Nitrogen applications to many managed systems are made to increase production and yield, and effects on biodiversity and ecosystem function in such systems are of lower concern. Atmospheric nitrogen deposition is the total of dry and wet deposition from the air and, in areas with little direct fertiliser nitrogen input to soils and waters, accounts for the majority of the annual nitrogen input. Dry deposition consists of gases, such as NO2 and NH 3, and fine particles (including fine particle NH4). Wet deposition (soluble forms in, for example, rain and snow) consists of NH4, NO-, and forms such as HNO-, (NH4)2SO4, and NH4NO3.
The deposition of reduced nitrogen (NHX) is highest near its emission source, usually an agricultural source, whereas oxidized nitrogen compounds (NO^,), and especially NO2, are transported over larger distances (so-called long-range trans-boundary pollutants). The deposition rate is also strongly influenced by the structure of the vegetation onto which it is deposited. For example, atmospheric nitrogen deposition onto vegetation with a rough canopy structure, for example, a pine forest, is considerably higher (>50%) than that onto smooth surfaces, for example, a calcareous grassland. When studying ecotoxicological effects of nitrogen, these differences must be considered.
The effects of increased nitrogen deposition on terrestrial ecosystems have been most studied in northern and
Western Europe, partly because nitrogen input into terrestrial ecosystems from atmospheric deposition was extremely high in Europe in the 1980s and early 1990s, reaching 30-40 kg ha-1 yr-1 in many areas of central Europe and exceeding 60kgha-1yr-1 in others. Although nitrogen deposition from the atmosphere declined after the 1990s as a result of the impact of new legislation, current deposition rates in, for example, the Netherlands are still very high (up to 42 kgN ha-1 yr-1) and among the highest in the world. For this reason, ecotoxicological effects of nitrogen are most clearly understood in this region.
Of all the reactive nitrogen forms, NH3, NH|, and NO-are the most important in terrestrial ecosystems. Since ecotoxicity is very much a process depending on the dose and concentration of the contaminant, the deposition rate and the distance to the source determines to a large extent the potential for toxicity. Nitrogen in the form of nitrate (NO-) is not considered toxic to plants, and in fact stimulates growth of many plant species. Since it does not bind strongly to soil particles, NO3- is easily leached to deeper layers and is therefore normally not present in high concentrations in the rooting zones of the soil. In contrast, NH 3 and NH4 can be very toxic in terrestrial ecosystems. In the soil, NH3 is readily converted into NHj which binds strongly to soil particles, resulting in accumulation of this form of nitrogen in high concentrations.
In general, three groups of plants can be distinguished, based on their nitrogen nutrition: those preferring NO-, those preferring NH4, and those using both NH4and NO-. NHj-tolerant species are usually found in acidic habitats and are likely to be adapted to NHj nutrition, since NH4 is the dominant form of nitrogen in soils at low pH. These species are mainly slow-growing perennial species. In contrast, species from more buffered habitats usually prefer NO- and are also usually faster-growing annual species. Despite these species preferences, NHj is readily taken up by both NHi-tolerant and -intolerant species, as NH4 uptake is energetically more favorable than NO 3 uptake. Uptake of the different nitrogen forms can occur via the roots and via the leaves (foliar uptake). Following uptake, NHj is either assimilated into amino acids and other organic nitrogen compounds or accumulated as free NH4. The assimilation of NH4 takes place in the cells, mainly via the glutamine synthetase/gluta-mate synthase pathway and produces protons.
Accumulated NH4 can cause severe toxicity symptoms in sensitive plant species, bryophytes, and lichens. The capacity to assimilate NH4, and the rate at which this takes place, determines whether these species are sensitive to NH4 or NH3 nutrition. Direct NH4 toxicity effects on vascular plants and mosses can be clearly shown in hydroponic experiments. Many characteristic species
from species-rich grasslands are sensitive to NHj (e.g., Antennaria dioica, Arnica montana, and Cirsum dissec-tum), and NH^ toxicity symptoms in these species are expressed as severe growth suppression and high mortality (Figure 2).
The mechanisms behind toxicity of NH3 and NHj differ between species. In general, elevated NH3 and NHj uptake decreases photosynthetic capacity and results in cation deficiency and chlorosis. At the same time, concentrations of the anions CP and SO4~ can increase, resulting in a severe charge imbalance in the plant cells. Finally, NH3 and NH| uptake results in disruption of cell membrane functioning. The decline of the brown moss Scorpidium scorpioides, for example, is directly related to serious disruption of cell pH regulation, reduction in nitrate reductase activity, and membrane dysfunction due to increased NHj nutrition. One of the key mechanisms of NH| toxicity in vascular plants is linked to NH^ uptake mechanisms; with the uptake of NH^, cations such as Ca2+, Mg2+, and K+ (but mainly K+) are actively excreted and uptake of cations (especially Mg2+) is prevented, causing cation deficiencies and chlorosis in the leaves. Other recorded negative effects in higher plant species are a reduction in fine roots, a decline in mycorrhizal associations, and strongly reduced germination and seedling establishment.
In the air, nitrogen oxides dissolve in small water droplets and form the acidic molecules HNO2 and HNO3. As a result, nitrogen oxides cause a drop in pH of the rain. Natural rain has a pH in the range of 5.0-5.6, whereas acidified rain has a pH between 3.0 and 5.0. Sulfur dioxide (SO2) also causes rain acidification, and was the dominant contributor in the late nineteenth century and the first half of the twentieth century, when acidification had devastating effects in many sensitive ecosystems in Europe and North America. As a result of strong reductions in sulfur emissions since the 1970s in these regions, much of the remaining acidification due to atmospheric deposition is caused by nitrogen-containing compounds. Nitrogen in the reduced (NHX) form also causes acidification in the soil through both uptake by plants and increased nitrification (the microbial conversion of NHj into NOf). Both processes produce protons and cause direct acidification of the soil.
Acidification seriously affects terrestrial ecosystems, especially slightly buffered systems. Following acidification, protons (H+) are exchanged on the soil complex for macronutrients such as calcium (Ca2+), magnesium (Mg2+), and potassium (K+), thereby buffering the soil against the acidifying processes. Subsequent leaching of Ca2+, Mg2+, and K+ leads to loss of the soil's buffering capacity by base cations and to nutrient imbalances for plant growth. Buffering by cations and leaching will continue until all base cations are exchanged and continuing acidification will lead to a shift in the buffer range of the soil from cation buffering (pH 4.5-6.0) to aluminum (Al3+) buffering (pH < 4.5). Consequently, Al3+ and micronutrients such as manganese (Mn2+) and iron (Fe2+, Fe3+) are exchanged for protons, resulting in increased concentrations of these free metal ions in the soil. Free metals are known to be highly toxic to most soil organisms. The decline of some characteristic plant species of species-rich acidic grasslands and heathlands in Western Europe is due to effects of aluminum toxicity and increased Al/Ca ratios in the soil, and to base cation depletion, as a result of increased acidification.
An increase in total nitrogen deposition drastically increases the availability of different forms (NH^, NOJ of nitrogen in the soil, either directly or indirectly through processes such as increased mineralization, litter turnover rates, and nitrification. Many natural and seminatural ecosystems (including forests, heathlands, and mat grass swards) are naturally nutrient poor (oligotrophic or meso-trophic) and nitrogen limited, which has resulted in a high plant diversity with characteristic species adapted to nitrogen-limited, nutrient-poor environments, and the low levels of competition typical of these conditions. An increase in nitrogen availability often results in a higher nitrogen uptake and higher plant productivity of competitive species for which elevated NH^ is not directly toxic. For example, grasses such as the highly competitive species Deschampsia flexuosa do not respond negatively to elevated NH| and may even increase in biomass, indicating that the increased NHj input may eventually lead to an increase in competition for light and nutrients and hence to changes in species composition.
The decline of many characteristic species and the increase in highly competitive species (mainly grasses) is therefore often attributed to elevated nitrogen deposition. Such changes in species composition due to elevated nitrogen input have been observed in pine forests, heathlands, nutrient poor grasslands, wetlands, and bogs in Western Europe, where highly competitive grasses such as Deschampsia flexuosa and Molinia caerulea have increased at the expense of less competitive forb species. With increasing nitrogen loads, nitrogen becomes readily available and other nutrients become limiting. In oligotrophic and meso-trophic ecosystems, shifts from nitrogen limitation toward phosphate (P) limitation are commonly found. P availability can limit plant production and determine species composition in communities such as calcareous grasslands and calcareous dune grasslands. With increased nitrogen availability, the nitrogen content in these ecosystems increases relative to P content, affecting plants' susceptibility to herbivores and pathogens, as well as mycorrhizal infection rates, and sensitivity to frost and drought.
One example of these indirect mechanisms leading to ecotoxicological effects is a major shift in vegetation composition in Dutch heathlands that was linked to high atmospheric nitrogen input. Here a dominance of the shrub Calluna vulgaris was replaced by a dominance of the grass Deschampsia flexuosa. Although D. flexuosa is more competitive than the slow-growing C. vulgaris, in mature Calluna stands, invasion does not occur because of competition for light. Other key ecosystem effects were linked to increased nitrogen content of the vegetation, which increased the plants' susceptibility to herbivores like the heather beetle (Lochmaea suturalis). When C. vul-garis plants die as a result of such attacks, they leave large empty patches in which fast-growing grasses can outcompete the young Calluna plants if nitrogen availability increases. Grass species only started to dominate Dutch heathlands when the canopy of C. vulgaris was opened up by processes such as frost damage and herbivory.
Although direct toxicity effects might damage plants, they do not necessarily result in immediate mortality. For example, in Western European and North American coniferous forests (in particular pine and Douglas fir), nitrogen pollution has led to reduced growth rates, a higher percentage of brown and dead needles, and accelerated needle loss. In addition, acidic precipitation causes direct damage to the leaves by decreasing the wax layer, making the trees more vulnerable to disease. Nitrogen deposition also reduces the availability of base cations. In addition, geochemical conditions in the soil are affected as the concentrations of free metal ions (such as Al3+) increase and pH levels decrease, causing serious deleterious effects to the roots. All these processes weaken the trees, making them vulnerable to drought, frost, diseases, and insects which are likely to ultimately kill them.
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