Previous work has concluded that the science evaluating the connection between specific drivers and specific services is limited (Carpenter et al. 2009; Norgaard 2010). In this section, we connect the existing work on N-driven changes in ecosystem structure and function with ecosystem services. Overarching requirements that must be met in order for an effective accounting framework to be developed and implemented include:
In the following section, we focus on several (but by no means all) key ecosystem services that directly link to management of N in the environment: (1) food, fuel and fibre production, (2) climate regulation, (3) maintenance of human health and (4) maintenance of biodiversity and aesthetics. For each ecosystem service, we evaluate whether there is enough information to construct an appropriate ecological production function (the biophysical relationship between ecosystems and services; Daily et al. 2009). We also review attempts to examine benefits and costs of N increases on each service. Types of economic costs that N enrichment can incur include mitigation, damage, remediation and substitution costs (Moomaw & Birch 2005). Others have argued that monetary value should not be the only metric of ecosystem services within a defensible framework, in part because we do not yet have approaches to give monetary value to all relevant services (Toman 1998), and thus such a framework would be incomplete (Norgaard 2010). We maintain that economic valuation is useful because it is easily understandable by society and is a common unit that allows for simple stacking of services when comparing management options (Dodds et al. 2009; Birch et al. 2011). We identify available data that could support an ecosystem services approach to N management as well as critical knowledge gaps that could prevent making useful connections between ecosystem processes, ecosystem services, and valuation of such services. We assemble cost information where available as the metric cost per unit N, which is increasingly available from a number of recent studies (Kusiima & Powers 2010; Birch et al. 2011). If costs per unit N were not available, but we had total damage costs, we calculated this metric based on total damage costs (from Appendix S1) divided by N fluxes to the affected parts of the ecosystem. Finally we apply and compare these cost estimates within an example accounting framework to illustrate how it can inform decisions.
N and food, fuel and fibre production
One suite of ecosystem services that has been greatly enhanced by N addition to the environment is food, fuel and fibre production. Because ecosystems are often limited by N availability, N additions to soils and surface waters can markedly boost biological production in these systems. Within the past century intensive agricultural production has yielded tremendous increases in human nutrition and well-being, largely as a result of the invention and large scale implementation of the Haber-Bosch process for N fixation (Galloway et al. 2008). The development and accelerated use of nitrogen fertilizers has driven large increases in food production for both humans and animals in affluent nations, and has shifted the balance between malnutrition and an adequate diet for a huge number of people in developing nations (Smil 2002). Increases in N-based fertilizers and modern agricultural practices have more than doubled the number of people who were fed from a hectare of agricultural land managed with organic residues and N2-fixers in the early 1900s (Evans 1980; Smil 2002).
The broad benefits of N fertilization on food and material production are well known, particularly for agriculture, but the damages to these services caused by increasing N in the environment are not as well understood. Several studies have quantified damage costs of N on food and fibre production. In Table 2, we focus on valuations of damages or benefits associated with mitigation, since remediation and substitution costs are only now becoming available for many systems (e.g. Jenkins et al. 2010; Birch et al. 2011). One of the most complete national analyses of N effects examined the consequences of US air pollution control policies (Chestnut & Mills 2005). Emissions of N and S oxides led to acidification and damage to materials that cost c. $133 million annually prior to the US Acid Rain Program, 1990 Clean Air Act Amendments (Chestnut & Mills 2005). Nitrogen oxides also contribute to ozone formation in the troposphere, which can reduce crop and forest production in ways that could offset any fertilization effects, particularly in areas where N loading is already high. Ozone reductions projected to result from the 1990 Clean Air Act Amendments were estimated to provide a total annual benefit to the US commercial timber industry of about $800 million, and improved yields were estimated to benefit grain crop producers by $700 million in 2010 (Chestnut & Mills 2005). Increases in N also fuel UV damages to crop production, fisheries and corals, since N2O is currently the most important contributor to the breakdown of stratospheric ozone (Ravishankara et al. 2009). We discuss UV damages further in the section on human health.
In aquatic ecosystems, increasing N loads can stimulate production, particularly in estuaries and near-coastal waters, with mixed effects. At low N loading, fisheries may be limited by N, whereas increasing N loads can lead to eutrophication, hypoxia, and anoxia with the potential to reduce fish production (Fig. 2; Breitburg et al. 2009). Also, the desirability of enhanced production of any given species is somewhat variable: for example, greater algal production could ultimately lead to fish kills; atmospheric N loading could stimulate the production of undesirable or exotic species (e.g. Suding et al. 2004) leading to questions about how various increases in production should be valued. Despite these complexities, greater understanding of how to value the net benefits or detriments of N loading to the environment would contribute significantly to our understanding and ability to implement an ecosystem services approach to management of the environment and natural resources.
Harmful algal blooms (HABs) and fish kills linked to N or other nutrients have caused substantial losses to the seafood industry. Whitehead et al. (2003) estimated that the lost consumer surplus due to a dinoflagellate (Pfiesteria sp.) related fish kill is between $37 million and $72 million in the month following a fish kill. Jordan et al. (unpubl. data) provide a more comprehensive estimate of the damage costs of eutrophication on fisheries production by estimating the damage to fisheries via reductions in the area of submerged aquatic vegetation (SAV) along Mobile Bay (Gulf Coast of USA). They estimate that a 20% loss of SAV damage cost in 2008 dollars to combined shrimp and crab fisheries is $764 ha−1 year−1 per unit SAV habitat. Using an empirical response function of the impacts of N loading on SAV extent (Latimer & Rego 2010), a 20% loss in SAV due to N would have an impact on crab and shellfish production of c. $56 per kg N (S. Jordan, pers. comm.). Production of shrimp and crabs in Gulf estuaries is large and sensitive to habitat loss (Jordan et al. 2009), and damage to this valuable fishery is one of the highest per kg N damages we identified (Table 2).
N and climate regulation
Nitrogen plays a key role in the maintenance of a stable climate, a crucial regulating ecosystem service, by influencing the production of several greenhouse gases (N2O, CO2 and CH4) and through its role as a mediator of aerosol production. Human alteration of the N cycle affects Earth’s climate system via direct and indirect pathways. Nitrogen availability provides a fundamental constraint on plant growth and net CO2 uptake across much of the world, now, and in response to rising atmosphere CO2 concentrations in the future (Hungate et al. 2003). As discussed above, N inputs from atmospheric deposition can enhance plant growth rates and may account for a significant fraction of current terrestrial C uptake in some systems (Liu & Greaver 2009; Thomas et al. 2010). Furthermore, additions of N to some soils can inhibit decomposition, slowing release of CO2 to the atmosphere and leading to an increase in soil C stocks (e.g. Janssens & Luyssaert 2009).
However, net greenhouse benefits of C storage by some ecosystems may be somewhat dampened by the production of other greenhouse gases. In a meta-analysis, nitrogen additions were found to stimulate CH4 production, decrease CH4 uptake and increase N2O production (Liu & Greaver 2009). Atmospheric N2O concentrations are increasing rapidly in response to N enrichment of terrestrial and aquatic systems, and are presently 16% greater than during pre-industrial times (Forster et al. 2007). Due to high per-molecule warming potential, small changes in N2O concentrations have a disproportionately large effect on the climate system. N enrichment directly increases N2O production by stimulating nitrification, the oxidation of ammonium to nitrate (Robertson & Tiedje 1987), and denitrification (Seitzinger et al. 2006). N2O is a byproduct of both of these microbially mediated transformations. N availability also affects the rate of N2O production, both by increasing the overall rate of each N transformation process and by affecting the fraction of nitrification or denitrification that produces N2O rather than nitrate or N2 (Beauchamp 1997). The net effect of N enrichment on CH4 emissions is a function of competing processes. Atmospheric NOx and resulting ozone maintain high concentrations of hydroxyl in the atmosphere, which serves to remove atmospheric CH4 (Isaksen et al. 2009). And in anaerobic soils, an abundance of nitrate can decrease rates of CH4 production by increasing soil and sediment redox potential (Reay & Nedwell 2004).
Nitrogen also influences the climate system through its link to ozone. In the lower atmosphere, N plays a key role in tropospheric ozone production (Skalska et al. 2010). In turn, ozone affects the climate system directly by acting as a greenhouse gas with roughly double the climate effect of N2O (Forster et al. 2007), and indirectly through effects on photosynthesis and plant uptake of atmospheric CO2. Ozone damage to plants, as discussed earlier in the section on production, also may decrease plant uptake of atmospheric CO2 by as much as 14–23% (Sitch et al. 2007), leading to more CO2-driven warming.
In addition to affecting the balance of greenhouse gases in the atmosphere, production of NOx and NHy increases the concentrations of atmospheric aerosols, which aside from their negative health effects can provide substantial cooling, both directly (due to high reflectivity) and indirectly (by mediating cloud formation). Sulphate aerosols and nitrate aerosols act similarly in these processes, with the role of nitrate aerosols expected to increase in the future (Adams et al. 2001).
The influence of reactive N continues into the upper atmosphere, where ozone acts to provide a small amount of cooling. In this portion of the atmosphere, N2O currently is the most important contributor to the breakdown of stratospheric ozone, both now and in future projections (Ravishankara et al. 2009). Regulatory actions stemmed the production of CFCs that were formerly the dominant driver of depletion of the protective stratospheric ozone layer, but N2O production has continued to increase. Thus, N2O is currently the dominant and largely unregulated driver of UV-related damages to ecosystems and human health. The global benefits of the Montreal Protocol in reducing the use of ozone-depleting chemicals were estimated to be $300 billion (2008 dollars) for the period 1987–2060, and this did not include the human health benefits, such as 333,500 avoided skin cancer deaths (Smith et al. 1997a,b). We were not able to obtain damage costs to individual services, but collective UV damages associated with CFCs are estimated to be $49,669 per metric ton (Talberth et al. 2006). The ozone-depleting potential of N2O is c. 0.017 relative to CFCs (Ravishankara et al. 2009) so damages would be $844 per metric ton of N2O. Based on these values, potential UV-related damages related to N2O production in the USA are c. $1.33 kg−1 N2O-N.
Clearly N has the potential to modulate the ecosystem service of climate regulation. However, the relative importance of various N effects on climate is poorly understood, as are interactions between effects. Birch et al. (2011) were not able to find economic valuation functions to monetize the effects of N on greenhouse gases and climate regulation in their analysis of the effects of decision about N management in the Chesapeake Bay watershed. Recently, Kusiima & Powers (2010) identified several efforts to provide preliminary values of the anticipated impacts of greenhouse gases of c. $4–10 per ton of CO2, equivalent to $1.2–3.1 per kg N.
More research is clearly needed to elucidate interactions between N enrichment and climate at multiple scales and in multiple systems. In order to implement an ecosystem services approach to managing N with respect to climate influences, one would need to understand the relative magnitude of different N effects on the climate system, as well as gain an understanding of interactions between various N effects, dominant feedback mechanisms and thresholds. In addition, one would need a way to value the climate regulating properties of N in a manner that made it possible to compare the worth of such services to the value of other N-related ecosystem services. Consider the net greenhouse gas implications of N reduction efforts. Wetland and riparian restoration may be conducted in order to reduce nutrient loading and eutrophication of surface waters, but these activities have the added benefit of substantial carbon sequestration and the cost of additional greenhouse gas production (CH4 and N2O). Jenkins et al. (2010) determined that existing markets yield an estimate of $70 ha−1 for wetlands in the Mississippi River alluvial valley (USA), but when accounting for additional benefits such as nitrogen mitigation, waterfowl recreation and other valued services, the wetland value estimate rose to $1035 ha−1. A framework that included a full accounting of different N reduction strategies and net benefits would allow for more optimal and efficient N management.
N and maintenance of human health
Tremendous benefits to human health and well-being have resulted directly or indirectly from human alteration of the N cycle, particularly in terms of nutrition, materials (e.g. wood, paper, fabric), and provision of heat, light and transportation. Many of these positive impacts are quite evident, and can be tracked through economic indicators. However, when N is transported downwind and downstream from sites where its use is primarily beneficial to humans, it can become a hazard to human health (Townsend et al. 2003). These detrimental impacts are more challenging to track and do not correlate with the benefits (Raudsepp-Hearne et al. 2010). In the atmosphere, NOx is an important precursor of tropospheric ozone and particulate matter, which can increase rates of asthma and other respiratory issues, particularly in children and other vulnerable populations (Delucchi 2000).
The provision of clean water for drinking and other domestic uses is a key ecosystem service, and unfortunately nitrate contamination in drinking water is a growing issue in the USA. The number of drinking water violations of the nitrate standard in community drinking water wells increased from c. 650 to 1200 between 1998 and 2008 (US EPA 2009). Excess nitrate in drinking water has been associated with a number of illnesses, including blue-baby syndrome and several types of cancers (Townsend et al. 2003; Ward et al. 2005), although there is disagreement in the literature on these points (Powlson et al. 2008). Communities across the USA are dealing with nitrate contamination in drinking water, and making choices between replacement, treatment and prevention. Many of these choices will be based on costs and tradeoffs between ecosystem services.
In addition to direct effects from N enrichment of air and drinking water, excess N in surface waters can also have indirect effects on human health through, for example, stimulation of HABs that produce toxins (Camargo & Alonso 2006), outbreaks of dangerous pathogens like Cryptosporidium, or simply unpleasant odours and tastes that are costly to remove. There is also some suggestion that N enrichment can exacerbate pathogens such as West Nile virus, pollen allergens, swimmer’s itch, malaria, and cholera (Townsend et al. 2003; Johnson et al. 2010). Even where nitrate concentrations are below the US EPA drinking water standard (10 mg nitrate-N L−1), nitrate and eutrophication can increase treatment costs of safe drinking water. Some treatment processes designed to remove the products of eutrophication can introduce harmful byproducts into drinking water (Cooke & Kennedy 2001).
The costs of human health problems related to N have been evaluated in a number of studies. In a detailed review of the valuation of air quality regulations on humans and ecosystems, the mortality and illness associated with reactive N forms as precursors to PM and ozone were the most substantial of the measured effects (Table 3; Chestnut & Mills 2005). A number of US and EU studies have also examined the cost of NOx and NHy effects on respiratory health; most recently the ExternE project determined the health impacts of reactive N in air to be $28 per kg of NOx-N and $16 per kg NH3-N (Bickel & Friedrich 2005). The health impacts for NH3-compounds are uncertain, and costs could be lower (van Grinsven et al. 2010). These values are similar to estimates used in a Chesapeake Bay N assessment (Table 2; Birch et al. 2011).
Table 3. Abatement costs of reducing nitrogen from various individual sources and from integrated projects. For comparison, the price of N fertilizer was c. $0.44 per kg N (1980–2000) and in 2008 was c. $1.21 per kg N (Bruulsema & Murrell 2008)
|Cost||$ kg−1 N||Location||Reference|
| Electric utilities/NOx||$4.80||Chesapeake Bay, USA||Birch et al. (2011)|
| Industrial/NOx||$22.00||Chesapeake Bay, USA||Birch et al. (2011)|
| Mobile sources||$14.00||Chesapeake Bay, USA||Birch et al. (2011)|
| Non-agricultural/NH3||NE||Chesapeake Bay, USA||Birch et al. (2011)|
| Agriculture/NO3||$10.00||Chesapeake Bay, USA||Birch et al. (2011)|
| Urban and mixed land use/NO3||$96.00||Chesapeake Bay, USA||Birch et al. (2011)|
| Point Sources||$18.00||Chesapeake Bay, USA||Birch et al. (2011)|
| Agricultural drainage water/NO3||$2.71||Mississippi Basin, USA||Jaynes et al. (2010)|
| Current expenditures towards meeting Chesapeake TMDLs – (1985–2009)||$8.76||Chesapeake Bay, USA||US EPA (2009); Blankenship (2011)|
| Projected costs to meet Chesapeake TMDLs – (2010–2025)||$14.27||Chesapeake Bay, USA||US EPA (2009); Blankenship (2011)|
| Projected costs of using wetlands to control nutrient damages||$4.40–5.62||Mississippi Basin, USA||Kusiima & Powers (2010)|
| Estimated cost for achieving a 45% reduction in nitrate-N load||$2.50||Cedar River Watershed, Iowa, USA||Helmers & Baker (2010)|
Few studies comprehensively address all impacts of N on drinking water and human health, but a number of pieces of the puzzle are available. A study using limited data on the link between colon cancer and nitrate determined a health cost of $0.1–3.4 per kg N leaching to groundwater in the EU (van Grinsven et al. 2010). Several studies have examined the impacts of HABs in coastal areas, which may be associated with N (Table 2). Hoagland et al. (2002) found that impacts of illnesses resulting directly from shellfish poisoning in the USA were difficult to estimate but used mortality, hospital visits and lost worker hours to estimate that one paralytic shellfish poisoning event cost c. $6 million. Corso et al. (2003) found that the total cost of a single Cryptosporidium outbreak during 1993 in Milwaukee, Wisconsin (USA) was $96.2 million: $31.7 in health costs and $64.6 million in lost worker productivity. We could not obtain or develop damage cost per kg N estimates for health effects of harmful algae or pathogens, in part because the causes are sometimes unclear. More research is needed to better test exposure-response relationships and the transferability of the limited number of these relationships (van Grinsven et al. 2010).
Another way to look at the problem is to consider the cost to treat contaminated water. Approximately 15% of the US populations or 45.4 million people use water from private wells for domestic use (Hutson et al. 2004), and a recent survey indicated that 4.4% of private drinking water wells in the USA were higher than the US EPA nitrate standard for human consumption (DeSimone 2009). Based on these estimates, c. 2 million people are on well water containing nitrate above the US EPA standard for human consumption. If it costs $560 per person to treat well water (US EPA 2009), then we estimate that the cost to treat nitrate contaminated well water is c. $1.12 billion dollars. We multiply this number by the c. 7 Tg of nitrogen that moves from land to water in the USA and [value of 23 Tg inputs to land from Fig. 1 and assuming that transport to groundwater is equivalent to the transport factors of 0.3 from Smith et al. (1997a), yielding an estimate of c. $0.16 per kg N for groundwater contamination of drinking water. This value is greater than the estimate of Dodds et al. (2009), indicating a need for further review of these damage costs. Our review here indicates that there are tremendous health impacts and consequences of nitrogen pollution, and including the full range of these impacts, not the just the well-studied impacts in air, will better inform decisions related to the management of N.
N and maintenance of biodiversity and aesthetics
Excess N can affect the integrity, resilience and beauty of the natural world by reducing biodiversity. This loss of biodiversity can occur through a number of different mechanisms. N additions can cause shifts in primary producer communities in both terrestrial and aquatic systems, leading to decreased biodiversity (e.g. Deegan et al. 2002; Dupréet al. 2010). A recent global analysis further supports the notion that N deposition is the main driver of altered species composition in a range of ecosystem types and in some cases this includes an increase in invasive species (Bobbink et al. 2010). Species that tend to show increases in abundance are often non-native invasives with high vegetative and population growth rates, which have the potential to drive local populations of rare native species to extinction (Bobbink et al. 2010). In some, but not all types of wetlands, increased productivity is associated with decreased plant diversity (Bedford et al. 1999); moreover, rare or more ecologically valuable species may be replaced by generalists and invasive species (Morris 1991). Furthermore, the loss of plant species due to N deposition can be detrimental to insect herbivores that depend on them, as exemplified by checker spot butterflies in serpentine grasslands of California (Weiss 1999). The provision of habitat for organisms which influence the integrity, resilience, spiritual value and beauty of the natural world is an important service (e.g. Losey & Vaughan 2006).
Ecosystem acidification via atmospheric N deposition is another driver of species changes. Following deposition, nitrate can leach out of soils, carrying with it a loss of base cations (K, Ca, and Mg). Soil acidification can also lead to mobilization of inorganic Al (Reuss 1983) with detrimental effects on tree health, including aluminium interference with calcium uptake, cold tolerance, and aluminium toxicity to roots (Parker et al. 1989; Cronan & Grigal 1995). These leaching processes usually result in lower pH in soil solution and streamwater, and higher concentrations of inorganic monomeric Al. Low pH and inorganic monomeric Al are directly toxic to fish (Baker & Schofield 1982), and fishless lakes in the Adirondacks have significantly lower pH and acid neutralising capacity than lakes with fish (Gallagher & Baker 1990). Leaching of Al from soils into sensitive aquatic systems also has been shown to reduce fish diversity (Nierzwicki-Bauer et al. 2010). These shifts in fish abundance and diversity have implications for sport fishing and recreation, as well as cultural and existence values (Banzhaf et al. 2006).
High rates of N loading to surface waters can contribute to excessive productivity, or eutrophication, characterized by algal blooms that prevent swimming, fish consumption and/or other human use (Van Dolah 2000), hypoxia (Breitburg et al. 2009), shifts in species composition (Vaas & Jordan 1990) and food webs, and water with unpleasant tastes and odours (Pretty et al. 2003). These factors negatively affect fish production, biodiversity, water quality, recreation potential, aesthetics and human health. Dodds et al. (2009) conservatively estimate that the costs of freshwater eutrophication, including costs to recreation, waterfront real estate, and spending on recovery of threatened and endangered species in the USA are c. $2.2 billion per year.
Nitrogen can also influence how humans experience nature. Nitrogen is a component of regional haze, which can affect visibility and decrease aesthetic enjoyment of places where people live, work and recreate, including parks and other rural areas (Malm 1989). Visibility damages associated with reactive N in the Chesapeake Bay watershed were $120 million (Birch et al. 2011). Damage costs of HABs to recreation and tourism range from < $1–28 million (Hoagland et al. 2002). Damages by reactive N to recreational use within the Chesapeake Bay estuary were estimated to be $730 million per year (Birch et al. 2011).
Some studies have estimated the value of improving the quality of natural resources by asking people what they are willing to pay. Banzhaf et al. (2006) estimated that New York state residents are willing to pay $45–100 each year to reduce the number of acidified lakes and improve forest health in the Adirondacks, which translates to $300–700 million for all state residents. A key challenge to this approach was that the effects needed to be explained to and understood by the respondents. Thus, in addition to accounting for human well-being in such a decision framework, an effort must be made to reach out to and educate the public to ensure that they are aware and have sufficient knowledge about the connections between N reductions and benefits.