Trophic structure, species richness and biodiversity in Danish lakes: changes along a phosphorus...

Freshwater Biology (2000) 45, 201–218

Trophic structure, species richness and biodiversity inDanish lakes: changes along a phosphorus gradient

ERIK JEPPESEN, JENS PEDER JENSEN, MARTIN SØNDERGAARD, TORBEN LAURIDSEN andFRANK LANDKILDEHUSNational Environmental Research Institute, Department of Lake and Estuarine Ecology, PO Box 314, DK-8600 Silkeborg,Denmark

SUMMARY

1. Using data from 71, mainly shallow (an average mean depth of 3 m), Danish lakeswith contrasting total phosphorus concentrations (summer mean 0.02–1.0 mg P L− l),we describe how species richness, biodiversity and trophic structure change along atotal phosphorus (TP) gradient divided into five TP classes (class 1–5: B0.05, 0.05–0.1,0.1–0.2, 0.2–0.4, \0.4 mg P L−1).2. With increasing TP, a significant decline was observed in the species richness ofzooplankton and submerged macrophytes, while for fish, phytoplankton and floating-leaved macrophytes, species richness was unimodally related to TP, all peaking at0.1–0.4 mg P L−1. The Shannon–Wiener and the Hurlbert probability of inter-specificencounter (PIE) diversity indices showed significant unimodal relationships to TP forzooplankton, phytoplankton and fish. Mean depth also contributed positively to therelationship for rotifers, phytoplankton and fish.3. At low nutrient concentrations, piscivorous fish (particularly perch, Perca fluviatilis)were abundant and the biomass ratio of piscivores to plankti-benthivorous cyprinidswas high and the density of cyprinids low. Concurrently, the zooplankton was domi-nated by large-bodied forms and the biomass ratio of zooplankton to phytoplanktonand the calculated grazing pressure on phytoplankton were high. Phytoplanktonbiomass was low and submerged macrophyte abundance high.4. With increasing TP, a major shift occurred in trophic structure. Catches of cyprinidsin multiple mesh size gill nets increased 10-fold from class 1 to class 5 and the weightratio of piscivores to planktivores decreased from 0.6 in class 1 to 0.10–0.15 in classes3–5. In addition, the mean body weight of dominant cyprinids (roach, Rutilus rutilus,and bream, Abramis brama) decreased two–threefold. Simultaneously, small cladoceransgradually became more important, and among copepods, a shift occurred fromcalanoid to cyclopoids. Mean body weight of cladocerans decreased from 5.1 mg inclass 1 to 1.5 mg in class 5, and the biomass ratio of zooplankton to phytoplanktonfrom 0.46 in class 1 to 0.08–0.15 in classes 3–5. Conversely, phytoplankton biomassand chlorophyll a increased 15-fold from class 1 to 5 and submerged macrophytesdisappeared from most lakes.5. The suggestion that fish have a significant structuring role in eutrophic lakes issupported by data from three lakes in which major changes in the abundance ofplanktivorous fish occurred following fish kill or fish manipulation. In these lakes,studied for 8 years, a reduction in planktivores resulted in a major increase in clado-

Correspondence: Erik Jeppesen, National Environmental Research Institute, Department of Lake and Estuarine Ecology, PO Box314, DK-8600 Silkeborg, Denmark.E-mail: [email protected]

© 2000 Blackwell Science Ltd 201

E. Jeppesen et al.202

ceran mean size and in the biomass ratio of zooplankton to phytoplankton, whilechlorophyll a declined substantially. In comparison, no significant changes were ob-served in 33 ‘control’ lakes studied during the same period.

Keywords : biodiversity, fish, phytoplankton, species richness, zooplankton

Introduction

In recent years, food-web structure and interactionsin the pelagic of lakes have been subject to intensedebate. With a few exceptions (e.g. Hrbacek et al.,1961; Brooks & Dodson, 1965; Brooks, 1969), the pre-vailing view a decade or two ago was that food websare primarily regulated via the available resources(‘bottom–up’ control). Since then it has become evi-dent that food webs may be regulated via fish (calledpredatory control or ‘top–down’ control) (Carpenter,Kitchell & Hodgson, 1985; Gulati et al., 1990; Carpen-ter & Kitchell, 1993). However, opinions vary on therelative importance of resource and predatory controlalong a nutrient gradient. On the basis of the studiesof Fretwell (1977) and Oksanen et al. (1981) of terres-trial environments, Persson et al. (1988) claimed thatherbivory on phytoplankton depends on the numberof food web links and that the zooplankton grazingpressure is high in lakes with an even number oflinks (e.g. lakes with only zooplankton and phyto-plankton or lakes with predatory fish, planktivorousfish, zooplankton and phytoplankton) and low inlakes with an odd number of links. Predatory controlwill, therefore, be strongest in food webs with aneven number of links (two, four, etc.) and resourcecontrol highest in food webs with an odd number oflinks. There are several examples supporting the hy-pothesis (Hansson, 1992; Persson et al., 1992; Wurts-baugh, 1992), but also many exceptions among bothlakes and streams (Leibold, 1990; Flecker &Townsend, 1994; McIntosh & Townsend, 1994;Mazumder, 1994; Brett & Goldman, 1996). Among thereasons for the deviations are behavioural changes ofprey, aiming at reducing the predation risk (McIntosh& Townsend, 1994), and changes in the compositionof primary producers towards grazing-tolerant/graz-ing-resistant forms at high grazer density. Moreover,species at each food-web level may differ in sensitiv-ity towards potential predators. In addition, somespecies cover more than one trophic level or showontogenetic shifts in food preference. In nature, foodwebs are therefore rarely simple.

Empirical analyses are alternative approaches forevaluating food-web interactions. Based on statisticalanalyses of experimental data, McQueen, Post & Mills(1986) and McQueen et al. (1989) concluded that re-source control is highest at the bottom of the foodweb and predatory control strongest at the top of theweb. They also suggested that predatory control ishigh in oligotrophic lakes and low in eutrophic lakes,particularly because eutrophic lakes are typicallydominated by cyanobacteria that are difficult forgrazing zooplankton to handle. Elser & Goldman(1990) and Carney & Elser (1990) then developed theintermediate control hypothesis, arguing that thezooplankton grazing pressure on phytoplankton is (a)low in oligotrophic lakes due to low nutrientavailability and because the zooplankton is domi-nated by copepods that are less efficient grazers thanlarge-bodied cladocerans; (b) high in mesotrophiclakes in which the zooplankton is dominated by theefficient grazer Daphnia spp.; and (c) low in eutrophiclakes in which the phytoplankton is dominated bygrazing-resistant species, such as cyanobacteria. Incontrast, Sarnelle (1992) argued that changes in thebiomass of zooplanktivorous fish have their greatestimpact on the phytoplankton (via the zooplankton) ineutrophic lakes. Likewise, Jeppesen et al. (1997, 1999)recorded dominance by small-bodied zooplanktonand a low potential grazing pressure on phytoplank-ton in eutrophic lakes, irrespective of whether thephytoplankton was dominated by cyanobacteria orby edible green algae, thus suggesting major top–down control of zooplankton.

Changes in nutrient loading also result in changesin community structure at each trophic level orwithin different taxonomic groups. To quantify suchchanges, numerous indices have been developed (fora review see Washington, 1984). The most commonlyused are species richness and various indices of di-versity, and in particular the Shannon–Wiener Index.Contrasting results have been obtained for the differ-ent taxonomic units among studies. Some authorshave found a unimodal relationship between species

© 2000 Blackwell Science Ltd, Freshwater Biology, 45, 201–218

A detailed study of Danish lakes along a phosphorus gradient 203

richness and trophic state, peaking at some interme-diate level in the gradient (Stockner & Benson, 1967),whereas others record a monotonical decline (Patalas& Patalas, 1966) or even no significant changes (Eck-mann & Rosch, 1998). Diversity indices, and particu-larly species richness, are sensitive to sample size(MacArthur & Wilson, 1967), invasion barriers, pH,latitude, lake area, heterogeneity, predation and dis-turbance (Hall, Cooper & Werner, 1970; Magnuson,1976; Fryer, 1985; Rørslett, 1991; Keller & Conlon,1994) apart from ecological stress factors such aseutrophication. Variables such as heterogeneity andpH may, in turn, be influenced by eutrophication,reflecting a decline in submerged macrophyte cover-age and enhanced phytoplankton production, respec-tively. The validity of some of the commonly useddiversity indices has been questioned by several au-thors (Hurlbert, 1971; Chutter, 1972; Goodman, 1975).Consequently, Hurlbert (1971) developed an alterna-tive index, the probability of inter-specific encounter(PIE) index, incorporating both species richness andevenness components. The biological relevance of thePIE index is, therefore, more apparent than that of thecommonly used indices (Washington, 1984), but sofar it has not been used for aquatic ecosystems. Inaddition, most studies of species richness and diver-sity have focused on one or two taxonomic groups,whereas only a few have compared the response ofseveral trophic levels/taxonomic groups to changes introphic state (e.g. Reed, 1978).

Using survey data from 71, mainly shallow, Danishlakes, we took an empirical approach to elucidatehow trophic structure, species richness, biodiversityand the extent of top–down control are affected bychanges in trophic state. We also included case stud-ies from three lakes in which the biomass of cyprinidschanged markedly as a consequence of fish kill ormanipulation.

Methods

Samples for the analysis of water chemistry and phy-toplankton and zooplankton communities were takenfortnightly during summer (1 May–1 Oct). Totalphosphorus and chlorophyll a were analysed on adepth-integrated sample from the photic zone at amid-lake station according to Søndergaard, Kris-tensen & Jeppesen (1992) and Jespersen & Christof-fersen (1987), respectively.

Zooplankton densities were determined usingdepth-integrated water samples taken with a Patalassampler and pooled from one to three stations. De-pending on the total phosphorus (TP) level, between4.5 and 9 L of the pooled sample were filteredthrough an 80-mm net and fixed in Lugol’s iodine(1 mL, 100 mL tap water) and 0.5–1 L was settledovernight in Lugol’s solution. Rotifers and small nau-plii were counted from the settled samples, whilst allother zooplankton were counted from net samples.At least 100 individuals of the dominant zooplanktonspecies were counted. Length–weight relationships,according to Dumont, Van De Velde & Dumont(1975) and Bottrell et al. (1976), were used to estimatebiomass. If possible up to 50 individuals weremeasured.

Phytoplankton was counted on Lugol-fixed sedi-mented water samples (pooled sample from a mid-lake station in the photic zone) using an invertedmicroscope. Biovolume was calculated by fitting thedifferent species and genera to geometric forms(Utermohl, 1958; Edler, 1979; Rott, 1981). A factor of0.29 was used to convert biovolume (mm3) to biomass(mg dw) (Reynolds, 1984).

The composition and relative abundance of thepelagic fish stock in the lakes were determined bystandardized fishing (J.P. Muller et al., unpublished)with multiple mesh (6.25, 8, 19, 12.5, 16.5, 22, 25, 30,33, 38, 43, 50, 60, 75 mm) gill nets. The length anddepth of each section of mesh was 3 and 1.5 m,respectively. Fishing was conducted in each lake be-tween 15 August and 15 September, as previous trialfishing indicated that the distribution of the fish pop-ulations was most even during this period (Muller etal., unpublished). Moreover, young-of-the-year fishwere also large enough to be included in the catch bythis time. The nets were set in the late afternoon andretrieved the following morning. Catch per unit effort(CPUE) of planktivorous fish was calculated as meancatch per net.

Species richness of submerged (isoetids, angio-sperms, mosses and charophytes) and floating-leavedmacrophytes was determined in early August, moreor less corresponding to peak biomass. We dividedeach lake into zones (typically 15–22). In each zone,we conducted observations from a boat using a waterglass or by diving at 10 randomly selected locationsat 25–50 cm depth intervals from the shore to themacrophyte boundary. At each point, we measured



coverage divided into the following categories: B1,1–5, 5–25, 25–50, 50–75, 75–95, \95% and theninterpolated these data to a coverage for the wholelake (Jeppesen et al., 1998). For the sake of clarity, thedata were divided amongst five TP categories:B0.05, 0.05–0.1, 0.1–0.2, 0.2–0.4 and \0.4 mg P L−1.The data analysis comprised stepwise regression onloge-transformed data. In some cases a unimodal rela-tionship to TP appeared. We then included [loge(TP)]2

as an independent TP variable. To estimate changesin biodiversity, we used the Shannon–Wiener Index(H%, bits)

H %n= − %s

i=1

ni

nln

ni

n, H %w= − %

s

i=1

bi

bln

bi

b

where ni and bi are the number and biomass of the ithspecies, n and b are the total number and biomass ofall species, and s is the total number of species. Wealso used the Hurlbert ‘encounter’ index (PIE, bits)

PIE=n

n−1�

1− %s

i=1

�ni

n�2�

Results

The lakes studied were, overall, small and shallow(Table 1). Spearman’s correlation on loge-transformeddata revealed a strong significant positive relation-ship between TP and total nitrogen (TN), making itdifficult to discriminate between the effect of the twonutrients on trophic structure. We selected TP as theindependent nutrient variable, as it has been shownmost often to be the best indicator of trophic state intemperate lakes world-wide (OECD, 1982) as well asin Danish lakes (Kristensen, Jensen & Jeppesen, 1990).Total phosphorus was significantly negatively corre-lated with mean depth, but not related to lake surfacearea. As trophic structure and dynamics are substan-tially affected by lake depth (Keller & Conlon, 1994;Jeppesen et al., 1997), we included depth as an inde-

pendent variable in our regression analysis to avoidbias introduced by variations in depth along the TPgradient. We also included lake area, as species rich-ness and biodiversity have been shown to be sensi-tive to the size of the area studied (MacArthur &Wilson, 1967).

Species richness and diversity

Multiple regressions revealed that fish species rich-ness was significantly unimodally related to TP (PB0.0001) and positively related to mean depth(PB0.0001). Lake area contributed significantly (PB0.001) if mean depth was excluded, otherwise it didnot (P\0.14). The mean number of fish species in-creased from six in class 1 to nine in classes 3–4,followed by a decrease to seven in class 5. The Shan-non–Wiener Index for fish abundance and biomasswas significantly (PB0.001 and PB0.02, respec-tively) unimodally related to TP (Fig. 2), and forabundance only significantly (PB0.01) related tomean depth. Hurlbert’s index (PIE) was significantlyunimodally related to TP (PB0.003), and positivelyrelated to mean depth (PB0.008) and peaked (0.6bits) in class 3. In contrast to species richness, lakearea did not contribute significantly to the indexvariations in PIE when TP was included as the inde-pendent variable.

Zooplankton species richness declined consider-ably (PB0.0001) with increasing TP (Fig. 1), from amean of 44 in class 1 to 32 in class 5. The decline wasparticularly noticeable for the number of cladoceranspecies (14 to eight), although a pronounced declinewas also observed in the species number of copepodsand rotifers (PB0.0001 and PB0.007), respectively.Multiple regressions revealed that only for rotifersdid mean depth contribute significantly (PB0.0001)to the variation in species richness, the contributionof TP then being insignificant (P\0.25). The Shan-non–Wiener Index, based on both abundance and

Table 1 Frequency distribution of some morphometric and chemical data for the 71 study lakes

MaximumMedianMean 75% Percentile Minimum25% Percentile

Lake area (km2) 2.4 0.4 420.051.80.213.4Mean depth (m) 16.50.941.22.4

TP (mg P L−1) 0.21 0.15 0.08 0.31 0.02 0.995.890.392.831.17TN (mg N L−1) 1.892.08

Chlorophyll a (mg L−1) 87 52 30 131 5 399



Fig. 1 Box-plot showing the species richness ofphytoplankton, total zooplankton and cladocerans, fish,submerged macrophytes and floating-leaved macrophytes infive different TP classes. The full line represents medianvalues. Also shown are 10, 25, 75 and 90% percentiles of thevariables.

diversity indices for the selected taxa when TP wasincluded as an independent variable.

Phytoplankton species richness was significantly(PB0.002) unimodally related to TP, changing froman average of 84 in class 1 to 95 in class 4, followed bya decline to 81 in class 5. Mean depth contributedsignificantly (PB0.05) in a positive manner to thevariation in species richness when TP was included,while lake area did not (P\0.3). The Shannon–Wiener Index, for both abundance and biomass, wassignificantly unimodally related to TP (PB0.002 andPB0.05, respectively). Lake area contributed signifi-cantly and positively to the variation in the diversityindices (PB0.0001 and PB0.02, respectively),whereas mean depth did not (PB0.5). PIE was, how-ever, not related to TP (P\0.5).

Marked changes occurred in species richness ofsubmerged and floating-leaved macrophytes accord-ing to TP. The number of submerged macrophytespecies declined from an average of 11.7 species inclass 1 to 0.5 in class 5. In addition, the maximumdepth distribution at which submerged macrophyteswere recorded declined significantly (Fig. 3; PB0.001). Species richness of floating-leaved macro-phytes tended to be unimodally (though notsignificantly) related to TP, ranging from an averageof 2.1 in class 1 over 3.1 in class 3 to 2.5 in class 5.

Trophic structure

The catch of planktivorous fish by weight (CPUEw)increased significantly (PB0.0001) with TP (Fig. 3).From the lowest to the highest TP class, CPUEw

increased 10-fold. The increase is mainly attributed toroach at low and bream at high TP (Fig. 4). TheCPUEw of both bream and roach was significantlypositively related to TP (PB0.001). Mean depth didnot contribute significantly when TP was included asan independent variable (P\0.2).

With increasing TP, the contribution of piscivorousfish to total CPUEw decreased substantially (PB0.0001), from 62% in class 1 to 10–15% in classes 3–5(Fig. 3). Amongst the piscivores, perch dominated inthe nutrient-poor lakes, whereas northern pike (Esoxlucius L.) and pike perch (Lucioperca lucioperca L.)were most abundant in nutrient-rich lakes (Fig. 4).The latter was introduced into Denmark and occursin only a few lakes. However, the CPUEw of perchwas negatively related to TP and mean depth, while

biomass, was significantly (PB0.001) unimodally re-lated to TP (Fig. 2) and positively related to meandepth (PB0.0001). The indices peaked in class 4(1.9–2 bits). Likewise, PIE was significantly unimo-dally related to TP (PB0.0001). It rose with meandepth (PB0.0001) (Fig. 2) and peaked in class 4 (0.7bits). Lake area did not contribute significantly (P\0.2) to the variation in species richness or any of the



CPUEw of pike was unrelated to both variables. Ben-thivorous ruffe (Gymnocephalus cernuus L.) was sig-nificantly (PB0.0001) unimodally related to TP,being highest in class 3. Only for perch did meandepth contribute significantly to a multiple regressionincluding TP (PB0.05) as perch biomass increasedwith depth.

Besides the changes in biomass, marked changesoccurred in the size/age structure in populationswithin the fish community. For example, the averagebody weight of perch decreased considerably (PB0.01) from 56 g in class 1 to 27 g in class 5 (Fig. 5). Incontrast, pike mean weight increased significantly(PB0.02) with TP, from 650 g in class 1 to 1200–1500 g in classes 4 and 5. No changes were observedfor pike–perch or ruffe (P\0.8 and P\0.2, respec-tively). Amongst the cyprinids, the body weight ofboth the dominant roach and bream declined signifi-cantly (PB0.004 and 0.03, respectively) with TP. Forroach this was 132 g in class 1 to 35–56 g in classes 4and 5, and for bream from 466 g to 188–267 g. Noadditional (P\0.15) effect of mean depth on bodyweight was found for any species (P\0.15) when TPwas included as an independent variable.

In accordance with the major changes in fish com-munity structure, the contribution of Daphnia spp. tothe total biomass of cladocerans decreased signifi-cantly (PB0.0001) with TP (Figs 6 and 7) from 63–70% in classes 1 and 2 to 38% in class 5. The meanindividual body weight of cladocerans also decreasedaccordingly from 5.1 mg in class 1 to 1.5 mg in class 5.The decline in body weight of cladocerans reflectednot only the low relative abundance of Daphnia spp.,but also a reduction in body weight of the remainingDaphnia spp. (PB0.01) and of small cladocerans (PB0.002). The contribution of cyclopoids to the totalabundance and biomass of copepods increased (Fig.6) from an average of 49% in class 1 to 77% in class 5.Daphnia spp. and calanoid copepods contributed 70%of the total biomass in class 1 and 30% in class 5,while the contribution of rotifers to total biomass wasrather low (10–15%) or constant between classes (Fig.6). Multiple regressions revealed that CPUE ofplankti-benthivorous fish by numbers contributedpositively and significantly (PB0.0001) to the varia-tion in the percentage of Daphnia spp., to mean speci-men biomass of cladocerans and to the biomasspercentage of calanoids to total copepods. TP or TN

Fig. 2 Box-plot showing the Shannon–Wiener diversity index based on abundance (H %n) (upper) and biomass (H %w) (middle) andthe Hurlbert PIE index (lower) for fish, zooplankton and phytoplankton. For further details see the Fig. 1 legend.



Fig. 3 August biomass of zooplanktivorous fish (CPUE, catchin multiple mesh-size gill nets, 14 different mesh sizes6.25–75 mm in late summer) versus summer mean lake waterconcentrations of total phosphorus. Also shown are thepercentage of carnivorous fish, summer mean (1 May–1 Oct)of zooplankton:phytoplankton biomass ratio, epilimnionchlorophyll a concentration, Secchi depth and the maximumdepth of submerged macrophytes versus total phosphorus.Mean9SD of the five total phosphorus groups is shown(from Jeppesen et al., 1999).

Fig. 4 Box-plot showing mean fresh weight of various fishspecies caught in multiple mesh-size gill nets in late summer(15 August–15 Sept). Note the different scale on the Y-axis.For further details see the Fig. 1 legend.

0.0001), suggesting lower fish predation pressure indeep rather than in shallow lakes.

The changes in the zooplankton appear to cascadeto the phytoplankton. Consequenty, the meanzooplankton:phytoplankton biomass ratio duringsummer decreased significantly with increasing TP(PB0.0001) from an average of 0.46 in class 1 to0.08–0.15 in classes 3–5 (Fig. 3). Mean depth did notmake a significant contribution (PB0.06). With in-creasing TP, changes occurred in the phytoplanktoncommunity. Dinophytes, chlorophytes, diatoms andchrysophytes dominated at low TP, whereascyanophytes and diatoms were dominant at interme-diate TP, with chlorophytes and cyanophytes beingdominant at high TP (Fig. 8).

did not add significantly to any of the regressions(P\0.2). In addition to TP, the average body weightof cladocerans and the percentage biomass ofcalanoids amongst copepods and the biomass per-centage of Daphnia increased with mean depth (PB



Fig. 5 Box-plot showing average body weight of differentfish species (caught in multiple mesh-size gill nets) in fivedifferent TP classes. For further details see the Fig. 1 legend.

In three eutrophic lakes surveyed intensively for 8years, we observed major changes in fish abundanceduring the study period due either to fish kills or fishmanipulation. Data from these lakes may, therefore,help elucidate the structuring role of fish. In LakeEngelsholm, cyprinids were removed by netting fromApril 1992 to September 1994. In total, 19.2 t freshweight, or 438 kg ha−1, were removed, of whichbream constituted 85%. The highest proportion(11.5 t) was removed in 1992, 2 t in 1993 and 5.7 t in1994. Cyprinids were calculated at 660 kg ha−1 in1990 and at 138–260 kg ha−1 in 1992–96 (Møller,1998). In Lake Arreskov, a major fish kill occurred inautumn and winter 1991 (Sandby, 1998). An addi-tional 4 t of cyprinids were removed in 1995 andduring 1993 and 1995 the lake was stocked with 0+

pike, amounting to a total of 141 individuals per ha.Cyprinid biomass was calculated at 172 kg ha−1 in1987 and 71 kg ha−1 in 1995 (Sandby, 1998). In Lake

Fig. 7 Box-plot showing summer average specimen weight ofvarious genera and groups of zooplankton in five differentTP classes. For further details see the Fig. 1 legend.

Fig. 6 Time-weighted summer mean biomass and percentagecontribution of zooplankton to total biomass in five differentTP classes.



Fig. 8 Time-weighted summer mean biovolume andpercentage contribution of various phytoplankton in fivedifferent TP classes.

reduction in the planktivorous fish biomass resultedin an increase in mean body weight of cladoceransand the zooplankton:phytoplankton biomass ratio,while chlorophyll a decreased (Fig. 9). In comparison,no significant (P\0.3) changes with time were ob-served in any of these variables in a reference set of33 lakes studied during the same period, suggestingthat the changes in the three lakes reflect variations infish abundance rather than inter-annual fluctuationsdetermined by variations in, for example, climate.

Discussion

Species richness and diversity

We observed a different response in species richnessand biodiversity for the various taxonomic groups.For most of the variables included, species richnessand diversity showed a unimodal relationship withTP (Figs 1 and 2) and were also variably related tolake area and/or depth. A different pattern was foundfor species richness of zooplankton and submergedmacrophytes, which declined monotonically with TP.

Numerous studies have shown that species rich-ness of fish in lakes, in accordance with the theory ofisland biography and habitat diversity (MacArthur &

Hejrede, 72 kg ha−1 of cyprinids were removed dur-ing 1991–94. In addition, a major, but unquantified,fish kill occurred under the ice in winter 1995–96(County of Storstrøm, 1996). In all three lakes, the

Fig. 9 Time-weighted summer mean body weight of cladocerans, zooplankton:phytoplankton ratio and chlorophyll a in threelakes during 8 years in which major changes have occurred in cyprinid biomass due to fish removal or fish kill, and in 33other reference lakes (box-plot, see legend of Fig. 1). Arrows show years of major interventions/kills.



Table 2 Correlation analyses of morphometric and nutrientdata based on loge-transformed data

Mean depth TNSurface area

−0.36*TP 0.64***ns– 0.49** nsSurface area

– nsMean depth 0.48**

*PB0.05; **PB0.01; ***PB0.001.

This contrasts with the findings of Dodson (1991,1992) who, by compiling data from 32 European and66 North American lakes, found that species richnessin both sets of lakes increased with lake size. Inaddition, it was unimodally related to phytoplanktonproduction, which in other studies has been shown tobe linked with TP (OECD, 1982). Furthermore, Dod-son (1992) found that species richness was higher inareas rich in lakes. Our study includes only a fewoligotrophic lakes so we cannot exclude the possibil-ity that a unimodal pattern would emerge if morenutrient-poor lakes had been included. Another ex-planation of the high species richness in the lower TPclasses, however, could be that the abundance ofsubmerged macrophytes increases habitat hetero-geneity and then, presumably, the species richness. Inshallow mesotrophic lakes in which the area andvolume occupied by submerged macrophytes may behigh, the presence of macrophytes may have a greatimpact on overall structural complexity. This con-trasts with deep lakes in which the macrophytes arerestricted to near-shore areas. In contrast to speciesrichness, all three selected diversity indices were uni-modally and positively related to TP. The higherspecies richness in the lower TP classes does, there-fore, not result in higher diversity.

For phytoplankton, both species richness and theShannon–Wiener Index were unimodally related toTP, whereas the importance of lake depth and lakearea varied: diversity increasing with lake area andspecies richness with lake depth. As for zooplankton,different patterns may be found in the literature(Margalef, 1978). For example, studies of the histori-cal development in species richness and diversity ofdiatoms in Lake Washington also revealed a unimo-dal relationship with increased eutrophication, whileMargalef (1980) provided evidence of higher diversityin oligotrophic than in eutrophic lakes and marinewaters. The Shannon–Wiener Index for phytoplank-ton, based on abundance, varies from 1.0 to 4.5 bits,but typically falls in the range 2.4–2.6 (Harris, 1986).Our data are at the lower end of this range, whichmay reflect an overall lower diversity in lakes (andshallow coastal estuaries) compared with the openocean (Margalef, 1978, 1980).

Cross-analysis of data from 641 lakes revealed thatlake area contributed most significantly to the varia-tion in macrophyte species richness in Scandinavianlakes (Rørslett, 1991). Residual analyses showed lowspecies richness in lakes with low pH and a unimodal

Wilson, 1967), is strongly linked to lake area (Magnu-son, 1976; Browne, 1981; Keller & Crisman, 1990;Bachman et al., 1996). We found fish species diversityto be independent of lake area, which may, however,be ascribed to the small variation in lake area charac-terizing our study lakes (Table 2). How the speciesrichness and diversity of the fish community changealong a trophic gradient is not clear. While severalstudies indicate a decline in species richness withincreasing eutrophication (for a review see Larkin &Northcote, 1969; Lee, Jones & Jones, 1991), othershave not recorded any changes (Bachman et al., 1996;Eckmann & Rosch, 1998). In our study of mainlyshallow lakes, species richness and the three selecteddiversity indices were all unimodally related to TP.Depth added significantly to species richness anddiversity when based on abundance, which might beexplained by a greater number of available niches indeep lakes.

Most studies have reported an increase in micro-crustacean zooplankton species richness with lakesize (e.g. Patalas, 1972; Fryer, 1985; Dodson, 1991,1992). Others have found a closer correlation withlake depth (Keller & Conlon, 1994), which may beexplained by the larger heterogeneity (Keller & Con-lon, 1994) and perhaps also the lower fish predationpressure in deeper lakes (Keller & Conlon, 1994;Jeppesen et al., 1997). In our study of mainly shallowlakes, however, lake depth contributed significantlyonly to the variation in the species richness of rotifers,but not of cladocerans or copepods. The higher sensi-tivity of rotifers to lake depth may reflect the fact thatrotifer species, unlike most microcrustaceans, areadapted to life in the often oxygen-poor hypolimnion(Hofmann, 1985), implying that the abundance ofspecies is probably higher in stratified than in non-stratified lakes. We found that the species richness ofall selected zooplankton taxa decreased monotoni-cally with TP and was not independent of lake size.



relationship to eutrophication, with mesotrophic–eu-trophic lakes supporting more species than olig-otrophic and hypertrophic lakes. In accordance withthe latter, case studies of cultural eutrophication ofmesotrophic lakes have typically shown a loss inspecies richness (Ozimek & Kowalczewski, 1984;Kowalczewski & Ozimek, 1993, Sand-Jensen, 1997).We also found a significant decrease in macrophytespecies richness in our study. The low contribution ofoligotrophic lakes to our dataset may explain whywe, unlike Rørslett (1991), found a monitonical de-cline with increasing TP. Hence, a recent study ofDanish lakes including more oligotrophic lakesshowed a tendency towards a unimodal relationshipbetween species richness and TP (Vestergaard &Sand-Jensen, unpublished). Compared with the othertaxa in our study, the percentage decline in speciesrichness with increasing TP was particularly high forsubmerged macrophytes. This may reflect the concur-rent significant decrease in maximum depth distribu-tion and macrophyte-covered area with eutrophica-tion and, accordingly, a loss of habitat heterogeneityfor the plants (Fig. 3). For floating-leaved plants,which are less sensitive to increased turbidity, weobserved a unimodal relationship with TP.

The general unimodal response in species richnessand/or diversity for the different taxonomic groups toincreasing TP or trophic state is to be expected (Dod-son, 1992): only a few species occur in distilled waterand hypertrophic conditions lead to loss of speciesdue either to competitive exclusion (Tilman, 1982) oradverse conditions (e.g. high pH, Hansen, Chris-tensen & Sortkjær, 1991; and low oxygen in the hy-polimnion during summer or under ice duringwinter, Magnuson, 1976). Various other factors alsoplay a role, however, and may obscure clear-cut rela-tionships. We have already mentioned morphometry,physical structure and biological complexity in gen-eral. Variation in top-down control also plays a role,and it is generally believed that high predation orgrazing pressure result in loss of diversity of preyorganisms (Paine, 1969), though the effect seems todepend on nutrient state (Proulx & Mazumder, 1998).Increased fish predation may have contributed to thedecline in species richness and diversity of zooplank-ton at high TP, but it does not explain the decline inthe same variables for phytoplankton as the grazingpressure on phytoplankton was low at high TP. Thevariation in disturbance of various kinds may also be

important and it is generally accepted that speciesrichness and diversity are highest at intermediatedisturbances (Connell, 1978).

Trophic structure

The increase in CPUEw and the decline in the percent-age of piscivores with increasing TP agree with theresults from other studies in both temperate andsubtropical regions (Bays & Crisman, 1983; Hanson &Peters, 1984; Persson et al., 1988; Quiros, 1990; Bach-man et al., 1996). The change in the fish communityfrom dominance by perch and pike in mesotrophiclakes to exclusive dominance by cyprinids in eu-trophic lakes is well-documented in other studies ofNorth European lakes (Svardson, 1976; Leach et al.,1977; Persson, 1983; Persson et al., 1988). The superi-ority of roach in eutrophic lakes is attributed to ahigh potential growth rate along with a higher preda-tion efficiency on cladocerans (Persson, 1983) and anability to exploit smaller zooplankton prey (Stenson,1979; Lessmark, 1983). Moreover, cyprinids are omni-vorous, while large perch are piscivores. Finally, theloss of habitat complexity with the disappearance ofsubmerged macrophytes, and thus increased turbid-ity, disfavours percids and, furthermore, augmentsintra-specific competition among them (Persson et al.,1988).

The enhanced dominance of cyprinids was accom-panied by a decline in average size of both cyprinidsand perch. The decline probably reflects enhancedcompetition for food, which in turn may be mediatedby reduced predation of piscivores on young fish.Thus, the fraction of large piscivorous individualsamong the perch population declined. In addition,the average weight of pike increased considerably,probably as a result of enhanced cannibalism due toreduced structural complexity (less macrophytes)and, consequently, the loss of refuges (Grimm &Backx, 1990). As large pike control small prey fish lessefficiently than small pike (Grimm & Backx, 1990), theincapability to control young planktivores decreased.

It is generally believed that ruffe biomass increaseswith increasing TP (Biro, 1977; Hartman & Numann,1977; Persson, 1983; Bergman, 1991). Our study cov-ered a larger TP gradient than in previous work,however, and showed a significant unimodal rela-tionship peaking in class 3 (0.1–0.2 mg P L−1). Theruffe is benthivorous (Johnsen, 1965; Bergman, 1991)



and a competitor to young benthivorous perch. It hasbeen argued that an increase in ruffe abundance withincreasing TP reflects the fact that the foraging abilityof ruffe (unlike that of perch) is largely independentof light (Bergman, 1991). Thus, ruffe may forage effec-tively in turbid eutrophic lakes (Bergman, 1991;Bergman & Greenberg, 1994). However, in eutrophiclakes ruffe faces other competitors. The observed de-cline at high TP is probably caused by the increase inthe abundance of bream (Fig. 3). The bream is sup-posedly a superior competitor in these lakes, bothbecause it is an efficient predator on benthic inverte-brates and because it, unlike ruffe, feeds efficiently inthe pelagic and therefore may alternate between thebenthic and pelagic feeding mode. That ruffe is aninferior competitor at high cyprinid density is evidentfrom several biomanipulation experiments showinghigh abundance of ruffe during the first 1–2 yearsafter cyprinid removal, followed by a major declinewhen young perch, after having initially taken advan-tage of the ‘empty niche’ in the pelagic, reached thebenthivorous state (E. Jeppesen and M. Søndergaard,unpublished). The enhanced importance of pike–perch in eutrophic lakes is well-known from otherEuropean lakes (Svardson, 1976) and in accordancewith their adaptations (high light sensitivity of theireyes, Ali, Ryder & Anctil, 1977) for foraging effi-ciently in turbid water.

Confirming most other studies, zooplanktonbiomass increased with increasing TP (McCauley &Kalff, 1981; Hanson & Peters, 1984). In contrast to anumber of studies (Brooks, 1969; Bays & Crisman,1983), however, we did not find any differences in theproportion of major taxa to biomass. Rotifers typi-cally constituted 10–15% of biomass in all TP classes,this being consistent with results from oligo-mesotrophic Norwegian lakes (Hessen, Faafeng &Andersen, 1995). The results from the Danish andNorwegian lakes suggest that changes in the contri-bution of rotifers are generally not to be expectedover a substantial TP gradient (3–1000 mg P L−1) innorth temperate lakes, which contradicts studies fromsubtropical lakes, in which an increase in rot-ifer:zooplankton biomass was found with increasingtrophic state (Bays & Crisman, 1983). Some studieshave found an increasing share of cladocerans withTP at the expense of copepods (Patalas, 1972;Rognerud & Kjellberg, 1984; Straile & Geller, 1998). Inour study, the contribution of cladocerans and cope-

pods to total biomass in the five TP classes rangedfrom 50 to 58% and 34 to 42%, respectively, and thecontribution of cladocerans tended to decrease andcopepods to increase slightly, though not signifi-cantly, with increasing TP. Major changes occurred inthe contribution of calanoids and cyclopoids to cope-pod biomass, however, as well as in the contributionand mean individual weight of large cladocerans. Ashift from calanoids to cyclopoids with increasingtrophic state has been suggested by several authors(Gliwicz, 1969; Patalas, 1972; Bays & Crisman, 1983)and has been related to changes in food size spectra(Pace, 1986) or predation (Hessen et al., 1995;Jeppesen et al., 1997). Owing to their ability to makeevasive jumps, cyclopoids are often less vulnerable tofish predation than calanoids (Winfield et al., 1983),although there are exceptions (Brooks, 1969). More-over, juvenile cyclopoids might be superior tocalanoids at high food concentrations (Santer & Vanden Bosch, 1994), and increased predation on juvenilecalanoids by adult cyclopoids may further reducecalanoid abundance (Straile & Geller, 1998). Severalauthors record an increased share of small cladocer-ans with increasing trophic state (e.g. McNaught,1975; Patalas, 1972). This is often attributed to en-hanced fish predation (Brooks, 1969; Lyche, 1990).Accordingly, we found an increase in the contribu-tion of Daphnia spp. among cladocerans and of clado-ceran mean body weight after being released fromfish predation after fish removal and fish kills (Fig. 9).World-wide, biomanipulation experiments in eu-trophic lakes have shown similar results providedthat fish biomass was substantially reduced (e.g.Benndorf, 1990; Hansson et al., 1998).

The changes in the zooplankton communitystructure and biomass appeared to cascade downto phytoplankton. Assuming that during summercladocerans ingest organic carbon corresponding to100% of their biomass per day (compared with cope-pods 50% per day and rotifers 200% per day) (Hansenet al., 1992) and assuming that they exclusively feedon phytoplankton, then grazing amounted to 59% ofphytoplankton biomass per day in class 1 and 16–19% per day in classes 4 and 5. The latter is so lowthat zooplankton probably have little effect on phyto-plankton growth, while in class 1 and 2 lakes it seemsprobable that zooplankton apply a considerable graz-ing pressure to phytoplankton. In addition, the ob-served reduction in the individual weight of clado-



cerans and Daphnia spp. presumably further ad-versely affected the grazing capacity of phytoplank-ton, because small-bodied cladocerans are lessefficient grazers on large-sized phytoplankton thanthe large-bodied forms (Gliwicz, 1977, 1990). How thealteration in fish community structure and biomass,mediated by the changes in trophic state, affects her-bivory on phytoplankton has been the subject ofextensive discussion (DeMelo, France & McQueen,1992; Carpenter & Kitchell, 1992). Our data, as well asthose of Leibold (1990) and Sarnelle (1992), suggest amore significant effect in eutrophic lakes than inmesotrophic lakes.

Several authors have argued that filamentouscyanobacteria prevent grazer control by zooplanktonin eutrophic lakes, which may explain a decrease inherbivory from mesotrophic to eutrophic lakes (Elser& Goldman, 1990; Carney & Elser, 1990). Supportingthis view, controlled laboratory experiments haveshown that dense cyanobacterial assemblages can pre-vent the growth of large cladocerans (e.g. Lampert,1981; Dawidovicz, Gliwicz & Gulati, 1988; Gliwicz,1990). Furthermore, in field experiments Daphnia spp.have occasionally failed to respond to minor, and in afew cases even to major, reductions in planktivorousfish populations (e.g. Van Donk et al., 1990; Riemannet al., 1990; Moss, Stansfield & Irvine, 1991; DeMelo,France & McQueen, 1992). The results from Danishlakes indicate, however, that the role of cyanobacteriain the decline in herbivory from mesotrophic to eu-trophic lakes is considerably less significant than thatof fish. First, the zooplankton:phytoplankton ratio andthe size of cladocerans are low not only in eutrophiclakes dominated by cyanobacteria but also in the mostnutrient-rich lakes dominated by edible green algae.Second, a reduction in fish predation on zooplankton,due either to the removal of cyprinids or to fish kills,caused a major increase in the abundance and averagesize of cladocerans and a decrease in phytoplanktonbiomass in eutrophic lakes, irrespective of whethercyanobacteria or chlorophytes dominated previously(Jeppesen et al., 1997; Søndergaard, Jeppesen &Jensen, 1998; Jeppesen et al., 1999; authors’ unpub-lished data). Finally, whereas in Danish lakes CPUEw

contributed highly significantly to the variation in thezooplankton:phytoplankton ratio and cladoceranmean size, the percentage contribution of cyanobacte-ria to the total biomass of phytoplankton did not(Jeppesen et al., 1997).

The strong evidence of a decrease in herbivory,mediated by planktivorous fish from mesotrophic tohypertrophic lakes implies that the cascading effect offish manipulation on phytoplankton in shallow lakeswill be greatest under hypertrophic conditions,though the benefit of such a manipulation may betransient as planktivorous fish are highly successful insuch lakes (Fig. 3; Kitchell et al., 1977; Leach et al.,1977; Persson et al., 1988; Jeppesen et al., 1990). There-fore, a return to a turbid state and to low zooplanktongrazing seems likely, even though it may be some-what delayed in shallow lakes with an extensivegrowth of submerged macrophytes (Meijer et al.,1994). The data in Fig. 3 suggest that the likelihood ofobtaining high zooplankton grazing in shallow lakesincreases markedly when summer mean TP dropsbelow about 0.1 mg P L−1. Therefore, the prospects oflong-term success of biomanipulation in shallow lakesare most probably greatest below this threshold(Jeppesen et al., 1990). The results shown in Figs 3–8may leave the impression that the shift from clear toturbid conditions occurred gradually with increasingTP. However, several case studies of shallow lakeshave shown a step-wise shift and, within certainnutrient regimes (typically 0.5–0.15 mg P L−1;Jeppesen et al., 1990), lakes may shift between the twostates (Moss, 1990; Jeppesen et al., 1990; Scheffer,1990), this being the case in Danish lakes (Jeppesen etal., 1990). Lake-specific differences in nutrientthresholds, and the fact that we also included some-what deep lakes, may explain the lack of step-wiseshifts in the present study.

In summary, species richness and diversity of mostof the selected taxonomic groupings were unimodallyrelated to TP in the nutrient range from 0.02 to 1.0 mgP L−1 and, moreover, often related positively to meandepth and/or lake area. Only the species richness ofzooplankton and submerged macrophytes deviatedby showing a monotonical decline with increasing TP.Major changes occurred in the community structureand size distribution of fish, which apparentlycascaded to the lower trophic levels. The share ofpiscivores declined markedly and, moreover, thesize of the predatory fish changed towards thosesize classes less likely to control cyprinids (largerpike and smaller perch). Enhanced predation ap-peared to lead to a marked reduction of the zooplank-ton:phytoplankton biomass ratio and the grazing



pressure upon phytoplankton. The latter may havebeen exacerbated by a decline in the size of grazingzooplankton.

Acknowledgments

The authors thank the counties for access to datafrom the Survey Programme of Danish Lakes andMartin R. Perrow for valuable comments. The studywas supported by the Danish Natural Science Re-search Council (Grant 9601711) and the research pro-gramme ‘The Role of Fish in Ecosystems, 1999–2001’,funded by the Ministry of Agriculture, Fisheries andFood. They thank Anne Mette Poulsen, KatheMøgelvang and Henrik F. Rasmussen for skillfultechnical assistance.

References

Ali M.A., Ryder R.A. & Anctil M. (1977) Photorecep-tors and visual pigments as related to behavioralresponses and preferred habitat of perches (Percaspp.) and pikeperches (Stizostedion spp.). Journal ofthe Fisheries Research Board of Canada, 34, 1475–1480.

Bachman R.W., Jones B.L., Fox D.D., Hoyer M., BullL.A. & Canfield D.E. (1996) Relations betweentrophic state indicators and fish in Florida (U.S.A.)lakes. Canadian Journal of Fisheries and Aquatic Sci-ences, 53, 842–855.

Bays J.S. & Crisman T.L. (1983) Zooplankton andtrophic state relationships in Florida lakes. CanadianJournal of Fisheries and Aquatic Sciences, 40, 1813–1819.

Benndorf J. (1990) Conditions for effective biomanip-ulation; conclusions derived from whole-lake ex-periments in Europe. Hydrobiologia, 200/201,187–203.

Bergman E. (1991) Changes in abundance of twopercids, Perca fluviatilis and Gymnocephalus cernua,along a productivity gradient. Canadian Journal ofFisheries and Aquatic Sciences, 48, 536–545.

Bergman E. & Greenberg L.A. (1994) Competitionbetween a planktivore, a benthivore, and a specieswith ontogenetic diet shifts. Ecology, 75, 1233–1245.

Biro P. (1977) Effects of exploitation, introductionsand eutrophication on percids on Lake Balaton.Journal of the Fisheries Research Board of Canada, 34,1678–1683.

Bottrell H.H., Duncan A., Gliwicz Z.M., Grygierek E.,Herzig A., Ilkowska-Hillbricht A., Kurazawa H.,Larsson P. & Weglenska T. (1976) A review of someproblems in zooplankton production studies. Nor-wegian Journal of Zoology, 24, 419–456.

Brett M.T. & Goldman C.R. (1996) Consumer versusresource control in freshwater pelagic food webs.Science, 275, 384–386.

Brooks J.L. (1969) Eutrophication and changes in thecomposition of zooplankton. In: Eutrophication,Causes, Consequences, Correctives, pp. 236–255. Na-tional Academy of Sciences, Washington, DC.

Brooks J.L. & Dodson S.I. (1965) Predation, body sizeand composition of plankton. Science, 150, 28–35.

Browne R.A. (1981) Lakes as islands: biogeographicdistribution, turnover rates, and species composi-tion in the lakes of central New York. Journal ofBiogeography, 8, 75–83.

Carney J.J. & Elser J.J. (1990) The strength ofzooplankton–phytoplankton coupling in relation totrophic state. In: Ecology of Large Lakes (ed. M.M.Tilzer), pp. 616–631. Springer, New York.

Carpenter S.R. & Kitchell J.F. (1992) Trophic cascadeand biomanipulation: interface of research andmanagement—a reply to the comment by DeMeloet al. Limnology and Oceanography, 37, 208–213.

Carpenter S.R. & Kitchell J.F. (Eds) (1993) The TrophicCascade in Lakes. Cambridge University Press,Cambridge.

Carpenter S.R., Kitchell J.F. & Hodgson J.R. (1985)Cascading trophic interactions and lake productiv-ity. BioSciences, 35, 634–639.

Chutter F.M. (1972) An empirical biotic index of thequality of water in South African streams and riv-ers. Water Research, 6, 19–30.

Connell J.H. (1978) Diversity in tropical rain forestsand coral reefs. Science, 199, 1302–1310.

County of Storstrøm (1996) Hejrede sø—overvag-ningsdata 1995. Teknik- og Miljøforvaltningen, 36 pp.(In Danish).

Dawidovicz P., Gliwicz Z.M. & Gulati R.D. (1988)Can Daphnia prevent a blue–green phytoplanktonbloom in hypertrophic lakes? A laboratory test.Limnologica, 19, 21–26.

DeMelo R., France R. & McQueen D.J. (1992) Bioma-nipulation: hit or myth? Limnology and Oceanogra-phy, 37, 192–207.

Dodson S.I. (1991) Species richness of crustaceanzooplankton in European lakes of different sizes.



Verhandlungen der internationale Vereinigung fur theo-retische und angewandte Limnologie, 24, 1213–1229.

Dodson S.I. (1992) Predicting crustacean zooplanktonspecies richness. Limnology and Oceanography, 37,848–856.

Dumont H.J., Van De Velde I. & Dumont S. (1975)The dry weight estimate of biomass in a selection ofCladocera, Copepoda and Rotifera from plankton,periphyton and benthos of continental waters.Oecologia, 19, 75–97.

Eckmann R. & Rosch R. (1998) Lake Constancefisheries and fish ecology. Archiv fur Hydrobiologie,53, 285–301.

Edler L. (1979) Recommendations on methods for marinebiological studies in the Baltic Sea. Phytoplankton andchlorophyll. Baltic Marine Biologists Working Group9, 38 p.

Elser J.J. & Goldman C.R. (1990) Zooplankton effectson phytoplankton in lakes of contrasting trophicstatus. Limnology and Oceanography, 36, 64–90.

Flecker A.S. & Townsend C.R. (1994) Community–wide consequences of trout introduction in NewZealand streams. Ecological Applications, 4, 798–807.

Fretwell S.D. (1977) The regulation of plant communi-ties by food chains exploiting them. Perspectives inBiology and Medicine, 20, 169–185.

Fryer G. (1985) Crustacean diversity in relation to thesize of water bodies: some facts and problems.Freshwater Biology, 15, 347–361.

Gliwicz Z.M. (1969) Studies on the feeding of pelagiczooplankton in lakes with varying trophy. Ekologiapolska, 17, 663–708.

Gliwicz Z.M. (1977) Food size selection and seasonalsuccession of filter feeding zooplankton in a eu-trophic lake. Ekologia polska, 17, 179–225.

Gliwicz Z.M. (1990) Why do cladocerans fail to con-trol phytoplankton blooms? Hydrobiologia, 200/201,83–97.

Goodman D. (1975) The theory of diversity–stabilityrelationships in ecology. Quarterly Review of Biology,50, 237–266.

Grimm M.P. & Backx J. (1990) The restoration ofshallow eutrophic lakes and the role of northernpike, aquatic vegetation and nutrient concentration.Hydrobiologia, 200/201, 557–566.

Gulati R.D., Lammens E.H.H.R., Meijer M.-L. & vanDonk E. (1990) Biomanipulation, tool for watermanagement. Hydrobiologia, 200/201, 1–628.

Hall D.J., Cooper W.E. & Werner E.E. (1970) Anexperimental approach to the production dynamicsand structure of freshwater animal communities.Limnology and Oceanography, 15, 839–928.

Hansen A-M., Christensen J.V. & Sortkjær O. (1991)Effect of high pH on zooplankton and nutrients infish-free enclosures. Archiv fur Hydrobiologie, 123,143–164.

Hansen A.M., Jeppesen E., Bosselmann S. & An-dersen P. (1992) Zooplankton i søer—metoder og art-skodeliste, 1148. Miljøministeriet, Miljøprojekt 205.In Danish.

Hanson J.M. & Peters R.H. (1984) Empirical predic-tion of crustacean zooplankton biomass and pro-fundal macrobenthos biomass in lakes. CanadianJournal of Fisheries and Aquatic Sciences, 41, 439–445.

Hansson L.A. (1992) Effects of competitive interac-tions on the biomass development of planktonicand periphytic algae in lakes. Limnology andOceanography, 73, 241–247.

Hansson L.A., Annadotter H., Bergman E., HamrinS.F., Jeppesen E., Kairesalo T., Luokkanen E.,Nilsson P.-A, ., Søndergaard M. & Strand J. (1998)Biomanipulation as an application of food chaintheory: constraints, synthesis and recommendationsfor temperate lakes. Ecosystems, 1, 558–574.

Harris G.P. (1986) Phytoplankton Ecology. Structure,Function and Fluctuation. Chapman and Hall,London.

Hartman J. & Numann W. (1977) Percids of LakeConstance, a lake under eutrophication. Journal ofthe Fisheries Research Board of Canada, 34, 1670–1677.

Hessen D.O., Faafeng B.A. & Andersen T. (1995)Replacement of herbivore zooplankton speciesalong gradients of ecosystem productivity and fishpredation pressure. Canadian Journal of Fisheries andAquatic Sciences, 52, 733–742.

Hofmann W. (1985) Dynamics of vertical zooplanktoncommunity strucure in the Pluss-see: cluster analy-sis. Verhandlungen der internationale Vereinigung furtheoretische und angewandte Limnologie, 22, 2983–2986.

Hrbacek J., Dvorakova V., Korinek V. & ProchazkovaL. (1961) Demonstration of the effect of the fishstock on the species composition of zooplanktonand the intensity of metabolism of the whole plank-ton association. Verhandlungen der internationaleVereinigung fur theoretische und angewandte Limnolo-gie, 14, 192–195.



Hurlbert S.H. (1971) The nonconcept of species diver-sity: a critique and alternative parameters. Ecology,52, 577–586.

Jeppesen E., Jensen J.P., Kristensen P., SøndergaardM., Mortensen E., Sortkjær O. & Olrik K. (1990)Fish manipulation as a lake restoration tool in shal-low, eutrophic, temperate lakes 2: threshold levels,long-term stability and conclusions. Hydrobiologia,200/201, 219–227.

Jeppesen E., Jensen J. P., Søndergaard M., LauridsenT., Pedersen L.J. & Jensen L. (1997) Top-down con-trol in freshwater lakes: the role of nutrient state,submerged macrophytes and water depth. Hydrobi-ologia, 342/343, 151–164.

Jeppesen E., Jensen J.P., Windolf J., Lauridsen T.,Søndergaard M., Sandby K. & Hald Møller P.(1998) Changes in nitrogen retention in shalloweutrophic lakes following a decline in density ofcyprinids. Archiv fur Hydrobiologie, 142, 129–152.

Jeppesen E., Søndergaard M, Kronvang B., Jensen J.P.,Svendsen L.M. & Lauridsen T. (1999) Lake andcatchment management in Denmark. Hydrobiologia,395/396, 419–432.

Jespersen A.-M. & Christoffersen K. (1987) Measure-ments of chlorophyll a from phytoplankton usingethanol as extraction solvent. Archiv fur Hydrobiolo-gie, 109, 445–454.

Johnsen P. (1965) Studies on the distribution and foodof the ruffe (Acerina cernua L.) in Denmark, withnotes on other aspects. Meddelelser fra DanmarksFiskeri– og Havundersøgelser. New Series, 4, 137–156.

Keller W. & Conlon M. (1994) Crustacean zooplank-ton communities and lake morphometry in precam-brian shield lakes. Canadian Journal of Fisheries andAquatic Sciences, 51, 2424–2434.

Keller A.E. & Crisman T.L (1990) Factors influencingfish assemblages and species richness in subtropicalFlorida lakes and a comparison with temperatelakes. Canadian Journal of Fisheries and Aquatic Sci-ences, 47, 2137–2146.

Kitchell J.F., Johnson M.G., Minss C.K., Loftus K.H.,Greig L. & Olver C.H. (1977) Percid habitat; theriver anology. Journal of the Fisheries Research Boardof Canada, 34, 1959–1963.

Kowalczewski A. & Ozimek T. (1993) Further long-term changes in the submerged macrophyte vegeta-tion of the eutrophic Lake Mikolajskie (NorthPoland). Aquatic Botany, 46, 341–345.

Kristensen P, Jensen J.P. & Jeppesen E. (1990) Simpleempirical lake models. In: Nitrogen and Phosphorusin Fresh and Marine Waters. C-abstracts, pp. 125–145. Miljøstyrelsen, Copenhagen.

Lampert W. (1981) Inhibitory and toxic effects ofblue–green algae on Daphnia. Internationale Revueder gesamten Hydrobiologie, 66, 285–298.

Larkin P.A. & Northcote T.G. (1969) Fish as indices ofeutrophication. In: Eutrophication, Causes, Conse-quences, Correctives. Proceedings of a Symposium, 11–15 June 1967, Madison, WI, pp. 256–273. NationalAcademy of Sciences, Washington, DC.

Leach J.H., Johnson M.G., Kelso J.R.M., Harmann J.,Numann W. & Entz B. (1977) Responses of percidfishes and their habitats to eutrophication. Journal ofthe Fisheries Research Board of Canada, 34, 1964–1971.

Lee G.F., Jones P.E. & Jones R.A. (1991) Effects ofeutrophication on fisheries. Review of Aquatic Sci-ences, 5, 287–305.

Leibold M.A. (1990) Resource edibility and the effectsof predators and productivity on the outcome oftrophic interactions. American Naturalist, 134, 922–949.

Lessmark O. (1983) Competition between perch (Percafluviatilis) and roach (Rutilus rutilus) in southSwedish lakes. Dissertation, University of Lund,Sweden.

Lyche A. (1990) Cluster analysis of plankton commu-nity structure in 21 lakes along a gradient of tro-phy. Verhandlungen der internationale Vereinigung furtheoretische und angewandte Limnologie, 24, 586–591.

Magnuson J.J. (1976) Managing with exotics – a gameof chance. Transactions of the American Fish Society,105, 10.

Margalef R. (1978) Diversity. In: Phytoplankton Manual(ed. A. Sournia). UNESCO, Paris.

Margalef R. (1980) Perspectives in Ecological Theory.University Chicago Press, Chicago, IL.

MacArthur R. & Wilson E.O. (1967) The Theory ofIsland Biogeography (foliage height diversity), pp. 8–67. Princeton University Press, Princeton, NJ, 109,111, 112.

Mazumder A. (1994) Patterns of algal biomass indominant odd- vs. even-link lake ecosystems. Ecol-ogy, 75, 1141–1149.

McCauley E. & Kalff J. (1981) Empirical relationshipsbetween phytoplankton and zooplankton biomassin lakes. Canadian Journal of Fisheries and AquaticSciences, 38, 458–463.



McIntosh A.R. & Townsend C.R. (1994) Inter-popula-tion variation in mayfly anti–predator tactics: differ-ential effects. Ecology, 75, 2078–2090.

McNaught D.C. (1975) A hypothesis to explain thesuccession from calanoids to cladocerans duringeutrophication. Verhandlungen der internationaleVereinigung fur theoretische und angewandte Limnolo-gie, 19, 724–731.

McQueen D.J., Post J.R. & Mills E.L. (1986) Trophicrelationships in freshwater pelagic ecosystems.Canadian Journal of Fisheries and Aquatic Sciences, 43,1571–1581.

McQueen D.J., Johannes M.R.S., Post J.R, Steward T.J.& Lean D.R.S. (1989) Bottom–up and top–downimpacts on freshwater pelagic community structure.Ecological Monographs, 59, 289–310.

Meijer M.-L., Jeppesen E., Van Donk E., Moss B.,Scheffer M., Lammens E., Van Nes J.A., de JongBerkum G.J, Faafeng B.A. & Jensen J.P. (1994) Long-term responses to fish-stock reduction in small shal-low lakes: Interpretation of five-year results of fourbiomanipulation cases in the Netherlands and Den-mark. Hydrobiologia, 275/276, 457–466.

Moss B. (1990) Engineering and biological approachesto the restoration from eutrophication of shallowlakes in which aquatic plant communities are impor-tant components. Hydrobiologia., 200/201, 367–378.

Moss B., Stansfield J. & Irvine K. (1991) Developmentof daphnid communities in diatom- andcyanophyte-dominated lakes and their relevance tolake restoration by biomanipulation. Journal of Ap-plied Ecology, 28, 586–602.

Møller P.H. (1998) Engelsholm sø. In: Sørestaurering iDanmark: metoder, erfaringer og anbefalinger (eds M.Søndergaard, E. Jeppesen & J.P. Jensen), pp. 145–153. Miljøstyrelsen, Miljønyt. In Danish.

OECD (1982) Eutrophication of Waters. Monitoring, As-sessments and Control. OECD, Paris.

Oksanen L.D., Fretwell S.D., Arruda J. & Niemela P.(1981) Exploitation of ecosystems in gradients ofprimary productivity. American Naturalist, 118, 240–261.

Ozimek T. & Kowalczewski A. (1984) Long–termchanges of the submerged macrophytes in eutrophicLake Mikolajskie (North Poland). Aquatic Botany, 19,1–11.

Pace M.L. (1986) An empirical analysis of zooplanktoncommunity size structure across lake trophic gradi-ents. Limnology and Oceanography, 31, 45–55.

Paine R.T. (1969) The Pisaster–Tegula interaction:prey patches, predator food preference, and inter-tidal community structure. Ecology, 50, 950–961.

Patalas K. (1972) Crustacean plankton and the eu-trophication of St. Lawrence Great Lakes. CanadianJournal of Fisheries and Aquatic Sciences, 29, 1451–1462.

Patalas J. & Patalas K. (1966) The crustacean planktoncommunities in Polish lakes. Verhandlungen der in-ternationale Vereinigung fur theoretische und ange-wandte Limnologie, 16, 204–215.

Persson L. (1983) Effects of intra- and interspecificcompetition on dynamics and size structure of aperch Perca fluviatilis and a roach Rutilus rutiluspopulation. Oikos, 41, 126–132.

Persson L., Andersson G., Hamrin S.F. & Johansson L.(1988) Predation regulation and primary productionalong the productivity gradient of temperate lakeecosystems. In: Complex Interactions in Lake Commu-nities (ed. S.R. Carpenter), pp. 45–65. Springer, NewYork.

Persson L., Diehl S., Johansson L., Andersson G. &Hamrin S.F. (1992) Trophic interactions in temperatelake ecosystems: a test of food chain theory. Ameri-can Naturalist, 140, 59–84.

Proulx M. & Mazumder A. (1998) Reversal of grazingimpact on plant species richness in nutrient-poor vs.nutrient-rich ecosystem. Ecology, 79, 2581–2592.

Quiros R. (1990) Factors related to variance of residu-als in chlorophyll – total phosphorus regressions inlakes and reservoirs of Argentina. Hydrobiologia,200/201, 343–355.

Reed C. (1978) Species diversity in aquatic microsys-tems. Ecology, 59, 481–488.

Reynolds C.F. (1984) The Ecology of Freshwater Phyto-plankton. Cambridge University Press, Cambridge.

Riemann B., Christoffersen K., Jensen H.J., Muller J.P.,Lindegaard C. & Bosselmann S. (1990) Ecologicalconsequences of a manual reduction of roach andbream in a eutrophic, temperate lake. Hydrobiologia,200/201, 241–250.

Rognerud S. & Kjellberg G. (1984) Relationship be-tween phytoplankton and zooplankton biomass inlarge lakes. Verhandlungen der internationale Vereini-gung fur theoretische und angewandte Limnologie, 22,666–671.

Rott E. (1981) Some results from phytoplankton count-ing intercalibrations. Schweiziche Zeitschrift fur Hy-drologie, 43, 34–62.



Rørslett B. (1991) Principal determinants of aquaticmacrophyte richness in northern European lakes.Aquatic Botany, 39, 173–193.

Sandby K. (1998) Arreskov Sø. In: Sørestaurering iDanmark: metoder, erfaringer og anbefalinger (eds M.Søndergaard, E. Jeppesen & J.P. Jensen), pp. 105–115. Miljøstyrelsen, Miljønyt 28. In Danish.

Sand-Jensen K. (1997) Eutrophication and plant com-munities in Lake Fure during 100 years. In: Fresh-water Biology. Priorities and Development in DanishResearch (eds K. Sand-Jensen & O. Pedersen), 1997.GEC Gad, Copenhagen.

Santer B. & Van den Bosch F. (1994) Herbivorosnutrition of Cyclops vicinus : the effect of a purealgal diet on feeding, development, reproductionand life cycle. Journal of Plankton Research, 16, 171–195.

Sarnelle O. (1992) Nutrient enrichment and grazereffect on phytoplankton in lakes. Ecology, 73, 551–560.

Scheffer M. (1990) Multiplicity of stable states infreshwater systems. Hydrobiologia, 200/201, 475–486.

Stenson J.A.E. (1979) Predatory-prey relations be-tween fish and invertebrate prey in some forestlakes. Report from the Institute of Freshwater Research,Drottningholm, 58, 166–183.

Stockner J.G. & Benson W.W. (1967) The succession ofdiatom assemblages in the recent sediment of LakeWashington. Limnology and Oceanography, 12, 513–552.

Straile D. & Geller W. (1998) Crustacean zooplanktonin Lake Constance from 1920 to 1995: response toeutrophication and re-oligotrophication. Archiv furHydrobiologie, 53, 255–274.

Svardson G. (1976) Interspecific population domi-nance in fish communities of Scandinavian lakes.Report from the Institute of Freshwater Research. Drot-tningholm, 56, 144–171.

Søndergaard M., Kristensen P. & Jeppesen E. (1992)Phosphorus release from resuspended sediment inthe shallow and wind exposed Lake Arresø, Den-mark. Hydrobiologia, 228, 91–99.

Søndergaard M., Jeppesen E. & Jensen J.P. (Eds)(1998) Sørestaurering i Danmark: metoder, erfaringer oganbefalinger. Miljøstyrelsen, Miljønyt 28. In Danish.

Tilman D. (1982) Resource Competition and CommunityStructure. Monographs in Population Biology, vol. 17.Princeton.

Utermohl H. (1958) Zur Vervollkomnung der quanti-tativen Phytoplankton-Methodik. Mitteilungen derinternationale Vereinigung fur theoretische und ange-wandte Limnologie, 9, 1–38.

Van Donk E., Grimm M.P., Gulati R.D. & KleinBreteler J.P.G. (1990) Whole-lake food-web manipu-lation as a means to study community interactionsin a small ecosystem. Hydrobiologia, 200/201, 275–291.

Washington H.G. (1984) Diversity, biotic and similar-ity indices. A review with special relevance toaquatic ecosystems. Water Research, 18, 653–694.

Winfield I.J., Peirson G., Cryer M. & Townsend C.(1983) The behavioural basis of prey selection byunderyearling bream (Abramis brama) and roach(Rutilus rutilus). Freshwater Biology, 13, 139–149.

Wurtsbaugh W.A. (1992) Food–web modification byan invertebrate predator in the Great Lake (U.S.A.).Oecologia, 89, 168–175.

(Manuscript accepted 5 June 2000)


Trophic structure, species richness and biodiversity in Danish lakes: changes along a phosphorus...

Documents

Transcript of Trophic structure, species richness and biodiversity in Danish lakes: changes along a phosphorus...