Free cupric ion concentration and Cu(I1) speciation in a eutrophic lake

Transcription

1 Limnol. Oceanogr., 38(6), 1993, , by the American Society of Limnology and Oceanography, Inc. Free cupric ion concentration and Cu(I1) speciation in a eutrophic lake HanBin Xue and Laura Sigg Institute for Water Resources and Water Pollution Control (EAWAG), Swiss Federal Institute of Technology, CH Diibendorf Abstract The free cupric ion concentrations, [Cu2+], in the water column ofeutrophic Lake Greifen were evaluated by means of ligand exchange with catechol and cathodic stripping voltametry of the Cu-catechol complexes. Total dissolved Cu, [Cu],, was in the range x 10-8 M, while [Cu2-I] ranged from lo-l6 to lo-l4 M at different times and depths. These values of [CU~-~] are much lower than literature values for the open sea. Equilibrium models of titration data indicate the presence of very strong ligands with conditional stability constants of 1013~y-10t4~9 and corresponding concentrations of nm. Measured [Cu2+] was especially low during the algal bloom in spring and was higher in winter. p[cu] and ratios of [Cu], to [Cu2 t ] thus exhibit a seasonal pattern similar to that of algal productivity, suggesting that the Cu-complexing ligands are produced by algae. Several trace metals such as Cu, Zn, and Fe are essential to biological activity as nutrients, but are toxic at elevated concentrations. The biological availability of trace metals in terms of nutrient limitation and toxicity is related to free aquo ion concentrations and not to concentrations of total metals or metal complexes (Sunda and Guillard 1976; Anderson and Morel 1982; Sunda and Ferguson 198 3). Trace metals tend to be complexed by organic and inorganic ligands in natural waters; Cu forms stronger organic complexes than most other divalent metals. The sensitivity of some organisms to very low concentrations of free cupric ions, [Cu2+], has been demonstrated (Sunda and Guillard 1976; Brand et al. 1986). There is, therefore, much interest in the determination of [Cu2+] and of the concentrations and stability constants of natural ligands. Cu complexation in natural waters has been extensively studied, mostly in seawater (Sunda and Hanson 1987; Sunda and Huntsman 1991; Moffett et al. 1990; Van den Berg 1984a). Numerous techniques have been used to measure [Cu2+] and stability constants and concentrations of Cu-complexing ligands. Each method has its own advantages and limitations. Voltametric methods have been used because they distinguish between labile species Acknowledgments We thank W. Sunda, W. Stumm and B. Wehrli for discussions, H. Ambiihl and collaborators of the limnology department of EAWAG for data on algal productivity, and D. Kistler and A. Kuhn for sampling. (usually inorganic species and weak organic complexes) and nonlabile complexes (Coale and Bruland 1988, 1990). Fixed potential amperometry avoids many of the uncertainties related to reduction of labile organic complexes (Waite and Morel 1983; Hering et al. 1987). Ion-selective electrodes have been used in some cases to measure ambient [Cu2+] but are usually not sensitive enough. Bioassays (Sunda and Ferguson 1983; Anderson et al. 1984; Hering et al. 1987) provide direct information on the biological relevance of the actual concentrations, but generally lack sensitivity and represent very involved experiments. Sorption of hydrophobic Cu complexes onto Cl8 SEP-PAK cartridges could not completely separate complexed Cu, because only a fraction of the Cu chelates is sufficiently hydrophobic to be extracted in this manner (Hanson et al. 1988). Ligand-exchange techniques are based on the competition for Cu complexation between natural ligands and added ligands, such as catechol or EDTA, and the subsequent specific determination of concentrations of these complexes. [Cu2+] is then determined by equilibrium calculations. These methods allow us 1200 to work at the natural levels of total Cu concentration, [CU],, and to determine indirectly very low values of [Cu2+]. Ligand-exchange techniques, combined with cathodic stripping voltametry, CSV (Van den Berg 1984a), sorption on Cl8 SEP-PAK cartridges (Sunda and Hanson 1987), liquid-liquid partitioning (Moffett and Zika 1987), or chemiluminescence (Sunda and Huntsman 199 1) have been used

2 Cu speciation in a eutrophic lake 1201 to measure high levels of Cu complexation in seawater. By these techniques, the presence of strong ligands and low [Cu2+] has been demonstrated in seawater. Very little information is available on ambient Cu complexation in freshwater, which often contains high levels of dissolved organic matter (Van den Berg et al. 1987). In this work we apply CSV to Cu-catechol complexes (Van den Berg 1984a) to measure very low [Cu2+] in a highly eutrophic lake. The working conditions in these samples are carefully evaluated. Temporal and spatial variations of [Cu2+] and Cu complexation in the water column of a small eutrophic lake are presented. Conditional stability constants and concentrations of Cu-complexing ligands are evaluated, and the sources of strong organic ligands that affect Cu speciation are discussed. Theory of ligand exchange and calculation of [Cu2+] The determination of [Cu2+] and complexation parameters for natural organic ligands is based on the competition between natural organic ligands and the ligand catechol, which is added in excess (Van den Berg 1984a). The concentration of Cu-catechol complexes at equilibrium depends on the concentrations and stability constants of the natural ligands in the sample. After equilibrating the sample with catechol, the concentration of Cu-catechol complexes is determined by differential pulse cathodic stripping voltametry (DPCSV); the reduction current of Cu-catechol complexes adsorbed onto the surface of a hanging mercury drop electrode (HMDE) is measured. [Cu2+] is calculated from equilibrium relationships with the catechol complexes. The total concentration of dissolved Cu in the original water sample is given by D&r = [CU2+] + [CU]i, + Z[CULi]. (1) [Cu]i, and [CULi] represent Cu concentrations as inorganic complexes and complexes with natural organic ligands, respectively. Equation 1 can be rewritten as D-d. = [CU2+](1 + (Yin + ZKi[L,]) (2) where CYin = Z!l &cob[c032-]i + Z,&,JOH-1 + l&m-1 + Pso4[SQ12-I. pi represents the stability constant of the ith Cu complex with a specified inorganic ligand i, ai, can be calculated from the major ion composition of lake water, Ki represents the conditional stability constant of the Cu complex with a natural organic ligand Li, and [Li] the concentration of free ligand (uncomplexed); the product KJL,] is defined as the complexing coefficient of ligand Li with Cu. After adding catechol to a sample of natural water and equilibrating the ligand exchange between catechol and natural organic ligands, the dissolved Cu is distributed as follows: [CU], = [CU2+] + [CU], + Z[Cu(cat)J + Z[CuLJ (3) where Z[Cu(cat)J is the concentration of Cucatechol complexes. The catechol complexes are determined selectively by DPCSV, together with [Cu]i,: [CU2+] + [CU], + Z[Cu(cat)J = i,ls (4) where ip (ampere, A) is the peak current and S (A M- l) expresses sensitivity. Cu-natural organic complexes do not contribute significantly to the peak current at the reduction potential of Cu-catechol in seawater (Van den Berg 1984a) and in lake water (this study), probably due to their stronger stability or (and) weaker adsorption onto the electrode surface. The free cupric ion in solution in any case must equilibrate with all species of catechol complexes. Therefore, [Cu2+] in the presence of catechol can be calculated from Z[Cu(cati)]: with Z[Cu(cat)J Icu2+l = &,Jcat2-] + P2cat[cat2-]2 = Wu(cat) A CY cat a! cat = Pkat[cat2-l + P2cat[cat2-12e (5) P 1 cat and P2cat represent the stability constants of Cu(cat) and Cu(cat)22- complexes. The concentration of free deprotonated catechol, [cat -], can be calculated from mass balance and the acid dissociation equilibria of catechol: X[Hicat] = [H,cat] + [Heat-] + [cat -] = [cat], - {X[Cu(cat)J + Z[M(cat)]} (6)

3 1202 Xue and Sigg where X[M(cat)] stands for the total concentration of catechol complexes with metal ions other than Cu 2+. For major cations, only the complexation constant of Mg with catechol is available; the complexation constant with Ca is expected to be similar. The complexes with Mg and Ca are both negligible at ph 8. The concentrations of catechol complexes with Cu and other trace metals are also negligible compared to the high [cat],. Only the acid dissociation equilibria of catechol must thus be taken into account. Equation 3 can be rewritten as [CU], = [CU2+](1 + ai, + O!,at + I: Ki[Li]). (7) From Eq. 7 the sum z KJLJ can be obtained for known [Cu], and [Cu2+] calculated according to Eq. 5 - t1 + % + &at)- (8) Concentration of free ligands and the corresponding Cu-complexing coefficient might remain constant after addition of catechol if the ambient [Cu] is much lower than that of ligands. [Cu2+] in the original sample (in the absence of catechol) can thus be calculated from the above complexing coefficients for equilibria between Cu and natural ligands and from the mass balance: [cu2+] = D& 1 + ai, + X KJLi] (9) Experimental Lake Greifen (Switzerland) is highly eutrophic with a surface area of 8.5 km2 and a volume of 150 x 1 O6 m3. It has an average depth of 17.7 m and a maximum of 32.2 m. Water samples were collected from the deepest point of the lake. Its tributaries are strongly loaded with nutrients and pollutants from sewage and agriculture. Go-Flo sampling bottles (General Oceanics, 5 liters) were used to collect samples from different depths. Under N2 pressure, the samples were transferred to polyethylene bottles (Sigg et al ). Sample aliquots were filtered after transport to the laboratory. The filtration device and filtering membranes (0.45 pm) were washed with M HN03 and rinsed with bidistilled wa- ter. Aliquots of the filtered samples were immediately acidified to 0.01 M HN03. Total dissolved Cu was measured by differential pulse anodic stripping voltametry (DPASV). Total dissolved Cu and Cu in unfiltered acidified samples were also measured by graphite furnace atomic absorption spectrometry (AAS). The filters were digested with aqua regia; Cu in the digested solutions was measured by graphite furnace AAS. Dissolved organic C (DOC) was measured by combustion at 680 C on a Shimadzu TOC 500 instrument. Samples from the lake contained 3.5 mg liter- DOC. The original ph (measured in situ with a combined sensor, Ziillig) was in the range Total chloride was 7 x 10m4 M, sulfate 2 x 1O-4 M, alkalinity 3-4 x loa M, and Ca l-2 x 1O-3 M (Sigg et al. 1991). DPCSV measurements were carried out within 3 d after collection of samples. Water samples used in each series of CSV measurements were stored in the dark at 4 C and filtered just before use. A sample from lake Cristallina (Tessin, Switzerland), a small acidic lake (ph - 5) with low DOC (SO.5 mg liter- ) was measured for comparison. DPCSV measurements of Cu-catechol complexes -DPCSV was performed with an HMDE, an Ag/AgCl reference electrode, and a graphite counter electrode held in a Metrohm VA 663 stand combined with a Metrohm E506 polarecord. Catechol was added to the samples after 5 min of purging with Suprapure N2. Equilibration was allowed for 5 min under N2. A new Hg drop was made and the stirrer switched on simultaneously. The Cu-catechol complexes were collected for 3 min at the electrode without applied potential. After the collection period, the stirrer was turned off and 15 s later the voltage scan was started in the negative direction. Scanning parameters were initial potential of 0 V (vs. the Ag/AgCl reference electrode), pulse height of 50 mv, and scan rate of 5 mv s-l. DPCSV measurements were carried out at C. The reduction peak potential for Lake Greifen water was mv. In the absence of catechol, no significant current peak was measured, which means that under the conditions used, dissolved inorganic species of Cu do not contribute to the measured signal. Thus, the theoretical analysis of

4 Cu speciation in a eutrophic lake 1203 the data could be simplified, and the first two terms on the right-hand side of Eq. 3 were ignored. DPASV measurement of total Cu concentration- [Cu], was determined by DPASV in acidified filtrates. Equipment and electrodes were the same as those used in DPCSV. After purging a sample with Suprapure N2 for 10 min, Cu was deposited at the Hg electrode at a potential of -0.4 V for 3 min under stirring and for 30 s of rest time. Anodic scanning was performed at a rate of 2.5 mv s- l and a pulse height of 50 mv. Titration curve with Cu and determination OfDPCSVsensitivity-To obtain a Cu titration curve, we spiked a series of subsamples with different Cu concentrations. We pipetted 25- ml subsamples into a series of 50-ml highdensity polyethylene beakers; 150 ~1 of HEPES buffer (a solution containing 1 M N-2-hydroxyethylpiperazine-N -2-ethanesulfonic acid and 0.5 M NaOH) was added to each beaker to give a final concentration of 6 x 1 O-3 M HEPES and a final ph of 8.OkO.05. Cu was added to all beakers but one, giving a concentration range between 7.8 x 1 Oeg and 5.5 x 1 O-7 M in 1 O-l 5 steps. The series was allowed to equilibrate at C for h. DPCSV analyses were performed in the same polyethylene beakers to minimize adsorption effects. After purging with Suprapure N2 for 5 min, ~1 of catechol was pipetted into the beaker to give the desired concentration, 1 x 10m3 M in most experiments. After a 5-min equilibration, CSV measurement was carried out. A titration curve was obtained in terms of peak current (ij as a function of [Cu], for a single sample of lake water. The sensitivity had to be calibrated for each individual sample; it was determined from the portion of the titration curve at high concentrations of Cu2+, where stronger organic ligands were saturated, and essentially all of the Cu was complexed by catechol. The titration with Cu of a blank solution containing only KNO, (4 x 1 O-3 M) and HEPES buffer (5 x 1 O-3 M), as well as catechol (2.5 X 10B5 M), gave a linear relationship between the peak current and the Cu concentration in the range of 0.1 x 1O-g-1 x 1O-8 M added total Cu (TCu); the blank value obtained from this experiment was x 1O-g M TCu. The kinetics of ligand exchange between added catechol and natural ligands were examined in a time-dependent experiment. DPCSV peak currents were followed with time after addition of 1 X 10m3 M catechol to a lake-water sample under N2 to protect catechol from oxidation. The measurements yielded re- producible peak currents (&3%) over a period of 2 h after mixing and indicated that equilibrium of the ligand exchange was reached in a few minutes. An equilibration time of 5 min was thus used in all measurements. To minimize adsorption of Cu on beaker walls, we carried out tests of Cu-spiked lake waters (10 pg liter-l) by DPCSV in different beakers consisting of different materials, namely Teflon, high-density polyethylene, lowdensity polyethylene, and glass. No obvious differences in reduction current were found among different materials at the same incu- bation time (20 h). The S.D. of these results from the five beakers was +2.9%, which implies there was no serious adsorption on the beaker walls, probably due to competition by the strong natural ligands. High-density polyethylene beakers (50 ml) were chosen as polarographic cells (Miiller 1989). HEPES buffer had negligible complexing effects (Good et al. 1966). No complexing effect was detected in the titration of a blank solution (mentioned above) as well as in the experiments with model solutions (see below). Effects of HEPES buffer on the p[cu] measurement by DPCSV of catechol complexes were also examined in the presence and absence of the buffer at the same ph. ph was kept constant by adding small amounts of M HCl during measurement in the absence of HEPES. NO significant effect was found at the HEPES concentration used in the lake water. For p[cu] measurement in the absence of the buffer, the ph of lake water increased up to 0.5 units over time due to exposure to the atmosphere and purging with N2; these ph variations would obviously affect the results. We thus routinely used HEPES to ensure constant ph. p[cu] was measured by this method in a model solution containing 1 x 1 O-8 M CU, 2 X 10B7 M EDTA, 2 x lop3 M Ca, 0.01 M KN03, and 6 x 1O-3 M HEPES buffer, at ph 7.7. Catechol concentrations of l-5 x 1 O-5 M were used in different experiments. Measured

5 1204 Xue and Sigg 6 P a P 2-1: / J 0 I, I, 1, I I I t----)--t Potential Volt Fig. 1. Voltametric current signals of Cu-catechol complexes as a function of catechol concentration in terms of DPCSV current peaks and peak current (i,) vs. total catechol concentration. A. The current peaks (vs. potential) of Cu-catechol in Lake Greifen water collected at 5-m depth on 8 August 1990; 1, 2, 3, 4, 5, and 6 correspond to 5 x 10e4, 1 x 10-3, 2 x 10-3, 5 x 10-3, 8 x lo-), and 1 x 10 2 M catechol without addition of Cu. B. Lake Cristallina water collected on 7 July 1990 without addition of Cu-0; Lake Greifen water collected at 5-m depth on 21 March 1990 without addition of Cu-O; Lake Greifen water collected at 5-m depth on 21 March 1990 with addition of 1.6 x 1O--8 M Cu--o. p[cu] was O. 1, which compares well with the calculated value of The method was also applied to an algal nutrient medium of known composition and the results were compared with the calculated p[cu]. This medium contained 3.2 X 10e7 M Cu, 3.4 x 1O-6 M EDTA, 2.4 x 1O-4 M Ca, 3.0 x 1O-4 M Mg, 7.7 x 1O-7 M Zn, 2.2 x 1O-5 M Fe, and 3.1 x 1O-5 M citrate at ph 7.6. The calculated speciation gave p[cu] = 10.1, with most of the Cu complexed by EDTA; p[cu] = 9.98 was obtained experimentally by titration with Cu and determination of the catechol complexes at two different concentra- tions of catechol. The measured p[cu] was thus in good agreement with the calculated value. Results Total dissolved Cu (by DPASV) was between 5 x 1O-g M and 2.8 x 10e8 M. Particulate Cu (AAS measurements) was in the range < x 1 O-g M. AAS measurements of total dissolved Cu and TCu, which are less precise than the DPASV measurements in this range, gave similar values. The ratio of dissolved Cu to TCu was to ~0.95. The distribution coefficients Kd (ratio of solid phase concentration to dissolved phase concentration) can be

6 Cu speciation in a eutrophic lake 1205 Table 1. p[cu] measured with different concentrations of catechol (M) Lake Greifen 21 Mar 21 May 8 Aug 24 Ott Depth [W 1 (ml WW 5x p[cu], Catechol (M) 8X 10-d 1x x10 5 2x 10-s 5x x10 5 8~10.~ Lake Cristallina* 27 Jul * All data wcrc obtained by DPCSV in ph 7.5 solution bukcred with HEPES. Original ph of Lake Cristallina water was derived from measurements in the settling particles (Sigg et al. in press) and are in the range lo-120 x lo3 liters kg-l, comparable with those for Lake Constance and Lake Zurich (Sigg 1987). Thus, dissolved Cu predominates in Lake Greifen, in which the suspended matter concentrations vary between 0.5 and 4 mg liter-l. The depth profiles of dissolved Cu during summer stratification showed generally higher Cu concentrations in the upper 10 m than in deeper waters. These differences can be explained by inputs from sewage, which flows directly into the upper water layers. The suitable catechol concentration must be chosen as a tradeoff between analytical sensitivity and minimizing interferences of catechol adsorption; 2.5 x 1 O-5 M catechol was recommended for determining complexation of Cu in seawater (Van den Berg 1984a), and 2 x 1 O-4 M was used in the Scheldt estuary (Van den Berg et al. 1987). Preliminary experiments showed, however, that there was no detectable current reduction for Lake Greifen water at such low concentrations of catechol, due to much stronger complexation by organic ligands. Figure 1 shows the DPCSV signals in Lake Greifen water at different catechol concentrations and the peak current as a function of the added concentration of catechol for different water samples. The peak current increased with added catechol and approached a maximum after the added catechol was higher than a certain concentration in water from both Lake Greifen and Lake Cristallina. However, the critical concentration for detectable current signals was dependent on the individual sample. The curve for Lake Cristallina water with organic C CO.5 mg liter-l was similar to that for seawater. Reproducible p[cu] (20.05) and good sensitivity (0.36 A M-l) were obtained in the catechol concentration range of 1 x 1O-5-1 x 1O-4 M for Lake Cristallina water. The curve for Lake Greifen water (DOC 3.5 mg liter - l) was shifted toward higher concentrations of catechol by more than two orders of magnitude compared to the curve for Lake Cristallina. The range of 5 x 1O-4-1 x 1 O-3 M catechol yielded acceptable sensitivity ( A M-l) and reproducible p[cu] (+O. 1 units) for Lake Greifen water (Table 1). Therefore, a concentration of 1 x 1O-3 M catechol was chosen for determining Cu complexation in Lake Greifen water. Replications with the same concentrations of 1 x 1O-3 M catechol gave errors in the peak current of ~20% at 1 O-lo A (ambient [Cu] in Lake Greifen water) and < 5% at 1 O-8 A, corresponding to errors of 40.1 units for pcu. Figure 2 shows an example of the current peaks increasing with addition of Cu and two titrations of a Lake Greifen water with Cu. DPCSV sensitivities were extracted from the portion of titration curves at high [Cu2+]; the same sensitivity was obtained for the two titrations with different concentrations of catechol (Fig. 2B). The sensitivities ranged between 0.25 and 0.40 A M-l for all samples examined. The titration curve with Cu (Fig. 2) gives the concentration of Cu-catechol complexes

7 1206 Xue and Sigg a c.a I I I I I I I I I I Potential Volt [CuIT (M) lo Fig. 2. DPCSV current peaks (vs. potential) increasing with addition of Cu, and titration curves with Cu, in terms of peak current (i,) as functions of [Cu], for Lake Greifen water collected at 5-m depth on 8 August A. The current peak without addition of Cu-l; with addition of 0.79 x lo-*, 1.26 x lo-*, 1.57 x lo-*, and 3.15 x lo-* M Cu, in the presence of 1 x 1O-3 M catechol-2, 3, 4, and 5. B. i, measured with 5 x 1O-4 M catechol-0; i,, measured with 1 x 1O-3 M catechol-0. The slopes of the lines at high concentrations of Cu, where the stronger organic ligands were saturated, express the sensitivities of DPCSV measurements for the corresponding water. (from the peak current) as a function of total added Cu; [Cu2+] was calculated for each point from Eq. 5. The ambient [Cu2+] was calculated with Eq. 8 and 9 from the reduction current and total dissolved Cu in the original sample, where ain = 57 was used, as calculated from the major ion concentrations, with pi corrected for an ionic strength of 0.01 M (Martell and Smith 1974). ain is negligible compared to acal (1 X lo -1 X 107) or 2 Ki[Li]. Table 2 gives the [Cu2+] of the samples collected at different dates and depths in Lake Greifen. It is surprising that [Cu +] values are so low compared to available values for surface water in the open sea. The average p[cu] at 5 m is , with exception of the two win- ter samples; [Cu2+] values are, thus, two orders of magnitude lower than most published values for surface seawater (Table 3). Only the higher p[cu] ( ) for the Scheldt (Van den Berg et al. 1987) and Tamar estuaries ( , Van den Berg et al. 1990), which were also determined by DPCSV, overlie the range for Lake Greifen. The data show that [Cu2+] values are 6-7 orders of magnitude lower than the dissolved Cu values. Thus, almost all of the ambient dissolved Cu in the lake is complexed with organic ligands. Figure 3 exemplifies titration curves in terms of [Cu2+] as a function of [Cu],. The relationship between log[cu2+] and log[cu], is illustrated for several sampling dates in Fig. 4. There

8 Cu speciation in a eutrophic lake 1207 Table 2. [Cu],, [CL+], p[cu], conditional stability constants, and ligand concentrations of Lake Greifen water. (Conditional stability constants and ligand concentrations calculated by FITEQL program.) 1990/1991 Depth [WT [Cue ] (m) (nw (lo-i5 M) PKM log K, log Kz (2) 21 Mar Apr May Aug Ott Jan * :7* * Obtained at ph 7.8, close to the original ph of the lake water at that time, but the constant and ligand concentration were calculated on the basis of the results determined at ph 8 to comparc with other samples. is an obvious shift between the two curves trations. Most data series could be converged represented in Fig. 4A, which indicates a dif- with this optimization process. The computed ference in the complexing characteristics of the conditional complexation constants and lilake water between the two samplings in April gand concentrations are also listed in Table 2. and May; the other points at different sampling dates also reflect changes in complexing char- 500 acteristics. The two samplings in August and I - I October (Fig. 4B) differ in the lower part of the xl O-l4-0 I curve (corresponding to the stronger ligands) and overlap in the upper part (weaker ligands) The direct results of these titration experiments are values of p[cu] and of ZZ Ki[L,] (Eq. s 2 and 9)-the product of the stability constants and the ligand concentrations. These param- r eters can be estimated with reasonable preci- cu=r sion: +O.l log units for p[cu] and +0.2 for 2 log@ Ki[Li]). It is however much more difficult to extract individual conditional stability constants and ligand concentrations from Z Ki[Li] The complexing characteristics of these samples are certainly caused by a complex mixture of different ligands. Approximation of this complex system by fitting to a simple two- xl O-8 ligand model represents a simplification, which [CUIT (M) allows us to evaluate the range of stability constants and of ligand concentrations that may be present; two ligands represent the minimal number required to fit the data. The FITEQL program (Westall 1982), with a two-ligand model, was used to estimate conditional stability constants and ligand concen- Fig. 3. Titration curves of Lake Greifen water with Cu(I1) in terms of [Cuz+] as a function of [Cu],. The samples were collected at 5- (A) and 30-m (0) depths on 8 August [Cu2+] was calculated for each titration point in the absence of catechol on the basis of ligand exchange between the added catechol (1 x 1O-3 M) and natural organic ligands. Insert shows an enlarged view of the lowest concentration range.

9 Sunda and Hanson FL! 1987 % Sunda and Ferguson 1983 Sunda and Ferguson 1983 Van den Berg 1984b Van den Berg et al Van den Berg et al Table 3. Comparison of ambient p[cu], ligand concentrations (nm), and conditional stability constants. Sampling location PH P[C4 CL a log K log Kz Technique References Lake Greifen* Shelf water off North Carolina Lower Newport River estuary Sargasso Sea North Pacific ? l l OkO &0.4 CSV of Cu-catechol complexes Chemiluminescence and ligand competition with EDTA Chemiluminescence and ligand competition with EDTA Ligand exchangeiliquidliquid partition DPASV This study Sunda and Huntsman 1991 Sunda and Huntsman 1991 Moffett et al ti fz Coale and B&and a 1988, 1990 & Coast of Peru Southeastern Gulf of Mexico Mississippi River plume South Atlantic Scheldt estuary Tamar estuary l Bioassay SEP-PAWligand competition with EDTA Bioassay CSV of Cu-catechol complexes CSV of Cu-catechol complexes CSV of Cu-catechol complexes

10 -11 Cu speciation in a eutrophic lake 1209 I I I I i B i I I cl - d A og [CUIT log [CUIT Fig. 4. Titration curves of Lake Greifen water with Cu(I1) in terms of log[cu*-+] vs. log[cu],. The samples in panel A were collected at 5-m depth on 23 April (0) and 21 May (0) 1990; those in panel B were collected on 8 August (0) and 24 October (0) 1990 and 9 January 1991 (Cl). Points representing log[cu*+] vs. log[cu], in the original water on the other sampling dates are plotted in panel A for comparison [21 March (A), 8 August (Cl), 24 October m, and 9 January (A)]. The curves fitted to the water data were computed from the conditional stability constants and ligand concentrations for strong and weak organic ligands listed in Table 2. B -6.0 Lo@, values for the stronger ligands range between 13.9 and 14.9 (avg 14.3) and log Kz for the weaker ligands between 11.8 and 12.9 (avg 12.3). These values must be considered as estimates of the order of magnitude of the stability constants involved, which are also influenced by the data-fitting procedure and the assumption of a two-ligand model. The total concentrations are nm for the stronger ligands (CL,) and nm for the weaker ligands (CL,). The titration data could be fitted well with these two ligands, as shown by the solid curves in Fig. 4A and B. These curves were calculated with the stability constants and ligand concentrations given in Table 2 using the MICROQL program (Westall 1979). Discussion The reliability of the method under these lake-water conditions must be discussed first; the exchange kinetics with the natural organic ligands, the effect of different catechol concentrations, and competitive effects must all be considered. Equilibrium of ligand exchange between added catechol and natural ligands is required to determine [Cu2+] values by CSV of catechol complexes. If equilibrium was not attained, Z[Cu(cat)J would be underestimated, and consequently [Cu2+] would be underestimated. According to the kinetics of ligand exchange, the rate of the reaction depends on the catechol concentration; the high concentration of cat-

11 1210 Xue and Sigg echo1 used should promote rapid equilibrium. The kinetic experiment indicated that equilibrium of the ligand exchange was reached within a few minutes. Reliable results for p[cu] in the original water sample should be independent of the catechol concentration used, if equilibrium is reached for ligand exchange between added catechol and natural ligands. Different concentrations of catechol were thus used, but the applicable range was limited in these samples (Fig. 1B). For Lake Cristallina water, catechol concentrations between 1 x 10H5 M and 1 x 10m4 M, close to that recommended by Van den Berg (1984a) for seawater, were optimum for determining [Cu2+]. For Lake Greifen water, the optimal catechol concentration range was between 5 x 1O-4 and 1 x 1O-3 M, and the measured [Cu2+] decreased at catechol concentrations > 2 x 10m3 M. The data listed in Table 1 show that reproducible results (_+ 0.1 p[cu] units) were obtained within the applicable range. High concentrations of catechol could introduce inaccuracies in DPCSV measurement due to adsorption problems at the electrode surface. The use of ligand competition could also introduce biases in determining [Cu2+]. In DPCSV technique, reaction between catechol and metal ions other than Cu can decrease the free catechol concentration, resulting in low values for [Cu2+]. Calculations for our water conditions show that this effect was not significant. Another problem is competition between Cu and other metal ions in complexing with natural organic ligands. If it occurs significantly, then the extent of Cu complexation, and therefore [Cu2+], will be affected by the free ion concentration of the other metals. If the added catechol also complexes these metals, the concentration of these metals bound to natural organic ligands will decrease. The resulting increase in free natural organic ligands would lead to underestimates of [Cu2+]. However, Cu generally forms stronger organic complexes than do other divalent trace metals. In addi- tion, if such competitive effects are significant, then the measured [Cu2+] should change with the concentration of added catechol. Within the given range of catechol concentration in our experiments, the values of [Cu2+] are independent of catechol concentration (Table 1). Therefore the above-mentioned biases were not a problem. Determination of [Cu], in Lake Greifen water by DPCSV of catechol complexes appeared to be impossible due to strong complexation by organic ligands, unlike the situation in seawater. Van den Berg (1984b) used a high catechol concentration (8 x 1 O-4 M) to overcome interference from natural organic ligands by determining dissolved Cu concentration by standard additions. For the Lake Greifen samples, a quite high concentration of catechol (1 x 10e3 M) could not completely outcompete the natural organic ligands for complexing Cu, resulting in too low a measured concentration of dissolved Cu compared to the value measured by DPASV or AAS. DPCSV measurement with a much higher concentration of catechol would have introduced interferences in DPCSV reduction current as discussed above, and therefore, was not feasible. Even in Lake Cristallina water that had a much lower concentration of natural organic ligands, the dissolved Cu concentration measured by DPCSV was lower than those obtained from DPASV or AAS. Determination of total dissolved Cu by DPCSV must thus be used with caution. An average p[cu] (15.3) determined in samples from Lake Greifen is 2-3 orders of magnitude higher than available values for surface water in the open sea as shown in Table 3. This table also gives conditional stability constants and ligand concentrations for different waters determined by different techniques for comparison. The very low values of [CU~-~] in Lake Greifen appear to be related to high concentrations of strong ligands. The conditional stability constants for these strong ligands are also higher than most values determined in surface seawater. The stability constants in Lake Greifen are in the same range as those for the Scheldt estuary (Van den Berg et al. 1987), while the ligand concentrations are somewhat higher. These stability constants and ligand concentrations are however, as mentioned above, dependent on the methods used, especially for data fitting. It must be realized that natural waters contain a wide range of different ligands with different stability constants. The values obtained from a simplified two-ligand model represent an average of different ligands with different stability constants. The results ob-

12 . Cu speciation in a eutrophic lake 1211 tained also depend, in the case of ligand exchange methods, on the concentration of added ligand and [Cu], used in their determination (Van den Berg et al. 1990). With the catechol ligand exchange technique, the ligands that can be detected depend on the value of a,,1 (Eq. 7); a cat must be of the same order of magnitude as x KJL,]. Using other techniques, such as titration with Cu at higher concentration levels, may give very different results, since different ligands with a range of stability constants are present in natural waters. Depending on the concentration range of Cu used in a particular method, different kinds of ligands are detected; the strongest ligands are detected at low concentrations-close to the natural level. The very low concentrations of Cu2+ and very high complexation stability constant indicate that there must be natural ligands with strong complexation properties in the lake. What kinds of organic compounds can provide such strong complexation and control Cu2+ at such low levels? The synthetic ligands EDTA and NTA (with stability constants log K = 18.8 for CuEDTA and log K = 13.0 for CuNTA) are known to be present in Lake Greifen at concentrations of1 x lo-*medtaand0.5 x 10-8MNTA (Ulrich 1991). However, the competition of 1 x 10e3 M Ca at ph 8 results in a log conditional stability constant for CuEDTA of 11, which is much less than those for natural organic ligands. EDTA and NTA are, thus, not significant in complexing Cu in the lake. Organic ligands of known structure (such as EDTA and similar chelating ligands) have stability constants with Cu in the range of log K = 12-l 8. The stability constants estimated in the lake waters are however conditional for ph 8 and Ca - l-2 x 10B3 M; this means that these ligands must have a high selectivity for Cu over Ca. Examples of such ligands can be found, e.g. ethylenediiminodibutanedioic acid (EDDS), which has a stability constant with Cu log K = 18.4 (Mar-tell and Smith 1974); a conditional stability constant in the presence of Ca is calculated as log K = Natural ligands may achieve a similar selectivity with chelating structures. Sulfidic groups in the organic material may contribute to the strong complexation of Cu. The natural organic ligands that strongly complex Cu are probably either directly or indirectly (e.g. humic materials) produced by the biota. The occurrence of Cu chelators in algal cultures has been demonstrated by McKnight and Morel (1979), Van den Berg et al. (1979), and Zhou and Wangersky (1985, 1989). An earlier study on Cu binding by an algal exudate from a green alga gave the conditional stability constant and its ph dependence (Xue and Sigg 1990). On the basis of these values, the stability constant at ph 8 would be > 1013, which is close to those we estimated for the strong ligands in Lake Greifen water. This comparison implies that algal production of extracellular ligands is a potential source of strong Cucomplexing ligands in the lake. We have found that proteins containing sulfidic groups were released by a species of algae (Chlamydomonas) in response to added Cu (unpubl. work). Analysis of organic material (Hollander 1989) showed that organic compounds in Lake Greifen originate mainly from lacustrine algae. Studies on the molecular weight distribution and analytical fractionation of dissolved organic matter in the lake (Gloor et al ; Schneider et al. 1984) indicate that the structural and molecular weight distribution of the organic material varies seasonally and that these variations parallel those of algal exudates during algal blooms in summer. Algal blooms in Lake Greifen occurred in March-April and August-October 1990, as shown in Fig. 5 (H. Ambiihl pers. comm.), which plots seasonal variations of assimilated 14C, chlorophyll (avg value from 0 to 5 m, data from the Limnology Department, EAWAG), p[cu], and [CU], : [Cu2+] at 5-m depth. [Cu], : [Cu2+J is equivalent to the complexing coefficient for organic ligands x KJL,] (Eq. 9) when the terms (1 + ain) are negligible on the right side of Eq. 9; it indicates the extent of complexation independent of [CU],. p[cu] and log([cu],/[cu2+]) exhibit the same pattern as the variation of chlorophyll over time. Both have peaks in March-April, minima in May, and low values in fall-winter 1990-l A side-by-side comparison of titration curves for log[cu2+] vs. log[cu], also shows stronger complexation in the April sample compared to the May sample (Fig. 4A), as indicated by the shift in the curve along the x-axis (Sunda and Huntsman 199 1). A shift between the titration curves of 8 August, 24 October, and 9

13 1212 Xue and Sigg Fig. 5. Variations of chlorophyll, assimilated 14C, p[cu], and log([cu],,/[cu*+]) over time in Chlorophyll and assimilated 14C represent averages from the values of O- 5-m depth; p[cu] and log([cu],/[cu*+]) are the measured values at 5-m depth. [Cu], : [Cu*-+] is equivalent to Z K,[L,] (Eq. 9). Points do not align on the time axis due to different sampling dates. January is also seen (Fig. 4B); lake water sampled in January has comparatively the lowest complexing ability. p[cu] in January is almost two orders of magnitude less than the peak value in April, indicating greater variations in p[cu] than in [Cu],. Particulate P as a planktonic indicator also appeared to correlate with complexation characteristics. Particulate P concentrations in the surface water in March and April ( pg liter-*) were higher than in May (12.1 pg liter-l) or January (7.1 pug liter- ). The above facts provide some evidence that biologically produced organic ligands play an important role in Cu complexation. These findings in a eutrophic lake can be compared with studies of Cu complexation in the oceans. The typical p[cu] levels found in the oceans (- 12-l 3, Table 3) are often attributed to the occurrence of biologically produced ligands (Sunda 1990; Moffett et al. 1990). Sunda (1990) suggested that an optimum level of p[cu] for phytoplankton in the oceans can be established by this mechanism, so that [Cu2+] is neither at a toxic nor at a growth-limiting level. In a study of Cu complexation in the Sargasso Sea, maximum Cu complexation was reported in the region of the chlorophyll maximum (Moffett et al. 1990). The distribution of Cu ligands in the North Pacific (Coale and Bruland 1990) also suggests a biological source. Similar mechanisms may be acting in lakes. Our results in Lake Greifen indicate that even lower [Cu2+] may exist in lakes than in the oceans; the optimum level for the freshwater biota, however, is not known. This very low [Cu2+] in Lake Greifen appears to be related to the very high productivity of phytoplankton in this eutrophic lake. Comparison with other lakes with different trophic states and biological productivity would show whether this is typical. These processes provide examples of the profound influence that aquatic organisms can exert on the distribution and chemistry of trace metals in natural waters by production of extracellular organic ligands, as well as by intracellular uptake and adsorption onto cell surfaces. In turn, interactions of metal ions with biological molecules may have important effects on the growth and physiology of organisms (Sunda 1990). Conclusions Our experiments indicate that ligand exchange with catechol and CSV of the Cu-catechol complexes can be used to measure ambient [Cu2+] in a eutrophic lake. The measured values of p[cu] at different times and depths are in the range of l 6.0; [Cu2+] values are thus orders of magnitude lower than those of [CU],. Most of the dissolved CU is, thus, complexed with organic ligands. The results indicate the presence of high concentrations of very strong ligands. We suggest that these organic ligands are biologically produced and that the very low [Cu2+] is thus related to high productivity of phytoplankton in eutrophic Lake Greifen. References ANDERSON, D. M., J. S. LIVELY, AND R. F. VACCARO Copper complexation during spring phytoplankton blooms in coastal waters. J. Mar. Res. 4: ANDERSON, M. A., AND F. M. M. MOREL The influence of aqueous iron chemistry on the uptake of iron by the coastal diatom Thalassiosira weissflogii. Limnol. Oceanogr. 27: BRAND, L. E., W. G. SUNDA, AND R. R. L. GUILLARD Reduction of marine phytoplankton reproduction rates by copper and cadmium. J. Exp. Mar. Biol. Ecol. 96: COALE, K. H., AND K. W. BRULAND Copper com-

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