Human impacts on the nitrogen cycle

Agricultural and industrial nitrogen (N) inputs to the environment currently exceed inputs from natural N fixation (Galloway 2003). As a consequence of anthropogenic inputs, the global nitrogen cycle (Fig. 1) has been significantly altered over the past century. Global atmospheric nitrous oxide (N₂O) concentrations have increased from a pre-industrial value of ~270 ppb to ~319 ppb in 2005 (Alley et al. 2007). Human activities account for over one-third of N₂O emissions, most of which are due to the agricultural sector (Alley et al. 2007). This article is intended to give a brief review of the history of anthropogenic N inputs, and reported impacts of nitrogen inputs on selected terrestrial and aquatic ecosystems. This review is not meant to be exhaustive, but to present some of the more salient impacts of human activities on the N cycle.

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History of anthropogenic nitrogen inputs

Approximately 78% of earth’s atmosphere is N gas (N₂), which is an inert compound and biologically unavailable to most organisms. In order to be utilized in most biological processes, N₂ must be converted to reactive N (Nr), which includes inorganic reduced forms (NH₃ and NH₄⁺), inorganic oxidized forms (NO, NO₂, HNO₃, N₂O, and NO₃^-), and organic compounds (urea, amines, and proteins) (Galloway 2003). N₂ has a strong triple bond, and so a significant amount of energy (226 kcal mol-1) is required to convert N₂ to Nr (Galloway 2003). Prior to industrial processes, the only sources of such energy were solar radiation and electrical discharges (Galloway 2003). Utilizing a large amount of metabolic energy and the enzyme nitrogenase, some bacteria and blue-green algae convert atmospheric N₂ to NH₃, a process known as biological nitrogen fixation (BNF, Schlesinger 1997). The anthropogenic analogue to BNF is the Haber-Bosch process, in which fossil fuel H₂ is reacted with atmospheric N₂ under high temperature and pressure to produce NH₃ (Smil 2001). Lastly, N₂ is converted to NO by energy from lightening, which is negligible in current temperate ecosystems, or by fossil fuel combustion (Galloway 2003).

Until 1850, natural BNF, cultivation-induced BNF (e.g., planting of leguminous crops), and incorporated organic matter were the only sources of N for agricultural production (Smil 2001). Near the turn of the century, Nr from guano and sodium nitrate deposits was harvested and exported from the arid Pacific islands and South American deserts (Smil 2001). By the late 1920s, early industrial processes, albeit inefficient, were commonly used to produce NH3 (Galloway 2003). Due to the efforts of Fritz Haber and Carl Bosch, the Haber-Bosch process became the largest source of nitrogenous fertilizer after the 1950s, and replaced BNF as the dominant source of NH3 production (Smil 2001). From 1890 to 1990, anthropogenically created Nr increased almost ninefold (Galloway 2003). During this time, global population more than tripled, partly due to increased food production.

Since the industrial revolution, an additional source of anthropogenic N input has been fossil fuel combustion, which is used to generate energy (e.g., to power automobiles). During combustion of fossil fuels, high temperatures and pressures provide energy to produce NO from N₂ oxidation (Galloway 2003). Additionally, when fossil fuel is extracted and burned, fossil N may become reactive (i.e., NO_x emissions, Galloway 2003). During the 1970s, scientists began to recognize that N inputs were accumulating in the environment and affecting ecosystem functioning (Galloway 2003).

Impacts of anthropogenic inputs on the nitrogen cycle

Between 1890 and 1990, global reactive nitrogen (Nr) creation had increased nearly 50% (Galloway and Cowling 2002). During this period, atmospheric emissions of Nr species reportedly increased 250% and deposition to marine and terrestrial ecosystems increased over 200% (Galloway and Cowling 2002). Additionally, there was a reported fourfold increase in riverine dissolved inorganic N fluxes to coasts (Galloway and Cowling 2002). N is a critical limiting nutrient in many systems, including forests, wetlands, and coastal and marine ecosystems (Vitousek and Howarth 1991, Rabalais 2002); therefore, this change in emissions and distribution of Nr has resulted in substantial consequences for aquatic and terrestrial ecosystems.

Atmosphere

Atmospheric Nr inputs mainly include oxides of N (NO_x), ammonia (NH₃), and nitrous oxide (N₂O) from aquatic and terrestrial ecosystems (Schlesinger 1997), and NO_x from fossil fuel and biomass combustion (Galloway et al. 2003).

In agroecosystems, fertilizer application has increased microbial nitrification (aerobic process in which microorganisms oxidize ammonium [NH₄+] to nitrate [NO₃-]) and denitrification (anaerobic process in which microorganisms reduce NO₃^- to atmospheric nitrogen gas [N2]). Both processes naturally leak nitric oxide (NO) and nitrous oxide (N₂O) to the atmosphere (Schlesinger 1997). Of particular concern is N₂O, which has an average atmospheric lifetime of 114-120 years (Houghton et al. 2001), and is 300 times more effective than CO₂ as a greenhouse gas (Schlesinger 1997) . NO_x produced by industrial processes, automobiles and agricultural fertilization and NH₃ emitted from soils (i.e., as an additional byproduct of nitrification, Schlesinger 1997) and livestock operations are transported to downwind ecosystems, influencing N cycling and nutrient losses. Galloway et al. (2003) cited six major effects of NO_x and NH₃ emissions: 1) decreased atmospheric visibility due to ammonium aerosols (fine particulate matter [PM]); 2) elevated ozone concentrations; 3) ozone and PM impacts on human health (e.g., respiratory diseases, cancer); 4) increases in relative forcing and global climate change; 5) decreased agricultural productivity due to ozone deposition; and 6) ecosystem acidification (e.g., Houdijk et al. 1993) and eutrophication.

Biosphere

Terrestrial and aquatic ecosystems receive Nr inputs from the atmosphere through wet and dry deposition (Galloway 2003). Atmospheric Nr species can be deposited to ecosystems in precipitation (e.g., NO₃^-, NH₄⁺, organic N compounds), as gases (e.g., NH₃ and gaseous nitric acid [HNO₃]), or as aerosols (e.g., ammonium nitrate [NH₄NO₃]; Galloway 2003). Aquatic ecosystems receive additional nitrogen from surface runoff and riverine inputs (Rabalais 2002).

Increased N deposition can acidify soils, streams, and lakes and alter forest and grassland productivity. In forest and grassland ecosystems, Nr inputs have produced initial increases in productivity followed by declines as critical thresholds are exceeded (Galloway 2003). Nr effects on biodiversity, carbon cycling, and changes in species composition have also been demonstrated. In highly developed areas of near shore coastal ocean and estuarine systems, rivers deliver direct (e.g., surface runoff) and indirect (e.g., groundwater contamination) N inputs from agroecosystems (Rabalais 2002). Increased N inputs can result in freshwater acidification and eutrophication of marine waters.

Terrestrial ecosystems

Impacts on productivity and nutrient cycling

Much of terrestrial growth in temperate systems is limited by N; therefore, N inputs (i.e., through deposition and fertilization) can increase N availability, which temporarily increases N uptake, plant and microbial growth, and N accumulation in plant biomass and soil organic matter (Aber et al. 1989). Incorporation of greater amounts of N in organic matter decreases C:N ratios, increasing mineral N release (NH₄⁺) during organic matter decomposition by heterotrophic microbes (i.e., ammonification, Aber 1992). As ammonification increases, so does nitrification of the mineralized N. Because microbial nitrification and denitrification are “leaky”, N deposition is expected to increase trace gas emissions (Matson et al. 2002). Additionally, with increasing NH₄⁺ accumulation in the soil, nitrification processes release hydrogen ions, which acidify the soil. NO₃^-, the product of nitrification, is highly mobile and can be leached from the soil, along with positively charged alkaline minerals such as calcium and magnesium (Schlesinger 1997). In acid soils, mobilized aluminum ions can reach toxic concentrations, negatively affecting both terrestrial and adjacent aquatic ecosystems.

Anthropogenic sources of N generally reach upland forests through deposition (Aber et al. 2003). A potential concern of increased N deposition due to human activities is altered nutrient cycling in forest ecosystems. Numerous studies have demonstrated both positive and negative impacts of atmospheric N deposition on forest productivity and carbon storage. Added N is often rapidly immobilized by microbes (Nadelhoffer et al. 1999), and the effect of the remaining available N depends on the plant community’s capacity for N uptake (Bauer et al. 2004). In systems with high uptake, N is assimilated into the plant biomass, leading to enhanced net primary productivity (NPP) and possibly increased carbon sequestration through greater photosynthetic capacity. However, ecosystem responses to N additions are contingent upon many site-specific factors including climate, land-use history, and amount of N additions. For example, in the Northeastern United States, hardwood stands receiving chronic N inputs have demonstrated greater capacity to retain N and increase annual net primary productivity (ANPP) than conifer stands (Magill et al. 2004). Once N input exceeds system demand, N may be lost via leaching and gas fluxes. When available N exceeds the ecosystem’s (i.e., vegetation, soil, and microbes, etc.) uptake capacity, N saturation occurs and excess N is lost to surface waters, groundwater, and the atmosphere (Aber et al. 1989, Bauer et al. 2004, Magill et al. 2004). N saturation can result in nutrient imbalances (e.g., loss of calcium due to nitrate leaching) and possible forest decline (Aber 1992).

A recently published, 15-year study of chronic N additions at the Harvard Forest Long Term Ecological Research (LTER) program has elucidated many impacts of increased nitrogen deposition on nutrient cycling in temperate forests. Magill et al. (2004) found that chronic N additions resulted in greater leaching losses, increased pine mortality, and cessation of biomass accumulation. Additionally, Bauer et al. (2004) reported that chronic N additions resulted in accumulation of non-photosynthetic N and subsequently reduced photosynthetic capacity, supposedly leading to severe carbon stress and mortality. These findings negate previous hypotheses that increased N inputs would increase NPP and carbon sequestration.

Impacts on plant species diversity.

Many plant communities have evolved under low nutrient conditions; therefore, increased N inputs can alter biotic and abiotic interactions, leading to changes in community composition. Several nutrient addition studies have shown that increased N inputs lead to dominance of fast-growing plant species, with associated declines in species richness (Huenneke et al. 1990, Tilman 1997, Wilson and Tilman 2002). Other studies have found that secondary responses of the system to N enrichment, including soil acidification and changes in mycorrhizal communities have allowed stress-tolerant species to out-compete sensitive species (Houdijk et al. 1993, Egerton-Warburton and Allen 2000). For example, both Aerts and Berendse (1988) and Bobbink et al. (1992) found evidence that increased N availability has resulted in declines in species-diverse heathlands. Heathlands are characterized by N-poor soils, which exclude N-demanding grasses; however, with increasing N deposition and soil acidification, invading grasslands replace lowland heath.

In a more recent experimental study of N fertilization and disturbance (i.e., tillage) in old field succession, Wilson and Tilman (2002) found that species richness decreased with increasing N, regardless of disturbance level. Competition experiments showed that competitive dominants excluded competitively inferior species between disturbance events. With increased N inputs, competition shifted from belowground to aboveground (i.e., to competition for light), and patch colonization rates significantly decreased. These internal changes can dramatically affect the community by shifting the balance of competition-colonization tradeoffs between species. In patch-based systems, regional coexistence can occur through tradeoffs in competitive and colonizing abilities given sufficiently high disturbance rates (Hastings 1980). That is, with inverse ranking of competitive and colonizing abilities, plants can coexist in space and time as disturbance removes superior competitors from patches, allowing for establishment of superior colonizers. However, as demonstrated by Wilson and Tilman (2002), increased nutrient inputs can negate tradeoffs, resulting in competitive exclusion of these superior colonizers/poor competitors.

Aquatic ecosystems

As reviewed by Rabalais (2002), aquatic ecosystems also exhibit varied responses to nitrogen enrichment. NO₃^- loading from N saturated, terrestrial ecosystems can lead to acidification of downstream freshwater systems and eutrophication of downstream marine systems (Rabalais 2002). Freshwater acidification can cause aluminum toxicity and mortality of pH-sensitive fish species (Rabalais 2002). Because marine systems are generally nitrogen-limited, excessive N inputs can result in water quality degradation due to toxic algal blooms, oxygen deficiency, habitat loss, decreases in biodiversity, and fishery losses (Rabalais 2002).

Acidification of freshwaters

Atmospheric N deposition in terrestrial landscapes can be transformed through soil microbial processes to biologically available nitrogen, which can result in surface-water acidification, and loss of biodiversity. NO₃^- and NH₄⁺ inputs from terrestrial systems and the atmosphere can acidify freshwater systems when there is little buffering capacity due to soil acidification (see above discussion, Rabalais 2002). As described by Driscoll (2001), N pollution in Europe, the Northeastern United States, and Asia is a current concern for freshwater acidification. Lake acidification studies in the Experimental Lake Area (ELA) in northwestern Ontario clearly demonstrated the negative effects of increased acidity on a native fish species: lake trout (Salvelinus namaycush) recruitment and growth dramatically decreased due to extirpation of its key prey species during acidification (Mills et al. 2000).

Eutrophication of marine systems

Urbanization, deforestation, and agricultural activities largely contribute sediment and nutrient inputs to coastal waters via rivers (Rabalais 2002). Increased nutrient inputs to marine systems have shown both short-term increases in productivity and fishery yields, and long-term detrimental effects of eutrophication. Tripling of NO₃^- loads in the Mississippi River in the last half of the 20th century have been correlated with increased fishery yields in waters surrounding the Mississippi delta (Grimes 2001); however, these nutrient inputs have produced seasonal hypoxia (oxygen concentrations less than 2-3 mg L^-1, "dead zones") in the Gulf of Mexico (Rabalais 2002, Galloway 2003). In estuarine and coastal systems, high nutrient inputs increase primary production (e.g., phytoplankton, sea grasses, macroalgae), which increase turbidity with resulting decreases in light penetration throughout the water column. Consequently, submerged vegetation growth declines, which reduces habitat complexity and oxygen production. The increased primary (i.e., phytoplankton, macroalgae, etc.) production leads to a flux of carbon to bottom waters when decaying organic matter (i.e., senescent primary production) sinks and is consumed by aerobic bacteria lower in the water column. As a result, oxygen consumption in bottom waters is greater than diffusion of oxygen from surface waters .

Integration

The above system responses to reactive nitrogen (Nr) inputs are almost all exclusively studied separately; however, as discussed by Galloway et al. (2003), research studies are increasingly indicating that nitrogen loading problems are linked by multiple pathways transporting nutrients across system boundaries. This sequential transfer between ecosystems is termed the nitrogen cascade (Galloway and Cowling 2002; see illustration from United Nations Environment Programme [1]). During the cascade, some systems accumulate Nr, which results in a time lag in the cascade and enhanced effects of Nr on the environment in which it accumulates. Ultimately, anthropogenic inputs of Nr are either accumulated or denitrified; however, little progress has been made in determining the relative importance of Nr accumulation and denitrification, which has been mainly due to a lack of integration among scientific disciplines (Galloway 2003, Davidson and Seitzinger 2006).

Most Nr applied to global agroecosystems cascades through the atmosphere and aquatic and terrestrial ecosystems until it is converted to N₂, primarily through denitrification (Galloway 2003). Although terrestrial denitrification produces gaseous intermediates (nitric oxide [NO] and nitrous oxide [N₂O]), the last step—microbial production of N₂^—is critical because atmospheric N₂ is a sink for Nr (Davidson and Seitzinger 2006). Many studies have clearly demonstrated that managed buffer strips and wetlands can remove significant amounts of nitrate (NO₃^-) from agricultural systems through denitrification (e.g., Tate et al. 2000, Jackson et al. 2006). Such management may help attenuate the undesirable cascading effects and eliminate environmental Nr accumulation (Galloway et al . 2003).

Human activities dominate the global and most regional N cycles (Galloway et al. 2004). N inputs have shown negative consequences for both nutrient cycling and native species diversity in terrestrial and aquatic systems. In fact, due to long-term impacts on food webs, Nr inputs are widely considered the most critical pollution problem in marine systems (Rabalais 2002). In both terrestrial and aquatic ecosystems, responses to N enrichment vary; however, a general re-occurring theme is the importance of thresholds (e.g., nitrogen saturation) in system nutrient retention capacity. In order to control the N cascade, there must be integration of scientific disciplines and further work on Nr storage and denitrification rates (Davidson and Seitzinger 2006).

Bibliography

• Aber, J. D. 1992. Nitrogen cycling and nitrogen saturation in temperate forest ecosystems. Trends in Ecology & Evolution 7:220-224.

• Aber, J. D., C. L. Goodale, S. V. Ollinger, M. L. Smith, A. H. Magill, M. E. Martin, R. A. Hallett, and J. L. Stoddard. 2003. Is nitrogen deposition altering the nitrogen status of northeastern forests? Bioscience 53:375-389.

• Aber, J. D., K. J. Nadelhoffer, P. Steudler, and J. M. Melillo. 1989. Nitrogen saturation in northern forest ecosystems. Bioscience 39:378–386.

• Aerts, R., and F. Berendse. 1988. The Effect of Increased Nutrient Availability on Vegetation Dynamics in Wet Heathlands. Vegetatio 76:63-69.

• Alley et al. 2007. IPCC Climate Change 2007: The Physical Science Basis. Contribution of Working Group I in the Third Assessment Report of Intergovernmental Panel on Climate Change. Report Summary for Policy Makers (SPM).

• Bauer, G. A., F. A. Bazzaz, R. Minocha, S. Long, A. Magill, J. Aber, and G. M. Berntson. 2004. Effects of chronic N additions on tissue chemistry, photosynthetic capacity, and carbon sequestration potential of a red pine (Pinus resinosa Ait.) stand in the NE United States. Forest Ecology and Management 196:173-186.

• Bobbink, R., G. W. Heil, and M. Raessen. 1992. Atmospheric Deposition and Canopy Exchange Processes in Heathland Ecosystems. Environmental Pollution 75:29-37.

• Davidson, E. A., and S. Seitzinger. 2006. The enigma of progress in denitrification research. Ecological Applications 16:2057-2063.

• Driscoll, C. T., G. B. Lawrence, A. J. Bulger, T. J. Butler, C. S. Cronan, C. Eagar, K. F. Lambert, G. E. Likens, J. L. Stoddard, and K. C. Weathers. 2001. Acidic Deposition in the Northeastern United States: Sources and Inputs, Ecosystem Effects, and Management Strategies. Pages 180-198.

• Egerton-Warburton, L. M., and E. B. Allen. 2000. Shifts in arbuscular mycorrhizal communities along an anthropogenic nitrogen deposition gradient. Ecological Applications 10:484-496.

• Galloway, J. N. 2003. The Global Nitrogen Cycle. Pages 557-583 in H. D. Holland and K. K. Turekian, editors. Treatise on Geochemistry. Pergamon, Oxford.

• Galloway, J. N., J. D. Aber, J. W. Erisman, S. P. Seitzinger, R. W. Howarth, E. B. Cowling, and B. J. Cosby. 2003. The nitrogen cascade. Bioscience 53:341-356.

• Galloway, J. N., and E. B. Cowling. 2002. Reactive nitrogen and the world: 200 years of change. Ambio 31:64-71.

• Galloway, J. N., F. J. Dentener, D. G. Capone, E. W. Boyer, R. W. Howarth, S. P. Seitzinger, G. P. Asner, C. C. Cleveland, P. A. Green, E. A. Holland, D. M. Karl, A. F. Michaels, J. H. Porter, A. R. Townsend, and C. J. Vorosmarty. 2004. Nitrogen cycles: past, present, and future. Biogeochemistry 70:153-226.

• Grimes, C. B. 2001. Fishery production and the Mississippi River discharge. Fisheries 26:17-26.

• Hastings, A. 1980. Disturbace, coexistence, history, and competition for space. Theoretical Population Biology 18:363-373.

• Houdijk, A. L. F. M., P. J. M. Verbeek, H. F. G. V. Dijk, and J. G. M. Roelofs. 1993. Distribution and decline of endangered herbaceous heathland species in relation to the chemical composition of the soil. Plant and Soil 148:137-143.

• Houghton, J. T., Y. Ding, D. J. Griggs, M. Noguer, P. J. van der Linden, X. Dai, K. Maskell, and C. A. Johnson. 2001. IPCC Climate Change 2001: The Scientific Basis. Contribution of Working Group I in the Third Assessment Report of Intergovernmental Panel on Climate Change Cambridge University Press.

• Huenneke, L. F., S. P. Hamburg, R. Koide, H. A. Mooney, and P. M. Vitousek. 1990. Effects of soil resources on plant invasion and community structure in Californian serpentine grassland. Ecology 71:478-491.

• Jackson, R. D., B. Allen-Diaz, L. G. Oates, and K. W. Tate. 2006. Spring-water nitrate increased with removal of livestock grazing in a California oak savanna. Ecosystems 9:254-267.

• Magill, A. H., J. D. Aber, W. S. Currie, K. J. Nadelhoffer, M. E. Martin, W. H. McDowell, J. M. Melillo, and P. Steudler. 2004. Ecosystem response to 15 years of chronic nitrogen additions at the Harvard Forest LTER, Massachusetts, USA. Forest Ecology and Management 196:7-28.

• Mills, K. H., S. M. Chalanchuk, and D. J. Allan. 2000. Recovery of fish populations in Lake 223 from experimental acidification. Canadian Journal of Fisheries and Aquatic Sciences 57:192-204.

• Nadelhoffer, K. J., M. R. Downs, and B. Fry. 1999. Sinks for N-15-enriched additions to an oak forest and a red pine plantation. Ecological Applications 9:72-86.

• Rabalais, N. N. 2002. Nitrogen in aquatic ecosystems. Ambio 31:102-112.

• Schlesinger, W. H. 1997. Biogeochemistry : An analysis of global change, San Diego, CA.

• Smil, V. 2001. Enriching the earth : Fritz Haber, Carl Bosch, and the transformation of world food production. MIT Press, Cambridge, MA.

• Tate, K. W., G. A. Nader, D. J. Lewis, E. R. Atwill, and J. M. Connor. 2000. Evaluation of buffers to improve the quality of runoff from irrigated pastures. Journal of Soil and Water Conservation 55:473-478.

• Tilman, D. 1997. Community invasibility, recruitment limitation, and grassland biodiversity. Ecology 78:81-93.

• Venterink, H. O., M. J. Wassen, A. W. M. Verkroost, and P. C. de Ruiter. 2003. Species richness-productivity patterns differ between N-, P-, and K-limited wetlands. Ecology 84:2191-2199.

• Vitousek, P. M., and R. W. Howarth. 1991. Nitrogen Limitation on Land and in the Sea - How Can It Occur. Biogeochemistry 13:87-115.

• Wilson, S. D., and D. Tilman. 2002. Quadratic variation in old-field species richness along gradients of disturbance and nitrogen. Ecology 83:492-504.