Water toxicity and metal contamination assessment of a polluted river: the Upper Vistula River (Poland)

Applied Geochemistry 19 (2004) 153–162 www.elsevier.com/locate/apgeochem

Water toxicity and metal contamination assessment of a polluted river: the Upper Vistula River (Poland) C. Gue´guena,b,*, R. Gilbina,c, M. Pardosa, J. Dominika,b a

Institute F.-A Forel, University of Geneva, 10 route de Suisse, CH-1290 Versoix, Switzerland Centre d’Etudes en Sciences Naturelles de l’Environnement, University of Geneva, 10 route de Suisse, CH-1290 Versoix, Switzerland c Present address: IRSN—Laboratory of Experimental Radioecology, Cadarache, Bat 186, BP3, F-13115 Saint-Paul-Lez-Durance Cedex, France

b

Received 14 January 2003; accepted 13 May 2003 Editorial handling by R. Fuge

Abstract In aquatic systems, the bioavailability of an element to microorganisms is greatly influenced by its chemical speciation. The goal of this work was to assess metal toxicity to a green algae (Pseudokirchneriella subcapitata) and a bacterium (Vibrio fisheri) as a function of size fractionation and chemical speciation (using the program MINTEQA2) in contaminated water of the Upper Vistula River. Water samples were collected at 1 reference site, 4 polluted sites and one polluted site on the Vistula’s main tributary, the Przemsza River. Toxicity measurements were performed on unfiltered samples and, total dissolved (< 1.2 mm), and truly dissolved (< 1 kDa) fractions. Trace metal (Cd, Co, Cr, Cu, Mn, Pb, Zn) concentrations were measured in these samples and also in the colloidal fraction (1 kDa–1.2 mm). At the reference site, the low metal concentrations were in agreement with the absence of measurable toxicity. In the polluted section of the river, free metal concentrations were largely below the potential toxic levels for bacteria, which was in agreement with the absence of toxicity. Although Zn2+ was at potentially toxic-level concentrations in total dissolved and truly dissolved fractions in the polluted riverine section, toxicity for algae was observed, only in truly dissolved fractions from two stations. The absence of toxicity in most samples was related to metal association with particles and with low molecular weight ligands as well as the presence of organic ligands (phenol). The reason for toxic effects in two ultrafiltered samples is not clear, but may be related to the elimination of the colloidal organic fraction and thus the eradication of its protective effect occurring in natural samples. # 2003 Elsevier Ltd. All rights reserved.

1. Introduction The potential effects of toxic compounds in the environment are determined not only by their intrinsic toxicity and concentration but also by their physicochemical forms. Whereas the total metal concentration * Corresponding author at: IARC-Frontier, University of Alaska Fairbanks, 930 Koyukuk Drive, PO Box 757335, Fairbanks, AK 99775-7335, USA. Fax: +1-907-474-2679. E-mail address: [email protected] (C. Gue´guen).

is easier to measure, it is not a reliable indicator of toxicity (Campbell, 1995). The availability of an element to living organisms depends on its chemical speciation (e.g. De Haan et al., 1993). In natural water systems, trace metals can be partitioned between different physical states such as free or complexed, associated with colloids or with particles. It is often assumed that metals are mainly bioavailable in free ionic and labile form for microorganisms (Free-Ion activity Model, FIAM: Tessier and Turner, 1995), whereas the particle-bound or ligand-complexed metals are viewed as not directly

0883-2927/03/$ - see front matter # 2003 Elsevier Ltd. All rights reserved. doi:10.1016/S0883-2927(03)00110-0

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available for microorganisms (Campbell, 1995). In this context, toxicity bioassays with microorganisms can be useful tools to link the biological response with chemical speciation in natural waters (Stauber and Davies, 2000). The goal of this paper was to determine metal toxicity to two aquatic organisms, a green alga (Pseudokirchneriella subcapitata) and a bacterium (Vibrio fisheri) in water samples collected in a highly polluted river. The toxicity was related to the metal concentration using tangential flow ultrafiltration for size fractionation, the presence of metal binding ligands and the free trace metal concentrations calculated by MINTEQA2. Complexation experiments (fluorescence quenching) were also conducted to study the potential effects of phenol in metal binding. The study was carried out on the Vistula River, the largest Polish river (1047 km length) and the second largest in the Baltic Sea watershed. The water quality of the Vistula River deteriorated during the last 40 years (Niemirycz, 1997) but has improved in the last few years. Mining and heavy industries located in the drainage basin of the Upper Vistula River, upstream from Cracow (Southern Poland) cause trace metal contamination of soil and sediment (Helios-Rybicka, 1983; Suschka et al., 1994; Verner et al., 1996), water (Gue´guen et al., 2000; Gue´guen and Dominik, 2003; Kasza and Wojtan, 2002; Suschka et al., 1994) and aquatic plants (Lewander et al., 1996). The river is also polluted by phenols, which originate from infiltration of surficial contaminants (Helios-Rybicka, 1996) and from tanneries. In the earlier studies carried out in the region of Cracow, no toxicity was found in the Vistula River using P. subcapitata, V. fisheri (Pardos et al., 2000) and the snail Lymnaea stagnalis (Wagner, 2000) bioassays. However a clear toxic effect was observed for Hydra attenuata (Pardos et al., 2000). In the present study the authors focus on a more contaminated section of the Upper Vistula River and particularly the toxicity related to metal speciation; the influence of anthropogenic inputs on metal partitioning is discussed elsewhere (Gue´guen and Dominik, 2003).

2. Materials and methods 2.1. Sample collection Sampling was carried out in November 1998 in the Upper Vistula River at a low water flow, upstream from Cracow (Fig. 1). At that time, the river at Lipowiec and Gora was partly covered by ice. Surface water samples (10 l) were collected with a polypropylene (PP) drum at 6 stations. From upstream to downstream, the first station was Lipowiec (the reference station) which is located close to the Vistula River source; then, Gora, Bobrek, Metkow and Tyniec. Station Chelmek is located on the

Przemsza River, the most important tributary of the Upper Vistula River. The confluence with the Przemsza River is located between stations Gora and Bobrek. Industrial complexes, particularly coal and Pb–Zn mining, influence the region downstream from Lipowiec. Ultra clean conditions were maintained during all stages of sample collection, transport, handling, processing and analysis (EPA, 1996). The samples were immediately filtered in the field with a 1.2 mm polypropylene filter (Calyx, MSI). The tangential flow ultrafiltration (TFF) was performed on the total dissolved fraction ( < 1.2 mm) using a 1 kDa regenerated cellulose cartridge (PrepScale, Millipore) in the laboratory within 4 h (Gue´guen et al., 2002). The colloidal (1 kDa–1.2 mm) and truly dissolved fraction (< 1 kDa or < 1 nm) fractions were isolated. As the performance and retention characteristics of ultrafiltration cartridges may vary with the operating conditions (Guo, 1995), the effective cut off was checked using molecular probes (Gue´guen et al., 2002). The membrane cut off was evaluated at about 3 kDa. Cartridges were cleaned between each sample collection with NaOH, HCl and a large volume of ultrapure water. The concentration factors (ratio of the volume of the initial sample to the volume of the fraction retained by the ultrafiltration membrane) applied for TFF were between 2 and 4. 2.2. Chemical measurements For metal measurements, water samples were collected in a polypropylene acid-washed container and acidified (pH < 2) with ultrapure HNO3 (Merck, Geneva, Switzerland) and stored in darkness at 4  C until analysis. Water samples were digested using a microwave digestion system (ETHOS, Milestone, Sorisole, Italy). Fifty ml of sample was mineralised using 3 ml ultrapure HNO3 and 2 ml ultrapure H2O2. A multiplestage heating program was used: 2 min at 200 W, 10 min at 800 W and 10 min at 250 W, each at 20  105 Pa pressure. Appropriate batch blanks were prepared with each set of samples. In the unfiltered samples and in the total dissolved ( <1.2 mm), colloidal (1 kDa–1.2 mm) and truly dissolved ( < 1 kDa) fractions, concentrations of major elements (Ca, Mg) and trace metal concentrations (Cd, Co, Cr, Cu, Mn, Pb and Zn) were determined by ICP-AES (Perkin Elmer) and ICP-MS (HP 4500, Agilent), respectively. Certified standard reference materials (SRM 1643d from NIST) were analyzed to check the accuracy of metal determinations. The quantification was within 5% of its referred value for the riverine SRM (n=6). Blank concentrations were negligible as compared to metal concentrations in samples. The particulate metal content was calculated by difference between the unfiltered sample and the total dissolved sample. Mass balance of the ultrafiltration process was close to 100% for major and trace metals (91–106%).

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For organic C measurements, water samples were filtered with pre-combusted GF/F filters and acidified in the field to pH 2. Concentrations of dissolved organic C (DOC) and colloidal organic C (COC) were determined with a Shimadzu TOC 5000A analyser within two 3 weeks. Alkalinity, NH+ 4 , Cl , NHO3 , phenol and PO4 were determined by colorimetric measurements in the total dissolved fraction within 2 h of collection (Bartram and Ballance, 1996). Conductivity, salinity, and pH were measured within 2 h of collection by using a portable pH meter and conductimeter (WTW Instruments, Geneva, Switzerland). 2.3. Chemical speciation modeling The equilibrium concentrations of the metal species considered were calculated using the computer program MINTEQA2 (Allison et al., 1991). In the version 3.11, it includes a sub-model for computing the complexation of some metal cations with DOM (Allison and Perdue, 1994; Allison et al., 1991; Dobbs et al., 1989; Susetyo et

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al., 1991). This sub-model is based on the work of Dobbs et al. (1989) who considered DOM as a complex material consisting of many different types of monoprotic acid sites. The electrostatic interactions are not taken into account explicitly. The concentration of the binding sites is normally distributed with respect to their log K values for proton or metal binding. A database available for proton and metal interaction with Suwannee River DOM and their mean log K values is included (log K=4.9, 3.3 and 3.5 for Cu-DOM, Cd-DOM and Zn-DOM, respectively). The composite ligand component DOM represents a complex mixture of ligands without distinction between humic and fulvic fractions. The present study is restricted to the complexation of three metals (Cu, Zn, Cd), each in the +2 oxidation state. Manganese(II), Fe(III), Ca(II), Mg(II) and Na(I) were also introduced into model calculations. Their respective concentrations were detailed in Gue´guen and Dominik (2003). The total dissolved concentrations of K+ and SO24 taken from the literature (Baradziej, 1998) have been included in the MINTEQA2 calculations.

Fig. 1. Location of sampling sites () in the Upper Vistula River (Southern Poland).

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For each complex species considered, the stability constants were fitted to an interpolation function that has the form of a Debye–Hu¨ckel equation. The model’s inputs comprised the total dissolved concentrations of metal and ligands (DOM, carbonate, Cl , NO3 , PO34 ) in the solution, pH, temperature, alkalinity and ionic strength. The constants for the inorganic side reactions of metal with CO23 , OH , Cl , NO3 and PO34 were taken from Sillen and Martell (1964). Although Eh measurements are relatively straightforward, the assumption of a reversible chemical equilibrium, and the lack of interfering reactions at the electrode surface are hardly met in natural water, particularly in polluted environments. However, as the Upper Vistula River was partly covered by ice during the water sampling, the suboxic conditions are expected and were consistent with high concentrations of dissolved Mn (Fig. 2). Colloidal Fe can be significant in metal transport in some acid mine drainage areas (Schemel et al., 2000; Kimball et al., 1995), Also in the study, Fe was mostly found in the colloidal fraction in the Upper Vistula River (Gue´guen and Dominik, 2003). However, Fe was only a minor constituent of colloids that were mainly composed by colloidal organic C and Ca (Gue´guen and Dominik, 2003). 2.4. Metal–phenol complexation The complexation properties of phenol were studied in Milli-Q water (Millipore system, >18 M
) by fluo-

rescence in the presence of one paramagnetic trace metal, Cu. Such a metal ion is especially suited to act as a quencher. This property has been used for quantifying metal complexation, by humic substances (Ryan and Weber, 1982; Susetyo et al., 1991). The study of Cuphenol complexation was performed at lex=250 nm with concentrations of Cu (80 mg/l) and phenol (1000 mg/l), close to those measured at most polluted stations in the Upper Vistula River. Fluorescence measurements were performed with a Fluorolog 212 SPEX fluorimeter equipped with double monochromators for both excitation and emission. Measurements were performed in a 1 cm path length fused silica cell thermostatically controlled at 20  C. The bandwidths for excitation and emission were both 4 nm, while wavelength increment and integration time were 0.5 nm and 0.5 s, respectively. All spectra, both in emission and excitation modes, were corrected for instrumental factors (De Souza Sierra et al., 1994). Then, the Milli-Q water blank was subtracted from each spectrum to remove the Raman scattering band of water. 2.5. Bioassays Bioassays were performed on two aquatic organisms, the green microalga P. subcapitata (formerly known as Selenastrum capricornutum Printz) and the bacterium V. fisheri (formerly Photobacterium phosphoreum, Microtox1 test). These two unicellular aquatic organisms are widely used and have been standardized for the study of the toxicity of polluted water samples. Data for in vitro

Fig. 2. Concentration of trace metals (mg/l) in the (&) particulate, (&) colloidal and (&) truly dissolved fractions. Stations were noted according to their initial.

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25%, . . . of osmotic-adjusted sample in 2% NaCl standard medium), was measured with the model 500 Azur Microtox1 luminometer. The effective concentrations inducing 50% effect (EC50) were calculated for each bioassay. For the algal test, the EC50 was calculated using the bootstrap method in the Regtox Excel macro, which is available from http://perso.wanadoo.fr/eric.vindimian/ (E. Vindimian, INERIS, France). The comparison of the toxicity was based on 50% effects (EC50) calculated within 95% confidence intervals from 500 non-parametric bootstrap simulations with Lambda=0.01, Delta Lambda=5, and Convergence Criterion=0.00001. The effect for zero doses and the maximum effect were imposed in the model at 0 and 100, respectively. For the Microtox1 bioassay, the 30 min EC50 with 95% confidence intervals, were calculated using a linear regression package available from the manufacturer of Microtox1 (Microbics Corporation, 1992). The results were expressed in percentage (v/v) of natural sample added to culture medium. EC50 value is inversely related to toxicity. The observed toxicity was reported as no toxicity (EC50=100%), slight toxicity (100% > EC50 > 95%) and high toxicity (EC50 < 95%).

toxicity are available on a large panel of toxicants (including trace metals), which allows comparisons with the present study. These tests are performed using simple synthetic media, for short-term exposure, that limit chemical speciation changes during the tests. All water samples were kept frozen until use. Toxicity bioassays were performed at each station on the unfiltered and total dissolved samples, in triplicate, on a geometric range of dilutions with standard media (USEPA medium and 2% NaCl for P. subcapitata and V. fisheri, respectively). As the measurement of toxicity for P. subcapitata is based on changes in algal cell number, the presence of natural particles in the unfiltered sample may disturb the cell counting. Consequently, only the total dissolved and truly dissolved fractions were tested. Bioassays using the green microalgae P. subcapitata (initial concentration of 104 cells per mL) were performed in microplates during 72 h at 24  C under 60–80 mmol/m2/s, cool white lighting (Environment Canada, 1992). To eliminate false negative results due to low nutrient concentrations in samples, 1% (v/v) of stock nutrient solution (standard culture medium 1000  concentrated; USEPA, 1989) was added to samples prior to use in preparing the test dilutions. Thus all test solutions contained at a minimum the nutrient concentration in the stock culture medium. EDTA was not added to the culture medium to avoid competition between EDTA and other ligands for metal complexation (Koukal et al., 2003). The growth, inhibition or stimulation was measured in each dilution of sample (100, 50, 25. . . of sample in standard medium) by cell counting with a Coulter Counter Z1 and compared to growth in standard medium (USEPA, 1989). Bioassay using the luminescent bacterium V. fisheri was performed according to the standard Microtox1 test (Microbics Corporation, 1992). Inhibition of the bacteria luminescence in each dilution of the unfiltered, total dissolved and truly dissolved samples, (100, 50,

3. Results 3.1. Physicochemical parameters As the sampling was carried out during winter the temperature of the water was low (from 0.0 to 4.9  C) (Table 1); pH was comparable for all stations (7.2–7.8); the samples were also characterized by DOC concentrations which ranged from 4.8 to 7.8 mg/l and detectable phenol concentrations in three samples (up to 0.7 mg/l). Important increases in NH+ 4 concentrations, occasionally accompanied by lower NO3 values, suggest

Table 1 Chemical parameters of water samples in the Upper Vistula River Station

Cond. Salinity pH T ( C) (mS/cm) (%)

Alkalinity Hardnessa NH+ NO3 PO34 Cl phenol SPMb DOCc COCd 4 (mmol/l) (mmol/l) (mmol/l) (mg/l) (mg/l) (mg/l) (mg/l) (mg/l)

Lipowiec Gora Chelmek Bobrek Metkow Tyniec

0 1.4 4.9 3.7 3.1 2.1

0.5 1.9 4.2 3.5 2.8 2.0

a b c d

75 2350 1313 3570 1480 800

0.0 2.4 1.3 3.5 1.5 0.9

7.66 7.23 7.75 7.51 7.60 7.48

Hardness was calculated as Ca2++Mg2+. SPM=suspended particulate matter. DOC=dissolved organic carbon. COC=colloidal organic carbon.

2.22 10.28 10.43 14.07 8.18 6.35

30 260 210 220 110 50

130 48 160 80 160 190

<8 14 22 9 16 9

6 1321 491 2500 724 590

<0.2 0.7 <0.2 0.6 0.3 <0.2

5.3 45.6 13.9 99.5 39.2 5.4

5.2 7.8 6.2 6.3 6.6 4.8

4.9 6.9 5.8 2.4 5.6 1.8

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suboxic conditions especially at stations Gora and Bobrek. Phosphate and carbonate (alkalinity measurements) concentrations also increased downstream from the reference point, Lipowiec. Significant increases of hardness (Ca2++Mg2+) from 2.22 meq/l at Lipowiec to 14.07 meq/l at Bobrek and salinity (0.0, 2.4 and 3.5% at Lipowiec, Gora and Bobrek, respectively) were also observed. 3.2. Metal distribution Detailed results of metal concentrations and origin have been described previously (Gue´guen and Dominik, 2003). Therefore only a brief overview of trace metal behaviour is presented below. The concentrations of trace metals in the polluted section of the Upper Vistula River (Table 7 of Gue´guen and Dominik, 2003) were much higher than in the ‘‘world average river’’ (Meybeck, 1988; Meybeck and Helmer, 1989). The metal concentrations measured in this study were similar to those previously reported by Cydzik et al. (1996) and Gue´guen et al. (2000). Total metal concentrations were generally lower in the unpolluted sections and higher in the polluted sections than yearly mean concentrations reported by Kasza and Wojtan (2002), which is consistent with the low flow conditions during the sampling in winter 1998. The metal concentrations in the unfiltered fractions were lower at Lipowiec than at the other stations (Fig. 2). They largely increased from Gora and Chelmek located in the Przemsza River. At Bobrek, located downstream from the confluence with the Przemsza River, metal concentrations were generally the highest. At Lipowiec, metals were mainly found in the total dissolved fraction, except for Pb, which was predominantly associated with particles. Downstream from Lipowiec, Cu, Mn, and Co strongly increased especially in the total dissolved, while Cr was mainly in the

particulate fractions. At Chelmek and downstream from the Vistula–Przemsza confluence Cd and Pb concentrations in the particulate fraction increased strongly, while Zn concentrations increased considerably also in the total dissolved fraction. Differences between Lipowiec and the other stations were also noted in colloid/truly dissolved partitioning. The colloid fraction as a percentage of the total dissolved metal concentration was considerable at Lipowiec (10% for Co and Zn, 33% for Cu and 55% for Cr), but was much lower downstream (45%). For Pb, the colloidal fraction was quite similar at the reference point (16%) and the polluted sites (7–25%). The fractions of Cd and Mn associated with colloids were negligible at all stations. 3.3. Metal–phenol complexation In the presence of Cu, the fluorescence signal of phenol (lex=250 nm) decreased (Fig. 3), meaning that the phenolic compounds have complexed Cu in solution. The metal complexation depends on the chemical behavior of the metal being studied. Since, Cu and Zn have mostly A-type character, their affinity for ligands are comparable. Therefore, Zn–phenol association may also occur. 3.4. Toxicity bioassays The EC50 values for each individual metal added in the ionic form were obtained under the same conditions (particularly EDTA-free standard culture medium in algal bioassays) as for the natural samples. The EC50 values found in this study were in the same order of magnitude as these reported in the literature (Table 2) although for the algal bioassay they are closer to the lower limit of the cited ranges, probably due to the lack of metal–EDTA complexes in the study. Regardless of the sampling station, the unfiltered samples and the total dissolved fractions were non-toxic for V. fisheri, whereas only the total dissolved fraction from Bobrek was possibly slightly toxic (EC50 > 99%) for P. subcapitata (Table 3). After colloid removal by ultrafiltration, the truly dissolved fractions from Chelmek and Tyniec were toxic for P. subcapitata (EC50=69 and 35%, respectively). It can also be noticed that the toxicity at Bobrek was similar before and after colloid removal.

4. Discussion Fig. 3. Fluorescence spectra (lex=250 nm) of phenol and phenol+Cu. As Cu is a paramagnetic metal, the complex. Cuphenol is observed by fluorescence quenching of phenol signal. The phenol concentration after 3 months was similar.

Increasing water hardness generally decreases metal toxicity, possibly due to Ca competition of the cell surface (Jayaraj et al., 1992; Rai et al., 1981). An increase in salinity may also affect the toxicity partly because of

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changes in chemical speciation and in the physiological characteristics of organisms (Hall and Anderson, 1995). Compared to the salinity and hardness increases, the variations of DOC and nutrient concentrations at different stations were small, except for NH+ 4 . Therefore, these parameters can be regarded as only marginally influencing relative changes in the pollutant speciation and the bioavailability for microorganisms in the studied system. 4.1. Toxicity and size partitioning At the reference point, Lipowiec, the metal concentrations in the unfiltered samples and total dissolved Table 2 EC50 (mg/l) for P. subcapitata and V. fisheria Metal

Organisms

EC50 (mg/l)

Reference

Cd

P. subcapitata

65–74 17–452 7965 16,500 40–900 13,756 16,000 12–19 12–917 580 150 1900–4600 122 160 55–79 45–2600 488 1200

This study [1,2] This study [3] This study This study [3] This study [2] This study [3] [2] This study [3] This study [1,2] This study [3]

V. fisheri CrVI

P. subcapitata V. fisheri

Cu

P. subcapitata V. fisheri

Pb

P. subcapitata V. fisheri

Zn

P. subcapitata V. fisheri

a

For bacteria, the exposure duration was 30 mm. In this study the AAP medium for P. subcapitata tests was EDTA-free. [1] Blaise et al. (1998); [2] USEPA (2001); [3] Blaise et al. (1994).

159

fractions were the lowest, well below the EC50 for both organisms, which was in agreement with the absence of toxicity. Downstream, the metal concentrations increased dramatically. Still the EC50 values for V. fisheri were mostly higher than the measured metal concentrations, in agreement with the absence of toxicity. Although the concentrations of Pb in the unfiltered samples from Chelmek and Bobrek were close to potentially toxic-levels for V. fisheri, no toxicity was observed. The size fractionation showed that Pb was mainly associated with particles, presumably less or non-available to bacteria. Toxicity of metals bound to particles could be more relevant for other organisms than the unicellular organisms tested (e.g. fish or benthic organisms that could eat particles). But this contamination pathway is not consistent with the Free Ion Activity Model, which is actually considered to be valid for unicellular organisms and direct contamination through water, eg. via fish gills (Campbell, 1995). Similarly, at Cu and Zn concentrations above the toxic-levels (EC50) in the majority of the samples, no toxicity for P. subcapitata was observed in the total dissolved fractions, except for a slight toxicity for the sample from Bobrek (Table 3). 4.2. Toxicity and chemical modeling (MINTEQA2) For P. subcapitata, EC50 values were lower than the measured truly dissolved concentrations of Cu (Gora and Bobrek) and Zn (Chelmek, Bobrek, Metkow and Tyniec). However, only the truly dissolved fractions from Chelmek and Tyniec (and to a lesser degree Bobrek) were toxic. While this might be explained by their high metal concentrations, no correlation between metal concentrations and toxicity was observed, suggesting that metal speciation in the truly dissolved fraction may be more important than the absolute metal concentration. According to the FIAM model (Tessier and Turner, 1995), the biological effects to microorganisms

Table 3 Measured concentrations (mg/l) of Cu and Zn in the total dissolved and truly dissolved fractions, calculated concentrations of free metal ion concentrations (mg/l) in the Upper Vistula River compared to the observed toxicity for P. subcapitata in standard culture medium (without EDTA, 72-h test) exposed to filtered (Tot. diss.) and ultrafiltered (Truly diss.) water samplesa Station

Lipowiec Gora Chelmek Bobrek Metkow Tyniec

Cu

Zn

Phenol (mg/l)

Total diss.

Cu Truly diss.

Cu2+

Total diss.

Zn Truly diss.

Zn2+

0.90.1 46.3 0.2 17.8 0.5 90.0 2.3 25.6 0.2 18.3 0.1

0.60.0 43.71.0 16.80.5 82.12.0 23.80.3 18.10.2

0.09 5.59 0.63 4.05 0.75 1.84

16.20.1 17.30.1 145.50.4 129.20.9 168.71.1 99.90.3

13.20.1 15.10.2 142.22.3 129.80.9 162.01.1 91.21.2

11.9 13.0 57.2 71.5 86.7 69.1

<0.2 0.7 <0.2 0.6 0.3 <0.2

Toxicity Total diss.

Truly diss.

– – – + – –

– n.m. ++ + – ++

a Observed toxicity for P. subcapitata: (–) no toxicity, (+) slight toxicity, and (++) high toxicity EC50 for P. subcapitata measured in this study.

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should mainly be related to the free metal ion concentrations (Sunda and Lewis, 1978; Luoma, 1983; Martell et al., 1988; Shuttleworth and Unz, 1991; Campbell, 1995). The concentrations of free metal ions in the total dissolved and truly dissolved fractions were thus calculated by MINTEQA2 and compared to the observed toxicity (Table 3). Calculations for colloid-free water (equivalent of ultrafiltered fraction) yield the free metal ion concentrations comparable to those in total dissolved fractions. Consequently, no significant differences in results of bioassays carried out on total dissolved and truly dissolved (ultrafiltered) fractions were expected. Copper2+ concentrations were lower than the EC50 for algae, and according to the MINTEQA2 model the main forms were hydroxycomplexes. However, Zn2+ concentrations were at potentially toxic levels for algae, except at Lipowiec and Gora. Three situations were observed:  No toxicity occurrence: at Gora, the Zn2+ concentrations were lower than the EC50 for algae, whereas at Metkow, they were at potential toxic-levels. For both stations, elevated phenol concentrations were measured, but as no thermodynamic stability constants for phenol are available in MINTEQA2, this ligand was not included in the model. Thus, the calculated free metal ion concentrations were possibly overestimated.  A weak toxicity in the total dissolved and truly dissolved fractions: at Bobrek, where Zn2+ concentration was relatively high. Due to the large concentration of phenol, which constitutes a potential ligand for trace metals, the concentrations of free and labile metals were again overestimated. Nevertheless, a slight toxicity occurred, suggesting that ligand concentrations were not sufficient to complex metal and prevent toxicity.  High levels of toxicity occur in the truly dissolved fractions, but not in the total dissolved fractions, at Chelmek and Tyniec where Zn2+ concentrations were at potentially toxic levels in both fractions. Toxicity for algae was observed only in the truly dissolved fraction, but not in the total dissolved fraction. That is perhaps the most puzzling result in this study. The truly dissolved and total dissolved fractions have comparable concentrations of free metal ions, and thus a possibility of antagonistic effects occurring only in the total dissolved fraction can be dismissed. The absence of toxicity in the total dissolved fraction could possibly be explained by a sorption of the colloidal organic matter to the algal cell surface (Pempkowiak and Kosakowska, 1998), thus protecting them from the

pollutants in water. These authors suggested that the sorption of organic colloids may decrease membrane permeability for Cd by decreasing the number of available adsorption sites. A similar mechanism was proposed by Campbell et al., (1997, and references therein) and recently demonstrated by Koukal et al. (2003) for Cd and Zn in the presence of humic acids. Indeed, this seems to be a plausible reason for the absence of Zn toxicity in the total dissolved fraction at Tyniec and Chelmek. For example at Tyniec, the station where the COC is the lowest of this study, the surface area of colloids and algae were calculated. Assuming that the colloids are mainly humic substances (Mw 1000–3000 g/mol, Buffle, 1988), the colloid surface was estimated at 3–9.2 m2/l. The surface area of P. subcapitata (taken as an half ellipse was calculated at 0.013 m2/l for 106 cells /ml (number of cells obtained after incubation for 72 h), meaning that there were enough organic colloids to cover the membranes of algae during the bioassays.

The goal of the study was to determine metal toxicity in water samples collected in the Vistula River, where the main contamination by trace metals is well documented. However, the study showed that organic pollutants (eg. Phenol) were present in combination with metals. The identification of toxic agents in such mixtures requires knowledge of (1) single toxicity of each contaminant (2) synergistic and antagonism effects and (3) chemical speciation that controls the extent of metals bioavailability. The authors choose to discuss the results obtained on the third point, assuming that other contaminants were negligible with regard to toxicity. On the other hand, results on single species cannot be related to effects on communities: the extrapolation of single-species laboratory studies to natural communities is one of the problems encountered when carrying out risk assessment of chemicals in the environment (Jak et al., 1996). It is realized that the work presented here is only the first step and that more samples from the Upper Vistula River are needed, plus more complete chemical characterization of samples (especially organic pollutants), to more rigorously define the interactions between metal phase speciation and toxicity. Some inferences can, however, be made from the data at hand. The bioassay’s sensitivity towards compounds in natural samples is variable according to the species tests. In this study, algae appeared to be more sensitive to contaminants present in samples from the Vistula River than bacteria. As shown above, a large fraction of Cd, Cr and Pb was associated with particles.

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5. Conclusion

References

Standardized laboratory bioassays for water quality assessment and evaluation of potential impact of contamination on aquatic ecosystems are certainly valuable complements to the chemical quality evaluation. This study shows that metal concentrations potentially harmful for aquatic microorganisms as established by exposing organisms to metals added to the culture medium do not reflect the toxic effects induced by exposure to contaminated water samples. This is related to numerous factors influencing metal bioavailability in natural aquatic environment. Toxicity to the algae was observed only in some samples in which metal concentrations were high. Although Cu and Zn concentrations were at potentially toxic levels (above the EC50 for algae), total dissolved fractions were non-toxic, suggesting that metals were complexed and thus unavailable for algae. Unambiguous toxicity to P. subcapitata was only observed in the truly dissolved fraction at two stations. The concentrations of truly dissolved ligands, the potential complexation with organic pollutants (phenol) and interactions with colloids by sorption processes to algal cells are suggested as parameters determining the presence or absence of toxic effects. As V. fisheri is generally more sensitive to organic micro-pollutants than to metals, the absence of toxicity to these bacteria suggests that metals rather than other pollutants caused the toxic effects observed in algal bioassays. This study shows that metal size fractionation and calculation of the free metal concentrations and use of two different microorganisms in bioassays are useful tools to interpret the occurrence of toxicity. The protective role of colloids in natural systems deserves further study.

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Acknowledgements We gratefully acknowledge P. Arpagaus, R.L. Thomas, E. Szalinska and Z. Slusarczyk for their help during the field sampling in Poland. We thank all our colleagues from the IZWiOS at the Technical University of Cracow for providing laboratory facilities and technical assistance in Poland. We are indebted to C. Belin for the fluorescence facility use at the Molecular Physico-chemistry Laboratory, University of Bordeaux (France). Comments of Dr Jamieson and an anonymous referee greatly improved the quality of the manuscript. This work was funded by Swiss National Foundation 20-50796-97.

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