Aquatic Ecology 38: 587–598, 2004. © 2004 Kluwer Academic Publishers. Printed in the Netherlands.
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Long-term decline and recent recovery of Fucus populations along the rocky shores of southeast Sweden, Baltic Sea Jonas Nilsson*, Roland Engkvist and Lars-Eric Persson University of Kalmar, Department of Biology and Environmental Science, S–391 82 Kalmar, Sweden; *Author for correspondence (e-mail:
[email protected]) Received 13 January 2003; Accepted in revised form 29 March 2004
Key words: Environmental monitoring, Eutrophication, Fucus serratus, Fucus vesiculosus
Abstract The Fucus populations on rocky shores along 300 km of the coastal waters of southeast Sweden in the Baltic proper have been studied since 1984 at a number of fixed sites as part of monitoring programmes. This paper describes changes in distribution and abundance of F. vesiculosus and F. serratus during the period 1984–2001. Sheltered sites showed a consistent temporal and spatial pattern of Fucus spp. distribution over a coastline of 300 kilometres. The depth penetration and abundance of Fucus spp. increased during the 1980s. Around 1990 the development reversed as a consequence of grazing and in 1997 many sites were almost devoid of Fucus spp. Since 1998 both abundance and depth penetration have increased again, possibly as a result of local measures against eutrophication. Exposed sites, on the other hand, lost their Fucus populations at the beginning of the 1990s, and they have not recovered. Extended field studies lead us to deduce that the fixed sites referred to above were representative of the Fucus populations in the area investigated. Major declines, both at sheltered and exposed sites, are attributed to grazing by the isopod Idotea baltica. The populations of I. baltica may have been favoured by the continuing eutrophication of the Baltic, a series of mild winters in the 1990s, and a contemporary decline in some potential predators.
Introduction Only a few perennial marine macroalgae have adapted to the low salinity conditions in the Baltic Sea. On rocky shores along the Swedish coast of the Baltic proper, two of the most conspicuous macroalgae species are Fucus vesiculosus 共L.兲 and Fucus serratus 共L.兲. Both these fucoids are belt forming and, in monocultures or in mixed stands, can cover large areas from a few decimetres below the surface to a depth of several metres 共Kautsky et al. 1992; Malm et al. 2001兲. The large-scale geographic distribution of the two fucoids in the Baltic is determined by salinity 共Russell 1988兲. F. vesiculosus occurs from the southern Baltic to the Gulf of Bothnia and the inner part of the Gulf of Finland. F. serratus is only common along the shores of southeast Sweden, although
scattered populations have been found up to the middle of the Baltic proper 共Malm et al. 2001兲. The degree of mixing between the two species is correlated with wave exposure 共Malm 1999兲. These autogenic ecosystem engineers give the otherwise barren rocky bottom a structure of great complexity 共cf. Jones et al. 1994兲, which serves as shelter and nursery for many invertebrate and fish species. The Fucus communities have an important ecological function in the littoral ecosystem, and many authors have stressed the significance of this biotope 共e.g., Haage 1976; Kangas et al. 1982; Kautsky et al. 1992兲. A profound decline in abundance and depth distribution of F. vesiculosus has been reported from several areas around the Baltic Sea during recent decades. In the northern Baltic Sea, the deepest occurrence of F. vesiculosus thalli has changed from
588 11.5 m in 1943/44 to 8.5 m in 1984, and the depth of maximum development has decreased from 5–6 m in 1943/44 to 3–4 m in 1984 共Kautsky et al. 1986兲. At the end of the 1970s and the beginning of the 1980s, F. vesiculosus disappeared from large areas of the southwest of Finland 共Haahtela 1984; Mäkinen et al. 1984; Rönnberg 1984兲 and southern Finnish coastal waters 共Kangas et al. 1982兲. Similar negative developments were also reported from coastal areas in Poland 共Plinski and Florczyk 1984兲, Estonia 共Martin 2000兲 and Sweden 共Lindvall 1984; Kautsky et al. 1992; Engkvist et al. 2000兲. The ultimate cause of these changes is commonly ascribed to the effects of eutrophication and pollution 共Schramm 1996兲. One proximate factor, grazing by the isopod Idotea baltica Pallas 共Engkvist et al. 2000兲, has never been evaluated in conjunction with long-term field observations of Fucus populations. Scientists have often failed to demonstrate a single underlying factor responsible for synchronized fluctuation patterns in populations over large geographic areas. Population dynamics may be synchronized by climate, so a holistic approach is often used, rather than focusing on local weather or oceanographic parameters such as precipitation, temperature and salinity. Several recent studies have been reported on the impact of large-scale climatic forcing on different ecological systems 共reviewed in Stenseth et al. 2002兲. The North Atlantic Oscillation 共NAO兲 is frequently considered, as it is the dominant signal of interannual variability in atmospheric circulation for Northern Europe 共Hurrel 1995兲. NAO is highly correlated with weather effects 共Marshall et al. 2001兲, and in the Baltic it may influence factors such as salinity, water temperature, sea ice severity and sea level fluctuations 共Hänninen et al. 2000兲. Koslowski and Loewe 共1994兲 demonstrated that winter ice severity in the Western Baltic depends on the NAO winter index. Periods of weak ice production are strongly correlated with the occurrence of strong westerlies 共high index兲, and strong ice production with weak westerlies 共low index兲. In the Baltic, mild winters mean there are no or minimal ice scraping effects in the filamentous algal zone 共Kiirikki and Lehvo1997兲. Mild winters could also lead to increased survival and a subsequent increased recruitment of benthic crustaceans 共Beukema 1992兲. Previous studies of changes in Fucus populations in the Baltic have often compared data from sites that have been revisited only once or twice. The time elapsed from the first visit varies from a few years
共e.g., Kangas and Niemi 1985; Rönnberg et al. 1985兲 to several decades 共Kautsky et al. 1986; Rönnberg and Mathiesen 1998兲. Observations at intervals of several years or more could easily miss both large temporal and spatial variations, and any conclusions about development, history and causalities could accordingly be incorrect. In Sweden, monitoring of Fucus populations has been carried out on a national basis since 1993 in the Askö area of the northern Baltic proper 共e.g., Kautsky 2001兲, although even in the early 1980s a regional coastal environmental monitoring programme was initiated in the county of Kalmar along the Swedish part of the Baltic proper. One part of the Kalmar County programme was annual monitoring of macroalgal communities, paying special attention to documenting the status and trends of abundance and vertical distribution of Fucus populations at a number of fixed sites. A similar programme was started in the county of Blekinge in 1990. From 2001 and on, 27 sites were visited annually along a coastline of over 300 km. Data from these surveys have been used in part to verify hypotheses about grazing as a structuring factor of Fucus populations in the Baltic Sea 共Malm et al. 1999; Engkvist et al. 2000兲. Besides these fixed sites, other sites in surrounding areas were visited occasionally from 1983 to 1993. Seventy-two of these were revisited in 2000. This paper: 共1兲 presents a continuous data series on regional distributional changes in F. vesiculosus and F. serratus for the period 1984–2001; 共2兲 tests the extent to which Fucus data from fixed transects in the environmental monitoring programmes are representative of surrounding coastal areas; and 共3兲 evaluates the influence of bottom-up effects 共eutrophication and climate兲 and top-down effects 共fish predation and grazing兲 as potential causes of the observed changes.
Materials and methods Study area and fixed sampling sites Hard bottom surveys were conducted annually in September–October at fixed sites in Blekinge and Kalmar Counties along the Swedish coast of the Baltic proper 共Figure 1兲. The sites are oriented in all directions and include both sheltered and wave-exposed shores. However, exposed sites are present only in Blekinge County. In order to reflect conditions along a wide section of the coast, the chosen sites are lo-
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Figure 1. Baltic coast in southeast Sweden. Fixed sheltered sites are indicated by filled circles and fixed exposed sites by filled triangles. Approximate location and number of revisited sites are indicated by unfilled symbols.
cated in areas where they are not directly affected by point sources. The longest time series comprises 11 sheltered sites visited every year from 1984 to 2001. From 1990–1991, a finer sampling grid was adopted and previously unmonitored areas were added to the programme. In 1990 sampling started in the southern part of the area, Blekinge County, with five exposed and four sheltered sites, and in 1991 seven sheltered sites were added in Kalmar County. To maintain the long term perspective of the study and still take advantage of the increasing quantities of data, analyses were conducted of three overlapping periods: 1984– 2001, 1990–2001 and 1991–2001. The exposed sites
共n ⫽ 5兲 were visited only every second year during 1990–1998, and once a year thereafter. Fixed sites were graded according to wave exposure with fetch as variable 共Figure 1兲. Effective fetch describes the average distance in kilometres that a wave can collect energy before it reaches the shore 共Håkansson 1981兲. The index was calculated from information from nautical charts 共1:50 000兲, meaning that even smaller islands, skerries and shallows could be taken into account. Effective fetch 共Lf兲 was calculated from the formula Lf ⫽ 共⌺i cos␥i 兲 / 共⌺ cos␥i兲, where i is the distance 共in km兲 from the given site to the nearest wave-breaking site which was measured
590 for every deviation angle ␥i , where ␥i ⫽ 0°, ⫾ 6°, ⫾ 12°, ⫾ 18°, ⫾ 24°, ⫾ 30°, ⫾ 36°, ⫾ 42°. Nineteen sites had a fetch below 10, three lay between 30 and 40, one scored 64 and four scored between 90 and 100. The first 22 were labelled sheltered and the last five, all situated in the southern part of the area, as exposed. Investigations were conducted by scuba diving – from 1984–1989 following a range line and from 1990 following a measuring tape – along fixed transects from the shore to a depth at which no hard substrate remained. A diver documented the profile with a still camera or video camera and a second diver recorded the cover and recruitment of Fucus. Grazing intensity was measured on a 3-point scale where 0 denoted few grazing marks and 2 denoted obvious grazing marks in more than 50 % of remaining Fucus spp. plants. In the analysis all data on F. vesiculosus and F. serratus are pooled and are hereafter referred to as Fucus spp. The Fucus spp. cover was estimated using a percentage scale: 0, 5, 10, 25, 50, 75 and 100%. All estimations and measurements were made in a 3–5 m corridor on both sides of the measuring tape. An interval where Fucus spp. covered 25% or more of the substrate was considered a belt. The grazer community was sampled according to Engkvist et al. 共2000兲 and Engkvist et al. 共2004兲. F. vesiculosus with holdfasts was detached and carefully placed in 1 mm net bags. To measure the density of I. baltica, samples were analysed in terms of weight of Fucus and numbers of animals. Depth was measured with a digital depth gauge 共Aladdin or Mares兲 having an accuracy of ⫾ 0.1 m. All measurements were adjusted to mean water level using daily water Table data 共Swedish Meteorological and Hydrological Institute, SMHI兲. Revisited sites Since 1984 a large number of sites have been visited occasionally, mostly as checkpoints for small, potentially harmful outlets. On each visit the extension of the Fucus belt was measured by scuba divers. The visits were clustered around groups of years, namely 1983–1985, 1987–1988, 1990–1991 and 1992–1993. During 2000, seventy-two of these sites were revisited 共Figure 1兲 and the extension of the Fucus belt was measured again. Six of the revisited sites were considered wave-exposed as they scored a fetch of more than 75. The remaining 66 sites scored 10 or less and were categorized as sheltered.
To be able to analyse changes in the distribution of Fucus between year groups with standard statistics 共ANOVA兲, each site within a year group was considered as an independent sample from one population of Fucus spp. From each year group, including 2000, six sheltered sites were selected randomly, so that a single site occurred only once, so that we could examine whether the fixed transects provided a good indication of conditions in a larger area. As for the fixed sites, revisited exposed sites were present only in Blekinge County 共Figure 1兲. Six exposed sites were visited between 1987 and 1990 and they were all used as replicates for comparison with the results from the revisit in 2000 共paired t-test兲. Data analysis The statistical analysis used two variables of Fucus spp. distribution. Depth distribution of the Fucus belt was calculated by subtracting the upper depth limit from the lower, meaning that a belt that had disappeared scored zero. From 1990, the first year a measuring tape was used, an index was calculated to describe the total cover of Fucus spp. at each fixed site. The length of an interval was multiplied by the percentage cover within the interval. The indexes for all intervals were then summarized in a Fucus cover index for each site. Spearman Rank Order Correlation was used to decide if there was covariation between observed grazing intensity and change in Fucus spp. cover between two consecutive field observations in any of the sites. For sheltered sites data were used for the period 1990–1994 from a total of 22 sites and a total of 91 observations 共n ⫽ 91兲, and for the period 1998–2001 from a total of 22 sites 共n ⫽ 77兲. From wave exposed sites data were used for the period 1991–1994 from a total of 5 sites and these sites were observed a total of 16 times 共n ⫽ 16兲. Oceanographic and meteorological data were provided by the Swedish Meteorological and Hydrological Institute 共SMHI兲 and were used to evaluate changes in Fucus cover. The long-term development of the Fucus belt at sheltered sites for the period 1984–1996 was correlated with the NAO winter index 共December through March兲 with different time lags 共Hurrell 1995; Stenseth et al. 2002兲. The correlation period was chosen because the Fucus populations was almost extinct at many sites in 1996. Any change in growth conditions thereafter can only take effect after prolonged time lags as F. vesiculosus eggs travel only 0.5–2 m from the mother plant 共Serrão et al. 1996兲,
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Figure 2. Yearly mean values of the vertical distribution of Fucus vesiculosus belts at 11 sheltered sites 1984–2001. Values of the winter NAO index are superimposed shifted one year to the right 共see text兲. Figure 3. Yearly mean values of Fucus spp. Cover index at 16 sheltered sites 1990–2001.
and it will take some years before a new Fucus belt is established. The statistics were calculated using SPSS for Windows 10.01 and Statistica 5.5. When repeated measures 共RM兲 ANOVA was used, data were checked for sphericity and where necessary corrected according to Greenhouse and Geisser 共1959兲 or replaced with a multivariate procedure. All measurements of dispersal are ⫾ 1 SE.
Results Fixed sheltered sites The longest time series covered 1984 to 2001 and consisted of 11 sheltered sites from the northern part of the area, Kalmar County 共Figure 1兲. The extension of the Fucus vesiculosus belt was extremely variable over time 共Figure 2兲 but was consistent between sites 共RM ANOVA, p ⬍ 0.001兲. In 1986 and 1989 the belt was wider than in 1996–1999 共post hoc according to Tukey, p ⬍ 0.008兲. In 1987–1988 the belt was at its widest, and thus wider than in all the years from 1990–2001 共Tukey, p ⬍ 0.004兲. The increase that is suggested for 2000–2001 in Figure 2 is not revealed in the analysis. Nevertheless, there was no difference between the starting year 1984 and 2000–2001, giving an apparent oscillation around the 1984 value. Further sites were added to the sampling programme from 1990. The Fucus belt did not change 共RM ANOVA p ⬎ 0.05兲 during the period 1990–
2001 at sheltered sites, including the 11 analysed above 共n ⫽ 16兲. The RM ANOVA did not, however, meet the assumptions necessary for the analysis 共no sphericity兲. This problem arises when the development between times across sites is correlated with the development at a later time point at the same sites, which is a dependency between sites and times. For the same time period and the same sites the cover index varied 共RM ANOVA p ⬍ 0.05, corrected for lack of sphericity, Figure 3兲. The Figure suggests a minimum Fucus spp. occurrence around 1997. A total of 22 sheltered sites were visited each year between 1991 and 2001. During this period the cover index changed significantly 共p ⫽ 0.023, multivariate test according to Wilks, used in lieu of sphericity, Figure 4兲. The Figure suggests a minimum Fucus spp. occurrence around 1997. Fixed exposed sites Fucus spp. diminished over time at the exposed sites with no sign of recovery 共RM ANOVA, p ⬍ 0.001, Figure 5兲. The cover index decreased from 37.1 ⫾ 8.8 to 10.1 ⫾ 7.7 共post hoc according to Tukey, p ⬍ 0.05兲. It may be worth noting that at one of the five sites 共the one with the lowest fetch in the exposed group兲, the change in the cover index had a slightly different pattern, reaching a maximum of 61 in 1993 and a minimum of 41 in 2001. The mean at the other
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Figure 6. Mean values of the Fucus spp. belt per year group at the revisited sheltered sites. N⫽ 6 at each year group, i.e., total N⫽30. Figure 4. Yearly mean values of Fucus spp. Cover index at 22 sheltered sites 1991–2001.
6兲. There was no difference between the year group 1983–1985 and 2000 共Tukey, p ⬎ 0.58兲, indicating that start and end values were also similar in this series. The results seem to be equivalent to the 1984– 2001 time series at the fixed sheltered sites. Revisited exposed sites The Fucus spp. belt at the revisited exposed sites 共Figure 1, n ⫽ 5兲 diminished from 4.4 ⫾ 0.90 m in the year group 1987–1990 to 0.78 ⫾ 0.50 m in 2000 共pairwise t-test p ⫽ 0.012兲. This result seems to be consistent with the result from the fixed exposed sites. Field observation of grazing effects and grazer populations
Figure 5. Yearly mean values of Fucus spp. Cover index at 5 exposed sites 1990–2001.
four sites changed from 34.4 ⫾ 10.8 in 1990 to 2.8 ⫾ 1.14 in 2001. Revisited sheltered sites The Fucus spp. belt at the revisited sheltered sites 共Figure 1兲 共n ⫽ 6, N ⫽ 30兲 reached a maximum width in 1987–1988 of 2.17 ⫾ 0.20 m and a minimum in 1992–1993 of 0.83 ⫾ 0.19 m 共p = 0.030, ANOVA and post hoc according to Tukey p ⫽ 0.022, Figure
The decline of Fucus spp. coincided with field observations of large numbers of the grazing isopod Idotea baltica or with the characteristic grazing marks that can be found under large densities of grazing isopods 共Engkvist et al. 2000, 2004兲. At the majority of sites the most obvious grazing damages were found in the deepest part of the Fucus belt. Both adult and juvenile plants were grazed both in upper and lower parts of the thalli. At the exposed sites dense mats of recruits were occasionally found close to or up to tens of metres away from the few remaining patches of adult Fucus. However, field observations indicate that these recruits were heavily grazed and did not survive the following season.
593 During the period of the most rapid decline 共1990– 1994兲 there was a strong correlation between observed grazing intensity and change in Fucus spp. cover, measured as change in cover index at sheltered sites 共r ⫽ 0.885, p ⬍ 0.05兲. A heavy grazing intensity was accompanied by a large reduction of the cover index. The correlation between grazing intensity and change in cover index was equally strong at wave exposed sites 共r ⫽ 0.902, p ⬍ 0.05兲. During the period of recovery at sheltered sites 共1998–2001兲 we found no correlation between grazing intensity and change in Fucus spp. cover 共r ⫽ 0.157, p ⫽ 0.19兲. Grazing intensity was much weaker in this period and obvious grazing marks were only occasionally observed. At sheltered sites in Kalmar County grazer events started in 1989 共Engkvist et al. 2000兲, continuing to 1994. The mean abundance of I. baltica varied between 0.2 and 84 ind/100 g of Fucus wet weight, the highest abundances occuring at the most affected sites. At exposed sites in Blekinge the populations of Idotea peaked from 1992 onwards, with mean abundances between 4.8 and 73.4 ind/100 g of Fucus 共Engkvist et al. 2004兲. Oceanographic and meteorological data There were no regional trends in Secci depth, salinity and nutrients over the period 1990 to 2001 共Tobiasson et al. 2002兲. The 1990s were characterized by relatively mild winters. There was a major exception to this pattern during the winter of 1995/1996, when extensive, thick sea ice occurred from December to April. Between 1990 and 1996, in the period when the most negative changes in Fucus cover occured, the area was struck by 24 storms 共more than 20 ms–1兲. The three worst storms originated from the southwest and occurred on 26 January and 27 February 1990 with average wind speed of 27 ms–1 and on 22 January 1993 with a wind speed of 26 ms–1. Maximum low water was 64 cm below MWL for two hours on 12 May 1993. NAO index The NAO index for 1983–1995 correlated negatively with yearly mean values of the extension of the Fucus belt for 1984–1996 共r⫽ ⫺ 0.757, p ⬍ 0.01兲, i.e., with a time lag of one year 共Figure 2兲. Correlation with a time lag of two years was also significant 共–0.715, p ⬍ 0.01兲. All other combinations were in-
significant. The time lag of one year was used here since it had the closest correlation with the Fucus spp. belt, and since the climate in the winter most likely has the strongest impact on the ephemeral plant and animal communities in the subsequent summer and autumn. Grazing effects do not become manifest until late autumn, when the grazers born in the summer have become adult and capable of grazing on Fucus spp. Monitoring is carried out in early autumn and cannot detect changes in the Fucus spp. belt until the year after the impact. During the period of low index values 共1985–1989兲, the Fucus belt increased or remained wide, while during high index periods 共1990–1996兲, the belt decreased or remained narrow.
Discussion In the late 1970s and early 1980s a large-scale decrease in F. vesiculosus distribution was documented in the coastal waters of south and southwest Finland 共Kangas et al. 1982; Rönnberg 1984兲, with a slight recovery a few years later 共Rönnberg et al. 1985; Kangas and Niemi 1985兲. The decline was most apparent at sheltered sites in the outer archipelago and, in contrast to the findings of the current study, there were only small changes at wave-exposed sites. In the same period the geographical distribution of F. vesiculosus also declined in several other coastal areas in the Baltic Sea 共Schramm 1996兲. The most obvious changes were observed in eutrophicated or polluted sheltered and semi-exposed localities, such as the innermost part of the Gulf of Gdansk 共Plinski and Florczyk 1984兲 and in the recipient area of several pulp mills on the Swedish east coast 共Lindvall 1984; Kautsky et al. 1992兲. From 1993 to 2000, F. vesiculosus disappeared from about 60% of the shores in the Archipelago Sea in southwest Finland 共Helminen et al. 2000兲. However, only minor changes in Fucus distribution were observed during contemporary studies conducted in the Askö archipelago in Sweden 共Kautsky 2001兲. The current study clearly demonstrates substantially weakened Fucus spp. populations at sites subjected to wave exposure. Within a few years, dense belts of F. vesiculosus and F. serratus virtually disappeared from tens of square kilometres of rocky shores. At three of the five exposed sites, where the two species formed mixed stands from the beginning, there was a consistent pattern in that F. vesiculosus declined before F. serratus 共Engkvist et al. 2004兲. At
594 sheltered sites depth penetration and abundance increased during the 1980s. In 共approximately兲 1990, the development reversed, however in 1997 many sites were almost devoid of Fucus spp. Similarly to exposed sites, we attribute the major declines of Fucus at sheltered sites to grazing by the isopod I. baltica. Consequently, the direct cause of the decline of Fucus spp. along the rocky shores of southeast Sweden from 1990 and onwards 共Engkvist et al. 2000, 2004兲 was mass occurrences of I. baltica. The structure of F. vesiculosus populations in the Baltic has previously been attributed mainly to physical factors 共e.g., Kautsky and van der Maarel 1990兲, pulp-mill effluents 共Lindvall 1984兲 and eutrophication 共Schramm 1996兲. However, no single oceanographic or meteorological factor can explain the reduction of Fucus in our study area. Ice scraping can be excluded since the decline started during the mild winters in the early 1990s, years before the cold winter of 1995/ 1996. The periods of the most rapid decline in canopy cover did not coincide with the most severe storms and both exposed and sheltered sites were similarly affected. Nor is desiccation a plausible explanation, as the deepest parts of the belt disappeared initially at many sites 共Engkvist et al. 2000兲. A decreased water transparency could also be excluded, as the Secci depth in the coastal waters of southeast Sweden was generally constant over the period 1990 to 2001 共Tobiasson et al. 2002兲. F. vesiculosus is more tolerant to abiotic stress than F. serratus 共Malm and Kautsky 2003兲. The observation by Engkvist et al. 共2004兲 that the F. vesiculosus canopy declined years before F. serratus at exposed sites could therefore be a further indication against abiotic factors as structurers of the populations in this area. After the mild winters in 1990 to 1995, reflected by a high NAO index 共Figure 2兲, the Fucus cover decreased or remained narrow, probably due to synergistic effects of increased water temperatures, weak ice production and high sea levels. The weak ice production resulted in minimal ice scraping; mild winters are also often associated with a high sea level and therefore minimal desiccation and freezing effects in the upper algal zone and the combined effect of these factors may have favoured the growth of filamentous algae 共Kiirikki and Lehvo 1997兲. We also believe that these winter conditions resulted in increased winter survival and earlier recruitment of Idotea the following year. In favourable years these early recruits can produce a brood in the same autumn 共Jazdzewski 1970; Salemaa 1979; Kangas et al. 1982兲.
Figure 7. Stock biomasses of juvenile cod 共subdivisions 25 to 32兲 and adult herring 共subdivisions 25 to 29 and 32 excluding Gulf of Riga兲 in the Eastern Baltic.
Bottom-up and top-down forces may simultaneously influence marine macroalgal communities 共Hauxwell et al. 1998兲 and probably also Baltic communities 共Worm et al. 2001兲. In our study area, not only mild winters but also eutrophication may have contributed by promoting excess growth of filamentous algae, forming the nutritional base for large grazer populations 共Jansson 1967兲 and by improving the quality of Fucus as food for grazers 共Hemmi and Jormalainen 2002兲. The abundance of adult I. baltica peaks in September–October and this maximum is usually followed by a rapid decline 共Kangas et al. 1982兲. The most intense grazing of Fucus occurs in this period. The experiments reported by Engkvist et al. 共2000兲 demonstrated density-dependent grazing effects of Idotea on F. vesiculosus. At isopod densities of more than 80 animals per 100 g of F. vesiculosus, biomass decreased by 50% in two weeks. These results were also corroborated by field data. Fish predation has been identified as an important source of mortality that could contribute to the seasonality of epifauna in vegetated marine habitats 共e.g., Russ 1980兲. In accordance with Russ’s 共1980兲 findings, and many others, we suggest that fish predation could regulate the autumn abundance of adult Idotea and perhaps be the force that can maintain the abundances of Idotea below the threshold at which grazing effects are manifested. Reduced predation pressure from fish due to depleted stocks of predators such as cod 共Gadus morhua L.兲 and herring 共Clupea harengus L.兲 共Figure 7, Anon. 2002兲 might allow the isopods to continue grazing throughout autumn and
595 early winter, with devastated Fucus populations as a consequence. At most of the sheltered sites the cover of mainly F. vesiculosus increased from 1998, and the distribution in 2001 was again quite similar to that in the mid-1980s. This was unexpected because re-establishment is considered to be a slow process, mainly due to the weak dispersal capacity of fucoids where gametes settle approximately 0.5–2 m from the fertile plant 共Serrão et al. 1996兲. In the current study a subsequent recruitment was observed even at sites where Fucus had virtually disappeared and where the nearest fertile plants were more than 20–50 metres away. These recruits grew into new, dense stands in a few years. This successful recruitment may indicate that F. vesiculosus has a stronger dispersal capacity than has previously been believed, even though we cannot completely exclude other factors, such as seeding from drifting fertile plants 共Engkvist et al. 2000兲 or germination from a propagule bank 共Worm et al. 2001兲. Whatever the reason, the current re-establishment is astonishing as it occurred at sites in some of the most eutrophicated parts of the archipelago. Indirect effects of eutrophication, such as increased herbivory 共Malm et al. 1999兲, increased amounts of filamentous algae 共Worm et al. 2001兲 or deposited matter 共Berger et al. 2003兲, should inhibit recruitment and therefore delay a recovery. The recovery also deviates from our hypothesis that fish predation should be the important regulator of grazer abundance. In spite of continuously low predatory fish stocks we found much weaker grazing intensity in the period of recovery and so far we cannot explain this decreased grazing pressure with certainty. The recovery of Fucus at sheltered sites, however, coincides with local efforts to reduce discharges of nitrogen and phosphorus to coastal waters and could therefore be a result of local measures against eutrophication. In the 1990s, nitrogen discharges from several sewage treatment plants in Kalmar and Blekinge County were reduced, as was the atmospheric deposition of nitrogen 共e.g., Tobiasson et al. 2002兲. There is a general opinion that perennial macroalgae such as fucoids have been disfavoured during the continuing eutrophication process in the Baltic Sea, and, as mentioned above, there is no doubt that the distribution of F. vesiculosus has decreased significantly, even though the underlying causality is seldom unambiguous. Unfortunately, there is a lack of historical data on Fucus distribution in the Baltic, with very few studies comparing distribution patterns
of Fucus communities before large-scale eutrophication started in the 1950s 共Cederwall and Elmgren 1990兲. Earlier surveys carried out in southeast Sweden 共Svedelius 1901; Sjöstedt 1920; Levring 1940兲 are all qualitative, primarily based on a trawling technique. One quantitative ecological study is the work carried out in the 1940s in the Öregrund archipelago of the northern Baltic Sea 共Wærn 1952兲. The diving profiles in Wærn’s study have been revisited twice 共Kautsky et al. 1986; Eriksson et al. 1998兲. Compared to the study in 1943/44, both recent surveys reported a significantly decreased vertical distribution of the Fucus belt and the deepest plants. As the sites were situated in areas not directly affected by local land run-off, it was suggested that the changes indicated large-scale eutrophication of the northern Baltic Sea. However, compared to the study in 1986, the study conducted by Eriksson and colleagues did not indicate any further changes in Fucus distribution, except at depths between 1–2 m where the percentage coverage of F. vesiculosus decreased from 20% to less than 1%. The lack of other detectable changes was interpreted as an indication that no further eutrophication-related distributional changes had taken place during the previous decade. Although in itself this statement may not be false, it does not mean that no changes had occurred during the ten years. As seen in the current study, there can be large fluctuations in a period of ten years or more, meaning that single observations at such intervals could easily overlook rapid dynamics and thus lead to incorrect conclusions. Sampling of Fucus population variables at a limited number of fixed transects, such as the cover index or depth of belt, may incorrectly be taken as representative of Fucus populations in surrounding areas. Published data on the conformity between scattered samples of Fucus variables and populations are scarce. However, in the current study revisited sites exposed to equivalent wave exposures as fixed transects showed a development of Fucus variables comparable to the fixed sites. Therefore, one can assume that the fixed sites are good indicators of the state of the Fucus spp. populations in the coastal waters of southeast Sweden. This analysis was conducted to detect regional trends in distribution and consequently is not a comment on distributional changes at single sites. The great loss of Fucus spp. at the exposed sites could lead to major ecosystem changes. A common phenomenon in eutrophicated shallow coastal areas in
596 the Baltic is a mass occurrence of different kinds of opportunistic filamentous algae, which in most places cover or replace perennial macroalgae such as F. vesiculosus 共Kangas et al. 1982; Plinski and Florczyk 1984; Vogt and Schramm 1991; Eriksson et al. 1998兲. When F. vesiculosus is lost on exposed shores the biomass of fast-growing filamentous species may increase as much as fivefold 共Kiirikki 1996兲. Pihl et al. 共1994兲 believe that the increased dominance of filamentous algae might cause significant changes in fish community composition. They proposed that large fishes would be replaced by smaller fish species, more adapted to the new structure. An increasing amount of filamentous algae changes the physical environment by modifying habitat complexity, which is shown to negatively affect the predation efficiency of larger fishes 共Isaksson et al. 1994兲. The removal of fucoids, and thereby the shelter provided by these marine ecosystem engineers, could have a dramatic influence on the habitat choice of several fish species such as young cod and perch 共Perca fluviatilis L.兲, species that usually avoid habitats lacking large structures as they depend on this complexity for cover 共e.g., Keats et al. 1987兲. In conclusion, the current study has demonstrated large-scale temporal and spatial changes in Fucus populations along the Swedish part of the Baltic proper. The study also showed that fixed monitoring sites could be used to detect such synchronized fluctuation patterns as they were representative of surrounding coastal areas. These changes can only be detected by using continuous, long-term field observations. In the light of the current and other studies of Fucus populations in the Baltic Sea 共e.g., Kangas et al. 1982; Schramm 1996; Eriksson et al. 1998兲, it is obvious that many factors interact in structuring the populations, making it difficult to distinguish between bottom-up and top-down effects 共cf. Menge 2000兲. We found that grazing by Idotea baltica was the proximate cause of the decrease in Fucus abundance. The heavy grazing intensity of I. baltica may, however, ultimately been favoured by factors such as eutrophication, mild winters and reduced predation pressure. The recovery of Fucus in sheltered areas may be a first sign of advantageous effects of local measures against eutrophication.
Acknowledgements This research was funded by the University of Kalmar and the Swedish Environmental Protection Agency. We would like to thank Stefan Tobiasson and Björn Fagerholm for help with field sampling and also the late Anders Johansson and Ingemar Andersson of the County Administrative Boards of Kalmar and Blekinge.
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