Nutrient export fluxes in coastal systems, primarily as nitrogen (N), phosphorus (P) and silicon (Si), have a significant impact on water quality and control the nature and magnitude of coastal productivity. In coastal areas, nutrients are delivered by rivers, groundwater discharge and atmospheric deposition. The growing impact of anthropogenic activities has profoundly affected the quality of marine waters over the last 50 years. Such alterations are well documented and have been linked to perturbations in nutrient export fluxes from the continent . In areas of restricted water exchange, the export of excess N and P to coastal waters may cause coastal eutrophication, a blooming of suspended and bed-anchored algae (including toxic species), alteration of community structures, degradation in the ecosystem function and modifications of marine food webs  . The continuing changes in land use and global urbanisation of coastal margins thus pose a continual threat to coastal waters.
CONTINENTAL NUTRIENT SOURCES
On a global scale, riverine inputs of N and P to coastal seas have possibly increased by factors of 2 to 3 , , . Agriculture, in the form of fertilizers, leachates and animal wastes, is the largest contributor of N and P in aquatic systems . Other major inputs include point-source discharges of wastewater from urban sewer networks , and industrial wastes. The direct discharge of P exchanged with soils and sediments also contributes significantly to the budget of this element.
Riverine Si fluxes, originating predominantly from weathering, have generally been altered little by human activity . However, human management of rivers has, in some cases, altered the Si fluxes extensively , often leading to a reduction in diatom blooms as a result of damming.
The direct discharge of groundwater into the ocean, termed submarine groundwater discharge (SGD), has been recently recognized as an additional pathway of nutrients from the land to coastal waters , . On a global scale, SGD rates vary between 0.01-10 % of river runoff . However, the concentrations of nutrients in groundwater are typically higher than those in coastal and river waters , , , , . Therefore, in terms of fluxes, such high concentrations can compensate for the relatively low SGD rates. At the local scale, SGD of nutrients is a prominent transport pathway, particularly in enclosed bays, karstic and fractured systems (e.g., Hawaii ), or at locations where rivers are small or non-existent (e.g., Yucatan peninsula ).
Atmospheric deposition is a significant source of N compounds to the coastal zone, particularly in summer and autumn, but is only a minor source of Si and P , , . Nitrogen delivered by the atmospheric pathway can be either in the oxidized or reduced form . For instance, atmospheric deposition amounts to 30% of the total land based nitrogen input to the North Sea, mainly as oxidized N, and 50% to the Baltic Sea . The N:Si:P ratio for wet deposition in the North Sea is 503:2:1 .
Nutrients are significantly altered by biogeochemical processes during their transport along the land-ocean transition zone, especially in estuarine systems. Figure 1 summarizes the major N sources and transformation processes in an estuary. Estuaries are usually turbid, and hence primary production is often limited by light availability. Light conditions generally improve towards the coastal zone and primary production becomes a dominant process in controlling the biogeochemical cycles of nutrients .
Figure 1. Important sources and sinks of N in an estuary (source: Tappin, 2002) .
N species in aquatic environments include dissolved (nitrate, nitrite, ammonium, organic N) and particulate (organic N) constituents . The removal of N occurs by deposition and permanent burial in sediments and, most importantly, loss to the atmosphere by bacterial denitrification. This process is coupled to organic matter decomposition and reduces nitrate to gaseous N2/N2O under anoxic conditions. Part of the nitrate pool originates from coupled nitrification/denirification, in which the ammonium produced from organic matter degradation is first oxidized to nitrate, and subsequently denitrified . In temperate and tropical estuaries the estimated loss of nitrate N via denitrification varies widely, and also varies in time and space within estuaries , . Because denitrification requires low oxygen concentrations, this process is particularly important in muddy sediments , . It is also quantitatively more important in ecosystems characterized by relatively long residence times . In groundwater systems, the nitrate supplied either by infiltrating water or produced through nitrification , is also commonly removed through denitrification. As in surface estuaries, a set of conditions, namely the presence of labile organic matter, a low redox potential and sufficient time for reaction, are prerequisite for effective denitrification to occur. However, field studies often report only limited nitrate removal prior to discharge to coastal waters primarily due to a lack of labile dissolved organic matter , , ), as is the case in many shallow groundwater aquifers or sandy nearshore sediments, or due to high groundwater velocities , .
P species in aquatic systems include dissolved (inorganic, organic P) and particulate (inorganic, organic P) constituents . The retention of P in land-ocean transition zone is often attributed to adsorption on solid particles, which are constantly trapped in estuarine sediments , or form part of the solid matrix in coastal aquifers. However, in the case of very large rivers that discharge directly in the continental shelf, P retention in the mixing zones between freshwater and seawater will be limited . Adsorption onto solids such as iron and aluminum oxides is particularly effective , , and thus may be also coupled to the redox conditions . For instance, removal of P is very efficient in subterranean estuaries characterized by zones of iron oxide accumulation, (“Iron Curtains” , ). The behavior of P in estuarine systems is also influenced by the strong physico-chemical gradients, which result from the variations in pH, ionic strength and solution composition between the freshwater and seawater end-members (e.g. , , ). The removal of P can occur through bacterial reduction of phosphate to gaseous phosphine. However, little is known on the rate of phosphate-phosphine transformation and its contribution to overall P cycling , .
Tidal and marginal sediments are considered important sinks of N and P, although a quantitative estimation of their role remains uncertain , , . On the global scale, it is generally accepted that intertidal sediments are more efficient for P burial than for N , .
Relevant Si species in the aquatic environments include dissolved Si (DSi), mainly as undissociated monomeric silicic acid, Si(OH)4, and particulate Si (biogeneic silica, BSiO2), which includes the amorphous silica in both living biomass and biogenic detritus in surface waters, soils and sediments. The main transformation processes are the uptake of DSi and the biomineralisation as BSiO2 in plants and organisms, as well as the dissolution of BSiO2 back to DSi. Over sufficiently long time scales, BSiO2 may undergo significant chemical and mineralogical changes , even including a complete diagenetic transformation of the opaline silica into alumino-silicate minerals .
The major producers of BSiO2 in marine environments are diatoms. However, other organisms such as radiolarians, sponges and chrysophytes may be important local sources of BSiO2 . Large quantities of DSi are also fixed on land by higher plants, forming amorphous silica deposits, known as phytoliths . Their role in the Si cycle has only recently been studied , . In general, riverine Si fluxes have been much less altered by human activity than those of N and P. However, increased damming of major rivers has promoted siliceous phytoplankton blooms , , and therefore, reduced Si fluxes to the coastal zone. For example, the damming of the Danube has reduced the DSi concentration by more than 50 % .
Nutrient budgets and fluxes have been established at the local (major rivers) and regional (coastal seas) scales across Europe. Two examples are provided below: (1) a N budget of the continental inputs to the North Sea and (2) a N budget of major riverine inputs and transformations along the Western Scheldt river-estuarine system . Other nutrient budgets have been established, among others, for the Western shelf of the Black Sea and the Baltic Sea . At a smaller scale, detailed estimates of the nutrient sources, transport and transformations are also available for the Seine and Humber continuums , .
The continental inputs of nitrogen to the North Sea originate from rivers, atmospheric inputs and, to a much smaller extent, direct discharges and dumping . The riverine contributions are summarized in Table 1 and, collectively, amount to almost twice that of atmospheric inputs . Figure 2 shows the spatial distribution of the total atmospheric N deposition to the North Sea. The spatial pattern results from the distribution of the source areas and precipitation rates. On average, deposition amounts to 0.9 ton N per km2, with deposition up to 50% higher than average around territorial waters of Belgium, the Netherlands and Germany. Approximately 60% of total atmospheric N deposition results from combustion (nitrogen oxides) and approximately 40% from agricultural activities (ammonia) .
Table 1: Annual river inputs (ton per year) of nitrogen for all relevant rivers around the North Sea (source: Radach and Lenhart, 1995) . P and S fluxes are also shown River N P Si Firth of Forth 20 186 11 Tyne/Tees 14735 593 9309 Humber 60636 5891 17928 Thames 26214 3786 14931 Ems 25736 614 6805 Noordzeekanaal 10877 1767 3912 Lauwer 333 143 25 Lake IJssel/Kornwerderzand 12320 461 3588 Lake IJssel/Den Oever 21232 80 5170 Meuse 91159 4400 34402 Rhine 191543 14194 69623 Scheldt 31670 2116 15077 Yzer 267 109 37 Elbe 126314 3822 34520 Jade 8 3 2 Schleswig-Holstein river 8 3 2 Weser 52862 3420 18470 Danish rivers 1227 513 136
Figure 2. Total atmospheric nitrogen deposition to North Sea, 1999, in ton N per km2(source: Hertel et al, 2002)
The Western Scheldt Estuary
The Scheldt River and its tributaries drain 21,580 km2 in northwestern France, northern Belgium and southwestern Netherlands . The Scheldt estuary is a a macrotidal system, with an average residence time in brackish waters of 1 to 3 months. The mixing zone of fresh and salt waters extends over a distance of 70 to 100 km. The area of tidal influence goes up to 160 km from the river mouth and includes the major .
The hydrographical basin includes one of the most heavily populated regions of Europe, where highly diversified industrial activity has developed. As a consequence, the whole catchment was heavily polluted until the mid 1970s, when water degradation culminated due to the continuous increase of nutrient and organic mater inputs. The level of wastewater treatment, especially in the upstream zones, was an important factor contributing to this degradation. The estuary was particularly affected by domestic and industrial inputs from the great Brussels, Antwerp and Gent areas . Since then, better management of industrial and domestic wastewater point sources has led to a progressive improvement of the environmental conditions in the estuary. Billen et al. (2005) and Soetaert et al. (2006) provide two recent comprehensive reviews of this long term evolution.
A mass budget for nitrogen has been established for the saline estuary (km 0 to100) and for the tidal river network (km 100 to 160) of the Western Scheldt for the summer months . Three periods have been analyzed (1990, 2002 and 2010). This allows for the assessment of the influence the secondary and tertiary wastewater treatment in the catchment on the N dynamics. Figure 3 shows that the tidal river and the estuary contribute almost equally to the overall biogeochemical cycling of N, despite the very different volumes involved. For the simulated periods, the large decrease in N input (> 55 %) expected between 1990 and 2010 will not lead to a significant decrease of N export to the coastal zone during the summer period.
ADD FIGS HERE Figure 3. Mass budget for (A) ammonium and (B) nitrate in the tidal rivers (right) and in the saline estuary of the Western Scheldt (left) in the summer of 1990, 2002 and 2010. Processes: resp=aerobic respiration; nitrif=nitrification; denit=denitrification; npp=net primary production. Transport fluxes are positive seawards. All fluxes are given in kmol day-1. Top arrow: Riverine and lateral inputs; Left arrow: export to the coastal zone. (source: Vanderborght et al., 2007)
- Vanderborght, J-P, I. Folmer, D. Rodriguez Aguilera, T. Uhrenholt, and P. Regnier (2007), Reactive-transport modelling of a river-estuarine coastal zone system: application to the Western Scheldt, Marine Chemistry 106, 92-110.
- Garnier, J., G. Billen, E. Hannon, S. Fonbonna, Y. Videnia, and M. Soulie (2002), Modelling the transfer and retention of nutrients in the drainage network of the Danube river, Estuarine, Coastal and Shelf Science, 54, 285-308.
- Tappin, A.D. (2002), An Examination of the Fluxes of Nitrogen and Phosphorus in Temperate and Tropical Estuaries: Current Estimates and Uncertainties, Estuarine, Coastal and Shelf Science 55, 885-901.
Will list all references (currently in footnotes)
- [ELOISE Nutrient Dynamics in European Water Systems ONLINE]
- [ELOISE Nutrient Dynamics in European Water Systems in pdf format]
- [Case studies]
- [LOICZ Land-Ocean Interactions in the Coastal Zone]
C. Spiteri and P. Regnier Department of Earth Sciences Faculty of Geosciences Utrecht University, P.O. Box 80.021 3508 TA Utrecht, The Netherlands
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Please note that others may also have edited the contents of this article.