Eutrophication of the Seas along Sweden`s West

Transcription

Effective July 1, 2011, this publication
is handled by the Swedish Agency for
Marine and Water Management.
Telephone +46 (0)10 698 60 00
[email protected]
www.havochvatten.se/publications
Eutrophication of the
Seas along Sweden’s
West Coast
report 5898 • november 2008
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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Phone: + 46 (0)8-505 933 40
Fax: + 46 (0)8-505 933 99
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Internet: www.naturvardsverket.se/bokhandeln
The Swedish Environmental Protection Agency
Phone: + 46 (0)8-698 10 00, Fax: + 46 (0)8-20 29 25
E-mail: [email protected]
Address: Naturvårdsverket, SE-106 48 Stockholm, Sweden
Internet: www.naturvardsverket.se
ISBN 978-91-620-5898-2.pdf
ISSN 0282-7298
© Naturvårdsverket 2008
Digital publication
SWEDISH ENVIRONMENTAL PROTECTION AGENCY
Report 5898 • West Coast Eutrophication
Contents
1 INTRODUCTION
5
2 PHYSICAL SETTING
2.1 Geography and Bathymetry
7
7
2.2 Circulation and Water Masses
8
2.3 Swedish Coastal Waters
12
3 NUTRIENT SOURCES AND TRENDS
3.1 Sources
13
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
4.1 Phytoplankton Production
26
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
5.1. Effects of Countermeasures Taken
43
43
5.1.1. Swedish sources
43
5.1.2. Transboundary sources
45
5.2. Responses to Reductions in Nutrient Inputs
5.2.1 Nutrient concentrations and ratios
46
46
3
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
6.1. The Objective of “Zero Eutrophication”
52
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
4
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
5
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.
6
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 topographically 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
7
(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
Figure 2.2. General circulation in the North Sea
continues northeast along the
(OSPAR Commission 2000).
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 35 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
8
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
Figure 2.3. Modelled flux of water through Skagerrak (across a transect
between Oksøy in Norway and Hanstholm in Denmark). Time series are
Kattegat surface
for mean flux for the 1st and 4th quarters of the year from 1955 to 2004. The
water (KSW, which
unit of the flux is Sverdrup (106 m3 s-1) and the negative sign indicates flux
contains the Baltic
into Skagerrak. Data obtained with the NORWECOM model (Skogen and
Søiland 1998) driven by archived meteorological data for the modeled time
outflow), water
period.
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
9
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.
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).
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
10
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 shortterm 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 intermediate 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.
11
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.
12
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
Agricultural
land
Point sources
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.
13
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
14
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).
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).
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 lessdense 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).
15
Figure 3.2. Mean nitrate concentrations in January–April at Helgoland in the German Bight
(Aure and Magnusson 2008).
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
16
likely change to one with predominant N limitation as nitrogen is depleted and
phosphorus is efficiently recycled.
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.
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.
17
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).
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).
18
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.
120000
6000
TN
5000
TP
80000
4000
60000
3000
40000
2000
20000
1000
0
1988
1990
1992
1994
1996
1998
2000
2002
2004
2006
Figure 3.5. Trend in nutrient loads from Danish rivers into the Kattegat from 1989 to 2006.
(http:/www.dmu.dk; Ærtebjerg 2007).
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.
19
0
2008
Total phosphorus (tons yr-1)
Total nitrogen (tons yr-1)
100000
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).
24
0.9
TN
TP
0.8
0.7
16
0.6
12
TP (µg L-1)
TN (µg L-1)
20
0.5
0.4
8
1970
1975
1980
1985
1990
1995
2000
2005
2010
Figure 3.6. Arkona Sea nutrient concentrations (µg L-1) (SMHI and Danmarks Miljøundersøgelser data).
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
20
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).
21
Total nitrogen (µmol/l)
40
30
20
10
0
1980
1985
1990
1995
2000
2005
2010
3,0
2,5
Total phosphorus (µmol/l)
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 concentrations 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 concentrations 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.
2,0
1,5
1,0
0,5
0,0
1980
1985
1990
1995
2000
2005
2010
Figure 3.7. Trends in TN (top) TN (bottom) and since 1982 in surface
waters (<15 m) at Anholt in the Kattegat (SMHI data).
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).
22
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).
23
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
Fig 3.9. Budget of biologically active nitrogen for the Kattegat (units kt yr-1)
through the Belt Sea
(Spokes et al. 2006).
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.
24
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.
25
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
14
C 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 phytoplankton 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
26
(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 phytoplankton 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.
27
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 (Ndominated) 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.
28
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 nearundetectable 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.
29
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 Nlimited 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
30
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
31
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 ecologicallyinsignificant 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
32
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 permanent stratification and strong fronts. These are ideal conditions for the establishment 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
33
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)
34
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).
35
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
Figure 4.2. Oxygen depletion near the sea bed in the
varying temperatures of Skagerrak
southern Kattegat (blue: <4 mg l-1, red: <2 mg l-1) during
September in an extreme year (2002, top) and a normal
surface water during this time of the
recent year (2006, bottom). Source: NERI.
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
36
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 intermittent 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 1015 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
37
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 improvement 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.
6
4
-1
oxygen (ml l )
5
3
2
1
0
1900
1910
1920
1930
1940
1950
1960
1970
1980
1990
2000
2010
Figure 4.3. Annual mean oxygen concentrations in the Gullmarfjorden with observations weighted according to their
month of sampling (SMHI data).
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,
38
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.
Figure 4.4. Estimated sedimentation in the Gullmar Fjord 1950 to 2007
(O. Lindahl presentation).
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
39
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 benthichabitat 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)
40
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 biogeochemical 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 faunallydepauperate 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
41
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).
42
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).
43
Table 5.1. Estimated anthropogenic loads to the western seas between 1995 and 2005 (SEPA 2007).
Anthropogenic P emissions (t yr-1)
Total diffuse
Point
emissions
Total
2005
1995
2005
1995
2005
1995
2005
10
10
40
40
60
40
100
80
60
50
410
370
350
280
760
650
40
20
120
105
450
340
980
835
Arable land
Logging
1995
2005
1995
2005
Öresund
30
30
<5
<5
Kattegat
340
310
10
10
70
80
<5
<5
10
5
80
85
440
420
10
10
80
65
530
495
Skagerrak
Total
Difference %
4.5%
Urban runoff
0.0%
1995
18.8%
6.6%
24.4%
Anthropogenic N emissions (t yr-1)
Inland sources
Point
emissions to
the sea
Total
1995
1995
1995
2005
2005
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
900
900
700
400
1,600
1,300
27,900
22,600
5,500
2,800
33,400
25,400
Skagerrak
Total
Difference %
19.0%
49.1%
24.0%
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
44
14.8%
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
Figure 5.1. Estimated nitrogen loading to the Belt Seas of Denmark
Kattegat (Figure 3.5). At least
(Conley et al. 2007).
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
Figure 5.2. Changes in the concentration of forms of nitrogen
in atmospheric deposition at Anholt, outer Kattegat (Ærtebjerg
have resulted in a substantial
2007).
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).
45
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.
46
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
47
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
48
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. According 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 reestablish 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
49
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 evapotranspiration (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
50
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.
51
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.
52
In addition to these four targets, the SEPA strategy includes eight associated specifications. 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).
Proposed 2020
Objective/Interim
Target
Objective/
Target
Objective/Interim Target Set in
1995
Achievements by 2005
and outlook for 2010
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.
53
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
54
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. Catchmentbased 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
55
– 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, N2fixing 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
56
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
57
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 erosionsensitive 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
58
must be bolstered and coordinated to support adaptive management effectively.
Various hydrodynamic and ecosystem models exist, but there are not decisionsupport 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 recommendations 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
59
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.
60
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 overenrichment 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 diffusesource 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.
61
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
62
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 sedimentwater 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.
63
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78
Eutrophication of the
Seas along Sweden’s
West Coast
report 5898
swedish epa
isbn 978-91-620-5898-2
Issn 0282-7298
The Swedish Environmental Protection Agency has convened 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 recommended future strategies to counteract eutrophication in
the marine areas concerned.
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

Eutrophication of the Seas along Sweden`s West

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