A two-day Discussion Meeting of the Royal Society, ‘The Nitrogen Cycle’, held in London in June 1991 (Stewart & Rosswall, 1992) reviewed the considerable progress made in understanding the N cycle in agricultural, forest and aquatic systems. The meeting included some discussion of the concerns which were already being expressed at that time over nitrate in water supplies, and the impacts of nitrogenous gases on tropospheric chemistry, the greenhouse effect and the ozone layer. Since then, disquiet over the impacts of nitrogenous compounds on the environment has increased, and numerous papers have been published on many aspects of the problem. We now have much better understanding of the size and scale of the perturbation of the N cycle, and several review papers have highlighted the complexity of the formidable issues that are challenging environmental scientists (Vitousek, 1994; Galloway et al., 1995; Vitousek et al., 1997).

Presentations and discussions about the impacts of nitrogenous compounds dispersed via the atmosphere, and of ozone as a secondary pollutant

Air quality thresholds for O3 for the protection of human health and vegetation set by the European Union (EU) have been exceeded in Europe regularly in the 1990s. Because target reductions for oxides of nitrogen (NOx) set for the year 2000 are unlikely to be achieved, these O3 exceedances are likely to continue into the next millenium. Improvements of plant tolerance towards O3 are being investigated but very little work has been done to explore NOx tolerance and plant acclimation to NO2 and NO. However, it is clear that within the populations of some plant species there is wide variation, and some individuals can fix NOx and use the nitrogen directly from the atmosphere, rather than rely upon, for example, root uptake of nitrate. It is possible that individuals capable of fixing NOx could be selected for a range of species, and genotypes with high rates of uptake could be of value as crops or for forestation in polluted areas (e.g. landscaping in the vicinity of motorways) to reduce tropospheric concentrations of NOx significantly and also to decrease the potential for O3 production.

Emissions of reactive oxidized nitrogen (NO and NO2), collectively known as NOx, from human activities are c. 21 Tg N annually, or 70% of global total emissions. They occur predominantly in industrialized regions, largely from fossil fuel combustion, but also from increased use of N fertilizers. Soil emissions of NO not only make an important contribution to global totals, but also play a part in regulating the dry deposition of NO and NO2 (NOx) to plant canopies. Soil microbial production of NO leads to a soil ‘compensation point’ for NO deposition or emission, which depends on soil temperature, N and water status. In warm conditions, the net emission of NOx from plant canopies contributes to the photochemical formation of ozone. Moreover, the effect of NOx emissions from soil is to reduce net rates of NO2 deposition to terrestrial surfaces over large areas.

Increasing anthropogenic emissions of NOx have led to an approximate doubling in surface O3 concentrations since the last century. NOx acts as a catalyst for the production of O3 from volatile organic compounds (VOCs). Paradoxically, emission controls on motor vehicles might lead to increases in O3 concentrations in urban areas.

Removal of NO and NO2 by dry deposition is regulated to some extent by soil production of NO; the major sink for NO2 is stomatal uptake. Long-term flux measurements over moorland in Scotland show very small deposition rates for NO2 at night and before mid-day of 1–4 ng NO2-N m−2 s−1, and similar emission rates during afternoon. The bi-directional flux gives 24-h average deposition velocities of only 1–2 mm s−1, and implies a long life-time for NOx due to removal by dry deposition.

Rates of removal of O3 at the ground are also influenced by stomatal uptake, but significant non-stomatal uptake occurs at night and in winter. Measurements above moorland showed 40% of total annual flux was stomatal, with 60% non-stomatal, giving nocturnal and winter deposition velocities of 2–3 mm s−1 and daytime summer values of 10 mm s−1. The stomatal uptake is responsible for adverse effects on vegetation. The critical level for O3 exposure (AOT40) is used to derive a threshold O3 stomatal flux for wheat of 0·5 μg m−2 s−1. Use of modelled stomatal fluxes rather than exposure might give more reliable estimates of yield loss; preliminary calculations suggest that the relative grain yield reduction (%) can be estimated as 38 times the stomatal ozone flux (g m−2) above the threshold, summed over the growing season.

The authoritative talk by Professor Fowler (Fowler et al., 1998), emphasized the huge increase in the rate of NOx (NO and NO2) emissions into the atmosphere due to fossil fuel combustion, from 1 Tg N y−1 to over 20 Tg N y−1 during the 100 yr between 1880 and 1980. He went on to predict that this rate of emission from anthropogenic sources would increase to 46 Tg N y−1 by the year 2025. In addition, NO can also be released from the soil following microbial action, a process that is very dependent upon soil temperature, nitrogen availability and water content. Later in the meeting, Professor Raven (Raven & Yin, 1998) pointed out that terrestrial plants, though not necessarily each individual species, have over the past 450 million yr coped with large changes in nitrogenous compounds in the environment. Nevertheless, this is no basis for complacency about the current situation because the rates of change caused by man's activities are probably unprecedented. Furthermore, the fact that terrestrial plant life in some form can continue, despite massive changes in environmental chemistry, does not necessarily indicate that the systems on which we ourselves are dependent will be conserved.

The global emission of ammonia (NH3) 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 NH3 emission is estimated to come from anthropogenic sources. NH3-N emissions are of the same order as the NOx-N emissions on both global and European scales. Emitted NH3 returns to the surface mainly in the form of dry deposition of NH3 and wet deposition of ammonium (NH4+). In countries with high NH3 emission densities, dry deposition of NH3 from local sources and wet deposition of NH4+ from remote sources dominate the deposition. In countries with low NH3 emission densities only wet deposition of NH4+ from remote sources dominates the deposition. Surface exchange of NH3 is essentially bi-directional, depending on the NH3 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 NH3 doubles for each increase of 5°C. However, exchange of NH3 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 NH3 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.

Human activity has greatly perturbed the nitrogen cycle through increased fixation by legumes, by energy and fertilizer production, and by the mobilization of N from long-term storage pools. This extra reactive N is readily transported through the environment, and there is increasing evidence that it is changing ecosystems through eutrophication and acidification. Rothamsted Experimental Station, UK has been involved in research on N cycling in ecosystems since its inception in 1843. Measurements of precipitation composition at Rothamsted, made since 1853, show an increase of nitrate and ammonium N in precipitation from 1 and 3 kg N ha−1 yr−1, respectively, in 1855 to a maximum of 8 and 10 kg N ha−1 yr−1 in 1980, decreasing to 4 and 5 kg N ha−1 yr−1 today. Nitrogen inputs via dry deposition do, however, remain high. Recent measurements with diffusion tubes and filter packs show large concentrations of nitrogen dioxide of c. 20 μg m−3 in winter and c. 10 μg m−3 in summer; the difference is linked to the use of central heating, and with variations in wind direction and pollutant source. Concentrations of nitric acid and particulate N exhibit maxima of 1·5 and 2 μg m−3 in summer and winter, respectively. Concentrations of ammonia are small, barely rising above 1 μg m−3.

Taking deposition velocities from the literature gives a total deposition of all measured N species to winter cereals of 43·3 kg N ha−1 yr−1, 84% as oxidized species, 79% dry deposited. The fate of this N deposited to the very long-term Broadbalk Continuous Wheat Experiment at Rothamsted has been simulated using the SUNDIAL N-cycling model: at equilibrium, after 154 yr of the experiment and with N deposition increasing from c. 10 kg ha−1 yr−1 in 1843 to 45 kg ha−1 yr−1 today, c. 5% is leached, 12% is denitrified, 30% immobilized in the soil organic matter and 53% taken off in the crop. The ‘efficiency of use’ of the deposited N decreases, and losses and immobilization increase as the amount of fertilizer N increases. The deposited N itself, and the acidification that is associated with it (from the nitric acid, ammonia and ammonium), has reduced the number of plant species on the 140-yr-old Park Grass hay meadow. It has also reduced methane oxidation rates in soil by c. 15% under arable land and 30% under woodland, and has caused N saturation of local woodland ecosystems: nitrous oxide emission rates of up to 1·4 kg ha−1 yr−1 are equivalent to those from arable land receiving >200 kg N ha−1 yr−1, and in proportion to the excess N deposited; measurements of N cycling processes and pools using 15N pool dilution techniques show a large nitrate pool and enhanced rates of nitrification relative to immobilization. Ratios of gross nitrification[ratio ]gross immobilization might prove to be good indices of N saturation.

Dr Willem Asman concluded that the major global sources of atmospheric NH3 are excreta from domestic animals and fertilizers. A question raised was: how reliable are the emission estimates and extrapolations? The answer was that emission estimates are surrounded by uncertainty, which is a major handicap to sound modelling of NH3 dry deposition and, consequently, to obtaining good estimates of critical load exceedences.

Major uncertainties in emission estimates seem to be related to the use of simple emission factors, many of which are highly empirical or have been derived from measurements carried out under conditions which deviate considerably from those following modern practices of handling and applying manure and fertilizers. An example is provided by the commonly used emission factors for synthetic fertilizers (see e.g. Bouwman et al. (1997)), which are much higher than recent micrometeorological assessments seem to suggest. Thus, emission from urea, the most widespread fertilizer used in the world (currently around 55% of world N consumption) can be completely avoided if the fertilizer is incorporated into the upper soil layers. Similarly, a growing crop can reduce losses to well below 10% of the applied amount of urea-N, i.e. to less than half of the generally used emission factors of 15% for Europe and 25% for the tropics. The emission factor for NPK-fertilizer is set at 4%, whereas that for pure calcium-ammonium-nitrate, the same N compound as is present in NPK-fertilizers, is assumed to be only 2%.

Dry deposition of NH3 and NOx (NO and NO2) can affect plant metabolism at the cellular and whole-plant level. Gaseous pollutants enter the plant mainly through the stomata, and once in the apoplast NH3 dissolves to form NH4+, whereas NO2 dissolves to form NO3− and NO2−. The latter compound can also be formed after exposure to NO. There is evidence that NH3-N and NOx-N can be reversibly stored in the apoplast. Temporary storage might affect processes such as absorption rate, assimilation and re-emission. Once formed, NO3− and NO2− can be reduced, and NH4+ can be assimilated via the normal enzymatic pathways, nitrate reductase (NR), nitrite reductase and the glutamine synthetase/glutamate synthase (GS/GOGAT) cycle. Fumigation with low concentrations of atmospheric NH3 increases in vitro glutamine synthetase activity, but whether this involves both or only one of the GS isoforms is still an open question. There seems to be no correlation between fumigation with low concentrations of NH3 and in vitro GDH activity. The contribution of atmospheric NH3 and NO2 deposition to the N budget of the whole plant has been calculated for various atmospheric pollutant concentrations and relative growth rates (RGRs). It is concluded that at current ambient atmospheric N concentrations the direct impact of gaseous N uptake by foliage on plant growth is generally small.

High loads of nitrogen to spruce and beech forests can result in a complete inhibition of NO3− uptake by the roots of the trees. This conclusion is based on (a) a comparison of a field site continuously exposed to high loads of N and a N-limited site, (b) the results of N fertilization of a N-limited field site, and (c) laboratory experiments under controlled environmental conditions. From fertilization experiments in the field it appears that NH4+ uptake might become inhibited subsequent to an excessive uptake of NH4+. Apparently, the inhibition of NO3− uptake by high loads of N to forests is a consequence of an accumulation of organic amino compounds in the roots originating from phloem transport from the shoot to the roots. These amino compounds seem to signal the N demand of the shoot to the roots. At present this function cannot be attributed to an individual organic amino compound in beech or spruce, but Gln is a likely candidate in both species among other compounds, e.g. Glu in spruce or Asp in beech trees. Direct inhibition of NO3− uptake by NH4+ can be excluded from the present studies. The mechanism(s) by which elevated levels of particular organic amino compounds interact with NO3− uptake remains to be elucidated. This (these) mechanism(s) seem to affect NO3− influx rather than NO3− efflux. As a consequence of this (these) mechanism(s), spruce and beech trees can prevent, within a certain physiological window, N over-nutrition when the roots are exposed to excessive amounts of inorganic N. However, inhibition of NO3− and NH4+ uptake by the roots makes more N available for leaching into the ground water and, in addition, for soil microbial processes that result in the production and re-emission of volatile N compounds into the atmosphere.

At the ‘Höglwald’ site, continuously exposed to high loads of N, >20% of the N input from throughfall into the spruce and beech plots is re-emitted as NO and N2O. However, the NO to N2O ratio is highly dependent on the tree species, with a preference for NO in the spruce and a preference for N2O in the beech plot. Since at least part of the NO emitted from the soil will be converted inside the canopy in the presence of ozone to NO2 that might then be absorbed by the leaves, the portion of the N in the throughfall that will be released from the forest by gaseous N emission is higher in the beech than in the spruce plot. Leaching of NO3− into the ground water is high in the spruce, but minute in the beech plot. However, this positive effect of beech on ground water quality is achieved at the expense of an enhanced release of radiatively active N gases into the troposphere.

Because nitrogen is the mineral nutrient needed in largest amounts by plants, it is usually also the limiting factor for plant growth in terrestrial ecosystems (Vitousek & Howarth, 1991). Consequently, the deposition of oxidized and reduced N compounds will almost invariably have large effects in these systems, and because N availability not only regulates plant growth but also that of organisms at other trophic levels, disturbances of several ecosystem processes might occur. The alternations introduced by deposition of atmospheric N compounds are both of a quantitative and of a qualitative nature. Moreover, N deposition can have phytotoxic as well as growth-stimulating effects.

This short commentary gives a personal view of some of the possible consequences of N deposition on plants. It refers particularly to the oral presentations given by Professor Heinz Rennenberg and Dr Marta Peréz-Soba, and to the discussions held after their talks where appropriate. Separate attention is given to four different consequence of anthropogenic N-deposition: N-availability; N-form, N-uptake by the shoot, and the period of N-uptake. Finally, I have tried to adopt an ecosystem perspective and discuss briefly the concept of critical loads of N.

Tropospheric ozone can affect crop yield and has been reported to cause reductions in growth and biomass of forest tree species in laboratory and glasshouse studies. However, linkages between growth and ambient ozone concentrations in the field are not well established for forest trees. Ambient ozone concentrations have been shown to cause foliar injury on a number of tree species throughout much of the eastern USA. Symptom expression is influenced by endogenous and exogenous factors and, therefore, ozone-exposure/tree-response relationships have been difficult to confirm. Clearly defined, cause-effect relationships between visible injury and growth losses due to ozone have not been validated. Generalizations of sensitivity of forest trees to ozone are complicated by tree development stage, microclimate, leaf phenology, compensatory processes, within-species variation and other interacting stresses. In general, decreases in above-ground growth at ambient ozone levels in the eastern USA appear to be in the range of 0–10% per year. However, these conclusions are based on a small number of tree species, with the vast majority of studies involving individual tree seedlings in a non-competitive environment. Comparative studies of small and large trees indicate that seedlings are not suitable surrogates for predicting responses of mature trees to ozone. Process-level modelling is a promising methodology that has been recently utilized to assess ozone effects on a stand to regional scale, indicating that ozone is affecting forest growth in the eastern USA. The extent and magnitude of the response is variable and depends on many edaphic and climatic factors. It is imperative when conducting assessment exercises, however, that forest biologists constantly keep in mind the tremendous variability that exists within natural systems. Scaling of single site/physiological response phenomena from an individual tree to an ecosystem and/or region necessitates further research.

In comparison with the effects of extended drought periods or severe nutrient stress, those of ozone are generally much milder, at least with respect to growth. However, there is substantial evidence from experiments, in the main using young saplings, that O3 does impose a stress on forest trees under European conditions. Decreased chlorophyll contents and photosynthetic rates, changes in carbon allocation, increased antioxidant activity, and reductions in biomass due to O3 have often been recorded, particularly in fast-growing species. Furthermore, 3 appears to weaken the trees' resilience to a range of biotic and abiotic stresses. Interactions between O3 and climatic stress, in particular drought and frost hardiness, are likely to result in potentially detrimental effects.

A link between the occurrence of O3 and forest damage is not unequivocally established in Europe, and the problem remains of extrapolating and/or scaling up from studies on seedlings to predict responses to O3 of mature trees and forest stands, because we know so little about acclimation to O3. An accurate assessment is also lacking of the magnitude of the O3 effect on European trees both in terms of the forest areas affected and its extent. In this review we suggest that C allocation is the key factor underlying the responses of trees to O3. Stomata also play a key role, since the acquisition of C must be achieved while an effective control over water consumption is retained.

Current estimates of forest yield losses attributable to ozone pollution amount to c. 10% over Europe as a whole. This figure is derived from a synthesis of all European studies using trees for which AOT40 exposure values are available. However, the choice of 40 nl l−1 as the threshold concentration for demonstrable effects has led to debate, and this value might not be low enough to predict ozone effects in Scandinavia, where concentrations are lower than in southern Europe, and chronic injury resulting from cumulative exposure is observed. This ‘level I’ approach provides a useful means for mapping physiologically effective concentrations but has significant shortcomings in that it is unable to take environmental conditions into account. In order to produce a mechanism capable of predicting yield losses resulting from ozone pollution at specific sites and for individual species, the effective ozone dose, a product of conductance and concentration, must be calculated. Process-based models relating the environment (temperature, humidity/saturation deficit, incident light and soil moisture content) to conductance are available for a number of species (oak, Scots pine, Norway spruce, Sitka spruce, beech and poplar). Effective ozone doses could therefore be calculated, and the relationships between effective dose and yield loss could be determined by revisiting existing data for which only concentrations or AOT40 values are available at present. Yield loss must be the growth parameter with which ozone damage is expressed, since visible injury does not necessarily represent severe injury to the plant, whilst visibly unaffected plants may be significantly compromised in terms of biomass accumulation. For conifers, premature needle drop, although indicative of ozone pollution, might not represent a significant reduction in growth since older needles are, functionally, relatively unimportant.

When revisiting old experiments the question should also be asked ‘what is an appropriate control treatment’. It is unrealistic to expect ambient O3 concentrations on a global scale to return to pre-industrialization levels, and therefore the use of a charcoal-filtered treatment is probably unrealistic. In addition, charcoal filters will remove some of the NOx, and therefore on nutrient deficient soils (typical of many forest soils) the ‘control’ treatment might represent a reduced supply of nitrogen, making it difficult to disentangle the ozone effect per se.

The stage has therefore been reached where ‘level II’ ozone exposure mapping (i.e. based on effective dose at the physiological level) is possible for a number of species. As financial considerations become increasingly important, the scientific community must aim to provide better estimates of O3-induced yield losses so that long-term environmental audits can be performed.

Evidence that enhanced reactive nitrogen deposition is affecting semi-natural terrestrial ecosystems comes from historic increases in plant tissue N concentrations, correlations between tissue N concentrations and present-day total atmospheric N deposition, changes in plant amino-acid composition and effects on N assimilation. The ecological significance of such changes in biomarkers is uncertain. This paper explores the ecological significance of reactive atmospheric N deposition through a review of previous experimental findings and new experimental evidence from an acidic and a calcareous grassland, both showing phosphorus limitation, and a N-limited Calluna vulgaris (L.) Hull heathland in upland Britain. Nitrogen addition in the range 0–20 g N m−2 yr−1 initially (years 0–4) increased the growth of Calluna and a decline in some subordinate species. In subsequent years, shoot extension was not stimulated, but winter injury was observed from 1993 onwards, suggesting a strong interaction between N supply and climatic conditions. By contrast, the grasslands showed a small decrease in the cover of higher plants in later years (6–7) of the experimental treatments (0–14 g N m−2 yr−1) and no growth stimulation. All N treatments reduced the bryophyte cover in the acidic grassland. There were marked effects on below-ground processes, including a sustained stimulation of N mineralization in the grassland soils, and an increase in the bacterial utilization of organic substrates in the heathland, as measured in BIOLOG plates. The results strongly suggest the importance of atmospheric N deposition on microbially driven processes in soils, and are discussed in relation to the scale of potential ecosystem changes and their reversibility by pollution abatement.

Although there is a great deal of concern about the effects of human activities on biodiversity, until recently there has been very little interest in the effects of ozone on the species that constitute the major part of the flora, the diverse herbaceous and shrubby species of natural and semi-natural communities. However, many wild species have been shown to be at least as sensitive to ozone as crops that show significant yield losses, so there is a pressing need for an evaluation of the risk to wild species posed by ozone. This review attempts to assess progress and highlight problems. It begins with a comment on semantics, discusses the difficulties involved in measuring relative ozone resistance and then proceeds to consider the effects of ozone on growth and resource allocation. The evidence for evolution of resistance is appraised and then the potential effects of several interactions (cutting/grazing, competition, soil water deficit and nutrition) are considered. The review ends with some remarks on observation of oxidant-induced changes in ecosystems.

Nitrogenous inorganic compounds impact plants as nutrients, signals and toxins. We are dissecting a regulatory network that controls nitrate assimilation at the level of nitrate reductase (NR) activity. The identification of protein kinase cascades, protein phosphatases and 14-3-3 proteins as regulators of NR are giving clues about how plants sense their nutrient availability, and use the information to signal changes in their metabolism and developmental strategies to cope with supplies. We hope that understanding these controls might lead to the design of transgenic plants with deregulated signalling networks, which would make them more efficient in using nitrogen fertilizers, and improving quality and yield of crops. There are circumstantial indications that gaseous anthropogenic nitrogenous emissions might also have complex regulatory influences on plant growth and development.

Man's activities pose a number of threats to the functioning, structure and diversity of natural and semi-natural ecosystems. One of the main threats is the increase in concentrations in air pollutants in this century (Wellburn, 1988; Tamm, 1991). This paper is a commentary on the effects of tropospheric ozone (O3) and airborne nitrogen deposition (both oxidized (NOx) and reduced (NHy) compounds) on natural and semi-natural ecosystems, based upon the oral presentations and the discussions during the Symposium, extended with a personal overview and some suggestions about future challenges for research. The most important effects of these air pollutants on natural and semi-natural vegetation are summarized and evaluated in ecological terms, with respect to the functioning and structure of unaffected systems. Air pollutants are transported over both short and long distances (as far as a few thousand km) before being deposited on surface water, vegetation or soil. In this way, vegetation over a large area or in remote regions can be influenced by airborne pollutants (see Fowler et al. (1998); Asman, Sutton & Schjørring (1998)).

As a result of increasing anthropogenic nitrogen deposition, N availability in many forest ecosystems, which are normally N-limited, has been enhanced. We discuss the impacts of this increased N availability on the ectomycorrhizal (ECM) symbiosis which is generally regarded as an adaptation to nutrient limited conditions. Nitrogen deposition can influence fruit-body formation by ECM fungi, the production and distribution of the extraradical mycelium in the soil and the formation of ectomycorrhizas.

Available data from long-term N deposition studies indicate that the most prominent effects might be discernible above-ground (i.e. on the formation of fruit bodies). ‘Generalist’ species, forming a symbiosis with a wide range of tree species, seem to be less affected by increased N availability than ‘specialist’ species, especially those living in symbiosis with conifers. However, the importance of below-ground investigations to determine the impacts of N deposition on the ECM symbiosis must not be underestimated. Culture experiments show an optimum N concentration for the formation of extraradical mycelium and mycorrhizas. Often, negative effects only become visible at comparatively high N concentrations, but the use of a few easily cultivated species of ECM fungi, which are adapted to higher N concentrations, undermines our ability to generalize.

So far, N deposition experiments in the field have only shown minor changes in the below-ground mycorrhizal population, as estimated from the investigation of mycorrhizal root tips. However, effects on the ECM mycelium, which is the main fungal component in terms of nutrient uptake, cannot be excluded and need further consideration.

Because the photoassimilate supply from the plant to the fungal partner is crucial for the maintenance of the ECM symbiosis, we discuss the possible physiological implications of increasing N inputs on the allocation of C to the fungus. Together with ultrastructural changes, physiological effects might precede obvious visible changes and might therefore be useful early indicators of negative impacts of increasing N inputs on the ECM symbiosis.

The global cycles of carbon and nitrogen are being perturbed by human activities that increase the transfer from large pools of non-reactive forms of the elements to reactive forms that are essential to the functioning of the terrestrial biosphere. The cycles are closely linked at all scales, and global change analyses must consider C and N cycles together. The increasing amount of N originating from fossil fuel combustion and deposited to terrestrial ecosystems as nitrogen oxides could increase the capacity of ecosystems to sequester C, thereby removing some of the excess carbon dioxide from the atmosphere and slowing the development of greenhouse warming. Several global and ecosystem models have calculated the amount of C sequestration that can be attributed to N deposition, based on assumptions about the allocation of N among ecosystem components with different C[ratio ]N ratios. They support the premise that, since industrialization began, N deposition has been responsible for an increasing terrestrial C sink, but there is great uncertainty whether ecosystems will continue to retain exogenous N. Whether terrestrial ecosystems continue to sequester additional C will depend in part on their response to increasing concentrations of atmospheric carbon dioxide, widely thought to be constrained by limited N availability. Ecosystem models generally support the conclusion that responses to increasing concentrations of carbon dioxide will be greater, and the range of possible responses will be wider, in ecosystems where increased N inputs originate as atmospheric deposition. The interactions between N deposition and increasing carbon dioxide concentrations could be altered considerably, however, by additional factors, including N saturation of ecosystems, changes in community composition, and climate change. Nitrogen deposition is also linked to global change issues through the volatile losses of nitrous oxide, which is a potent greenhouse gas, and the role of nitrogen oxides in the production of tropospheric ozone, which could interact with plant responses to elevated carbon dioxide. Any consideration of the role of N deposition in global change issues must also balance the projected responses against the serious detrimental impact of excess N on the environment.