OR/14/047 Nutrients and wetlands

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Farr, G, and Hall J. 2014. Atmospheric deposition and groundwater dependent wetlands: implications for effective catchment management and future Water Framework Directive groundwater classification in England and Wales. British Geological Survey Internal Report, OR/14/047.

The atmospheric deposition of nutrients, mainly nitrogen and its various chemical species, and its effects upon GWDTEs comprise the main focus of this report. The effects of elevated nitrogen deposition and a reduction in plant species richness is well documented (e.g. Stevens et al. 2010)[1]. The majority of GWDTEs considered within this report are low nutrient systems and exposure to prolonged or elevated levels of nutrients may cause significant ecological damage.

It is beneficial at this early stage to provide a brief description of the nitrogen cycle, also outlining sources of non-atmospheric nitrogen, the various species of nitrogen and the processes that facilitate the changes from one form of nitrogen to another. The description of the nitrogen cycle will be discussed in the following subchapters and will follow a source-pathway-receptor approach; the receptors in this example are GWDTEs.

The nitrogen cycle, simplified in Figure 1 illustrates the pathways and receptors for atmospheric nitrogen and inorganic and organic fertilizers in the environment. Future work requires an improved understanding and quantification of the N cycle, particularly relatively unstudied processes such as dry deposition, N fixation and decomposition/rnineralisation (Adams, 2003)[2].

File:OR14047fig1.jpg
Figure 1    Simplified Nitrogen Cycle (BGS).

Sources of atmospheric nitrogen

Atmospheric pollutants are diverse, and include nutrients as well as other pollutants such as sulphur, base cations, heavy metals and gases. This report focuses on atmospheric nutrients, primarily nitrogen and its species (both oxidized and reduced) that, in excess, can have a negative impact on GWDTEs.

Atmospheric nitrogen can arise from a variety of natural and anthropogenic sources and can be deposited as both wet and dry deposition (EA, 2005)[3]. Nitrogen can originate from activities occurring both locally and over large areas. Natural sources can include forest soils, that can emit about 10–13% of N compounds that were originally deposited as NH3/HN+ and HNO/NO-, back to the atmosphere as N oxides (Horvath et al. 2006)[4]. Lightning can also fix nitrogen from the atmosphere (Environment Agency, 2005)[5] although this is not a major contributor to atmospheric deposition.

Dentrification is the process by which bacteria reduce nitrogen, resulting in the release of gaseous nitrogen (N2) back into the atmosphere. Dentrification can occur within anaerobic areas of many wetlands which means that GWDTEs can themselves be a source of nitrogen. Drewer et al. (2010)[6] show that peatlands can be both sources and sinks of nitrogen (and other green house gases) and calculate nitrogen budgets for two peatlands in Northern Europe. Anthropogenic addition of nitrate to wetlands may even act as a catalyist and enable increased levels of N2O flux from wetlands (e.g. Liu and Greaver, 2009)[7] and Moseman-Valtierra, 2011)[8].

In addition to atmospherically derived nitrogen there are many anthropogenic and natural terrestrial sources of nitrogen. It is important to consider all sources of nitrogen that can potentially cause significant damage as this will improve future N budgets or source apportionment studies. Nitrates in groundwater are a widespread issue across the UK, with the application of fertilisers, sewage sludge and crop residues coupled with changes in landuse allowing both diffuse and point sources of nutrients to enter controlled waters (i.e groundwater and surface waters). Monitoring of nitrate levels in groundwater and surface water is established across England and Wales, with reporting undertaken for every groundwater and surface water body. Other anthropogenic sources of nitrogen in groundwater include: leaking sewers, application of sewage sludge to land, landfills and septic tanks (BGS, 1996)[9]. Terrestrial sources are often referred to as ‘diffuse pollution’ although ‘point sources’ such as non-mains waste water treatment and waste disposal can also contribute to the nitrate in controlled waters. In reality many dispersed point sources can appear to come from one single source of diffuse pollution (EA, 2005)[3].

Oxidised and reduced nitrogen

Atmospheric nitrogen can be divided into two broad categories; oxidised and reduced (Table 1). When nitrogen (N) is oxidised it gains an oxygen molecule/s forming either nitric oxide (NO), nitrogen dioxide (NO2), nitrous acid (HONO) or nitric acid (HNO3) and if it is reduced it forms ammonia (NH3). Oxidised and reduced nitrogen can be further divided on their sources; oxidised nitrogen tends to be sourced from anthropogenic combustion processes (e.g. power generation and traffic), whereas reduced nitrogen originates primarily from agricultural processes.

Table 1    Sources of oxidised and reduced nitrogen, adapted from RoTAP (2012)[10].
Oxidised Nitrogen Sources
Nitrogen oxides (NO) Combustion of fossil fuels from traffic and urban sources and industrial emissions. NO and NO2 are collectively known as NOx
Nitric oxide (NO)
Nitrogen dioxide (NO2)
Nitrous acid (HONO)
Nitric acid (HNO3) Also from nitrogen gas and water vapor during lightening strikes (not a major contributor to atmospheric nitrogen)
Nitrate (NO3-) Wet deposition and via surface and groundwater
Reduced Nitrogen Sources
Gaseous ammonia NH3 Agriculture, livestock, poultry, manure management (cattle) also synthetic fertilizer application
Aerosol NH4+ Associated with SO42- from emissions
Wet deposited NH4+ Agriculture: the effects of wet deposited NH+ are thought to be less than that of dry deposited NH3

Nitrogen Oxides (NOx)

NO and NO2 are collectively known as NOx and are formed when nitrogen (N) is oxidised forming nitrogen oxides (NOx). The primary source for air emissions of nitrogen oxides (NOx) are combustion sources e.g. road transport, public electricity and heat generation sector and industry (see RoTAP, 2012)[10].

Ammonia (NH3)

Ammonia (NH3) emissions are primarily sourced from the agricultural sector, specifically manure management, degradation of urea from livestock (cattle) but also from synthetic fertiliser applications (RoTAP, 2012)[10]. The sources of ammonia (NH3) can be both diffuse, sourced from large agricultural areas, and also from point sources such as pig and poultry farms, however many point sources can also produce diffuse pollution. The diffuse nature makes monitoring emissions for ammonia (NH3) more uncertain than for the combustion generated nitrogen dioxides (NOx). This also means that any modeled or spatial data will also be susceptible to the same uncertainties (RoTAP, 2012)[10]. This uncertainty will also apply to the 5 x 5 km grid square of atmospheric nitrogen deposition data used later on within this report (see Modelling of atmospheric deposition in the UK).

Pathways for atmospheric nutrients

Once emitted to the atmosphere compounds are formed and transported often over long distances, subsequently being deposited in the form of pollutants such as particulate matter (sulphates, nitrates) and related gases (nitrogen dioxide, sulphur dioxide and nitric acid). Once in the atmosphere there are two processes by which deposition can occur, that is via ‘WET’ or ‘DRY’ deposition, both of which can be considered as direct pathways at GWDTEs. Wet deposition is the portion dissolved in cloud droplets and is deposited during rain or other forms of precipitation (EPA, 1999). Dry deposition includes both gas and particle transfer to surfaces during periods of no precipitation (EPA, 1999). Both the wet and dry deposition can be deposited directly upon GWDTEs.

Indirect pathways for atmospheric deposition involve: surface water, surface water runoff and groundwater to a GWDTE. The cumulative effect of atmospheric nutrient deposition across a groundwater body (or catchment of a GWDTE) must be considered for any successful source apportionment study and will be influenced by landuse, vegetation, soils, rainfall and topography.

Understanding the contribution of atmospheric loading and terrestrial loading on a catchment scale will be important for implementing effective and targeted management plans for both the WFD and HD. A general rule of thumb is that terrestrial loading at lowland habitats far exceeds loading from atmospheric sources.

Receptors and factors affecting atmospheric nutrient deposition and loss

Atmospheric deposition does not discriminate and its effects are felt by a variety of receptors including: soils, freshwater and vegetation (see RoTAP, 2012)[10] and also seawater where nutrients can contribute to algal blooms. Responses and changes to atmospheric deposition occur in soils, freshwater and vegetation and affect a wide range of ecosystems (RoTAP, 2012)[10]. Atmospheric deposition is an important source of N in semi-natural upland ecosystems (Helliwell et al. 2007)[11] as many upland systems maybe exposed to less terrestrial nitrogen sources due to their topographical setting and surrounding low intensity land use. In the context of this report vegetation at GWDTEs must be considered as the principal receptor because most GWDTE are defined in terms of vegetation characteristics and it is change within the vegetation that is used to determine if a GWDTE is in unfavourable condition.

Vegetation

There is strong evidence that the effect of nitrogen deposition on vegetation in general (and not just GWDTEs) has already been reflected by a significant reduction in total plant species, diversity and frequency of sensitive plant species since the 1980s (RoTAP, 2012)[10]. The effects of atmospheric N deposition on species diversity is not straight forward and for any given habitat it will depend on abiotic conditions including: buffering capacity, soil nutrient status and soil factors that influence the nitrification potential and nitrogen immobilisation rates (Bobbink et al. 1998)[12]. Maskell et al. (2010)[13] found a strong negative correlation between atmospheric nitrogen deposition and plant species richness in selected habitats (heathland acid, calcareous and mesotrophic grassland) in the UK. Maskell et al. (2010)[13] also highlights the complexity and interactions of land management and grazing and their influence on the susceptibility of sites to nitrogen deposition. Nitrogen deposition has also been shown to have a cumulative impact (e.g Dupre et al. 2010)[14]. The national 5 x 5 km deposition maps (see Modelling of atmospheric deposition in the UK) are based on annual mean deposition rates; the difficulty of quantifying the effect of cumulative deposition should be considered especially during any future source apportionment study. Furthermore Stevens et al. (2011)[15] highlight the ability of certain species to be impacted even at low levels of nitrogen deposition — even below that of the critical loads (for explanation see Critical loads).

Changes in vegetation can also result from the failure to implicate suitable grazing regimes, abandonment of sites (i.e no management) or historic management decisions such as the stabilization of many dune systems across coastal areas in the UK. It is important to consider how the effects on vegetation of land management changes and vegetation management can be distinguished from the effects of atmospheric (and terrestrial) impacts during any source apportionment study.

Vegetation is the primary receptor for atmospheric deposition at many GWDTEs. CSM or common standards monitoring (see JNCC, 2004)[16] and repeat surveying of vegetation is used to identify indicator species that are related to nutrient enrichment. The first six year report on common standards monitoring Williams (2006)[17] states: It is often very difficult to determine the effects of air pollution natuural or semi natural habitats, given the complex interactions between pollution impacts, management and abiotic influences. As a result, the impacts of air pollution, and the identification of air pollution as an adverse activity affecting condition, are considered to be substantially under-reported in this assessment.

There are however some concerns regarding this approach and these are raised by Emmett et al. (2011)[18]. and also summarized in Chapter 12 within this report. Different habitats are assigned nitrogen critical loads (ie, thresholds for the impacts from atmospheric deposition; these are discussed in more detail in Critical loads and recent data analysis (Stevens et al. 2011)[15] and Emmett et al. 2011)[18] show that for many habitats across large areas of the UK, nitrogen deposition exceeds the critical loads.

Soils

Topsoil nitrogen concentration has decreased in many habitats despite continued total atmospheric nitrogen deposition remaining the same over the last 20 years (RoTAP, 2012)[10]. The reasons for this are not known but could be associated with changes in the C:N ratios such that the nitrogen signal is diluted by increased C fixation by plants or that microbioal activity has been effected by N deposition, thus increasing the availability of N to plants, RoTAP (2012)[10]. Nitrogen (N) is however immobile in soil organic matter, relatively little is leached to freshwaters (RoTAP, 2012)[10] and it is therefore important to consider cumulative nitrogen (N) deposition rather than present day deposition (Emmett et al. 2011)[18]. The importance of abiotic conditions including soil nutrient status and buffering capacity all affect the ability for NO3-/NH4 nitrification and mobilisation (Bobbink et al. 1998)[12] and thus the impact it can have on any receiving ecosystem.

A study into four UK upland catchments (Helliwell et al. 2007)[11] describes how nitrogen concentrations in acid sensitive upland soils were greater in areas dominated by mineral soils with a small C-pool rather than peaty soils (large C –pool). Helliwell et al. (2007)[11] conclude that if nitrogen deposition remains at current levels then it is possible that upland catchments with small C –pools will be more susceptible to NO3- leaching, thus having a direct impact on habitats and freshwater systems that receive water from these upland catchments. Helliwell et al. (2007)[11] describe how the geomorphology (slope, altitude and bare rock) of upland catchments may provide a control for winter NO3- leaching and how in the summer significant relationships between the C pool and surface water NO3- were observed. The implication for this is that any GWDTEs that receive an element of surface water flow from an upland catchment may also be indirectly impacted by the ability of the soils and other geomorphological characteristics to limit (or enhance) leaching of NO3- during the year. Source apportionment studies or models to understand atmospheric nitrogen deposition across groundwater bodies would need to consider soil types, slope, altitude and areas of bare rock within the analysis.

Seasonal variation and climate change

The natural variability of rainfall (intensity and amount) varies seasonally across England and Wales, with winter periods traditionally being wetter than summer periods. This natural variability of rainfall has a direct influence on wet atmospheric deposition, and this is factored into the Concentration Based Estimated Deposition (CBED: RoTAP, 2012)[10] data for England and Wales (see Modelling of atmospheric deposition in the UK). During winter biological uptake and transformation of nitrogen is greatly reduced (Helliwell et al. 2007)[11] and this also generally corresponds with periods of greater rainfall and wet deposition.

Nitrogen loss from wetlands can also vary throughout the year as seasonal patterns of organic carbon (important for dentrifying bacteria) loss changes depending upon plant types and their ability to create varying amounts of litter (Weisner et al. 1994)[19].

The potential effects of climate change on air pollution impacts on soils and vegetation are potentially very wide-ranging and are discussed in more detail in the RoTAP (2012)[10] report. The RoTAP report summarises the three main potential impacts of climate change on atmospheric deposition, they include;

(i) changes in the tolerances of plant species to soil acidification and N enrichment under different climate conditions;
(ii) increased frequency of climatic stresses to which air pollution increases sensitivity (e.g. drought); and
(iii) increased uptake, weathering and leaching of N and base cations and increased base cation weathering due to climate-induced changes in plant growth and hydrological conditions

Attenuation of nitrogen in wetlands

Attenuation of nitrogen in wetlands is a complex subject and although it must be mentioned it is beyond the scope of this project to deal with it in detail. The following is a very short description of some key issues related to the attenuation of nitrogen in wetlands, and a detailed review of the literature is needed to expand further upon this subject.

Nitrogen can be both retained, attenuated and lost (i.e. cycled) within many GWDTEs and the key processes associated with this are; nitrification, denitrification and uptake by plants. Dentrification is the primary mechanism for nitrogen retention (Saunders and Kalff, 2001)[20]. and occurs in anoxic environments when bacteria use the oxygen in nitrate for respiration and release N gas back to the atmosphere (Woods Hole Group, 2007)[21]. Denitirification also depends upon the release of organic carbon from plant litter and living macrophytes, which is used directly by denitirfying bacteria within wetlands and also indirectly by stimulating a lower redox potential (Weisner et al. 1994)[19].

In upland systems nitrogen is generally tightly cycled and retained, with minimal release to surface water or groundwater. However nitrogen saturation can occur in some systems if deposition exceeds the retention capacity of soils and biota in the system (Helliwell et al. 2007)[11]. The ability of wetlands to retain nitrogen has been highlighted by several studies: Chapman and Edwards, (1999)[22] and Davies et al. (2005)[23] suggest that the dominance of NO3 in inorganic N leaching in semi natural systems is due to the retention of NH4 via uptake, adsorption or nitrification within the ecosystems. Jansson et al. (1998)[24] describe the ability of wetlands in the Baltic sea drainage basin to retain 5–13% of atmospheric and terrestrial nitrogen, preventing eutrophication in the Balitic sea; however the potential of damage to the actual wetlands is not discussed in detail.

Modelling of atmospheric deposition in the UK

The deposition data used within this report, and also for the APIS (Air Pollution Information System) website (www.apis.ac.uk) is calculated on a 5 x 5 km grid using the CBED (Concentration Based Estimated Deposition) methodology. Maps are produced of wet and dry deposition of sulphur, oxidised and reduced nitrogen, and base cations using measurements of air concentrations of gases and aerosols as well as concentrations in precipitation from the UK Eutrophying and Acidifying Pollutants (UKEAP) network (Hall et al. 2014 [in press]). Site based measurements are interpolated to generate maps of concentrations for the UK. The ion concentrations in precipitation are combined with UK Met Office annual precipitation data to generate wet deposition. Gas and particulate concentration maps are combined with spatially distributed estimates of vegetation-specific deposition velocities (Smith et al. 2000)[25] to generate dry deposition. Examples of these maps are presented in Figure 2.

More detail on CBED can be found in RoTAP (2012)[10]; some of the key points are listed below:

  • Dry deposition of oxidized nitrogen is generated using data calculated from and interpolated between 30 sites
  • The use of vegetation-specific deposition velocities enables different deposition values to be derived for deposition to different land cover types; for critical load exceedances, values for moorland are applied to all non-woodland habitats, and deposition values for woodland are applied to all woodland habitats
  • Wet deposition mapping requires the use of an orographic enhancement factor which accounts for the natural variability in annual rainfall conditions which directly influences wet deposition
  • Deposition data used for calculating critical load exceedances are 3-year annual averages of the sum of wet plus dry deposition to moorland and woodland
File:OR14047fig2.jpg
Figure 2    CBED 5 x 5 km nitrogen deposition to moorland for 2010–12: (a) oxidized nitrogen; (b) total (oxidized + reduced) nitrogen.
Contains OS data © Crown Copyright and database right [2015].

There are several different models that can be used for air pollution modeling for both long (>50 km) and short range (<20 km) predictions, the main output being to provide an estimate of a concentration of deposition of a pollutant. The APIS (Air Pollution Information System) website (www.apis.ac.uk) is one of the main portals to this information and further details of modeled concentration and deposition values in the UK can be found in RoTAP, 2012[10] (see Critical loads) and at (https://pollutantdeposition.defra.gov.uk)

References

  1. ↑ STEVENS, C, DUPR`E, C, DORLAND, E, GAUDNIK, C, GOWING, D J G, BLEEKER, A, DIEKMANN, M, ALARD, D, BOBBINK, R, FOWLER, D, CORCKET, E, MOUNTFORD, J O, VANDVIK, V, AARRESTAD, P A, MULLER, S, and DISE, N B. 2010. Nitrogen deposition threatens species richness of grasslands across Europe. Environmental Pollution, 158(9), pp.294–2945.
  2. ↑ ADAMS, M. 2003. Ecological issues related to N deposition to natural ecosystems: research needs. Environment International, JUN, Vol. 29, No. 2–3, pp.189–199 ISSN 0160-4120.
  3. ↑ 3.0 3.1 ENVIRONMENT AGENCY. 2005. Attenuation of nitrate in the sub-surface environment. Science Report SC030155/SR2.
  4. ↑ HORVÁTH, L, FÜHRER, E, and LAJTHA, K. 2006. Nitric oxide and nitrous oxide emission from Hungarian forest soils; linked with atmospheric N-deposition. Atmospheric Environment, Vol. 40 pp.7786–7795.
  5. ↑ ENVIRONMENT AGENCY, 2005. Attenuation of nitrate in the sub-surface environment. Science Report SC030155/SR2.
  6. ↑ DREWER, J, LOHILA, A, AURELA, M, LAURILA, T, MINKKINEN, K, PENTTILÄ, T, DINSMORE, K J, MCKENZIE, R M, HELFTER, C, FLECHARD, C, SUTTON, M A, and SKIBA, U M. 2010. Comparison of greenhouse gas fluxes and nitrogen budgets from an ombotrophic bog in Scotland and a minerotrophic sedge fen in Finland. European Journal of Soil Science, 10, Vol. 61, No. 5, pp.640-650 ISSN 13510754. DOI 10.1111/j.1365-2389.2010.01267.x.
  7. ↑ LIU, L, and GREAVER, T L. 2009. A review of nitrogen enrichment effects on three biogenic GHGs: the CO2 sink may be largely offset by stimulated N2O and CH4 emission. Ecology Letters, OCT, Vol. 12, No. 10, pp.1103–1117 ISSN 1461-023X. DOI 10.1111/j.1461-0248.2009.01351.x.
  8. ↑ MOSEMAN-VALTIERRA, S, GONZALEZ, R, KROEGER, K D, TANG, J, CHAO, W C, CRUSIUS, J, BRATTON, J, GREEN, A, and SHELTON, J. 2011. Short-term nitrogen additions can shift a coastal wetland from a sink to a source of N2O. Atmospheric Environment, 8, Vol. 45, No. 26, pp.4390-4397 ISSN 1352–2310.
  9. ↑ BRITISH GEOLOGICAL SURVEY, 1996. Identification and quantification of groundwater nitrate pollution from non-agricultural sources. R&D technical report P32.
  10. ↑ 10.00 10.01 10.02 10.03 10.04 10.05 10.06 10.07 10.08 10.09 10.10 10.11 10.12 10.13 RoTAP. 2012. Review of transboundary Air Pollution: Acidification, Eutrophication, Ground Level Ozone and Heavy Metals in the UK. Contract Report to the Department for Environment, Food and Rural Affairs. Center for Ecology & Hydrology.
  11. ↑ 11.0 11.1 11.2 11.3 11.4 11.5 HELLIWELL, R C, COULL, M C, DAVIES, J J L, EVANS, C D, NORRIS, D, FERRIER, R C, JENKINS, A, and REYNOLDS, B. 2007. The role of catchment characteristics in determining surface water nitrogen in four upland regions in the UK. Hydrology & Earth System Sciences, 01, Vol. 11, No. 1, pp.356–371 ISSN 10275606.
  12. ↑ 12.0 12.1 BOBBINK, R, HORNUNG, M, and ROELOFS, J G M. 1998. The effects of air-borne nitrogen pollutants on species diversity in natural and semi-natural European vegetation. Journal of Ecology, Vol. 86, No. 5, pp.717–738 ISSN 1365-2745. DOI 10.1046/j.1365-2745.1998.8650717.x. Cite error: Invalid <ref> tag; name "Bobbink 1998" defined multiple times with different content
  13. ↑ 13.0 13.1 MASKELL, L C, SMART, S M, BULLOCK, J M, THOMPSON, K, and STEVENS, C J. 2010. Nitrogen deposition causes widespread loss of species richness in British habitats. Global Change Biology. 16, p.671–679.
  14. ↑ DUPRE, C, STEVENS, C L, RANKE, T, BLEEKERS, A, PEPPLER-LISBACH, C, GOWING, D J G, DISE, N B, DORLAND, E, BOBBINK, R, and DIEKMANN, M. 2010. Changes in species richness and composition in European acidic grasslands over the past 70 years: the contribution of cumulative atmospheric nitrogen deposition. Global Change Biology 16, 344–357. https://onlinelibrary.wiley.com/doi/10.1111/j.1365-2486.2009.01982.x/pdf
  15. ↑ 15.0 15.1 STEVENS, C J, SMART, S M, HENRYS, P, MASKELL, L C, WALKER, K J, PRESTON, C D, CROWE, A, ROWE, E, GOWING, D J, and EMMETT, B A. 2011. Collation of evidence of nitrogen impacts on vegetation in relation to UK biodiversity objectives. JNCC Report, No. 447. https://jncc.defra.gov.uk/pdf/jncc447_web.pdf
  16. ↑ JNCC, 2004. Common Standards Monitoring. Introduction to the Guidance Manual. https://jncc.defra.gov.uk/pdf/CSM_introduction.pdf
  17. ↑ WILLIAMS, J M, ED. 2006. Common Standards Monitoring for Designated Sites: First Six Year Report. Peterborough, JNCC.
  18. ↑ 18.0 18.1 18.2 EMMET, B A, ROWE, E C, STEVENS, C J, GOWING, D J, HENRYS, P A, MASKELL, L C, and SMART, S M. 2011. Interpretation of evidence of nitrogen impacts on vitiation in relation to UK biodiversity objectives. JNCC Report No. 449.
  19. ↑ 19.0 19.1 WEISNER, S E B, ERIKSSON, P G, GRANELI, W, and LEONARDSON, L. 1994. Influence of macrophytes on nitrate removal in wetlands. Ambio 23:363–366.
  20. ↑ SAUNDERS, D L, and KALFF, J. 2001. Nitrogen retention in wetlands, lakes and rivers. Hydrobiologia. 442. pp.205–212.
  21. ↑ WOODS HOLE GROUP. 2007. Natural attenuation of nitrogen in wetlands and water bodies. For Massachusetts Department of Environmental Protection. https://webmail.nerc.ac.uk/eea/docs/dep/water/resources/a-thru-m/,DanaInfo=www.mass.gov+attenufr.pdf
  22. ↑ CHAPMAN, P J, EDWARDS, A C, and CRESER, M S. 2001. The nitrogen composition of streams in upland Scotland: some regional and seasonal differences. Science of the Total Environment. 265, pp.65–83.
  23. ↑ DAVIES, J J L, JENKINS, A, MONTEITH, D T, EVANS, C D, and COPPER, D M. 2005. Trends in surface water chemistry of acidified UK freshwaters, 1988–2002. Environmental Pollution. 137, p.27–39.
  24. ↑ JANSSON, A, FOLKE, C, and LANGAAS, S. 1998. Quantifying the nitrogen retention capacity of natural wetlands in the large-scale drainage basin of the Baltic Sea. Landscape Ecology, AUG, Vol. 13, No. 4, pp.249–262 ISSN 0921-2973. DOI 10.1023/A:1008020506036.
  25. ↑ SMITH, R I, FOWLER, D, SUTTON, M A, FLECHARD, C, and COYLE, M. 2000. Regional estimation of pollutant gas deposition in the uk: model description, sensitivity analysis and outputs. Atmospheric Environment, 34, 3757–3777.
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