Ecosystem change in the large and shallow Lake Säkylän Pyhäjärvi, Finland, during the past ~400 years: implications for management
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Lake Säkylän Pyhäjärvi has been an important fishing site and drinking water source for the local population for centuries. The lake has undergone significant changes: (1) the water level was lowered in the 1600s and in the 1850s; (2) planktivorous coregonid fish were successfully introduced in the early 1900s; (3) nutrient input from intensified agriculture has increased since the 1950s and (4) the effects of the current variable climate on the lake and its catchment have become more evident since the 1990s. We determined the phases of oligotrophication, eutrophication and recovery and elucidated the ecosystem changes by combining palaeolimnological records with detailed neolimnological data. The sedimentary diatom and cladoceran assemblages first showed a relatively eutrophic period followed by oligotrophic periods, linked with the artificial changes in water level and consequent shifts in macrophyte abundance. The oligotrophic period in the early 1900s is thought to represent the target trophic state for the lake. After the 1950s, introduction of vendace resulted in higher planktivory reflected by an increased relative abundance of small-bodied pelagic cladocerans. Signs of eutrophication occurred due to increased nutrient load. During the last 10 years, signs of recovery have been recorded. A complex history such as that of Lake Pyhäjärvi illustrates the difficulties in selecting management targets, and the risk of setting false targets, for lakes based solely on monitoring data—both neolimnological and palaeolimnological approach are needed.
KeywordsEutrophication Palaeolimnology Diatoms Cladoceran subfossils Monitoring Lake recovery Restoration targets Climate change adaptation
Lakes around the globe have undergone increasing eutrophication due to sewage loading, input of diffuse nutrients from agricultural activities, fish stockings and other pressures related to human activities (Millennium Ecosystem Assessment, 2005). It is expected that the pressure on lakes and their ecological state will increase in the future due to both an increasing world population, demanding a higher food production and to climate change (Millennium Ecosystem Assessment, 2005). Such changes will further challenge the protection and use of ecosystem services of the lakes such as water supply, industry, recreation and fisheries, not least because nutrient loading and climate warming can have a synergistic effect (Jeppesen et al., 2010; Moss et al., 2011; Jeppesen et al., 2014). Thus, adaptation and management actions are often needed. Predicting future climate change is, however, difficult due to the high uncertainty regarding large-scale dynamics, mechanisms and timing of the changes (IPCC, 2014; Refsgaard et al., 2014). It is moreover expected that the frequency and magnitude of extreme weather events will increase (Knutti & Sedláček, 2013; IPCC, 2014), which will further challenge the function and quality of lake ecosystems (Jeppesen et al., 2015). Detailed analyses of the long-term dynamics of lake ecosystems as well as key factors and interactions affecting water quality and ecosystem services are, therefore, required to allow implementation of mitigating management measures including climate change adaptation tools.
Although many lakes have been intensively monitored for multiple years, the existing long-term records usually cover only the recent 40–50 years or often less. If longer records are needed, palaeolimnological studies may help to extend the time series (Smol, 1985, 1991; Battarbee et al., 2005; Sayer et al., 2010). Analyses of cladoceran subfossils and of diatoms have been used to reconstruct long-term changes in lake ecosystems and their response to, for instance, eutrophication (summarised in Davidson & Jeppesen, 2013). Cladocerans are functionally adapted to different microhabitats (pelagic, macrophyte-associated, benthic). Changes in the contribution of key cladoceran taxa may therefore yield information about habitat alterations and changes in lake trophic structure (Frey & Pamini, 1986; Jeppesen et al., 2001). In addition, cladoceran subfossils are useful indicators of changes in predation by planktivorous fish and fish stockings (Amsinck et al., 2007), the latter being reflected in changes in the size of Daphnia ephippia (Verschuren & Marnell, 1997; Jeppesen et al., 2002) and the contribution of Daphnia to the total sum of Daphnia and Bosmina ephippia (Leavitt et al., 1994; Jeppesen et al., 2003). Changes in cladoceran community can also be caused by climate change (Smol et al., 2005; Thienpoint et al., 2015) though such effects may be difficult to separate from those of other factors. Also, the climate change effect may be indirect and for example channelled through food webs. Fossil diatom assemblages have, among other purposes, been extensively used for creating quantitative inference models for, for example, total phosphorus (e.g. Anderson, 1993; Bennion et al., 2001; Hall & Smol, 2010), but although qualitatively diatoms can be very good at tracking nutrients, no matter how carefully the quantitative inference is done—it cannot be trusted alone (Davidson & Jeppesen, 2013). Also, for both neolimnological and palaeolimnological approaches, reduction of the detailed species level data to annual or decadal scale or indexes may potentially lead to loss of valuable ecological information on species’ seasonality, habitat structure and food web interactions (Sayer et al., 2010).
Restoration of eutrophicated lakes in Europe aims to maintain or restore good ecological status according to EU’s Water Framework Directive where ‘good ecological status’ is defined for all lake types based on a set of variables to be compared to reference systems (European Union, 2000; Phillips, 2014). The current status is evaluated relative to past ‘pristine status’, i.e. the time when the aquatic system was not influenced by human activities. ‘Pristine status’ may be difficult to define or quantify as aquatic systems in rather densely populated Europe have been influenced by various anthropogenic pressures for thousands of years (Battarbee & Bennion, 2010; Bouleau & Pont, 2015) and are constantly changing in reaction to new pressures such as those related to the ongoing climate change.
Lake Säkylän Pyhäjärvi (61°00′N, 22°18′E, 45 m.a.s.l) in Southern Finland (hereafter “Pyhäjärvi”) is an example of a large shallow lake affected by human activities in its catchment and by lake fishery for centuries. The Pyhäjärvi Restoration Programme was voluntarily established 20 years ago by local municipalities, companies and associations after the lake had shown visible symptoms of eutrophication and monitoring data had indicated that the external nutrient load was too high to maintain an oligotrophic state (Sarvala & Jumppanen, 1988; Malve et al., 1994; Ekholm et al., 1997). In recent years, climate change-related alteration of both nutrient loading (Ventelä et al., 2011) and the fish community (Jeppesen et al., 2012) has affected the lake.
In this paper, we describe the aquatic ecosystem history and water quality of Pyhäjärvi during the past ~400 years using a combination of palaeolimnological and neolimnological records. We further aim to define a target ecological status for use in lake management processes.
Materials and methods
History of the lake
Pyhäjärvi was isolated from the Litorina Sea (one of the former phases of the Baltic Sea) about 5600 years B.P. (Eronen et al., 1982). Archaeological findings (Lehtomäki-Hilander, 2000) suggest that the lake catchment was already then colonised by humans and used as a fishery. There are also indications of pre-industrial production of iron from the lake ore (Uotila, 2011), which is formed when dissolved iron hydroxide in the groundwater moves through the sediment and precipitates around the oxic-anoxic interface. The outflow, River Eurajoki, was modified for the first time for industrial use in the late 1600s when an iron works was established at the outflow in the northern part of the lake (Korvenmaa, 1989). There are indications that this increased river outflow may have resulted in a lowering of the water level, but no quantitative data are available (Uotila, 2011; G. Haggren, personal communication). Increased use of firewood for iron production may have impacted the land use characteristics of the catchment area (Laitakari, 1965). A major lowering of the water level by 2 m occurred during 1852–1854 in connection with reclamation of arable land (Laitakari, 1925; Veira, 1974). The pre-lowering mean depth was ~7.5–8 m. Further modifications of the outflow were officially permitted in 1910, but the first real regulation permit was given in 1936. The upper and lower limits of the permit were, however, often exceeded during extreme flood and drought events. Therefore, new regulations were established by the Water Court in 1975. The present regulation amplitude between the permitted lower and upper water levels is rather narrow, 58 cm, but continuous monitoring of river flows and improved hydrological forecasts have made it possible to mostly fulfil the authorised water level targets (Marttunen et al., 2000). The lake’s current mean depth is 5.5 m.
The lake and its fishery have both been strongly influenced by a long history of stocking events. Three commercially important species, whitefish (Coregonus lavaretus), vendace (Coregonus albula) and signal crayfish (Pacifastacus leniusculus), were introduced. Whitefish was stocked in 1908–1909 and soon after it became the most important species for the commercial fishery (Sarvala et al., 1998a). Vendace was introduced for the first time in 1925, but extensive stocking of the species in 1948–1952 resulted in the establishment of a permanent population (Helminen et al., 1993). The third commercially significant introduced species was signal crayfish, which was imported from Sweden to replace the plague-ravaged noble crayfish (Astacus astacus). In the beginning, the population developed slowly, but since the 2000s the catches have increased (Karjalainen et al., 2015) and crayfish catch has become an economically important income source of fishermen.
Shifts have also occurred in the nutrient loading to the lake. South-western Finland is an important food production area. After the Second World War, the efficiency of the agriculture increased and so did soil fertilisation. In addition, the municipalities developed sewage systems, and flush toilets became increasingly common from the 1950s and onwards. Currently, industrial and municipal waste waters are treated in waste-water treatment plants and diverted from the drainage area of Pyhäjärvi, i.e. they are led to the outlet below the lake.
Pyhäjärvi has been subjected to a large-scale restoration programme since 1995 (Ventelä & Lathrop, 2005; Kirkkala, 2014). A multitude of measures have been implemented in the catchment in order to reduce the external load (Kirkkala et al., 2012; Kirkkala, 2014). Pyhäjärvi has also been intensively biomanipulated for decades in the form of removal of planktivorous fish by commercial fishery. Fish removal has traditionally been undertaken in winter and the annual harvest rate of vendace has approached the total production (Sarvala et al., 1998a). Since 2013, an increasing proportion of fish removal by biomanipulation consisted of the natural planktivore smelt (Osmerus eperlanus), which was fished in late autumn with open-water seine nets. The restoration project has also subsidised the harvest of commercially unwanted fish since 1995. In 2002–2006, the EU provided funds for this fishing, which was especially intensive in 2002–2004 and was partly, together with lower external nutrient load, responsible for the observed water quality improvement (Ventelä et al., 2007). The annual biomanipulation effort varied in 1995–2014 from 50 kg/ha (in 2002) to 8 kg/ha (in 2008) (A-M. Ventelä, unpublished data).
The data used in this study originate from national, areal and project-funded monitoring programmes; yet, the monitoring protocol and methods are comparable. The water chemistry and hydrology of Pyhäjärvi have been monitored since the 1960s as part of a national programme. Since 1971, monitoring has been more intensive and involves use of automatic meters continuously measuring the river flows (Ekholm et al., 1997). The nutrient concentrations in the water of Pyhäjärvi and the main rivers Yläneenjoki and Pyhäjoki have been measured since 1980, first as part of a statutory monitoring programme (Sarvala & Jumppanen, 1988) and since 1992 by the regional authorities approximately at a 2–3 week interval during the open-water season. The water chemistry and hydrology data used in this study are derived from the Finnish Environment Institute’s Oiva data service (www.ymparisto.fi/oiva).
Water samples for nutrient and chlorophyll a analyses were taken from the surface to the bottom at a single sampling point situated in the deepest area of the lake. In our study, we used average open-water season (May–October) values for the 0–5 m layer, representing >80% of the total lake volume. The methods of analyses followed Finnish standard laboratory procedures (Ekholm et al., 1997) and are accepted by Finnish Accreditation Service (FINAS).
The annual total phosphorus (TP) and total nitrogen (TN) loads from the major rivers were calculated based on TP and TN concentrations in water samples and continuously measured stream flow data (Ekholm et al., 1997). We used the sums of the TP and TN loads of River Yläneenjoki and River Pyhäjoki as total external nutrient load.
Sediment coring, subsampling, dating and core correlation
Two sediment cores were taken through the ice from the deepest part of the lake in March 2004 using a HON-Kajak-type corer (Renberg, 1991) for the most recent sediments (0-30 cm) and a modified Kullenberg corer (PP corer; Putkinen & Saarelainen, 1998) for the deeper Sections (0–687 cm, of which the uppermost 130 cm were used here). The Kajak core was sectioned into 1 cm slices for all analyses, whereas subsamples from the PP core were taken at different intervals for diatom and cladoceran analyses (see below).
Sediment samples from the HON-Kajak corer (0–30 cm) and the topmost sediment (0–15 cm) from the PP cores were analysed for 137Cs activity using an Ortec gamma spectrometer equipped with a four-inch NaI(Tl) crystal, two amplifiers and a 2048-channel pulse height analyser at the Geological Survey of Finland. The HON-Kajak and the PP cores were correlated based on their 137Cs activity profiles using the 1986 Chernobyl nuclear fallout peaks as reference point. The resulting composite core was analysed for 210Pb activity using an ORTEC OCTETE PC alpha spectrometer (Äikäs et al., 1994) at the Institute of Geological Sciences, Polish Academy of Sciences, Warsaw, Poland. Radiometric dates were calculated using the Constant Rate of Supply (CRS) model of dating (Appleby, 2001) and applying the 137Cs date of 1986 as reference point.
Sedimentary diatom, cladoceran and Gloeotrichia analyses
Cladoceran and Gloeotrichia remains were analysed from 41 subsamples with varying frequency (each 1 cm slice down to 30 cm and approximately every tenth below that) (Figs. 2b, c, 4). Each sample was homogenised, and subsamples of approximately 5 g wet sediment were boiled in 30 ml 10% KOH and kept cold (4°C) until all taxonomic analyses had been performed. All subfossils larger than 80 µm were identified using an Olympus SZX12 stereomicroscope with 100× magnification and a Leitz Labovert FS inverted light microscope with 320× magnification. The subfossils were divided into two size fractions: >140 µm and 80–140 µm. Daphnia spp. ephippia and Gloeotrichia colonies were counted from the entire sample, whereas the remaining Cladocera taxa were subsampled due to high densities. As several different subfossil parts (e.g. carapace, headshield, postabdomen, postabdominal claw) of each taxon were counted, only the most abundant and most representative subfossil part was used for data analysis. Whenever possible, all taxa were identified to species level. In addition, the dorsal length of all enumerated Daphnia spp. ephippia was measured. Cladocerans were identified using standard keys such as Frey (1959), Margaritora (1985), Hann (1990) and Flössner (2000) and Gloeotrichia using Tikkanen & Willén (1992). The cladoceran assemblages were expressed as percentage abundance. Classification of cladoceran habitat preferences was based on Hann (1990) and Røen (1995). The Palaeolimnological Predation Index (PPI) is based on a relationship between the ephippia ratio of Daphnia/(Daphnia + Bosmina) and fish abundance (catch per unit effort in gillnets) (Jeppesen et al., 2003). The macrophyte indices are Chydoridae/(Chydoridae + Bosmina) carapaces excluding C. sphaericus (Jeppesen et al., unpubl.; Bjerring et al., 2008) and the sum of the relative abundance of the macrophyte-associated species.
In order to explore possible shifts in diatom assemblages and to identify zones with similar species compositions, constrained cluster analyses with the paired group method were conducted using Euclidean, Bray–Curtis and chord distance similarity measures. Changes in species richness were analysed through rarefaction analysis (Birks & Line, 1992), which produces realistic estimates of richness without any bias associated with the variability of individual sample count size. Cluster analyses and rarefaction analyses were performed using the software PAST (Hammer et al., 2001).
Past lake water TP concentrations were inferred using a quantitative diatom-based TP inference model comprising 68 southern Finnish lakes with a TP range from 3 to 89 µg TP/l (Kauppila et al., 2002).
For phytoplankton analyses, composite samples have been collected from 0 to 5 m depth six to 24 times during May–October each year since 1980. Two sites were sampled in the first decade, and samples from ten locations have been bulked since the early 1990s. Additional phytoplankton data collected as part of a national monitoring programme were available from the years 1963–1965, 1968, 1971 and 1977. Phytoplankton was counted using the Utermöhl technique (Utermöhl, 1958) and biomass was estimated separately for each cell using standard methodology (Tikkanen & Willén, 1992).
Zooplankton has been sampled weekly or at 2–3 week intervals during the open-water season (May–October) since 1987, and less complete data were available for 1977, 1982 and 1986. Samples were taken during day from 0 to 5 m depth with a 0.5- or 1-m-high tube sampler (volume 3.4 or 6.8 l) at ten locations selected using a stratified random design. The samples were concentrated with a 50-µm mesh net and combined in the lab to one composite sample per date. In the lab, subsamples were enumerated until 50–200 individuals of the dominant crustacean species had been counted and measured. Zooplankton counts and measurements were converted to zooplankton biomass using carbon-length regressions (Sarvala et al., 1998b).
The weight of 0+ vendace in autumn (S0+) was used as a neolimnological index for planktivory (NPI) (transformed as: NPI = 100–S0+) (Ventelä et al., 2011).
Sediment dating and core correlation
The distributions of the most abundant diatom taxa in the composite core are displayed together with the rarefaction-estimated number of species at 300 counted valves and the diatom-inferred past lake water TP concentrations (Fig. 2a). Four distinctive local diatom zones (Z1–Z4; Fig. 2a) were identified based on the clustering results (not shown).
Z2 (30–80 cm; ~1700–1930): In this zone, the amount of periphytic diatoms such as T. flocculosa, Pseudostaurosira brevistriata, Navicula exilis, Nitzschia subacicularis, N. gracilis and N. dissipata var. media increased (Fig. 2a). The diatom-inferred TP concentrations decreased to 15–20 µg TP/l, and the number of diatom taxa increased to 70-80. Gloeotrichia abundance decreased (Fig. 2b). Large-bodied Daphnia and Bosmina increased, especially Bosmina longispina. The abundance of the predatory Leptodora kindtii decreased (Fig. 4). The palaeolimnological predation pressure index indicated an overall low predation pressure and the ephippial size of Daphnia spp. was large (Fig. 5a). Both the number of macrophyte-associated taxa (Fig. 5c) and the macrophyte abundance index (Fig. 5g) showed an abrupt increase at the onset of zone 2 followed by a declining trend from mid-zone 2. The major artificial lowering of the lake water level was undertaken in the later part of this period.
2004–2014: The mean TP concentration in the lake during 2004–2014 was 19.6 ± 2.0 µg/l, decreasing from the previous period. Total phytoplankton biomass decreased since 2003; in 2011-2013 the cyanobacteria species P. agardhii constituted relatively high proportions of the total biomass (Fig. 9). In 2004 and 2014, cyanobacteria annual biomass means were exceptionally low. Bosmina spp. and Daphnia spp numbers were high (Fig. 7a, b), but as in Z3–Z4 their size was closely linked to the level of planktivory (Fig. 8b; R 2 = 0.68, P < 0.0001, Fig. 8d; R 2 = 0.53, P < 0.0001).
External load and water quality
The annual mean concentration of TP in the lake increased in the 1980s and the early 1990s. The highest mean concentration (22 µg/l) was observed in 1994. The variation increased during the 2000s (17–23 µg/l) (Fig. 10f). The annual mean concentration of TN remained rather stable (400–500 µg/l). The means of TN for the periods 1980–1994 (Z3), 1995–2003 (Z4) and 2004-2014 were 457.4 ± 43.6, 474.2 ± 38.4 and 452.4 ± 32.0 µg/l (Fig. 10g), respectively. The average annual chlorophyll a concentration varied, but in the 1990s and 2000s the level was higher than in the 1980s. The means of chlorophyll a for 1980–1994, 1995–2003 and 2004–2014 were 5.6 ± 1.6, 7.5 ± 1.6 and 7.35 ± 1.59 µg/l, respectively (Fig. 10h). Secchi depth decreased during the monitoring period 1980–2014. Average Secchi depth for the periods 1980–1994, 1995–2003 and 2004–2014 were 3.1 ± 0.5, 2.8 ± 0.2 and 2.5 ± 0.3 m, respectively (Fig. 10i). In 2014, the water quality parameters indicated more oligotrophic conditions than in the previous years. TP, TN and chlorophyll a concentrations were low, and in July 2014 exceptionally high Secchi depths (up to 4.8 m) were recorded, and also the annual mean Secchi depth was relatively high.
The results from Z1 (AD ~1600–~1700), the period before the major water level drawdown, suggest that the lake was meso-eutrophic. The diatom assemblages were dominated by planktonic and meroplanktonic species, and based on the low macrophyte abundance index and the low number of truly macrophyte-associated cladoceran taxa macrophyte abundance was presumably low. Judging from the size of Daphnia and the paleolimnological predation index, the predation pressure on cladocerans was moderate to low. This concurs with the fact that there was no specialised planktivore among the native fish species (Sarvala et al., 1998a).
The human-induced water level decrease apparently had a profound effect on the aquatic ecosystem of Pyhäjärvi during Z2 (~1700–~1930). A major shift in the diatom and cladoceran communities, indicating an increase in macrophytes and improved water quality, occurred at 80-cm sediment depth. This corresponds to a change in sediment properties as the magnetic susceptibility increased from ~1.5 to ~4 SI units between 90 and 80 cm (data not shown; magnetic susceptibility was measured at 1 cm intervals directly from unopened core tubes using a Bartington MS2 magnetic susceptibility meter and the MS2C Core Logging Sensor), indicating that the sediment became permanently more mineral-rich. It is quite likely that this was a result of lake water lowering, changing the system into one that constantly transports more mineral matter to the deep than before. However, it is unclear whether the oligotrophication was caused by the possible lowering in the late 1600s or by the lowering in the 1850s. The 210Pb dating of the core covered only part of this period (>1863 AD) and partly contradicts the earlier study by Räsänen et al. (1992) who ascribed the oligotrophication to the 1852–1854 water level lowering undertaken to develop by then still traditional and low-intensive agriculture. However, dating uncertainties occurred also in Räsänen et al. (1992) for sediment depths below 53 cm. The possible early water level modification in connection with the establishment of the Eurajoki iron mill in the late 1600s is based not on data but on qualitative archaeological and historical evidence and it is thus difficult to estimate its role in oligotrophication.
Notwithstanding uncertainty over the timing of oligotrophication (biotic community changes and lowering of DI-TP), it likely occurred in this period as a response to the water level drawdown. This is supported by an increase in macrophyte abundance as evidenced by the enhanced proportion of periphytic diatoms and of both macrophyte- and sediment-associated cladocerans, with potential positive feedback on water clarity via reduced phytoplankton development, as seen in other studies (e.g. Coops et al., 2003; Bucak et al., 2012; Jeppesen et al., 2015). Accordingly, a decline in the proportion of planktonic diatom forms suggests a reduction in phytoplankton production and improved water clarity. The increase in littoral taxa may also reflect a change in sedimentation patterns, associated with the lower water level, which would increase the share of littoral diatoms and cladocerans at the deep water coring site. The cladoceran record suggests continued low planktivory, a typical response to oligotrophication (Persson et al., 1988; Jeppesen et al., 2000). All these results suggest that during Z2 Pyhäjärvi was in a meso-oligotrophic clear-water state with relatively high abundance of macrophytes and relatively low fish planktivory, this being consistent with early limnological observations in the mid-1910s (e.g. Secchi depth 4.0–4.3 m; Järnefelt, 1927). The connection of Pyhäjärvi with an aquifer (Rautio & Korkka-Niemi, 2011, 2015) might also have contributed to the oligotrophication. Planktivorous whitefish were successfully introduced in 1908, but our study did not find any clear effects of this introduction on the cladoceran community in the Z2 sediment record. Salo et al. (1989) and Räsänen et al. (1992), however, found a reduction in B. coregoni mean body size in sediment samples representing year ~1921, which they ascribed to whitefish stocking.
During Z3 (~AD 1930–1994), some major changes occurred in the catchment and the lake: the introduction of vendace in the 1950s, resulting in high vendace abundances since the 1960s (Sarvala et al., 1998a), intensification of agriculture and increased use of chemical fertilisers after the Second World War (Valkama et al., 2009). Finally, in the early 1990s, the effect of climate change on the lake and its catchment became more evident, including changes in winter precipitation and nutrient load (Ventelä et al., 2011) and in lake temperature (Jeppesen et al., 2012). These changes apparently had a pronounced impact on the lake ecosystem. DI-TP increased to 20–25 µg/l. Similarly, the monitoring TP record increased from 11 µg/l (in 1980) to 19.3 µg/l (in 1994). The eutrophication can also be seen as a rapid increase of F. crotonensis, a species indicative of eutrophic waters, and as a clear increase in cyanobacteria and diatoms. Algal blooms were observed, arousing public concern (Kirkkala, 2014). Anabaena akinetes were also recorded in the entire 30-cm core in a separate study by Räsänen et al. (2006), but the akinete record does not allow estimates of past Anabaena abundances. Similar problems are not likely with Gloeotrichia, which were counted as whole colonies.
A marked decline of large-bodied Daphnia and B. longispina is evidenced by both palaeolimnological and neolimnological samples. The abundance of the macrophyte-associated species of cladocerans decreased, also indicating more eutrophic conditions with higher phytoplankton growth, further facilitated by low grazing of phytoplankton by the small-bodied zooplankton (Jeppesen et al., 1997). The fish stockings, especially of vendace, had a strong effect on the zooplankton through increased planktivory, clearly reflected as changes in cladoceran body sizes (Fig. 5a) (see also Sarvala et al., 1998b).
The large-scale restoration of Pyhäjärvi started in 1995 (the beginning of Z4) (Ventelä & Lathrop, 2005; Kirkkala, 2014). The most important activities have been the nutrient load reduction measures in the catchment and the intensive long-term biomanipulation conducted by local professional fishermen. The biomanipulation effort was intensified after 2002 (Ventelä et al., 2007). In this zone, the diatoms associated with eutrophic conditions dominated but decreased in the topmost layers, showing detectable recovery of the lake after restoration. The neolimnological monitoring data show that total phytoplankton biomass more than doubled from 1995 to 2002 but then declined steeply, which is consistent with the TP changes and the initiation of restoration measures. The phytoplankton changes were mainly due to variations in biomass and species dominance of cyanobacteria (e.g. Anabaena spp,) and diatoms (e.g. Tabellaria spp., Aulacoseira spp.). The highest Gloeotrichia biomass recorded by monitoring data was in 2001 with highly visible blooms in shore areas, which attracted great public attention and concern, and the users of the lake were dissatisfied with the water quality (Stedman et al., 2007). Both Gloeotrichia and Anabaena spp had lower biomasses later in the 2000s.
Both palaeo- and neolimnological results reveal a high predation pressure: small-sized ephippia dominated in the palaeo data, and the mean carapace length of Daphnia and Bosmina was inversely related to the fish predation pressure in the neolimnological data. Our results suggest that it took some time before the intensive lake management efforts started to take effect, but the deterioration trend finally ceased in the 2000s.
For the most recent period, 2004–2014, only neolimnological data are available. High latitude lakes have been shown to be sensitive to climatic forcing (Itkonen et al., 1999; Sorvari et al., 2002), and the impact of climate-related environmental variables clearly increased also in Pyhäjärvi: the length of the ice cover period decreased (Ventelä et al., 2011) and the maximum lake temperature increased at the same time as the vendace stock declined (Jeppesen et al., 2012). Also the external phosphorus load varied strongly; the highest (wet 2008) and lowest (dry 2009) annual loads during the monitoring period were observed in consecutive years. From 2004 to 2014, the mean TP concentration was similar to that in 1995–2004, which likely reflects the loading reduction in the catchment and the successful biomanipulation.
After 2002, phytoplankton biomass and composition varied strongly but are indicative of lake recovery. The proportion and biomasses of species and taxa indicative of eutrophic conditions were lower than in the 1990s. Gloeotrichia biomass was low and the cyanobacteria community was dominated by P. agardhii, a species that does not form visible blooms disturbing the recreational use of the lake, and which is mainly concentrated in the deeper parts of the water column. In 2014, the phytoplankton composition and biomass reflected a, perhaps temporary, improvement in ecological status. The increase in zooplankton abundance was connected with the low abundance of planktivores resulting from sufficiently large biomanipulation catches.
Based on our study results, we conclude that the external nutrient load is closely linked with precipitation, indicating that the current catchment measures are not sufficiently able to reduce the nutrient load peaks derived from an increased water flow. More detailed studies and new practical tools, in particular better buffering of the system to hydrological extremes, are needed to compensate for the climate change-induced higher nutrient loading, as highlighted also in other studies (Trolle et al., 2011; Jeppesen et al., 2014). Controlling the nutrient load during very rainy periods and extreme weather events may turn out to be a strongly demanding and almost impossible task.
Biomanipulation may be an effective in-lake management tool also in a variable climate. The current trend of increasing temperatures and a decreasing planktivore biomass (Jeppesen et al., 2012) indicates a change towards a food web structure with high numbers of large cladocerans and of other herbivorous zooplankton, with a resulting positive effect on water quality, a development that may be furthered by fish manipulation. However, the phytoplankton community is also strongly affected by climate change (Adrian et al., 2009; Pätynen et al., 2014) and may, as indicated by the development in recent years in Pyhäjärvi, become dominated by cyanobacterial species like Planktothrix, which are less edible to zooplankton (Rücker et al., 1997). This may diminish the trophic cascade effect of the biomanipulation (Søndergaard et al., 2007; Ger et al. 2014).
Climate-related changes will challenge the protection and use of ecosystem services such as water supply, industry, recreation and fisheries, not least because nutrient loading and climate warming effects act in synergy (Jeppesen et al., 2010; Moss et al., 2011; Jeppesen et al., 2014). Adaptation and management actions are needed, but to be sufficiently efficient they should be based on the knowledge of the key functions of the ecosystem.
Decrease in eutrophic taxa
Increase in abundance and size
Decrease in biomass
Varying, but relatively low (biomanipulation, low stock of planktivores)
Increase in climate-related effects
Both eutrophic and benthic taxa
Varying, affected by planktivory
Abundant, highest biomass in 2001
Variation in planktivore stocks
Initiation of Pyhäjärvi restoration programme
Elevated nutrient levels
Increase in eutrophic taxa, high abundance of periphytic taxa
Decrease in Daphnia and Bosmina, dominance of pelagic taxa
Decrease in macrophyte abundance
Stocking and consequent high abundance of whitefish and vendace, enhanced planktivory
Increased nutrient input
Increase in TP concentration
Oligotrophic and periphytic taxa
Sediment- and macrophyte-associated taxa, significant increase in Daphnia and Bosmina
Increase in macrophytes to relatively high abundance
Introduction of whitefish since 1908, but low level of planktivory
Lowering of water level
Decrease in TP concentration, meso-oligotrophic clear-water state
Planktonic and meroplanktonic diatoms, mesotrophic taxa
Low abundance of Daphnia and Bosmina, high abundance of pelagial Chydorus
Natural meso-eutrophic lake
In spite of the current climate-driven challenges to restoration of Pyhäjärvi, some signs of recovery emerged in both the palaeo- and neolimnological records following the initiation of restoration. Particularly, the clearly improved water quality recorded in 2014 offers hope that the target state of the meso-oligotrophic period of the early 1900s may be reached. However, if the challenges induced by the climate change become too difficult to meet in the future, the restoration target should be the best possible adaptation and mitigation of the negative effects. A similar revision of targets may be necessary in other similar lake restoration projects irrespective of whether the current targets are set by the local communities or, for example, the European Union.
Our case study shows that long-term data records are needed to understand the realistic management targets. Together the palaeo- and neolimnological records provide a more detailed insight into the recent changes as well as a more reliable hindcast of the changes occurring before initiation of monitoring than monitoring data alone, as highlighted also by Battarbee et al. (2005). Also, neolimnological data are needed to complement the information given by palaeolimnological methods (Davidson & Jeppesen, 2013). A complex history such as that of Lake Pyhäjärvi illustrates the difficulties in selecting management targets, and the risk of setting false targets, for lakes based solely on monitoring data.
Erik Jeppesen, Martin Søndergaard, Liselotte Sander Johansson and Susanne Lildal Amsinck were supported by the MARS project (Managing Aquatic ecosystems and water Resources under multiple Stress) funded under the 7th EU Framework Programme, Theme 6 (Environment including Climate Change), Contract No. 603378 (http://www.mars-project.eu), ‘CLEAR’ (a Villum Kann Rasmussen Centre of Excellence project) and CRES. Anne-Mari Ventelä and Teija Kirkkala were supported by the Pyhäjärvi Restoration Program and the Academy of Finland funded projects CARE (203591) and Lakes in Trouble (256240). Funding for field and laboratory studies came from many sources over the years, in particular a series of grants from the Academy of Finland (including Nos. 09/109, 09/074, 1071004, 1071149, 1071255, 35619, 44130, 52271, 104483, 201414, 256240), EU’s Financial Instrument for Fisheries Guidance (Pyhäjärvi Biomanipulation Project 2002–2006), European Regional Development Fund, Ministry of Agriculture and Forestry (biomanipulation funding since 2010) and Pyhäjärvi Protection Fund. Numerous persons have participated in the data acquisition; special thanks are due to Vesa Saarikari, Marjo Tarvainen, Sanna Autio and Kristiina Vuorio. We also thank Kari Uotila for providing archaeological information. Finally, we thank Thomas Davidson, John Smol and an anonymous reviewer for their useful comments and suggestions.
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