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The influence of ecological processes on the accumulation of persistent organochlorines in aquatic

Olof Berglund

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Department of Chemical Ecology and Lund University, Sweden Lund 1999 DISCLAIMER

Portions of this document may be illegible in electronic image products. Images are produced from the best available original document. The influence of ecologicalprocesses on the accumulation of persistent organochlorinesin aquatic ecosystems

Olof Berglund

Akademisk avhandling, som for avlaggande av filosofie doktorsexamen vid matematisk-naturvetenskapliga fakulteten vid Lunds Universitet, kommer att offentligen forsvaras i Bla Hallen, Ekologihuset, Solvegatan 37, Lund, fredagen den 17 September 1999 kl. 10.

Fakultetens opponent: Prof. Derek C. G. Muir, National Research Institute, Environment Canada, Burlington, Canada. Avhandlingen kommer att forsvaras pa engelska.

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1 7 I lf tO CIC 103 The influence of ecological processes on the accumulation of persistent organochlorines in aquatic ecosystems

Olof Berglund

Dissertation Lund, 1999 A doctoral thesis at a university in Sweden is produced either as a monograph or a collection of papers. In the latter case, the introductory part constitutes the formal thesis, which summarises the accompanying papers. These have either already been published or are in manuscripts at various stages (in press, submitted or in ms).

© Olof Berglund ISBN 91-7105-114-7 SE-LUNBDS/NBKE-99/1016+144pp

2 This thesis is based on the following papers, which are referred to by their Roman numerals.

I. Berglund, O., Larsson, P., Ewald, G., and Okla, L. Bioaccumulation and differential partitioning of PCBs in freshwater, planktonic food webs. (■Submitted)

II. Berglund, O., Larsson, P., and Broman, D. Organochlorine accumulation and stable isotopes in an Atlantic salmon (Salmo salar) population from the Baltic - effects of omnivory and reproductive strategies. (Submitted)

III. Berglund, O., Larsson, P., Ewald, G., and Okla, L. The effect of lake trophy on lipid content and PCB concentrations in planktonic food webs. (Manuscript)

IV. Berglund, O., Larsson, P., Ewald, G., and Okla, L. Influence of trophic status on PCB dynamics in lake sediments and biota. {Manuscript)

V. Berglund, O., Larsson, P., Bronmark, C., Greenberg, L., Eklov, A., and Okla, L. 1997. Factors influencing organochlorine uptake in age-0 brown trout {Salmo trutta) in lotic environments. Can. J. Fish. Aquat. Sci. 54: 2767-2774.

Paper V is reprinted with permission from the publisher.

3 Till Mor och Far.

4 Contents:

Introduction 7 OC accumulation in aquatic organisms 14 Lipid content and composition 15 Elimination 16 Size, age, and growth rate 16 Biomagnification ofOCs 17 Stable isotopes and effects 19 Trophic status and OC accumulation: lakes 21 Biomass 21 Sedimentation 23 Volatilisation 25 Recycling 25 Stratification 26 Trophic status and OC accumulation: streams 27 Spiralling and biomass 27

Heterotrophy vj . autotrophy 29 Watershed influence 29 Conclusions 30 Future research 32 77ze dynamics ofOCs in streams 32 77ze dynamics of lipids in ecosystems 33 References 34 Tack 43

5 6 Introduction The contamination of aquatic ecosystems by persistent organochlorines (OCs) has been of major concern for decades. Due to their persistence to chemical and microbial breakdown, OCs are today distributed worldwide, even in regions were they have never been applied (Muir et al. 1990; Tanabe et al. 1983). The OCs are transported in the atmosphere and redistributed both on a global and local scale via volatilisation and subsequent washout by precipitation, dry deposition, or gas exchange (Bidleman 1988; Simonich and Hites 1994, 1995). The process of global redistribution of OCs has been termed "global distillation”. The global distillation theory predicts that compounds will be fractionated latitudinally, depending on the ambient temperature and the volatility of the compound (Wania and Mackay 1993). Compounds with low volatility will remain in warmer regions, while compounds of higher volatility will be transported to colder regions, far from the source. Polychlorinated biphenyls (PCBs) and , both considered "semivolatile", are predicted to deposit in temperate regions (Wania and Mackay 1993), where, as an effect, concentrations in organisms may be higher than in the source areas (Larsson et al. 1995). Today, the use and production of most OCs have been banned or severely restricted in most industrialised countries. The main source of OCs to aquatic systems in temperate regions is, therefore, considered to be atmospheric deposition (Swackhamer and Hites 1988; Larsson and Okla 1989; Muir et al. 1990). The atmospherically transported OCs reach the freshwater systems via different routes. A fraction is deposited directly on lake or river surfaces via wet and dry deposition, or air-water gas exchange over the water surface. The major part of the OCs, however, is deposited on land in the watershed areas of the rivers and lakes. Once deposited, several different ecological processes in the watershed area, the streams and rivers, and the lakes affect the transport and fate of the OCs, before reaching the oceans (Fig. 1). After deposition on the ground or soil surface, the OCs can follow several different transport pathways before reaching the freshwater ecosystems. The nature of the watershed area will, therefore, have a major influence on the transport of the OCs to the rivers and lakes. In the catchment area, rain-, or meltwater is being transported downhill towards the stream channels or ground water. Because of the lipophilic properties of OCs, they readily adsorb to particles and the partitioning is dependent on organic carbon content, lipid content and size and structure of the particles (Nau-Ritter et al. 1982; Eisenreich 1987; Swackhamer and Skoglund 1993). The partitioning coefficient for OCs between water and particles is often normalised on an organic carbon or lipid basis, and considered proportional to the partitioning of OCs between octanol and water (K0VI) (Mackay 1982; Chiou 1985). The partitioning coefficients for OCs are approximately 10s - 107 , which means that lipids at equilibrium contain approximately a million times the

7 atmospheric deposition

WATERSHED AREA vegetation soil v organic carbon y LAKES nutrient regime biomass overland flow microbial recycling subsurface flow

spiralling volatilisation

groundwater flow ^ OCEANS ^ salmon vectors z STREAMS biomass carbon source sedimentation v food web >

Figure 1. Ecological and physical procecess affecting the fate and transport of OCs from the atmosphere to the oceans. concentration of OCs than water. Hence, much of the OCs deposited on the ground via the atmosphere will sorb to soil particles and vegetation where they partition to the organic carbon and lipids. Therefore, the retention, transport and fate of the deposited OCs is intimately linked to the fate and transport of organic carbon. Much of the transport of organic carbon to streams and rivers in the watershed is water mediated; both in the form of suspended particles and dissolved organic carbon. Rain- or meltwater in the watershed can reach the streams and rivers following three major pathways (Fig. 2); overland flow, sub ­ surface flow and groundwater flow (Allan 1995). The dominating pathway may greatly influence the amount of OCs reaching the streams. Overland flow occurs when the water infiltration capacity of the soil is exceeded. When exceeded, the rainfall in excess of the infiltration capacity accumulates on the soil surface and moves as an irregular sheet of overland flow. Overland flow is likely in arid areas, when the soil surface is frozen, and where human activities make the land surface less permeable, like urbanisation and agriculture. Water that penetrates the soil reaches the groundwater, from which it discharges to the stream slowly and over a long period of time. If the soil structure includes a relatively impermeable layer underlying highly permeable topsoil, water will accumulate at that layer and move downhill, reaching the streams laterally through their banks (Allan

8 Precipitation

Figure 2. Pathways of water moving downhill. Overland flow (1) occurs when precipitation exceeds the infiltration capacity of the soil. A relatively impermeable layer will cause water to move laterally through the relatively permeable topsoil as shallow sub-surface flow (2). Water that permeates the two layers adds to groundwater flow (3) that may reach the streams, lakes or the oceans. (Redrawn from Allan 1995).

1995). This movement is termed sub-surface flow, and is slower than overland flow but faster than that of groundwater. The water pathway to the streams determines the amount and type of organic carbon within the water. During overland flow, great amounts of particulate matter transport from the catchment area occur. In the slower sub-surface flow, transport is reduced, but the transport of dissolved materials is enhanced. In ground water flow no particle transport occurs and little dissolved materials are transported (Allan 1995). Given the lipophilicity of the OCs and their affinity to organic carbon and particles, it is likely that the overland flow with concomitant particles will transport greater amounts of the OCs deposited by rain, than the sub-surface flow and groundwater flow. Water reaching the streams via the two latter pathways has been filtered through the top soil containing , other organic matter and root structure to which OCs are likely to adsorb. In streams and rivers, the structure, and transport of nutrients, carbon and chemicals are governed by the uni-directional flow of water. The food web is almost exclusively benthic, with organisms adapted to the fast current, and pelagic food webs only play a role in larger, slow ru nning rivers (Allan 1995). The food webs derive much of their energy from heterotrophic pathways, with as a carbon source, especially near the stream source where the current is high and the streams are shaded by the riparian vegetation dominated by forest The dominating or primary producers in the

9 a

b TTOTSlf s ^ s 7r tt

Figure 3. Nutrient or OC spiralling in a stream ecosystem, (a) The three components of spiralling length, S. The nutrient atom or OC molecule travels an average longitudinal distance, Sw, dissolved in the water compartment, W, plus an average distance, SP, in the periphyton compartment, P, plus an average distance, Sc, in the compartment, C, before returning to the water. (Redrawn from Newbold et al. 1981). (b) Schematic drawing showing the effect of periphyton biomass on the spiralling of OCs in streams. The hypothetical OC partitions into periphyton in proportion to the amount of periphyton per unit area (wide vertical arrows); OC spiralling length (S) is proportional to periphyton biomass. (Redrawn from Stewart et al. 1993).

streams are periphyton, attached to the streambed. The transport and cycling of chemicals in the stream is also governed by the uni-directional flow. The cycling of nutrients and carbon in streams has been described with the "spiralling" hypothesis (Newbold et al. 1982,1983). The spiralling hypothesis predicts that the probability of uptake of a chemical in biota increases with decreasing current speed and increasing retentiveness of the stream system. The retentiveness of the stream increases with increased periphyton biomass and decreased downward transport of biota. Therefore, the probability of uptake of OCs in stream biota can be expected to be higher in slow current, eutrophicated streams with high periphyton biomasses (Fig. 3). The concentrations of OCs in stream food webs, like lake food webs, have been shown to increase with increasing (Galassi et al. 1994; Hill and Napolitano 1997). The ecosystem structure, number of species, and trophic levels

10 RAIN AND AIR PARTICLES AIR

rain air particle air/water exchange dissolved deposition

dry wet' 1

water . „ water r ' r ' ' outflow 63- suspended suspended water COLUMN sediment sediment outflow Inflow water particles

sediment/water sediment <5 sediment diffusion resuspension deposition r; SEDiMEm-d t

sediment burial

Figure 4. Processes of input and output of OCs in lakes (Redrawn from Mackay 1991). may differ greatly between streams and rivers of diffemt size, order and trophic status (Power etal. 1988; Eklov etal. 1998). Thus, several ecological processes, like nutrient status and food web structure influence the accumulation of OCs in stream organisms. The sedimentation of particles in streams and rivers is lower than in lakes, and often seasonal. The streambed sediments that build up during periods of low flow are resuspended and transported downstream during periods of high flow. Thus, OCs that are not associated to non-moving organisms and plants will be transported with the water and sediment further downstream. The major part of the annual stream flux of OCs has been shown to coincide with the short periods of high water discharges (Bremle and Larsson 1997). The input of OCs from rivers or tributaries to lakes can be of considerable magnitude. Mackay (1989) estimated the annual riverine input of PCBs to Lake Ontario to be 74% of the total input to the lake. The other dominating input of OCs to lakes was atmospheric deposition. The major loss routes of OCs from lakes are sedimentation, volatilisation and riverine output (Fig. 4). The magnitude and relative importance of these fluxes vary greatly between lakes depending on morphometry, like lake depth and volume, and ecological variables. Ecological variables greatly influence both the volatilisation and sedimentation of OCs. The most important ecological variable seems to be the lake biomass, and perhaps in particular the plankton biomass. Volatilisation is influenced by the concentration of suspended particles in the water, as only OCs in the freely dissolved phase are subjected to

11 volatilisation (Millard et al. 1983; Baker and Eisenreich 1990; Larsson et al. 1990). Due to the partitioning of OCs between particles, dissolved organic carbon, and water, high concentration of suspended particles or plankton biomass may reduce the pool of OCs available for volatilisation. Settling particles can purge the water column of OCs associated with the particles (Baker et al. 1991). The settling rates have been shown to be a function of lake trophic status (Cobelas et al. 1993; Tartar! and Biasci 1997). Lakes with high plankton biomasses have high sedimentation rates and, thus, a high withdrawal of OCs from the water column that may ultimately result in incorporation into the sediment. Ecological variables not only influence the OC load and fluxes to, from, and within the lake, but also the OC concentrations in different organisms. OCs have been suggested to biomagnify in food chains (Oliver and Niimi 1988; Evans et al. 1991; Kiriluk et al. 1995; Russel et al. 1995; Kidd et al. 1998a). Due to the more efficient transfers of OCs relative to the transfer of energy (carbon), between trophic levels, the OC concentrations increases with increasing trophic level. Therefore, the number of trophic levels in a lake may influence OC concentrations in the top predators (Rasmussen et al. 1990). Species composition and intra-specific physiological differences may also influence OC concentrations in aquatic organisms. The most important physiological variable seems to be lipid content, which may differ between, and within species affecting OC uptake (Larsson et al. 1996; Kucklick and Baker 1998). Also, size, age and growth rate of individuals in populations are of importance (Bergner 1985; Larsson et al. 1992; Sijm et al. 1992; Madenjian et al. 1993; Stow and Carpenter 1994), all of which can vary greatly between different types of lake ecosystems. The amount of OCs leaving the lake with river outflow is also very much dependent of ecologically governed lake processes and fluxes. If sedimentation and/or volatilisation are high, the river outgoing mass of OCs may be less than the river ingoing mass. The lake may, thus, act as a sink of OCs in the water system, reducing or retarding the amount of OCs reaching the and oceans. However, lake sediments can also, under certain circumstances act as sources rather than sinks of OCs to the water (Larsson 1985,1986; Zeng et al. 1999), adding to the masses of OCs transported out to the oceans by rivers. Interestingly, OCs and nutrients are not only transported downstream in rivers to the oceans, but have also been shown to be transported upstream from the oceans, contaminating lakes and areas far from the source of . Using migrating salmon as vectors, OCs and nutrients are incorporated into the lake ecosystems via the roe and salmon carcasses after spawning, increasing concentrations and altering composition in the entire ecosystem in pristine lake areas (Kline et al. 1990,1993; Bilby et al. 1996; Ewald et al. 1998). The pollutants in the salmons are acquired out in the open oceans via food or water.

12 Therefore, the matrix differs both in quantity and composition from the matrix in remote freshwaters that do not hold migrating salmon populations. In our studies we have used PCBs and DDTs as model compounds representing the lipophilic persistent contaminants. These two, well known compound classes, do not only pose a serious environmental problem themselves but, because of their worldwide distribution and high background levels, they are ideal model compounds for other chemicals with similar physical chemical properties. Thus, using PCBs and DDTs, we need not add new tracer compounds to the systems when determining the fate and transport of similar pollutants of more limited use and distribution. Also, because of their recalcitrance and accumulation potential they are suitable tracers when studying processes and fluxes in ecosystem analyses.

plankton amphipods alewives salmonids mysids sculping smelt Figure 5. PCS concentrations in different organisms in Lake Ontario. Trophic level increases from left to right. (Redrawn from Oliver and Niimi 1988).

As described, several ecological processes influence both the transport and loading of OCs to the aquatic ecosystems, as well as the accumulation of OCs in the organisms in the aquatic food webs. The papers in this thesis are mainly focused on two of the ecological processes influencing OC concentration in aquatic biota; food chain bioaccumulation, and trophic status of the aquatic system. Because of their lipophilic properties, the OCs bioaccumulate in organism lipid tissue. Top predators have been observed to accumulate greater concentrations than their prey, which have led to the theory of "biomagnification", with increasing concentrations with increasing trophic level (Oliver and Niimi 1988; Rasmussen et al. 1990; Evans et al. 1991; Kiriluk et al. 1995; Russel et al. 1995; Kidd et al. 1998a, Fig. 5). This process have led to public concern for the conservation of top

13 oligotrophic

eutrophic

total phosphorus chlorophyll a plankton biomass

Figure 6. Principal relationship between lake trophic status, and OC concentrations in lake biota. The figure is based on the results from Larsson et al. (1993), investigating PCS concentrations in pike (Esox lucius) against total phosphorus and chlorophyll a, and Taylor et al. (1991), investigating concentrations in net plankton against net plankton biomass. predatory wild life like birds of prey and seals, and for the public health hazards consuming like salmonids. Recent studies have observed negative relationships between OC accumulation in aquatic organisms and the trophic status of lakes (Taylor et al. 1991; Larsson et al. 1992). Plankton and pelagic fish in oligotrophic lakes were found to contain higher concentrations than in eutrophic lakes (Fig. 6). This relationship may cause intricate problems when attempting to remedy the problem of . When decreasing the input of nutrients to the aquatic systems an undesired side effect may be that the concentrations of OCs in the aquatic organisms increase. The aims of the papers in this thesis were, therefore, to investigate (i) the theory of biomagnification of OCs in aquatic food webs (Paper I, IT), and (ii) the influence of trophic status on OC accumulation in aquatic food webs (Paper m, IV, V).

OC accumulation in aquatic organisms Aquatic organisms accumulate OCs by passive diffusion via partitioning with water, or via food through gastrointestinal uptake (Clark and Mackay 1991; Gobas 1993a, 1993b). The relative importance of these two pathways has been, and still is, a moot topic. In natural , OCs partition between water, dissolved organic carbon (DOC), and particulate

14 organic carbon (POC). Only the OCs in the truly dissolved state are available for partitioning between water and aquatic organisms. The equilibrium-partitioning model predicts that OCs will, driven by the concentration gradient, partition between compartments until equilibrium is reached. Fugacity, a thermodynamic quantity related to the chemical potential or activity, characterising the tendency of a substance to "escape” from one compartment to another has been used to predict the equilibrium partitioning behaviour of OCs (Mackay 1979,1989; Mackay and Wania 1995). The accumulation via food has been suggested to be of greater importance than that via water for aquatic consumers (Madenjian et al. 1993). When consuming OC contaminated food, the primary consumers, like , reach equilibrium concentrations in shorter time than when only being exposed to OCs dissolved in water (Wyman and O'Connors Jr 1980). Further, aquatic consumers have been shown to accumulate OCs to concentrations higher than predicted from equilibrium partitioning with water only (Connolly and Pedersen 1988). The accumulation of OCs in aquatic organisms is affected by several, often related factors. Chemical factors include the solubility, lipophilicity, and molecular configuration of the compound. Ecological factors include lipid content and composition, metabolism and excretion, age, size, growth rate, and food choice. The accumulation of OCs in aquatic organisms is quantified using the bioaccumulation factor (BAF), which is the observed ratio of organism concentration (C,) to the concentration in water (Cw) expressed in equivalent units. BAF = C/CW The factor (BCF) is the ratio at true equilibrium, and often considered directly proportional to the octanol-water-partitioning coefficient (K0VJ) of the compound; BCF =C/CW~ Kow

Lipid content and composition The accumulation of OCs in aquatic biota is dependent on the lipid content in the organisms (Harding et al. 1981; Shaw and Connell 1982; Harding 1986; Barron 1990; Stange and Swackhamer 1994). The OCs are extremely lipophilic compounds and strongly associate with the lipids of organisms. The proportional relationship between lipid-normalised BAF and the Km of the OC has been demonstrated in several studies (Mackay 1982;Chiou 1985; Swackhamer and Skoglund 1993). Due to the lipophilicity of OCs, lipid content explains a major part of the variability in wet or dry weight normalised PCB concentrations in aquatic organisms (Swackhamer and Skoglund 1993; Kucklick and Baker 1998; Larsson et al 1996; Paper I, n,).

15 The lipid composition or quality in aquatic organisms may also affect the accumulation of OCs (Swackhamer and Hites 1988; Bierman 1990; Ewald and Larsson 1994; Bremle and Ewald 1995). Ewald and Larsson (1994) demonstrated that triglycerides accumulated more PCB than phospholipids. Lipid composition generally differs between trophic levels (phytoplankton - zooplankton - fish) and also within trophic levels between species (salmon (Salmo salar) - pike (Esox luciusj) (Ewald and Larsson 1994; Stange and Swackhamer 1994; Napolitano et al. 1996). In studies comparing OC accumulation at different trophic levels, differences in lipid composition between the organisms should, thus, be considered when interpreting the results.

Elimination The bioaccumulation of a chemical is dictated by the relationship between the rate at which the chemical is taken up by the organism and the rate at which is eliminated (LeBlanc 1995). When uptake is from water only, an exponential model can describe the concentration in an organism: dC/dt = ktCv - kjCi where Ci is the concentration in the organism, Cw is the concentration in water, k, the uptake rate, and k} the elimination rate. Because of the sequestration of lipophilic OCs, within the body, in compartments distant from the site of elimination, and the reduced ratio of elimination sites: body mass, elimination rates for OCs would be expected to decrease with increasing body mass (Brown et al. 1982, LeBlanc 1995). Studies have shown that the depuration of OCs decrease with increasing fish or plankton size and increasing lipophilicity (Km) of the compound (LeBlanc 1995; Sijm and van der Linde 1995; Fisk et al. 1998).

Size, age, and growth rate The size and age of organisms generally increase with increasing trophic level (Sprules and Bowerman 1988). Several studies show a positive relationship between size or age (Madenjian et al. 1993; Stow and Carpenter 1994) and OC concentrations in aquatic organisms. Therefore, the increasing size or age of organisms with increasing trophic level may explain the concomitant increase of OC concentrations, independent of food chain effects on accumulation. The time required to reach a certain size may also influence OC concentrations. The growth dilution hypothesis predicts a dilution of OCs in fast growing aquatic organisms (Borgmann and Whittle 1992; Sijm et al. 1992; Stow and Carpenter 1994). Adding the effect of growth rate in the uptake equation gives:

16 bcf0>baf ,>baf 2

Figure 7. Principal relationship between OC concentrationin water (C„) and OC concentration in biota (C|) with different growth rates (G). Increasing growth rates (G0 < G, < Gj results in decreasing C; at a given C„, hence; BCF0 > BAF, > BAF2.

dC/dt=klCvl-(kJCl+Gl) where G, is the net growth rate of the organism. When large amounts of energy from the food intake is allocated to new tissue development, the OCs will be diluted in the larger body mass, compared to when much energy is allocated to metabolism, respiration, food search, predator avoidance, or migration. Consequently, as e. g. juvenile organisms have a higher growth rate compared to adults the concentrations of OCs may be lower due to growth dilution (Fig. 7).

Biomagnification of OCs The concept of biomagnification of OCs has been investigated since high concentrations of DDT were observed in top predators, five decades ago. Aquatic top predators were identified as especially vulnerable for biomagnification of OCs (Rudd 1964). This concentrative process was attributed to the presumed efficient transfer of OCs from one trophic level to the next, in combination with the reduction in biomass associated with each progressive trophic level, due to losses in the biomass conversion between trophic levels (Rudd 1964). This would yield greater concentrations (but not total amounts) of OCs associated with the next trophic level, than that present in the previous trophic level. There are indications of an efficient trophic transfer of OCs, 50-80% may be transferred between trophic levels (Jackson and Schindler 1996; Madenjian etal. 1998).

17 Food chain accumulation has been defined as the increase in the organism chemical concentration exceeding that that could be expected from exposure to the chemical in the water phase only (Thomann 1989). An equation for the uptake of a chemical from the water and food for the zth trophic level (z = 2, 3,4...n) is given by: dC/dt = + aM., yu_i CM - (kjCj+G^) where a is the chemical assimilation efficiency, and y the specific consumption rate. Generally, if organism BAFs are above that predicted from Kow, the chemical is considered to have bioaccumulated via food (Oliver and Niimi 1988; Haffner et al. 1994). Also, a biomagnification factor (BMF, concentration in predator/concentration in prey) ratio greater than one is considered indicative of food chain effects (Clark and Mackay 1991; Russell et al. 1995). Examining OC concentrations at different trophic levels in aquatic food chains has tested the biomagnification theory. These studies have found increasing OC concentrations with increasing trophic level, mainly in fish (Oliver and Niimi 1988; Evans et al. 1991; Russell et al. 1995), aquatic mammals (Muir et al. 1988), and birds of prey (Focardi et al. 1988). Rasmussen et al. (1990) concluded that concentrations of PCB in top predatory fish from different lakes increased with increasing length of food chains, a result attributed to biomagnification. The concept of biomagnification in aquatic food chains have recently been questioned (LeBlanc 1995). Some studies have failed to find increasing OC concentrations in food chains (Hope et al. 1998; Kucklick and Baker 1998), in particular for planktonic food chains (Harding et al. 1997). Also, increasing concentrations of OCs with trophic level have been attributed to differences in lipid content (Bentzen et al. 1996; Hope et al. 1998; Kucklick and Baker 1998), depuration rates (LeBlanc 1995; Sijm and Van der Linde 1995), size (Bergner 1985), and exposure duration (Harding et al. 1997), rather than to biomagnification. LeBlanc (1995) argued that the apparent increase of lipid content with trophic levels would cause an increase of lipophilic OCs by bioconcentration alone driven by passive diffusion over a fugacity gradient. In Paper I we showed that PCBs were not biomagnified in pelagic, planktonic food chains with 3 trophic levels (Fig. 8). Lipid normalised PCB concentrations were lower in zooplankton than in both phytoplankton and fish, which did not differ in concentrations. We conclude that the concept of biomagnification did not apply to the lower trophic levels in the investigated lakes. Other factors than food are likely to govern the accumulation of OCs in these organisms. Lipid content was of major importance, explaining 30 to 60 % of the variation in PCB concentrations, and was the investigated variable with the highest degree of explanation. Also, we suggest a difference in elimination of PCBs between the

18 9.0

6.0 ------■------■------■------phytoplankton zooplankton fish Trophic level

Figure 8. Bioaccumulation factors (BAFLW) for three trophic levels; phytoplankton, zooplankton and fish, in 19 Swedish lakes. Lines connect the BAFs for each lake. (Paper I). three trophic levels. The elimination of high molecular weight PCBs decreases with increasing organism size (Paper I).

Stable isotopes and food chain effects Stable isotopes are useful measures in describing the feeding relationships in a food web and allow the identification of multiple food sources consumed by an organism (Peterson and Fry 1987; Fry 1991; Kling et al. 1992). The enrichment of heavy isotopes of nitrogen (15N) and carbon (13C) relative to the light isotopes (14N and 12C) or the fractionation of these isotopes through the food web has been well-documented (Minegawa and Wada 1984; Peterson and Fry 1987; Owens 1987). Ratios of stable isotopes of nitrogen (3!5N) and carbon (d 13C) can be applied to verify current understanding of trophic interactions in an aquatic food web based on traditional methods, and can be used to describe the distinctive food web pathways leading up to the top trophic level. 31SN have successfully been used to determine trophic position (Fig. 9), while d 13C, which also increases with trophic level, perhaps is most useful in determining the carbon source of the food web. 315N has been used to examine biomagnification of OCs in food webs. These studies find a correlation between 315N and OC concentrations, which could indicate biomagnification of OCs in aquatic food chains (Broman et al. 1992; Kiriluk et al. 1995; VanderZanden and Rasmussen 1996; Kidd et al. 1998a). Concentrations of OCs in aquatic species are significantly related to their 31SN both for freshwater and marine food webs (Broman et al. 1992; Kiriliuk et al. 1995; Kidd et al. 1995; Kucklick et al. 1996; VanderZanden et al. 3 Predators

Omnivores £ 2 i. £ H Algae 1 Terrestrial plants

-6 -4 -2 0 2 4 6 8 10

disN (%c)

Figure 9. d 15N values of biota from different trophic levels in lakes. (Redrawn from Fry 1991).

1997; Kidd et al. 1998a). The relationship between d 15N and OC concentration is compromised by a commonly found positive relationship between d lsN and lipid content (Kiriluk et al. 1995; Bentzen et al. 1996; Kucklick and Baker 1998). However, many studies show that even when lipid normalised, food web structure and trophic level (Haffher et al. 1994; Kiriluk et al. 1995; Bentzen et al. 1996; Kidd et al. 1998b) may affect OC concentrations. In studies examining food web bioaccumulation, most relationships between d 15N and OCs are based on means or comparing different populations and species (Broman et al. 1992; Kiriluk et al. 1995; Kidd et al. 1995; Kucklick et al. 1996; Kidd et al. 1998a, 1998b). These organisms can differ greatly in inter-specific life history parameters other than diet, strongly influencing OC uptake, such as age, size, and lipid composition, thereby complicating interpretation of the relationships between trophic position and OC concentration. In Paper II we demonstrated that the lipid content explained most of the variation of OC concentrations in a salmon population. When divided into reproductive strategies d 15N increased the explanation, but still, the lipid content had the main influence on salmon OC accumulation, and the influence of food chain bioaccumulation and omnivory seemed to play a minor part. In the entire population, d 15N and OC concentrations were positively related, but the relationship was weak (Fig. 10). The relationship between lipid content and OC concentration was stronger. To our knowledge, few studies have been able to find any relationship between trophic position and OC uptake in a single population of an omnivorous fish species, even though the d 15N have varied from 3%o to 6%o, equivalent to one or two trophic levels (Kiriliuk et al. 1995; Kidd et al. 1998a; 1998b). In the investigated fish populations, size, age, and especially lipid content have instead explained the variation in OC concentrations.

20 10000

1000 p* w 100: o

10 11.5 12 12.5 13 13.5 14

Figure 10. Relationship between ]£PCB concentrationsand d15N in Atlantic salmon (Salmo solar) from the Baltic Sea. (Paper II)

Trophic status and OC accumulation: lakes Recent studies in temperate regions have demonstrated a negative relationship between lake trophy and OCs (Taylor et al. 1991; Larsson et al. 1992). The concentrations of OCs like PCB and DDT in aquatic organisms decreased with the trophic status of the lakes, measured as total phosphorus concentrations, or plankton biomass. In these studies, the external input of OCs to the lakes was assumed similar, as it was dominated by atmospheric deposition (Bidleman 1988; Larsson and Okla 1989). Hence, the observed differences in OC concentrations are likely a result of internal lake processes and ecological variables, or watershed influence. In paper m and IV, we wanted to determine which lake processes differ between oligotrophic and eutrophic lakes to the extent that it may influence the accumulation of OCs, producing the observed negative relationship.

Biomass The effect of eutrophication on lake ecosystems, through excessive input of nutrients, is primarily an increase of the overall biomass of plankton communities, especially that of the phytoplankton (Pace 1986; Mazumder et al. 1988). The recorded lower OC concentrations in organisms of eutrophic systems have been attributed to the increased biomass of these systems (Taylor et al. 1991; Richer and Peters, 1993). Taylor et al. (1991) suggested a ’’biomass dilution ” responsible for the negative relationship. Lakes of high trophic status have high plankton biomasses to which the lipophilic OCs are partitioned. This partitioning would reduce the OC concentrations in the water, and thereby

21 the bioavailability, giving a lower concentration per weight unit of plankton. may also occur under rapid growth conditions when the growth rate of phytoplankton exceeds the partitioning rate of OCs between water and phytoplankton, which prevents equilibrium to be reached (Swackhamer and Skoglund 1993). In Paper III we found no relationship between the lipid normalised PCB concentrations in planktonic food webs and the plankton biomass in lakes. Hence, we could find no support for the ’’biomass dilution ” theory. Instead, lipid content explained most of the variation in PCB concentrations in biota regardless of lake trophy. Due to the lipophilicity of OCs, lipid content and, consequently, the accumulation capacity explains a major part of the variability in wet or dry weight normalised PCB concentrations in aquatic organisms (Swackhamer and Skoglund 1993, Kucklick and Baker 1998, Larsson et al. 1996). The lipid content of both phyto-, and zooplankton can vary considerably, and seems to be related to the nutrient regime of the aquatic system. Algal culture experiments and field studies have recorded an increased phytoplankton lipid content under nutrient stress (Shiffin and Chisholm 1981; Groeger and Kimmel 1988; Parrish and Wangersky 1990; Reitan et al. 1994; Arts et al. 1997). This lipid accumulation is considered being partially a result of a steady lipid synthesis combined with reduced cell division rate and protein synthesis caused by the reduced availability of nutrients (Siron et al. 1989; Sukenik and Livne 1991). The species composition of both the phytoplankton and the zooplankton community differs between eutrophic and oligotrophic lakes. In eutrophic lakes, blue green algae dominates the phytoplankton community, while the species composition in oligotrophic lakes is more diverse, composed of several different species (Wetzel 1983). The zooplankton composition also differs, with large cladocerans and calanoids common in oligotrophic lakes, and smaller zooplankton species, cyclopoids and rotifers dominating in eutrophic, due to differences in fish pressure, and the microbial community (Watson and Kalff 1981; Beaver and Crisman 1982; Pace 1982; Bays and Crisman 1983). We found a negative relationship between the lipid content of phytoplankton and the trophic status of the lakes (Paper Id, Fig. 11). This could be a result of different phytoplankton community compositions with lipid-rich species more prominent in oligotrophic lakes (Larsson et al. 1998), and/or intra-specific changes in lipid content or synthesis in phytoplankton, due to nutrient stress. The phytoplankton summer composition in the different lakes indicated that the first might be the case. In the more eutrophic lakes, the phytoplankton communities were dominated by blue-green algae, low in lipid content, whereas the communities of the oligotrophic lakes were more diverse, composed of several other taxons, all more lipid-rich than blue-green algae. However, we could not exclude intra-specific differences in phytoplankton lipid content as the phytoplankton sample from

22 one oligotrophic lake, totally dominated by blue-green algae, had higher lipid content than phytoplankton samples from the eutrophic lakes, also dominated by blue-green algae. Further, a negative relationship between phytoplankton lipid content and lake trophy was also found in the spring sampling of the same lakes (Paper IV). In spring, the phytoplankton communities in all lakes were dominated by diatoms, and in several lakes by the same species. Hence, we conclude that both inter-, and intra-specific differences in phytoplankton lipid content, caused by different nutrient regimes, influences the OC accumulation.

3 0.01

0.001 1000

total phosphorus (pg/L) Figure 11. Relationship between phytoplankton lipid content and lake trophy, measured as Tot-P, in 19 Swedish lakes. (Paper HI).

Sedimentation Due to the hydrophobic nature of OCs like PCBs they associate with particles in the water, and the sorption is mainly governed by the hydrophobicity of the compounds and the lipid or organic carbon content of the particles (Nau-Ritter et al. 1982; Eisenreich 1987; Swackhamer and Skoglund 1993). The main removal process of PCBs from waterbodies is, therefore, considered to be sorption to settling particles and deposition to the sediments, and burial in aquatic sediments have been suggested as the final sink for OCs (Eisenreich 1987, Baker et al. 1991). However, under certain conditions sediments have been shown to act as sources rather than sinks of OCs to the water and atmosphere (Larsson 1985, 1986; Zeng et al. 1999). The sedimentation rates of autochthonous particles (dead algae) are greater in eutrophic lakes than in oligotrophic lakes (Cobelas et al. 1993; Liukkonen et al. 1993; Weyhenmeyer et al. 1995; Tartar! and Biasci 1997). In eutrophic lakes following algal blooms, the sedimentation of dead algae may purge the water column of the particle reactive OCs. This has led to an alternative hypothesis explaining the relationship between OCs and

23 lake trophy. It is suggested that the increased sedimentation of organic matter in eutrophic lakes, with concomitant purging of the water column from OCs will enhance sediment contamination (Pavoni et al. 1990; Millard et al. 1993), and reduce water concentrations. Due to the great amounts of settling algae, the microbial degradation in eutrophic lakes is insufficient to mineralise the entire pool of settling organic carbon (Liukkonen etal. 1993). Hence, the lipophilic OCs associated with the settling algae may not be released back to the

Oligotrophia Eutrophic

rl8.5%

■ fish Q phytoplankton □ macrozooplankton □ water □ microzooplankton B surface sediment Figure 12. Pie charts showing the mean within lake distribution of XPCB between compartments in six oligotrophic and five eutrophic lakes in Sweden. (Paper IV). water following the crash of an . Instead, the OCs will accumulate in the sediment. In Paper IV we investigated the sediment accumulation of PCBs in lakes in a trophic gradient. The mass of PCB was positively related to lake trophy, both in the surface sediment and in the deeper sediment. The within lake distribution of PCBs revealed that approximately 10% more of the total PCB amounts were associated with the surface sediments in the eutrophic lakes than in the oligotrophic lakes (Fig. 12). Also, more PCBs were buried in the sediment in eutrophic lakes than in oligotrophic lakes. We concluded that the sediment accumulation and burial of PCBs were greater in more eutrophic lakes, likely due to greater sedimentation of dead phytoplankton to which the PCBs were associated.

24 Volatilisation Jeremiason et al. (1994) concluded that in Lake Superior, a large oligotrophic lake, the dominant loss process of PCBs was volatilisation rather than sedimentation. In a constructed mass balance, the annual net volatilisation of PCBs from Lake Superior was estimated to be 1900 kg, while the loss due to burial in the sediment was 110 kg. Only PCBs in the freely dissolved phase are subjected to volatilisation from the water surface to the atmosphere (Millard et al. 1983; Baker and Eisenreich 1990; Larsson et al. 1990). The size of the freely dissolved pool of PCBs is governed by the concentrations of suspended particles and dissolved organic carbon in the water. Thus, the relative importance of volatilisation can be expected to be lower in eutrophic lakes, with high biomasses, than in oligotrophic, with lower biomasses. In paper IV we found that the total amount of PCBs were greater in the eutrophic lakes than in the oligotrophic lakes, even though atmospheric deposition dominated the input of OCs and the lakes, thus, should receive similar amounts. However, as expected, given the differences in the concentrations of suspended particles, the pool of freely dissolved PCBs was greater in the oligotrophic than in the eutrophic lakes of our study (Fig. 12). Hence, we expect the losses of PCB by volatilisation, over the time scale of decades, from the time of PCB introduction, to be greater in the oligotrophic lakes compared to the eutrophic, which could in part explain the differences in total PCB amounts between the two lake types. Also, the withdrawal of particle associated PCB from the water column via sedimentation and burial may increase total amounts in eutrophic lakes, as it will lower the concentrations of PCB in the water column. This may enable more PCB to be partitioned to the water from air, via air-water gas exchange. The relative amount of freely dissolved ZPCB was greater in oligotrophic lakes, but not the absolute amount. Since we found no difference between relative amounts in biota this difference must be linked to the differences in the sediment accumulation and recycling of PCB in oligotrophic and eutrophic lakes. In oligotrophic lakes, PCB is recycled from settling particles in the water column back to the freely dissolved phase. Only freely dissolved PCBs are available for uptake in organisms via equilibrium partitioning (Schrap and Opperhuizen 1990). The larger fraction of freely dissolved PCBs in the epilimnion of oligotrophic lakes may therefore increase the potential for these compounds to be taken up by the pelagic biota via equilibrium partitioning.

Recycling Recently, it has been demonstrated that internal lake processes are of great importance and magnitude in redistributing PCBs between water, biota and sediments in lakes (Gevao et al.

25 1997). There appears to be an intense recycling of organic carbon and PCBs at or immediately above the sediment-water interface. The annual recycling of PCBs have been shown to exceed both the atmospheric input and the sediment accumulation and burial (Baker et al. 1991; Sanders et al. 1996; Larsson et al. 1998). However, the site of mineralisation may differ between oligotrophic and eutrophic lakes. In oligotrophic lakes, most of the mineralisation of organic carbon takes place in the water column (Baker et al. 1991). Of the organic matter that reaches the lake floor, 80-90% is degraded just above the sediment-water interface. Presumably more labile organic compounds are preferentially degraded, altering the organic chemical composition of particles as they settle through the water column (Baker et al. 1991). Due to the higher sedimentation rates, the shallower depth and, hence, the shorter particle settling times (Sanders et al. 1996), most of the mineralisation in eutrophic lakes take place just at the sediment surface. Assuming similar morphometry, the turnover of organic carbon in oligotrophic lakes is generally more efficient than in eutrophic (Wetzel 1983). hi oligotrophic lakes much of the mineralisation of labile organic carbon takes place in the water column, and sediment accumulation rates are lower than in eutrophic lakes (Wetzel 1983). Thus, the OCs associated with settling particles are recycled back to the water more efficiently in oligotrophic lakes. In the water, the freely dissolved OCs will again partition between water, suspended solids, and dissolved organic carbon. In eutrophic lakes a greater fraction of OCs will be buried in the sediment, and withdrawn from the water column and pelagic organisms.

Stratification hi the temperate regions, lakes with limited surface area, and sufficient depth are dimictic, i. e., the water circulate freely twice a year in the spring and autumn and undergo thermal stratification in summer and in winter (Wetzel 1983). Considerable amounts of the water column PCBs are withdrawn from the epilimnion during summer stratification presumably due to loss by sedimentation (Baker et al. 1985; Jeremiason et al. 1998). However, all of the PCBs are not incorporated into the sediment but rather redistributed in the lake when the stratification is broken and the lake water undergo mixing at the autumn overturn. The formation of a thermocline also prevents resuspended sediment from reaching the epilimnion in the profundal regions (Weyhenmeyer 1996). Thus, the epilimnion of stratified lakes does not receive any addition of PCB sediment recycling. In shallow, eutrophic, well mixed lakes, plankton compete with great amounts of resuspended sediment for the partitioning of PCBs (Weyhenmeyer et al. 1995). In deep, oligotrophic, stratified lakes resuspended particles remain in the hypolimnion (Weyhenmeyer 1996).

26 Therefore, the partitioning ofPCBs during stratification is only occurring between living plankton and water (and DOC) in the epilimnion. In paper IV we demonstrated that the total £PCB amount in the epilimnion decreased from April to August in a majority of the stratified lakes. These results support the hypothesis of a seasonal PCB sink caused by the settling of dead algae with associated PCBs from the epilimnion towards the lake floor. The PCBs released to the water as a result of the mineralisation of the labile fractions of organic carbon is less in eutrophic lakes. A fraction may partition within the sediment to particles and interstitial pore water. An incomplete mineralisation in the water column due to higher sedimentation rates following algal blooms, and the generally shallower depths in eutrophic lakes may, therefore, attribute to the greater PCB amounts found in the sediments. As the sediment accumulation and burial were greater in eutrophic lakes than in oligotrophic, this mechanism may explain the negative relationship between OC concentrations in biota and lake trophy (Taylor et al. 1991; Larsson et al. 1993).

Trophic status and OC accumulation; streams In Paper V we demonstrated a positive relationship between total phosphorus concentrations in streams and OC concentrations in brown trout (Salmo trutta) (Fig. 13). These results are contrary to the negative relationship between trophic status and OCs in biota suggested for pelagic, lake ecosystems. Lotic systems differ from lentic systems in several respects, which may to differences in uptake of OCs in biota in relation to trophic status. The uni-directional flow of water governs the processes and constrains on organisms in lotic systems. Stream food webs are almost exclusively benthic, with the pelagic compartment playing only a minor role in energy and chemical fluxes (Allan 1995). Furthermore, lotic food webs derive much of their energy from terrestrial organic matter (detritus). These energy pathways are referred to as heterotrophic and the consumers of detritus are and detrivores, in contrast to autotrophic pathways linked to higher trophic levels by herbivores, which dominate in lakes (Allan 1995).

Spiralling and biomass The most obvious feature of streams is the uni-directional flow, which affects the biotic and abiotic environment. The cycling of nutrients in streams has been described by the "spiralling" hypothesis (Newbold et al. 1983). This hypothesis is concerned with the joint processes of cycling and transport, where the spiralling length is dependent on rate of cycling and retentiveness. The circular, cycling behaviour of chemicals in lake systems is replaced by a helix shaped, spiralling behaviour in lotic systems due to the uni-directional

27 flow (Fig. 3). If spiralling length is short, then a nutrient atom cycles through the biota a greater number of times as it travels the length of the stream, than it would if spiralling length was long. Therefore, the probability of uptake in biota of a chemical increase with decreased spiralling length (Newbold et al. 1983).

1 00 OH------______i'---- 'i 'i 'i '» iiii"------______t'---- i' 'i i i It N,i i : r = 0.71 j

1 100E ' i # .= • * %.» h .."4• *« •* . : «•••:t

1 I------i1------—1i----—1i iI iI 1I 1I 1I 1I------1—1------i—1----i1 i Mili 111 l- 10 100 1000 Tot-P (ng/L)

Figure 13. Scattergram showing the correlation between Tot-P in water and concentrations of XDDT in O h brown trout (Salmo trutta) from 52 stream sites in southern Sweden. (Paper V).

The most prominent effect of eutrophication in lakes is the increased phytoplankton biomass and . The effects of excessive nutrient load in streams and rivers are not as clear as in lakes. Because of the continuous flow of water, nutrients may not exert a primary limitation on biomass, instead, light and current conditions may, to a greater extent than in lakes, limit (Moore 1977). However, several investigations have shown increased periphyton and microbial biomass and productivity as a response to increased phosphorus loads (Elwood et al. 1981; Meyer and Johnson 1983; Perrin et al. 1987; Bothwell 1989). The formation of dense algal mats may cause P-limitation in cells within the mat, while those closer to the surface of the periphyton matrix remain P-replete (Elwood etal. 1981; Bothwell 1989). Periphyton provides a large total surface area for adsorption and uptake of lipophilic, organic pollutants. Therefore, it potentially plays a very significant role in retarding the net movement of OCs downstream. It has been proposed that the movement of contaminants downstream should be analogous to the movement of phosphorus that is taken up actively or adsorbed to periphyton (Stewart et al. 1993). Consequently, according to the "spiralling" theory, the spiralling length should be negatively related to periphyton biomass. If a positive relationship exists between phosphorus and periphyton biomass, the

28 reduction of spiralling length would increase the probability for the pollutant to be taken up by biota in streams with high phosphorus concentrations. Thus, the observed positive relationship between OC concentrations in trout and phosphorus concentrations in the streams described in Paper V, could be explained by higher uptake of pollutants by trout in eutrophic streams due to shorter spiralling length.

Heterotrophy vs. autotrophy Peterson et al. (1985) showed that with increased phosphorus levels, stream ecosystems shift from heterotrophy, where microbial of allochthonous detritus is the dominant energy source, to autotrophy where autochthonous, photosynthetic, periphytic algae dominate as an energy source. The OC concentrations are higher in periphyton than in detritus of terrestrial origin, due to higher levels of pollutants in the surrounding medium (water vs. air), while the partitioning coefficient of OCs between air-plants and water- periphyton is similar (Reischl et al. 1989; Hermanson and Hites 1990; Bacci et al. 1990, Stange and Swackhamer 1994.). However, the equilibrium of OCs between air and water is partitioned towards water, and governed by the solubility and vapour pressure of the compounds. Partitioning coefficients between air-water (KAV/) for DDT and PCB are in the range 1 10"3 - 5 10"3 (Baker and Eisenreich 1990; Mackay 1991), which means that water contain 1000 to 5000 times the amount of OCs in air at equilibrium. Consequently, the OC concentrations in aquatic primary producers are higher than levels in their terrestrial counterparts. Differences in levels of OCs at the base of the food web represented by allochthonous detritus, or autotrophic periphyton are likely to persist through the transfer in the food chain, via grazing/collecting invertebrates, to trout, feeding on these invertebrates. Consequently, a higher degree of autotrophy in streams, caused by eutrophication, may lead to increased levels of OCs in top predators in the lotic food chain. In lakes, autochthonous phytoplankton is the dominating energy source in both eutrophic and oligotrophic ecosystems, except for the littoral zones where allochthonous detritus may dominate. Therefore, no obvious shift of origin in energy source is exhibited along a trophic gradient in lakes. Hence, no such shift can affect the levels of OCs in pelagic plankton and predators.

Watershed influence Other mechanisms, not related to periphyton biomass and production, may also influence the relationship between total phosphorus and OCs in streams. The relationship between nutrients and land use in the catchment area has been shown in other studies (Allan 1995).

29 OCs reach the stream mainly via rain deposition either on the stream surface directly, or in the watershed area with subsequent transport to the stream. Considering the small area of streams compared to the watershed areas, input via the watershed area dominates the OC budget of running waters, as opposed to lakes, where deposition via rain on the lake surface, may be a considerable input source to the system. PCBs and DDTs have an organic carbon to water partitioning coefficient (K^) of approximately 106 , which indicates that they are strongly associated with dissolved, or particulate organic carbon; in extension, sources of organic carbon and OCs in streams are probably the same. The major sources of carbon in streams are from leaves and needles of riparian vegetation, which, after senescence, fall into the river, and from the forest floor and soil in the watershed area (Allan 1995). Most of the water transporting organic carbon reaches the stream via groundwater or sub-surface pathways (Fig. 2). In agricultural land, however, the infiltration capacity of the soil may be exceeded during heavy rainfall or snowmelt. Agricultural and urban land, therefore, lead to increased overland flow, which may be an important source of organochlorines to running waters. Concentrations in groundwater and sub-surface flow are low, as the organochlorines are "filtered" through adsorption and microbial processing of the organic carbon with which they are associated, as the water makes its way to the stream. When transported via overland flow, the organochlorines in rain or meltwater are not "filtered" on their way to the streams, and there is a washout of particulate organic matter (POM) and dissolved organic matter (DOM) from the soil and interstitial soil water, with which organochlorines are associated (Paper V). There are similarities in the behaviour of phosphorus and OCs. The compounds are associated with organic carbon in waters, are readily taken up by biota in aquatic ecosystems, and are highly adsorptive. As mentioned, the spiralling processes of phosphorus and organic carbon (Newbold et al. 1982) may also be valid for OCs (Stewart et al. 1993; Sallenave et al. 1994). Hence, it is probable that transport from watershed areas to streams is similar for both types of compounds. Therefore, the positive relationship between the OC concentration in stream organisms and concentration of phosphorus, could be explained by a higher input of both "compounds" to running waters, caused by higher run-off via overland flow of both phosphorus and pollutants from watershed areas with a higher proportion of agricultural land.

Conclusions Several ecological processes affect the transport, distribution and uptake of OCs in aquatic systems. In this thesis I have focused on two processes, food chain accumulation and trophic status of the aquatic system. Contrary to earlier theories, PCBs did not seem to be

30 biomagnified in planktonic, freshwater food chains (Paper I). Concentrations in primary and secondary consumers were similar or lower than in primary producers, hence, the biomagnification process did not elevate the PCB concentrations above what could be predicted from partitioning with water alone. The PCBs were differentially fractionated along the food chain, the average lipophilicity increased from phytoplankton to fish. The increasing concentrations of OCs with trophic level, attributed to biomagnification, is most evident for compounds with log Km > 6 (Oliver and Niimi 1988; Russel et al. 1995). The size of aquatic organisms increases with trophic level (Sprules and Bowerman 1988). Given the fact that the elimination rates decreases with increasing organism size and increasing K0VJ of the compound (LeBlanc 1995; Sijm et al. 1995; Fisk et al. 1998), the OC accumulation patterns in aquatic food webs, attributed to biomagnification may, therefore, be caused by differences in depuration with different organism size. The lipid content is the major factor determining OC accumulation in aquatic organisms, independent of trophic level (Paper I, H, and IH). Although we found a weak relationship between trophic level, measured as d 15N, and OC concentrations, lipid content exerted the main influence on OC uptake in a specific population of omnivorous fish (Paper H). It is essential, therefore, that lipid content and quality, size, and age of the organisms is considered when comparing OC accumulation in aquatic food webs. In lake ecosystems, the major effect of eutrophication is an increased biomass of both plankton and fish (Paper IV). The increased plankton biomass results in higher particle sedimentation rates in eutrophic lakes. This higher particle sedimentation also results in higher sedimentation and burial of particle associated OCs in eutrophic lakes (Paper IV). Although there appears to be an intense mineralisation of organic carbon and recycling of OCs from settling solids and the sediment surface, this may be relatively more pronounced in oligotrophic lakes. An incomplete mineralisation in the water column due to the higher sedimentation rates following phytoplankton blooms, and the generally shallower depths in eutrophic lakes, may explain the greater amounts of OCs in the sediments. Eutrophication also causes inter-, and intraspecific differences in the phytoplankton communities of lakes. The observed negative relationship between summer dry weight PCB concentrations in phytoplankton and lake trophic status was explained by interspecific differences in lipid content of the phytoplankton (Paper IQ). The eutrophic lakes were dominated by blue-green algae, low in lipids, while in the oligotrophic lakes the phytoplankton communities were more diverse, composed by several taxons, all more lipid-rich than blue-green algae. In spring, when diatoms dominated in all lakes there were intraspecific differences in lipid content, the diatoms in oligotrophic lakes had higher lipid content than diatoms in eutrophic lakes.

31 In streams, we found a positive correlation between OC concentration in biota and trophic status (Paper V). This is the opposite of that observed in lakes, where a negative relationship has been determined. The differences found between lakes and running waters may be due to several processes, none of which are mutually exclusive. Differences may relate to how compounds cycle in lentic and lode ecosystems, e. g. cycling vs. spiralling. It may also be connected with differences in uptake between pelagic, autotrophic food webs in lakes, and benthic, heterotrophic food webs in rivers. Furthermore, the catchment area should have a greater influence on the input of pollutants to running waters than to lakes.

Future research The dynamics of OCs in streams The fate and transport of lipophilic contaminants in lode environments is a field of research of priority. Streams and rivers dominate as recipients of outiets of pollutants from anthropogenic activities, such as industrial wastewater, plants and agricultural drainage water. Thus, lode systems have a history of severe contamination and can also be expected receive large amounts of different new pollutants in the foreseeable future. To understand the influence of ecological processes on the transport and uptake of these pollutants is, therefore, of priority. The knowledge is needed not only to control exposure to stream living organisms, and the pollutant’s effects on the stream ecosystems, but also, to control and predict the amounts of pollutants reaching the estuaries and oceans. There are several features of lode environments that potentially may influence the transport and fate of lipophilic contaminants. Of special interest is the role of trophic status, e. g. periphyton biomass, on the retention and uptake of contaminants as predicted by the spiralling hypothesis. Here, it is important to determine whether or not the positive correlation between Tot-P and OCs (Paper V) is caused by an increased periphyton biomass. On an ecosystem level, the question is: does an increased periphyton biomass increase the retention of contaminants in a river stretch, and does the increased retention affect uptake in organisms? On a larger scale, how does the trophic status of a river influence the net transport of contaminants from the river watersheds to the estuaries and oceans? The lode ecosystem shift from heterotrophy to autotrophy, caused by excessive nutrient loads is also of interest, when predicting contaminant exposure for stream living organisms. The degree of heterotrophy versus autotrophy can readily be determined by stable isotope analysis, using d 13C. Thus, it would be possible to determine the influence of different carbon sources, and even different ratios of carbon sources, on the levels of pollutants in stream organisms. It has been suggested, that a fertilisation of streams and rivers would be beneficial for fish harvests, as fertilisation increases growth rates of

32 salmonids in streams. If ecological processes cause the positive relationship between OC concentrations and phosphorus, fertilisation of streams may cause contaminant concentrations to increase in fish. Parameters and processes of specific interest are: * phosphorus and periphyton biomass - is phosphorus limiting for periphyton growth? ^periphyton biomass and retention of contaminants - the spiralling theory * heterotrophy versus autotrophy - influence of contaminant levels in the primary carbon source

The dynamics of lipids in ecosystems Lipids play a key role in the fate of OCs in aquatic ecosystems. The lipid content of organisms and trophic transfer of lipids seem to govern the distribution of OCs in the ecosystem. Indeed, from an ecological perspective the production and cycling of lipids is of profound importance for the understanding of ecosystem structure and dynamics. Lipids are the major energy source for the aquatic consumers, which adapt feeding strategies in search to optimise energy intake. Also, allocation of lipids or energy within organisms differs between ecosystems and organisms. Organisms allocate lipids within the body for e. g. reproduction, migration or hibernation, depending on life-phase or season. Eutrophication seems to influence the content and composition of lipids in phytoplankton, both intra-, and interspecifically. These differences can have great importance on an ecosystem level, as the lipids or energy produced at the phytoplankton level govern the production at consumer levels above phytoplankton. Fatty acids synthesised by phytoplankton can be traced in food webs. Thus, with fatty acid profiles and stable isotopes, it is possible to quantify relative importance of different species and groups of primary producers as energy sources in the aquatic food web. Also, the breakdown of lipids by the microbial community can influence the cycling of OCs in aquatic ecosystems. The degree of mineralisation of the labile organic carbon may influence the recycling to withdrawal ratio of OCs in the sedimentary process in lakes. Thus, the influence of trophic status on the net mineralisation of organic carbon may have effect on the fluxes of OCs between the water - particle - sediment compartments in lakes. Parameters and processes of specific interest are: * nutrient stress and lipid production in phytoplankton - content and quality * transfer of lipids in the food web - selective feeding and allocation * breakdown and recycling of lipids - sedimentation and the

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42 Tack Tack till Pelle och Goran och John. Doctoral theses from the Department of Chemical Ecology and Ecotoxicology, Department of Ecology, Lund University, Sweden.

1. A nders SOdergren . Transport, distribution, and degradation of organochlorine residues in limnic ecosystems (defended at Dept of Limnology). May 23,1973. 2. GORAN Bengtsson . Ecological significance of amino acids and metal ions, a microanalydcal approach (defended at Dept of Zooecology). May 24,1982. 3. Carita Brinck . Scent marking in mustelids and bank voles, analysed of chemical compounds and their behavioural significance (defended at Dept, of Zooecology). May 17,1983. 4. A nders Tunlid . Chemical signatures in studies of bacterial communities. Highly sensitive analyses by gas chromatography and mass spectrometry. October 3, 1986. 5. A nders Thur Sn . Phtalate esters in the environment: analytical methods, occurrence, distribution and biological effects. November 4,1988. 6. Peter Sundin . Plant root exudates in interactions between plants and soil micro­ organisms. A gnotobiotic approach. March 16,1990. 7. A nders V aleur . Utilization of chromatography and mass spectrometry for the estimation of microbial dynamics. October 16,1992. 8. Hans Ek . Nitrogen acquisition, transport and metabolism in intact ecto- mycorrhizial associations studied by 15N stable isotope techniques. May 14, 1993. 9. Roland Lindquist . Dispersal of bacteria in ground water - mechanisms, kinetics and consequences for facilitated transport. December 15,1995. 10. A lmut G erhardt . Effects of metals on stream invertebrates. February 17,1995. 11. O lof Regnell . Methyl mercury in lakes: factors affecting its production and partitioning between water and sediment. April 21,1995. 12. Per Woin . Xenobiotics in aquatic ecosystems: effects at different levels of organisation. December 13,1996. 13. GORAN Ewald . Role of lipids in the fate of organochlorine compounds in aquatic ecosystems. October 18,1996. 14. Johan K nulst . Interfaces in aquatic ecosystems: Implications for transport and impact of anthropogenic compounds. December 13,1996. 15. G udrun Bremle. Polychlorinated biphenyls (PCB) in a . April 25,1997. 16. Christer Bergwall . Denitrification as an adaptive trait in soil and groundwater bacteria. November 27,1998. 17. A nna Wallstedt . Temporal variation and phytotoxicity of Batatasin-m produced by Empetrum hermaphrodytum. November 27,1998. 18. D arius Sabaliunas . Semipermeable membran devices in monitoring of organic pollutants in the aquatic environment. April 28,1999. 19. Cecilia A grell . Atmospheric transport of persistent organic pollutants to aquatic ecosystems. May 21,1999. 20. O lof Berglund . The influence of ecological processes on the accumulation of persistent organochlorines in aquatic ecosystems. September 17,1999. This thesis is based on the following papers

I. Berglund, O., Larsson, P., Ewald, G., and Okla, L. Bioaccumulation and differential partitioning of PCBs in freshwater, planktonic foods webs. (Submitted) n. Berglund, O., Larsson, R, and Broman, D., Organochlorine accumulation and stable isotopes in an Atlantic salmon (Salmo salar) population from the Baltic Sea - effects of omnivory and reproductive strategies. (Submitted)

III. Berglund, O., Larsson, R, Ewald, G., and Okla, L. The effect of lake trophy on lipid content and PCB concentrations in planktonic food webs. (Manuscript)

IV. Berglund, O., Larsson, P., Ewald, G., and Okla, L. Influence of trophic status on PCB dynamics in lake sediments and biota. (Manuscript)

V. Berglund, O., Larsson, P., Bronmark, C., Greenberg, L., Eklov, A., and Okla, L. 1997. Factors influencing organochlorine uptake in age-0 brown trout (Salmo trutta) in lotic environments. Can. J. Fish. Aquat. Sci. 54: 2767-2774.

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