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Aquatic Ecology 37: 137–150, 2003. © 2003 Kluwer Academic Publishers. Printed in the Netherlands.
Zooplankton, phytoplankton and the microbial food web in two turbid and two clearwater shallow lakes in Belgium Koenraad Muylaert 1,*, Steven Declerck 2, Vanessa Geenens 1, Jeroen Van Wichelen 1, Hanne Degans 2, Jochen Vandekerkhove 2, Katleen Van der Gucht 1,4, Nele Vloemans 1,4, Wouter Rommens 3, Danny Rejas 2,6, Roberto Urrutia 1,7, Koen Sabbe 1, Moniek Gillis 4, Kris Decleer 5, Luc De Meester 2 and Wim Vyverman 1 1 Dept. Biology, University Gent, K.L. Ledeganckstraat 35, Gent, 9000, Belgium; 2Lab. Aquatic Ecology, KULeuven, Debériotstraat 32, Leuven, 3000, Belgium; 3Lab. Botany, KULeuven, Kasteelpark Arenberg 31, Heverlee, 3001, Belgium; 4Dept. Microbiology, University Gent, K.L. Ledeganckstraat 35, Gent, 9000, Belgium; 5Institute for Nature Conservation, Kliniekstraat 25, Brussel, 1070, Belgium; 6Current address: Laboratory of Limnology, Universidad Mayor de San Simón, Cochabamba, Bolivia; 7Centro EULA, Universidad de Concepcion, Concepcion, Chile; *Author for correspondence (e-mail:
[email protected]; phone: +32 9 264 53 66; fax: +32 9 264 53 34)
Received 21 November 2001; accepted in revised form 19 November 2002
Key words: Alternative stable states, Biomanipulation, Eutrophic shallow lakes, Macrophytes, Phytoplankton, Zooplankton Abstract Components of the pelagic food web in four eutrophic shallow lakes in two wetland reserves in Belgium (‘Blankaart’ and ‘De Maten’) were monitored during the course of 1998–1999. In each wetland reserve, a clearwater and a turbid lake were sampled. The two lakes in each wetland reserve had similar nutrient loadings and occurred in close proximity of each other. In accordance with the alternative stable states theory, food web structure differed strongly between the clearwater and turbid lakes. Phytoplankton biomass was higher in the turbid than the clearwater lakes. Whereas chlorophytes dominated the phytoplankton in the turbid lakes, cryptophytes were the most important phytoplankton group in the clearwater lakes. The biomass of microheterotrophs (bacteria, heterotrophic nanoflagellates and ciliates) was higher in the turbid than the clearwater lakes. Biomass and community composition of micro- and macrozooplankton was not clearly related to water clarity. The ratio of macrozooplankton to phytoplankton biomass – an indicator of zooplankton grazing pressure on phytoplankton – was higher in the clearwater when compared to the turbid lakes. The factors potentially regulating water clarity, phytoplankton, microheterotrophs and macrozooplankton are discussed. Implications for the management of these lakes are discussed. Abbreviations: CPUE – catch per unit effort, HNF – heterotrophic nanoflagellates, SPM – suspended particulate matter Introduction Because of intensive exchange of nutrients between their water columns and sediments, shallow lakes are sensitive to eutrophication (e.g., Ekholm et al. (1997)). When nutrient loading is low, shallow lakes
are relatively clear with low phytoplankton biomass and dense stands of submerged macrophytes. Under the influence of eutrophication, these clearwater lakes become more turbid. Phytoplankton blooms develop and submerged macrophytes disappear due to a lack of light. Due to the loss of submerged macrophyte
138 stands, this transition is usually associated with a loss of structural diversity and, as a result, a decrease in biodiversity at the higher trophic levels takes place (e.g., Hanson and Butler (1994) and Scheffer (1998)). At the same time, the transition results in a loss of important socio-economical functions like the recreational use of the lake for swimming or its use as a source of drinking water. While oligotrophic lakes are generally clear and hypertrophic lakes frequently turbid, shallow lakes at intermediate nutrient concentrations may exhibit either clearwater or turbid states (Scheffer et al. 1993). Many, predominantly biological, feedback mechanisms stabilise these alternative states. As a consequence, during eutrophication, the clearwater state is often maintained until nutrients reach high levels. Conversely, turbid shallow lakes often do not respond to even major nutrient reductions and remain turbid. The stability of each state depends on the nutrient levels: at increasing nutrient levels, there is a decrease in the stability of the clearwater state and an increase in the stability of the turbid state. Several studies have demonstrated that it is possible to reconvert turbid shallow lakes to clearwater lakes (e.g., Meijer et al. (1999)) when reductions in nutrient loading are accompanied by the application of management strategies that alter the structure of the food web itself (biomanipulation, Shapiro and Wright (1984)). Like many low-lying countries in Western Europe, Belgium has numerous wetlands. Due to the advent of industrial activity as well as an intensification of agriculture, many of these wetlands have received high nutrient inputs during the previous century and as a result many are severely eutrophied. Shallow lakes are the dominant lake type in Flanders and, where many of these lakes were of the clearwater type and possessed rich submerged vegetations in the past, most are now eutrophic and turbid (e.g., Denys (1997)). These shallow lakes have a potentially high ecological value if submerged vegetations are restored. This can be attained by applying appropriate management strategies. However, to develop effective management strategies, a good knowledge of the functioning of the ecosystem is required. In contrast to many other Western European countries, relatively few studies have been carried out in Belgium with respect to the functioning of shallow lakes. To achieve a better knowledge on shallow lakes in Belgium and to provide a framework for developing management strategies aimed at restoring the clearwater state in shallow lakes, a research study was
carried out during 1998–2000 in the framework of the Flemish Impulse Program for Nature Development (VLINA). In four lakes from two important wetland reserves in Belgium, all components of the pelagic food web were monitored during two years. In contrast to many previous studies of shallow lakes, this study also included microheterotrophs like bacteria, heterotrophic nanoflagellates and ciliates, important components of the microbial food web. Here we present an initial account and general overview of the results and aim at evaluating the potential factors regulating water clarity in these lakes. Based on these conclusions, some recommendations for the development of management strategies aimed at restoring the clearwater state in the lakes were formulated.
Materials and methods Study sites Four shallow lakes from two nature reserves were selected for this study (Figure 1). Both reserves are important wetlands in Belgium and are protected by national and international agreements. Two lakes, lake Visvijver and lake Blankaart, are situated in the ‘Blankaart’ wetland reserve (Diksmuide, Province of West-Vlaanderen), in the South-West of Belgium, close to the coast. It is positioned close the river Ijzer and is frequently flooded by this river in winter, resulting in the two lakes being connected. The lakes were created by peat digging and are on average about 1 m deep. Lake Blankaart is the larger of the two (surface area 32 ha) and receives surface water loaded with nutrients and sediments via two rivulets. Nutrient inputs through these rivulets are very high due to intensive agriculture and stock breeding in the catchment. Lake Visvijver is a small lake (0.6 ha) receiving no surface water inputs. Both lakes were dredged in 1995 to remove sediment and associated nutrients. Simultaneously, the fish community was manipulated in lake Visvijver, but not in lake Blankaart. These management activities resulted in the transition to a clearwater state in lake Visvijver but had no effect on lake Blankaart (Peeters et al. 1996). The two other lakes, lake Maten 12 and lake Maten 13, are situated in the ‘De Maten’ wetland reserve in the North-Eastern part of Belgium (Genk, Province of Limburg). The ‘De Maten’ reserve contains a total of 32 small lakes with a surface area ranging from 0.5–9 ha (lake Maten 12: 3.2 ha, lake Maten 13: 3.3 ha).
139
Figure 1. The location of the ‘Blankaart’ (B) and ‘De Maten’ (M) wetland reserves in Belgium (left) and detailed maps at identical scales of the four lakes studied (right). Scale bar is 500 m.
The ‘De Maten’ lakes were created by peat digging around the year 1400 and have been used for fish culture until 1991. All lakes in this pond system are directly or indirectly connected by a system of overflows and are fed by surface water via two main rivulets (Cottenie et al. 2001). Between the two ‘De Maten’ lakes studied, water flows from lake Maten 13 to lake Maten 12. Sampling and analyses All lakes were sampled monthly during winter and biweekly during summer from January 1998 to December 1999, resulting in 34 sampling occasions in each lake. Subsurface samples were collected during daytime at a fixed sampling station in each lake. Temperature, pH and Secchi depth were measured at the time of sample collection. Samples for bacteria, heterotrophic nanoflagellates (HNF), ciliates and phytoplankton were fixed in the field using Lugol’s solution, formalin and thiosulphate (Sherr et al. 1989). Macrozooplankton was sampled using a SchindlerPatalas trap. The trap was deployed at different depths to include the entire water column. Macrozooplankton samples were fixed with sucrose-saturated formalin (Haney and Hall 1973). Rotifers were collected quantitatively by filtering 0.5–3 l water through a 30 m pore size nylon mesh. The mesh containing the rotifers was stored in 30 m filtered lake water fixed with 5% w/v formalin. A whole water sample was transferred to the laboratory (cooled and darkened) to be subsampled for dissolved nutrients and suspended particulate matter (SPM).
Bacteria were stained with DAPI, filtered onto 0.2 m pore size membrane filters and counted using epifluorescence microscopy (Porter and Feig 1980). Counts were converted to biomass assuming a cellular carbon content of 2 10 −11 mg C cell −1 (Lee and Fuhrman 1987). HNF were also stained with DAPI, filtered onto 0.8 m pore size membrane filters and enumerated using epifluorescence microscopy (Sherr et al. 1993). Three size classes were discerned during enumeration (< 4 m, 4–10 m and > 10 m). Ciliates were enumerated using inverted microscopy after staining with Bengal rose to facilitate the distinction between living organisms and detritus. Ciliates were identified up to genus level where possible or were assigned to a given size class (< 20 m, 20–40 m or > 40 m). For conversion of HNF and ciliate abundances to biomass, in each lake and for all genera or size classes, 25–100 individuals from different samples were measured. For each lake, an annual average biovolume was calculated for each genus or size class. Biovolume was converted to biomass using a conversion factor of 15 10 −11 mg C m −3 for HNF (Fenchel 1982) and 22 10 −11 mg C m −3 for ciliates (Putt and Stoecker 1989). Phytoplankton was identified up to genus level and enumerated using inverted microscopy. For each lake, of each genus, 25–50 cells from different samples were measured. For each lake, an average biovolume was calculated for each genus and biovolume was converted to biomass using the formulations given by Menden-Deuer and Lessard (2000). Macrozooplankton was enumerated using a dissection microscope. Cladocerans were identified up to species level while copepods were
140 identified to the order level. For each taxon in each sample, 30 individuals were measured to convert abundances to biomass using published length-weight regressions (Bottrell et al. 1976). For quantification of rotifers, the organisms were washed off the nylon mesh and enumerated using a dissection microscope. All individuals were identified to species level where possible or to genus level. Abundances were converted to biomass using published data on biomass content of the taxa encountered (Ruttner-Kolisko 1974; Dumont et al. 1975; Pontin 1978). Samples for dissolved nutrients were filtered over a GF/F filter and stored frozen until analysed using a Skalar autoanalyser according to the methods described in Greenberg et al. (1992). No data on total phosphorus or nitrogen are available. Suspended particulate matter (SPM) concentrations were determined gravimetrically after filtration of a known volume of water onto a preweighed GF/F filter. Twice each year (May and August), the percentage cover by submerged macrophytes was determined along several transects in the lakes and extrapolated to the entire lake surface. Fish biomass and community composition were estimated once in September 2000 using multi-mesh size gill nets (Jeppesen et al. 1999). We used 42 m long, 0.75–1.5 m wide nets subdivided in equal sections of 3 m length with mesh sizes ranging from 6.25 to 70 mm. Nets were placed in the littoral as well as the open water around sunset and were sampled about 16 h later. Four nets were placed in lake Blankaart and the ‘De Maten’ lakes while only two nets were used in the smaller lake Visvijver. Ten individuals of each species were weighed and measured and catches were converted to ‘catch per unit effort’ (CPUE), which is the total wet biomass per net. Statistical analyses We compared biomass and biomass ratios of different components of the food web between the two lakes in each wetland reserve using t-tests. As in each wetland reserve, the two lakes studied are interconnected at least during part of the year and have comparable nutrient levels, our null hypothesis was that biomass or biomass ratios of components of the food web are similar in the two lakes of each wetland reserve. As within-lake seasonal variation often exceeded between-lake differences, paired t-tests were used in which data from the same sampling dates were compared. All data were log-transformed before analysis
to reduce skewness and approximate normal distribution.
Results Concentrations of dissolved inorganic nitrogen (the sum of nitrate, nitrite and ammonium) were higher in the lakes of the ‘Blankaart’ wetland reserve than the ‘De Maten’ reserve (Table 1, Figure 2). Of the two lakes of the ‘Blankaart’ reserve, dissolved inorganic nitrogen concentrations were highest in lake Blankaart, where concentrations up to 20 mg N l −1 were measured. Dissolved inorganic nitrogen concentrations were nearly an order of magnitude lower in the lakes of the ‘De Maten’ wetland reserve and were lowest in lake Maten 13. In the ‘Blankaart’ lakes, and to a lesser extent in the ‘De Maten’ lakes, nitrogen concentrations decreased over summer (Figure 3). Dissolved phosphorus concentrations were also maximal in the ‘Blankaart’ lakes, with lake Visvijver having the highest phosphorus levels (Table 1, Figure 2). Dissolved phosphorus concentrations were close to an order of magnitude lower in the lakes of the ‘De Maten’ reserve. Of the two ‘De Maten’ lakes, Maten 13 had the lowest phosphorus levels, with concentrations approaching detection limit on several occasions during the study. In contrast to nitrogen, phosphorus did not display a clear seasonal trend in the lakes (Figure 3). pH was generally more than one unit higher in the ‘Blankaart’ lakes than the ‘De Maten’ lakes (Table 1). SPM concentrations and Secchi depth were measured as indicators of water clarity (Table 1, Figures 2 and 4). In both wetland reserves, the two lakes studied differed strongly with respect to water clarity. SPM concentrations were highest and Secchi depth lowest in lake Blankaart and lake Maten 12. Lake Visvijver and lake Maten 13 had lowest SPM concentrations and the lake bottom was, apart from a single sampling occasion, always visible. In each wetland reserve, the difference in SPM concentrations between the two lakes was highly significant (t-test: ‘Blankaart’ as well as ‘De Maten’ p < 0.0001). Given these strong differences in water clarity among the lakes studied, we will refer to lake Blankaart and lake Maten 12 as the turbid lakes and lake Visvijver and lake Maten 13 as the clearwater lakes. Submerged macrophyte vegetations were only observed in the clearwater lakes (Table 1, Figure 2). During this survey, percentage cover by submerged
141 Table 1. Averages for some important variables measured in the four lakes studied (1998 and 1999). Dissolved nitrogen is the sum of nitrate, nitrite and ammonia. For phytoplankton, ciliates, micro- and macrozooplankton and fish, the average percentage contribution of some important taxonomical/functional groups to annual average total biomass is given next to the average biomass of these groups.
Variable Abiotic variables pH SPM (mg l −1) Secchi depth (m) Nutrients dissolved phosphorus (g l −1) dissolved nitrogen (g l −1) Phytoplankton (g C l −1) Chlorophyta Cryptophyta Euglenophyta Incertae Sedis Cyanobacteria Bacillariophyta Dinophyta Chrysophyta Xanthophyta Microheterotrophs (g C l −1) Bacteria HNF Ciliates oligotrich ciliates other ciliates Microzooplankton (g C l −1) Asplanchna other rotifers copepod nauplii Macrozooplankton (g C l −1) cyclopoid copepods calanoid copepods Bosmina Daphnia Ceriodaphnia Submerged macrophytes (% cover) Fish (CPUE in g net −1) Planktivorous fish Benthivorous fish Piscivorous fish
‘Blankaart’ Blankaart
Visvijver
‘De Maten’ Maten 12
Maten 13
8.5 43 0.37
8.5 6 lake bottom
7.2 19 0.36
7.1 9 lake bottom
280 5880 1190 757 (46%) 137 (17%) 71 (6%) 38 (5%) 86 (8%) 95 (17%) 5 (< 1%) 2 (< 1%) 1 (< 1%)
460 750 160 21 (10%) 94 (63%) 1 (< 1%) 31 (16%) 9 (5%) 3 (4%) 2 (< 1%) 1 (< 1%) 0 (< 1%)
45 207 430 125 (30%) 41 (14%) 148 (24%) 41 (16%) 32 (5%) 16 (5%) 10 (3%) 6 (2%) 11 (2%)
18 99 180 69 (23%) 55 (35%) 20 (9%) 11 (15%) 2 (1%) 4 (3%) 16 (10%) 4 (4%) 1 (< 1%)
262 22 177 124 (68%) 53 (32%) 916 75 (4%) 349 (41%) 492 (55%) 1478 850 (62%) 0 (0%) 454 (19%) 172 (13%) 0 (0%) 0 1714 566 (33%) 1129 (66%) 20 (1%)
134 2.4 47 24 (73%) 23 (27%) 517 7 (2%) 289 (49%) 221 (49%) 524 316 (53%) 18 (9%) 35 (9%) 28 (17%) 122 (9%) 55 0 0 0 0
170 9 233 57 (40%) 176 (60%) 338 26 (8%) 187 (56%) 125 (36%) 344 131 (29%) 26 (11%) 150 (45%) 21 (9%) 8 (2%) 0 420 259 (62%) 0 (0%) 161 (38%)
130 3.4 53 29 (57%) 24 (43%) 255 8 (4%) 122 (42%) 125 (53%) 240 37 (22%) 68 (30%) 32 (21%) 90 (11%) 6 (3%) 43 202 192 (95%) 10 (5%) 0 (0%)
macrophytes was around 50% in both clearwater lakes. In lake Visvijver, the dominant macrophyte was Chara globularis, while floating beds of filamentous algae covered a large part of the lake surface towards the end of summer. In lake Maten 13, Drepanocladus fluitans, Polygonum amphibium and Nitella trans-
lucens dominated the submerged vegetation, with Utricularia vulgaris also being commonly observed. Phytoplankton biomass was significantly higher in the turbid than in the clearwater lakes (t-test: ‘Blankaart’ as well as ‘De Maten’ p < 0.0001), with lake Blankaart having highest biomass (Table 1, Figure 4). The phytoplankton community in the turbid lakes was
142
Figure 2. a–d: Median values (horizontal line) and 25 (top of box) and 75 percentiles (bottom of box) for some important chemical and physical variables. e: Fish biomass (as catch per unit effort or CPUE) and community composition as measured in September 2000. f: The average (May and August in 1998 and 1999) cover by submerged macrophytes.
dominated by chlorophytes (mostly coenobial taxa like Scenedesmus) while, in the clearwater lakes, cryptophytes were the dominant group of algae. Highest biomass of cyanobacteria was observed in lake Blankaart, where a Planktothrix bloom developed towards the end of the summer in 1998. In the turbid lakes, phytoplankton biomass increased gradually over summer (Figure 3). In the clearwater lakes, blooms developed rather irregularly between spring and autumn. During summer, phytoplankton biomass tended to be low in the clearwater lakes. Bacterial biomass was, like phytoplankton biomass, significantly higher in the turbid than in the clearwater lakes (t-test: ‘Blankaart’ p < 0.0002 and ‘De Maten’ p < 0.01), with maximal average biomass occurring in lake Blankaart (Table 1, Figure 4, Figure
5). Differences in heterotrophic protistan biomass between turbid and clearwater lakes were more pronounced than differences in bacterial biomass (Table 1, Figure 4). Biomass of ciliates was much higher than that of HNF in all lakes. Combined biomass of heterotrophic protists (HNF and ciliates), as well as biomass of HNF and ciliates separately, was on average higher in the turbid than in the clearwater lakes. This difference was significant for ciliates as well as HNF (t-test: ‘Blankaart’ and ‘De Maten’ p < 0.0001, both for ciliates and HNF). Oligotrich ciliates were relatively more important in the lakes of the ‘Blankaart’ reserve than in those from the ‘De Maten’ reserve. In contrast to phytoplankton, bacteria and heterotrophic protists, biomass of micro- as well as mac-
143
Figure 3. Temporal dynamics of suspended particulate matter (SPM) concentration, dissolved phosphorus concentration, dissolved inorganic nitrogen concentration (nitrate + nitrite + ammonia) and phytoplankton biomass in the lakes from the ‘Blankaart’ (left) and the ‘De Maten’ (right) wetland reserves in 1998 and 1999. Note different scales for the ‘The Blankaart’ lakes and the ‘De Maten’ lakes.
rozooplankton seemed not to be clearly related to water clarity (Table 1, Figure 4). Biomass of macrozooplankton was significantly higher in the turbid lake in the ‘Blankaart’ reserve (t-test: p < 0.01) but did not differ significantly between the two lakes from the ‘De Maten’ reserve (t-test: p = 0.059). Bio-
mass of microzooplankton also was significantly higher in the turbid lake in the ‘Blankaart’ reserve (ttest: p < 0.05) but did not differ significantly between the two lakes from the ‘De Maten’ reserve (t-test: p = 0.46). In the clearwater lakes, total zooplankton biomass was generally highest during spring. In the tur-
144
Figure 4. a–c, e, g, i: Median values (horizontal line) and 25 (top of box) and 75 percentiles (bottom of box) for some important biological variables. d, f, h, j: Percentage contribution of some important groups to total biomass of phytoplankton (d), ciliates (f), microzooplankton (h) and macrozooplankton (j).
145 bid lakes, total zooplankton biomass peaked during summer (Figure 5). Of all lakes, the biomass of macrozooplankton was on average highest in lake Blankaart. Biomass of microzooplankton was comparable to macrozooplankton biomass in all lakes except lake Blankaart, where macrozooplankton biomass was about twice as high as microzooplankton biomass. Copepod nauplii and rotifers contributed equally to total microzooplankton biomass in all lakes. The carnivorous rotifer Asplanchna contributed on average only 2–8% of total microzooplankton biomass in the lakes studied. The percentage contribution of daphnids (Daphnia and Ceriodaphnia) to total macrozooplankton biomass was slightly higher in the clearwater lakes than in the turbid lakes, but this difference was only significant in the ‘Blankaart’ reserve (t-test: ‘Blankaart’ p < 0.05 and ‘De Maten’ p = 0.88). To estimate the grazing impact of macrozooplankton on phytoplankton, the macrozooplankton to phytoplankton biomass ratio was calculated as an indicator of macrozooplankton grazing pressure on phytoplankton (Table 2). In each wetland reserve, the biomass ratio of total macrozooplankton (t-test: ‘Blankaart’ p < 0.001 and ‘De Maten’ p < 0.005) or daphnids (t-test: ‘Blankaart’ p < 0.0005 and ‘De Maten’ p < 0.05) over phytoplankton was significantly higher in the turbid when compared to the clearwater lake. While the macrozooplankton to phytoplankton ratio differed only about 5-fold between the clearwater and turbid lakes, the daphnids to phytoplankton ratio was 16 and 24 times higher in the clearwater than the turbid lake in the ‘Blankaart’ and ‘De Maten’ reserve, respectively. Fish biomass (as CPUE) tended to be higher in the turbid than the clearwater lakes, with lake Blankaart having highest fish biomass (Table 1, Figure 2). No fish were caught in lake Visvijver, where a period of anoxia preceded a massive fish kill in 1998. Planktivorous and benthivorous species like white bream (Blicca bjoerkna), roach (Rutilus rutilus) and bream (Abramis brama) dominated in lake Blankaart while the exotic and piscivorous brown bullhead (Ameiurus nebulosus) and (to a lesser extent) roach and rudd (Scardinius erythrophtalmus) were the most important species observed in lake Maten 12. In lake Maten 13 rudd and tench (Tinca tinca) dominated the fish community with the exotic pumpkinseed (Lepomis gibbosus) and topmouth gudgeon (Pseudorasbora parva) being common too.
Discussion Dissolved inorganic nutrient concentrations in the lakes from the ‘Blankaart’ wetland reserve exceeded those from the lakes in the ‘De Maten’ reserve by nearly an order of magnitude. In both wetland reserves, dissolved nutrient concentrations in the two lakes studied were comparable. This is not surprising, given the fact that the lakes in both reserves were at least during some part of the year interconnected. Flooding of the small dike separating the clearwater lake Visvijver from the turbid lake Blankaart occurs regularly in winter, resulting in an exchange of water between both lakes. Lake Maten 13 discharges into lake Maten 12 by an overflow system and both lakes are fed by the same rivulet. Despite the fact that the two lakes in both wetland reserves had similar nutrient concentrations, a clearwater and a turbid lake were present in each wetland reserve. This observation supports the alternative stable states theory (Scheffer et al. 1993) in that water clarity may differ strongly in shallow lakes having similar water chemistry. The fact that a clear and a turbid lake occurred in both wetland reserves despite large differences in nutrient concentrations between the two reserves illustrates that alternative stable states may occur over a wide range of nutrient loadings. In general, it is assumed that at total phosphorus concentrations exceeding 100–150 g l −1, a stable clearwater state cannot be achieved (Jeppesen et al. 1990; Hosper and Jagtman 1990). Although unfortunately no data on total phosphorus were available for this study, with an average dissolved phosphorus concentration of 460 g l −1, this limit must have been exceeded in lake Visvijver. Water clarity in the four lakes studied was related to food web structure rather than nutrient concentration. Biomass of phytoplankton and microheterotrophs, the zooplankton to phytoplankton ratio, occurrence of submerged macrophytes and the fish community differed markedly between the clearwater and turbid lakes. Phytoplankton biomass was nearly twice as high in the turbid lakes when compared to the clearwater lakes. Being a component of the seston, phytoplankton contributes to total SPM and may therefore be a direct cause of high suspended matter concentrations in turbid lakes. When phytoplankton biomass was converted to dry weight (assuming a carbon to dry weight ratio of 0.5, Reynolds (1984)), however, phytoplankton accounted only for on average 5.9% (maximum 22%, in lake Blankaart)
146
Figure 5. Temporal dynamics of bacterial biomass, biomass of heterotrophic protists (HNF + ciliates), biomass of microzooplankton (rotifers + copepod nauplii) and macrozooplankton biomass in the lakes from the ‘Blankaart’ (left) and the ‘De Maten’ (right) wetland reserves in 1998 and 1999. Note different scales for the ‘Blankaart’ lakes and the ‘De Maten’ lakes.
or 5.1% (maximum 14%, in lake Maten 13) of total suspended matter in the turbid lakes. Therefore, although high SPM concentrations are associated with high phytoplankton biomass, phytoplankton cannot be the only factor responsible for the high suspended
sediment levels in the turbid lakes. Submerged macrophytes covered about half the lake surface in the clearwater lakes but were absent in the turbid lakes. When forming a dense vegetation, submerged macrophytes may reduce water column turbulence and in-
147 Table 2. Average ratios of zooplankton biomass over phytoplankton biomass in the four lakes studied.
Ratio
‘Blankaart’ Blankaart
Visvijver
‘De Maten’ Maten 12
Maten 13
Macrozooplankton/phytoplankton Daphnids/phytoplankton
1.6 0.15
6.7 2.4
0.9 0.09
4.4 2.1
crease water clarity through a reduction of sediment resuspension and an increase in SPM sedimentation rates (e.g., James and Barko (1990) and Jones (1990)). In turbid lakes, no submerged macrophytes are present to reduce turbulence in the water column and, as a result, even low wind velocities may create turbulence that is strong enough to cause resuspension of bottom sediments (Kristensen et al. 1992). Differences in the fish community may also explain the difference in water clarity between the clear and turbid lakes. Benthivorous fish species attained high biomass in the turbid lakes. Moreover, the piscivorous brown bullhead, which attained high biomass in the turbid lake Maten 12, also has a benthic lifestyle. Through their foraging activities, fish communities from the turbid lakes may contribute substantially to sediment resuspension (cf. Meijer et al. (1990) and Havens (1991)). Next to the presence of macrophyte stands, which prevent resuspension and promote sedimentation of suspended matter, low biomass or complete absence of benthivorous fish may be an additional cause of the low SPM concentrations in the clearwater lakes. Total nutrient concentrations were similar in the clear and turbid lakes in both wetland reserves studied, while phytoplankton biomass differed about twofold. Therefore, some factor other than nutrients must have been regulating phytoplankton biomass in the clearwater lakes. In shallow lakes, macrozooplankton are generally considered to be the dominant grazers on phytoplankton. Macrozooplankton is in turn preyed upon by planktivorous fish. Planktivorous fish influence macrozooplankton biomass as well as community composition, as certain macrozooplankton groups like Bosmina and copepods are able to avoid fish predation (Verity and Smetacek 1996; Slusarczyk 1997) while the relatively large daphnids are highly sensitive to fish predation (e.g., Brooks and Dodson (1965) and Schriver et al. (1995)). In the four lakes studied, however, no clear relation could be found between CPUE of planktivorous fish and total macrozooplankton biomass or the contribution of daphnids to macrozooplankton biomass. Neither total bio-
mass nor community composition, however, is a good indicator of zooplankton grazing pressure on phytoplankton. We calculated the ratio of zooplankton biomass over phytoplankton biomass as an indicator of zooplankton grazing pressure on phytoplankton. This ratio was significantly higher in the clearwater than in the turbid lakes. Most zooplankton groups, like rotifers, copepods or Bosmina, graze selectively on specific phytoplankton size classes and are therefore generally considered incapable of efficiently controlling phytoplankton biomass (Hansson and Carpenter 1993; Havens 1993; Cyr and Curtis 1999). Only daphnids like Daphnia and Ceriodaphnia have high clearance rates on a wide size-range of particles and capable of regulating phytoplankton biomass (Hall et al. 1976). The biomass ratio of daphnids to phytoplankton was also significantly higher in the clearwater when compared to the turbid lakes. Moreover, the difference between clearwater and turbid lakes with respect to this ratio was much higher for daphnids than for total macrozooplankton. This suggests than grazing by daphnids may play an important role in regulating phytoplankton biomass in the clearwater lakes. Like for phytoplankton biomass, microheterotrophic components of the microbial food web (bacteria, HNF and ciliates) attained significantly higher biomass in the turbid than in the clearwater lakes. To a large extent, these microheterotrophs depend directly or indirectly on phytoplankton as a food source. Exudates produced by phytoplankton are an important substrate for aquatic bacteria in shallow lakes (Kamjunke et al. 1997). HNF and ciliates feed on bacteria (e.g., Sanders et al. (1989)) and small phytoplankton (e.g., Weisse et al. (1990)). The low biomass of bacteria, HNF and ciliates in the clearwater lakes may therefore reflect lower food availability. Top-down regulation, however, may also explain the lower biomass of the components of the microbial food web in the clearwater lakes. Daphnids are nonselective filter feeders whose filter apparatus is capable of retaining particles down to about 1 m in size (Brendelberger 1991). Phytoplankton, bacteria, HNF
148 and ciliates comprise potential food items for daphnids (Jürgens 1994). Therefore, if grazing by daphnids affects phytoplankton populations, it is also likely to affect bacterial, HNF and ciliate populations. In both wetland reserves, a clearwater and a turbid lake occurred next to each other in an interconnected system and with virtually identical nutrient loading. This indicates that, in both wetland reserves, restoration of the turbid lakes to the clearwater state should be possible. The differences between the turbid and the clearwater lakes are well in agreement with the theory of the alternative stable states (Scheffer et al. 1993) and suggest that the restoration of the clearwater state in the turbid lakes can be accomplished by changes in food web structure rather than by water quality improvement alone (Perrow et al. 1997; Hansson et al. 1998; Meijer et al. 1999). Such changes in food web structure can be achieved by the application of biomanipulation (Shapiro and Wright 1984; Gulati et al. 1990). In practice, this method most often involves a strong reduction of the planktivorous and benthivorous fish biomass via large scale fish removals, often combined with stocking of piscivorous fish (e.g., pike). The removal of planktivorous fish is meant to allow for populations of large-bodied filter feeding Cladocera to develop and gain control over phytoplankton by grazing. The reduction of the biomass of benthic fish results in a decrease of sediment resuspension and associated nutrient release from the lake sediments. A decrease in phytoplankton biomass and the amount of suspended matter results in an increase in water clarity, resulting in opportunities for the establishment of macrophytes. These macrophytes are now considered to stabilize the clearwater state through several positive feedback mechanisms (e.g., Hansson et al. (1998)). The difference in water clarity between the studied ponds is to a large extent due to differences in food web structure. Biomanipulation, for instance, has been successfully applied in ponds of the Blankaart system. Although lake Visvijver originally was a turbid pond lacking submerged vegetation, the lake switched to the clearwater state after it was dredged in 1995 and biomanipulated during three subsequent years (Declerck et al. 2000; Declerck 2001). A similar result was obtained in lake Kasteelvijver, another lake in the Blankaart reserve. Although a large proportion of superficial sediments of lake Maten 12 is made up by peat, which is a potentially important source of organic matter and humic acids that can increase water turbidity, the difference in water clarity
between the ponds in de Maten reserve probably also results from differences in water level regime. In contrast to lake Maten 12, water levels in lake Maten 13 are highly variable and often quite low. The pond water level has repeatedly fallen dry during the last decades. By impacting the fish community, such events may have acted in a similar way as biomanipulation. In addition, low water levels may have stimulated the germination of submerged vegetations. Both the successful biomanipulation of lakes in the Blankaart system and the impact of water levels on food web structure in lake Maten 13 suggest that fish removal or temporal drainage may be a very promising techniques for the restoration of the clearwater state in the turbid lakes of the two wetland reserves. This is certainly true for lake Maten 12, as this lake is very similar to lake Maten 13 with respect to morphometry, size and water quality. In Lake Blankaart, several small scale in situ enclosure experiments have resulted in strong increases in water clarity when fish were excluded (Declerck et al. 1997; Declerck 2001). In these experiments, a rapid growth of populations of large cladocerans was observed, resulting in strong reductions in phytoplankton biomass due to grazing. Such results indicate that planktivorous fish suppress large zooplankton in lake Blankaart and that biomanipulation may contribute to the restoration of the Blankaart system. However, a large proportion of the increase of water transparency may also have resulted from an enclosure artefact causing sedimentation of algae and suspended matter. Moreover, a large scale food web manipulation experiment that was carried out in a 1 ha part of the lake that was separated by a net and in which a 60% removal of the benthic and planktivorous fish biomass was combined with the introduction of pike fingerlings, did not result in any increase of water transparency (Declerck et al. 2001). Extrapolation of the results of the successful food web manipulations done in lake Visvijver should therefore be made with care, as lake Blankaart differs from lake Visvijver in many aspects. Lake Blankaart is considerably larger than lake Visvijver and is more exposed to wind. Wind can be a very important cause of sediment resuspension in shallow lakes (Douglas and Rippey 2000) and wind-induced waves can hamper the development of submerged vegetation (Hosper 1997). Furthermore, in contrast to lake Visvijver, lake Blankaart is a very open system that is connected with several rivulets. These rivulets allow an easy immigration of fish. Moreover, via the rivulets, there is a continuous inflow of water rich in nutrients, sus-
149 pended matter and pesticides. If a large scale biomanipulation of lake Blankaart would be undertaken, these factors would hamper the development of submerged vegetation and would continuously have a destabilising effect on the clearwater state, decreasing the probability of a successful restoration on a long term. We therefore advocate that large-scale food web manipulations in Lake Blankaart should be preceded by a drastic reduction in nutrient loading of the upstream rivulets in the Blankaart catchment area. This can be achieved by an alteration of the agricultural practice, encompassing measures to reduce erosion and a restriction of the use of fertilizers and pesticides. Moreover, the input of suspended matter can be reduced by the re-meandering and dredging of the rivulets and the implementation of buffer strips (Declerck et al. 2001).
Acknowledgements The research presented in this paper was carried out in the framework of the Flemish VLINA project 97/03 ‘Restoration of nature values in shallow lakes: investigating the structure and functioning of the microbial loop and the trophic cascade in a series of model systems’. KM and KS are postdoctoral fellows of the Belgian Fund for Scientific Research. HD acknowledges a fellowship from the Flemish Institute for the Promotion of Scientific-Technological Research in Industry (IWT). During data analysis and writing, SD, KvdG, NV, WR, MG and LDM were financially supported by EU project BIOMAN (EVK2CT-1999-00046).
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