Understanding the fate of anthropogenic N additions in soils, and especially the factors that regulate its retention and export, has considerable implications for the functioning of recipient terrestrial ecosystems and the environment beyond. In particular, changes in soil N retention that accompany increasing N deposition determine the export of N to adjacent aquatic ecosystems, where it can have substantial, harmful effects on water quality owing to eutrophication and acidification. Also, the loss of gaseous N to the atmosphere has implications for air quality, since it acts as one of the more potent greenhouse gases alongside CO2. In general, little is known about the capacity of soils to act as N sinks, or the factors that regulate their ability to do this. However, since most ecosystems are N-limited, it is generally thought that plants will act as sinks for large amounts of anthro-pogenic N and that this may slow down, or even halt, these deleterious effects of N deposition (Aber et al. 1989). It also appears as if some soils have a very high capacity to accumulate and retain N in stable organic matter pools, thereby reducing N export and harmful effects of pollution.
This has been shown to be the case in semi-natural grasslands in Britain, for example, where significant amounts of N added over several years have accumulated in soil organic matter, thereby reducing the effects of N pollution on these systems and also the export of N to the environment beyond (Phoenix et al. 2004).
It has been proposed that one of the most likely routes of N storage in soils is the rapid immobilization of N inputs by soil microbes and its subsequent transfer to plants and/or stable, non-microbial organic matter pools (Zogg et al. 2000). A number of studies support this notion. For example, a field study of alpine tundra showed that soil microbes act as a large, albeit short-term, sink for simulated anthropogenic N inputs, especially towards the end of the growing season (Fisk and Schmidt 1996). Similarly, 15N labelling studies in temperate hardwood forests show that soil microbes act as a large initial sink for inputs of15N–NO3, and that subsequent micro-bial turnover and cell death leads to the transfer of this micromicro-bial N to plants
and/or stable organic matter pools where it is retained (Zogg et al. 2000) (Fig. 6.11). Immobilization into the microbial biomass also represents a significant and immediate sink for added N in semi-natural temperate grasslands (Jackson et al. 1989; Bardgett et al. 2003) (Fig. 6.12), and its sub-sequent transfer to non-microbial, recalcitrant organic matter pools repres-ents an important long-term N sink in these ecosystems (Phoenix et al.
2004). These studies all point to the importance of microbial immobiliza-tion for N retenimmobiliza-tion in polluted ecosystems and suggest that there is need for
Time since addition of 15NO3– (weeks) 16 8
0 100
Inorganic N Forest floor
Microbial biomass Organic matter
Root biomass
Percentage recovery of added 15N
Fig. 6.11 Schematic diagram of total recovery of added 15NO3 in inorganic-N, microbial, soil organic matter, and root biomass pools, throughout a four-month growing season in a northern hardwood forest. (Adapted from Zogg et al. 2000.)
Grassland type
Improved Unimproved
0 20 40 60 80
Shoot Root Soil microbes
Proportion of added 15N taken up into different pools
Fig. 6.12 Uptake of 15N derived from 15N–NH4 by different pools, 50 hours after labelling two grasslands: a productive, fertilized grassland (improved) and an unproductive, unfertilized grassland (unimproved.) Data are expressed as % added. (Data from Bardgett et al. 2003.)
further studies to determine the mechanisms that regulate the capacity of microbes to sequester N.
Not all ecosystems accumulate N when subject to N deposition; rather, some reach N saturation very rapidly resulting in significant amounts of N export in drainage waters. This has been shown to be the case in hardwood forests of northern United States, for example. Here, 8 years of experi-mental NO3additions (at 2.5 times ambient N deposition) to a number of sites dramatically increased leaching losses of NO3and DON, indicat-ing rapid N saturation of this ecosystem (Pregitzer et al. 2004). This is in contrast to the fore-mentioned studies of UK semi-natural grasslands where significant amounts of added N were retained in soil organic matter, thereby reducing N losses (Phoenix et al. 2004). It is not entirely clear why some ecosystems are more susceptible to N saturation than others, but one suggestion is that it depends on initial levels of N availability (Aber et al.
1998). For example, Gundersen et al. (1998) showed that forests with the highest risk of N saturation and NO3leaching had surface organic hori-zons with C : N ratios of25, whereas a survey of 181 forests across Europe by MacDonald et al. (2002) showed that forests with organic horizons of C : N 25 and soil pH 4.6, and atmospheric N inputs 3 g N m2yr1, were most susceptible to N saturation. Similarly, it is well established that fertile grasslands that receive regular additions of fertilizer N are especially prone to N saturation and N export, which contrasts with low fertility grasslands where microbes sequester a large proportion of added N, which is then transferred to live plant and stable organic matter pools (Bardgett et al. 2003) (Fig. 6.12). In other words, inherently N-rich soils receiving high atmospheric N deposition are very susceptible to N saturation and large losses of N via leaching (Pregitzer et al. 2004).
The ability of soils to retain N is likely to be strongly affected by changes in plant species composition that accompany increased N deposition, which potentially lead to significant feedbacks to the decomposer community and soil N cycling. A good example of how N-driven shifts in plant community structure influence N cycling comes from moist meadow communities of alpine tundra of the Rocky Mountains, USA. These communities are espe-cially susceptible to N deposition, since they receive relatively large inputs of N from melting snow, an important reservoir of wintertime N deposi-tion (Bowman 1992). Furthermore, these plant communities are cohabited by both the N-limited grass Deschampsia caespitosa and the N-insensitive forb Acomastylis rossii, which exert markedly different influences on soil N cycling: typically, rates of N mineralization in soils under Deschampsia are 10-fold greater than beneath Acomastylis (Steltzer and Bowman 1998).
Therefore, the replacement of Acomastylis as a dominant in moist meadow by Deschampsia under elevated N deposition, leads to a strong positive feedback, increasing rates of N mineralization and leaching of N from soil, leading to greater N export from alpine catchments (Welker et al. 2001).
This kind of positive feedback is probably mirrored in many situations where N deposition promotes the dominance of fast-growing plant species that produce N-rich plant litter, which is easily decomposed by microbes in soil, leading to rapid N mineralization and potential for N loss (Fig. 6.13).
Furthermore, the potential for N export is likely to be further promoted by associated shifts in soil food-web structure towards bacterial dominance. As discussed in Chapter 4, the encouragement of bacterial-dominated food webs by nutrient enrichment promotes fast and ‘leaky’ nutrient cycles with a high potential for nutrient export to adjacent ecosystems. As a result of these feedbacks to increasing N, it has been proposed that responses to N deposition will be non-linear (Fig. 6.14) since the export of N from soils will be accelerated as a result of species replacement, owing to enhanced N cycling in soil (Welker et al. 2001).
Before leaving the subject of N deposition, it is important to highlight one other area of research that is receiving increasing attention, that is, the question of how atmospheric N deposition influences the production of DON and its loss from polluted soils. Until recently, fluxes of DON were largely ignored, with emphasis being placed on the cycling and loss of inorganic N. However, a number of studies reveal that several ecosystems facing N deposition, including grassland and forest, are subject to high amounts of DON loss in leachates relative to inorganic N (McDowell et al. 1998; Phoenix et al. 2004; Pregitizer et al. 2004). A good example of this is in hardwood forests of North America, where experimental N additions increased DON export six-fold (Pregitizer et al. 2004). Similarly, Phoenix et al. (2004) were surprised by the large amount of DON exported in leachate from semi-natural grasslands subject to N deposition. This
Self
Low rates of N cycling High retention of soil N
Fast growing, competitive species.
High rates of N cycling High rates of N leaching from soils.
Disturbance, exhaustion of soil
Fig. 6.13 Conceptual model showing the generalized pattern of plant species self-replacement in communities with low and high N availability. The model shows that in the face of greater N deposition, N conserving species of infertile environments will be replaced by more competitive species, with traits that promote more rapid N cycling in soil. Under such conditions, the new state would be preserved, even in the absence of additional N inputs, until N reserves in soil are exhausted when species adapted to low N conditions will re-establish. (Redrawn with permission from Oxford University Press; Welker et al. 2001.)
high loss of DON from polluted soils has been attributed to effects of N deposition on the production of organic substrates and their processing by soil microbes, and to incomplete degradation of lignin as a result of inhibition of lignolytic enzymes, leading to increased soil concentrations of phenolics and greater production of DON in polluted soils (Waldrop et al. 2004). Further studies are clearly needed in this area to determine the mechanisms controlling the production and loss of DON from polluted soils and the significance of this form of N loss for adjacent ecosystems.