Cédric L Garnier
email: c.garnier@lptc.u-bordeaux1.fr

Lesven Ludovic

Billon Gabriel

Magnier Aurélie

Mikkelsen Øyvind

Pižeta Ivanka

L G Lesven

A Billon

Ø Magnier

I Mikkelsen

Pižeta

G Lesven

Ø Mikkelsen

Voltammetric procedure for trace metal analysis in polluted natural waters using homemade bare gold-disk microelectrodes

Keywords: Voltammetry, Bare gold microelectrode, Environmental monitoring, Trace metals

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Introduction

One of the new interests in environmental analytical chemistry is the development of in situ or at least in-field measurement and monitoring in the aquatic systems, including sediments, where the sampling and the porewater extraction is quite a difficult task because of the partial or total anoxia of the medium and the difficulty in preserving it in the sample until measurement. Up to now, only electroanalytical methods offer the ability to achieve this challenge for trace metals measurements. Research in this area has progressed quickly thanks to the parallel work of several groups. Wang and co-workers [1][2][3][4] investigated the behaviour of microelectrodes (carbon fiber and gold), bare or covered with mercury, for application in remote monitoring of trace lead, copper, mercury or selenium with potentiometric stripping voltammetry as the method of choice. Luther and co-workers demonstrated that the use of mercury-plated gold microelectrodes allows the in situ determination of, among others, Mn(II), Fe(II) and S(-II) in sediment porewaters [5][START_REF] Cai | MTTF (ed) Environmental electrochemistry: analyses of trace element biogeochemistry[END_REF][START_REF] Glazer | MTTF (ed) Environmental electrochemistry: analyses of trace element biogeochemistry[END_REF][8][9][10]. In the same period, Buffle and co-workers focussed their micro-captors on Ir microelectrodes also plated with mercury for Zn, Pb, Cd and Cu analysis [11][12][13][14][START_REF] Tercier-Waeber | Rozan MTTF (ed) Environmental electrochemistry: analyses of trace element biogeochemistry[END_REF]. In this case, a monitoring system is introduced in the water column, though the voltammetric measurements are not really performed in situ because of the necessity for removing oxygen from the sample before analysis. Another attempt with Ir microelectrodes with only traces of mercury was successful for Pb and Cd determination and measurements in box-core sediments [16]. Preparation for in-field measurements has been carried out on gold electrodes in acidic solutions [17,18], solid dental amalgam electrodes [19,20] and alloys [21]. Kirowa-Eisner's team [22][23][24][25][26][27] researched solid bare silver and gold electrodes of macro scale, contributing to a better understanding and application of underpotential deposition (UPD) effects in stripping voltammetry [28].

In this paper, a voltammetric study on 25-μm and 125-μm diameter gold-disk microelectrodes is presented for Cd, Cu, Hg, Mn, Pb and Zn determination with detection limits in the nmol L -1 range. A detailed calibration and stability study on these electrodes in seawater and polluted freshwater sampled in the River Deûle (northern France) has been performed, and an application to in-field analyses is presented and compared with inductively coupled plasma-mass spectrometry (ICP-MS) measurements. The same electrodes have been used in situ to analyse sediment concentrations of Mn and Pb.

Experimental

Equipment

In the laboratory, voltammetric measurements were carried out with a voltammetric analyser comprising a μAutolabII coupled to a VA663 stand (EcoChemie, Metrohm) controlled by the GPES 4.9 software. An Ag/AgCl/3 mol L -1 KCl electrode and a carbon rod were used as the reference and the auxiliary electrode, respectively. All potential values in this paper are given versus this Ag/AgCl reference electrode. A Teflon PFA cell (Metrohm) was used to minimize sample contaminations.

In the field, a pump (CO/TECH) is used, as well as two homemade Teflon cells, a homemade reference electrode (Ag/AgCl/3 mol L -1 KCl in agarose matrix) and a magnetic stirrer, Radiometer). A combined microelectrode (Radiometer) linked to a pH meter (Meterlab PHM210, Radiometer) is used to measure the sample pH. An ICP-MS (X Series, Thermo Elemental) was used in a clean room to measure the total dissolved metal concentrations.

Reagents

Standard solutions (100 μg L -1 , 1, 10 and 100 mg L -1 of Cd, Cu, Hg, Mn, Pb and Zn) were prepared by dilution of 1,000 mg L -1 metal solutions (CertiPur, Merck: nitrate salts in HNO 3 ) in Milli-Q water (Milli-Q Plus apparatus, Millipore, resistivity 18.2 MΩ cm). All other reagents (HNO 3 , KOH, C 2 H 5 OH) are of analytical grade. Gold wires used to design the microelectrodes (diameter 25-and 125-μm, purity 99.99 %) were obtained from Goodfellow Company (UK).

Voltammetric measurements were performed in marine and river water samples. Unpolluted seawater was sampled in the Authie Bay (northern France), filtered at 0.45-μm and stored in dark at 4 °C so as to minimize the development of organisms. On the other hand, river water was sampled in the River Deûle (North of France), just upstream from the former highly polluting metal industry factories. As the final goal of this study is to use the microelectrodes in situ at the water-sediment interface in aquatic river systems, this second sample was stored without any treatment in the dark at natural pH (i.e. 8.1) and at ambient temperature.

Dissolved trace metal concentrations in both natural samples were measured by ICP-MS after filtration and HNO 3 acidification.

Gold-disk microelectrode preparation

The protocol used for microelectrode preparation was based on that developed by Nyholm and Wikmark [29] and more recently by Billon and van den Berg [30].

Briefly, one extremity of a 4-to 5-cm gold wire (25-or 125-μm diameter) is attached to a shielded electrical copper wire using silver epoxy. The second extremity is heat-sealed in a polyethylene pipette tip (200 μL) by using a homemade tubular oven (diameter ca. 1 cm) at 500 °C. A second tip enclosing the electrical wire is stuck to the first one with a non-conductive epoxy glue, leaving out the copper wire contact.

An electroactive gold-disk surface is obtained by polishing the heat-sealed extremity with different sand papers and subsequently with diamond paste (diameter 1 μm) to a mirror finish. The surface of the gold microelectrode is checked after every polishing step with an optical microscope (Reichert, Austria) at magnification from 110× to 320×.

In order to clean the gold surface, the microelectrode was rinsed under ultrasound for 60 s, first in ultra-pure ethanol, then in a 10% ultrapur HNO 3 solution and finally in Milli-Q water. To validate all the preparation steps, the microelectrode response is checked by using cyclic voltammetry: 10 cycles from 0.65 V to -1.1 V (i.e. the working potential window for electrochemical measurements on our gold micro-surfaces) were performed in seawater at a rate of 0.1 V s -1 . The electrode response indicating a good manufacture is generally characterised by voltammograms displaying no hysteresis between the cycles and two oxygen reduction waves which must disappear after N 2 purging.

Voltammetric parameters

Trace metals analysis was carried out using either square wave or differential pulse anodic stripping voltammetry (SWASVor DPASV), depending on the signal stability (see "Results and discussion"). Samples were purged with N 2 before each measurement due to the high sensitivity of a gold electroactive surface towards oxygen.

The voltammetric parameters were first optimised to improve the sensitivity and the reproducibility of the measurements. These parameters are summarized in Table 1. Overall, a 120-s deposition time was chosen to obtain a good compromise between the detection limit and the duration of the experiment. In addition, before the accumulation step, a 30-s conditioning step at 0.65 V is used to clean the surface of the electrode, i.e. to keep it electroactive. The microelectrode was kept on a potential of 0.65 V between each measurement.

Results and discussion

Applications of in situ measurements and determinations of electroactive species of trace metals are envisaged either in seawater or in river water matrices and their corresponding sediments. As in situ determinations exclude standard addition methods, but rather rely on calibration lines, experiments for the characterisation and calibration of the homemade microelectrodes were classified by the matrix of the investigated solution in order to minimize the quantification errors.

Filtered seawater sample

Unlike spectroscopic techniques (e.g. ICP-MS, AAS and GFAAS) electrochemical analyses are not really affected by the complex seawater matrix, which even could simplify the metal speciation by formation of chloride-metal compounds. In addition, the high ionic strength (I≈0.5) coupled to the buffer capacity that is linked to the presence of hydrogeno-carbonates ensures stable conditions during measurements [START_REF] Buffle | Analytical chemistry. Ellis Horwood 32. Ružić I[END_REF]. Unlike the spectroscopic techniques which directly measure total metal concentrations, electro-chemical techniques can distinguish the labile metal fraction, sum of free and inorganically bound metal, which is more available for the biota than the metal fraction complexed by organic compounds. Moreover, as our aim is to use microelectrodes in natural systems, it appears more realistic to perform experiments in marine samples than in obviously simpler artificial ones. For all these reasons, the microelectrode suitability for trace elements analysis has been tested in filtered seawater sample.

The first step was to determine which elements could be detected by the prepared electrodes and to locate their corresponding peak potentials.

25-μm gold-disk microelectrode

Calibration Mn, Zn, Pb, Cd, Cu and Hg peaks obtained by SWASV with the 25-μm gold-disk microelectrode are shown in Fig. 1 for successive additions of standard solutions in filtered seawater at natural pH. These 6 trace metals could thus be analysed in natural waters with different sensitivities. In spite of very low current scale (peaks height lower than 0.01 nA for the first additions), for a 120-s deposition time without stirring the signal to noise ratio allows analysis of metal concentrations below 1 μg L -1 .

Compared to classical measurements carried out on mercury or liquid amalgam surfaces, all the peak potentials observed for the different metals deposited on the gold microelectrode are shifted to more positive values. In fact, the linear dependence of current on concentration of monolayer forming, with the effect of UPD, is used ( [28] and references cited therein). This shift allows easier analysis of Mn and Zn at natural pH. Moreover, the order of Pb and Cd oxidation peaks is swapped compared to mercury electrodes (which matches with the results of Bonfil et al. [27]), which may be due to different goldmetal affinities. These two peaks are quite close, and in the case of high concentrations of both metals, it could complicate their analysis. Zn is characterised by two successive peaks. The presence of two peaks for zinc can be explained by either a difference between the stability of the Au-Zn monolayer and the other subsequent Zn-Zn deposits, or from stepwise oxidation, or from quasireversible electrode reaction with two waves as is possible for certain values of α (electron transfer coefficient) [32], which could be a subject of another investigation.

To calibrate the microelectrode over a wide concentration range, additions of the 6 metals were carried out from 0.2 to 50 μg L -1 (to 100 μg L -1 for Mn) in a decade mode [33]. For each addition, the SWASV measurements were repeated twice. As the baseline did not significantly change with time during the experiments, mathematical treatment consisted in subtracting the blank curve (which here represents the average of 3 successive voltammograms obtained in seawater before any metal addition) from the obtained voltammograms. Subsequently, the peak height and/or area are determined using a linear baseline. Concerning Zn, the area of the sum of the two peaks and the height of the more positive one were determined. At low concentrations, due to Mn and Cu peaks shape, i.e. position in the steep baseline, it was not possible to correctly integrate and determine the area of these peaks, only their height. The obtained results, i.e. the slopes of the calibration lines representing the sensitivity of the measuring systems, are summarized in Table 2 (data for peak areas are not shown), and the calibration of Pb is shown as an example in Fig. 2. For each metal, the linearity range is determined by the detection limit (i.e. the lowest concentration leading to a measurable and repetitive peak) and the upper concentration limit, above which the microelectrode response is no longer linear, caused by the saturation of the electrode surface. The response of the 25-μm microelectrode for Hg (peak potential 0.42 V), although linear (R 2 0.995) from 0.5 to 20 μg Hg L -1 is not shown in this paper due to its difficult reproducibility. This phenomenon is certainly due to the formation of a strong Au amalgam [34], leading to a Hg coating of the gold surface that changes its reactivity, making it inappropriate for DPASV and SWASV techniques, in contrast to stripping chronopotentiometry [18].

The results indicate that the sensitivity when considering peak height (expressed in A L mol -1 ), decreases in the following order: Cu, Pb, Zn, Cd and Mn. Yet, except for Mn, all these trace metals are detectable at concentrations as low as 0.2 μg L -1 (1-3 nmol L -1 ) and up to 20-30 μg L -1 (150-460 nmol L -1 ) with linearity coefficient ranging between 0.996 and 0.999. Similar experiments were carried out in UV-digested seawater samples, but have not shown any relevant differences in voltammogram shape and peak evolutions (data not shown). This is probably due to the low concentrations of natural organic matter found in marine waters, and the quite high metal concentrations added, when compared to the concentration present at first, resulting in a low metal fraction in the form of inert organic compounds (i.e. a low metal complexing capacity).

Compared to previous results obtained with the mercury-plated gold wire and the disk electrodes (both of 25-μm diameter) [30], the 25-μm gold-disk microelectrode presented here showed a poorer detection limit for Pb than the gold/mercury-wire (0.07 nmol L -1 ) but a much better detection limit than the gold/mercury-disk (10 nmol L -1 ), in spite of a lower deposition time (120 s compared to 300 s). So, a bare gold surface seems to be more sensitive to trace metals than a similar one plated with mercury, while during the accumulation step there is no formation of amalgam and diffusion into the body of the electrode, which in turn might not be depleted in the stripping step. The comparison of Pb sensitivities normalized to the exposed electroactive surface area (gold disk 9.8 A L mol -1 mm -2 , gold/mercury-wire and disk 6 A L mol -1 mm -2 [30]) confirmed this assumption. On the other hand, the saturation during the accumulation step occurred at a lower concentration with the gold-disk microelectrode. This is due to the difference in the electroactive geometry, which is a volume in the mercury plated microelectrode and only a monolayer surface in the bare gold one. However, detection limits obtained with the bare gold microelectrodes are low enough to determine trace metal concentrations in heavily polluted aquatic systems like waters or interstitial waters from several rivers in northern France. Calibration of the microelectrode was also carried out by DPASV. DPASV measurement peaks, although less noisy than SWASV ones, are less sensitive and so lead to a more difficult determination of trace metal concentrations below 1 μg L -1 . Coupled with the higher duration of DPASV measurements, SWASV was chosen as the most convenient technique for trace metals analysis in marine waters using this type of microelectrode.

Effect of salinity Variations of Cu and Pb sensitivities were studied as a function of salinity by dilution of the seawater sample with milli-Q water. Results obtained are depicted in Fig. 3. The observed variations seemed independent of the analysed metal. For a salinity range from 30 to 12, the microelectrode sensitivity is quite constant (variation below 6%). At the salinity value of 6, the sensitivity diminished by a factor close to 2. Moreover, the baseline, the peak shapes and the peak positions are really modified (data not shown) but always remain repeatable. As an addition of NaCl or NaNO 3 to the solution did not bring the signals to the initial shape and sensitivity, this phenomenon should not only be attributed to a drop of the ionic strength but also to metal speciation modifications, mainly controlled by inorganic anions and leading to variations of their electrochemical labilities and so to variations of their behaviour at the gold surface. Therefore, trace metal analysis in river water should need some

Salinity peak height sensitivity

Fig. 3 Sensitivities variations of Cu (triangles) and Pb (diamonds) (expressed in A L mol -1 ×10 3 ) versus salinity. Experimental setup as in Fig. 1 modifications of the voltammetric parameters from those optimised for marine waters.

Electrode stability in time As the aim of this work is to prove the suitability of the gold microelectrodes for the monitoring of trace metals concentrations in natural systems, their stabilities were studied over a long period. The seawater sample with 5 μg L -1 of Pb was analysed for 16.5 h. Evolutions of Pb peak heights are depicted in Fig. 4. Compared to the average values (dotted line), measurements did not show any important variations (standard deviation inferior to 5%) and so confirmed the ability of this type of microelectrode for extended in situ analysis.

125-μm gold-disk microelectrode

Calibration of the 125-μm microelectrode for Zn, Pb, Cd and Cu was carried out by applying SWASV in seawater samples using a similar procedure as for the 25-μm electrode. As predicted by the theory and the preliminary results, which showed that concentration limits of saturation were higher for the 125-μm microelectrode than for the 25-μm one (surface ratio of the two electrodes is 25), the analysed range of metal additions was extended to 200 μg L -1 . The obtained results are summarized in Table 3.

Peak positions of the 4 studied metals are not modified by the use of this larger gold-disk surface. Compared to the results obtained with the 25-μm gold-disk, for all the studied metals, the linearity range is shifted to higher concentrations by one order of magnitude (i.e. from 0.2 to 30 μg Pb L -1 for the 25-μm electrode, and from 1 to 200 μg Pb L -1 for the 125-μm electrode). The analysed peaks are higher, and so easily measurable by voltammetric equipment, leading to higher sensitivity values. It has to be pointed out that this increase of sensitivity depends on the metal, ranging from a factor of 2 (Pb) to 7.5 (Zn) for the peak height (see Tables 2 and3). Calibration lines of the 25-μm and 125-μm microelectrodes for Pb are compared in Fig. 5. In this figure, the peak heights have been normalised to surface area of the electrodes and so expressed in A mm -2 . These results showed that for the 25-μm gold-disk the definition of the Pb peak is easier. Moreover, as expected, for the smaller electroactive area the sensitivity normalized to the electrode surface is higher (for Pb, 9.8 A L mol -1 mm -2 versus 4.8 A L mol -1 mm -2 , for the 25-μm and the 125-μm electrode, respectively) [START_REF] Bard | Electrochemical methods, fundamentals and applications[END_REF].

So, the 125-μm microelectrode could be used to study more polluted samples, as for example sediments, necessitating more robust equipment for analysis.

Unfiltered river samples

The River Deûle is well known for its high degree of metal pollution in water column as well as in sediments, as a result of the intensive industrial activities [START_REF] Vdović | [END_REF]. Total dissolved metal concentrations in the water sampled on 16 June 2005 were measured by ICP-MS after sample filtration and acidification. Results obtained are 0.30±0.02 μg Cd L -1 , 1.60±0.08 μg Cu L -1 , 4.0± 0.2 μg Mn L -1 , 4.8±0.3 μg Pb L -1 and 30±2 μg Zn L -1 . Considering these values, analysis carried out in the River Deûle samples has focussed on the measurements of Cu, Pb and Zn.

As observed before, in samples with low salinity such as river water, voltammograms are quite different from the marine ones. Moreover, an adsorption of organic compounds occurred, as indicated by a peak in the first cathodic scan of the cyclic voltammetry procedure. With time, this adsorption greatly modifies the sensitivity of the gold electroactive surface, leading to incorrect measurements. From the cyclic voltammogram in seawater sample (data not shown), it is possible to define the sign of the surface charge as negative below 0.1 V and positive above. As the standby potential defined in our experiments to clean the surface is 0.65 V (see Table 2), the electroactive surface is kept positive, which facilitates the adsorption of organic matter that is globally negative at natural pH, since it is mostly deprotonated [START_REF] Buffle | Analytical chemistry. Ellis Horwood 32. Ružić I[END_REF]. The solution adopted to minimize the interaction of organic matter on measurement consists in keeping the electrochemical cell off between two measurements, allowing the electrode to position to the potential of the solution (open circuit potential) and by just maintaining a conditioning step of 30 s at +0.65 V. This led to a clear attenuation of this undesirable adsorption, giving correct voltammetric analysis of unfiltered river sample within the time of measurement.

Calibration

The 25-μm gold-disk microelectrode was calibrated in the unfiltered river sample, applying SWASV and DPASV techniques for Cu, Pb and Zn, by standard additions from 1 to 20 μg L -1 (to 50 μg L -1 for Zn) in a decade mode.

Owing to the baseline stability, the mathematical treatment of the obtained voltammograms was similar to that used for the seawater sample. The results obtained for DPASV are summarized in Table 4. Zn, Pb and Cu peaks are depicted in Fig. 6 for successive additions of standard solutions in unfiltered sample at natural pH. The signal stability proved the ability to use the 25-μm gold-disk microelectrode for analysis of these 3 trace metals in a complex medium such as this unfiltered polluted river water. Regarding the shape of the voltammograms obtained in the pure sample, prior to standard additions (see Fig. 6, lowest scans), peaks corresponding to the 3 metals are clearly detectable and could be analysed after calibration of the microelectrode. Their peak potentials were shifted to more negative values by 110-280 mV, when compared to seawater matrix. This indicates the existence of more inert complexes.

Considering the complexity of the analysed solution and particularly the existence of binding organic compounds and adsorptive particles, some chemical forms of these 3 metals-named inert ones-might not be analysable by voltammetry. Indeed, the fraction of metal measurable by a voltammetric technique, referred to as labile, usually corresponds to the sum of free metal, labile inorganic complexes and some organic ones. In this sense, during additions of standard metal solutions, a concave curvature in the microelectrode response (i.e. area or peak height versus added metal concentration) means that the added

E (V) I (nA)

Zn

Cu Pb

Fig. 6 Evolution of Zn 2+ , Pb 2+ and Cu 2+ peaks measured by DPASV without stirring for different additions of standard solutions in unfiltered and deaerated river sample (Zn 0, 2.5, 5, 7.5, 10, 12.5 and 25 μg L -1 ; Pb, Cu 0, 1, 2, 3, 4, 5 and 10 μg L -1 ). Working electrode 25-μm gold-disk microelectrode metal is inertly complexed. At the end of the curve, a linear section allows determination of the microelectrode sensitivity [33]. In contrast, if the microelectrode response is linear over the whole concentration range scanned, the analysed metal should only be in the labile forms. For Cu, the microelectrode response is linear up to saturation (20 μg L -1 of added metal). This metal seems to be only bound by inorganic and organic labile ligands, at least in the sensitivity range of the electrode. It is probable that the strongly complexed Cu is one or two orders of magnitude lower than the first standard addition (1 μg L -1 ) and thus covered by noise. So, Cu concentrations measured by the microelectrode should be quite close to the total ones measured by ICP-MS. In contrast, although the response of Pb seems linear below 5 μg L -1 of added Pb (see Fig. 6), a curvature appears at higher concentrations, which indicates that Pb is not fully labile but partially complexed by inert organic ligands or adsorbed on particles. Above 10 μg L -1 , the microelectrode response remains linear, allowing the determination of its sensitivity. In the case of Zn, the microelectrode response is quite similar to that obtained with Pb, although the identification of the curvature is more complicated. So, for Pb and Zn, the concentrations determined by voltammetry should be lower than the total ones, corresponding to labile and a priori more bioavailable fractions.

Zn and Cu sensitivities obtained in this sample (see Table 4) are lower by a factor close to 2 compared to the sensitivities obtained using DPASV (data not shown) in the seawater sample. Sensitivities for Pb is increased by a factor close to 2. These differences were not further investigated but confirm the necessity to calibrate the microelectrode for each type of water analysed.

Comparison of the results obtained by SWASV and DPASV showed that DPASV leads to more repeatable and less noisy signals, albeit less sensitive. This may be due to the presence of natural organic compounds and particles which interact with trace metals, a phenomenon which could disturb a rapid reoxidation step as in SW mode. DPASV was therefore chosen for the monitoring measurements carried out on river samples. Initial concentrations of Cu, Pb and Zn were determined from the voltammograms obtained from the sample prior to standard addition of metal ions. Results obtained are 1.70± 0.01 μg Cu L -1 , 0.70±0.01 μg Pb L -1 and 20.4±1.3 μg Zn L -1 . As expected, Cu concentration measured by voltammetry is close to the total one measured by ICP-MS, whereas fractions analysed for Pb and Zn are respectively 80% and 30% lower than total concentrations in the sample.

Dissolved trace metals in the River Deûle

Some modifications of the electrochemical equipment were carried out for the monitoring experiments. Due to the high microelectrode sensitivity toward oxygen, samples have to be deaerated before their voltammetric analysis. So, measurements have not been carried out directly in the river, but by means of a homemade transportable measurement system placed in the boot of a car, parked near the River Deûle. A pump (CO/TECH) immerged in the river feeds two on-line homemade Teflon cells (average internal volume 65 mL). The first cell is used to perform pH measurements. The second cell is used to perform voltammetric measurements with the 25-μm gold-disk microelectrode. Homemade reference electrode (Ag/ AgCl/3M KCl in agarose matrix) and auxiliary electrode (Pt wire) are devised to minimize their sizes compared to the commercial ones, and to be usable for analysis in sediments. These 3 electrodes are connected to the μAutolabII electrochemical instrument, controlled by the GPES 4.9 software installed on a laptop. Both cells are slowly but continuously stirred (magnetic stirrer, Radiometer) to prevent the sedimentation of particles. Continuous stirring of the on-line cells has not led to either significant increase of the noise of the signal, nor noticeable increase of the DPASV peaks, and so is not to be compared with classical stirring in the voltammetric cell used in the laboratory which affects the rate of reduction at the electrode. A measurement sequence, carried out on a (new) river sample, consists in: (i) rinsing the cells for 5 min with the solution from the immerged pump, (ii) stopping the pump-flow and purging the solution in both cells for 10 min, (iii) stopping the purge and performing 2 cyclic voltammetry sequences to clean the gold-disk surface and checking the absence of oxygen, (iv) performing a DPASV analysis with the 30-s purge before each measurement (deposition time 120 s).

This system was used to monitor Cu, Pb and Zn concentrations in the River Deûle (just upstream of the metal industry factories), for 6.5 h on 21 July 2005. During that period, 15 samples were taken and studied. For each sample, 20 mL was filtered and acidified for ICP-MS measurements. The measured pH values ranged between 7.7 and 8.0, and were not modified by purging.

The DPASV measurements for each sample were repeated twice. The voltammograms were more noisy than those obtained in the laboratory, an expected phenomenon considering the conditions of on-site measurements, and so have necessitated the use of a Savitzky-Golay smoothing function (GPES 4.9 software, EcoChemie) before the determination of Cu, Pb and Zn peaks. Considering the baseline shape and the peaks positions, very close to those obtained during the calibration step, the sensitivities summarized in Table 4 have been used to calculate Cu, Pb and Zn concentrations. These values are displayed in Fig. 7 and compared to the ICP-MS results.

Total metal concentrations measured by ICP-MS were higher than those measured on 16 June 2005. This could be explained by the strong wind observed on that day, blowing in the opposite direction to the flow, which could more polluted waters coming from metal industry factories site move upstream. Over the 6.5 h, total Zn (Fig. 7a) and Cu (Fig. 7c) concentrations decreased (respectively from 38 to 35 μg Zn L -1 and from 1.9 to 1.5 μg Cu L -1 ), while total Pb (Fig. 7b) concentrations increased (from 7.5 to 9.1 μg Pb L -1 ).

As expected for Cu, the labile concentrations measured by the microelectrode are closer to the total ones than in the case of Zn and Pb. Indeed, the average labile fractions (i.e. F M =M labile /M total ×100) of Zn, Pb and Cu are 60%, 29% and 87%, respectively. These labile fractions appear to fluctuate (see in Fig. 7a,b and c the labile metal concentrations variations versus the total ones) and to be independent of the total metal concentrations. When these fractions are normalized to their maximal values (i.e. F′ M =F M /F M max ), as depicted in Fig. 7d, they show a similar tendency, even for Cu. These could indicate that the reactivity of the system (i.e. interactions of organic compounds and particles with the analysed trace metals): (i) is different in the absolute between the 3 metals, as the amplitude of the labile fraction ranges from 20% to 100% depending on the metal, (ii) varies independently of the total metals concentrations, and (iii) its time variation is quite similar for the 3 metals. On this subject, it has to be pointed out that the period in which the metal fraction as inert forms is maximum (Fig. 7d), corresponds to the classical time of high biologic activity. But, to confirm or reject this relationship, additional experiments should be carried out.

This experiment has shown the ability of the gold-disk microelectrodes for the monitoring of dissolved trace metals in a polluted aquatic environment, without filtration of the samples. The metal concentrations measured by this technique are not the total dissolved ones, like those determined by ICP-MS, but labile ones corresponding to the fraction of metal not bound by inert organic compounds or particles, a fraction which is a priori more bioavailable for the biota.

Finally, in order to prove that in situ analysis is really possible in complex and abrasive media with such a gold microelectrode, a preliminary study was carried out in a sediment. For that purpose a sediment core was sampled in the Deûle Canal, just upstream from the former Metaleurop smelter plant. About 3 cm of the overlaying water was also kept to avoid any disturbance at the interface. Reference electrode and counter electrode were introduced directly into the water and the 25-μm gold-disk microelectrode was carefully introduced into the sediment using a micromanipulator. Since oxygen is consumed in the first 2 mm under the sediment-water interface (personal data, not shown here, obtained with an O 2 microsensor, Unisense) due to bacterial activities, deployment of the gold microelectrode does not suffer from broad oxygen reduction waves at deeper depths. Figure 8 displays a typical voltammogram in the sediment for a deposition time of 240 s. The presence of Mn 2+ , probably associated with inorganic ligand [30], is bound to the reduction of manganese hydroxides(III/IV) by the bacterial activity, which consumes these species as electron acceptors for the degradation of organic matter. The presence of lead at high concentration level (10 μg L -1 ) was also detected using this microelectrode. Metaleurop, a former Zn and Pb smelting plant, had indeed dramatically polluted this zone: lead is present up to 1% in the sediment, just in front of the discharge banks [37].

Concentrations of manganese (175±10 μg L -1 ) and lead (18±2 μg L -1 ) were estimated by using ex situ calibration curves carried out in the river water medium. However, under these conditions, the calculation of the true concentration, assuming first of all that the oxidation peak is bound to the studied metal, remains a complex problem. To partly solve such difficulties, several voltammograms have been recorded in the sediment by varying the deposition time (Fig. 8, inset). This experiment at least allows one to check if the peaks height is a linear function of the deposition time. Sediment is a non-well-defined system made of particles and water, in which adsorption of organic compounds, minerals colloids and clusters (as for instance FeS) at the surface of the electrode may lead to a nondefined current or, even worse, a blocking of the electrode might occur. This experiment has pointed out that the redissolution peaks at -0.9 V and -0.17 V can be attributed to manganese and lead species, respectively. But it also appears that the peak detected at -0.55 V is not a sure proof of the presence of labile zinc in the porewater.

Conclusion

Homemade bare gold-disk microelectrodes of 25-μm and 125-μm diameter have successfully replaced mercury and mercury-coated electrodes in adapted voltammetric procedures for trace metals analysis in polluted natural waters. Calibration curves were performed in seawater and in river water samples, the SWASV method being more appropriate for seawater matrix and DPASV for river water. Compared with a mercury electrode, the oxidation peak potentials of Mn, Zn, Pb, Cd, Cu and Hg are shifted to more positive values, having analytical linearity range from 0.2 μg L -1 to 20 μg L -1 and from 1 μg L -1 to 200 μg L -1 for the 25-μm and 125-μm electrode, respectively, depending on the analysed metal. The electrode sensitivity varied from 0.23 to 4.75 nA L μmol -1 , and from 1.48 to 58.53 nA L μmol -1 for the 25-μm and 125-μm electrode, respectively, depending on the analysed metal and should be determined in any significantly different sample matrix. When compared with the same electrode coated with mercury, this bare gold-disk surface appears to be more sensitive.

The electrodes were successfully applied in an on-site monitoring system to determine and follow the labile concentrations of Cu, Pb and Zn in polluted river water. Due to the drawback of its high sensibility to oxygen, this electrode should be appropriate for in situ measurements in anoxic conditions like those found in sediments, as proved by the preliminary study undertaken in the sediments of the Deûle River (France).

The sensitivity improvement implies that such electrodes could be applied for detection of natural concentrations of trace elements; the unexpected findings like Zn double peak and more negative peak positions in river water than in seawater, are the subject of future investigations of such bare metal electrodes. Repeated field measurements and larger result databases should give the expected evidence of the biological activity in the measured natural and polluted waters and porewaters.

Fig. 1

 1 Fig. 1 Evolution of Mn 2+ , Zn 2+ , Pb 2+ , Cd 2+ , Cu 2+ and Hg 2+ peaks measured with the 25-μm gold-disk microelectrode using SWASV without stirring for different additions of standard solutions in filtered and deaerated seawater sample (Mn: 1, 2, 5, 7.5, 10, 15 and 20 μg L -1 ; Zn, Pb, Cd, Cu, Hg: 1, 2, 3, 4, 5, 7.5 and 10 μg L -1 )

Fig. 2

 2 Fig. 2 Calibration of the 25-μm gold-disk microelectrode for Pb up to 60 μg L -1

Fig. 4

 4 Fig. 4 Time variations of Pb peak height measured by SWASV (diamonds peak height, dotted line peak height average value) in seawater sample with 5 μg L -1 of Pb. Experimental setup as in Fig. 1

Fig. 5

 5 Fig. 5 Comparison of a the 25-μm (diamonds) and 125-μm (triangles) gold-disk microelectrode calibration lines (additions in decade mode from 0 to 200 μg L -1 of Pb) and b peaks shapes (additions 1, 2, 3, 4, 5 and 10 μg Pb L -1 ) measured in seawater sample with SWASV

Fig. 7 Fig. 8

 78 Fig. 7 Variations in time of a Zn, b Pb and c Cu concentrations measured by DPASV in the River Deûle on 21 July 2005. d Variations in time of the normalized Zn, Pb and Cu labile fractions (see explanation in the text). (solid line DPASV average values and standard deviation, dotted line ICP-MS values, squares Zn, diamonds Pb, triangles Cu)

Table 1

 1 Optimised voltammetric parameters for Zn, Cd, Pb, Cu, Mn and Hg analysis in seawater sample by SWASV and DPASV techniques

		SWASV	DPASV
	Purge	30 s	30 s
	Accumulation step
	Conditioning potential	0.65 V	0.65 V
	Conditioning time	30 s	30 s
	Deposition potential	-1.1 V	-1.1 V
	Deposition time	120 s	120 s
	Equilibration time	5 s	5 s
	Measuring step
	Potential window	-1.1 to 0.65 V	-1.1 to 0.65 V
	Frequency	25 Hz	/
	Step potential	5 mV	10 mV
	Amplitude	20 mV	75 mV
	Modulation time	-	0.05 s
	Interval time	-	0.60 s

Table 2

 2 Results of the 25-μm gold-disk microelectrode calibrations for Mn 2+ , Zn 2+ , Pb 2+ , Cd 2+ and Cu 2+ in seawater sample applying SWASV (all the potential values are given versus the Ag/AgCl, [KCl]=3 mol L -1 reference electrode)

		Mn	Zn	Pb	Cd	Cu
	Linearity (μg L -1 )	1 -80	0.2-30	0.2-30	0.2-20	0.2-20
	Peak potential (V)	-0.93	-0.57 and -0.34	-0.15	-0.02	0.25
	Sensitivity (A L μg -1 ×10 12 )	4.2	7.0	23.3	8.8	27.9
	Sensitivity (A M -1 mm -2 )	0.05	0.9	9.8	2.0	3.6
	Sensitivity standard deviation (%)	0.5	1.2	0.9	0.6	0.5
	R 2	1.000	0.996	0.998	0.999	0.999

Table 3

 3 Results of the 125-μm gold-disk microelectrode calibrations for Zn 2+ , Pb 2+ , Cd 2+ and Cu 2+ in seawater sample by applying SWASV (all the potential values are given versus the Ag/ AgCl, [KCl]=3 mol L -1 reference electrode)

		Zn	Pb	Cd	Cu
	Linearity (μg L -1 )	1-100	1-200 5-100 -100
	Peak potential (V)	-0.58 and -0.37 -0.15 -0.02 0.25
	Sensitivity	22.6	282.5 68.8	334.6
	(A L μg -1 ×10 12 )
	Sensitivity	0.12	4.8	0.63	1.7
	(A M -1 mm -2 )
	Sensitivity standard	0.7	1.1	1.1	0.9
	deviation (%)
	R 2	0.999	0.996 0.998 0.997

Table 4

 4 Results of the 25-μm gold-disk microelectrode calibrations for Zn 2+ , Pb 2+ and Cu 2+ in an unfiltered river sample by applying DPASV (all the potential values are given versus the Ag/AgCl, [KCl]=3 mol L -1 reference electrode)

					0.25
					0.20
					0.15
					0.10
		Zn	Pb	Cu
	Linearity (μg L -1 ) Peak potential (V) Sensitivity (A L μg -1 ×10 12 ) 2.0 0 -25 -0.87 and -0.60 -0.30 0.03 10-20 0-20 23.7 8.9	0.05	-1.0	-0.8	-0.6	-0.4	-0.2	0.0	0.2
	Sensitivity (A M -1 mm -2 )	0.2	10.0	1.1
	Sensitivity standard	6.5	1.9	0.6
	deviation (%)
	R 2	0.984	1.000 0.999

Acknowledgements This work was supported by the Region Nord Pas-de-Calais [Stardust (INTERREG III)], the national PNETOX program and the European Community (FEDER) as well as by COGITO, Program of Integrated actions by ÉGIDE. We are very grateful to Jean-François Barthe for his technical support during the sampling and the monitoring, as well as for the ICP-MS measurements. We thank anonymous referees for their valuable suggestions.


Véronique Lenoble
email: lenoble@univ-tln.fr

Cédric Garnier

Armand Masion

F Ziarelli

Jean-Marie M Garnier

Combination of 13 C/ 113 Cd NMR, potentiometry, and voltammetry in characterizing the interactions between Cd and two models of the main components of soil organic matter

Keywords: Cadmium complexation, Soil organic matter, Complexation modeling, 13 C/ 113 Cd NMR, Voltammetry

Combination of 13 C/ 113 Cd NMR, potentiometry, and voltammetry in characterizing the interactions between

Cd and two models of the main components of soil organic matter

Introduction

Natural organic matter (NOM) is ubiquitous in the environment and consists of a complex mixture [START_REF] Aiken | Humic substances in soils, sediment and water[END_REF], depending on the origin and age of the material [START_REF] Buffle | Complexation of trace metals in natural waters[END_REF][START_REF] Gu | [END_REF]. NOM is also known to play important roles in the fate of many contaminants due to its complexing properties. A better comprehension of NOM structural and functional properties can greatly improve our understanding of the underlying mechanisms responsible for heavy metal complexation [4]. This, in turn, may enhance our predictive capabilities regarding the behavior of NOM and environmental inorganic pollutants in natural ecosystems.

Reactivity of soil organic matter (SOM) towards cadmium, a quite toxic element, is of particular interest, especially in the case of polluted soils and their (phyto-)remediation [5][6][7], but also to better quantify the bioavailability of Cd towards the plants. Many studies have investigated implications of SOM in Cd transport and bioavailability, e.g., in a rhizosphere soil [8], but with a less important combination of analytical tools than in the study presented here.

Many microorganisms, especially in soils, produce extracellular polysaccharides (EPS) with a high molecular weight, consisting of polysaccharides, proteins, and nucleic acids [9]. These biopolymers are known to complex heavy metals [START_REF] Sterritt | Immobilisation of ions by bio-sorption[END_REF][START_REF] De Stephano | [END_REF][12] and represent the main components of SOM in natural systems [13,14].

No single analytical tool can provide structural or functional information about NOM because of its heterogeneous, complex nature. Thus, a combined application of various analytical techniques is more suitable. Among characterization methods, spectroscopic techniques appear the most useful, since they are non-destructive, usually require no or little sample preparation, and they provide valuable information on molecular structure and chemical or functional NOM properties [15]. On the other hand, voltammetry is a suitable technique to study the characterization of NOM metal-binding sites, as it is a sensitive method that does not need any physical or chemical sample preparation.

The EPS structure can be characterized by using spectroscopic techniques, e.g., FTIR and 13 C solid-state NMR spectroscopy. NMR analysis is a direct probe of the studied nuclei surroundings [16] and the chemical shifts of the targeted nuclei have a high sensitivity to variations in the local chemical surroundings. Regarding Cd complexation, 113 Cd NMR spectroscopy has already been used as a probe of Cd 2+ complexation by a known functional group [17,18], by natural organic matter [19,20], and by biomaterial [21].

Differential pulse anodic stripping voltammetry (DPASV) is another technique that is able to recognize metal-ligand complexation by directly measuring the non-organic metal fraction [22,23]. The metal fraction analyzed by DPASV referred to as labile, usually corresponds to the sum of free metal and labile inorganic complexes. Therefore by knowing the total metal concentration, the organically bound metal fraction can be calculated.

With voltammetry results fitted by using Scatchard [24] and Ružić [25] linearization methods one can estimate if the number of SOM apparent binding sites is one or more. However, as these methods often generate important errors in terms of the ligand binding properties [22], a non-linear fitting has to be applied afterwards. Nowadays, more complex and flexible fitting programs have been developed that include models with various chemical interactions [26][27][28][29]. One such program is PROSECE (Programme d'Optimisation et de Speciation Chimique dans l'Environnement), software that allows the determination of complexation properties using a discrete model distribution of binding sites [22,30,31]. PROSECE be considered as an alternative to the MODEL VI or NICA-DONNAN approaches involved in WHAM and FITEQL software, respectively [27,29].

In this study, two exopolysaccharides (EPS) used as models for exudates of roots and of soil bacteria were studied for their reactivity towards Cd and their implications in terms of Cd speciation and bioavailability in soils. The selected EPS were of bacterial (Phytagel) and plant (polygalacturonic acid, hereafter named PGA) origin. Both have well-known structures [32,33], but their reactivity towards trace metals (i.e., in order to study their role in metal bioavailability) has not been investigated much. By combining spectroscopic techniques, potentiometry, and voltammetry (which has scarcely been presented in the literature) and by fitting of the measured data to the corresponding model, conditional binding parameters (e.g., logK, pK a , and binding site concentrations) are proposed in this work. These parameters can further be used in any geochemical model currently lacking these kinds of values.

Experimental

Chemicals

All reagents were of Normapur quality. Deionized water was obtained with a Milli-Q water purification system. Polygalacturonic acid and Phytagel were of Analytical Grade and were both provided by Sigma.

Procedures for EPS characterization and EPS-Cd complexation study

Potentiometry experiments were conducted on 50 mL of 250 mg L -1 PGA and Phytagel solutions (I = 0.01 M, NaNO 3 ), which corresponds to 12.5 mmol DOC L -1 and 7.85 mmol DOC L -1 of carbon content, respectively. Spectroscopic analyses of cadmium complexation was performed by addition of cadmium (as 1 mg Cd/mL in 2% HNO 3 solution, Inorganic Ventures Labs) to PGA and Phytagel solutions, with a fixed carboxylic group to metal quantity ratio of 10/1 or 10:1 (0.48 mmol COOH L -1 and 0.044 mmol Cd L -1 in 50 mL), in a solution of pH 6. After 5 days of contact time, the solution was freeze-dried and analyzed by FTIR, 13 C and 113 Cd solid-state NMR spectroscopy.

Voltammetric cadmium titrations were carried out at different pH values (3, 8, and 10) in order to determine proton and cadmium competition for the binding sites. For each EPS (3.33 mmol DOC L -1 , I = 0.01 M, NaNO 3 ), and at the various studied pH values, standard additions in a logarithmic addition mode [22,31] were performed; total concentration ranged from 10 -9 to 10 -7 mol L -1 . For each metal addition an equilibration time of 1 h was set up, which appears sufficient to reach complexation equilibrium. Each measurement was duplicated, and analytical uncertainty is around 3%.

Analytical tools

Total organic carbon in solution was measured with a Shimadzu TOC 5000 with an accuracy of 0.1 mg C L -1 .

Fourier transform infrared (FTIR) analyses were carried out with KBr pellets (100 mg KBr + 1 mg sample) on an Equinox55 (Bruker) spectrometer. Spectra were recorded in transmission mode with 16 scans for each spectrum, between 4,000 and 400 cm -1 , with a 4-cm -1 resolution.

All solid-state cross polarization nuclear magnetic resonance spectroscopy with magic angle spinning (CPMAS NMR) spectra were obtained on a Bruker Avance 400-MHz NMR spectrometer operating at a 13 C and 113 Cd resonance frequency of 101.6 MHz and 88.8 MHz, respectively. 13 C and 113 Cd CPMAS experiments were performed with a commercial Bruker double-bearing probe with zirconium dioxide rotors of 4-mm outer diameter. The CP technique [34] was applied during MAS of the rotor at 10 kHz. A ramped 1 H pulse starting at 100% power and decreasing to 50% was used during contact time to circumvent Hartmann-Hahn mismatches [35,36]. For 13 C and 113 Cd CPMAS, contact times were 2 ms and 5 ms, respectively, and the number of scans were 2,048 and 50,000 (delay of 2 s and 5 s, respectively) [37]. To improve the resolution, dipolar decoupling on the proton channel was applied with TPPM-15 sequence during acquisition. The 13 C and 113 Cd chemical shifts were referenced to tetramethylsilane and 0.1 M Cd (ClO 4 ) 2 in aqueous solution, respectively [37].

EPS acidic functions distribution was analyzed by potentiometric titrations [26,[38][39][40][41]. Acid-base titrations were carried out following the procedure detailed elsewhere [30]. Briefly, experiments were conducted in thermostated cells at 25±0.2°C, by additions of HNO 3 (0.20 M, from HNO 3 69% J.T. Baker) until pH 2 was reached, then additions of KOH (0.10 M, from KOH 0.5033 M Sigma-Aldrich) up to a value of 11.9 were performed. NaNO 3 concentration in KOH and HNO 3 standard solutions was 0.10 M. The microtitration stand (Metrohm) was equipped with two Titrino 716 titrators controlled by Tinet2.4 software. The combined pH-micro-electrode used (Ross 8103SC, Orion) was calibrated by pH buffer solutions (HANNA 4.01, 7.01, and 10.01) before each titration and controlled after.

Voltammetric measurements were performed on a Metrohm-EcoChemie stand controlled by the GPES 4.9 software [23]. This apparatus is composed of a potentiostat/galvanostat (PGSTAT12) connected to the measurement stand (VA663), comprising a static mercury drop electrode (SMDE), a reference (Ag/AgCl/KCl 3 mol L -1 ) and an auxiliary glassy carbon electrode. After each DPASV measurement, carried out in a thermostated Teflon ® cell (25±0.2°C), pH was monitored by a combined micro-electrode (Mettler, Inlab422) linked to a pHM713 (Metrohm). The voltammetric measurements involved N 2 purging through the cell for 120 s, a deposition time of 600 s at a potential of -0.9 V, then a scan in differential pulse mode towards oxidative potentials, from -0.9 to 0.1 V.

Results and discussion

NOM characterization 13 C NMR and FTIR spectra (Fig. 1, thin lines) confirm the known structure of EPS (see Table 1).

A Phytagel unit is composed of four tetrasaccharides (i.e., a 24-carbon moiety), and PGA is composed of a monosaccharide (i.e., a six-carbon moiety). Thus, each carbon should contribute to the signal with a proportion of 4.2% for Phytagel and 16% for PGA. The 13 C NMR carbon distributions are presented in Table 1. The proportions derived from the fitting are close to the theoretical values, thus showing that the applied CPMAS pulse sequence using a 1 H ramp pulse during cross polarization [42] and an optimized contact time (2 ms) allows a (semi-)quantitative interpretation of the 13 C NMR spectra.

PGA acide-base titration is presented in Fig. 2; the titration of Milli-Q water that does not present any acidic sites. Acid-base titrations results were fitted (Fig. 2) by PROSECE to determine the distribution of the discrete acidic sites [30]. In the literature, distribution of natural organic matter acidic sites is reported to range from two [40] to six sites [38]. For each studied EPS, the best PROSECE fitting of the acid-base titration (see Fig. 2 and Table 2) was obtained with four classes of acidic sites, with two preponderant in percentage, expressing the presence of various kinds of surroundings for these sites.

All models applied to NOM aim at defining groups of acidic sites, distinguishing those with acidic pK a in the range usually attributed to carboxylic groups (pK a <7, thereafter named acidic sites A) and those with basic pK a in the range usually attributed to phenolic groups (pK a >7, thereafter named acidic sites B) [26,30,38,39]. For the studied EPS (Table 2), acidic sites A and also acidic sites with higher pK a , in the range of acidic sites B, were present. The observation is consistent with others NOM acidic sites identification [26,30,38,40,41]. Furthermore, this result has already been observed for acidic binding sites of simpler structures (e.g., EDTA and glucuronic acid), whose acidic sites showed pK a values in the range of acidic sites B [START_REF] Ringbom | Complexation in analytical chemistry[END_REF].

Studies on PGA and gellan (Phytagel-like polymer family) structures revealed a helix-based structure of the chains [32,33], leading to an overall conformation that forms a three-dimensional network. Therefore, the observed acidic sites distribution could be linked to the helix structure, leading to different but repetitive surroundings of the carboxylic groups.

Results also emphasize the importance of the carboxylate and hydroxyl groups and of the hydrogen bonds in the interchain interactions, as the pK a of such sites (e.g., C4 on PGA) is around 11-12 [START_REF] Makridou | [END_REF].

Although the total acidic sites density is almost the same for the two studied exopolysaccharides (14.6 and 13.2 meq g DOC -1 ; Table 2), the distribution is completely different: Phytagel acidic sites are mainly acidic sites B, with a ratio A/B of 0.15, whereas for PGA, the proportion of acidic sites B is in smaller excess (ratio A/B of 0.76). These results can be explained by the different environments of the carboxylic groups inside the three-dimensional polymer structure [32,33,[START_REF] Flory | Principles of polymer chemistry[END_REF] and the OH distribution. Therefore characterization highlighted the presence of various classes of acidic binding sites for each studied compound.

Exopolysaccharide-Cd complexation

It is well known that only a minor fraction of the measured acidic sites display a high affinity towards metals [START_REF] Buffle | Complexation reactions in aquatic systems: an analytical approach[END_REF]. Both our spectroscopic and voltammetric complexation investigations aimed at determining the nature of these high energy sites, i.e., whether sites A or B are involved in Fig. 1 13 C NMR spectra (a) and FTIR spectra (b) of PGA alone (thin line) and PGA-complexed Cd (thick line). Analyses were performed with a fixed EPS carboxylic group to metal quantity ratio of 10:1 (0.48 mmol COOH L -1 and 0.044 mmol Cd L -1 in 50 mL) solution at pH 6, freeze-dried after 5 days of contact time and analyzed cadmium binding. The obtained results should allow the precise characterization of the Cd-binding sites of the studied SOM exopolysaccharides.

Spectroscopic study of cadmium complexation

Cadmium complexation by the studied EPS was assessed by 13 C and 113 Cd NMR and FTIR measurements. 13 C NMR spectrum of PGA in the presence of Cd (Fig. 1a, thick line) revealed a 4-ppm downfield shift of the 172-ppm peak with respect to the spectrum without Cd, indicating the change in the surroundings of the carboxyl group(s). This change can also be seen by FTIR analysis (Fig. 1b, thick line). The main difference is the presence of COOH in PGA without Cd (1,739-cm -1 band) which almost disappears when in the presence of cadmium, being replaced by COO - (1,593-cm -1 band). The sharp band at 1,381 cm -1 corresponds to the nitrate of the added cadmium salt.

In contrast, no differences were observed between 13 C NMR spectra of Phytagel and Phytagel-Cd (data not shown). This observation is consistent with the titrations results. Indeed, to avoid Cd(OH) 2 precipitation, the experiment was performed at pH 6, i.e., a pH value too low to allow quantitative deprotonation of acidic sites B, which account for the vast majority of the measured sites. The proportion of sites reacting with Cd was small (around 13%, i.e., sum of % of site 1 and site 2, Table 2). Therefore, Cd complexation though occurring, did not lead to any significant NMR downfield shift. The same conclusions can be drawn from the FTIR data. 113 Cd NMR measurements showed (Fig. 3) downfield shift and broadening of signals of PGA-Cd and Phytagel-Cd compared with unreacted solid-state mixtures of the Cd salt with the studied exopolysaccharides, confirming strong interactions between Cd metal and organic materials.

Additionally, Fig. 3 shows a different shift of signal for Cd in PGA and Phytagel, -38 and -88 ppm respectively, which can be due to a different cadmium complexation on the two studied EPS.

Therefore, spectroscopic analyses added to characterization and also proved that cadmium complexation is different when considering the two studied EPS.

Voltammetric study of cadmium complexation

Voltammetric cadmium titrations were carried out at pH values 3, 8, and 10 to determine proton and cadmium competition for the binding sites. These analyses allow a more precise definition of the binding sites involved in cadmium complexation.

Experiments conducted at pH 3 showed no cadmium complexation by the studied exopolysaccharides (results not shown). Therefore, the concerned EPS complexing sites seemed not to be deprotonated at this pH, which correspond to pK a values above 3. This observation corroborates the potentiometric titrations results.

Fig. 3 113 Cd NMR analysis for PGA-complexed Cd (thick line), Phytagel-complexed Cd (dashed line), Phytagel, and cadmium simple contact (dotted line). Experimental conditions for complexation analysis are 0.48 mmol EPS L -1 and 0.044 mmol Cd L -1 dissolved in 50 mL; for the simple contact experiment, the same proportion of EPS and Cd (both as solids) were mixed but not dissolved A acidic sites with acidic pK a usually attributed to carboxylic functionality, B acidic sites with basic pK a usually attributed to phenolic functionality Cadmium titration results obtained with PGA at pH 8 and 10 are presented in Fig. 4. As data points are clearly below the non-complexation line (dashed line) it underlines that interactions between exopolysaccharides and cadmium dominate Cd speciation. So, natural exudates of bacteria and of roots could largely control cadmium distribution in soils. Moreover, interactions increased between pH 8 and pH 10, suggesting Cd 2+ /H + competition towards EPS binding sites and the influence of acidic sites B.

Fitting of metal titrations results using Scatchard [24] and Ružić [25] linearizations showed a curvature, i.e., that it is not a one-site complexation that is occurring (data not shown). Therefore, there is more than one class of binding site, which agrees with the different cadmium surroundings highlighted by our 113 Cd NMR measurements. This also agrees with a previous study on cadmium complexation by bacterial exopolysaccharides involved in the adhesion to surfaces and protection from environmental stress ([14 and references therein).

The experiments carried out were fitted by PROSECE using a discrete distribution of binding sites to characterize exopolysaccharides/Cd reactivity. Inorganic chemical composition of the solution was taken in account to calculate the inorganic speciation of Cd, using thermodynamic stability constants from MINEQL and MINTEQ databases [START_REF] Westall | MINEQL: a program for the calculation of chemical equilibrium composition of aqueous systems, RM Parsons laboratory technical note 18[END_REF][START_REF] Eary | Version 4.00 of the MINTEQ Geochemical Code[END_REF]. Best fitting was obtained defining two classes of sites (L1 and L2), both acidic and complexing, each defined by a binding site density (μeq g DOC -1 ), an acidic constant (pK a ), and a stability constant (logK) towards Cd, i.e., six unknown parameters for each EPS. The values of these six parameters were optimized using PROSECE, by fitting simultaneously the experimental data (i.e., H + and labile Cd concentrations vs. total Cd ones) obtained at pH 8 and 10, taking account of the EPS acidic parameters modeled previously (see Table 2). The obtained results are summarized in Table 3, and results of fitting by the proposed model of the experiments carried out on PGA are shown in Fig. 4.

For each studied EPS, a "strong" (L1) and a "weak" (L2) site can be defined in terms of the obtained complexation constant values. According to the pK a values obtained, these sites were described as follows: the "strong" binding sites are a part of the acidic sites B, the "weak" binding sites are a part of the acidic sites A.

Therefore, the different binding sites involved in Cd complexation, highlighted by characterization and spectroscopic measurements, were identified by voltammetry and fitting. However, differences appeared between PGA and Phytagel in terms of the sites densities, the stability constants, and the acid constants.

There was an important effect of pH variation on PGA complexation (Fig. 4), particularly at low Cd concentrations. This has been simulated (Table 3) by a small proportion (0.41 μeq g DOC -1

) of acidic sites B (strong Cd-binding site L1 with logK 10.4) and an important proportion (1.87 μeq g DOC -1

) of acidic sites A (weak Cd-binding site L2 with logK 7.1).

In contrast, with Phytagel there were almost no significant pH effects on the complexation on the scanned Cd concentrations (data not shown), and fitting (Table 3) showed a major proportion (1.32 μeq g DOC -1

) of acidic sites B (strong Cd-binding site L1 with logK 9.0). The difference of stability constant between PGA and Phytagel acidic site B has to be noted, and is probably due to the different surroundings highlighted by the spectroscopic techniques. Regarding to their quite basic pK a values (Table 3), these strong sites (L1) should not be efficient toward cadmium speciation in neutral to acidic conditions occurring in soil environments, due to a high H + /Cd 2+ competition effect. Yet these exopolysaccharides present a second class of sites, much more concentrated but weaker complexants (L2) with acidic pK a values (around 4). Therefore, it can be concluded that natural exudates must conserve an important role in cadmium speciation in soils.

As already mentioned, only a few studies reported in the literature have dealt with Cd complexation by exudates and they do did not involve the same analytical tools or fitting models employed here [14,[START_REF] Karlsson | [END_REF]. Nevertheless, these studies also pointed out the presence of two binding sites [14,[START_REF] Karlsson | [END_REF]. Lamelas et al. [14] obtained complexation parameters for the two binding sites that were different from ours (logK 2.4 and 2.95 ) when dealing with H + and Cd 2+ binding properties of bacterial exopolysaccharides. However, the two sets of results are barely comparable as the fitting model involved continuous distribution of organic matter binding sites. The complexation parameters found by Karlsson et al. [START_REF] Karlsson | [END_REF] for Cd complexation on functional groups of an organic soil are in the same range as our results: one strong complexant site in terms of its Cd 2+ -SOM association (logK = 11.2-11.6) and with basic functionality regarding H + -SOM association (pK a = 9.96), which the authors attributed to thiol functionality; and one weak complexant site (logK 3.2) and with acidic functionality (pK a = 3.5) [START_REF] Karlsson | [END_REF]. However, as the analytical tools are different (ion-selective electrode and EXAFS) the comparison is once again not easy.

The characterization of our model EPS has therefore led to a better understanding of the binding capacities of SOM main components towards cadmium and thereby to a better knowledge of the role of these key SOM compounds.

Conclusion

Owing to the combination of various analytical techniques and fitting, the characterization of the Cd-binding sites of the main components of soil organic matter (exudates of roots and of soil bacteria) was precisely investigated in this work.

Potentiometric results revealed the presence of acidic sites with acidic functionality in terms of H + -EPS association (sites A; pK a = 3.7-4.0) and acidic sites with basic functionality (sites B; pK a = 11.3-11.7), for both studied EPS. The ratio A/B is different for the two exopolysaccharides, probably reflecting the presence of various kinds of binding sites for the studied EPS. Spectroscopic techniques ( 13 C and 113 Cd NMR, FTIR) corroborated this result and especially proved that cadmium surroundings are different when considering complexation by the two studied exopolysaccharides.

Furthermore, voltammetric experiments combined with fitting allowed the characterization of each EPS binding site, including fractions of the defined acidic sites A and B. These binding sites present various intrinsic properties (various pK a ranges), and complexation ability towards cadmium is different (logK = 7.1-8.2 for sites A; logK = 9-10.4 for sites B). The binding sites could thereby be defined and differences between the two exopolysaccharides could be fitted by the proposed model. This combination of analytical tools and fitting should be used whenever characterization of SOM complexation properties is concerned. It allows one to better take into account the intrinsic reactivity of the main compounds of soil organic matter, mimicking SOM usually found in a plant-soil system, through their complexation ability. Furthermore, this study allowed an easier comprehension of cadmium speciation at the plant/soil interface. The parameters determined in this study can therefore be considered as trustworthy data for biogeochemical models (e.g., for metaltransfer modeling), to improve our knowledge and understanding of metal transfer from a soil to a plant.

Fig. 2

 2 Fig. 2 PGA potentiometric measurements (open circles) as compared to Milli-Q titration (no acidic sites; dotted line). Experiments were conducted in a thermostated cell (25±0.2 °C), on 50 mL of 250 mg L -1 PGA solution (I=0.01 M, NaNO 3 ), which corresponds to 12.5 mmol DOC L -1 of carbon content. Solid line results of fitting

Fig. 4

 4 Fig. 4 Labile (pCd labile , expressed as -log[Cd labile ]) vs. total Cd (pCd T , expressed as -log[Cd T ]) concentration in a solution of 3.33 mmol DOC L -1 of PGA. The metal fraction referred to as labile corresponds to the sum of free metal and labile inorganic complexes. Various experiments are presented: pH 8 (open circles), pH10 (open triangles), and non-complexed Cd (dashed line) as well as the results of fitting for pH 8 (solid line) and pH 10 (dotted line)

Table 1

 1 Assignment of13 C CPMAS spectra of PGA and Phytagel, along with their schematic representations

	Chemical shift (ppm)	Proportion	Assignment a
	PGA
	68-75	53.1	C2p, C3p, C5p
	81	16.0	C4p
	96-102	17.0	C1p
	172	13.9	C6p
	Phytagel
	18	4.1	C6r
	62	6.5	C6g
	68-77	54.6	C2 3 5g / C2 3 5ga / C2 3 5r
	83	15.0	C4g / C4ga / C4r
	104	16.4	C1g / C1ga / C1r
	176	3.4	C6ga

a p PGA, g glucose, ga glucuronic Acid, r rhamnose Phytagel schematic representation PGA schematic representation

Table 2

 2 Acidic sites distribution and total acidic sites density for Phytagel and PGA

		Phytagel		PGA
	Total acidic sites	14.6			13.2
	density (meq g DOC -1 )
		pK a	%	Type pK a	%	Type
	Site 1	4.0	10.9 A	3.6	43.1 A
	Site 2	6.6	2.0	A	8.2	2.2	B
	Site 3	9.3	3.6	B	10.3 7.7	B
	Site 4	11.2 83.4 B	11.7 47.0 B

Table 3

 3 PROSECE fitting results (site densities, stability constants, and pK a ) for PGA and Phytagel at the various studied pH values

	PGA		Phytagel
	L1	L2	L1	L2

Anal Bioanal Chem (2008) 390:749-757

Acknowledgments The authors are grateful to Spectropôle (Marseille University) for their kind help during this work. This work was financially supported by the TRANSPLAM (metals soil-plant transfer) project, which as a part of the French inter-organisms program TOXNUC. The authors also thank Ivanka Pieta for her valuable comments, as well as the referees for their advice.