 2009)

Ammonia comes from the breakdown and volatilisation of urea. Emissions and deposition vary spatially, with "emission hot-spots" associated with high-density intensive farming practices. Other agriculture-related emissions of ammonia include biomass burning or fertiliser manufacture. Ammonia is also emitted from a range of non-agricultural sources, such as catalytic converters in petrol cars, landfill sites, sewage works, composting of organic materials, combustion, industry and wild mammals and birds (Sutton et al. 2000, Wilson et al. 2004).


At the turn of the 21st century, total ammonia emissions in the UK were estimated to be 283 kt N yr-1 (Sutton et al. 2000) with 228 kt coming from agricultural sources (Pain et al. 1998). In 2010 the agricultural sector was responsible for 89% of UK NH3 emissions (CEIP, 2010; Defra, 2011). National NH3 emissions in the UK are mapped at a 5 km grid resolution, using the AENEID model (Dragosits et al. 1998) for agricultural sources, and at a 1 km or 5 km grid resolution for non-agricultural sources and are freely available from the National Atmospheric Emission Inventory (

Emissions trends have mostly been downward since peak in late 1980s and early 1990s but have now flattened. As the climate warms, volatilisation of ammonia emissions will lead to a further rise in ammonia concentrations.

High emission areas with intensive dairy farming can be distinguished from low emission areas with extensive sheep and beef farming or “hot-spot” patterns associated with intensive pig and poultry farming. Emissions from agricultural sources vary temporally with agricultural practice. Seasonal variation is also associated with climate; volatilisation being highest when it is warmer. Some non-agricultural emission sources (e.g. seabird colonies) contribute only small amounts to the overall NH3 emissions in the UK but, due to their location, are often the dominant emission source in remote and otherwise “clean” areas. Larger seabird colonies have been shown to emit similar amounts of NH3 to large intensive poultry farms (Sutton et al. 2000, Wilson et al. 2004).

Atmospheric Interactions

Atmospheric ammonia has impacts on both local and international (transboundary) scales. In the atmosphere ammonia reacts with acid pollutants such as the products of SO2 and NOX emissions to produce fine ammonium (NH4+) containing aerosol. While the lifetime of NH3 is relatively short (<10-100 km), NH4+ may be transferred much longer distances (100->1000 km) (Asman et al. 1998, Fowler et al. 1998). Hence NH3 emissions contribute to international transboundary air pollutant issues addressed by the UNECE Convention on Long Range Transboundary Pollution.

In addition to the transboundary effects, NH3 has substantial impacts at a local level: emissions occur at ground level in the rural environment and NH3 is rapidly deposited (see Nitrogen deposition). As a result some of the most acute problems of NH3 deposition are for small relict nature reserves located in intensive agricultural landscapes (Sutton et al. 1998).

Ammonia can be volatilised, emitted into the atmosphere when the surface concentration exceeds that of the surrounding air. Losses of NH3 by volatilisation from the application of nitrogen (N) fertilisers range from negligible amounts to >50% of the applied fertiliser N, depending on fertiliser/manure type (e.g. urea higher volatilization rates than ammonium nitrate), application practice (e.g. injection, surface application) and environmental conditions (Peoples et al. 1995, Freney 2005). Solubility and dissolution processes primarily drive the magnitude of NH3 emissions, higher in warm drying conditions and smaller in cool wet conditions.

Concentrations and deposition

Ammonia concentrations are monitored across the UK (UK pollutant deposition),  and show large spatial variability, reflecting a combination of the large number of ground level sources, primarily related to livestock farming, and the very reactive nature of gaseous NH3. Concentrations of NH3 range from 10 μg m-3 in areas of intensive livestock production, especially dairy and beef production, to 0.1 μg m-3 in the Scottish Highlands, especially in the north-west of Scotland and in the Hebrides (RoTAP, 2012).

These concentrations can be used to estimate deposition although deposition varies with ecosystem type and meteorology. Due to the varying affinity and compensation points of ammonia for different habitats, expressed in differences in mean deposition velocities, the rates of ammonia deposition vary greatly between habitat types.

Maps of concentrations and depositions across the UK are mapped using the FRAME model and calibrated using the measured NH3 values at monitoring stations. This means that maps of NH3 dry deposition need to be interpreted with care, noting whether they refer to inputs to specific habitat types (e.g. woodland, shrublands and croplands) or net dry deposition averaged over entire grid squares. For the purpose of assessing critical loads exceedance, deposition values for the relevant habitats need to be used, rather than grid averages.

Areas at risk from ammonia/nitrogen impacts include those close to point sources and areas within intensive agricultural regions which see elevated ammonia concentrations.


Effects of ammonia have been established from transect studies downwind of significant NH3 sources (van Herk 1999; Pitcairn et al. 1995, 1998; Wolseley et al. 2006) and a field release (Sheppard et al 2011). Ammonia can be taken up through the leaves via stomata, increasing the potential for nutrient N uptake. The consequences of foliar uptake and processing of an alkaline gas for cellular functions, appear to drive the deleterious effects of NH3 on terrestrial plants. Alkalinity is also thought to be a key driver for NH3 effects on epiphytic lichens (van Herk 2001). Atmospheric NH3 also impacts as NH4+, when the NH3 deposits to plant surfaces, dissolves and is washed into the soil where it can increase soil acidity and interfere with base cation uptake (Pearson and Stewart 1993, Fangmeier et al. 1994, Krupa 2003). Effects represent the combined effects of uptake through shoots as NH3/NH4+ and roots as NH4+.

Negative effects on vegetation occur via direct toxicity, when uptake exceeds detoxification capacity and, via N accumulation, which increases the likelihood of detrimental interactions with other abiotic and biotic stressors. Ammonia can also enrich a system with nitrogen putting under-storey species at risk as they become shaded by the expansion of nitrophiles (N loving plants) that use the additional N to increase productivity and expand the over-storey. Nitrogen enrichment affects competition for resources, favouring fast growing, tall species with rapid N assimilation rates. Mosses and lichens are most at risk, they have limited detoxification capacity relative to their uptake potential and a large surface area relative to mass (Pearson and Stewart 1993).

Many lichen species are sensitive to even small increases in NH3 concentrations above c. 1µg m-3 (Wolseley et al. 2006). Current evidence suggests that the absence of acidophytic lichens (lichens loving acid conditions) from twigs and trunks of acid-barked trees, growing in NH3 rich environments, is due to NH3 neutralizing the bark pH (van Herk 2001). Sheppard et al. (2004) found that monthly NH3 concentrations > 20 µg m-3 decimated Cladonia portentosa populations in less than one year and that after three years the concentration had fallen to < 3 µg m-3. Wet deposited NH4+ caused only restricted damage.

In mosses, NH3 exposure can increase both the N and amino acid content of ectohydric pleurocarpous mosses. Elevations in N and amino acid content have been proposed as a well coupled indicator of NH3-N deposition (Pitcairn et al. 2006). Moss species differ with respect to their N uptake, and presumably their tolerance (Pitcairn et al. 2006). Some Sphagnum (bog mosses) appear to be very sensitive, especially those that lack the red-orange pigments, carotenoids, that protect against oxidative stress (Sheppard et al 2011). Overall dry deposited ammonia-N drives species composition change and reduces species cover and diversity, much faster than the same unit of N in wet deposition (Sheppard et al 2011).

Attributing both specific effects in the field and indicators can be challenging because ammonia is a form of nitrogen which is an essential plant growth nutrient. In addition, some of the effects are difficult to separate from those caused by management, or lack of shading of the under-storey.

A summary of effects on vegetation are:

  • Eutrophication leading to changes in species assemblages; increase in N loving species (e.g. grasses) and species that can up regulate their carbon assimilation at the expense of species that are conservative in their N use.
  • Shift in dominance from mosses, lichens and ericoids (heath species) towards grasses like Deschampsia flexuosa, Molinia caerulea and ruderal species, e.g. Chamerion angustifolium, Rumex acetosella, Rubus idaeus.
  • Increased risk of frost damage in spring (van der Eerden et al 1991)
  • Increased winter desiccation levels in Calluna and summer drought stress
  • Increase in N loving epiphytes, e.g. Xanthoria parietina, at the expense of epiphytes that prefer acid bark.
  • Increased incidence of pest and pathogen attack, e.g. heather beetle outbreaks.
  • Direct damage and death of sensitive species, e.g. lichens and mosses, Sphagnum, Pleurozium schreberi.
  • Reduced root growth and mycorrhizal infection leading to reduced nutrient uptake, sensitivity to drought and nutrient imbalance with respect to N that is taken up via the foliage (Perez Soba 1995 for Scots pine).
  • Increase in soil pH follows acidification
  • Ammonia excess will lead to increases in nitrification and denitrification, contributing to greenhouse gas emissions.
Asman, W.A.H.; Sutton, M.A.; Schjoerring, J.K. 1998 Ammonia: emission, atmospheric transport and deposition New Phytologist 139 27-48
Bobbink, R.; Hicks, K. ; Galloway, J. ; Spranger, T. ; Alkemade, R. ; Ashmore, M. ; Bustamante, M. ; Cinderby, S.; Davidson, E. ; Dentener, F. ; Emmett, B. ; Erisman, J. W.; Fenn, M. ; Gilliam, F. ; Nordin, A.; Pardo, L. ; Vries, W. 2010 Global assessment of nitrogen deposition effects on terrestrial plant diversity: a synthesis Ecological Applications 20 30-59
Dragosits, U.; Sutton, M.A.; Place, C.J.; Bayley, A. 1998 Modelling the spatial distribution of ammonia emissions in the UK Environmental Pollutution (Nitrogen Conference Special Issue) 102 195-203
Fangmeir, A.; Hadwiger-Fangmeir, A.; Van der Eerden, L.J.M.; Jager, H.J. 1994 Effects of atmospheric ammonia on vegetation - a review Environmental Pollution 86 43-82
Fowler, D.; Sutton, M.A.; Smith, R.I.; Pitcairn, C.E.R.; Coyle, M.; Campbel, G.; Stedman, J. 1998 Regional mass budgets of oxidized and reduced nitrogen and their relative contribution to the N inputs of sensitive ecosystems Environmental Pollutution (Nitrogen Conference Special Issue) 102 337-342
Freney, J. R. 2005 Options for reducing the negative effects of nitrogen in agriculture Science in China Series C-Life Sciences 48 861-870
Krupa, S.V. 2003 Effects of atmospheric ammonia (NH3) on terrestrial vegetation: a review Environmental Pollution 124 179-221
Pain, B.F.; Weerden, T.J.; Chambers, B.J.; Phillips, V.R.; Jarvis, S.C. 1998 A new inventory for ammonia emissions from U.K. agriculture (Ammonia Special Issue) Atmospheric Environment 32 309-313
Pearson, J.; Stewart, G.R. 1993 The Deposition of Atmospheric Ammonia and Its Effects on Plants New Phytologist 125 283-305
Peoples, M. B.; Herridge, D. F.; Ladha, J. K. 1995 Biological Nitrogen-Fixation - an Efficient Source of Nitrogen for Sustainable Agricultural Production Plant and Soil 174 3-28
Perez-Soba, M ; Dueck, T.A.; Puppi, G.; Kuiper, P.J.C. 1995 Interactions of elevated CO2, NH3 and O3 on mycorrhizal infection, gas exchange and N metabolism in saplings of Scots pine. Plant and Soil 176 107-116
Pitcairn, C.E.R.; Leith, I.D.; Sheppard, L.J.; Sutton, M.A.; Fowler, D.; Munro, R.C.; Tang, S.; Wilson, D. 1998 The relationship between nitrogen deposition, species composition and foliar nitrogen concentrations in woodland flora in the vicinity of livestock farms. Environmental Pollution 102 41-48
Pitcairn, C.E.R.; Fowler, D.; Grace, J. 1995 Deposition of fixed atmospheric nitrogen and foliar nitrogen content of bryophytes and Calluna vulgaris Environmental Pollution 88 193-205
Pitcairn, C.E.R.; D., Fowler ; Leith, I.D.; Sheppard, L.J.; Tang, Y.S.; Sutton, M.A.; D., Famulari 2006 Diagnostic indicators of elevated nitrogen deposition Environmental Pollution 144
Sheppard, L.J.; Leith, I.D.; Crossley, A.; Dijk, N.; Fowler, D.; Sutton, M.A.; Woods, C. 2008 Stress responses of Calluna vulgaris to reduced and oxidised N applied under 'real world conditions' Environmental Pollution 154 404-413
Sheppard, L.J.; Leith, I.D.; Mizunuma, T. ; Cape, J.N.; Crossley, A.; S., Leeson ; Sutton, M.A.; Fowler, D.; Dijk, N. 2011 Dry deposition of ammonia gas drives species change faster than wet deposition of ammonium ions: evidence from a long-term field manipulation Global Change Biology 17 (12) 3589-3607
Sheppard, L.I. ; Leith, I.D.; Crossley, A. 2004 Effects of enhanced N deposition on Cladonia portentosa; results from a field manipulation study. Lichens in a Changing Pollution Environment, English Nature Workshop 51-62
Sutton, M.A.; Milford, C.; Dragosits, U.; Place, C.J.; Singles, R.J.; Smith, R.I.; Pitcairn, C.E.R.; Fowler, D.; Hill, J.; ApSimon, H.M.; Ross, C.; Hill, R.; Jarvis, S.C.; Pain, B.F.; Phillips, V.C.; Harrison, R.; Moss, D.; Webb, J.; Espenhahn, S.E.; Lee, D.S.; Hornung, M.; Ullyett, J.; Bull, K.R.; Emmett, B.A.; Lowe, J.; Wyers, G.P. 1998 Dispersion, deposition and impacts of atmospheric ammonia: quantifying local budgets and spatial variability Environmental Pollutution (Nitrogen Conference Special Issue) 102 349-361
Sutton, M.A.; Dragosits, U.; Tang, Y.S.; Fowler, D. 2000 Ammonia emissions from non-agricultural sources in the UK. Atmospheric Environment 34 855 - 869
Van den Berg, L. J.L.; Peters, C. J.H.; Ashmore, M. R.; Roelofs, J. G.M. 2008 Reduced nitrogen has a greater effect than oxidised nitrogen on dry heathland vegetation Environmental Pollution 154 359-369
Van der Eerden, L.J.M.; Dueck, T.A.; Berdowski, J.J.M.; Greven, H.; Dobben, H.F. 1991 Influence of HN3 and (HN4)2 SO4 on heathland vegetation Atic Bot Neerl 40 281-296
Wiedermann, M. M.; Gunnarsson, U.; Nilsson, M. B.; Nordin, A.; Ericson, L. 2009 Can small-scale experiments predict ecosystem responses? An example from peatlands Oikos 118 449-456
Wilson, L. J.; Bacon, P. J.; Bull, J. ; Dragosits, U.; Blackall, T. D.; Dunn, T. E.; Hamer, K. C.; Sutton, M. A.; Wanless, S. 2004 Modelling the spatial distribution of ammonia emissions from seabirds in the UK Environmental Pollution 131 173-185
Wolseley, P. A.; James, P. W.; Theobald, M. R.; Sutton, M. A. 2006 Detecting changes in epiphytic lichen communities at sites affected by atmospheric ammonia from agricultural sources Lichenologist 38 161-176