995

Limnol. Oceanogr., 50(3), 2005, 995–1010q 2005, by the American Society of Limnology and Oceanography, Inc.

Lead and copper speciation in remote mountain lakes

Anja Ploger, Eric Fischer, Hans-Peter Nirmaier, Luis M. Laglera, Damiano Monticelli,2 andConstant M.G. van den Berg1

Department of Earth and Ocean Sciences, University of Liverpool, Liverpool L69 3GP, United Kingdom

Abstract

We determined the chemical speciation of lead and copper in remote mountain and high-latitude lakes atdifferent times of the year, providing background data for regions in Europe least affected by anthropogeniceffects. The lakes are characterized by low ionic strength, clear waters, and oligotrophic conditions; are fedpredominantly by atmospheric precipitation; and are ice-covered during a large part of the year. Lead- andcopper-complexing ligands were determined by cathodic stripping voltammetry with ligand competition usingCalcein blue and salicylaldoxime, respectively. Water column averages of the dissolved lead concentrationsranged from 0.2 to 0.9 nmol L21, with generally lowest levels during the prolonged winter ice cover, whereasdissolved copper concentrations, varying from 1 to 1.7 (60.1) nmol L21, showed no significant seasonalvariations. The lead speciation was controlled by complexation with ligands at concentrations mostly below 2nmol L21, with values for the stability constant (log K ) of 12.5–13.7; calculated ionic lead concentrations9PbL

(2log[Pb21] values) were 11.9–14.5. Copper complexation was controlled by ligands at concentrations of 12–21 nmol L21, with values for log K of 13.5–14.0 and 2log[Cu21] values of 14.4–15.1. The concentration of9CuL

the copper-binding ligands, but not those for lead, varied seasonally, with about 50% higher concentrationsduring open water conditions compared to periods of ice cover. The data were consistent with the presence ofonly one class of ligands for copper and lead. The stability constant of the ligands is similar to that of fulvicacid; however, evidence regarding the actual nature of the ligands is still lacking. The lake data show that (1)competition between calcium and lead causes a reduction in the stability constant of approximately onelog-unit for each order of magnitude of Ca21, and (2) lead scavenging in the lakes is moderated by complex-ation.

Copper and to a lesser extent lead are known to occurcomplexed with organic matter in the ocean (Hanson et al.1988; Capodaglio et al. 1990) and in lacustrine systems (Xueand Sigg 1993; Achterberg et al. 1997; Taillefert et al. 2000).These studies indicate the importance of complexation withorganic ligands for elements such as lead and copper. Al-though both copper and lead are involved in biological pro-cesses, only copper is an essential micronutrient that can beeither biolimiting or toxic to aquatic organisms dependingon concentration and chemical speciation (Brand et al. 1986;Morel et al. 1991).

Most studies of the distribution of trace metals or tracemetal speciation in freshwater have focused on lowland oreutrophic lakes or both, often with an emphasis on the redox-

1 Corresponding author (Vandenberg@liv.ac.uk).2 Present address: Dip. Scienze Chimiche e Ambientali, Univer-

sita dell’Insubria, Via Valleggio 11, 22110 Como, Italy.

AcknowledgmentsLluıs Camarero and Marc Ventura assisted with sample collection

at Lake Redo and provided data for temperature, pH, and O2. Hans-Joerg Thies and Ulrike Nickus did the same for Lake GKS. LeifLien collected samples at Ovre Neadalsvatn. Evzen Stucklık wasthe site operator at Ladove and further assistance with sampling andlake data was given by Ota Strunecky, Martin Blazo, Linda Ned-balova, and Jan Turek. Bjoern Rosseland and Gunnar Raddum as-sisted with sampling of Lake Ferguson. Atte Korhola was the siteoperator for the Finnish lakes and further assistance with samplingand lake chemistry was given by Sanna Sorvari, Milla Rautio, LauraForsstrom, and Raino-Lars Albert.

This research was financially supported by the EMERGE projectof the EU (EVK1-CT-1999-00032).

driven cycling of elements at oxic–anoxic boundaries (Bal-istrieri et al. 1992a, 1994, 1995; Achterberg et al. 1997; Xueet al. 1997; Xue and Sunda 1997; Xue and Sigg 1999; Tail-lefert et al. 2000), whereas comparatively few studies havebeen carried out on relatively pristine lakes (e.g., Nojiri1985; Aldrich et al. 2001; Skjelkvale et al. 2001). Prelimi-nary data on a mountain lake suggest that lead is fully com-plexed by organic matter (Fischer and van den Berg 2001)and copper similarly complexed in an alpine lake (Averyt etal. 2004).

In this study we report on the dissolved organic speciationof lead and copper in remote alpine and subarctic lakes lo-cated across Europe. The lakes selected for this study arelocated at high altitude or high latitude (subarctic) lakes,situated above the tree line, without an inlet stream or withonly a minor or seasonal inlet stream. The lakes are unusualin that they receive all inputs either directly from atmospher-ic precipitation or indirectly via runoff from the catchment.Because of their remoteness they are not affected by localanthropogenic inputs and the catchments are relatively un-disturbed. Rainfall composition influences the chemistry ofthese lakes significantly, as the catchments tend to be small,with poorly developed soil layers or no soil at all and ig-neous rocks as the predominant bedrock, resulting in verylittle weathering (Catalan and Camarero 1992). Other char-acteristics of these lakes are generally a low primary pro-ductivity and low levels of organic matter, low nutrient con-centrations, and, for the mountain lakes, a rather high solarradiation, making them quite fragile with little bufferingagainst changes in for instance pH or metal ion concentra-tions.

996 Ploger et al.

Sampling and analytical methods

Study sites—The water columns of five lakes were studiedin detail, along with the surface waters of a further 19 lakesthat were sampled just once. All lakes were oligotrophic orultraoligotrophic, located above the local tree line and werenot subject to local sources of contamination. The catch-ments were relatively undisturbed and covered by poorlydeveloped soils or bare rocks. Basic morphological param-eters are summarized in Table 1.

Sampling—Lake Redo was studied most frequently (fourtimes), Ladove was sampled twice, and the other lakes weresampled once. Samples were taken from the water columnin the deepest part of each lake, in winter through a hole inthe ice and in summer using a rubber or aluminum boat. Thesamples were collected either via a length of silicone tubingattached to a battery-powered peristaltic pump and loweredto the appropriate depth (Redo, Gossenkoellesee [GKS], La-dove Pleso, Lake Ferguson) or using a Niskin bottle (Redo1998 for depths greater than 10 m and all of Ovre Neadals-vatn). During 1997 and 1998 the samples from GKS andRedo were filtered in-line through 0.45-mm pore size filters(Millipore) into 500-ml low-density polyethelene (LDPE)bottles. Pump-collected samples were pressure-filtered in-line, whereas Niskin bottle-collected samples were suction-filtered within an hour of sampling (Lake Redo), or 2 d later(Ovre Neadalsvatn). During 2000/2001 water samples werefiltered in-line with a 0.2-mm pore size Sartorius filtrationcartridge fitted to the outflow of the pumped water, exceptwhen samples had been bottle collected (samples deeper than10 m and some of the Finnish lakes, as indicated in theresults).

The Finnish lakes were sampled in the middle of eachlake at 1 m depth, either using a peristaltic pump with sili-cone tubing and in-line cartridge filtration (0.2 mm), or witha Limnos water sampler followed by 0.45-mm filtration laterin the laboratory.

At each depth one unfiltered and two filtered samples weretaken for the analysis of dissolved metal concentrations (op-erationally defined as metals that pass through the respectivefilters) and metal speciation studies, respectively. The unfil-tered samples were used here as a check on possible samplecontamination during the filtration step and to obtain totalacid-soluble metal concentrations (not reported here). Sam-ples for metal determinations were acidified to pH 2.2 withpurified HCl and stored at room temperature. Samples forspeciation studies (0.2- or 0.45-mm filtered) were frozen assoon as possible and stored at 2208C until analysis.

LDPE sample bottles were cleaned by soaking in hot de-tergent (24 h) followed by soaking in 5 mol L21 nitric acidand 1 mol L21 purified hydrochloric acid (1 week each), andstored filled with Milli-Q (MQ) water acidified to pH 2.2with purified hydrochloric acid.

Analytical methods—Instrumentation and reagents: Tracemetal concentrations and speciation were determined by vol-tammetry using a Metrohm VA 663 electrode stand con-nected via an IME-663 module to a computer-controlled po-tentiostat (m-Autolab, PSTAT 10 or PGSTAT 10, Eco

Chemie). Dissolved lead was measured by anodic strippingvoltammetry (ASV) with either a rotating, mercury-film,glassy carbon disk electrode, or a hanging mercury dropelectrode (HMDE). Dissolved copper and the speciation oflead and copper were determined by cathodic stripping vol-tammetry (CSV) using the HMDE as the working electrode.The reference electrode was double junction Ag/AgCl, 3 molL21 KCl, and the counterelectrode was a glassy carbon rod.

Water was purified by reverse osmosis (Milli-RO, Milli-pore) followed by ion exchange, giving MQ water. Hydro-chloric acid and ammonia were purified by sub-boiling dis-tillation using a quartz condenser. Standard metal solutionswere prepared by appropriate dilution of atomic-absorptionstandard solutions (Spectrosol, BDH) acidified to pH 2 withHCl. A stock solution of 0.01 mol L21 salicylaldoxime (SA)(BDH) was prepared in 0.1 mol L21 HCl. The SA was re-crystallized by addition of MQ water to a concentratedsolution in methanol acidified with HCl. The pH buffer con-taining 1 mol L21 HEPES (N-2-hydroxyethylpiperazine-N9-2-ethanesulfonic acid, Aristar grade) and 0.55 mol L21

NaOH was cleaned using Chelex resin. Preparation of allother reagents was as before (Fischer and van den Berg1999, 2001).

The concentration of fulvic acid was determined by fluo-rimetry: 389-nm excitation, 468-nm emission detection, 10-nm slit width, with calibration against aqueous solutions ofSuwannee River fulvic acid reference material (IR101F-1,International Humic Substances Society) on a Perkin ElmerSL5 luminescence spectrometer.

Determination of metal concentrations: All samples wereultraviolet (UV)-digested (45 min, using either a 125- or a600-W high-pressure mercury vapor lamp, in polytrifluo-rochloroethylene capped acid-cleaned silica tubes) to decom-pose organic substances prior to voltammetric metal analy-sis.

Lead in the 1997/1998 samples was determined by square-wave (SW) ASV using a rotating, glassy carbon disk elec-trode (RDE), in situ plated with mercury in the presence ofthiocyanate (Fischer and van den Berg 1999). The Ladovesamples from September 2000 were analyzed by 10 HzSWASV using an HMDE, preceded by 5-min deposition at20.9 V. In all other samples lead was determined by 50 HzSWASV using an RDE with a preplated mercury film: Theelectrode was conditioned daily in a 0.01 mol L21 ammo-nium acetate (NH4Ac) buffer solution by cycling the poten-tial 50 times between 20.8 and 10.8 V. The mercury filmwas preplated from a buffered (0.01 mol L21 NH4Ac) solu-tion containing 30 mmol L21 Hg after a 10-min purge withnitrogen, 10-min deposition at 20.4 V, followed by a scanfrom 20.3 to 0.0 V. Samples were partly neutralized withpurified ammonia and buffered with NH4Ac (final concen-tration 0.01 mol L21) at a pH of 4–4.5, and mercury wasadded to a final concentration of 3 mmol L21, followed by a5-min purge with nitrogen gas. The conditioning potentialwas 20.1 V, the deposition potential 21.5 V, conditioningtime 20 s, deposition time 10 min, equilibration time 10 s,and the scan was from 20.9 to 20.25 V. ASV scans obtainedby the latter method are shown in Fig. 1.

Copper was determined by CSV in the presence of 30

997Lead and copper speciation in lakes

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998 Ploger et al.

Fig. 1. Anodic stripping voltammograms for a Redo lake watersample with a preplated Hg film, pH 4.5, 5-min deposition at 21.5V: (A) original scan of the sample containing 0.26 nmol L21 Pb;(B) after addition of 0.5 nmol L21 Pb; (C) after addition of 1.5 nmolL21 Pb.

Fig. 2. Variation of log aPbCB as a function of calcium concen-trations (log scale). The straight line represents the linear regressionthrough the data.

mmol L21 SA (Campos and van den Berg 1994) and HEPESto buffer the pH at 7.8: deposition potential 21.1 V, depo-sition time 30 s, 10 s conditioning at 20.1 V, and scans bySWCSV (10 Hz) from 0 to 20.65 V.

Complexing ligand titrations: The samples were thawedovernight at 48C, and carefully swirled to ensure dissolutionof any particulates that might have formed by the freezingprocess. Complexing ligand titrations were carried out in 28-ml polystyrene tubes (Bibby, Sterilin) that were cleaned bysoaking in 1 mol L21 HCl, and then conditioned twice withtitrations prior to first use to minimize metal ion adsorptiononto the walls.

Lead speciation—Lead-complexing ligands were deter-mined by CSV in the presence of Calcein blue (CB) (8-[N,N-bis{carboxylmethyl}aminomethyl]-4-methylumbelliferone)(Fischer and van den Berg 2001). The CB concentration was500 nmol L21 and the pH was adjusted to 7 with TES (N-tris[hydroxymethyl]methyl-2-aminoethanosulfonic acid)buffer (0.01 mol L21 final concentration). The equilibrationwas overnight and the concentration range was usually be-tween 0 and 10 nmol L21 of added Pb but had to be adjustedfor some lakes to a narrower or wider range. The concen-tration of PbCB in equilibrium with the natural ligand(s) wasdetermined by differential-pulse CSV: adsorption potential20.15 V, adsorption time 90–120 s, pulse frequency 10 s21,pulse height 25 mV, step height 5 mV.

The stability constant of lead with CB was calibratedagainst ethylenediaminetetraacetic acid (EDTA) in UV-di-

gested MQ water at a range of calcium concentrations (2–150 mmol L21) similar to the range encountered in the sam-pled lakes. Therefore, EDTA additions were made to MQwater with added calcium, containing 500 nmol L21 CB, 0.01mol L21 TES, and 10 nmol L21 Pb, and equilibrated over-night before the labile lead concentration was determined.The ratio X, of the reduction current in the presence ofEDTA over that in the absence of EDTA was used to cal-culate values for aPbCB (van den Berg 1985):

a 5 [(a9 1 a )X 2 a9 ]/(1 2 X) (1)PbCB Pb PbEDTA Pb

where aPb is the a coefficient for inorganic complexation oflead and aPbEDTA is the a coefficient of the EDTA complexof lead that was calculated using an ion-pairing model fordifferent calcium concentrations (van den Berg pers.comm.). It can be seen in Fig. 2 that the values for aPbCB

decreased with increasing calcium concentrations. The re-sulting linear equation:

21log a 5 (20.96 6 0.05)log[Ca ] 2 (0.63 6 0.20)PbCB

was used to calculate aPbCB values for each lake accordingto its calcium concentration (here and elsewhere ‘‘6’’ indi-cates the standard deviation). As the EDTA calibrations wereperformed at the same CB concentrations as used in the ti-trations (500 nmol L21), the resultant aPbCB did not requirerecalculation for different CB concentrations. The data forlake GKS (Fischer and van den Berg 2001) were recalcu-lated here using the same aPbCB values as those used for thisstudy to maintain consistency.

Conditional stability constants were calculated by linear-ization of the titration data (Ruzic 1982; van den Berg 1982)using the following equation (Fischer and van den Berg2001):

21 21[Pb ] [Pb ] 15 1 (2)

[PbL] [C ] K9 [C ]L PbL L

999Lead and copper speciation in lakes

where K is the conditional stability constant for complex-9PbL

ation of lead with natural ligands (L):

[PbL]K9 5PbL 21[Pb ][L9]

where [L9] is the concentration of L not complexed by lead,[PbL] the concentration of lead complexed with L, and[Pb21] the free ionic concentration of lead. CL, the total li-gand concentration, is defined as the sum of [PbL] and [L9].A plot of [Pb21]/[PbL] as a function of [Pb21] is linear inthe presence of a single, dominant ligand (Ruzic 1982; vanden Berg 1982), with a slope equal to CL and a y-axis in-tercept equal to (K CL)21. Linearity of the plots for these9PbL

titrations, which went well beyond the ligand concentration,suggested that the lead complexation was dominated by asingle complexing ligand for the detection window investi-gated. The sensitivity was estimated from the linear part ofthe titration curve where L had been saturated with lead.

Copper speciation—Copper-complexing ligands were de-termined by CSV using ligand competition against SA(Campos and van den Berg 1994). A 100-ml sample aliquotwas transferred to a Teflon bottle to which HEPES buffer(0.01 mol L21, pH 7.8) and SA (5 mmol L21) were added.Ten-milliliter aliquots were pipetted into 10 polystyrenetubes with previously added copper to give a concentrationrange of 0 to 80 nmol L21 copper. Aliquots were allowed toequilibrate overnight with the reagents at room temperature.The concentration of copper bound by SA was determinedby CSV using an adsorption potential of 20.1 V and a de-position time of 60 s.

The sensitivity was calculated from the high end of thetitration data and was corrected for its underestimation be-cause of incomplete formation of CuL using an iterative pro-cedure similar to that used before (Turoczy and Sherwood1997). A first estimate of the sensitivity was obtained fromthe last few points (3–4) of the titration. This estimate wasused to calculate first estimates for labile and organicallycomplexed copper for the metal additions. The data werethen linearized using an equation as for lead; plots of [Cu21]/[CuL] as a function of [Cu21] were straight for all samplestested, suggesting that the organic copper complexation wasdominated by a single type of ligand. Therefore, a one-ligandmodel was used. The initial estimates for CL and K9 werethen used to recalculate [CuL] and [Culabile] and correct thesensitivity in an iterative manner. This method takes intoaccount the systematic error that would otherwise occur be-cause of incomplete saturation of the ligands in that part ofthe titration used to calculate the sensitivity. The correctedvalues for S were 5–20% greater than the initial value.

The value of aCuSA was calibrated against EDTA analo-gously to the method described above for lead, by additionof EDTA to the lake water samples.

Results and discussion

General water columns conditions—Background infor-mation on the water chemistry of the sampled lakes is sum-marized in Table 2. Most of the lakes had a low ionic

strength with conductivities typically #20 mS cm21, and alow alkalinity, usually below 100 mmol L21. Some of theFinnish lakes and Lake Ferguson in Greenland had greateralkalinities, up to 532 mmol L21, because of more alkalinebedrock. The concentrations of calcium and magnesiumroughly covaried with the alkalinity, indicating that carbon-ate from bedrock or glacial moraines was the main sourceof the alkalinity. Lowest as well as highest concentrationsoccurred in the subarctic/arctic lakes (i.e., Finland, Norway,and Greenland). Calcium constituted the major cation in thealpine lakes, which is important for metal speciation consid-erations, whereas for about half of the subarctic lakes sodiumwas the major cation because of the proximity to the sea.

The dominant anions in these lakes were bicarbonate andsulfate. The majority of the lakes had circumneutral pH val-ues, 6.2–7.6, with the exception of three Finnish lakes, whichhad lower pH values of 5.6–5.8. Because of relatively lowproductivities in these lakes (leading to very transparent wa-ters) and poorly developed soils and vegetation in the catch-ments, dissolved organic carbon (DOC) values tended to below, especially in the high mountain lakes, ranging from 0.3to 2.3 mg L21. Nitrate values were very low in the subarcticlakes in Norway and Finland (#1 mmol L21), whereas highervalues were found in Redo, GKS, and Ladove (11–17 mmolL21) as well as in Lake Ferguson (19 mmol L21). This isconsistent with the available atmospheric nitrate depositiondata: Nitrate concentrations in rainwater were 4 mmol L21 ata station close to Ovre Neadalsvatn, 11 mmol L21 at Redo,13 mmol L21 at GKS, and 18 mmol L21 at a site near LakeLadove (MOLAR Water Chemistry Group 1999).

Depth profiles of temperature, pH, and dissolved oxygenat the time of sampling are shown, where available, in Fig.3. As a general pattern, the lakes were ice-covered everywinter for 6 months or more, from around November/De-cember until the spring thaw in June/July. Most of the lakeswere dimictic and during the early summer the lakes wouldat first tend to mix and then restratify for a short period untilthe autumn overturn. During the open-water season, thelakes were fully oxygenated except for waters close to thesediments in deep lakes where a minor degree of oxygendepletion occurred in the deepest part of the hypolimnion,near the sediments. The depletion was more pronounced dur-ing the long period of ice cover because of the gradualbreakdown of organic matter.

The pH profiles were relatively constant with depth. Av-erage water column values can vary considerably, however;e.g., in September 2000 the pH of Ladove was about 7.3,whereas the following July it was only 6.1, caused by meltwater draining into the lake, which tends to cause a tem-porary decline in pH, alkalinity, calcium, and major ion con-centrations. The transparency of mountain lakes is known tobe very high, leading to deep penetration of UV light, whichis thought to be a cause for DOC breakdown to the bottomof the shallower lakes (Sommaruga and Psenner 1997).

Filtration effects—Initially (1997–1998) water sampleswere filtered in-line through 0.45-mm membrane filters,whereas subsequent samples were filtered where feasible us-ing a cartridge with a 0.2-mm cutoff and an internal 0.45-mm prefilter. This change was made as filtration through the

1000 Ploger et al.

Table 2. Overview of the chemical composition of the sampled lakes. Fulvic acid concentrations were determined by fluorimetry;conductivity, alkalinity, pH, Ca, Mg DOC and NO3 data are from the EMERGE database of the project lakes and from B. O. Rosseland(pers. comm.), not necessarily sampled in the same year or season as this study, which explains apparent discrepancies between DOC andfulvic acid concentrations for some lakes.

LakeConductivity

mS cm21

Alkalinitymmol L21 pH

CammolL21

MgmmolL21

DOCmg CL21

NO3

mmolL21

Fulvic Acidmg L21

Lake RedoGossenkollesee

1220

45100

6.76.9

3371

14

0.40.4

1113

0.20.1

Ovre NeadalsvatnLadove PlesoLake Ferguson

61471

1765

532

6.26.67.1

1744

168

41

117

0.7*0.34.6*

01719

0.70.23.3

Finland:MallajarviSaanajarviMasehjavri

6131,009StuorramohkiSalmijarvi

9401819

51313

45156102

985

1974

6.77.27.06.95.86.36.8

17853747

52232

8291612

,488

0.71.92.31.80.30.31.5

0100100

0.67.83.2

12.40.10.21.1

VuobmegasvarriKohpejavriRidniToskaljarviSomaslompoloKaskasjarvi

1919

84632

9

96150

33345145

39

6.86.86.57.67.26.5

455515

1127515

1612

49545

8

1.01.51.80.70.31.4

001000

1.91.50.20.70.52.6

JuovvatsohkakPorevarriMuoddajavriHarrejavriAatsajokisaivo

541384716

4186181400117

5.67.37.37.66.9

359582

12530

866586221

1.21.01.11.52.0

00000

1.31.61.33.14.3

Tarjjunjargajavri 5 4 5.6 7 8 2.2 0 3.9

membrane filters was much slower, requiring sometimes sev-eral filter changes, and was more prone to contaminationbecause of the handling difficulties. This variation in the useof a 0.2-mm cartridge and 0.45-mm membrane filters affectsonly the lead data, as all samples used for copper had beencartridge filtered. It is likely that the lead concentration incartridge-filtered samples is reduced because of more leadbeing removed by the smaller pore size of the cartridge fil-ters. Our measurements (not shown) indicated a drop of 10–20% (47 samples) in the dissolved lead concentration be-tween 0.45- and 0.1-mm membrane filters. The differencebetween the 0.45-mm membrane and 0.2-mm cartridge mustbe less and the 10–20% effect can be used as an upper limit.This ,10–20% effect is much less than the differences seenbetween lakes. We considered this effect on the data set asunimportant as it does not affect the basic findings and con-clusions. We have no evidence for a difference in the lead-complexing ligand concentrations caused by 0.2- or 0.45-mm filtration.

Concentration and distribution of dissolved lead and cop-per—Vertical profiles of dissolved lead and copper concen-trations are shown in Figs. 4 and 5, whereas dissolved leadconcentrations in the Finnish lakes are summarized in Table3, together with average water column concentrations for allother lakes.

Lead: Average dissolved lead concentrations in the watercolumn ranged from 0.23 6 0.02 nmol L21 (Redo, Novem-ber 2000) to 0.89 6 0.19 nmol L21 (GKS), whereas theFinnish lakes surface waters showed a range from 0.28 nmolL21 (Stuorramohki) to 0.73 nmol L21 (Tarjjunjargajavri). At-mospheric inputs were clearly the cause for enhanced surfacewater levels of Redo in May (due to the inflow of meltwater), September, and November, whereas the winter (Feb-ruary) data showed nearly constant lead concentrations inthe water column due to the lack of inputs. The lead con-centration in Ovre Neadalsvatn was similarly low and con-stant under ice cover (March 1998). The influence of thespring melt on the lead concentration was noticeable also atLadove (as in Redo) where average dissolved Pb concentra-tions increased from 0.32 6 0.16 nmol L21 in September2000 to 0.88 6 0.22 nmol L21 the following July when snowand ice in the catchment were melting. Higher average dis-solved lead concentrations in GKS and Ladove could be re-lated to shorter water residence times (compared to Redo)or higher atmospheric inputs (or both) compared to OvreNeadalsvatn in northwest Norway and Lake Ferguson inGreenland, which are both much further away from Euro-pean continental lead sources than GKS in Austria and La-dove in Slovakia.

Lead concentrations in the majority of the Finnish lakeswere ;0.3–0.4 nmol L21, which is not as low as expected

1001Lead and copper speciation in lakes

Fig. 3. Depth profiles of temperature, pH, and oxygen for Lake Redo, Gossenkoellesee (GKS), Ladove, and Lake Ferguson at the timeof sampling.

given that this region in northwest Finland is considered tobe one of the cleanest in Europe (Ruhling 1992). On theother hand, in some of the smaller and shallower lakes thelead concentration may have been close to that of the at-mospheric water input because of a relatively short residencetime of the water. The lead concentrations in these lakes arelower than in some of the more central European lakes, andthey are at the low end of a survey of ;150 Finnish head-water lakes, which had a range of ,0.14 to 9.6 nmol L21

lead in the surface waters with a median value of 1.6 nmolL21 (Tarvainen et al. 1997). Lead concentrations in less re-mote lakes tend to be greater. For instance, seasonal surfacewater concentrations in Lochnagar, a Scottish mountain lakein relative proximity to industrial areas of Edinburgh, variedfrom 2.5 to 5.7 nmol L21 (mean 4.0 nmol L21) (Yang et al.2002). Hall Lake dissolved lead concentrations ranged from0.01 up to 4.9 nmol L21 at the sulfide maximum (Balistrieriet al. 1994). Lower lead concentrations have been found too,such as 0.24 nmol L21 in Lake Mashu, a deep oligotrophiclake in Japan (Nojiri et al. 1985), 0.01–0.39 nmol L21 Pb inLake Sammamish (Balistrieri et al. 1995), and lead concen-trations of 0.02–0.05 nmol L21 in the Great Lakes (Superior,Erie, and Ontario) (Nriagu et al. 1996), which are close to

open ocean concentrations, e.g., 0.050–0.100 nmol L21 leadin the Northwest Atlantic(Wu and Boyle 1997). Clearly thelake water residence time is an important factor determiningthe residual concentration of a scavenged element like lead.

Copper: Average dissolved copper concentrations in thewater columns of Redo and Ladove ranged from 1.0 to 1.76 0.1 nmol L21. The copper concentration was constant inthe water columns of these two lakes. The lack of increasedlevels at the surface suggests that atmospheric inputs of cop-per were minor or that that they were balanced by inputsfrom runoff and lake mixing. There were no significant sea-sonal variations, implying a balance between inputs and re-moval of copper from the water column on which the longperiod of ice cover had no apparent effect. Relatively con-stant copper concentrations in the water column are a com-mon feature also in other lakes (Nriagu et al. 1996), signif-icant changes only occurring at oxic–anoxic boundaries(Balistrieri et al. 1992b, 1994), whereas in our study thewater columns of the lakes remained fully oxic. This con-stancy is in contrast to the decrease in the lead concentrationdue to scavenging when the lakes were ice-covered, sug-gesting that the copper concentration was stabilized and less

1002 Ploger et al.

Fig. 4. Depth profiles of dissolved Pb and Pb-complexing ligand and dissolved Cu and Cu-complexing ligand concentrations for LakeRedo.

affected by adsorptive removal. This is supported by the factthat filtered and unfiltered copper concentrations were almostthe same, so very little copper was bound to particulates(data not shown).

The copper concentrations found throughout the watercolumns of these lakes are low compared to some otherfreshwaters. Dissolved copper in Hall Lake ranged from 0.57nmol L21 in anoxic bottom waters up to 18 nmol L21 in oxicsurface waters (Balistrieri et al. 1994), and in Lake Sam-mamish from 1.8 nmol L21 in anoxic bottom waters to 8.8nmol L21 at the surface (Balistrieri et al. 1992b). Mediancopper concentrations were 4.7 nmol L21 (range 1.0–36nmol L21) in Finnish headwater lakes (Tarvainen et al.1997); 6.5–13 nmol L21 in Esthwaite Water (a lowland lake)(Achterberg et al. 1997); 2.4 to 32 nmol L21 (mean 11.8nmol L21) in surface waters of Lochnagar, a Scottish moun-tain lake (Yang et al. 2002); 12 nmol L21, 14 nmol L21, and13 nmol L21 in Lake Superior, Lake Erie and Lake Ontario,respectively (Nriagu et al. 1996); and 1.1 nmol L21 in LakeMashu (Nojiri et al. 1985).

Lead- and copper-complexing ligands: The distribution oflead- and copper-complexing ligands in the water columns

of the lakes is plotted in Figs. 4 and 5. Table 3 gives thedata for the Finnish lakes and average water column con-centrations, calculated free metal ion concentrations, andconditional stability constants. Example titrations for leadand copper and the corresponding linearization plots areshown in Figs. 6 and 7, respectively.

Lead-complexing ligands: Lead was found to be fullycomplexed by organic complexing ligands with the concen-tration of the complexing ligands always in excess of thelead concentration. The increase in response for lead uponsaturation of the ligands was marked (Fig. 6A), indicatingthat the organic complexation was much stronger than theinorganic complexation of lead. The linearized plots of thetitrations were straight, indicating that a single ligand dom-inated the speciation. Lead-complexing ligand concentra-tions were lower than the copper-complexing ligand concen-trations (see below), and were in most cases less than 2 nmolL21, especially in the mountain lakes. Concentrations weregreater in some subarctic lakes: Ovre Neadalsvatn (5.0 61.0 nmol L21), Saanajarvi (4.8 nmol L21), lake 613 (4.3 nmolL21), Kaskasjarvi (11.3 nmol L21), Porevarri (3.8 nmol L21),and Harrejavri (11.8 nmol L21). Although above the tree line,

1003Lead and copper speciation in lakes

Fig. 5. Depth profiles of dissolved Pb and Pb-complexing ligands for GKS, Ovre Neadalsvatn (note the different scale), Ladove, andFerguson as well as depth profiles of dissolved Cu and Cu-complexing ligand concentrations for Ladove.

these lakes were at lower altitude than the mountain lakes(GKS, Redo, and Ladove). It is possible that relatively con-tinuous soil cover and vegetation in the catchment of thesubarctic lakes had caused increased levels of humic mate-rials, which may have contributed to the ligands. Further-more, the Finnish lakes were more shallow, e.g., Kaskasjarviand Harrejavri had a maximum depth of just 2 m, so the 1-m ‘surface’ samples were just 1 m above the sediments;another difference was that there was much less precipitationin Lapland (;400 mm yr21 compared to .1,200 mm yr21

in the mountains and at Neadalsvatn).The average stability constant (log K ) in the water col-9PbL

umns of the lakes was between 12.5 6 0.2 and 13.7 6 0.5,and 13.5 6 0.7 in the Finnish lakes. The calculated degreeof organic complexation of lead, aPbL, was ;103, much great-er than that for the very weak inorganic complexation oflead (aPb 5 2). As a result, the free ionic lead concentrationswere lowered by the organic complexation to 2log[Pb21]values of 11.9 6 0.5 to 14.5 6 0.2 (water column averages)and of 11.9 to 15.4 (13.3 6 0.9) for the Finnish lakes.

The range of detected complex stabilities is large (Table3), especially in the Finnish lakes, suggesting the presenceof different complexing ligands or sites in the different lakes,or competition by cations other than lead, Ca21 for instance.

The lead-complexing ligand concentrations in the watercolumns of GKS and Ovre Neadalsvatn, and in Redo Feb-ruary 1998 and September 2001, were greater in the surfaceand bottom waters than at intermediate depths, whereas sys-tematic variations were absent in the other profiles. Surfacewater increases could be from atmospheric input or runoff,whereas bottom water increases could have a benthic source.Such inputs would only be apparent when they are associ-ated with a temporary stratification of the lake.

Stronger lead complexation has previously been linkedwith increasing DOC concentrations in Paul Lake (Taillefertet al. 2000). A DOC profile was only available for GKS (notshown) and insufficient data are available to compare trends.There were no systematic seasonal variations in the ligandconcentrations in the lakes. In Redo, highest lead-complex-ing capacities (1.5 6 0.2 nmol L21) occurred during the win-

1004 Ploger et al.

Tabl

e3.

Dis

solv

edle

adan

dco

pper

conc

entr

atio

ns,

com

plex

ing

liga

ndco

ncen

trat

ions

(CL),

2lo

g[P

b21],

2lo

g[C

u21],

and

stab

ilit

yco

nsta

nts

for

the

Fin

nish

lake

san

das

wat

erco

lum

nav

erag

esan

dst

anda

rdde

viat

ions

,fo

rth

eot

her

lake

s.

Lak

eD

ate

Pb

Pb

(nm

olL

21 )

CL

(nm

olL

21 )

2lo

g[P

b21]

log

9K

PbL

Cu

Cu

(nm

olL

21 )

CL

(nm

olL

21 )

2lo

g[C

u21]

log

9K

CaL

Red

oF

eb98

Nov

00*

May

01*

Sep

01*

0.35

60.

070.

236

0.02

0.45

60.

190.

546

0.26

1.5

60.

20.

76

0.2

0.8

60.

10.

96

0.3

13.3

60.

112

.86

0.2

12.7

60.

412

.36

0.5

12.7

60.

112

.66

0.3

12.8

60.

312

.76

0.5

1.4

60.

31.

46

0.2

1.7

60.

2

20.2

65.

413

.36

3.4

20.8

66.

9

15.1

60.

214

.86

0.2

14.8

60.

2

14.0

60.

213

.96

0.2

13.8

60.

3G

KS

Ovr

eN

eada

lsva

tnL

adov

eP

leso

Lak

eF

ergu

son

Jul

97†

Mar

98S

ep00

*Ju

l01

*A

ug01

0.89

60.

190.

256

0.09

0.32

60.

160.

886

0.23

0.45

60.

15

2.0

60.

45.

06

0.1

0.8

60.

11.

26

0.4

1.5

60.

2

13.4

60.

114

.56

0.2

12.9

60.

311

.96

0.5

13.2

60.

5

13.3

60.

313

.26

0.1

12.7

60.

212

.56

0.2

13.7

60.

5

1.2

60.

21.

06

0.1

18.5

64.

112

.26

6.1

15.0

60.

214

.46

0.3

13.8

60.

213

.56

0.5

Fin

land

:M

alla

jarv

iS

aana

jarv

iM

aseh

javr

i61

31,

009

Sep

01*

Sep

01*

Sep

01*

Sep

01*

Sep

01*

0.33

0.31

0.38

0.37

0.38

0.5

4.8

1.1

4.3

1.1

13.5

12.0

14.3

12.7

13.3

13.2

13.2

14.6

13.6

13.6

Stu

orra

moh

kiS

alm

ijar

viV

uobm

egas

varr

iK

ohpe

javr

iR

idni

Sep

01*

Sep

01*

Sep

01*

Sep

01*

Sep

01

0.28

0.39

0.52

0.38

0.31

0.6

0.6

0.7

0.6

1.2

13.7

13.4

13.3

13.8

14.3

13.8

13.2

12.9

13.5

14.2

Tosk

alja

rvi

Som

aslo

mpo

loK

aska

sjar

viJu

ovva

tsoh

kak

Por

evar

ri

Sep

01S

ep01

Sep

01S

ep01

Sep

01

0.42

0.31

0.41

0.46

0.46

1.1

1.6

11.3 0.9

3.8

13.2

13.5

12.2

12.9

11.9

13.0

12.6

13.6

12.9

13.0

Muo

ddaj

avri

Har

reja

vri

Aat

sajo

kisa

ivo

Tajj

unja

rgaj

avri

Sep

01S

ep01

Sep

01S

ep01

0.35

0.42

0.49

0.73

1.9

11.8 1.5

1.6

13.6

11.9

13.7

15.4

13.0

12.8

14.0

15.5

*T

hese

sam

ples

wer

efi

lter

edus

ing

a0.

2m

mca

rtri

dge;

the

othe

rsa

mpl

esus

ing

a0.

45m

mm

embr

ane

filt

er.

†D

ata

from

Fis

cher

and

van

den

Ber

g20

01.

1005Lead and copper speciation in lakes

Fig. 6. Pb-complexing ligand titration of a sample from LakeFerguson (120-s deposition at 20.15 V, 0.5 mmol L21 CB, 0.01 molL21 TES, 0.49 nmol L21 total dissolved Pb): (A) peak current as afunction of total dissolved Pb; (B) linearization of the data.

Fig. 7. Results of a copper titration of a sample from Lake Redo(60-s deposition at 20.1 V, 5 mmol L21 SA, 0.01 mol L21 HEPES,1.1 nmol L21 total dissolved Cu): (A) peak height as a function oftotal copper; (B) linearization of the data.

ter (February 1998), whereas the other samplings, includingMay 2001 at the end of ice cover, yielded average [CL] rang-ing from 0.760.2 nmol L21 to 0.9 6 0.3 nmol L21. At La-dove the average lead-complexing capacity was higher inJuly 2001 (1.2 6 0.4 nmol L21) than in September 2000 (0.86 0.1 nmol L21).

The complex stability of the lead species in these moun-tain and subarctic lakes (Table 3) is high compared to otherwaters (Table 4). Competition by major cations (predomi-nantly calcium) is a factor of 103 greater in seawater, whichmay well explain the difference with those waters. Lowercomplex stability was obtained by ASV for lead species inPaul Lake (Taillefert et al. 2000): It will be interesting to

investigate whether the difference is due to the use of adifferent technique with a different detection window.

The K values are several orders of magnitude higher9PbL

than those for fulvic acid (Buffle et al. 1977); however, thisdoes not rule out humic and fulvic substances as the sourceof these ligands as it is known that large polyelectrolyticligands may well contain a large number of weak as well asa small number of much stronger complexation sites thatonly become apparent when determined at the low metal andligand concentrations in the lakes.

The ionic lead concentrations in the study lakes are mostlylower than in seawater (Table 4) in spite of its lower leadconcentration due to the high complex stability in the lakes.

1006 Ploger et al.

Tabl

e4.

Com

pari

son

ofdi

ssol

ved

lead

conc

entr

atio

ns,

2lo

g[P

b21],

lead

-com

plex

ing

liga

ndco

ncen

trat

ions

,an

dco

ndit

iona

lst

abil

ity

cons

tant

sfr

omth

eli

tera

ture

.

Sit

eM

etho

dN

atur

alpH

Ana

lyti

cal

pHP

b(n

mol

L2

1 )2

log[

Pb21

]C

Lx

(nm

olL

21 )

log

Kg

Ref

eren

ce

NE

Pac

ific

DPA

SV

— 8.1

0.01

7–0.

049

12.0

–13.

0(1

)0.

22–0

.56

(1)

9.4

–10.

0C

apod

agli

oet

al.

1990

Terr

aN

ova

Bay

(Ros

sS

ea)

DPA

SV

— —0.

025–

0.11

4—

(1)

0.25

–0.9

1(1

)9.

3–9.

9S

carp

oni

etal

.19

95

Sou

thS

anF

ranc

isco

Bay

DPA

SV

7.9

—0.

12–0

.20

12.5

(1)

0.89

60.

35(2

)12

.86

1.9

(1)

10.5

(2)

8.6

Koz

elka

etal

.19

97

Nar

raga

nset

tB

ayD

PAS

V7.

8–7.

9—

0.13

–0.3

211

.9–1

2.4

(1)

0.6–

1.0

(2)

5.1–

8.2

(1)

9.9

–10.

2(2

)8.

6–8.

8K

ozel

kaan

dB

rula

nd19

98

Pau

lL

ake*

DPA

SV

—6.

36

0.2

;0.

2–3.

7—

(1)

9.2

64

(1)

9.4

60.

8Ta

ille

fert

etal

.20

00

Ful

vic

acid

ISE

5.0

——

—4.

1B

uffi

eet

al.

1977

*B

elow

oxic

–ano

xic

tran

siti

on.

The Pb21 levels were lowest in the subarctic lakes becauseof greater stability of the PbL complexes: Average valuesfor 2log[Pb21] were 14.2 6 0.2 for Ovre Neadalsvatn, 13.26 0.5 for Lake Ferguson, and .13.5 for half of the Finnishlakes.

Copper-complexing ligands: Complex stability: Plots ofthe titration data according to Eq. 2 were straight, indicatingthat the speciation of copper was dominated by one class ofligand; this ligand was present in large excess to the dis-solved copper concentrations and at higher concentrationthan the lead binding ligand. Average water column concen-trations for [CL] at Redo and Ladove ranged from 12.2 66.1 to 20.8 6 6.9 nmol L21. The stability of the coppercomplexes was similar at Redo and Ladove, with averagewater column values for log K of 13.8 6 0.3 to 14.0 69CuL

0.2 for Redo, and 13.5 6 0.5 to 13.8 6 0.2 for Ladove.Ionic copper concentrations (2log[Cu21] values) were 14.46 0.3 to 15.1 6 0.2.

The average value for the conditional stability constantfor copper (log K 13.5–14.0) in these low-conductivity9CuL

lakes is about a log-unit less than that (15.8) in New Zealandalpine lakes (Averyt et al. 2004) at the same analytical pH(pH 7.8) but there is overlap in the data ranges. Althoughthe same voltammetric method was used, a higher detectionwindow (20 mmol L21 SA) was used in that work comparedto the 5 mmol L21 SA used here, so it is possible that stron-ger ligands were detected if these were present. On the otherhand, there is a large discrepancy between the calibratedcomplex stability of the CuSA complexes (aCuSA) that alonecould explain the difference in the calculated values forK : The constants from Campos and van den Berg (1994)9CuL

corrected to the ionic strength of the lakes (by using a sa-linity of 0.1) give a value for log aCuSA of 106.9 for 20 mmolL21 SA, whereas a value of 108.5 was used for the New Zea-land lakes study (Averyt et al. 2004). Averyt et al. (2004)recalibrated the constants for the CuSA species using a po-tentiometric method that could be based on activities ratherthan concentrations depending on the calibration method butthe method was not discussed in detail. They subsequentlyused a value of 200 for aCu (inorganic complexation of cop-per) to correct the constants to the Cu21 scale. According toour speciation model aCu should be between 2 and 3 for low-ionic-strength mountain lakes: Inorganic copper speciationis mostly controlled by carbonate ions, and the carbonateconcentration in these lakes tends to be very low (alkalinitiesbetween 4 and 400 mEq L21) and our measurements were atpH 7.8, explaining the low value for aCu, much lower thanthe value of 200 (Averyt et al. 2004) used for the recalibra-tion of aCuSA. It is possible that the high aCu value includesan activity coefficient but this is not indicated and the detailsof the method are apparently being published elsewhere (Av-eryt et al. 2004). The complex stability of CuSA and CuSA2

used in our work had been calibrated (Campos and van denBerg 1994) against EDTA using the same CSV method asused for the measurements of the samples, and is thereforeindependent of the value selected for aCu and does not re-quire correction using an activity coefficient.

It will be necessary to further evaluate the calibration ofthe CuSA complex stability in the future to come to an

1007Lead and copper speciation in lakes

Tabl

e5.

Com

pari

son

ofdi

ssol

ved

copp

erco

ncen

trat

ions

,2

log[

Cu21

],co

pper

-com

plex

ing

liga

ndco

ncen

trat

ions

,an

dco

ndit

iona

lst

abil

ity

cons

tant

sfr

omth

eli

tera

ture

.

Sit

eM

etho

dN

atur

alpH

Ana

lyti

cal

pHC

u(n

mol

L2

1 )2

log[

Cu21

]C

l x(n

mol

L2

1 )lo

gK

gR

efer

ence

Est

hwai

teW

ater

CS

Vof

Cu-

trop

olon

eco

mpl

exes

;6.

7–9.

97.

82.

8–13

.210

.9–1

4.1

(1)

6.8–

29.4

(1)

11.1

–14.

4A

chte

rber

get

al.

1997

Lak

eL

ucer

neC

SV

ofC

u-ca

tech

ol,

-oxi

ne,

&-

trop

olon

eco

mpl

exes

,an

dC

u-IS

E— 7.

89.

4414

.41

(1)

18.5

–19.

0(2

)10

0–1

10(3

)6.

0

(1)

14.3

(2)

10.8

–11.

0(3

)8.

6

Xue

and

Sun

da19

97

Lak

eG

reif

enC

SV

ofC

u-ca

tech

ol,

-oxi

ne,

&-

trop

olon

eco

mpl

exes

,an

dC

u-IS

E— 8.

012

.6–2

315

.16–

15.1

9(1

)25

.6–5

6.9

(2)

110

–350

(3)

10.0

(1)

14.6

–15.

8(2

)12

.5–1

3.2

(3)

8.6

Xue

and

Sun

da19

97

Lak

eS

empa

chC

SV

ofC

u-ca

tech

ol,

-oxi

ne,

&-

trop

olon

eco

mpl

exes

,an

dC

u-IS

E— 7.

913

.515

.11

(1)

19.0

(2)

100

(1)

15.5

(2)

12.6

Xue

and

Sun

da19

97H

umic

acid

Ion-

exch

ange

6.0

——

—(1

)6.

2–7.

0(2

)4.

9–5

.6Ta

gaet

al.

1991

Hum

icac

idA

SV

6.0

——

—(1

)6.

1(2

)5.

4F

ukus

him

aet

al.

1992

Hum

icac

idC

SV

ofC

u-ca

tech

olco

mpl

exes

— 7.8

——

(1)

30.0

(2)

200

(1)

12.5

(2)

10.3

Xue

and

Sig

g19

99F

ulvi

cac

idIS

E8.

0—

——

8.42

Buf

fie

etal

.19

77F

ulvi

cac

idC

SV

ofC

u-ca

tech

olco

mpl

exes

— 8.0

——

(1)

27.3

(2)

250

(1)

14.1

(2)

11.7

Xue

and

Sig

g19

99

agreement regarding the stability of the copper species inthese and other lakes.

Copper ligand distribution: The depth profiles of copper-complexing ligands in Redo in November 2000 and May2001 were concave (the ligand concentrations were higherat the surface and near the sediments), whereas the remain-ing profiles showed no systematic variation with depth. Thehigher surface and deepwater concentrations tended to mir-ror the same pattern seen for the lead-complexing ligands,albeit at a 103 greater concentration of the copper-com-plexing ligands. The average ligand concentrations for Redoand Ladove were higher during the open water season (La-dove 18.5 6 4.1 nmol L21, Redo 20.2 6 5.4 and 20.8 6 6.7nmol L21) than at the end of winter (Ladove 12.2 6 6.1nmol L21, Redo 13.4 6 3.4 nmol L21), whereas dissolvedcopper concentrations showed little variation. Similar to theligand concentrations, the DOC concentrations in the watercolumn of Redo were also lower during the winter, and low-est at the end of ice cover (Camarero et al. 1999), increasingagain as a result of algal activity during the summer. Highconcentrations of complexing ligands were probably causedby algal material in Esthwaite Water (Achterberg et al.1997), where the high ligand concentrations coincided witha phytoplankton bloom. A study of Lake Greifen has alsoprovided evidence for biological production of copper-com-plexing ligands (Xue and Sigg 1993). However, the moun-tain lakes studied here have very low ionic strength as wellas low nutrient concentrations, and productivity is low. It ispossible that the ligands in these lakes are from humic andfulvic materials, and not directly produced by algae.

These log K values and ligand concentrations are sim-9CuL

ilar to the L1 class of copper-complexing ligands found inother lakes (Table 5) despite the almost one order of mag-nitude higher dissolved copper concentrations and eutrophicnature of some of those lakes (e.g., Lakes Greifen and Sem-pach). So the lower productivity of Redo and Ladove andpotentially lower rates of release of organic ligands by phy-toplankton and other microorganisms is not reflected in low-er steady-state concentrations of copper-complexing ligandsof the L1 class.

The combination of high log K values and large excess9CuL

of ligands gives rise to low cupric ion concentrations in thelakes studied here. 2log[Cu21] values for lake Redo were14.4–15.6 (14.9 6 0.5) and for Ladove 13.7–15.3 (14.7 60.4). This strong degree of complexation may well be acause for the lack of seasonal variation in copper concentra-tions, as it would tend to keep the copper in solution.

Possible systematic method-related errors: To maintainconsistency of the data our measurements were at constantconcentration of added competing ligands (SA or CB) andat constant pH. In spite of this, the detection windows stillvaried between the lakes because of differences in major ionconcentrations (see below). It is in principle possible thatdifferent ligands are detected (if present) if the detectionwindow is greatly shifted (van den Berg and Donat 1992).The data treatment of the titrations with lead and copperindicated that only one ligand type dominated in each sam-ple. This is confirmed by the relative constancy of the com-

1008 Ploger et al.

Fig. 8. Stability of the lead complexes (log K ) as a function9PbL

of the calcium concentration (log[Ca9]) in the lakes.

plex stability within individual lakes when several depthswere studied. It is therefore likely that this study at leastfound the most important ligands in these lakes; however, itis possible that much weaker metal binding ligands (or othercomplexing sites on the same ligands) have not been de-tected.

It has been suggested that ligands could be present in thetype of buffers used for this study (Mash et al. 2003). Thepossibility of errors due to such ligands in the pH bufferswas eliminated by titration with copper of 0.01 mol L21 HE-PES buffer in UV-digested freshwater and with lead of 0.01mol L21 TES buffer in UV-digested MQ water. Both titra-tions were straight, indicating that ligands were not presentin the buffer solutions at significant levels. Any contributionof ligands from the pH buffers to the apparent ligand con-centration in these lake waters was therefore below the limitof detection.

Effects of major cation competition on metal complexa-tion: It is known that the pH tends to affect the stability ofcopper complexes with ligands in lakes (van den Berg andKramer 1979; Averyt et al. 2004); pH effects on fulvic acidshave been known of for a long time (e.g., Gamble 1970),and Ca21 is known to compete with lead for fulvic acidbinding (Pinheiro et al. 1999). The major ions likely to affectmetal complexation in the study lakes are H1 and Ca21. Ourmeasurements were at constant pH to simplify the measure-ments and to facilitate comparison, but the lake data have areasonable spread of calcium concentrations, which makesit possible to investigate the effect of calcium on the com-plex stability for lead (the copper speciation was determinedin only two lakes, so there is insufficient spread of calciumconcentrations). A plot of log K as a function of log[Ca21]9PbL

(Fig. 8) shows that log K decreases at a rate of 1.3 units9PbL

per order of magnitude of [Ca21] (in mmol L21):

log K 5 21.3 (60.4)log[Ca21] 1 15.1(60.7)9PbL

The plot has a slope close to unity, which would be ex-

pected for competition of a divalent metal ion as calcium forthe same binding site on the ligands as used by lead. As thiseffect occurred at calcium concentrations .10 mmol L21, thestability constant of the CaL species (K ) must have a value9CaL

of .105. There are no calcium complex stability data avail-able in the literature for ligands in natural waters, but thereare data for fulvic and humic acids, which may well be thesource of these ligands; these data are then sometimes ex-pressed on a weight basis as the molecular weight of thesecompounds is not accurately known, making comparison dif-ficult. Nevertheless, our data are consistent with literaturedata (Pinheiro et al. 1999) showing calcium binding by ful-vic acid at the same calcium concentrations.

Relation between lead complexation and the distributionof dissolved lead: The general shape of the lead profile inthe water columns of the lakes reflects its atmospheric inputfollowed by scavenging by adsorption on sinking particles(see for instance the profiles for lead in Redo [May andSeptember] and Ladove [September and July] and Neadals-vatn). The rate of scavenging along with the water residencetime determine the residence time of lead in the lake andtherefore its residual dissolved concentration. It is possiblethat the scavenging rate is moderated by the organic com-plexation. The general shape of the profile of lead is similarto that of the ligands in several of the plots in Fig. 4. Thelead and the ligands may to some extent have the samesource if they are both derived from a mixture of runoff andatmospheric inputs. However, their paths are likely to sep-arate unless there is an association that resists for instancethe removal of lead by scavenging. Plots of lead as a func-tion of the ligand concentration showed covariation betweenthe dissolved lead and the ligand concentrations (Fig. 9).Lines for 1 : 1 covariation have been drawn in the plots forcomparison, from which it can be seen that the actual slopestend to be less than unity, indicating that the complexationdoes not fully prevent the lead from being scavenged. How-ever, the covariation of lead with the ligand concentrationsuggests that the scavenging rate of dissolved lead is dimin-ished as a result of the organic complexation. This meansthat models attempting to calculate the residence time of leadin lakes should incorporate the change in scavenging bycomplexation reactions.

Source of the ligands: In principle the origin of the ligandscan be from humic or fulvic acids, such as from runoff, orproduced in situ by phytoplankton in the lakes. The produc-tivity of the majority of these low-ionic-strength lakes isvery low, and their transparency high (Sommaruga and Psen-ner 1997), which tends to be deleterious for phytoplankton(Sommaruga et al. 1997). It has been known for a long timethat humic and fulvic substances bind copper and other met-als (Buffle et al. 1977), and it is likely that these are themain source of the ligands in the mountain lakes (Xue andSigg 1999), rather than in situ production by phytoplankton,which may be a more important contributor to low-landlakes like Esthwaite Water (Achterberg et al. 1997). Thelarge difference in complexing ligand concentrations be-tween copper and lead is not necessarily a sign of differentligands: fulvic and humic acids tend to be relatively large

1009Lead and copper speciation in lakes

Fig. 9. (A) Concentration of dissolved lead as a function of theligand concentration for lead in Redo; (B) concentration of dis-solved lead as a function of the ligand concentration in Ladove andFerguson.

compounds that may well form weak as well as strong com-plexes depending on their configuration. The shape of thewater column profiles of the lead- and copper-complexingligands were similar, but at ;103 higher levels for copperthan for lead. This could suggest that different ligands bindthese two metals, but it is also possible that the same humicor fulvic acids bind copper and lead with a different numberof sites, or in differing configurations. The agreement incomplex stability of the copper species in the lakes in thisstudy and those in the New Zealand study (Averyt et al.2004) is also consistent with the same, possibly humic andfulvic, substances being responsible for complexation inthese waters.

Concentrations of fulvic acids in the mountain lakes canbe seen to constitute a major proportion of the DOC (Table2); calibration was against Suwannee River fulvic acid,

which may not be the correct model of the fulvic substancesin these lakes, and in some lakes the fulvic acid amountedto more than the DOC (the DOC values should be used asindicative values only as they were from a database on theselakes consisting of surface water samples collected at dif-ferent sampling times). The log K values of 13.5–149CuL

found here (Table 3) for what potentially are fulvic acids inthese lakes, are within the same range (12–14) as found forSuwannee River fulvic acid in 0.01 mol L21 KNO3 (Xue andSigg 1999). This similarity suggests that fulvic acids are areasonable model for the ligands in mountain lakes. How-ever, evidence regarding the actual nature of the ligands isstill lacking.

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Received: 6 May 2004Accepted: 16 December 2004

Amended: 18 January 2005