Soil Functional Responses to Excess Nitrogen Inputs at Global Scale

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Mark Adams, Phil Ineson, Dan Binkley, Georg Cadisch, Naoko Tokuchi, Mary Scholes and Kevin Hicks. There is little evidence that nitrogen (N) cycling in the.
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Mark Adams, Phil Ineson, Dan Binkley, Georg Cadisch, Naoko Tokuchi, Mary Scholes and Kevin Hicks

Soil Functional Responses to Excess Nitrogen Inputs at Global Scale There is little evidence that nitrogen (N) cycling in the highly weathered, low-phosphorus (P), acidic soils found in Southern Hemisphere continents will differ greatly from that in North America and Europe. Evidence from the ʻsouthʼ shows: the similarity in forms and temporal patterns in losses of N from different land uses; that the C:N ratios of the forest floor/litter layer from different continents are strongly predictive of a range of processes on a global scale; that generalizations based on Northern Hemisphere experience of the impact of N additions to ʻP-limitedʼ ecosystems are likely to fail for southern ecosystems where anatomical and physiological adaptation of native plants to low-P soils makes questionable the concept of ʻP-limitationʼ; that the greatest threats in the ʻsouthʼ are probably changes in land use that may greatly increase N inputs and turnover; that localized increases in N inputs produce similar effects to those seen in the ʻnorthʼ.

INTRODUCTION This review considers some of the key attributes that may be used on a global basis to predict the response to excess N addition of functions of terrestrial ecosystems. Previous reviews have focused on Europe and North America, here we place emphasis on southern ecosystems. Rather than attempt an overview of the whole N-cycle, we have selected functions where we have either new knowledge or where an insight of global relevance might be obtained. While it is obvious that the major processes of the N cycle are ubiquitous, differences in climates, management and soils modify the cycle sufficiently to suggest that increased N deposition will have differing effects in different parts of the world. It has been argued that these differences may be so great that the addition of N to tropical and temperate forests will initially have exactly the opposite effects on C storage in the two different systems (1)—increasing carbon storage in temperate forests as a result of alleviating N limitations but reducing storage in tropical forests owing to enhancement of P-limitation through acidification and other indirect effects of N additions. For much of the tropics and the southern continents, forms and amounts of N added to natural and agricultural systems will differ from those in the north, largely as a result of different patterns of industrial development and, especially, agricultural land uses. For example, Galloway (2, 3) has drawn attention to likely deleterious consequences of the sharply increasing fertilizer inputs of reactive nitrogen to agroecosystems in Asia. The starting point for our comparison is a consideration of how ecosystems differ between the ‘north’ and ‘south’ and how these differences may interact with increased N inputs. Here we devote more attention to soils than to other issues or factors. 530

NUTRITIONAL STATUS OF SOILS Nutrient Limitations A comparison of temperate and moist tropical forests by Matson et al. (1) highlighted the nature of nutrient limitations, if any, as a key difference between the systems. Terms such as ‘nutrient-limited’ should be used with care owing to significant species-specific differences in nutrient requirements and ability to acquire nutrients. In many parts of the ‘south’, soils are ‘poor’ for introduced species rather than for the native species. For example, Oxisols in the tropics that have been used extensively for agriculture are widely regarded as ‘infertile’ yet the same soils support native forests of high productivity and the soils (or rather the system) are, within that context, clearly ‘fertile’. Similarly, whilst forests have been a natural focus of much research in Europe and North America, a more global view requires consideration of a much wider variety of land uses. More specifically, while we might accept that many tropical soils are strongly weathered, contain low amounts of available P (not necessarily total P), are rich in oxides of Al and Fe, and are clearly P-deficient for crops or pastures, there is poor evidence to support arguments for P-limitation of native vegetation (forests, savannas, scrub) on soils that have not been previously used for agriculture. Thus, although our fundamental understanding of the processes within the N cycle can be applied to any terrestrial ecosystem, the relative importance of a particular function within the cycle may differ considerably between different regions of the world and according to land use and land use history. In the examples that follow, we place some emphasis on the interaction of water and nutrients as limiting factors. In many semi-arid savannas and grasslands, for example, it is the irregular supply of water that limits the rate of nitrogen mineralization (4, 5). The following issues help to focus predictions of the likely consequences of excessive additions of N to the increasingly developed ‘south’, and perhaps improve the predictability of responses of ecosystems in the ‘north’ where both deleterious and beneficial impacts have been noted (6). P-status and the N-cycle There is little evidence that N-cycling processes differ greatly among soils of widely different P-status. For example, N-mineralization was not limited by a shortage of P in Oxisols in Latin America (7). In the semi-arid savannas of South Africa, the rate of nitrification is the limiting step in the nitrogen cycle and is controlled mostly by water availability; in laboratory studies, additions of phosphorus to soil samples had no impact on nitrification (4). In the eucalypt forests that cloak most of the higher rainfall regions of Australia, and where P availability is poor by world standards, rates of nitrogen mineralization and nitrification are related to soil C:N ratios (8) as they appear to be throughout the temperate regions (9). Figure 1 illustrates the importance of the C:N ratio for the relative rates of a range of N-cycling processes, including: – nitrification in a range of Australian forest soils (8);

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– nitrification in forest soils in Germany (10); – nitrogen oxide fluxes in humid wet and semi-deciduous dry tropical forests soils in Puerto Rico (11); – nitrogen leaching in broadleaf and coniferous forests in Europe (12). Clearly, these relationships between the C:N ratio of the forest floor/litter layer and the production of nitrate (NO3- ), denitrification, and NO3- leaching are remarkably similar across all locations, despite what must be wide differences in P status. The rates of these processes are strongly regulated by C:N ratios less than 20, but largely independent of C:N ratios greater than 25 where they are thought to be more dependent on factors such as moisture and temperature status. We must bear in mind that the C:N ratios themselves reflect climate, productivity and species composition. Biological Nitrogen Fixation N2-fixation is frequently cited as being limited by shortages of P in soils (13). We might logically expect such limitations of N2-fixation on the low-P soils of Africa, Australia and South America. However, hard evidence supporting this hypothesis is limited. For example, in summarizing much of the work on N2-fixation in the Fynbos vegetation of South Africa, Stock and Allsopp (14) noted that soil type, especially soil physical properties, and consequent patterns of plant growth were the major influences on the large variation in nitrogen availability under introduced nitrogen-fixing Acacia spp. rather than the rate of fixation per se. We also lack definitive evidence that phosphate concentrations in soil solutions limit either nodule numbers and functioning, or the growth of free-living, root nodule bacteria (15). In semi-arid savannas, N2-fixation in Acacias appears to be most important during early stages rather than in maturing populations. In Australia, Hanson and Pate (16) demonstrated an increase in N2-fixation by Acacia spp. after application of P fertilizer, but additional (to rainfall) water was also applied. In a recent study of Acacia and Senna spp. in the Pilbara, northwestern Australia (see below), growth did not respond to added P and there was no evidence of an increase in N2-fixation (17). Binkley et al. (18) reported that the uptake and cycling of P was similar for matched plots of eucalypts and the legume Albizia facaltaria

Figure 1. Various nitrogen cycling functions (data points for the different parameters were ʻnormalizedʼ by calculating the percentage of the maximum value in each case) plotted against the C:N ratio of forest floor/litter layer: (ο) nitrate production in humus samples from northwestern German forest soils (100 = 68.1 ppm) (10); (�) nitrogen oxide fluxes (N2O and NO) from humid wet and semideciduous dry tropical forests soils in Puerto Rico (100 = 14.5 ng N cm-2 hr-1) (11); (Δ) nitrate leaching (100 = 40 kg N ha-1 yr-1) of forest floor at 33 temperate forest sites (2 broadleaf and the rest coniferous) in Europe from the Element Cycling and Output-fluxes in Forest Ecosystems in Europe (ECOFEE) database (12); (•) nitrification (100 = 0.25 µg g-1 d-1) in a range of Australian forest soils (8). Ambio Vol. 33, No. 8, December 2004

in Hawaii and whilst this comparison showed that supplies of P in soil were influenced by species composition, the similarity in amounts cycled and taken up owed much to the ability of the legume to exploit a larger available soil volume or the same volume more effectively. Ndufa et al. (19) showed that apart from an initial boost in performance, tropical shrub and tree legumes were not limited by P when growing on an Oxisol in Kenya. None of these examples from Africa, Australia or Hawaii support the hypothesis that N2-fixation of trees is limited by the availability of P. More likely, within-plant allocation of P to growth and to N2-fixation is linked to demand and to other factors that limit growth, such as water availability. Mechanisms that Determine P Availability In addition to examples from the old and P-poor soils in southern continents, we might consider remnants of old landscapes in Sweden and Finland. At Betsele, Sweden, forest growth was limited by phosphorus availability in low-lying areas (but not further upslope) where there was also high N availability and turnover but high capacity for fixation of P (20). The strong relationship at Betsele between productivity and topographic location, and the degree of independence of N and P availability, remind us that N and P availability are governed by substantially different processes (biological processes for N, bio- and geochemical processes for P (21)). At the present time, most evidence suggests that the nature of N-cycling processes are often independent of the P-status of native ecosystems on old soils. While Matson et al. (1) hypothesized that additions of N to the strongly P-fixing soils found in parts of the tropics may exacerbate P-limitation, owing to a reduction in soil pH, the adaptation of native plants to low P-availability is likely to be a strong buffer to any such change in either N-cycling processes or, for that matter, ‘P-limitation’ of native plants. Nevertheless, the combination of old and low-P soils and water-limited environments raises special issues for consideration in specific systems. Owing to the naturally low solubility of P in soil, it is always difficult to separate P availability from water availability, and, for example, sclerophylly has long been associated with water and P-poor ecosystems (22, 23). Some sclerophyllous species (e.g. members of the Proteaceae) are so well adapted to such conditions that they are sensitive to additions of inorganic P and N (24–26) and at the community level, both increases and decreases in growth have been observed in response to nutrient additions (14). To our knowledge, there are no rigorous field studies of the effects of chronic additions (as opposed to single or infrequent large additions) of N to these species. Even the most P-poor soils can provide adequate supplies of P to natural plant communities via mineralization of organic P during the short periods of moist soil or of more mesic temperatures after long periods of drought or extreme temperatures (27). In Eucalyptus marginata forests, growing on exceptionally low-P soils with extreme variability in seasonal moisture in southwestern Australia, phosphatase activity and indices of microbial and fungal activity and biomass were seasonally variable but not impaired beyond subsequent recovery. In nutrient-poor heathlands (28) and grasslands (29) the species composition of plant communities had a marked effect on the ability to mineralize organic P, with supplies of P seemingly limiting growth and other processes in N-enriched communities. Johnson et al. (29) demonstrated increased activity of root surface phosphatase in plots that had received long-term additions of N, at least for some species, but much further research is needed in this area. Differences among species in their ability to acquire the P needed for growth are clearly striking. At one extreme there are fast growing crops with high nutrient demands and poor capac-

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ity to acquire fixed P. Clearly, N additions to these agricultural systems will lead to increased incidence of N leaching. At the other extreme is the Proteaceae family that has evolved on old soils, producing cluster roots and excreting large quantities of organic anions and phosphatase enzymes in short ‘exudative bursts’ to acquire P from sources unavailable to other plants. Cluster roots are far more common than was originally thought; especially and curiously among leguminous species (30) and further research will undoubtedly find many more cluster-rooted genera and species in the south. Most of the rest of the native plant genera in the south are well-adapted to Ppoor soils via mycorrhizal associations, as are the majority of tropical plants. There is some evidence that the evolution of mycorrhizal symbioses over 400 million years has pre-adapted plants for the evolution of nodulation (31) and it will be interesting if the same applies to the cluster-rooted species. Some of the contradictory findings on the role of phosphorus in nitrogen fixation capacity may well be due to confounding of mycorrhizal and cluster root status (e.g. both mycorrhizae and cluster roots are inhibited by high concentrations of P in the root zone) with nodulation and/or activity of diazotrophic bacteria. Other Influences on Soil Nutrient Status The Pilbara in northwestern Australia is home to some of the world’s oldest soil parent materials (> 2500 Ma) and landscapes (32, 33). The ecosystems are largely fire-prone, and subject to extreme variation in rainfall (in any one year from zero to 800 mm) whilst evaporative demand is close to 2500 mm every year. The natural grasslands (Themeda, Astrebla, Aristida spp.) in the Pilbara have been grazed by native animals for millennia and by introduced herbivores for the past 100 years. There are, for literally tens of thousands of hectares, few leguminous species to replace the N (and P) lost through fire especially but these systems remain as productive as the water balance will allow. In general, these systems respond poorly or not at all to additions of P but do respond to additions of N in wetter years (34, 35). In African savanna ecosystems, where large amounts of N are also lost due to fire (4, 36), 15N natural abundance techniques applied in field studies suggested Acacias and herbaceous legumes make a significant contribution (approx 5–20 kg N ha-1 yr-1) to the Ncycle. The more humid native savanna systems of tropical Latin America do not respond to N without additions of P and in some cases, calcium (37). The interactions of fire and herbivory add some complexity to our understanding of water and phosphorus constraints of biological nitrogen fixation and notions of nutrient limitation. The inverse relationship in some ecosystems between fire frequency and ecosystem water balance, make this a difficult field in which to generalize. Summary: Soil Nutrient Status As was emphasized by Matson et al. (1), P status may influence ecosystem responses to N additions, but species composition, land use and land use history, including fires and herbivory, will modify any such response and it is too simplistic to assume that temperate and tropical systems will fall into a few convenient categories. Mechanisms that control ecosystem N and P limitation will obviously play a role in determining the consequences of increased N and P deposition (38). However, it is demonstrable that co-evolution of plants with low P soils in the south, including much of (wet and dry) tropical Australia and Africa, has produced a remarkable range of anatomical and physiological features that together render close to redundant the notion of ‘P-limitation’ in many native plant communities. 532

WILL SOIL AND LITTER QUALITY REGULATE RESPONSE TO N ENRICHMENT? In addition to differences among individual plant species in P nutrition, specific plant attributes and their interactions with N-enrichment may also impose other important ecosystem effects. The nitrogen content of litter is a major influence on its decomposition, with Melillo et al. (39) relating rates of mass loss to initial lignin-to-N ratio. With many residues from annual crops the C:N ratio alone is a reasonable predictor of decomposition rates but with litters from woody perennials, or when the C:N ratio rises above ca. 75, then the lignin: N ratio or the (lignin+polyphenol): N ratio become better indicators of decomposition rate. Increasing N loads can decrease the C:N ratio and polyphenol content of plant residues (40) and hence potentially increase decomposition, although this might be moderated to some extent due to retranslocation during senescence. Fog (41) reviewed the existing literature on the impact of increasing N input on decomposition processes, either when the extra N was added during plant growth or when added to the decomposing litter. The conclusion was that for nonrecalcitrant substrates, with low initial C:N ratios, the addition of N increased or did not change decomposition rates. In contrast, ‘poorer’ more recalcitrant substrates, showed the reverse trend with reduced decomposition under higher N inputs, believed to be the result of inhibition of specific groups of decomposer fungi. Generalizing these observations to the long-lived foliage of sclerophyllous species in the natural ecosystems in much of Africa, Australia and South America—we might suggest that the dominance of recalcitrant, high-lignin residues may result in a reduction in decomposition rates when N is added to excess. LOSSES OF N FROM SOILS Losses of N: Gaseous Denitrification to N2 is a major pathway for removal of excess N from terrestrial ecosystems. However, nitrous oxide (N2O) is frequently produced during the nitrification/denitrification series of reactions and exerts a strong global influence as one of the principal greenhouse gases (42). With an atmospheric concentration currently increasing at 0.25% yr-1 and a long atmospheric life of ca. 120 years, N2O production is one of the key negative consequences of excess N addition in both terrestrial and marine systems (42). Anthropogenic changes in the use of both N and land are central to the observed global increases in atmospheric N2O concentrations, with agricultural soils being the most important source of N2O (43). It has been demonstrated on numerous occasions that increased inputs of mineral N to terrestrial ecosystems will stimulate denitrification and that amounts of N2O emitted from agricultural soils correlate with the amounts of N fertilizer applied (44). Measured and modelled increases in N2O emissions from terrestrial systems are attributed to a combination of increased fertilizer usage, agricultural N2-fixation and N deposition (45). Of particular concern are tropical ecosystems where the combination of high intrinsic N2O production rates, coupled with projections of increasing N-fertilization and deposition, suggest significant increased fluxes of this gas in the future (1). The promotion of legume-rich, short-term fallows (e.g. in Kenya (46)) will further increase N2O fluxes. Phosphoruslimited agricultural systems in the tropics may shed excess N as NO and N2O, with rates of production being orders of magnitude higher than comparable N-limited systems (47). Fluxes from native plant communities on tropical soils may

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often be equivalent to those from agricultural soils in temperate regions. On the other hand, within semi-arid savannas N2O losses may be negligible (48) whereas losses of NO can be considerable (49). Predictably, this trend for increased NO emissions in drier conditions is also described for tropical dry forests (e.g. in Puerto Rico (11)). Losses of N: Leaching of Inorganic N An additional and important route for N leakage from ecosystems is as dissolved inorganic and organic nitrogen. NO3- leaching is often the first measurable response to increased N inputs, and is geographically widespread in Europe and North America, and can occur in the absence of visible damage to the vegetation itself (50). Indeed, there is good empirical evidence (51) that combining rates of N input with soil C:N ratios can provide reasonable predictions of the rate of leaching from European forests. Limited NO3- leaching is considered a normal feature of the N cycle in European forest ecosystems, even at low ‘background’ levels of N. Recent evidence from afforested catchments in South Africa shows that NO3- is leaching from pine plantations (52). These pine plantations are grown mostly for pulp using short rotations. Adjacent, high altitude grasslands—the natural vegetation—do not show any NO3- leaching. N deposition in these areas, from fossil fuel burning for power generation, is high, averaging 25 kg N ha-1 yr-1. In general and compared to background rates, there seems to be strong evidence that leaching losses of NO3- increase with increasing N inputs, and in some cases there may be a loss of seasonality in the pattern of NO3- leaching (9, 51, 53). The question remains as to the extent to which two-way information transfer between Europe and North America on the one hand and the rest of the world on the other can lead

system is initiated (12). For example, C:N ratios in coniferous stands of > 30, 25 to 30, and < 25 can be used to separate low, moderate, and high NO3- leaching loss, respectively. It would be interesting to see how widely applicable this scheme is within a global context, but we lack data for large areas. A C:N ratio of 20–25 has also been established as a ‘threshold’ for an increase in N oxide fluxes (11, 46); the C:N ratio may well be a globally applicable diagnostic feature of ecosystem N status. Populations and accompanying economic activity, especially agriculture, have increased significantly in recent years in many countries and regions outside Europe and North America. While the concentration of heavy industry and fossil fuel consumption in the ‘south’ seldom matches that of the more developed ‘north’, the green revolution spawned a dramatic and continuing increase in the use of fertilizers in agriculture, especially in Asia. For the Changjiang River, “the 2-fold increase in NO3- between 1980 and 1992 is more closely related to a doubling of fertilizer use in China during the same period” (58). In countries like Australia, overuse of fertilizers has led to significant reductions in the quality of groundwater, streamwater, lakes and estuaries. Caraco and Cole (58) used a global data set to examine NO3export from large rivers. One notable feature of rivers that drain drier regions (including the Orange, Zambesi and Nile rivers in Africa and the Murray in Australia) (Fig. 2) was noticeably reduced NO3- export vis-a-vis rivers in more ‘mesic’ regions (most from USA and Europe). Several mechanisms were postulated to account for the difference including ‘residence times’ that are greater (for a variety of reasons) in more xeric systems – thereby providing more opportunity for NO3- removal processes such as denitrification. As increased concentrations of NO3- in soil water, groundwater and streamwater are widely used as indicators of ‘N-saturation’, increased capacity of xeric rivers to retain or

Figure 2. Nitrate N export from large rivers with global distribution. (a, left) Relationship to population density in the catchment. (b, right) Relationship to export predicted on the basis of human activity and water runoff. (•) xeric systems (runoff < 0.1 m yr-1), (o) mesic systems (58).

to improved understanding and/or management of excess N inputs. Evidence of ‘nitrogen saturation’ in Japanese forests was recently presented by Ohrui and Mitchell (54) but such examples are still rare outside the USA and Europe. Experiments investigating the fate and impacts of inorganic N additions to forest ecosystems within the USA and Europe have shown the form of N added may be critical in deciding the fate of the N. The quantity of added N assimilated by trees may be unchanged (55) or clearly modified (56) by the form of added N. In general, NO3- additions leach more readily than equivalent inputs of NH4+ and this has been demonstrated both at the pot (57) and ecosystem scales (55). Some forests may be specifically ‘NO3- saturated’ rather than simply nitrogen saturated (50). As also illustrated by Figure 1, the C:N ratio of the forest floor (the whole organic litter layer without freshly fallen litter) is an important indicator of the point at which NO3 leaching from the Ambio Vol. 33, No. 8, December 2004

process N may well mask, or at least delay, signs of increasing additions of N. High elevations in tropical mountains are also more arid than their counterparts at middle latitudes (59) and, consequently, both scleromorphic characteristics of vegetation and xeric characteristics of streams are more likely in the tropics than may have been previously recognized. Losses of N: Leaching of Organic N Research in the ‘south’ has helped clarify the significance of leaching of organic N, an issue of considerable recent interest. Organic N dominates the N-load of rivers and streams draining from the forests of North America (except for hardwood forests in the northeast (60)), in much of Australia (61), and South America (for example in Chile (62)). For Australia, only after streamwater enters agricultural zones does inorganic N become

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a significant component of the total-N load (Fig. 3, from 61), a situation found worldwide. This scenario suggests that ecological attributes of remnants of riparian native vegetation, within otherwise agricultural landscapes, may be threatened by NO3-

Figure 3. Concentrations of total N and total P in the Murray River and tributaries, Australia. (X) forested subcatchments in Victoria, (•) agriculturally dominated reaches in Victoria, (∆) Murray River. Also shown are (ο) Murray River data for the ratio of dissolved inorganic nitrogen to soluble reactive phosphorus (DIN:SRP). The line indicates Redfield stoichiometry (15:1) (61).

-enriched drainage waters. There is some evidence for such a threat (63). However, the general case is still that dissolved organic nitrogen (DON) dominates the N content of drainage water from undisturbed systems (64). Coats and Goldman (65) concluded that over 90% of the total N load in basin streams was present as DON and Goodale et al. (66) estimated that between 28–87% of total dissolved N export was in the organic form. The observation that the high levels of NO3- found in streamwater draining forests in northeastern USA are seldom found in unpolluted and remote forested watersheds (62), led to the suggestion that dominance of NO3- is a consequence of human influence (67). Indeed, the data presented by Perakis and Hedin (62) suggest that concentrations of DON are relatively constant across a wide range of contrasting forest systems and that NO3dominates in more polluted regions. The significance attached to the work by Perakis and Hedin is perhaps a little surprising given that we have known for some time (61) that DON losses from unpolluted forests can rival rates of NO3-N losses from regions receiving more recent anthropogenic inputs. Evidence from microcosms along N deposition transects, together with experimental manipulations, has shown an increase in DON with increasing inorganic N deposition in highly organic soils (68, 69). In contrast, additions of NH4NO3 to a forest system in Switzerland showed an increase in NO3- leaching but not in DON flux (70) and Aber et al. (71) suggested that DON outputs did not respond to increased inorganic N loads. In many forests, DON fluxes from the forest floor into the mineral soil are largely dependent on the water flux (72) and high DON fluxes have been recorded just after snow melt. THE IMPORTANCE OF FIRE Many studies point to the importance of fires to C and N stocks and N availability (Wardle et al. (73); see especially Stock and Lewis (74) and Stock et al. (75) for influences on Fynbos vegetation; Attiwell and Adams (76) for influences on eucalypt forests). Fires are a primary mechanism contributing to the Nlimitation of fire-prone ecosystems (77) but have been traditionally and successfully used in tropical Latin America to manage native savannas to remove excess C and improve net N mineralization and fodder quality. The reduced frequency of fires in 534

a great many landscapes in Europe and North America, as well as elsewhere, is a recent phenomenon on ecological time-scales and likely to have a continuing role in enhancing a build-up of N in terrestrial ecosystems. Comparisons of ecosystem responses to added N between the old landscapes in Sweden and Finland and their counterparts in the Southern Hemisphere, could help identify beneficial, N-removing effects of fire vis-a-vis nitrogen additions from the atmosphere. It is clear both from experience and modeling that fire is a major management alternative for the removal of excess N that has accumulated from the atmosphere and that the management of fire will be crucial in determining how increasing N deposition impacts ecosystems in the developing countries. Only a few studies in southern ecosystems have focused on the effects of fire frequency on soil carbon and nutrients. In South African savannas, organic carbon increased by 15% in fire-protected areas while annual burning significantly decreased the light fraction carbon by 25%. (78). Soil nitrogen may decline with increased fire frequency (79), but whether nutrient availability increases or decreases under contrasting fire frequencies depends on fire intensity, soil type and changes in primary production. Frequent fires may reduce water availability for plants not only by reducing the mulching effect of the vegetation and litter but also by intensifying runoff (78). The natural abundances of isotopes of N in plants and soils have been used as indices of nitrogen enrichment of European ecosystems (80–82) with the relative abundance of 15N increasing with the N load and especially with increasing losses of 15Ndepleted NO3- in drainage waters. The use of 15N signatures as indices of nitrogen enrichment must be set against a more general picture, as outlined by Handley et al. (83), in which δ15N signatures increase with the residence time of N within ecosystems. The long-lived foliage of sclerophyllous species and tight within-plant cycling of N in the natural ecosystems in much of Africa, Australia, and South America, as well as in a surprisingly large proportion of tropical ecosystems, accounts for some of the strong 15N enrichment seen in these regions in the south (84). Recent evidence (79) illustrates the role of fire in enriching ecosystems with 15N at the expense of 14N and provides an alternate hypothesis to that proposed by Högberg (85) and Grogan et al. (86) to account for 15N enrichment of fire-prone vegetation. Given these and other processes that lead to 15N enrichment, it will be rather more difficult to use δ15N signatures as indices of nitrogen enrichment. This will be particularly the case in waterand P-poor ecosystems, and especially ecosystems where fire is a regular but unpredictable influence, without both a better knowledge of background δ15N signatures and a means of discriminating among processes contributing to 15N enrichment. CONCLUDING REMARKS An understanding of the interaction of climate, soil type and species composition on N cycling processes is fundamental to predicting the fate of N in ecosystems. Key diagnostic factors of ecosystem response to N addition are: i) limiting nutrient(s), if any, and differences among genera and species in nutrient requirements and modes of acquisition; ii) the C:N ratio of the soil organic horizons; iii) management regime (history of land use, species composition and fire); iv) the nature, mode and seasonality of N inputs; v) the form and seasonality of N outputs. For those more familiar with the ‘young’ ecosystems of Europe and North America, it is worth highlighting that floras co-evolve with soils and climate. “Ecosystems are not special creations; they evolve through time” (87). While the floras of much of Africa, Australia and South America are ‘different’, they are well-adapted to their environment. That adaptation includes phosphorus and nitrogen cycling, the features of which

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are commensurate with climates, soils and landscapes. Some N cycle characteristics are very similar on a global scale despite major geographical and system differences. We have insufficient evidence to make generalizations about how these low-P ecosystems, dominated by ‘P-sensitive’ genera and species, may respond to increased inputs of N (or P). Nevertheless the available evidence concurs with Köerner’s cautionary note (88) about the poor use and understanding of the term ‘limitation’. Nitrogen (or phosphorus) additions to these plant communities can result in wholesale changes in species composition—rather than relieving any ‘limitation’ of extant species. It will be increasingly important to recognize the role of history in understanding responses to added N and we note that recent studies have begun to address this (89, 90), albeit at stillshort time scales. Changes in species composition, as well as fire and other disturbance regimes (e.g. widespread planting of conifers on landscapes formerly dominated by Fagus and Quercus spp. in Europe, rapidly changing flora of the Hawaiian islands, elimination of fire from areas of high population density), will help determine the effects of excessive additions of N on the function of terrestrial ecosystems. Seasonality of climate (e.g. monsoonal climates, Mediterranean climates, the dry tropics) strongly regulates many features of the N cycle, especially soil N transformations. This seasonality will continue to provide a background of variability against which it will be rather difficult to assess the influence of N inputs on N cycling processes. We concur with many of the recommendations for future research proposed by Matson et al. (1), including that experimental studies of N enrichment are needed for tropical forests. However, we suggest the proposition that it is “too late to know how many temperate forests functioned in the absence of anthropogenic N”, needs careful examination. Many southern temperate forests remain little affected by anthropogenic N and are valuable experimental sites for process-based studies. Although N cycling may be measured at a range of scales, policies are implemented and enforced at the national or provincial/state levels. Increased nitrogen fertilization usage is expected in the developing world, especially in Africa, in the next 50 years. Multinational efforts to control N loss to the environments are surely needed, but these efforts will require commitments from individual countries and the policy-makers within those countries. References and Notes 1. Matson, P.A., McDowell, W.H., Townsend, A.R. and Vitousek, P.M. 1999. The globalization of N deposition: ecosystem consequences in tropical environments. Biogeochemistry 46, 67-83. 2. Galloway, J.N. 2000. Nitrogen mobilization in Asia. Nutrient Cycling in Agroecosyst. 57, 1-12. 3. Galloway, J.N., Aber J.D., Erisman, J.W., Seitzinger, S.P., Howarth, R.W., Cowling, E.B. and Cosby, B.J. 2003. The nitrogen cascade. BioScience 53, 341-356. 4. Scholes, R.J. and Walker, B.H. 1993. An African Savanna: Synthesis of the Nylsvley Study. Cambridge University Press, Cambridge, UK 5. Bennett, L.T. and Adams, M.A. 1999. Indices for characterizing spatial variability of soil nitrogen in semi-arid grasslands of northwestern Australia. Soil Biol. Biochem. 31, 735-746 6. Binkley, D. and Högberg, P. 1997. Does atmospheric deposition of nitrogen threaten Swedish forests? For. Ecol. Manage. 92, 119-152. 7. Cadisch, G., Giller, K.E., Urquiaga, S., Miranda, C.H.B., Boddey, R.M. and Schunke, R. 1994. Does phosphorus supply enhance soil-N mineralization in Brazilian pastures? Eur. J. Agron. 3, 339-345. 8. Attiwill, P.M., Polglase, P.J., Weston, C.J. and Adams, M.A. 1996. Nutrient cycling in forests of south-eastern Australia. In: Nutrition of Eucalypts. Attiwill, P.M. and Adams, M.A. (eds) CSIRO Publishing, Collingwood, Victoria, pp. 191-227. 9. Aber, J.D., Goodale, C.L., Ollinger, S.V., Smith, M-L., Magill, A.H., Martin, M.E., Hallett, R.A. and Stoddard, J.L. 2003. Is nitrogen deposition altering the nitrogen status of northeastern forests? BioScience 53, 375-389. 10. Kriebitzsch, W.U. 1978. Stickstoffnachlieferung in saure Waldböden Nordwestdeutschlands. Scr. Geobotanika 14, 1-16. 11. Erickson, H., Davidson E.A. and Keller, M. 2002. Former land use and tree species affect nitrogen oxide emissions from a tropical dry forest. Oecologia 130, 297-308. 12. Gundersen, P., Callesen, I. and de Vries, W. 1998. Nitrate leaching in forest ecosystems is related to forest floor C/N ratios. Environ. Poll. 102, 403-407. 13. Cadisch, G., Sylvester-Bradley, R., Boller, B.C. and Nosberger, J. 1993. Effects of phosphorus and potassium on N2 fixation (15N-dilution) of field-grown Centrosema acutifolium and C. macrocarpum. Field Crops Res. 31, 329-340. 14. Stock, W.D. and Allsopp, N. 1992. Functional perspective of ecosystems. In: The Ecology of Fynbos: Nutrients, Fire and Diversity. Cowling, R.M. (ed.). Oxford University

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We are pleased to acknowledge the financial support of the Natural Environment Research Council (NERC), UK, through their Global Nitrogen Enrichment (GANE) Thematic Programme, which funded the international workshop on responses of terrestrial ecosystems to nitrogen enhancement that led to this paper. We thank the Stockholm Environment Institute for their hospitality during the workshop and the unstinting support during the preparation of this manuscript. We would also like to thank Professor Mike Chadwick and Dr. Johan Kuylenstierna for many useful inputs and their support for the project. 92. First submitted 16 Febr. 2004. Revised manuscript received 1 Sept. 2004. Accepted for publication 2 Sept. 2004.

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Professor Mark Adams is presently with the Centre of Excellence in Natural Resource Management, at the University of Western Australia. His interests lie with the biogeochemistry of natural ecosystems and the ecophysiology of Australian native trees. His address: CENRM, University of Western Australia, 35 Stirling Highway, Crawley, WA 6009, Australia. [email protected] Professor Phil Ineson is the Chair in Global Change Ecology at the University of York, a joint post between the Department of Biology and the Stockholm Environment Institute at York. His research is largely concerned with the interactions between soils and global change, with a particular interest in trace gas fluxes and stable isotopes. His address: Stockholm Environment Institute-York, Sally Baldwin Building - 'D' Block, University of York, Heslington, York YO10 5DD, UK. [email protected] Dan Binkley is a professor in forest ecology at Colorado State University, where he also serves as the director of the Graduate Degree Program in Ecology. His research focuses on ecosystem dynamics, including forest productivity and plant/soil interactions. His address: Dept of Forest Sciences, Graduate Degree Program in Ecology and Natural Resource Ecology Laboratory, Colorado State University, Ft. Collins, CO 80523, USA. [email protected] Dr George Cadisch has been a Reader at Imperial College London but is presently a professor at the Institute of Plant Production and Agroecology in the Tropics and Subtropics at the University of Hohenheim in Stuttgart, Germany. His research focuses on nutrient cycling in agricultural tropical ecosystems in particular the impact of plant quality attributes and N2 fixation on N cycling and C sequestration. His address: Institute of Plant Production and Agroecology in the Tropics and Subtropics, University of Hohenheim, 70599 Stuttgart, Germany. [email protected] Mary Scholes is a professor in the School of Animal, Plant and Environmental Sciences at the University of the Witwatersrand, Johannesburg. She conducts research in nutrient cycling in savanna and plantations with specific emphasis on carbon, nitrogen and phosphorus. Her position is endowed by Sappi. Her address: University of the Witwatersrand, School of Animal, Plant and Environmental Sciences, P Bag 3, WITS, 2050, South Africa. [email protected] Dr. Kevin Hicks is a research associate at the Stockholm Environment Institute, based in the University of York. His research concentrates on air pollution impacts to terrestrial ecosystems and the links between scientific information and policy making. His address: Stockholm Environment Institute-York, Sally Baldwin Building - 'D' Block, University of York, Heslington, York YO10 5DD, UK. [email protected] Dr. Naoko Tokuchi is an associate professor of Field Science Education and Research Center, Kyoto University. She is interested in biogeochemistry and forest ecosystem dynamics. Her address: Field Science Education and Research Center, Kyoto University, 606-8502, Kyoto, Japan. [email protected]

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