eutrophication of the seas along sweden's west coast...eutrophication of the seas along...

Eutrophication of the Seas along Sweden’s

West Coast

report 5898 • novEmbEr 2008

rapport 5507 • novEmbEr 2006

Effective July 1, 2011, this publicationis handled by the Swedish Agency for Marine and Water Management.Telephone +46 (0)10 698 60 [email protected]/publications

Eutrophication of the Seas along Sweden’s West Coast

Report to the Swedish Environmental Protection Agency (Naturvårdsverket)

Panel for the Expert Evaluation of Eutrophication in the Western Swedish Seas Dr. Donald F. Boesch, Chair University of Maryland Center for Environmental Science, Cambridge Maryland, USA Dr Jacob Carstensen National Environmental Research Institute, Aarhus University, Roskilde, Denmark Dr. Hans W. Paerl Institute of Marine Sciences, University of North Carolina, Morehead City North Carolina, USA Dr. Hein Rune Skjoldal Institute of Marine Research, Bergen, Norway Dr. Maren Voss Leibniz Institute for Baltic Sea Research, Warnemünde, Germany

November 10, 2008

SWEDISH ENVIRONMENTAL PROTECTION AGENCY

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ISBN 978-91-620-5898-2.pdf ISSN 0282-7298

© Naturvårdsverket 2008

Digital publication

S W E D I S H E N V I R O N M E N T A L P R O T E C T I O N A G E N C Y R e p o r t 5 8 9 8 • W e s t C o a s t E u t r o p h i c a t i o n

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Contents 1 INTRODUCTION 5

2 PHYSICAL SETTING 7 2.1 Geography and Bathymetry 7 2.2 Circulation and Water Masses 8 2.3 Swedish Coastal Waters 12

3 NUTRIENT SOURCES AND TRENDS 13 3.1 Sources 13

3.1.1 Contributions of countries, the atmosphere, and the North and Baltic seas 13 3.1.2 Contributions from the Jutland Coastal Current 14 3.1.3 Point sources and atmospheric deposition 17 3.1.4 Trends in source inputs 19

3.2. Nutrient Status and Trends in Coastal Waters 20 3.2.1 Concentrations and dynamics 20 3.2.2 Trends 22 3.3.3 Budget aspects 24

4 ECOSYSTEM RESPONSES 26 4.1 Phytoplankton Production 26

4.1.1 Phytoplankton 26 4.1.2 Nutrient limitation 27 4.1.3 Climatic factors 30 4.1.4 Why N2 fixation does not compensate for N limitation 31

4.2. Macrophytes 34 4.3 Dissolved Oxygen 35

4.3.1 Status and trends 37 4.3.2 Organic matter supplies and metabolism 38

4.4. Benthos of Sediment Bottoms 40

5. REVERSING EUTROPHICATION 43 5.1. Effects of Countermeasures Taken 43

5.1.1. Swedish sources 43 5.1.2. Transboundary sources 45

5.2. Responses to Reductions in Nutrient Inputs 46 5.2.1 Nutrient concentrations and ratios 46


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5.2.2 Phytoplankton 47 5.2.3 Phytobenthos 48 5.2.4 Dissolved Oxygen 48

5.3. Other Significant Drivers Affecting Responses 49 5.3.1. Climate variability and change 49 5.3.2. Degraded state of the ecosystem 50

6. EVALUATION OF THE SWEDISH STRATEGY 52 6.1. The Objective of “Zero Eutrophication” 52

6.1.1. Interim targets and goals 52 6.1.2. Specific goals and strategies for west coast marine waters 54 6.1.3. The transgenerational reality 55 6.1.4. Climate change and other compounding forces 55

6.2. Measures and Their Implementation 56 6.2.1. Nitrogen controls are essential 56 6.2.2. Phosphorus reductions produce local benefits 57 6.2.3. Greater reductions of agricultural and atmospheric loads are needed 57 6.2.4. Multi-national cooperation is required 58

6.3. Integration of Monitoring, Modeling and Research for Adaptive Management 58 6.4. Transparency and Accountability 59

7. FINDINGS AND RECOMMENDATIONS 61

REFERENCES 64


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1 Introduction One of the most serious and challenging environmental problems facing Sweden is eutrophication of its surrounding seas as a result of excessive human emissions of plant nutrients. In 1999 the Swedish parliament (Riksdag) set fifteen environmental quality objectives for the nation, including the objective of Ingen Övergödning, translated as “Zero Eutrophication,” but more literally “No Over-Enrichment”. Specifically, the objective is: “Nutrient levels in soil and water must not be such that they adversely affect human health, the conditions of biological diversity or the possibility of varied uses of land and water.” It is further specified that: “The intention is for this environmental quality objective to be achieved within one generation.” In 2001, the Riksdag established interim targets, strategies and measures to facilitate reaching the national environmental quality objectives, which were revised in 2005. The Swedish Environmental Protection Agency (SEPA 2007) recently conducted a second in-depth evaluation of the Zero Eutrophication environmental quality objective, including progress in achieving the interim targets.

As part of its continued efforts to assess the state of eutrophication and progress toward its alleviation, the SEPA convened this international expert evaluation of eutrophication in the seas and coastal environments along the west coast of Sweden. It follows an earlier expert evaluation of eutrophication in all Swedish seas, which, while briefly addressing the western seas, focused largely on the Baltic Sea and its coastal environments (Boesch et al. 2006). That evaluation concentrated on the controversies regarding the controls of nitrogen versus phosphorus emissions. The SEPA used that report to develop standpoints to guide its actions to combat eutrophication (SEPA 2006).

This present expert evaluation was charged to evaluate the measures taken so far to achieve the Zero Eutrophication objective for the Danish Sounds, the Kattegat and the Skagerrak and the Swedish coastal environments bordering these waters and to recommend future strategies to counteract eutrophication there. These western seas have important differences from the Baltic Sea, including higher salinity and the influence of tides and the dynamic forces of the North Sea. As in the Baltic Sea, considerations have to be given to the sources and transport processes affecting nutrient delivery into these international seas, including from the Baltic and North Seas.

An expert panel was assembled by the SEPA to perform the evaluation. It consisted of five members, including one each from the neighboring countries of Denmark and Norway. Dr. Donald Boesch of the United States was invited by SEPA to chair the panel. The panel met from 8-13 August, 2008, in Marstrand, an island in the coastal archipelago along Sweden’s southern Skaggerak coast. Dr. Per Jonsson was the SEPA coordinator and Mats Blomqvist assisted the panel in accessing data and information and producing graphics. On 10 August, several Swedish experts met


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with the panel, presenting their recent findings and participating in discussion of issues before the panel. These included Drs. Suzanne Baden, Odd Lindahl, Leif Pihl, Johan Rodhe, and Rutger Rosenberg of Gothenburg University and Dr. Daniel Conley of Lund University. In addition to this consultation, the panel reviewed the findings of more than 100 scientific papers and reports, including very recent publications and national and regional assessments. A draft report was prepared while the panel worked at Marstrand and subsequently refined through correspondence.

The expert panel specifically considered: the status and sources of anthropogenic emissions of nutrients, including trans-boundary sources; the extent of eutrophication and the nutrients responsible for it; the effects of eutrophication on the ecosystem and natural resources; the confounding influence of other factors such as climate variability and change and fishing activities; the effectiveness of the present Swedish strategy to counteract eutrophication and prognosis for the future; and the adequacy of scientific research, monitoring and assessment to support its execution.


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2 Physical Setting 2.1 Geography and Bathymetry The Swedish west coast faces the Öresund, Kattegat and Skagerrak (Figure 1). These are three very different although connected water bodies. The Skagerrak is part of the deep Norwegian Trench which is a glacially excavated valley that runs along the coast of Norway and connects the North Sea with the deep Norwegian Sea to the north. The Skagerrak is about 700 m deep in the inner (eastern) part and there is a steep slope from the Swedish Bohuslän coast down into the deep Skagerrak. The Kattegat and Öresund in contrast are shallow sea areas that connect the Skagerrak with the Baltic Sea.

The Kattegat is a broad basin about 200 km long and 100 km wide. The boundary between Kattegat and Skagerrak is usually taken as a line from Skagen (north tip of Denmark) to the city of Göteborg on the Swedish coast. To the south, the Kattegat is connected with the Baltic Sea through the narrow strait of Öresund between Sweden and the Danish island of Zealand and through the belts around the island of Funen. This latter connection is broader and topo-graphically more complex, connecting through Samsø Belt, Little and Great Belts, and Fehmarn Belt, and finally across Darss Sill, with the Arkona Basin as the westernmost part of the Baltic Sea. The connection through Öresund has a sill depth of only about 8 m, while the connection through the belts is deeper at about 15 m at Darss Sill (Gustafsson 2006).

Figure 2.1. Seas along the Swedish west coast.

The western part of Kattegat is mostly shallow (<20 m deep), with the islands Læsø in north and Anholt in the south-central part. A deeper depression or trench cuts in from Skagerrak on the eastern side with depth >60 m to southeast of Læsø. The central Kattegat (between Læsø and Anholt) has a rugged topography with shallow (<20 m) areas also on the eastern side (e.g. Fladen Grund) except for a narrow, deeper trench running close to the Swedish coast. The southern Kattegat


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(south of Anholt) is mostly moderately shallow (20-40 m), while Laholm Bay on the Swedish side is particularly shallow (<20 m).

2.2 Circulation and Water Masses The North Sea circulation is basically counter-clockwise (Figure 2.2; Otto et al. 1990). Atlantic water from the inflow across the ridge between Scotland and Iceland flows into the North Sea across the northern boundary between Scotland and Norway. A part of this flow comes over the northern North Sea plateau while the rest flows in along the western slope of the Norwegian Trench. The inflow of Atlantic water over the plateau in the northwestern North Sea continues south with portions being deflected east by shoaling topography in the central North Sea (the Dooley current) and by the Dogger Bank in the southern North Sea. A portion of this water may also flow south and around the Dogger Bank. Much of the inflow in the Norwegian Trench continues into the Skagerrak where it circulates around and leaves on the northern side along the Norwegian Skagerrak coast. There is also some inflow of Atlantic water through the (English) Channel that continues northeast along the European continent south of the Dogger Bank. This flow is the main seawater that receives the input of fresh water from the large European rivers including the Seine, Scheldt, Rhine-Meuse, Weser, and Elbe. The fresh water lowers the salinity and gives the flow a distinct characteristic as a coastal current that flows north as the Jutland Current along the western and northern coasts of Denmark. The flow of Atlantic water into and through the North Sea is typically of the order 2 Sverdrup (1 Sv = 106 m3 s-1). There is a large seasonal variation (by a factor of 3-5 for flows through various parts of the North Sea), related to the general intensification of winds in the winter and calmer conditions during summer and also large interannual variability. Thus, the circulation may be particularly great

Figure 2.2. General circulation in the North Sea (OSPAR Commission 2000).


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during winters with a high North Atlantic Oscillation Index (NAO), with many passages of low pressures and strong southwesterly and westerly winds at the entrance region to the North Sea. Roughly half of the total flow through the North Sea circulates through the Skagerrak (Skjoldal 2007). A time series of modeled flux of water through the Skagerrak for the period 1955-2006 is shown in Figure 2.3. The Skagerrak experiences the confluence for four different water masses: outflowing Kattegat surface water (KSW, which contains the Baltic outflow), water from the Jutland Coastal Current (Jutland coastal water), water from the central North Sea (CNSW), and Atlantic water (AW). These water masses have different salinities and densities and when they meet in inner Skagerrak they can be layered one above the other. The least dense water is the Kattegat surface water, with an average salinity around 25 as it leaves Kattegat (Figure 2.4). The next less dense is Jutland coastal water which has salinities typically around 32-33 as it passes off Skagen. The Central North Sea water contains some fresh water mixed in from the coastal zones and typically has salinities between 34.5 and 35, while the Atlantic water has salinities >35. In the inner Skagerrak these water masses are typically stacked above each other, although there can be short-term and spatial variation in this pattern. The outflow from Kattegat continues north along the Swedish Bohuslän coast, overlying Jutland water, central North Sea and Atlantic water masses. Through mixing and entrainment, this buoyant coastal current increases its salinity as it continues north into the wide bight of the outer Oslofjord, where it is deflected and flows as the Norwegian Coastal Current along the Norwegian Skagerrak coast and then farther north along the Norwegian west coast. It has been shown that by the time the current passes Arendal, about half way along the Norwegian Skagerrak coast, most of the Jutland water can be accounted for as being present in the deeper part of the upper 30 m of the water column (Skjoldal et al. 1997, Aure et al. 1998). The Jutland Current is almost always present along the west coast of Denmark, although it can be temporarily halted or reversed by strong northerly winds. This

Figure 2.3. Modelled flux of water through Skagerrak (across a transect between Oksøy in Norway and Hanstholm in Denmark). Time series are for mean flux for the 1st and 4th quarters of the year from 1955 to 2004. The unit of the flux is Sverdrup (106 m3 s-1) and the negative sign indicates flux into Skagerrak. Data obtained with the NORWECOM model (Skogen and Søiland 1998) driven by archived meteorological data for the modeled time period.


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leads to an accumulation of water in the German Bight and a stronger flow as the winds slackens or shifts from the south. Thus, the flow into Skagerrak can have a pulsed character leading to spatial variation in the amount and thickness of the submerged Jutland water. The circulation in the Kattegat is typical estuarine with low-salinity water flowing out from the Baltic Sea as an upper layer, while saltier water flows south as a deeper layer. A large fraction of the deeper water is gradually entrained into the upper layer as it penetrates south through the Kattegat. The upper and deeper layer is usually separated by a pronounced density gradient, or pycnocline. The salinity of the Baltic water in the Arkona Basin is about 8 on average as it approaches the Belts and Öresund. Salinity increases to about 20 in the southern Kattegat. On the further passage north through Kattegat, the salinity of the upper layer increases to an average of about 25 south of Læsø. This corresponds to an entrainment of an amount of water about twice the Baltic outflow, resulting in an increase in the net volume flow by a factor of about 3. In the frontal area north of Læsø, where high salinity Skagerrak water is subducted as a bottom current, surface salinities can rapidly change 5 to 10 due to the mixing of different water masses.

The Baltic outflow is driven by the net freshwater input to the Baltic which is about 16,000 m3 s-1. When the Baltic outflow leaves the northern Kattegat at a salinity of 25 it has increased to a mean flow of about 60,000 m3 s-1 (0.06 Sv) due to

Figure 2.4. Annual (1998) mean of salinity and currents in the surface 5 m for the Skagerrak-northern North Sea region as modeled by Albretsen (2007).


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admixture and entrainment of saltier water. About 25 % of the outflow from the Baltic occurs through Öresund while the largest amount (about 75 %) exits through the Belts. Roughly one-quarter of the entrainment occurs in the passage through the Danish Belts, while the remaining takes place in the Kattegat. Superimposed on the net mean flow through Kattegat there is pronounced short-term variability. The instantaneous flow between the Baltic and Kattegat can be on the order of 0.1 Sv, driven by air pressure differences and effects of winds that lead to sea level changes both in the Baltic and in Kattegat. Thus, the outflow from the Baltic can occur as stronger pulses of 1-10 days duration, interspersed with periods of low or reversed flow in the direction from Kattegat into the Baltic Sea. The higher flow rates during periods of intensified outflows from the Baltic lead to a shallowing and strengthening of the pycnocline in southern Kattegat, while slackening or reversal of the flow leads to a deepening and weakening of the pycnocline. The pycnocline is typically located at about 15 m depth (Andersson and Rydberg 1993) and is very strong with a change in salinity of 5-15 units between the upper and lower layer. The source of the deep water that flows south in Kattegat and which is subsequently entrained into the outflowing Baltic water, is water from the North Sea circulation. The average salinity of the inflowing water at 40 m depth in the northern Kattegat is 33.9, decreasing to 33.3 at 40 m in the southern Kattegat (Gustafsson 2000). The water with salinity of about 34 is typically a mixture of Jutland coastal water and water from the central North Sea. Hydrographical data (including nutrients) shows that Jutland water is regularly present as an inter-mediate water layer below the pycnocline in Kattegat, with somewhat saltier water below. The magnitude of the Jutland Coastal Current is around 0.1 Sv as an annual average, based on the freshwater input to the southeastern North Sea (4.5 103 m3 s-1) diluted out to a salinity of 33 (Skjoldal 1993). Due to the seasonality in freshwater input and prevailing wind conditions, the Jutland Current is more voluminous in winter, with a flow of order 0.15 Sv. The inflowing deep water in the Kattegat is about 0.04 Sv as an annual average (to balance the outflow of 0.06 Sv, with about 0.02 Sv coming from the Baltic Sea). Thus, only a fraction of the Jutland coastal water circulates through the Kattegat, the majority being deflected north along the Swedish Bohuslän coast, underlying the outflowing layer of Kattegat surface water. The average residence time of water in Kattegat is typically 2-3 months if calculated on the basis of flushing time [the time needed for the net flow of 0.06 Sv to replace the volume of water in Kattegat (0.5 1012 m3)]. More detailed information on residence time for different parts of Kattegat and the Belts is presented by Gustafsson (2000). The residence time for the outflowing surface layer in Kattegat is typically about 1 month, while the inflowing deep layer can have residence time of several months. The residence time varies with local metorological conditions, both seasonally and inter-annually.


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The buoyant coastal current that flows north along the Swedish Bohuslän coast consists of the outflowing Kattegat surface water overlying Jutland water, floating on top of the central North Sea and Atlantic waters that circulate through Skagerrak (Rohde 1996). This deeper circulation is strong, typically of order 0.5-1 Sv, while the coastal current may be of order 0.15 Sv (0.06 Sv Kattegat outflow and 0.1 Sv Jutland Current). The transport time for the coastal current to flow along the Bohuslän coast is typically a few weeks (assuming a net current speed of 20 cm s-1). The average salinity of the surface layer of the coastal current is around 28 in northern Bohuslän (Koster), showing admixture of some of the underlying Jutland water into the surface layer. 2.3 Swedish Coastal Waters The Swedish west coast is characterized mostly by scattered skerries and small bays and fjords that communicate openly and effectively with the offshore waters both in Kattegat and in particular in Skagerrak. The coastal waters (defined as waters within the baseline) have been divided into regions based on typology according to the methodology given by the EU Water Framework Directive. The typology is based mainly on water salinity, exchange and residence time, and bottom substrate. The areas are from south to north: the coast along the Öresund, the coast along southern Kattegat including Skälderviken and Laholm Bay, the coast along northern Kattegat, the coast along Skagerrak, and the fjord systems north of Gøteborg including Havstensfjord and Gullmarfjord. In addition, the inner coastal water in many areas has been identified as a separate water type, being less exposed as habitats than the outer skerries and coast. The residence times of bottom water within the different regions are mostly of the order of some days (<9 days), except for the fjords where it is typically >40 days. The inner coastal water also typically exhibits short residence time (<9 days), although it can be somewhat longer (10-39 days) in some areas (SEPA 2008b). The openness of the coastal areas and short residence times of water within them mean that their general conditions are determined by the physicochemical properties of the offshore waters. Exceptions to this are some of the west coast fjords where narrow and shallow sills may limit the water exchange. This is particularly the case in Koljöfjord, which has shallow sills, but Gullmarsfjord and Havstensfjord also have relatively shallow sills that reduce the rate of renewal of the bottom water, e.g. for the Gullmar Fjord a mean residence time for the water below the sill of one month is reported (Lindahl 1989). This renewal takes place mainly in the winter period when the water outside the sill is coldest and relatively high salinities occur because freshwater discharge is low and mixing is high in the coastal water bodies. The longer residence time of deeper water makes these fjords more susceptible to local influence. Nevertheless, the fjords are also influenced from the outside in that the water above the sill may be effectively exchanged and the fjords act as sedimentation basins for fall-out from the production and organic load of the euphotic zone.


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3 Nutrient Sources and Trends 3.1 Sources 3.1.1 Contributions of countries, the atmosphere, and the North and Baltic seas The most important human sources of the nutrients responsible for eutrophication, nitrogen (N) and phosphorus (P) for the Kattegat and Skagerrak region are from land discharges and atmospheric deposition. The land-based inputs are driven by the amount of freshwater discharge that varies between 12 and 29 km3 yr-1 (Håkansson 2007). Most freshwater input comes from Sweden but the highest loadings of N and P come from Denmark. The largest single Swedish nutrient source is the River Göta Alv, the sixth largest river draining the greater Baltic Sea catchment; however, most of its load is transported northwards affecting the Skagerrak. Aside from inputs to the North Sea, only a small catchment of Germany drains into the Belt Sea region, with no nutrient discharge directly into the Öresund or Kattegat. From Sweden, the Kattegat receives 20,800 t yr-1 and the Skagerrak 1,800 t yr-1 of diffuse nitrogen loads (Table 3.1). Agricultural land covers only approximately 12% of the catchments, which are 55% forested. Of these diffuse loads, 55 % of the nitrogen emanates from agricultural land and 15% from N-deposition on lakes and other inland open waters that drain to the coast (i.e. indirect deposition) (Håkansson 2007). When N-retention is considered the overall input into the Kattegat is naturally reduced by 12,000 t yr-1 before it enters the coastal sea. Point sources deliver much less nitrogen to the Kattegat and Skagerrak with only 6,700 t yr-1 and 500 t yr-1, respectively (Table. 3.1). Table 3.1. Gross loads given in t N yr-1 for 2006 and normalized for mean runoff (Håkansson 2007). Diffuse sources Point sources Agricultural

land Forests

and clear cut areas

Open land

Deposition on water

Urban water

Unconnected dwellings and

WWTP Kattegat 20,800 8,100 2,400 5,900 700 6,700 Skagerak 1,800 900 400 100 100 500 Sum 41,400 7,200

Phosphorus gross loads from Sweden are 910 t yr-1 and 180 t yr-1 to Kattegat and Skagerrak and the diffuse input from agricultural land was again by far the largest share with 56% and 66%, respectively. Point sources only contributed 270 t yr-1 and 30 t yr-1. Assuming the retention estimates reliably reflect the natural processes the inputs into the Kattegat are roughly halved due to natural processes.


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From Denmark, the total nitrogen and total phosphorous loads discharged into the Kattegat, Öresund and Belt Seas were 51,800 and 1,500 t yr-1 in 2006 (Ærtebjerg 2007), but they should be considered in the framework of the overall decreasing trends (see below). Diffuse loads make up 60% for phosphorus and about 80-90% in the case of nitrogen; most comes from agricultural activities. Both nitrogen and phosphorus discharges show declining long-term trends with phosphorus loads decreasing more substantially than nitrogen (Carstensen et al. 2006). Nutrient yields (inputs per unit area) are still higher in Denmark than in Sweden. Norway contributes approximately 22,000 t yr-1 of N and 750 t yr-1 of P to the Skagerrak. These loads are mostly of anthropogenic origin 80% of the P and 50% of the N (based on 1993 data, Skjoldal et al. 1997). These nutrients largely enter in the Outer Oslofjord area and they contribute to very minor degree to the nutrient load in Swedish waters. As reviewed earlier, the outflow of water from the Baltic Sea at the surface is compensated with inflowing deep water through the Kattegat, Belt Sea and the Öresund. The exchange of water and nutrients imposes considerable variability in nutrient concentrations depending additionally on large scale climate variations (e.g. the NAO). In a model and data evaluation study Rasmussen and Gustafsson (2003) estimated that net transports were directed from the Baltic Sea towards the Skagerrak with high inter-annual and decadal variability. They also point out that there are decadal changes in these fluxes and that the Kattegat imported inorganic P from the Skagerrak. The exchange of water and nutrients between the Skagerrak and the North Sea is extremely high with an average of 4,300 kt TN yr-1 and over 400 kt TP yr-1, but there is a net export of 179 kt N and 15 kt P from the Skagerrak to the North Sea (Håkansson 2007). These fluxes are difficult to compare to the nutrient input from land but may contribute significantly to the nutrient budget. Atmospheric deposition brings another 40-45 kt N y-1 into the Kattegat/Skagerrak region (Håkansson 2007). A modelling study estimated the long-term mean input and demonstrated high variability in N deposition (Spokes et al. 2006). The mean input was estimated to be 70 mg N m-2 d-1, which is equivalent to a nitrogen concentration of 0.5 μm L-1 when mixed into a 10 m water column. 3.1.2 Contributions from the Jutland Coastal Current The open North Sea is dominated by exchange with the North Atlantic Ocean, but the coastal regions receive large amounts of nutrients from western European rivers. The riverine input of nutrients increased up to the 1980s, particularly for N (as nitrate), resulting in N/P (atomic) ratios of 30-35 for the total annual inputs of N and P to the southeastern North Sea in 1990 (Skjoldal 1993, NSTF 1994). The loadings of nutrients from these rivers have declined as a result of pollution reduction measures, beginning in the 1980s for P and 1990s for N (Figure 3.1). Flow-adjusted P loadings have declined by more than half, while equivalent nitrogen loads have declined by about 20%. As a result, the N:P ratios in the river


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discharges and in coastal waters near the rivers have increased to well above the Redfield ratio of 16 on a molar basis (McQuatters-Gallop et al. 2007; Philippart et al. 2007).

Consequently, while inputs of both nutrients are clearly declining, there remains surplus N to be exported in coastal currents along the coast to the north. As a result, the more distal portions of the shallow Wadden Sea, where N is imported from theses coastal water masses and phosphorus is efficiently recycled from sediments, continues to be affected by eutrophication (van Beusekom et al. 2005). While the nutrients from these riverine sources that potentially reach the Skagerrak and Kattegat region by transport with the Jutland Coastal Current (JCC) have likely declined, the decline in N is probably much less than the decline of P.

Around 1990 the Rhine and Elbe had nitrate concentrations of 500 μM or higher in winter. Inputs from these rivers lead to concentrations in the German Bight often exceeding 40 μM, more than twice the concentrations 20 years earlier (Figure 3.2). The European rivers are particularly enriched with nitrate, leaving an estimated surplus amount of 300,000 tons of nitrogen when phosphorus was depleted by the spring growth of phytoplankton (NSTF 1994). The N-enriched JCC flows into the inner Skagerrak, where, as described in Section 2.2, it submerges under the less-dense water flowing from the Kattegat. Around 1990 it was estimated that the JCC transported an annual amount of about 400,000 tons of nitrogen of anthropogenic origin into Skagerrak (NSTF 1994).

Figure 3.1. Trends in specific N and P loads (mean annual load/mean annual discharge) from the Rhine-Maas and Elbe-Weser rivers (van Beusekom et al. 2005).


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Most of the Jutland coastal water reaching the Skagerrak is advected north along the Swedish west coast and farther along the Norwegian Skagerrak coast. The winter (January-April) nitrate concentrations in the upper 30 m of the Norwegian Coastal Current (NCC) doubled between the 1970s and the early 1990s, apparently as a result of the increased concentrations in the Jutland coastal water (Skjoldal et al. 1997, Aure et al. 1998). This was associated with organic enrichment and lowered oxygen concentration in basins of fjords along the Norwegian Skagerrak coast, reflecting increased sedimentation and oxygen consumption rates (Skjoldal et al. 1997). There was also a declining trend in the oxygen concentrations at intermediate depths and salinities in the NCC in the autumn, starting around 1970 (Johannessen and Dahl 1996). The observed decrease in nitrate concentrations in the German Bight from the 1990s is reflected in the Norwegian Coastal Current where the mean nitrate concentration in the upper 30 m is now reduced roughly half way back to the situation in the 1970s (Aure and Magnusson 2008). It is likely that this same situation, reflecting transport of nutrients with the Jutland Coastal Current, has also affected the Swedish Skagerrak coast and, to some degree, the Kattegat. Some portion of the Jutland coastal water is advected south as an intermediate layer below the pycnocline in Kattegat (Section 2.3). In the southern Kattegat, high nitrate concentrations and high N/P just below the pycnocline in late April in most years suggest the presence of Jutland coastal water (Figure 3.3). This water would be entrained into the outflowing water and enriches the upper layer in spring and early summer with a surplus of nitrogen relative to phosphorus relative to the Redfield ratio. This could result in phosphorus limitation in spring and early summer, with the surplus nitrogen (mainly as nitrate) still used by phytoplankton nourished by recycling of P. Later in summer the situation would

Figure 3.2. Mean nitrate concentrations in January–April at Helgoland in the German Bight (Aure and Magnusson 2008).


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likely change to one with predominant N limitation as nitrogen is depleted and phosphorus is efficiently recycled.

3.1.3 Point sources and atmospheric deposition The nutrient inputs from Denmark and Sweden are dominated by diffuse inputs which come from agricultural land due to loss of nutrients applied in excess of the nutrients removed in crops and animal products. By the late 1970s all urban populations in Sweden have been connected to waste water treatment plants (WWTPs), which have constantly improved waste treatment and nutrient removal (Bernes 2005). The Rya WWTP in Göteborg is one such example, built in 1974 and amended with a nitrogen removal step in 1987. The nutrient release has decreased from over 600 t P and almost 3000 t N in the 1970s to less than 100 t P and 1500t N nowadays. Point sources from Sweden make up less than 15% of the total nitrogen inputs; however, a substantial improvement for diffuse nitrogen inputs has not yet been reached although progress has been made. This is not only the case in Sweden but is true for all riparian states along the Kattegat and Skagerrak region.

Figure 3.3. Nitrate concentrations (upper) and N/P ratios (based on nitrate and inorganic phosphate) at 20 and 50 m depth in southern Kattegat in late April from 1988 to 2007. Data from Institute of Marine Research, Bergen, Norway.


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Atmospheric deposition is an important source of N in coastal waters. It is suggested that N deposition from air may support ongoing phytoplankton blooms in summer but the N supplies and concentrations are considered too low to initiate a phytoplankton bloom (Carstensen et al. 2004, Spokes at al. 2006).

Figure 3.4. Total nitrogen (left, blue bars) and total phosphorus (right, yellow bars) loads of Swedish rivers and flow of rivers (red line) to the Kattegat, Skagerrak and Öresund from 1969 to 2007. Note the different scaling for the N and P load and the mean flows (SMHI database).


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3.1.4 Trends in source inputs Swedish river N loads into the Kattegat, and Skagerrak have increased through the 1970s to the mid 1980s, but do not show any significant trend over the past 20 years (Fig 3.4). In the Öresund region the situation is more dynamic with fewer clear trends, but overall high loads. Loads clearly depend on the flow rate so that changes in precipitation may be directly translated into changes of N and P inputs into coastal waters. P loads from Swedish rivers are still more variable and have not seen the trend development that can be seen for N inputs. Input into the Kattegat is much higher than that into the Skagerrak and Öresund. Data from 1995 to 2005 did not have statistically significant trends and suggested only slight decreases if any in N and P inputs into the Kattegat and Skagerrak (Håkansson 2007).

Nutrients inputs from Denmark show a drastic decrease from the late 1980s until present (Fig. 3.5). There is no similar trend apparent in the data from Swedish sources.

Trends in total N and P loads from the Baltic Sea can be investigated from changes in surface water concentrations at a station in the Arkona Sea (Figure 3.6). There has been a slight decline in TN concentrations and a more precipitous decline in TP since the mid-1980s. However, there was a substantial increase in TP observed over the past three years, which could be related to P releases from internal sources as a result of hypoxia (Vahtera et al. 2007). Chlorophyll concentrations have increased as well (not shown in the figure) for reasons that have not been investigated, but could be related to blooms of nitrogen-fixing cyanobacteria that respond to such increases in P.

0

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Figure 3.5. Trend in nutrient loads from Danish rivers into the Kattegat from 1989 to 2006. (http:/www.dmu.dk; Ærtebjerg 2007).


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Atmospheric deposition of nitrogen is rising globally (Duce et al. 2008), but declining within the region included in this evaluation. Atmospheric releases of ammonia from agriculture in Sweden have diminished by about 18% between 1995 and 2005. It was determined that 55% of the decline in the release of ammonia between 1990 and 2001 was linked to a reduction in the numbers of animals and 45% due to directed measures to reduce emissions (SEPA, 2007). Increasing trends in atmospheric deposition of nitrogen over the past decades have been reversed and should continue to decline as further controls are implemented. Atmospheric concentration and depositions of all nitrogenous compounds measured at the island of Anholt in the Kattegat has decreased from 1989 to 2006, corresponding to a reduction of 22% in annual depositions (Ærtebjerg 2007).

3.2. Nutrient Status and Trends in Coastal Waters 3.2.1 Concentrations and dynamics The surface TN and TP concentrations in the Skagerrak and Kattegat in winter are much higher than in the Baltic Proper. Close to the coast, nutrient levels are elevated over background levels. Concentrations follow a distinct annual cycle in the surface waters and inorganic nutrients are fully consumed during spring and summer. Below the thermocline, large inventories of nutrients are still present in summer, especially in the deeper parts of the Skagerrak. In the Kattegat there is close coupling between recycling from sediments into the water column due to the shallowness of the system. Along a transect away from a WWTP discharge in the Danafjord, a rapid decline in nutrient concentrations from over 30 to 5 µmol L-1 of NO3 (10 to 2 µmol L-1 of

1970 1975 1980 1985 1990 1995 2000 2005 2010

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Figure 3.6. Arkona Sea nutrient concentrations (µg L-1) (SMHI and Danmarks Miljøundersøgelser data).


21

NH4) over only 20 km has been described (Selmer and Rydberg 1993). Concentrations reach those typical for salinity of 25 in open waters, implying rapid uptake and mixing processes. However, the removal of nutrients may partly be achieved by deposition of organic matter in sediments, from which nutrients are partly released again. N:P ratios in coastal surface waters often deviate from the Redfield ratio of 16, but are adjusted over the annual cycle of production and recycling. Nutrient imbalances, however, may support growth of harmful algal blooms (HABs, Richardson 1997), but not cyanobacteria blooms as they do in the Baltic Sea (see Section 4). A surplus of nitrate or phosphate can also be found in the coastal water close to WWTP and river inputs. Moreover, dissolved organic matter, carried by outflows of the Baltic Sea or rivers, may add to the imbalances in nutrient supply depending on their state of degradation and susceptibility to biodegradation. Removal process of N and P can furthermore change nutrient availability. Denitrification and anammox rates are significant in Aarhus Bay, but may be lower along the Swedish west coast due to a different sediment type (Thamdrup and Dalsgard 2002). Indirect estimates from the Laholm Bay suggest rates of 1.87 mmol m-2 d-1 (Rydberg and Sundberg 1988). High removal rates of nitrate and ammonia in the surface water close to the outlet of the WWTP are suggested (Selmer and Rydberg 1993).


22

3.2.2 Trends Trends in nutrient concentrations in coastal waters are more difficult to evaluate than those for loads because the effects of currents and stratification column need to be considered. Data suggest that nutrient concentra-tions in the upper 15m (in summer that is the layer above the pycnocline) in the Kattegat increased in the 1970s and started decreasing slightly from the mid 1980s (Figure 3.7). Similar declines are noted in deeper waters. This decrease is more pronounced for P than for N. Both nutrients are still at concentra-tions greater than those preceding the acceleration of enrichment in the 1970s. Declines in TP and TN are also evident throughout the water column both in the Kattegat and south-eastern Skagerrak (Figure 3.8). Vertical time series make clear the importance of surface depletion and benthic regeneration of nutrients in the Kattegat, particularly for P.

Declines in nutrient and chlorophyll a concentrations and primary production have been observed in the Göta Älv estuary as a result of advanced wastewater treatment (Rydberg 2008). In Denmark, nutrient concentrations have significantly declined in coastal and open waters in response to measures taken reducing the inputs of nutrients from land (Carstensen et al. 2006).

1980 1985 1990 1995 2000 2005 2010To

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Figure 3.7. Trends in TN (top) TN (bottom) and since 1982 in surface waters (<15 m) at Anholt in the Kattegat (SMHI data).


23

Figure 3.8. Variations of the concentrations of total nitrogen and total phosphorus over time and with depth at stations in the southeastern Skagerrak and central Kattegat since the 1980s (SMHI database).


24

It is stressed that not only the absolute concentrations of nutrients are important, but also the ratios of nitrogen to phosphorus. Most nutrient sources, including river discharges, waste waters, and the Jutland Coastal Current, have increasing N:P ratios because phosphorus has been controlled more efficiently than the nitrogen. Atmospheric deposition is naturally greater for N than for P and the JCC can have N:P ratios up to 50. Although N and P deposition inputs from Denmark have large and different inter-annual variations, nutrient concentrations in coastal and open waters seem to be below the Redfield ratio most of the time (Carstensen et al. 2006). Increasing N:P ratio of nutrients in the input sources and in the coastal waters may potentially affect the composition of the phytoplankton community (see Section 4). The sediment pool of nutrients is very large (Rydberg and Sundberg 1988, Conley et al. 2007). Hypoxic and anoxic conditions in the bottom water have occurred and resulted in large scale die-offs of the benthic fauna. This again affects the nutrient sequestration at the sediment-water interface and may lead to higher storage of nutrients and organic matter in the sediments, which can readily be released by resuspension or events of hypoxia, and less denitrification (Conley et al. 2002). 3.3.3 Budget aspects Different numbers and extrapolations have been used to put the nitrogen sources in perspective. One budget for nitrogen (Figure 3.9) suggests that the Kattegat receives half as much N from direct atmospheric deposition as from land. However, the lateral transport through the Belt Sea to the Kattegat and Skagerrak is less than that estimated by Rasmussen and Gustafsson (2003). The nutrient fluxes to the Skagerrak, particularly for DIN, are clearly influenced by continental river water and average DIN flux can be as high as 350,000 kt yr-1 (Rydberg et al. 1996). While Rydberg et al. suggested that little of this nutrient load reaches the Swedish west coast or the Kattegat, Norwegian monitoring results by IMR in Norway have shown consistently elevated nutrient concentrations (particularly nitrate) in the coastal water masses along the Swedish and Norwegian Skagerrak coasts in the period from winter to early summer (Skjoldal 1993, Skjoldal et al. 1997, Aure et al.

Fig 3.9. Budget of biologically active nitrogen for the Kattegat (units kt yr-1) (Spokes et al. 2006).


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1998, Aure and Magnusson 2008). Nutrient concentrations in the rivers draining into the North Sea have been decreasing (van Beusekom et al. 2005) and this flux may further decrease. The extent to which these nutrients contribute to primary production is difficult to evaluate but should largely depend on the timing. In winter, primary production is low but the above mentioned concentrations have been observed in spring at highest river runoff when the spring bloom is starting. It may be assumed that the nutrients in surface waters and to a certain extent at the thermocline are fully consumed by phytoplankton (Rydberg et al. 2006). A budget provided in Håkansson (2007) shows very high nitrogen exchange rates of over 4000 kt N yr-1 from North Sea to the Skagerrak and vice versa and a net input of 231 kt N yr-1 from the Belt Sea and Öresund into the Kattegat. These numbers have to be considered when overall reductions are considered for the Swedish nutrient input from land based sources. Differences between the annual averages of net supply and export of nutrients to the Kattegat-Skagerrak region for the period 1985-2002 indicate a “change” or uptake of 214 kt N and 5 kt P y-1 (Håkansson 2007). In 2001-2002, the net supply of N from land, atmosphere and the Baltic Sea to the Skagerrak and Kattegat was estimated at 300 kt N yr-1 and of this about 225 kt N yr-1 are assumed to be exported to the North Sea and the rest removed by denitrification. This external load is approximately five times higher than all land based N-loads from Sweden for 2006. The overall reduction of land based sources is higher for TP than it is for TN. A lowering of primary production rates is to be expected and presumably also a change in the N:P ratios (Rydberg et al. 2000, addressed in section 4.1.2). The construction of a reliable budget is thus rather difficult due to major uncertainties inherent in the large and variable volumes of water exchanged between the Baltic Sea, Kattegat and Skagerrak and the North Sea to the west.


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4 Ecosystem Responses 4.1 Phytoplankton Production 4.1.1 Phytoplankton Planktonic primary production in the Kattegat-Skagerrak region of the Swedish west coast ranges from oligo- to mesotrophic (100-250 g C m-2 y-1) in open waters to meso- to eutrophic (200-350 g C m-2 y-1) in estuarine and near-shore regions, with monthly mean rates fluctuating from >700 mg to 2000 mg C m-2 d-1 based on 14C uptake measurements (Richardson and Heilmann 1995, Lindahl et al. 1998, Rydberg et al. 2006, Lindahl 2002). These rates show strong geographic gradients, with highest rates present in fjords, embayments and river mouth regions and lowest rates in open sea regions. These gradients appear to follow gradients in natural (oceanic) import and anthropogenic nutrient sources (Lindahl 2002, Johan Rodhe presentation to panel). The phytoplankton community is dominated by diatoms, which typically form spring blooms (late February-April) and large dinoflagellates (e.g. Ceratium), followed by flagellates and smaller dinoflagellates that dominated from late spring through summer and autumn months (Heilmann et al. 1994). In contrast to the Baltic Sea, filamentous and colonial cyanobacteria appear to largely be absent from the phytoplankton community in west coast waters. Likely reasons for this will be discussed later. While seasonal light availability and temperature regimes play important roles in determining phytoplankton successional patterns, nutrient availability and excesses determine the spatial distribution, magnitude and duration of phytoplankton biomass and blooms. Dominant nutrient sources, such as the Jutland Coastal Current, Baltic Sea outflow and local riverine inputs strongly modulate phyto-plankton primary production and biomass (as cell counts and chlorophyll a). This has been shown for both seasonal blooms and more sporadic blooms of potentially harmful taxa, such as the dinoflagellates (e.g. Dinophysis, Gymnodinium, Alexandrium, Prorocentrum), haptophytes (e.g. Chrysochromulina) (Aksnes et al. 1989) and prymnesiophytes (Prymnesium parva, Prymnesium spp.) (Håkansson 2007). The absolute loads and concentrations as well as ratios of nutrients supplied play roles in determining the structure and abundance (biomass) of phytoplankton communities. This suggests that the loading rates, concentrations and relative proportions of key nutrients (nitrogen, phosphorus and silicon) are important determinants of observed patterns in primary productivity, phytoplankton biomass, composition and successional patterns. Seasonal and inter-annual variability in nutrient supplies plays an important role in explaining variability in phytoplankton community biomass and compositional responses. This linkage can be shown both in terms of the extent to which major coastal currents are advected, dispersed and distributed in the west coast region


27

(i.e. Baltic outflows, JCC and other currents) and the manner by which fluctuations in freshwater discharge via rivers impacts nutrient delivery to estuarine and coastal waters. Typically, high rainfall seasons and years, during which the delivery of nitrogen and phosphorus loads is elevated, lead to relatively high rates of primary production and maximum phytoplankton standing stocks. The interactions of offshore, riverine, atmospheric and Baltic Sea nutrient inputs, together with vertical stratification, control the magnitude and temporal and spatial extent of phyto-plankton biomass and blooms. This is true for both diatoms and highly motile flagellate/dinoflagellate species. 4.1.2 Nutrient limitation Observational and experimental data indicate that the rates of supply, total loads and resultant concentrations of both nitrogen and phosphorus play key roles in determining the biomass and composition of planktonic primary producers. However, as is the case in most coastal marine ecosystems, the oversupply of nitrogen drives the overall eutrophication of Swedish west coast waters (Box 4.1). As with many other estuaries and continental shelf waters that have been studied, Swedish west coast waters exhibit a continuum of salinity (Figure 2.4) and nutrient gradients resulting from the interactions of freshwater runoff and coastal and oceanic circulation features. Spatial and temporal patterns of nutrient surpluses and depletions result in differential availabilities of N and P along these gradients (Figure 3.8). Typically, the more riverine and upper estuarine regions exhibit excess N relative to P supplies, while more saline coastal and seaward regions tend to have the lowest N supplies relative to P supplies. Loss of N due to denitrification (Rydberg and Sundberg 1988), more rapid turnover of available P in surface waters, and release of P from sediments due to iron sequestration by sulfide (Blomqvist et al. 2004) all contribute to the shift from N surplus to P surplus along the continuum.


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Box 4.1. Evidence for the Role of Nitrogen in Marine Eutrophication Many studies conducted over the past 50 years have shown that N enrichment is a primary causative agent of marine eutrophication (Dugdale 1967, Ryther and Dunstan 1971, Nixon 1995, Smetacek et al. 1991, D’Elia et al. 1986, Vollenweider 1992). Evidence includes:

• In situ evidence of the spatial and temporal relationship of N inputs vs. primary production responses (Paerl and Piehler 2008);

• Nutrient addition bioassays where N enrichment has been shown to stimulate primary production (Dugdale 1967, D’Elia et al. 1986, Fisher et al. 1992, 1999, Paerl and Bowles 1987, Pennock et al. 1994, Oviatt et al. 1995, Piehler et al. 2004);

• Paleoecological studies showing that historic increases in anthropogenic nutrient (N-dominated) loading led to eutrophication (Cooper and Brush 1993, Kemp et al. 2005);

• Uptake studies which have shown that at ambient concentrations and supply rates, N limitation is widespread (Harrison and Turpin 1982, Harrison et al. 1987, Syrett 1981);

• Correlative budgetary studies in which N supply rates were directly related to daily or annual rates of primary production in diverse coastal ecosystems (Nixon 1986, 1995);

• Stoichiometric analyses showing that, relative to carbon (C), phosphorus (P), and silicon (Si), N often falls below the nutrient supply ratio needed to sustain balanced plant growth (i.e. Redfield ratio of 105:16:1 for C:N:P; Redfield, 1958, Smith 1990);

• Case studies (e.g., Kaneohe Bay, Chesapeake Bay, Neuse River-Pamlico Sound, Long Island Sound, Narragansett Bay, Baltic Sea, coastal North Sea, northern Adriatic Sea, northern Gulf of Mexico) have shown that increasing N loads are directly linked to accelerated eutrophication (Smith et al. 1981, Nixon 1995, Fisher et al. 1999, Elmgren and Larsson 2001, Boesch et al. 2001, Boesch 2002, Paerl et al. 1998, 2004, Rabalais 2002).

Receiving waters exhibit varying sensitivities to N and other nutrient (P, Fe, Si) loads that are controlled by their size, hydrologic properties (e.g. flushing rates and residence times), morphologies (depth, volume), vertical mixing characteristics, geographic and climatic regimes and conditions. The magnitude and distribution of N in relation to other nutrient loads can vary substantially. In waters receiving very high N loads relative to requirements for sustaining primary and secondary production, other nutrient limitations may develop. This is evident in estuarine and coastal waters downstream of rivers draining agricultural regions highly enriched in N, such as the Po, Rhine, Yangtze and Mississippi, Ganges and Nile rivers (cf. Rabalais 2002, Nixon 2003). Excessive N loading may saturate in-shore primary production, leading to either P and Si co-limitation or exclusive P and Si limitation (Dortch and Whitledge 1992. Lohrenz et al. 1999, Conley 2000, Sylvan et al. 2006), but farther offshore or down drift, chronic N limitation remains (Smetacek et al. 1991, Rabalais et al. 1996). These more distant waters can support additional N-driven eutrophication (Smetacek et al. 1991, Codispoti et al. 2001).

Eutrophication can exert feedbacks on internal N cycling, altering the availability of N and subsequent eutrophication potential. Numerous studies have shown organic matter loading, sedimentation and the extent of bottom hypoxia can regulate key N transformations, including nitrification and denitrification (Henricksen and Kemp 1988, Smith and Hollibaugh 1989, 1998, Seitzinger and Giblin 1996, Heggie et al. 1999, Boynton and Kemp 2000, Fear et al. 2005). These feedbacks can significantly affect N availability, and hence subsequent eutrophication potential (Smith and Hollibaugh, 1998; Eyre and Ferguson, 2002). For example, in the Baltic Sea the extent of hypoxia formation is thought to control denitrification rates and hence the ability of the system to depurate itself of fixed N (Elmgren and Larsson 2001, Vahtera et al. 2007). Lastly, top down effects such as grazing, and removal of grazers by overfishing (Jackson et al. 2001) can significantly alter the flux, availability, utilization and manifestation of N and other nutrient inputs.


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Not surprisingly, bioassays that have been conducted along such gradients in this region show a greater tendency for P limitation toward the fresher end of this continuum, while N limitation tends to be more dominant in the more distal and more saline regions (Granéli et al. 1990, Elmgren and Larsson 2001, Ærtebjerg et al. 2003, Spokes et al. 2006). As with many other large coastal transitional systems (e.g. Chesapeake Bay, Neuse Estuary-Pamlico Sound system, the Dutch Delta region, the Nile Delta, Mississippi Delta, Danube River plume, and the Yangtze Delta), N and P co-limited conditions can also exist and at times prevail (Paerl et al. 1990, Rudek et al. 1991, D’Elia 1987, D’Elia et al. 1986, Fisher et al. 1992, Nixon 2003, Kemp et al 2005, Paerl and Piehler 2008). The Swedish west coastal region appears to fall in line with many other such coastal continua with fairly predictable spatial and temporal gradients in N and P limitation and co-limitation that reflect the combined influence of land-based, human-dominated inputs together with oceanic inputs of these nutrients. Patterns and trends in nutrient limitation can also be inferred from examining nutrient distributional data over time at key monitoring locations in the west coast waters (Figures 3.7 and 3.8) At locations near river mouths and in brackish estuaries, depletion of DIP is experienced earlier and more widely than depletion of DIN. More saline estuarine and near-shore locations demonstrate the highest incidence of depletion of both DIP and DIN; while offshore, mid-Kattegat and Skagerrak regions tend to show the highest incidences of strong DIN depletion, while DIP remains detectable at quite low concentrations (Johan Rodhe presentation to panel). At offshore stations, evidence suggests that DIN depletion tends to occur more rapidly than DIP depletion during and following the spring bloom, suggesting that N limitation develops during the course of the bloom. There is considerable variability in the timing and magnitude of these patterns. Most likely, this reflects the extent to which N is supplied by the external sources such as the Jutland Coastal Current, Baltic Sea outflow and the North Sea, as well as land runoff and atmospheric deposition. During summer months, both DIN and DIP remain depleted in the euphotic, upper mixed layer; however, in contrast to near-undetectable DIN concentrations, DIP concentrations remain detectable and DIN:DIP molar ratios are typically <5, indicating more effective recycling of P and significant and persistent N limitation. Bioassays conducted during this period have confirmed N limitation (Granéli et al. 1990; Spokes et al. 2006). During early spring periods of sufficient N and or P availability, silicon may play an increasingly important role in limiting growth of the dominant diatoms. Bioassays have not specifically indicated Si limitation; however very few bioassays have been conducted in these waters and they have largely focused on late spring and summer periods when diatoms would typically not dominate. The potential for Si limitation has been shown for the Baltic proper (Humborg et al. 2000) and additional bioassays during the early spring bloom period are needed to examine the importance of Si limitation and co-limitation, especially with N.


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4.1.3 Climatic factors Climatic variability plays an important role in determining the nature and extent of nutrient limitation of primary production in near-shore and off-shore waters. Firstly, variability in temperature influences the metabolism and growth optima of various phytoplankton groups. In particular, those groups that exhibit relatively slow growth rates at low temperatures, including some dinoflagellate and cyanobacterial species, may be favored by a general warming of the water column (Reynolds 2006, Paerl and Huisman 2008, BACC Author Team 2008). Presumably, these species may compete more effectively with diatoms under a regional warming scenario (although here are other physical constraints, including persistent mixing and high flushing rates that would prevent cyanobacterial dominance, see Section 4.1.4). Another product of warming will be intensification of thermal stratification. Density stratification of the waters of the Kattegat and Skagerrak is dominated by vertical salinity gradients, therefore increased surface water warming will likely play a relatively small role in enhancing stratification. However, stronger stratification would favor highly motile flagellate and dinoflagellate species that can migrate between the pycnocline and well-mixed surface waters. These species are capable of effectively sequestering DIP and (using alkaline phosphatases) organically-bound phosphorus at depth and storing assimilated P as polyphosphates for use in the lighted surface waters (cf. John and Flynn 2000, Reynolds 2006). The combination of effective P uptake and storage is likely to enhance the reliance on DIN (and potentially dissolved organic nitrogen, DON) availability to optimize bloom formation. Stated differently, the scenario of surface water warming, combined with stronger stratification (and calmer weather) should enhance N limitation, especially in off-shore waters. Climate warming models project elevated amounts and more episodic delivery of precipitation for northern Europe (Christensen et al. 2007), with potential impacts on the delivery of diffuse nutrients from the catchments to estuarine and coastal waters (Bernes 2003; Graham 2004, BACC Author Team 2008). The ramifications for nutrient limitation are uncertain; however, larger freshwater discharge events are likely to enhance delivery of nutrients to receiving waters. This would have a proportionately larger effect on N as opposed to P delivery, because DIN is more soluble and more effectively leached from soils than DIP (McDowell and Sharpley 2001, Toth et al. 2006). If so, P limitation should increase at riverine-estuarine locations and delivery of N to more distal waters should increase, possibly enhancing primary production, biomass and bloom formation in these largely N-limited waters. Accompanying this might be an increased potential for harmful (toxic, hypoxia generating and food-web altering) phytoplankton blooms, which are known to be strongly stimulated by increased N supplies in coastal marine waters (Anderson and Garrison 1997, Paerl 1997, Paerl and Whitall 1999). It would seem highly unlikely that the enhanced N load accompanying more frequent and intense storm (and runoff) events will cause these ecosystems to switch from N to P limitation or co-limitation, as stoichiometric analyses of these waters indicate


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low DIN:DIP ratios, except where enriched with N from the Jutland Coastal Current.

4.1.4 Why N2 fixation does not compensate for N limitation Geochemists have pointed out that, theoretically, nitrogen (N2) fixation should compensate for N-limitation in the world’s oceans and seas (c.f., Doremus 1982, Tyrell 1999) as well as inland waters (Schindler et al. 2008). According to this argument, P availability (which is assumed to control N2 fixation) is ultimately limiting primary production. In the world’s oceans, this argument operates over geological time scales and requires predictable and consistent biology (i.e., N2 fixation is solely and consistently controlled by new P inputs; Doremus 1982, Tyrell 1999). However, the theory does not seem to be compatible with biological time scales and the complex environmental controls of N2 fixation beyond phosphorus availability (Paerl 1990). In many estuarine and coastal systems, N2 fixation does not automatically “turn on” when P is adequate and N is limiting. Experimental data indicate that other factors, including N:P supply ratios, iron (Fe) limitation, organic matter availability, physical constraints such as turbulence, advective processes and residence time, irradiance, and potentially “top down” consumption processes control N2 fixation (Howarth 1988, Paerl 1990, Paerl and Fulton 2008). As a result, this argument has limited application to managing coastal eutrophication. Here we elaborate on these alternative restrictions on N2 fixation; most of them are applicable to Swedish west coast waters. Trace metal (Mo) and iron (Fe) limitation have been identified as potential factors controlling N2 fixation potentials in marine ecosystems (Howarth and Cole 1985, Rueter 1988, Paerl et al. 1994) because these metals are cofactors in the enzyme complex, nitrogenase, which mediates N2 fixation (Paerl 1990). Molybdenum was suggested as limiting N2 fixation under increasingly-saline conditions, based on the observation that sulfate (SO4

-2), which is abundant in seawater and is a structural analogue of the dominant source of Mo, molybdate (MoO4

-2), might competitively inhibit uptake of molybdate (Howarth and Cole 1985). Subsequent studies have found this not to be the case, even at very high salinities exceeding those found in the Kattegat-Skagerrak regions, i.e. molybdenum availability exceeds demands in these waters (Collier 1985, Paulsen et al. 1991). Therefore, there is little reason to believe that N2 fixation might be controlled by molybdenum availability in Swedish west coast waters. These waters have also been found to be quite rich in biologically-available iron (Croot et al. 2002). Accordingly, we conclude that it is unlikely that the paucity in N2 fixation in these waters is due to iron-limited conditions. For some time, it has been argued that salinity itself might be a barrier to the establishment of N2 fixers in coastal and open ocean environments, because growth of dominant freshwater N2 fixing genera, including Anabaena and Aphanizomenon, can be shown to be inhibited by salinities exceeding a few salinity units (Moisander et al. 2002a). However, salinity per se, is not a strong modulator of


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either the establishment or activities of all cyanobacterial diazotrophs. A wide variety of active N2-fixers, including the genus Nodularia, common to the Baltic Sea, has been observed in the plankton and benthos of estuarine, coastal and open ocean environments, and even hypersaline lakes and lagoons (Potts 1980, Paerl 2000, Moisander et al. 2002a). Therefore, salinity cannot explain the scarcity of N2 fixers in Swedish west coast waters. Turbulence exerts a strong impact on phytoplankton growth and structural integrity (Fogg 1982, Reynolds 1987). Increased levels of turbulence may inhibit growth of diazotrophs (Fogg 1982; Paerl 1990). Aquatic environments with persistent elevated turbulence may have a lower abundance of active N2-fixing heterocystous cyanobacteria. In laboratory experiments where shear rates representative of surface wind-mixed conditions were applied to bloom-forming cyanobacteria (Anabaena, Nodularia), Kucera (1996) and Moisander et al. (2002b) showed that rates of N2 fixation and photosynthesis can be suppressed by strong turbulence. The negative impacts of elevated shear could be due to: 1) breakage or weakening of cyanobacterial filaments, specifically at the delicate heterocyst-vegetative cell junction, causing O2 inactivation of nitrogenase in heterocysts (Fogg 1969), and 2) disruption of bacterial-cyanobacterial associations (Paerl 1990). In the Baltic Sea there are mid-summer periods of relaxed winds as well as stable fronts during summer months. These are often the times and locations where cyanobacterial blooms occur (Kononen et al. 1996) as was particularly evident during the warm, still conditions that prevailed during the summer of 2005 (Vahtera et al. 2007). Another major difference is that along the west coast the nutricline is located with the pycnocline around 15 m, whereas in the Baltic Sea the nutricline is deeper. This means that along the west coast there will be more pulses of nutrients being entrained into the surface layer as opposed to in the Baltic Sea, where the surface layer remains deficient in N for longer periods. These nutrient pulses favor diatoms and dinoflagellates such as Ceratium that are typically abundant around the pycnocline. The dinoflagellates can use their motility to exploit both nutrients and light. The horizontal movement of water is also much stronger along the west coast than in the Baltic Sea, due to both strong variations in barotrophic and baroclinic differences. Interestingly, despite the absence of planktonic N2 fixation in these turbulent systems, cyanobacteria and bacteria potentially capable of N2-fixation can be found in these systems, but they are most often confined to the benthos, submersed surfaces and in epiphytic communities (Paerl et al. 2000). Molecular studies, based on the analysis of the N2 fixing gene nifH, indicate that a diverse taxonomic potential exists for N2 fixation in these waters (Affourtit et al. 2001, Jenkins et al. 2004). However, N2 fixation activity is generally absent or present at ecologically-insignificant rates, and if it does occur, it is usually confined to sedimentary or biofilm habitats. A number of studies have suggested various physical and geochemical barriers to the establishment and dominance of N2-fixers in N-limited


33

estuaries, especially in the water column. The relatively turbulent properties of estuarine waters, which include strong wind mixing, horizontal advection, tidal mixing, and high rates of small-scale shear, may restrict the establishment and proliferation of diazotrophic cyanobacterial and bacterial communities (Moisander et al. 2002b; Paerl 1996). In particular, persistent vertical mixing of near-surface waters prevents dominance by buoyant filamentous diazotrophic cyanobacterial bloom genera (e.g., Anabaena, Aphanizomenon, Nodularia, Trichodesmium). Further, the growth of these N2-fixing species that typify Baltic Sea blooms is inhibited in higher salinities because nitrogenase activity is limited by higher sulfate concentrations (Stal et al. 2003). Residence time (flushing rates) can also play an important role in determining the degree to which diazotrophic cyanobacteria are present and dominate N-limited aquatic ecosystems. Even though N2 fixing cyanobacteria can form massive surface blooms in many lakes and quiescent marine ecosystems, growth rates of key bloom-forming genera (Nodularia, Aphanizomenon, Anabaena) are generally much lower (doubling times of 2-3 days) than those of non-N2 fixing eukaryotic groups such a diatoms, flagellates and even dinoflagellates (doubling times of 0.5-1 day). Therefore, in rapidly flushed estuaries and coastal sounds with low residence time that experience N limitation or P enrichment, bloom-forming cyanobacteria often do not compete effectively because growth rates cannot effectively keep up with flushing rates. As a result, they fail to exert dominance and more rapidly-growing taxa prevail. Relatively slow growth rates are often exploited by lake and reservoir managers to control and prevent cyanobacterial blooms, by keeping these systems flushed during periods of optimal cyanobacterial growth (summer), thereby promoting dominance by fast- growing and more desirable eukaryotic groups (Reynolds 1987). Water residence time in the surface layer of the Kattegat and Skagerrak is in the order of one month (Gustafsson 2000, Johan Rodhe presentation to panel). Such short residence times and highly dynamic horizontal advective conditions prevent the establishment and buildup of cyanobacterial bloom populations and may help explain their absence on seasonal and multi-annual time scales. In contrast, the Baltic Proper has a residence time on the order of 25 years, accompanied by per-manent stratification and strong fronts. These are ideal conditions for the establish-ment and persistence of cyanobacterial bloom populations (Kononen et al. 1996). In summary, estuarine and coastal waters have a diverse genetic potential for N2 fixation, which under favorable conditions (e.g., mid-summer stratified conditions in the Baltic Sea) can be readily expressed. However, more often, persistently wind mixed surface waters, readily flushed and nutrient-pulsed conditions in these environments represent physical and chemical barriers to N2-fixers, thus restricting their dominance and bloom potentials. This, combined with the fact that estuarine and coastal systems are frequent sites of active denitrification and phosphorus


34

sufficiency, helps explain why significant N2 fixation does not occur Swedish west coast waters even under prolonged nitrogen deficiency. 4.2. Macrophytes Macroalgae attached to rocky shores and bottoms, both in the intertidal zone and subtidally, provide important habitat within coastal ecosystems. Although Wennberg (1987) reported qualitative shifts in the macroalgal vegetation from the bladder wrack (Fucus spp.) to filamentous green algae (Cladophora and Enteromorpha) in the southern part of Laholm Bay during the 1970s and 1980s, there is surprisingly little scientifically rigorous documentation of changes in the macroalgal communities along the west coast of Sweden. Other long-term comparisons have shown an increase in filamentous algae, but no decrease in perennial brown algae (Johansson et al. 1998). Nonetheless, a narrowing of the depth distribution of dominant brown algae, increases of their epiphytes, and a decline in species richness in the lower littoral have been observed, consistent with declining light penetration and increased nutrients (Petersén and Snoeijs 2001, Eriksson et al. 2002). The increase in the ephemeral abundance of filamentous green algae (mainly Cladophora and Enteromorpha) in shallow bays along the Skagerrak and Kattegat coasts since the 1970s is, however, very well documented (Figure 4.1, Pihl et al. 1995, 1999). Similar manifestations of eutrophication have occurred around the world and commanded global attention when massive quantities of Enteromorpha and other drifting green macroalgae threatened to interfere with the 2008 Olympic sailing competition off Qingdao, China (Hu and He 2008). Masses of these algae drift into shallow waters and along shorelines, creating nuisance conditions for people seeking recreation and diminishing the habitat quality for important fishery species, such as plaice (Pihl et al. 2005) and important prey species (Wennhage and Pihl 2007). This increase in filamentous algae accumulating on soft bottoms seems

Figure 4.1. Percentage cover of filamentous algae in 400 shallow bays on the Swedish west coast, 1994-1996 (Pihl et al. 1999)


35

to be clearly linked to nutrient over-enrichment, but the sources of the nutrients (i.e., the degree to which local sources contribute versus the general increase in nutrient concentrations in coastal waters) are unclear. Nutrients regenerated from the sediments seem to be important in initiating and sustaining these macroalgal blooms (Sundbäck et al. 2003) and the buildup of nutrients in sediments maintains persistence of this new regime with its degraded ecosystem services (Troell et al. 2005). This might explain why there have not yet been any obvious improvements in this green algal mat phenomenon. Another benthic macrophyte community that has been severely affected by eutrophication is seagrass beds. The rooted marine vascular plant Zostera marina (eelgrass) declined in abundance as a result of eutrophication in a pattern consistent with the demise of seagrasses around the world. Along the Swedish Skagerrak coast, a 58 % decline in eelgrass was observed between the late 1980s and early 2000s (Baden et al. 2003), in part due to the reduction of depth at which it has sufficient light to live. Similar and contemporary declines in Danish eelgrass beds were closely related to increases in total nitrogen concentrations (Nielsen et al. 2002). Recent studies have also suggested that eelgrass communities may be affected by the dramatic reductions of top predators in the ecosystem as a result of overfishing (Moksnes et al. 2008). This has allowed an increase in smaller fish such as gobies that, in turn, reduce the populations of important grazers that keep the growth of epiphytes, which grow on the blades of eelgrass and shade them, in check. Eelgrass beds that succumb to overgrowth by epiphytes loose much of their habitat value for fishes, with reductions in both diversity and the juvenile populations of species such as cod (Pihl et al. 2006). While eelgrass may spread during dry years with low nutrient concentrations and high light levels, monitoring of Danish eelgrass beds under conditions of decreasing nutrient concentrations has shown that slow spreading of beds into deeper water has been observed in a few areas, but most eelgrass meadows have not reestablished according to the light potential (Ærtebjerg 2007). 4.3 Dissolved Oxygen One of the adverse effects of eutrophication is the serious depletion of oxygen from bottom waters, or hypoxia (Figure 4.2). Along the Swedish west coast and in the Kattegat hypoxia became apparent in the beginning of the 1980s, when oxygen concentrations in bottom waters over extensive areas in the southern Kattegat reached levels detrimental to benthic animals (species-specific effects typically starting in the range from 2 to 5 mg L-1; Vaquer-Sunyer and Duarte 2008).


36

The oxygen condition in the Kattegat is a balance between respiration in the sediments and bottom water and oxygen supply from photosynthesis in overlying waters and air-sea exchange. Because the Kattegat is almost permanently stratified around 15 m (Anderson & Rydberg 1993), oxygen has to be supplied to bottom waters through advection of surface Skagerrak water penetrating below the outflowing Baltic Sea water. Average residence time of the Kattegat bottom water is around 2 to 4 months (Gustafsson 2000, Johan Rodhe presentation to panel) with strong advective transport during winter slowing down during summer and intensifying again during autumn. These physical mechanisms result in the southern Kattegat having a natural oxygen minimum in September, but there can be strong inter-annual variations depending on the volume and degree of stratification of the bottom water. This implies that bottom waters during the low oxygen period (August-October) originate from the Skagerrak surface water in winter-spring. Due to the low and varying temperatures of Skagerrak surface water during this time of the year (2-6°C), there can be variations in the oxygen concentrations of approximately 1 mg L-1 in the water mass supplying oxygen to the Kattegat bottom waters. Conley et al. (2007) found quantitative evidence for three factors explaining variations in the summer-autumn oxygen concentrations of the Kattegat and Belt Sea: 1) temperature, through increased metabolism and lower oxygen saturation; 2) advective bottom water transport; and 3) nitrogen input from land that enhances primary production and export of organic material from the upper mixed layer. Other studies have also concluded that there is no single factor to which all variations in oxygen concentrations can be attributed (Rasmussen et al. 2003).

Coastal areas, such as the Laholm Bay and Skälderviken, which connect to the southern Kattegat, were also severely affected by low oxygen concentrations in

Figure 4.2. Oxygen depletion near the sea bed in the southern Kattegat (blue: <4 mg l-1, red: <2 mg l-1) during September in an extreme year (2002, top) and a normal recent year (2006, bottom). Source: NERI.


37

1980s (e.g. Rydberg et al. 1990). Hypoxia along the southern Swedish west coast is strongly linked to the conditions in the open waters of the southern Kattegat, where bottom waters, low in oxygen, are advected from the open to the coastal waters, typically during periods of easterly winds. The depletion of oxygen in the bottom layer can be further exacerbated in the shallow coastal region because the bottom water penetrates as a thin layer allowing respiration processes in the sediments and bottom waters to deplete the oxygen inventory in a thin section of the water column. Thus, hypoxia in the coastal areas of the southern Kattegat can be inter-mittent and more dynamic than the rather slow oxygen depletion and repletion processes of the open Kattegat. Higher primary production rates in the coastal zone (Carstensen et al. 2003) contribute organic matter that increases sediment respiration, intensifying hypoxia along the southern Swedish west coast (Figure 4.2). To the north along the Swedish west coast the coastal zone changes from shallow coastal embayments to fjords, many of these have a sill restricting the ventilation of bottom water. Long retention times of bottom waters in fjords naturally lead to hypoxia in the very deepest parts, but eutrophication has further lowered oxygen concentrations and increased the volume of hypoxia in areas such as Gullmarsfjord and Stigfjorden (Rosenberg 1990, Lindahl presentation to panel). Renewal of bottom waters typically occurs during strong wind events from north-easterly directions with infrequent major replenishments of oxygen (Erlandsson et al. 2006). Thus, the physical characteristics of the open southern Kattegat, the coastal embayment along the southern Swedish west coast, and the fjords on the Skagerrak coast are quite different in modulating the overall oxygen response to increased nutrient enrichment. 4.3.1 Status and trends Hypoxia in the open southern Kattegat and Öresund has become more prevalent since the 1970s when the first regular monitoring programs were established, and there are no signs of recovery despite reduced inputs of nutrients over the last 10-15 years (Conley et al. 2007). Extensive areas (~100-500 km2) are exposed to severe hypoxia (<2 mg l-1) in most recent years (2003-2006) (Ærtebjerg 2007), but sizes of these areas are much lower than in the catastrophic year of 2002 when >2000 km2 of the Kattegat and Öresund were exposed (Figure 4.4). The 2002 event was indeed an unfortunate combination of the factors leading to hypoxia in the open waters of the Kattegat: high temperatures, an almost complete stagnation of bottom waters, and high inputs of nitrogen (HELCOM 2003). However, with projected increases in temperature and precipitation due to climate change it is likely that the physical setting and conditions for this event may reoccur. The status and trends of hypoxia in the coastal embayments along the southern Swedish west coast is similar to the open Kattegat with some variation related to


38

how far into these embayments the hypoxia bottom water penetrates. The deep parts of the Öresund becomes hypoxic almost every year, while the shallower parts of the Öresund, Skälderviken and Laholm Bay only become hypoxic in the worst years. Oxygen concentrations in the Swedish fjords on the Skagerrak coast have also decreased over longer time-scales, with a tendency for some recovery in recent year (Erlandsson et al. 2006; Figure 4.3). In the Gullmarfjorden, annual means were around 3.5 ml L-1 up to around mid 1970s when oxygen levels started decreasing, reaching a level of about 2.0 ml L-1 in the late 1990s. These levels have then improved to about 3.0 ml L-1 in the most recent years. This recent improve-ment can be attributed to a combination of reduced primary production (Odd Lindahl presentation to panel) and the relatively favourable climatic conditions (expressed as lower NAO index) in recent years, which increased the salinity and density of the Skagerrak water replenishing the bottom waters of the Gullmarfjorden.

4.3.2 Organic matter supplies and metabolism The fate of organic matter produced in the surface layer can follow different pathways: remineralization in the surface layer, incorporation into higher trophic levels, or export to the bottom waters and sediments. Experimental studies using sediment traps have estimated annual sedimentation in the southern Kattegat to be 63 g C m-2 yr-1 (Olesen and Lundsgaard 1995). Carstensen et al. (2003) estimated the annual sedimentation for the entire Kattegat to be 55 g C m-2 yr-1 in an empirical model study, corresponding to 47% of the primary production. There are,

0

1

2

3

4

5

6

1900 1910 1920 1930 1940 1950 1960 1970 1980 1990 2000 2010

oxyg

en (m

l l-1

)

Figure 4.3. Annual mean oxygen concentrations in the Gullmarfjorden with observations weighted according to their month of sampling (SMHI data).


39

however, differences in the methods used to estimate the sedimentation rate as well as interannual variation. Therefore, this value should be interpreted with caution. Another approach to assess sedimentation rates is to apply an empirical relationship between measured primary production and sedimentation rate (Wassmann 1990). This relationship shows a progressively increasing export of production from the plankton that when applied to the primary production data from the Gullmarfjorden suggests that organic sedimentation rates may have increased four-fold since the 1950s (~25 g C m-2 yr-1) to the 1990s (~100 g C m-2 yr-1), even though primary production increased only three-fold (Lindahl 2002; Figure 4-4). The fraction of primary production that is exported from the surface layer similarly increased from ca. 30% to over 50%, although the extrapolation of the relationship from Wassmann (1990) to high primary production rates by virtue is uncertain. These studies indicate that sedimentation rates probably range from 50 g C m-2 yr-1 in the open waters to 100 g C m-2 yr-1 in the coastal regions. These findings of different primary production rates in inshore compared to offshore waters are also supported by Rydberg et al. 2006.

Seabed oxygen consumption rates are estimated to range from 10 to 20 mmol O2 m-2 day-1 from a variety of experimental and modelling studies (Rasmussen et al. 2003 and references therein). Converting sedimentation rates from the literature, assuming all sedimenting organic material is respired, gives somewhat higher values (~20 mmol O2 m-2 day-1), suggesting that up to 50% of the sedimented organic matter may be buried. However, these carbon and oxygen budgets are

Figure 4.4. Estimated sedimentation in the Gullmar Fjord 1950 to 2007 (O. Lindahl presentation).


40

uncertain. More investigations quantifying these rates over time and space would be useful for describing the effects of eutrophication on oxygen conditions. 4.4. Benthos of Sediment Bottoms The macrobenthos of the seas and coastal areas of western Sweden is probably the best studied worldwide. The responses of macrobenthos to organic enrichment of bottom sediments and ultimately to hypoxia in bottom waters are well characterized, leading to excellent documentation of the time course of changes related to eutrophication and associated hypoxia and a diagnostic ability to characterize stress on the communities based on faunal communities and in situ profiles of sediment structure (Figure 4.5).

Figure 4.5. Benthic infauna successional stages along a gradient of organic enrichment and oxygen depletion based on studies of the Gullmarsfjord, including typical sediment-profile images and benthic-habitat quality (BHQ, reflecting sediment structures) and benthic (BQI, reflecting species composition, abundance and richness) indices based on species composition and abundance (Rosenberg et al. 2004)


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The effects of eutrophication on macrobenthos were already clearly apparent in the mid-1980s when comparisons were made with the observations of the pioneering Danish marine biologist Johannes Petersen in 1911-12 (Pearson et al. 1985; Rosenberg, et al. 1987). Biomass was estimated to have increased in the Skagerrak, presumably as a result of trophic enrichment, but markedly decreased in the Kattegat, which by the mid-1980s had begun to experience stress of hypoxia resulting in the elimination of sensitive molluscs, echinoids and crustaceans. By the early 1980s, mass mortalities of benthic invertebrates in the inner Kattegat were observed (Rosenberg et al. 1992). Catches of Norway lobsters (Nephrops norvegicus) had initially increased, because they became more susceptible to trawl capture as they left their burrows under oxygen stress, but subsequently collapsed completely in the southern Kattegat as they succumbed to the lack of oxygen. Catches of bottom fish also declined dramatically in the Kattegat and Skagerrak after the 1980s (Håkansson 2003). While overfishing is largely responsible (Cardinale and Svedäng 2004), the increase of bottom water hypoxia seems also an important factor in this decline. Soft-bottom benthos in coastal regions has deteriorated in many of the fjords along the Swedish Skagerrak coast since the 1980s as a result of increased organic deposition (perhaps as a result of local increases in filamentous macroalgae as well as increased phytoplankton production) and hypoxia (Rosenberg and Nilsson 2005). Most of these fjords are not recipients of significant nutrient inputs from land, suggesting that these environments have experienced the effects of regional eutrophication impinging on the coastal zone. In addition to the effects on macrofauna, organic enrichment of sediments and severe oxygen depletion of overlying waters can dramatically affect the biogeo-chemical functioning of the seabed, including respiration, susceptibility to sediment resuspension, and nutrient recycling. Benthic animals inhabiting sediments are important regulators of this functioning through their burrowing and ventilation, bioturbation and tube-building activities. Organic enrichment of sediments and hypoxia in bottom waters tend to eliminate the deep burrowing animals that are particularly important in regulating the biogeochemical exchanges at the seabed. The ecosystem functions of organically-enriched and faunally-depauperate sediments are greatly altered, resulting in less efficient degradation of excess organic material, lower rates of denitrification, and storage of nutrients in the sediments. Under near-anoxic conditions such sediments can release large quantities of phosphorus, as sulfides outcompete phosphates for binding sites, and ammonia, as nitrification shuts down because of the lack of oxygen (Jordan et al. 2008). Nutrients released from the seabed in this way can refuel the fire of excess primary production, creating a positive feedback to maintain a vicious cycle of eutrophication. Benthic organisms are important food resources for bottom feeding fish, such as haddock and plaice. Although the effects of benthic food resources altered by


42

eutrophication have not been quantified for Swedish western seas, one can conclude that the effects on demersal fish are more likely negative than positive over the long term. Most Skagerrak-Kattegat fish stocks have been depleted by overfishing and climatic influences, and this diminution of food quantity and quality may pose a limiting factor in their recovery (Nielsen and Rosenberg 2003).


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5. Reversing Eutrophication 5.1. Effects of Countermeasures Taken 5.1.1. Swedish sources The Swedish EPA (2007) summarized the measures taken to reduce emissions of nitrogen and phosphorus and estimated loadings for a period circa 1995 and a period circa 2005. In 2005 atmospheric ammonia emissions for the whole of Sweden were estimated to be 52,400 t, a reduction of 16% from 1995. Agriculture accounts for 85% of ammonia emissions and most of the reduction resulted from the decline in the number of animals, particularly milk cows and pigs for slaughter, although the increase use of liquid manure has also contributed to reduced emissions. In 2006 total nitrogen oxide emissions in Sweden were approximately 170,000 t, a reduction of about 25% from 1995 levels. This is a continuation of a long-term trend resulting from measures to control emissions. Deposition of nitrogen on open fields in southern Sweden (Götaland) has declined by about 40% between 1988 and 2005, about 11% between 1995 and 2005. Estimated anthropogenic loads to western sea areas resulting in emissions to water are summarized in Table 5.1. These are broken down by source of diffuse and point emissions for phosphorus, but nitrogen loads reaching coastal waters are distinguished only between those arriving by river discharges from inland sources (including point sources discharging to rivers, agricultural and urban runoff, and atmospheric deposition falling on lands and lakes).


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Approximately 65% of the anthropogenic waterborne emissions of nitrogen to the sea for all of Sweden emanates from diffuse leaching from agricultural land and emissions from sewage treatment plants (SEPA 2007). Atmospheric deposition onto lakes accounts for an additional 18%. The greatest reductions in emissions of nitrogen between 1995 and 2005 have been achieved for point sources discharging to coastal waters, resulting in a reduction of 49% for the Swedish west coast. Leaching from agricultural lands has decreased by an estimated 13% nationally. Phosphorus emissions to the western seas have also declined primarily as a result of treatment of point sources effluents from sewage treatment plants and industry, resulting in a 24% reduction in these sources (Table 5.1). Some reductions in phosphorus from surface water draining urban areas and from agriculture have also been estimated, but these are relatively small in absolute terms. The Kattegat receives the largest loads of nitrogen from agriculture, although these loads have recently declined to 9,200 t (compared to 20,800 t from agricultural sources in the whole Kattegat catchment; SEPA 2008b). For Sweden as a whole, approximately one-third of the nitrogen used in agriculture is not taken up by crops and more than one-half of this surplus is presumed to leach into surrounding waters (SEPA 2007). Leaching is significantly higher in agriculturally intensive areas with well-drained soils in portions of Skåne, Halland and Västra Götaland, which drain

Table 5.1. Estimated anthropogenic loads to the western seas between 1995 and 2005 (SEPA 2007). Anthropogenic P emissions (t yr-1)

Arable land Logging Urban runoff Total diffuse Point emissions Total

1995 2005 1995 2005 1995 2005 1995 2005 1995 2005 1995 2005

Öresund 30 30 <5 <5 10 10 40 40 60 40 100 80

Kattegat 340 310 10 10 60 50 410 370 350 280 760 650

Skagerrak 70 80 <5 <5 10 5 80 85 40 20 120 105

Total 440 420 10 10 80 65 530 495 450 340 980 835

Difference % 4.5% 0.0% 18.8% 6.6% 24.4% 14.8%

Anthropogenic N emissions (t yr-1)

Inland sources

Point emissions to the sea

Total

1995 2005 1995 2005 1995 2005

Öresund 4,600 3,500 1,600 700 6,200 4,200

Kattegat 22,400 18,200 3,200 1,700 25,600 19,900

Skagerrak 900 900 700 400 1,600 1,300

Total 27,900 22,600 5,500 2,800 33,400 25,400

Difference % 19.0% 49.1% 24.0%


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to the western seas. The loads of phosphorus from agriculture are very uncertain, have declined only slightly and are more affected by preparation of the land and rainfall than by management of mineral fertilizers, the application of which in Sweden has declined by half since the 1970s. 5.1.2. Transboundary sources Reductions in nutrient loads from the northwestern European rivers (Figure 3.1) that contribute to the Jutland Coastal Current were discussed in Section 3.1.2. Evidence for declines in nutrient concentrations in the Baltic Sea outflow (Figure 3.6) was discussed in Section 3.1.4, which also presented documentation of the reductions of loads of both nitrogen and phosphorus from Danish rivers draining to the Kattegat (Figure 3.5). At least in the case of Denmark and the Rhine and Elbe rivers, these reductions have been achieved by specific measures taken to reduce nutrient pollution, initially by treatment of wastewaters to remove phosphorus and in some cases nitrogen followed by various steps to reduce diffuse source inputs, particularly from agriculture (Carstensen et al. 2006). The efforts in Denmark, which has intense animal agriculture and high nutrient yields to surface waters, have been especially aggressive and have resulted in a substantial reduction of nitrogen loads to the Danish Belt Seas from the high point in the mid-1980s (Figure 5.1). By 2002 reduction in total anthropogenic nitrogen transport to marine waters from Denmark has been estimated at 40%, from wastewater 69% and from agriculture by 31% in one assessment (Grant et al. 2006).

Figure 5.1. Estimated nitrogen loading to the Belt Seas of Denmark (Conley et al. 2007).

Figure 5.2. Changes in the concentration of forms of nitrogen in atmospheric deposition at Anholt, outer Kattegat (Ærtebjerg 2007).


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Atmospheric deposition of nitrogen onto the Kattegat and Skagerrak also emanates from multinational sources. Concentrations of nitrogen species collected in atmospheric deposition have declined over the past 20 years (Figure 5.2), essentially tracking the declines in deposition observed at Götaland in southern Sweden. These declines result from efforts to reduce emissions of nitrogen oxides for improvement of air quality as well as steps to reduce ammonia emissions from agriculture. 5.2. Responses to Reductions in Nutrient Inputs By the mid 1980s, nutrient inputs to the Kattegat and Skagerrak are believed to have increased by factors of 5-6 for nitrogen and 8 for phosphorus from the beginning of the 20th century (Conley et al. 2007, R. Rosenberg presentation to panel). As discussed above, significant strides have been made to reduce loadings to alleviate the undesirable effects of eutrophication in inland and marine waters and to improve air quality. The conceptual framework underpinning present nutrient management plans is that ecosystems will return to their original state once the nutrient pressure is released. This managerial framework is, however, challenged by emerging ecological theory that suggests that ecosystems respond in a non-linear manner to changing pressures leading to the existence of regime shifts between alternative stable states (see Section 5.3.2). In this section, we will investigate to what extent the marine ecosystems along the Swedish west coast are responding to decreasing nutrient inputs in a predictable manner. 5.2.1 Nutrient concentrations and ratios Nutrient concentrations have decreased in both coastal and open waters in recent years as a response to reduced inputs from the land and atmosphere and advection (SEPA 2008a, Ærtebjerg 2007, Carstensen et al. 2006). The difference in the timing of nitrogen versus phosphorus reductions has led to changes in the N/P ratio that are most pronounced in the coastal areas, albeit also observable in the open waters (SEPA 2008a, Ærtebjerg 2007). In recent years the N/P ratio has consistently been below the Redfield ratio of 16 on a molar basis (SEPA 2008a), typically around 5 for the open Kattegat (Ærtebjerg 2007), except in the Skagerrak where intrusions of N-rich water from the Jutland Coastal Current during summer months can raise this ratio. During the productive season (March-September) the open waters are potentially N-limited 100% of the time, whereas inorganic P is not always depleted from surface water, suggesting P limitation 80% of the period (Ærtebjerg 2007). The reduced concentrations of inorganic nitrogen and phosphorus have changed the nutrient ratios in favour of higher silicate availability, suggesting that silicate is potentially limiting only in estuaries with large N and P discharges. In general, nutrient levels have declined in response to reductions in nutrient inputs as anticipated, and further reductions will increase the periods of both N and P limitation and alleviate potential silicate limitation even further.


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5.2.2 Phytoplankton Despite decreasing nutrient levels in recent years, there is no univocal response for phytoplankton. Chlorophyll levels in the open waters of the Kattegat have remained at an almost constant level of about 2 µg L-1 according to Ærtebjerg (2007), whereas SEPA (2008a) found declines analysing phytoplankton biomass from a single station in the Kattegat, mainly due to reduced concentrations of dinoflagellates and nanoflagellates. The anticipated effects of nutrient reductions on phytoplankton biomass are therefore not consistently documented. Reduced pools of nutrients should conceptually lead to reduced phytoplankton biomass, but increasing turn-over rates of nutrients, reduced grazing of phytoplankton, and extended growing season could explain the lack of a clear phytoplankton biomass response. Reduced inorganic nutrient levels should especially lead to reduced spring blooms and production, but given the strong dynamics of this phenomenon and infrequent sampling during spring, quantitative evidence for this hypothesis is not straightforward. Assuming that the magnitude of the spring bloom has been reduced, then production should be relatively larger during the summer period. Seasonal patterns of primary production could indicate a shift in primary production from new production in spring to regenerated production during summer (Rydberg et al. 2006). That could explain the constant mean annual chlorophyll level that is observed. Increases in temperature will increase the turnover rate of nutrients in the surface layer, and mean surface water temperature in the Kattegat has increased by ~0.5 °C from the 1990s to the 2000s (Ærtebjerg 2007). Climate change may also have led to earlier development of the spring bloom that forms once the water column stabilizes, alleviating light as the limiting factor in the surface layer. An extended productive period will result in higher annual means. It should be noted that McQuatters-Gallop et al. (2007) found no decline in chlorophyll levels in the coastal North Sea following reductions in river nutrient loads and nutrient concentrations, which they attribute to the sea becoming warmer and clearer. Grazing in the open waters of the Kattegat and Skagerrak is entirely pelagic due to the permanent stratification, whereas filter feeders are also potential grazers of phytoplankton in the coastal zone. Based on established grazing rate-to-biomass relationships, the mesozooplankton is potentially capable of controlling the average phytoplankton biomass in the summer period, but, due to their relatively long reproduction times, the mesozooplankton is incapable of promptly responding to bloom situations by increases in biomass. Another issue is the palatability of the phytoplankton. Due to the rather turbulent environments in the Kattegat and Skagerrak, the phytoplankton community is dominated by larger species, typically diatoms such as Skeletonema spp. and Rhizosolenia spp. and dinoflagellates such as Ceratium spp. Changes in phytoplankton communities and reduced grazing are another factor that could explain why phytoplankton biomass apparently has not declined with nutrient concentrations. Another explanation for the constant


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chlorophyll levels is increased advective transports of large cyanobacteria blooms from the Baltic Sea in recent years. Particularly in 2006, large quantities of Nodularia spumigena spread from the Baltic Proper into the Öresund and the Kattegat. The frequency of phytoplankton blooms, particularly harmful algae blooms or HABs is believed to have increased with eutrophication (Hallegraeff 1993, 2003), and for the Kattegat it has been documented that years with higher nutrient inputs and concentrations are likely to have more summer blooms (Carstensen et al. 2004). These findings suggest that the bloom frequency along the Swedish west coast should decrease as a response to the measures taken, but this has not yet been documented. It should, however, be acknowledged that only very few of the blooms on the Swedish west coast are considered harmful and the biomass of the blooms seldom reach levels that can be characterised as a nuisance. 5.2.3 Phytobenthos Macroalgae and angiosperms are expected to increase their depth distribution with improving light conditions. Such improvements could therefore only be partially expected at present, because consistent reductions in phytoplankton biomass have not been consistently observed. Changes in the macroalgae community along the Swedish west coast have been reported (SEPA 2008a) with significantly decreasing depth distributions for two species only (Halidrys and Dilsea). The response of the entire macroalgal community has not been analysed. The experience from the coastal Danish monitoring program suggests little improvement in macroalgae and eelgrass depth distribution in some but not all coastal areas, despite significant declines in chlorophyll and improved light conditions (Ærtebjerg 2007). Rask et al. (1999) also reported improvements for eelgrass in the dry year of 1996 in which N inputs were less than one-half that for an average year. These results indicate a potential slow recovery, where colonisation of new suitable habitats takes considerable time. It is believed that these results can be projected to the Swedish west coast as well, and that a slow gradual recovery may have started. However, the recovery time is not known. 5.2.4 Dissolved Oxygen Oxygen conditions in the bottom waters have not improved despite reduced nutrient inputs and concentrations as described in Section 4.3. Increasing temperatures experienced in the region, perhaps related to climate change, counteract the anticipated improvements through reducing the supply of oxygen and increasing the metabolism. Another, perhaps even more important, factor is the reduced capacity of permanently removing nutrients in the sediment. Periods of low oxygen, particularly anoxia, reduces the nitrification-denitrification pathway of removing nitrogen. Changing the benthic community from deep-burrowing macrofaunal organisms to hypoxia-tolerant species affects the sediments ability to remove nutrients through burial and denitrification (Diaz and Rosenberg 2008). The reduced ability to remove nutrients provides a nutrient feedback to the water


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column that may sustain continued effects of eutrophication, i.e. a vicious cycle of hypoxia (Vahtera et al. 2007). However, it should be acknowledged that nutrient inputs do have an effect on oxygen conditions (Conley et al. 2007) and that present oxygen conditions would likely have been worse if nutrient inputs had not already been reduced, but there is a need for further reductions to counteract the effects of global warming. Accor-ding to the empirical relationships in Conley et al. (2007), a 1 °C temperature increase would require a compensating nitrogen reduction of 20 kt of nitrogen. Conley et al. (2007) also suggested that regime shifts may have occurred and proposed that reoccurring large events of hypoxia may have a cascading effect in decreasing the oxygen concentrations. Similar consequences have been observed in Chesapeake Bay and the Gulf of Mexico (Conley et al. in press). Thus, for the oxygen conditions along the Swedish west coast not to deteriorate even further, additional measures to reduce nitrogen inputs must be taken to prevent regime shifts and counteract temperature increases. Because dissolved oxygen conditions are the most important regulatory factor for benthic animal communities, it is not surprising that little recovery of these communities from regional eutrophication has been observed along the Swedish west coast (Rosenberg & Nilsson 2005). Nonetheless, benthic community recovery has been observed where direct organic loading has been abated or when physical conditions allow reoxygenation of basins (Rosenberg et al. 2002). Recovery may be delayed by the recruitment of larvae of deeply burrowing, long-lived species that characterize the healthy community. However, if these populations can re-establish they will advance the reversal of eutrophication by increasing denitrification and the sequestration of P in the sediments.

5.3. Other Significant Drivers Affecting Responses 5.3.1. Climate variability and change Air temperatures in the Baltic Sea basin have already risen over the past century, increasing by approximately 1°C in the northern areas of the Baltic Sea basin and by around 0.7°C in the southern areas (HELCOM 2007). Consequently, the warming is larger than the global mean temperature increase of 0.75°C reported by the Intergovernmental Panel on Climate Change. The projections for future climate change in the Baltic Sea region, with all of their caveats and uncertainties, indicate that atmospheric temperatures will continue to rise during the course of the 21st century in every sub-region of the Baltic Sea region by 4-5°C in winter and in summer alike. For most regions and seasons it cannot be firmly established whether there will be an increase or a decrease in precipitation. The winter projections show a particularly wide range of magnitude from the models, although


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they generally point towards increased winter precipitation. Precipitation is likely to increase by 20 % in winter and no significant changes in summer (HELCOM 2007). This would result in a decrease of summer river flows, while winter flows would tend to increase. The catchments would most probably be affected by the combination of both decreased summer precipitation and increased evapo-transpiration (BACC Author Team 2008). Nutrient transport from the catchments into the Kattegat and Skagerrak region would most probably increase with consequences for primary production and oxygen consumption in deep waters. Higher nutrient loads in combination with higher temperature may result in less dissolved oxygen in the waters and stronger stratification. Both effects would have the potential to increase hypoxia and negatively affect benthic fauna. Stronger freshwater inflow into the northern Baltic could reduce salt-water inflow from the North Sea. This may increase the residence time of water in the region and also aggravates the overall critical situation of a eutrophied system. In its Ingen Övergödning report, the SEPA (2007) indicated that new studies suggest that the effects of climate changes on eutrophication in the Baltic Sea may not be as dramatic as earlier supposed and, in fact, may alleviate some of its effects. Eutrophication of lakes and rivers and streams is, on the other hand, expected to increase due to climate changes. While increases in runoff are projected, this will primarily affect loading of organic materials and phosphorus and not nitrogen. And finally, extreme precipitation and runoff events have a relatively short and often local effect in the transport and concentrations of nutrients. 5.3.2. Degraded state of the ecosystem In managing the Swedish west coast to achieve a state of no Zero Eutrophication, it must be acknowledged that ecosystems often do not immediately respond to reduced pressure. The marine ecosystem response might come as a delayed response, it might come as a non-linear response where improvement is first seen once a certain threshold value for the nutrient input has been reached, or there may be several thresholds separating multiple stable states (Andersen et al. in press). Such thresholds are difficult to forecast or even empirically determine and the changes in the ecosystem may be irreversible if permanent physical changes result, populations of functionally critical species are extirpated, or important top-down controls by large predators are not restored. Reductions in populations of large predators, such as cod, can result in trophic cascades that allow excess algal production even if nutrient loads are reduced (Casini et al. 2008). The different responses described above are believed to fall into different categories. Nutrient concentrations, phytoplankton and phytobenthos will most likely respond to nutrient input reductions as lagged responses, albeit with large differences in the lag times. The recovery of oxygen conditions and benthic communities is more likely a non-linear response where significant improvements resulting from reducing nutrient inputs are resisted due to strong positive


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biogeochemical feedbacks in the sediments. Recovery may only be possible if hypoxic conditions do not occur for several years such that a healthy benthic community characterized by deep burrowing species establishes and the strong feedback of nutrients to the water column is curtailed.


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6. Evaluation of The Swedish Strategy 6.1. The Objective of “Zero Eutrophication” The Zero Eutrophication objective is to attain nutrient levels in soil and water that do not adversely affect human health, the conditions of biological diversity or the possibility of varied uses of land and water within one generation (SEPA 2007). The evaluation panel actually feels more comfortable with the more literal translation “No Over-Enrichment” for Ingen Övergödning rather than “Zero Eutrophication” because eutrophication is a process, implying the increase over time of nutrient inputs that drives organic matter production, rather than a condition. It can be the case that an ecosystem anthropogenically-enriched with nutrients has stable or even declining nutrient levels and is still considered unacceptably degraded (Nixon 1995). Moreover, eutrophication can be a natural process, for example associated with ecosystem aging and maturation (Wetzel 1983). Nonetheless, we have used the SEPA translation of “Zero Eutrophication” throughout the report, assuming the term implies the objective as stated. As a practical matter, it is not feasible to return to a completely pristine state in which there is no increase whatsoever of nutrients emanating from human activities in the environment and still support human populations and the food production required to sustain it. The Zero Eutrophication objective, as stated, allows the reality of a non-pristine condition as long as human health, conditions of biological diversity or varied uses are not impaired. This requires, however, the definition of the point at which adverse effects on human health, biodiversity and designated uses due to nutrient enrichment is reached, i.e. when it becomes over-enrichment. This is not a simple task. Other regions seeking to reverse coastal marine ecosystem degradation due to eutrophication, such as the Chesapeake Bay region and northern Gulf of Mexico (Kemp et al. 2005, Boesch 2006) have struggled to define environmental conditions, load reductions, and management practices effective in attaining them. There is much that could be gained by global comparative analyses of these strategies and their effectiveness and limitations. 6.1.1. Interim targets and goals To meet the Zero Eutrophication environmental quality objective, interim targets were formulated in 1995 for accomplishments through 2010. The Swedish Environmental Protection Agency (2007) evaluated progress in achieving the interim targets, generally through 2005, projected the likely progress to 2010, and proposed new interim targets for 2020 (Table 6.1). In recognition of the slower than anticipated response of the marine ecosystem discussed in Section 5, the SEPA also proposed adjustment of the overall objective so that the reduction of the load of nutrients needed to ultimately reach the objective is reached by 2020, rather than the full realization of the environmental quality objective.


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In addition to these four targets, the SEPA strategy includes eight associated speci-fications. Those of particular relevance to achieving Zero Eutrophication are:

1) Atmospheric deposition of nitrogen compounds must not exceed the critical load for eutrophication of soil and water anywhere in Sweden.

2) Nutrient conditions at the coast and in sea areas shall be essentially the same as the conditions that prevailed during the 1940s, and the emissions of nutrients to sea areas shall not lead to eutrophication. This was amended in the 2007 assessment to “the emissions of nutrients to sea areas lie at a level that the nutrient conditions at coasts and in seas can reach conditions that prevailed in the 1940s.”

3) Swedish coasts shall meet the demands of good ecological status with regard to nutritive salts as defined in the EU Water Framework Directive.

Table 6.1. Status of achievement of the Zero Eutrophication objective and its associated interim targets (SEPA 2007).

Objective/ Target

Objective/Interim Target Set in 1995

Achievements by 2005 and outlook for 2010

Proposed 2020 Objective/Interim Target

Objective Nutrient levels in soil and water must not be such that they adversely affect human health, the conditions of biological diversity or the possibility of varied uses of land and water. Aim is for the environmental quality objective to be achieved within a generation

No clear changes in eutrophication conditions can be seen and the situation continues to be serious. By 2020 will it only be possible to create the prerequisites necessary to fulfill the objective in the long term.

The load of nutrients shall be reduced by 2020, so that the objective can be reached in the long term.

Target 1 By 2010, Swedish waterborne anthropogenic emissions of phosphorus compounds (phosphorus) to lakes, rivers and streams and coastal waters will have decreased by at least 20% from 1995 levels. The largest reductions will be achieved in the most sensitive areas.

Emissions diminished by 14% and a further 150 tonnes in reductions remained to reach the interim target.

Target 2 By 2010, Swedish waterborne anthropogenic emissions of nitrogen compounds (nitrogen) to sea areas south of the Åland Sea will have been reduced by at least 30% compared with 1995.

Emissions diminished by 24%, there remained 3,500 tonnes of further reductions in emissions to sea areas to reach the interim target.

The interim targets for 2010 were not based on what reductions are required to achieve the objective, but on determination of a reasonable reduction within the timeframe. New interim targets are proposed based on Sweden’s assignment of reductions required to achieve relatively unaffected conditions in the Baltic Sea as determined by the MARE model and jointly allocated by Baltic countries in 2007.

Target 3: By 2010 emissions of ammonia in Sweden will have been reduced by at least 15% compared with 1995.

Emissions fell by 15% and are expected to decline further.

Emissions of ammonia in Sweden will have been reduced by 13% compared with 2005 levels.

Target 4: By 2010, emissions of nitrogen oxides to air in Sweden will have been reduced to 148,000 tonnes.

In 2006 emissions were about 179,000 tonnes and are forecasted to drop to 154,000 tonnes by 2010.

Emissions of nitrogen oxides to air in Sweden will have been reduced to 130,000 tonnes.


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In general, the evaluation panel is favourably impressed with the scope of and progress in implementation of the Swedish strategy to achieve the Zero Eutrophication objective. Most of the interim targets set for 2010 either have been or are projected to be achieved by the SEPA estimations. This is highly unusual in comparison to many other national and regional commitments to combat marine eutrophication. For example, as its 2010 goal for implementation approaches, the Chesapeake Bay Program in the United States is very likely to fall short of its targets of reducing N and P loads by 48 and 53%, respectively, by 2010 (Kemp et al. 2005) – measures projected to achieve just over one-half of the reductions are currently in place (Chesapeake Bay Program 2008). In Denmark, an adaptive management strategy was pursued with additional measures initiated as interim assessments indicated that targets were not going to be achieved (Grant et al. 2006, Carstensen et al. 2006). The original targets for P (80% reduction) were soon achieved by reductions from point sources, whereas N reductions from diffuse sources were more complicated to address. However, it is anticipated that targets of 49% reduction in nitrogen inputs will be achieved (Grant and Waagepetersen 2003). Phosphorus removal has been implemented in all municipal sewage treatment plants and nitrogen removal in more than three-quarters of those in Sweden (Bernes 2005, SEPA 2007). As a result, nitrogen loads directly to the waters of the Swedish west coast have declined by 49% and phosphorus loads by 24% between 1995 and 2005, alone (Table 3.4). Swedish atmospheric emissions of nitrogen oxides and ammonia have clearly declined and some reductions in atmospheric deposition in the Kattegat and southern Sweden have been observed, even though this is also strongly affected by emissions from other countries. Total phosphorus fertilizer applications have declined by two-thirds and nitrogen fertilizer applications have declined as well, although less markedly (Bernes 2005). Despite these achievements, nutrient loads delivered by rivers to the west coastal seas have not obviously declined (Figure 3.4), even though nitrogen loadings from inland sources have declined by an estimated 19% (Table 5.1). This could be related to nutrient saturation in landscapes of the catchments or other lag-time factors (Bernes 2005), but some of this discrepancy could also be the result of overestimation of the nutrient source reductions. 6.1.2. Specific goals and strategies for west coast marine waters Taking a national approach to the Zero Eutrophication objective and to the associated targets and specifications is certainly understandable and has been effective to this point. Also, it is clear why achievement of the Baltic Sea nutrient reduction requirements and multi-national commitments has largely driven the Swedish strategy. However, achieving the Zero Eutrophication objective for the west coast seas, which are equally important to Sweden, merits special considerations and approaches. The manifestations of eutrophication in the western seas are different in important respects, the Baltic NEST decision support system (Savchuk et al. 2007, Wulff et al. 2007) does not effectively assess reductions


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required to alleviate the deleterious effects in these areas, the areas of largest agricultural production drain to the west, and there are different sets of nations that must be engaged (for example, to continue to reduce inputs from the Jutland Coastal Current). The evaluation panel was struck by the importance of two important effects of eutrophication in the western seas, the alleviation of which could form the basis for a specific regional objective: hypoxia in the Kattegat and the degraded vegetated habitats in coastal zone (rocky macroalgal communities, eelgrass and filamentous algal mats). Both of these have a clear link with eutrophication and have major consequences for living resources and coastal economies. Resolving uncertainties about the effectiveness of nutrient reduction measures affecting diffuse inputs and optimizing the mix of diverse measures that can be taken requires a catchment approach within the specific landscapes of southern Sweden. Such approaches are commonly pursued elsewhere to abate nutrient pollution, but the evaluation panel saw surprising little of catchment-based models and implementation strategies – although we may be just unaware. Catchment-based approaches facilitate the use of geographic targeting of resources and actions to achieve maximum effects on delivery of nutrients to the sea. 6.1.3. The transgenerational reality The proposed modification of the Zero Eutrophication objective to specify the achievement of nutrient load reductions by 2020 that are required to eventually achieve the environmental quality objective is an appropriate one. It is at once a practical recognition that improvements in the marine environment have been slow to materialize, an incorporation of emerging scientific theory concerning thresholds and state changes, and a steadfast commitment to the objective. As with the mitigation of climate change, this will challenge the understanding of the public, skill of scientists to understand poorly understood and dynamic changes, and the commitment of governmental institutions to address intergenerational problems. Given this acceptance that patience is required (Bernes 2005), it is incumbent on scientists and managers to: (1) ensure the accuracy of the load reductions required to eventually achieve the good environmental status sought; (2) verify that the nutrient load reductions are actually being achieved; and (3) identify “leading indicators” of state changes that indicate that recovery is happening for use in monitoring. 6.1.4. Climate change and other compounding forces It is increasingly apparent that “conditions that prevailed during the 1940s” in Sweden’s west coast seas cannot be realized in the future. Although it might be possible to achieve the nutrient throughputs and concentrations that existed then, the marine ecosystems will be different because of other human effects – for example, the long-term effects of overfishing and the influence of invasive species


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– and because of important changes in the climatic conditions. Some of these changes are difficult, if not impossible, to predict, however understanding of the climate changes that are occurring or are likely to occur during this century is growing rapidly (Bernes 2005, BACC Author Team 2008). The recent background report on the Zero Eutrophication objective (SEPA 2007) did a commendable job in forecasting the potential effects of expected changes in temperature and precipitation on eutrophication in Sweden. Its focus was mainly on the Baltic Sea and similar assessments are advisable for the west coast seas and coastal environments. In any case, future environmental management will have to establish targets and monitor changes that challenge the notions of restoration and recovery.

6.2. Measures and Their Implementation 6.2.1. Nitrogen controls are essential The previous expert evaluation of eutrophication in Swedish seas (Boesch et al. 2006) covered the nation’s marine environments, but focused heavily on the Baltic Sea and the coastal environments of the east coast, particularly around Stockholm. The west coast seas and coastal environments were only briefly covered. That panel was unable to resolve the ongoing debate in Sweden regarding the efficacy of controlling nitrogen versus phosphorus inputs to reduce eutrophication. While both the limnologists and oceanographers on the panel could agree that more effort was needed to control phosphorus inputs, the former members held the position that the brackish Baltic ecosystem behaved much like many freshwater lakes in a way that made reduction of inputs of nitrogen ineffective or worse, counterproductive. They posited that whenever nitrogen concentrations were reduced sufficient to limit phytoplantonic growth but light and phosphorus were in sufficient supply, N2-fixing cyanobacteria would proliferate and alleviate the nitrogen shortage. The oceanographers, on the other hand, were more familiar with estuarine and marine waters where phosphorus is remineralized, nitrogen often depleted and limiting, and limitation by either nutrient can occur over space and time scales. The oceanographers found evidence of such joint or alternating limitation in the annual dynamics of production in the Baltic Proper and in the response of coastal waters around Stockholm to reductions in phosphorus and nitrogen point sources. Shortly after that evaluation, a multinational group of Baltic scientists published a synthesis that explained how elevated levels of both nitrogen, fueling the spring blooms and hypoxia, and phosphorus, remobilized from internal loads to support blooms of N2-fixing cyanobacteria during the summer, both played a role (Vahtera et al. 2007). Controls of both nutrients are required, they argued, in agreement with the oceanographers. The SEPA (2006) considered the expert evaluation as well as other evidence in concluding that, although greater emphasis on phosphorus reductions was required, efforts to reverse eutrophication in the Baltic require both nitrogen and phosphorus reductions. The P versus N debate continues, as


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exemplified by the recent publication by authors including the same limnologists participating in the 2005 expert evaluation arguing for a phosphorus-only control approach in lakes and even in coastal waters and citing that evaluation to support their case (Schindler et al. 2008). The earlier expert evaluation panel (Boesch et al. 2006) did, however, conclude unambiguously that reduction of nitrogen inputs to the waters of the Swedish west coast was required to address eutrophication problems there as there were no apparent risks of N2-fixing cyanobacterial blooms to alleviate nitrogen deficiency. The present evaluation reaffirms that conclusion and provides more detailed analysis of the spatio-temporal interplay of the two nutrients and an explanation of why N2-fixing cyanobacteria are not prevalent in the west coast seas and not likely to become so if nitrogen levels decline. 6.2.2. Phosphorus reductions produce local benefits At the same time, the earlier evaluation (Boesch et al. 2006) noted that reductions of phosphorus inputs would also have positive results in west coast waters because of the presently high levels of anthropogenic loading of both nutrients, but only if accompanied by nitrogen reductions. The present evaluation expands on that conclusion by considering phosphorus limitation seasonally and in the vicinity of nitrogen-rich riverine effluents. This explains why phosphorus removal in sewage treatment works discharging to the Göta Älv estuary produce local water quality and ecological benefits, while nitrogen removal would produce no apparent local benefits, but modest distant benefits (Erlandsson and Johannesson 2005; Isæus et al. 2005, Garde et al. 2008). By the same token, continued efforts to constrain or reduce point sources of phosphorus and improve poorly performing household waste treatment systems, and the recent discontinuance of phosphate-based detergents in Sweden would be expected to result in localized improvements within the coastal zone, but are unlikely to result in greater phosphorus limitation in the open Kattegat or Skagerrak. 6.2.3. Greater reductions of agricultural and atmospheric loads are needed With very substantial reductions in phosphorus and nitrogen loads from Swedish sewage treatment works having been achieved, most of the remaining reductions required to achieve the Zero Eutrophication objective and its interim targets must come from diffuse land-based sources, particularly agriculture, and atmospheric deposition of nitrogen (Table 5.1, SEPA 2007). Assessment of the effectiveness of the efforts to control these sources is beyond the scope of this evaluation, however it is worthwhile to note that while the inputs of phosphorus fertilizers have declined greatly the use of nitrogen fertilizers is as intense on a per-hectare basis as it was in the 1970s. Most of the decline in nitrogen leaching in agriculture has been due to a contraction in the area of arable land and the agricultural nitrogen surplus (the nitrogen applied in fertilizers and manure less the nitrogen removed in crops) averages 38% (Bernes 2005). This suggests that greater reductions in the nitrogen


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leaching from agricultural lands are possible and, in fact, anticipated as a result of both implementation of control measures and changes in the global agricultural economy (SEPA 2007). The SEPA lists an array of proposals for further measures and controls in agriculture: revise the agricultural action programme, reduce soil cultivation, make permanent or increase area used for catch crops, establish 100% of the land in sensitive areas as grassland, establish riparian strips for erosion-sensitive land, increase regionalization of wetland support, use ponds as phosphorus traps, lime filter drains, regulate drainage, and reduce the phosphorus content of animal feedstuffs. Further reductions in atmospheric emissions of ammonia and nitrogen oxides are anticipated and more aggressive European actions to reduce nitrogen oxide emissions are under negotiation (SEPA 2007). 6.2.4. Multi-national cooperation is required A significant portion of anthropogenic nutrients that affect the Kattegat and Skagerrak do not originate from Sweden but are transported by currents and outflows from the lower Baltic Sea, from Denmark through the Belt Seas and across the Kattegat, and from the catchments of the large rivers of western Europe, and via atmospheric transport from over an even larger footprint. Clearly Sweden’s actions alone are insufficient to achieve the load reductions and good environmental status required to meet the Zero Eutrophication objective. Sweden is party to two important regional marine conventions, OSPAR and HELCOM, which provide mechanisms to engage other nations in the necessary cooperative efforts to reduce nutrient inputs. It is in Sweden’s best interest to continue its leadership role as an early adopter of effective nutrient controls and in scientific research and assessment (e.g. the Baltic NEST decision-making tool). Multiple directives from the European Union, including the Nitrate Directive, Habitat Directive, Water Framework Directive and the Marine Strategy Directive, provide both mandates and impetus to achieve the Zero Eutrophication objective. However, it is important that Sweden takes a proactive role in the implementation of these directives to ensure common European standards for achieving ecological quality objectives that are consistent with the national Zero Eutrophication objective.

6.3. Integration of Monitoring, Modeling and Research for Adaptive Management Achieving the Zero Eutrophication objective in the Swedish western seas requires a long-term commitment, faced with uncertainties regarding the effectiveness of the measures taken and the response of marine ecosystems to nutrient load reductions. This calls for an adaptive management approach to optimize actions, forecast and verify outcomes, and adjust objectives (Boesch 2006). Key technical elements of adaptive management are models of system responses, strategic and continuous monitoring, and regular integrated assessment. Some components of these elements exist for the western seas, but all three elements


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must be bolstered and coordinated to support adaptive management effectively. Various hydrodynamic and ecosystem models exist, but there are not decision-support models analogous to the Baltic NEST model that has been recently used in optimally allocating nutrient load reductions in the Baltic Sea. Monitoring of many environmental variables is currently conducted by the SMHI, SEPA, county authorities, and the neighboring countries of Denmark and Norway, however it is not focused on the “leading indicator” approach described above, nor is it adequately coordinated and integrated. Concerns were voiced that monitoring resources may be redirected to address Water Framework Directive requirements of geographically specific recipient waters to the detriment of monitoring needed to address the pervasive eutrophication in the region. There are always the fiscal pressures to reduce investments in long-term monitoring, which are not seen as providing near-term results. Continued advanced research is also important, particularly on poorly known processes that are critical to understanding ecosystem responses to reductions in nutrient loading. Although there are various national assessments by Sweden, Denmark and Norway and periodic assessments for OSPAR, the evaluation panel found it difficult to pull together disparate information resources across national boundaries and from the catchments to the sea in order to address its charge during the limited time it had available. There is clearly a need for ongoing integrated assessments with periodic reporting of results in order to support adaptive management. By integrated assessments, we mean seamless analysis of the status and trends in the marine ecosystem; the processes responsible for changes; nutrient source delivery from catchments, the atmosphere, and point sources; measures taken to reduce nutrient inputs; comparisons of predictions and outcomes, and management recommenda-tions for achieving and accelerating the attainment of the environmental quality objectives. The evaluation panel heard of the recent Government decision to create a “marine research institute” at the University of Göteborg with particular emphasis on analyses to support policy formulation and encourages that initiative to address the need for such integrated assessment.

6.4. Transparency and Accountability There are many reasons for regular reporting of environmental conditions, nutrient loadings, and actions being taken to achieve the Zero Eutrophication objective and for periodic assessments of the effectiveness of actions being taken toward that end. Accountability to the elected government and citizens of Sweden, other parties to international agreements such as OSPAR and HELCOM, and the European Union is foremost among them. This is best served by processes, report and websites that are appropriately transparent so that other scientists and technical analysts, and even citizens, may access and understand the underlying data and analyses. The evaluation team found the Zero Eutrophication review report helpful in that regard because it provided or cited more extensive background information


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used to arrive at its conclusions and recommendations for new interim targets. Still, in this day of internet connectivity there are opportunities for even greater transparency. Finally, the evaluation team found the SEPA-published book Change Beneath the Surface: An In-Depth Look and Sweden’s Marine Environment truly exemplary as an attractive, accessible, scientific-based, and honest communications tool to a broader public audience.


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7. Findings and Recommendations Based on this evaluation of the physical setting, nutrient sources and trends, ecosystem responses, experience related to reversing eutrophication and the Swedish strategy, the Panel draws the following findings and recommendations:

1) Sweden’s progress in assessing and combating the deleterious over-

enrichment of the seas along its West Coast is commendable from both an international and a regional perspective. However, despite the fact that nutrient loadings and environmental concentrations have been reduced, the ecosystems have yet seen little improvement and in some cases, for example the proliferation of filamentous algae, seem to be undergoing progressive decline. While some of this recalcitrance is related to the time lags for ecosystem recovery, greater reductions of human nutrient inputs will be required in order to meet the Zero Eutrophication objective.

a. Because substantial reductions in nutrient loading from point sources have already been achieved and the opportunity for further reductions from these sources are limited, reductions in diffuse-source and atmospheric emissions from Sweden will be required.

b. Reducing Swedish nutrient emissions alone will be insufficient to achieve the Zero Eutrophication objective because a large share of the human nutrient inputs emanates from Denmark or are brought into the West Coast seas by flows from the Baltic and North seas and by atmospheric transport. Parallel efforts to reduce nutrient emissions from Baltic and North Sea nations will be required.

2) There is compelling evidence that both nitrogen and phosphorus are

important contributors to over-enrichment of the West Coast seas and management strategies should address both nutrients. In contrast to the Baltic Sea, where there remains some scientific disagreement about the efficacy of reducing nitrogen because of the prevalence of seasonal N2 fixation by cyanobacteria, there is abundant evidence and no significant disagreement that reductions of nitrogen loads are essential for reversing regional eutrophication. Further reductions in phosphorus loadings are likely to produce mostly localized benefits.

3) There are indications that the ecosystem is beginning to respond to nutrient

controls, as evidenced in measured declines in point and diffuse source emissions, atmospheric deposition, environmental nutrient concentrations and phytoplankton biomass and production. However, it is difficult to predict to what degree and how rapidly ecosystem recovery will proceed.

a. Achievement of environmental objective of Zero Eutrophication will be slow and be manifest by gains and setbacks. Complete recovery of some pre-existing conditions, such as bottom oxygen levels, may not be possible.


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b. Progress is delayed because nutrients are retained and recycled in ecosystem and biological processes (e.g. abundant filamentous algae and the paucity of deeply burrowing benthos) create positive feedbacks that make the ecosystem resistant to recovery.

c. Kattegat hypoxia, recovery of coastal vegetation (Zostera and algal mats) and benthic animal communities should be considered key ecological indicators of ecosystem recovery.

4) Other ecosystem interactions also affect recovery, including fishery

declines and associated top-down controls. In addition, climate change effects may make achievement of objective more difficult by increasing runoff and stratification, warming temperatures (affecting algal growth and composition), and altering boundary conditions with the North and Baltic seas. These will require periodically redefining achievable endpoints for ecosystem recovery.

5) Better coordination of monitoring, research and assessment is required in

order to pursue the Zero Eutrophication objective through adaptive management. Elements include:

a. Improved effectiveness and efficiency of local, subregional (counties), national and international monitoring to assess key indicators of pressures and responses.

b. Greater attention to quality assurance and standardization of techniques, particularly for biological measurements (e.g. benthic vegetation), and to the development of “leading indicators” of ecosystem recovery.

c. An appropriate balance in monitoring efforts devoted to achieving balance between national strategy and EU-directives.

d. Better integration of modeling, monitoring and research, including facilitating the use of monitoring results in research, supporting research on critical processes to help interpret monitoring, and integrated assessments using both models and monitoring results. More effective collaboration among SEPA, SMHI and universities in research, modeling, monitoring and assessment. The newly established “marine research institute” at the University of Göteborg holds promise in this regard.

e. Contextual relationships with broader ecosystem-based management efforts, particularly related to fisheries and agricultural landscapes.

6) Research and syntheses of knowledge on the causes and consequences of

eutrophication of West Coast seas have provided an adequate basis for determining the requirements for reversing eutrophication and the benefits of achieving the Zero Eutrophication objective. In many ways this science has led the world related to understanding coastal eutrophication. While there is a solid basis on which to act, strategic research designed to narrow


63

key uncertainties would be helpful and should be pursued. Among those critical questions which require more research for better understanding and quantification are the following:

a. The recycling of nitrogen and phosphorus, particularly the sediment-water column coupling, and the biogeochemical processes that regulate recycling. These affect nutrient limitation of phytoplankton production, rates of nutrient removal from the ecosystem (particularly with regard to denitrification), release of internally stored loads, positive feedbacks and associated thresholds that limit the extent and rate of ecosystem recovery.

b. The role of the Jutland Coastal Current in supplying nutrients to the West Coast seas. This requires further resolution, because of conflicting views in the literature, in order to determine the load reductions from North Sea-draining rivers required to achieve the Zero Eutrophication objective.


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Eutrophication of the Seas along Sweden’s West Coast

The Swedish Environmental Protection Agency has con-

vened a group of international experts for the evaluation

of eutrophication in the seas and coastal environments

along the west coast of Sweden.

Five highly qualified scientists have gone through the

scientific material on the situation in the Danish Sounds,

the Kattegat and the Skagerrak and reported their

findings to the Agency. The expert group has evaluated

the measures taken so far to achieve the environmental

quality objective “Zero Eutrophication” and has recom-

mended future strategies to counteract eutrophication in

the marine areas concerned.

report 5898

SWEdiSh Epa

iSbn 978-91-620-5898-2

iSSn 0282-7298

Swedish epa SE-106 48 Stockholm. visiting address: Stockholm - valhallavägen 195; Östersund - Forskarens väg 5 hus Ub; Kiruna - Kaserngatan 14. Tel: +46 8-698 10 00, fax: +46 8-20 29 25, e-mail: [email protected] internet: www.naturvardsverket.se orders ordertel: +46 8-505 933 40, orderfax: +46 8-505 933 99, e-mail: [email protected] address: Cm Gruppen, box 110 93, SE-161 11 bromma. internet: www.naturvardsverket.se/bokhandeln

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