Jia-Lan Wang

Enrica Alasonati

Mickaël Tharaud

Alexandre Gelabert

Paola Fisicaro

Marc F Benedetti
email: benedetti@ipgp.fr

Flow and fate of silver nanoparticles in small French catchments under different land-uses: The first one-year study

Keywords: Silver nanoparticles, surface water, single particle ICPMS, land-use, fate, concentration

This study focused on surface waters from three small creeks, within the Seine River watershed, which are characterized by different land-uses, namely forested, agricultural and urban. Silver nanoparticles (Ag-NPs) in these waters were detected and quantified by single-particle ICPMS during one-year of monthly sampling. Their temporal and spatial variations were investigated. Ag-NPs, in the three types of surface water, were found to range from 1.5 x10 7 to 2.3 x 10 9 particles L -1 and from 0.4 to 28.3 ng L -1 at number and mass concentrations, respectively. These values are in consistent with the very few previous studies.

In addition, the role of factors driving process and potential sources are discussed with correlations between Ag-NPs concentrations and biogeochemical parameters, like dissolved organic carbon concentration and divalent cations concentrations. For the forested watershed NOM controls the stability (number and mass) of the Ag-NPs as recently observed in the field in lake water in Germany. In the case of the agricultural and urban watersheds major cations such as Ca would control the number and mass of Ag-NPs.

Dilution processes are rejected as conductivity and Cl -ions do not show significant correlations with Ag-NPs or other major geochemical parameters. The specific exportation rates of Ag-NPs for artificial, agricultural and forested areas were calculated based on the monthly data for the full year and are equal to 5.5 ± 3.0, 0.5 ± 0.3 and 0.2 ± 0.2 gy - 1 km -2 , respectively. These data suggest a constant release of Ag-NPs from consumer products into freshwaters in artificial areas, for instance, from textiles, washing machines, domestic tap-water filters, outdoor paints.

These first data of Ag-NPs fluxes in surface waters of France enlarge the very limited database of field measurements. Moreover, for the first time, the influence of time, landuse and aquatic geochemistry parameters on Ag-NPs in real natural water samples is reported. It is also helpful to further understand the fate and the process of Ag-NPs in natural waters, as well as to the ecotoxicity studies in real-world environment.

Introduction

The first report of silver colloids goes back to 130 years ago in 1889 [START_REF] Lea | On allotropic forms of silver[END_REF]. For decades, nanosilver was mostly used under different names (Collargol, Argyrol, and Protargol) in therapeutic treatments [START_REF] Nowack | 120 years of nanosilver history: Implications for policy makers[END_REF]. More recently, the introduction and rapid development of nanotechnology opened up wider applications of silver nanoparticles, including textiles, cosmetics, food packaging, medical devices, catalysts, electronics, biosensors Moreover, silver nanoparticles are reported by Vance [START_REF] Vance | Nanotechnology in the real world: Redeveloping the nanomaterial consumer products inventory[END_REF] as the most frequently used nanomaterial and found in 435 consumer products (24% of total items). In a more general way, these silver nanoparticles may find their way into natural waters during their life-cycle. Even though wastewater treatment plants (WWTPs) are effective to remove silver nanoparticles, nano-sized silver/silver sulfides were found in the effluent of WWTPs [START_REF] Wang | Engineered silver nanoparticles in terrestrial environments: a meta-analysis shows that the overall environmental risk is small[END_REF]. For Europe input rate via WWTP residues to soils is estimated 0.37 tons.y -1 and 4.7 tons.y -1 to large rivers [START_REF] Wang | Engineered silver nanoparticles in terrestrial environments: a meta-analysis shows that the overall environmental risk is small[END_REF]. In addition to anthropogenic sources, silver nanoparticles could be also naturally formed from the silver ions in the water or in soils as recently reported by [START_REF] Huang | Significant contribution of metastable particulate organic matter to natural formation of silver nanoparticles in soils[END_REF] for soils [START_REF] Wimmer | New insights into the formation of silver-based nanoparticles under natural and semi-natural conditions[END_REF] for fresh waters. Johnson et al. (2005) have estimated for France an input of 11 tons y -1 of anthropogenic bulk silver from waste management to surface environments. Official registration in France recently reported 10 kg in 2017 (R-nano: https://www.r-nano.fr) that is a very small amount, so we would not expect so see too much engineered nano-Ag in waters. However, using the data of [START_REF] Wang | Engineered silver nanoparticles in terrestrial environments: a meta-analysis shows that the overall environmental risk is small[END_REF] for Europe and downscaling it to France, using Europe and France populations as the scaling factor, we can also estimate that 27 kg Ag-Nano y -1 and 425 kg Ag-Nano y -1 will be delivered to soils and waters, respectively. This extensive spreading of nano-sized silver and the lack of information regarding their long-term effect inevitably cause emerging concerns about their potential risks to aquatic ecosystems and/or humans. Despite a long history of use, more and more studies showed their adverse impacts on aquatic and terrestrial organisms, mammalian cells, and potentially on human health [START_REF] Bian | Silver nanoparticles promote procoagulant activity of red blood cells: A potential risk of thrombosis in susceptible population[END_REF]Dobrzyńska and Kruszewski, 2014;[START_REF] Gaillet | Silver nanoparticles: Their potential toxic effects after oral exposure and underlying mechanisms -A review[END_REF][START_REF] Johnston | A review of the in vivo and in vitro toxicity of silver and gold particulates: Particle attributes and biological mechanisms responsible for the observed toxicity[END_REF]. Musee [START_REF] Musee | Simulated environmental risk estimation of engineered nanomaterials: A case of cosmetics in Johannesburg City[END_REF] modeled their possible toxicity to the aquatic ecosystems of the city of Johannesburg, with a risk quotient (defined as the ratio of the estimated environmental exposure to the predicted noeffect concentration) mostly > 1.

The detection and quantification of nanosized silver, from different sources ranging from engineered and incidental to natural ones in the environment and named from here and after Ag-NPs, is however quite limited because of their expected low concentrations and the biogeochemical complexity of aquatic ecosystems. To circumvent this limitation, material flow analysis models have been developed to predict environmental concentrations since 2007 [START_REF] Boxall | Current and future predicted environmental exposure to engineered nanoparticles[END_REF]. The modeled concentrations of Ag-NPs in surface waters vary significantly, from 0.008 to 619 ng L -1 , depending on modeling parameters and studied sites (Dumont et al., 2015;[START_REF] Good | Implications of engineered nanomaterials in drinking water sources[END_REF][START_REF] Gottschalk | Modeling flows and concentrations of nine engineered nanomaterials in the Danish environment[END_REF][START_REF] Mueller | Exposure modelling of engineered nanoparticles in the environment[END_REF][START_REF] Musee | Simulated environmental risk estimation of engineered nanomaterials: A case of cosmetics in Johannesburg City[END_REF][START_REF] Brien | Nano-scale pollutants: Fate in irish surface and drinking water regulatory systems[END_REF][START_REF] Silva | Analytical chemistry of metallic nanoparticles in natural environments[END_REF][START_REF] Sun | Dynamic Probabilistic Modeling of Environmental Emissions of Engineered Nanomaterials[END_REF]. However, these predictions are currently not validated due to the lack of field measurements in aquatic environments which also limits the further advancement of modeling in other domains, such as their fate and toxicology in complex ecosystems. There is thus an urgent need for frequent and robust measurements of Ag-NPs concentrations on fluxes in continental surface waters.

A first issue is that to date, only a limited number of studies are available on the quantification of Ag-NPs in freshwaters. Some of them confirmed the presence of colloidal [START_REF] Wen | Colloidal and particulate silver in river and estuarine waters of Texas[END_REF] and more recently nano silver in wastewater effluent and river water [START_REF] Mitrano | Detecting nanoparticulate silver using single-particle inductively coupled plasma-mass spectrometry[END_REF][START_REF] Peters | Detection of nanoparticles in Dutch surface waters[END_REF][START_REF] Yang | Analysis of silver and gold nanoparticles in environmental water using single particle-inductively coupled plasma-mass spectrometry[END_REF]. However, most studies either focus on one sampling site or several sites along the same river for which a single sampling is performed. Moreover, those studies do not try to relate changes in concentration or size of Ag-NPs to bulk geochemical parameters (i.e., pH, dissolved organic carbon concentration, ionic strength, redox ) of the aquatic ecosystems. As for any other type of natural colloids (Benedetti et al., 2003b;[START_REF] Ilina | Speciation, Size Fractionation and Transport of Trace Elements in the Continuum Soil Water Mire Humic Lake River Large Oligotrophic Lake of a Subarctic Watershed[END_REF], it is also expected that the Ag-NPs flux will vary along the river and with the hydrological and climatic variations during the year. This knowledge is critical to build better material flow analysis models that can take into account such variations if they exist, and give outputs that can be compared to real field data like the one collected in this work.

The second issue is the identification, in the field, of the mechanisms and sources that could account for the variations of measured concentrations and fluxes of Ag-NPs among different type of waters. Indeed, most of the processes reported in literature are determined from laboratory experiments [START_REF] Shevlin | Silver engineered nanoparticles in freshwater systems Likely fate and behaviour through natural attenuation processes[END_REF] or small scale column experiments to understand the fate of Ag-NPs in soils solutions and their migration towards continental waters [START_REF] Yang | Influence of natural organic matter on transport and retention of polymer coated silver nanoparticles in porous media[END_REF]. This vast literature reports that organic matter (OM), pH, ionic strength, redox conditions and the concentration of charged cations (i.e., Ca 2+ , Al 3+ ) are critical parameters that will control the number and size of Ag-NPs in such designed experiments. Only recently, [START_REF] Wimmer | New insights into the formation of silver-based nanoparticles under natural and semi-natural conditions[END_REF] showed that in field, in lake water, OM promoted the formation of AgNPs. These factors were also shown to be important for the fate of trace metals and contaminants in a watershed under a high anthropogenic pressure like the Seine River [START_REF] Bonnot | Trace metals dynamics under contrasted land uses: contribution of statistical, isotopic, and EXAFS approaches[END_REF]. The direct evidence, from field data, of the role played by such drivers is still missing. It is therefore important to explore how they will be effectively related to changes in concentration, size and flux of Ag-NPs in river watersheds.

The third issue, generally addressed in material flow analysis models but not in today published field studies, is the impact of the land-use within a watershed. For instance, for the whole Seine River watershed, an increase in trace metal concentrations from « pristine » headwaters to heavily impacted urban-waters was reported [START_REF] Chen | Anthropophile elements in river sediments: Overview from the Seine River, France[END_REF][START_REF] Grosbois | The spatial and temporal trends of Cd, Cu, Hg, Pb and Zn in Seine River floodplain deposits (1994-2000)[END_REF][START_REF] Horowitz | Variations in trace element geochemistry in the Seine River Basin based on floodplain deposits and bed sediments[END_REF], with a significant impact of Paris conurbation due to wastewater treatment plants and combined sewer overflow inputs (Estebe, A., Mouchel, J.M. Thévenot, 1998;[START_REF] Garban | Origin, retention and release of trace metals from sediments of the river seine[END_REF]. In addition, a recent field monitoring of Ag-NPs was conducted on Meuse and Ijssel watersheds which are characterized by different land-uses [START_REF] Peters | Detection of nanoparticles in Dutch surface waters[END_REF]. Based on the measured concentrations of Ag-NPs from this study, a first order estimation of the Ag-NPs normalized export-rates can be calculated using an annual average discharge rate and watershed area (source Wikipedia, 2018). We calculated the export-rates normalized per km 2 or per capita and they are equal to 0.3 g km -2 y -1 (1 mg Ag-NPs capita - 1 y -1 ) and 2.0 g km -2 y -1 (6 mg Ag-NPs capita -1 y -1 ), for the Meuse and Ijssel rivers, respectively. Using only the 10 kg of nano-engineered /year declared in the R-nano database gives a rate of release equal to 0.15 mg Ag-NPs/ /capita for France . Potential explanations for such differences in two rivers are: their different processes governing the fate of the NPs along the rivers course, their differences in soils nature and in compositions, and may be the most importantly, their differences in land-use. For instance, the Meuse basin is comprised by croplands (39%), forests (29%), pastures (18%) and built up areas (12%) by the end of the 20th century [START_REF] Lambert | Effects of human land use on the terrestrial and aquatic sources of fluvial organic matter in a temperate river basin (The Meuse River, Belgium)[END_REF]; while rural areas characterize the Ijssel watershed with intensive agricultural dominating with croplands covering 70% of the basin [START_REF] Verwijmeren | Many rivers to cross[END_REF]. Putting aside the obvious effect of the land-use for the contaminants and Ag-NPs sources mobilization, one can wonder to what extent are the normalized fluxes of Ag-NPs governed by the land-use, or if these fluxes are only regulated by the above mentioned biogeochemical drivers. The answers to these questions will help to better constrain the hypothesis made in material flow analysis models. They will also provide an accurate range of Ag-NPs environmental concentrations in combination with realistic biogeochemical parameter values to be tested in future studies on Ag-NPs toxicity toward aquatic fauna.

To address those issues, the best approach is to sample waters from low to heavily human-impacted territories and to precisely assess the seasonal dynamics of Ag-NPs under such different land-uses. The present study relies on concentration analysis for the dissolved (< 1 kDa) and nano-particulate silver in three characteristic sub-basins (i.e., forested, agricultural, and urban), selected in a previous study on the identification of trace element geochemical background values in the Seine River watershed [START_REF] Bonnot | Trace metals dynamics under contrasted land uses: contribution of statistical, isotopic, and EXAFS approaches[END_REF]. Small size sub-watersheds (< 120 km 2 ) were chosen as the water samples will be prone to rapid modifications of their biogeochemical parameters as a function of hydroclimatic and land-use changes, in order to identify their role on the Ag-NPs concentration and fate and to inter-compare export rates specific to these three land-uses. Data to built the maps are from [START_REF] Bonnot | Trace metals dynamics under contrasted land uses: contribution of statistical, isotopic, and EXAFS approaches[END_REF]. Arrows pointing to the sampling location, exact coordinates are given in the supplementary material.

Materials and Methods

Sampling sites and protocol

The three sampling sites are located on the west of Paris. Their lithology and land uses, composed of artificial surfaces, agricultural areas and natural areas, are given in Fig. 1.

Detailed information on the watersheds area, distribution of land uses and discharge rates as measured during this study are given in Table S1. A snapshot of the water geochemistry was obtained by grab sampling of the three creek waters taken each month or each two months from October 2016 to December 2017 for Ag-NPs detection, using 1 L polyethylene containers. Before field sampling, the containers were previously washed with 1 N HCl for 2 days, then rinsed three times with MQ water in the laboratory. On each sampling site, one 1 L bottle was placed horizontally at the surface (about 10 cm) of creek until filled by water and it was done within 5 to 10 seconds depending of the flow rate. For the urban watershed, the water was collected from a bridge with a bucket and the 1L bottle filed directly with the bucket in less than 5 seconds.

To evaluate the sampling effect, three additional samplings were performed for each sampling site during three field missions. Sampling replicates were individually analyzed to evaluate the contribution of sampling to the overall uncertainty. The sampling effect was evaluated by performing ANOVA test for water sampling replicates, and was not significant.

The pH, dissolved oxygen and conductivity were measured in the field using Multi 3410 Multiparameter Meter. The sensors and electrodes were calibrated before field measurements in the lab and the calibration verified in the field prior to measurements.

Various sub-aliquots were filtered in the field through 0.22 µm (cellulose acetate) for analysis of the dissolved anions (non-acidified) and cations (acidified). Triplicate subsamples were acidified with 65% HNO 3 for trace elements and rare earth elements analysis. Glass vials previously washed with acid, rinsed and pyrolyzed at 500°C were used to store samples for the analysis of dissolved organic carbon (DOC) obtained from another aliquot filtered through 0.7 µm glass microfiber filters, and acidified with 85% H 3 PO 4 . All containers and materials used for sampling were rinsed three times with river waters in the field before collection.

Unacidified, unfiltered sub-samples were immediately analyzed back to laboratory for alkalinity determination and all the others above-mentioned samples were stocked at 4°C prior to analysis. The alkalinity was measured by titration with 10 -2 mol L -1 HCl (Titrando 809, Metrohm). The DOC was analyzed with a Total Organic Carbon Analyzer (TOC-L, Shimazu). The anions were quantified with ion chromatography (Dionex ICS 1100, Thermo Scientific®). The cations were analyzed with an ICP-AES (iCAP 6200 Series, Thermo Scientific®). So-called dissolved (i.e., < 0.22 µm and < 1 kDa) trace element concentrations were determined with a HR-ICP-MS (Element II, Thermo Scientific®) and the raw data were treated with uFREASI software [START_REF] Tharaud | UFREASI: User-FRiendly Elemental dAta procesSIng. A free and easy-to-use tool for elemental data treatment[END_REF]. Immediately after field sampling, in the laboratory, an aliquot (500 mL) of the bulk 1 L sample, taken for Ag-NPs analysis, was ultra-filtered through 1 kDa (regenerated cellulose) in order to remove all the particles present in waters to get access to the fraction of free Ag ions and Ag bound to organic matter or associated to nanoparticles smaller than a few nanometers ( 1.3 nm) [START_REF] Guo | Ultrafiltration and its Applications to Sampling and Characterisation of Aquatic Colloids[END_REF]. This ultra-filtrate was further used as the matrix to prepare all standards and solutions for sp-ICPMS analysis. It aims at conserving the same environment of Ag-NPs in waters and avoiding matrix effect during measurement.

Digestion of bulk water

Additionally in 2019, the total silver was quantified after acid digestion in 9 samples, which came from the urban watershed. A sub sample of 10 mL was taken for total silver analysis in triplicates and mineralized with a mixture of 3 concentrated acids (3 mL HCl, 1 mL HNO 3 and 0.5 mL HF) in a PFA closed vessel and heated for 24 H at 105 °C. After the digestion the mixture was evaporated at 85 °C to dryness. The residue was then dissolved in 10 mL of 2% HNO 3 . Blanks were run at the same time in triplicates and PFA vessels were cleaned with a concentrated HNO 3 at 120 °C before use and rinsed 3 times with MQ water.

Single-particles ICPMS (sp-ICPMS) measurements of Ag-NPs

Instrumental parameter and standard preparation

A high resolution ICPMS -Thermo Scientific ELEMENT II was used for Ag-NPs measurements in this study. It was joined with a quartz cyclonic spray chamber and a PFA MicroFlow nebulizer. The peristaltic pump rate was set at 12 rpm generating a sample uptake flow of 0.2 mL min -1 . Certified reference material Au nanoparticles of 60 nm (RM8013, NIST) was used to determine sample loss during sample uptake (ca. transport efficiency). The Au-NP stock suspension was diluted 2.5 x 10 5 times in the ultra-filtered matrix to achieve a proper particle concentration for sp-ICPMS measurements and avoid NPs coincidence at milli-second dwell-time. Dissolved calibration curves (Au: 50 to 5000 ng L -1 ; Ag: 50 to 1000 ng L -1 ) were prepared in the ultra-filtered matrix to determine the element sensitivity in the same matrix as the water samples and thus prevent matrix effect. The analysis of Ag-NPs in river waters was carried out with 1 ms dwell time, and 1 ms settling time [START_REF] Tharaud | TiO 2 nanomaterial detection in calcium rich matrices by spICPMS. A matter of resolution and treatment[END_REF]. 10 000 data points were acquired and the isotope 107 Ag and 197 Au were measured in low resolution mode and each measurement was performed in triplicate which allows to calculate means and standard deviations. The dilution of the natural water samples was unnecessary considering the expected low concentrations in natural waters. All suspensions and solutions were freshly prepared, with the minimum delay between sampling and analysis.

Data processing

According to the literature, the transport efficiency (TE) is needed to determine the concentration and size of Ag-NPs in river waters and can be obtained by three methods, based on 1) waste collection, 2) particle size, and 3) particle frequency [START_REF] Pace | Determining Transport Efficiency for the Purpose of Counting and Sizing Nanoparticles via Single Particle Inductively Coupled Plasma Mass Spectrometry[END_REF]. Here, the second method is selected since the reference material used in this study is better certified for the particle size than the particle concentration. Taking into account the matrix effect, TE was determined independently for each type of river water. The calculation method and the formula of TE are detailed in the supplementary materials.

Before data processing can be performed, NPs must be correctly distinguished from the dissolved background. Different methods exist to determine the threshold between dissolved and particulate fractions, such as average plus 3 [START_REF] Pace | Single particle inductively coupled plasma-mass spectrometry: a performance evaluation and method comparison in the determination of nanoparticle size[END_REF], iterative algorithms with 3 [START_REF] Tuoriniemi | Size discrimination and detection capabilities of single-particle ICPMS for environmental analysis of silver nanoparticles[END_REF] or simply by using the first minimum as the boundary between dissolved fraction and NPs [START_REF] Montaño | Improvements in the detection and characterization of engineered nanoparticles using spICP-MS with microsecond dwell times[END_REF]. However, for the size range expected for Ag-NPs [START_REF] Peters | Detection of nanoparticles in Dutch surface waters[END_REF], those approaches could lead to a large number of false negatives during NPs counting (i.e., underestimation of the representative number of NPs events). Instead, we used in this study another approach to prevent such artifacts: the subtraction method. It is analogous to the deconvolution method [START_REF] Cornelis | A signal deconvolution method to discriminate smaller nanoparticles in single particle ICP-MS[END_REF], but uses the ultra-filtered natural waters as the blank. Because the ultra-filtrate was supposed to best represent the realistic dissolved background in case of natural water samples. Basically, the histogram of frequency vs intensity of the ultra-filtrate through 1 kDa is subtracted from that of the raw water sample using the same bin-width (same range of intensity for each bin of both histograms). The resulting histogram is then used to calculate the concentration and size of Ag-NPs. Calculations of particle-number concentration from the signal frequency, and of particle size from the signal intensity, are well-described in literature. Here we used the procedure developed by Pace and co-workers [START_REF] Pace | Determining Transport Efficiency for the Purpose of Counting and Sizing Nanoparticles via Single Particle Inductively Coupled Plasma Mass Spectrometry[END_REF]. First of all, Ag-NPs are detected in all waters sampled throughout the year. The concentrations range, their average and their size and the results of previous studies are summarized in Table 1. The presence of Ag-NPs can be directly observed by comparing the signal of raw water and the matrix-blank with all particles removed (i.e., filtrate through 1 kDa) (Fig. S1). Ag-NPs number concentrations range from 1.5 x 10 7 to 2.3 x 10 9 particles L -1 , with an average of 9.1 x 10 8 particles L -1 . Their mass concentrations are in the range of 0.4 -28.3 ng L -1 , with an average of 7.9 ng L -1 . Compared with the scare database of Ag-NPs in surface waters (Table 1), the particle number concentrations found here are in a larger range and the average is 100 times higher than that of Dutch rivers [START_REF] Peters | Detection of nanoparticles in Dutch surface waters[END_REF]. It mostly results from the different data processing approach applied. The subtraction method in our study allows to better constrain the false negatives, where the small Ag-NPs are counted as the dissolved background. Regarding the mass concentration, our results cover the concentration range of the recent study in surface waters of Netherland. They are in agreement with previous measurements of colloidal Ag in Texas Rivers. Moreover, from the modeling view, Ag-NPs are expected in sur-face waters between 0.04 to 619 ng L -1 (Table 1). Our measured concentrations are within this global range, but a bit higher than most estimations for Europe (0.002 to 2.3 ng L -1 ) (Dumont and Williams, 2015). However, the existence of naturally occurring Ag-NPs is not included in the concentration estimation by modeling, resulting in an underestimation in that case. Also, for now, the challenges raised by all modeling studies in natural rivers come from the complex transformations undergone by silver nanoparticles on one hand, and their interactions with environmental components on the other hand. Thus, the field data obtained here will constitute a strong constraint to better refine the possible control factors in surface waters.

Results and Discussion

Ag-NPs concentrations in three waters

Among the three waters in this study, the largest Ag-NPs amount is found in the urban site, with an average concentration of 1.2 x 10 9 particle L -1 and of 12.3 ng L -1 (Table 1).

Ag-NPs mass concentrations in agricultural and forested waters are very close, considering the significant variations measured over the year. However, the forested water has a higher particle-number concentration than the agricultural water. This is mainly ascribed to their different size populations of Ag-NPs (Table 1). The size might be related to the water geochemistry affecting processes of Ag-NPs in this environment, which is to be discussed in more details in following sections.

Monthly sampling was performed on each catchment to investigate temporal variations of the Ag-NPs concentration over the year. The particle-number concentrations are presented in Fig. 2. It shows that the flux in the urban site is quite constant throughout the whole year, but the two other sites have larger temporal variations since average annual water flows are: 22 ± 9 L/sec and 51 ± 27 L/sec for the forested and agricultural creeks, respectively.

The larger variations in water flow rates reported for the agricultural and the forested sites account therefore for the temporal variability showed in Figure 2. Indeed, forested and agricultural sites, as small watersheds, react very quickly to the climatic and hydrological effects, whereas the urban watershed of greater size has diverse contributions (mixture of artificial, agricultural and natural areas) and its concentration variation can thus be slighter over time. Additionally, an increase of Ag-NPs from winter to spring is observed for forested and agricultural waters. The dissolved silver (< 1 kDa) is also quantified by sp-ICPMS in order to investigate the speciation of silver in three natural waters. The signal intensity in counts per second of ultra-filtered samples is converted into ng L -1 , using the dissolved Ag sensitivity calibration curve. Fig. 3 shows the concentrations of dissolved and nano Ag in three waters. It should be noted that values under the limit of quantification (2.5 ng L -1 ) are not discussed here. For most forested watershed water samples more nano silver is measured than dissolved, this is in agreement with results of [START_REF] Delay | Interactions and stability of silver nanoparticles in the aqueous phase: Influence of natural organic matter (NOM) and ionic strength[END_REF]. showing that NOM significantly influences the particle size distribution, the stability and the surface properties of Ag-NPs in the aqueous phase preventing it further dissolution, and that in presence NOM, Ag-NPs suspensions are also stable at variable ionic strength, which will facilitate their transport. In addition it was recently shown that natural organic matter content (i.e. 10 mgL -1 NOM) in lake water encourage the natural silver-based nanoparticle formation [START_REF] Wimmer | New insights into the formation of silver-based nanoparticles under natural and semi-natural conditions[END_REF].However, in agricultural and urban waters, the distribution changes from one sampling period to another. No conclusion has been made for these two sites, in which the transformation of silver is more complex and may be controlled by multiple parameters and processes (size and coating of Ag-NPs, pH, DOC, IS, dissolved oxygen, light irradiation, biofilm ) [START_REF] Desmau | Dynamics of silver nanoparticles at the solution/biofilm/mineral interface[END_REF][START_REF] Shevlin | Silver engineered nanoparticles in freshwater systems Likely fate and behaviour through natural attenuation processes[END_REF].

The total silver quantified in 2019 after acid digestion of the 9 remaining samples, from the urban watershed are given in Table S4. The total amount of Ag is higher than the addition of nanosized and dissolved fractions in almost all urban waters, except the sample of 2016-11-30. This is most probably related to the ageing water samples since those samples were stored in the refrigerator with no prior acidification for two or three years, part of the silver may have been lost to the container wall. Besides, when eliminating sample of 2016-11-30, the average of percentage of measured Ag (nano + dissolved) is equal to 70%. It means that a part of Ag is incorporated in larger particles or aggregates not detected by sp-ICPMS.

Field mechanisms and the sources accounting for the variations of measured concentrations

Correlation between elements and Ag-NPs concentrations (i.e., number or mass) can be interpreted within a mass balance approach or as the identification of driving parameter or process for small watersheds (Benedetti et al., 2003a). All related data in the following discussion (pH, DOC, conductivity, IS and elements concentrations in the dissolved fraction) are detailed in supplementary material (Table S2 andS3). The correlation between two variants is considered statistically significant when R 2 0.6 and p-value 0.01. There is a positive correlation between DOC concentrations and Ag-NPs concentrations in forested water (Fig. 4 andS2). been described in several studies [START_REF] Topuz | Agglomeration of Ag and TiO2nanoparticles in surface and wastewater: Role of calcium ions and of organic carbon fractions[END_REF][START_REF] Yu | Interactions between engineered nanoparticles and dissolved organic matter: A review on mechanisms and environmental effects[END_REF][START_REF] Zhang | Role of Secondary Particle Formation in the Persistence of Silver Nanoparticles in Humic Acid Containing Water under Light Irradiation[END_REF][START_REF] Wimmer | New insights into the formation of silver-based nanoparticles under natural and semi-natural conditions[END_REF][START_REF] Huang | Significant contribution of metastable particulate organic matter to natural formation of silver nanoparticles in soils[END_REF]. High NOM concentration inhibits the dissolution and aggregations of Ag-NPs when released into surface waters, thus making Ag-NPs more persistent in such natural matrices. The formation of Ag-NPs through reduction of Ag + by NOM can also lead to the increase of Ag-NPs concentration along with NOM [START_REF] Akaighe | Humic Acid-Induced Silver Nanoparticle Formation Under Environmentally Relevant Conditions[END_REF][START_REF] Sal Nikov | Silver ion reduction with peat fulvic acids[END_REF]Yin et al., 2012a[START_REF] Wimmer | New insights into the formation of silver-based nanoparticles under natural and semi-natural conditions[END_REF]. As reported in Fig. 3 there is a large reservoir of dissolved Ag, either as free Ag + or Ag bound to NOM to justify the potential increase of Ag-NPs by this process. For the agricultural and urban watersheds, no significant trends between DOC and Ag-NPs number or mass concentrations were found. This lack of trend does not mean that similar processes are not at work, but that the smaller range of NOM concentrations variations NOM = 2.2 mg L -1 and 1.7 mg L -1 versus 5.7 mg L -1 for agricultural and urban versus forested watersheds, respectively, prevented the building of a reliable correlation. Moreover, other potential processes which work more efficiently might be operating in those two watersheds. Indeed, we found that the particle-number concentration and the equivalent size for sphere of Ag-NPs correlate negatively with conductivity, and positively with Ca concentration, in agricultural and urban water samples, respectively (Fig. 5 and6). This means that Ag-NPs number concentration decreases with increasing charged ions concentration in the aquatic media. The agricultural and urban waters have higher concentrations of charged ions, while their average particle size is significantly higher than Ag-NPs from the forested waters (Table 1). The charge-induced aggregation, and even sedimentation, of silver nanoparticles has been reported in simplified natural systems with the presence of calcium ions [START_REF] Topuz | A systematic evaluation of agglomeration of Ag and TiO2nanoparticles under freshwater relevant conditions[END_REF]. These aggregates will interact with colloids or induce further sedimentation [START_REF] Labille | Heteroaggregation of titanium dioxide nanoparticles with natural clay colloids[END_REF], thus escape from the toplayer sampling since the sampling river surfaces are quite small and water column is shallow (< 0.3 m). In the case of the agricultural water samples the lack of increase in Considering negative correlation between Ag-NPs concentration and conductivity, together with size growing with Ca 2+ content, we can conclude that in the chargeabundant river, the aggregation or even sedimentation is favorable and the fate is more controlled by charge determining ions such as divalent cations (i.e., Ca 2+ and Mg 2+ ) or Al 3+ in extreme conditions with very low pH.

Another explanation for this correlation is that the observed changes are related to variable inputs of different sources with time for instance the urban watershed is partly covered by agricultural land. But the differences in bulk geochemistry, especially for major ions, are not that large for a single river (Table S3) and cannot totally explain the observed trend. However, correlations can also help to identify sources of elements and colloids (Benedetti et al., 2003a). For instance, for the forested site water samples, there is a andS3). Zr, is often considered as a geological background indicator because of its limited mobilization under most environmental conditions [START_REF] Aja | The aqueous geochemistry of Zr and the solubility of some Zr-bearing minerals[END_REF][START_REF] Hodson | Experimental evidence for mobility of Zr and other trace elements in soils[END_REF]. As for La and Ce, their speciation is generally controlled by natural organic matter [START_REF] Sonke | Lanthanide Humic Substances Complexation. II. Calibration of Humic Ion-Binding Model V[END_REF], which explains their good correlation with Ag-NPs number and mass concentrations.

Moreover, the Ce/La ratio is equal to 1.9 ± 0.1 for all water samples in the forested watershed, this value is equal to the geochemical background value of 1.9 given for French soils and suspended matter sediments [START_REF] Reimann | GEMAS: Establishing geochemical background and threshold for 53 chemical elements in European agricultural soil[END_REF]. Thus, silver nanoparticles in forest water are very likely to be of natural origin. In addition to the natural leaching from bedrock [START_REF] Qi | Concentration and distribution of trace elements in lignite from the Shengli Coalfield, Inner Mongolia, China: Implications on origin of the associated Wulantuga Germanium Deposit[END_REF], other potential pathways of naturally occurring Agcontaining NPs include the reduction of dissolved Ag + by natural organic matter [START_REF] Akaighe | Humic Acid-Induced Silver Nanoparticle Formation Under Environmentally Relevant Conditions[END_REF][START_REF] Sal Nikov | Silver ion reduction with peat fulvic acids[END_REF]Yin et al., 2012a;[START_REF] Wimmer | New insights into the formation of silver-based nanoparticles under natural and semi-natural conditions[END_REF], the oxidation and reduction of macroscale Ag objects [START_REF] Qi | Concentration and distribution of trace elements in lignite from the Shengli Coalfield, Inner Mongolia, China: Implications on origin of the associated Wulantuga Germanium Deposit[END_REF], and biological processes [START_REF] Klaus | Silver-based crystalline nanoparticles, microbially fabricated[END_REF][START_REF] Kroll | Extracellular polymeric substances (EPS) of freshwater biofilms stabilize and modify CeO2 and Ag nanoparticles[END_REF]. Moreover, the stability of Ag-NPs in the environment could be enhanced by surface sulfidation [START_REF] Levard | Sulfidation processes of PVP-coated silver nanoparticles in aqueous solution: Impact on dissolution rate[END_REF], via interactions with extracellular polymeric substances (EPS) [START_REF] Kroll | Extracellular polymeric substances (EPS) of freshwater biofilms stabilize and modify CeO2 and Ag nanoparticles[END_REF][START_REF] Li | Mechanistic understanding of reduced AgNP phytotoxicity induced by extracellular polymeric substances[END_REF] and with NOMs [START_REF] Topuz | Agglomeration of Ag and TiO2nanoparticles in surface and wastewater: Role of calcium ions and of organic carbon fractions[END_REF][START_REF] Yu | Interactions between engineered nanoparticles and dissolved organic matter: A review on mechanisms and environmental effects[END_REF][START_REF] Zhang | Role of Secondary Particle Formation in the Persistence of Silver Nanoparticles in Humic Acid Containing Water under Light Irradiation[END_REF] present in the natural matrix.

Le [START_REF] Pape | Trace element behavior and partition versus urbanization gradient in an urban river (Orge River, France)[END_REF] identified Antimony (Sb) as a typical anthropogenic element in the Orge river catchment, very close to our sampled watersheds. They concluded that Sb was of anthropogenic origin and carried by organic matter or sulfur containing species.

In the agricultural watershed we have a good correlation between Sb and Ag-NPs number concentration (Fig. 8). This could mean that both elements were trapped in or carried by the same species. Organic matter or sulfur containing species could originate from biosolids recycled from wastewater treatment plants (WWTPs) and used as fertilizers and soil amendments [START_REF] Healy | Metal concentrations in lime stabilised, thermally dried and anaerobically digested sewage sludges[END_REF][START_REF] Pape | Trace element behavior and partition versus urbanization gradient in an urban river (Orge River, France)[END_REF]. Several studies have confirmed the release of Ag-NPs in the effluent of WWTPs, at the level from < 12 ng L -1 up to 100 ng L -1 [START_REF] Li | Quantification of nanoscale silver particles removal and release from municipal wastewater treatment plants in Germany[END_REF][START_REF] Mitrano | Detecting nanoparticulate silver using single-particle inductively coupled plasma-mass spectrometry[END_REF]. . An historical pollution for Sb and As as well as Ag is reported and well known in the Seine river watershed [START_REF] Ayrault | Silver and thallium historical trends in the Seine River basin[END_REF][START_REF] Cloarec | Sedimentary record of metal contamination in the Seine River during the last century[END_REF]. We cannot exclude this old source that because of its dispersion via the atmosphere all over the catchment has probably reached the agricultural watershed under investigation. Therefore, it suggests that silver nanoparticles in agricultural water would be of anthropogenic origin.

In the case of urban river water samples, no clear correlations with the previously mentioned drivers such as organic matter or trace elements issued from anthropogenic activ- [START_REF] Gall | Investigating the metal contamination of sediment transported by the 2016 Seine River flood (Paris, France)[END_REF]. Thus new source of silver such as biocidal plastics and textiles [START_REF] Vance | Nanotechnology in the real world: Redeveloping the nanomaterial consumer products inventory[END_REF] are uncorrelated to those historical trace pollutant elements as seen in this study. In addition, the mixed contribution of three type of waters (forested 19%, agricultural 44% and artificial 37%) makes complex the interpretation of the source identification especially in the case of small watersheds that promptly react to hydrological and environmental changes. A weak trend is observed between Ag-NPs number concentration and the temperature of the sampling river (r 2 = 0.48, p = 0.0377). The higher the temperature the higher the number of Ag-NPs. Temperature and sunlight, affects the reductive formation of Ag-NPs [START_REF] Yin | Photoreduction and stabilization capability of molecular weight fractionated natural organic matter in transformation of silver ion to metallic nanoparticle[END_REF](Yin et al., , 2012b)). In summer, warmer water temperature months, increased exposition to UV-light would promote the reduction of Ag + by organic compounds leading to higher Ag-NPs concentrations [START_REF] Badireddy | Formation of Silver Nanoparticles in Visible Light-Illuminated Waters: Mechanism and Possible Impacts on the Persistence of AgNPs and Bacterial Lysis[END_REF][START_REF] Odzak | Influence of daylight on the fate of silver and zinc oxide nanoparticles in natural aquatic environments[END_REF][START_REF] Yin | Water chemistry controlled aggregation and photo-transformation of silver nanoparticles in environmental waters[END_REF]. Similar enhanced process by sunlight takes place for the reduction-formation Ag-NPs by natural EPS components [START_REF] Kroll | Extracellular polymeric substances (EPS) of freshwater biofilms stabilize and modify CeO2 and Ag nanoparticles[END_REF]. They thus result in the persistence of Ag-NPs in sunlight waters.

The issue of the land-use effect is still standing even after identifying potential sources for the forested watershed samples (natural) and the agricultural watershed (contamination). Calculation of normalized fluxes of Ag-NPs that can be obtained by the combination of the monthly water discharge data and the respective area of the three watersheds may help to evidence clear differences between each of the studied sites. The measured normalized (in km -2 y -1 ) export rate for Ag-NPs and dissolved silver are given in Fig. 9. It is clear in this figure that the flux from the urban site is the highest for nano and dissolved silver. If the monthly export rates of Ag-NPs are averaged, an annual export rate of 2.3 ± 1.3 g km -2 y -1 is obtained. This value is ten times higher than the export rate calculated for the agricultural and forested watersheds whose respective values are 0.4 ± 0.3 and 0.2 ± 0.2 g km -2 y -1 . The difference between the urban site and the two others is also seen for dissolved Ag export rates that are respectively equal to 2.6 ± 3.1, 0.6 ± 0.5 and 0.2 ± 0.1 g km -2 y -1 for urban, agricultural and forested sites. These large difference cannot be due to due to the constant release of Ag-NPs from consumer products into freshwaters as reported in several studies, [START_REF] Benn | Nanoparticle silver released into water from commercially available sock fabrics[END_REF][START_REF] Mitrano | Presence of nanoparticles in wash water from conventional silver and nano-silver textiles[END_REF]; [START_REF] Farkas | Characterization of the effluent from a nanosilver producing washing machine[END_REF]; [START_REF] Aznar | Quantification and size characterisation of silver nanoparticles in environmental aqueous samples and consumer products by single particle-ICPMS[END_REF][START_REF] Kaegi | Release of silver nanoparticles from outdoor facades[END_REF]. The amount of nanosized silver registered in the R-nano French declaration of use or production site (https://www.r-nano.fr/) is equal to 10kg in 2017 and was equal to 1 kg in 2013.

Effect of Land-use on exportation fluxes

In addition for France, in the https://www.nanotechproject.org/cpi/ data base of products with nano-sized silver, only 33 products are listed. Another approach is to use the data of [START_REF] Wang | Engineered silver nanoparticles in terrestrial environments: a meta-analysis shows that the overall environmental risk is small[END_REF] for

Europe and downscale it to France, using Europe and France population as the scaling factor, we can estimate that 27 kg Ag-Nano y -1 (i.e., 0.042 g km -2 y -1 ) and 425 kg Ag-Nano y -1 (i.e. 0.66 g km -2 y -1 ) could be delivered to soils and waters, respectively. These are very small numbers for the French territory. Other sources of Ag-NPs must be found to account for the fluxes in all three watersheds. In table S4 we have compiled the known input data of bulk Ag and Nano Ag to soils for the world, Europe, France and Switzerland with available data from (Johnson et al., 2005;Muller andNovack, 2008 and[START_REF] Wang | Engineered silver nanoparticles in terrestrial environments: a meta-analysis shows that the overall environmental risk is small[END_REF]. We can clearly see that bulk Ag, issued mostly from WWTPs, represents a much bigger source of Ag to soils and consequently to rivers or aquatic ecosystems than the input of nano-Ag [START_REF] Wang | Engineered silver nanoparticles in terrestrial environments: a meta-analysis shows that the overall environmental risk is small[END_REF].

our year of survey and could explain part of the discrepancy as we are missing in our calculations all the material arriving during the flood events. Thus despite the increased use of Ag-NPs consumer products [START_REF] Benn | Nanoparticle silver released into water from commercially available sock fabrics[END_REF][START_REF] Farkas | Characterization of the effluent from a nanosilver producing washing machine[END_REF][START_REF] Kaegi | Release of silver nanoparticles from outdoor facades[END_REF] their contribution to the general flow is still limited (i.e., 7% Table XX), more long term surveys are needed to confirm this result."

Conclusion

Our results clearly demonstrate that material flow analysis models should generate export rates that are land-use dependent in order to have improved predicting capabilities.

Ag-NPs in three surface waters were found in the range of 1.5 x10 7 to 2.3 x 10 9 particles L -1 at number concentration and 0.4 to 28.3 ng L -1 at mass concentration. In addition, some driving process factors and potential sources were identified by using correlations between Ag-NPs concentrations and other parameters, like the control of natural organic matters of Ag-NPs in the forested water. Whereas, various contributions and more complex process were found in agricultural and urban waters. Besides, the export rate of Ag-NPs from artificial, agricultural and forested areas were respectively, 5.5 ± 3.0, 0.5 ± 0.3 and 0.2 ± 0.2 g km -2 y -1 .

In addition, our study also demonstrates that the dissolved compartment should be taken into account in material flow analysis and toxicity models since it is the most reactive as well as the most toxic. Future studies should be focused on high resolution samplings during a given specific month when climatic factor varies a lot and when farming and/or hunting activities are active to better constrain the sources of the Ag-NPs as well as their fluxes.

Figure 1 :

 1 Figure 1: Characteristics of three sampling catchments (Agriculture, Forested and Urban) in the Seine River watershed of France.

Table 1 .

 1 Concentration and size of Ag-NPs measured in surface waters from the literature and present study measurements. PEC: Predicted Environmental Concentrations in surface water values taken from Peters et al., 2018. *10% most exposed rivers according to Dumont et al., 2015. n.g. not given; n.a. not analyzed.

Figure 2 .

 2 Figure 2. Flux of Ag-NPs measured in three waters for one-year sampling (except two contaminated cases). Error barres represent the standard deviation (SD) of three measurements.

Figure 3 .

 3 Figure 3. Mass concentration of nano (dark color) and dissolved (light color) silver measured in three waters. Forested, agricultural

Figure 4 .

 4 Figure 4. Mass concentration of Ag-NPs as function as DOC concentration in water samples from the forested watershed.

Figure 5 .

 5 Figure 5. Number concentration of Ag-NPs as function as conductivity in water samples from the agricultural watershed.

  2+ could be due to fast sedimentation or formation of nanoparticles outside the range of our analytical detection window.

Figure 6 .

 6 Figure 6. Size of Ag-NPs as function as the concentration of dissolved Ca2+ (< 0.22 µm) in water samples from the urban watershed.

Figure 7 .

 7 Figure 7. Mass concentration of Ag-NPs as function as the concentration of total Zr and La in water samples from the forested watershed.

  the Ag-NPs mass concentration and trace elements (Zr, Y, La and Ce) in the fraction < 0.22 µm (two examples are given in Fig.7

Figure 8 .

 8 Figure 8. Number concentration of Ag-NPs as function as the concentration of total Sb in water samples from the agricultural watershed.

  . A recent evaluation on historical trace pollutants load (Sb, Ag, Ni, Cu, Zn, As, Pb) in the Seine River watershed reports lower values in sediments reflecting the progressive decontamination of the catchment (Le

Figure 9 .

 9 Figure 9. Export rate of nano (a) and dissolved (< 1 kDa) (b) Ag in three catchments throughout the sampling year. Error bars

Acknowledgment

Material flow analysis, expressed in mass per capita per year or in mass per km 2 per year, relies on land-use or product use to simulate concentrations of Ag-NPs in aquatic systems. As two of the three watersheds are experiencing different land-uses (i.e., forestry, agricultural practices and artificial areas), we can recalculate the actual specific export rate of Ag-NPs for agricultural practices and artificial areas. The forested watershed (0.2 ± 0.2 g km -2 y -1 or 80 mg.capita -1 .y -1 ) is one end-member that will allow the calculation of the agricultural end-member since the agricultural basin is only covered by those two types of activities (i.e., 72% agriculture and 22% forestry). The calculated rate of export of Ag-NPs is therefore equal to 0.5 ± 0.3 g km -2 y -1 or 5mg.capita -1 .y -1 for land only used for agricultural purposes. Now the same calculation can be done for the artificial areas using the urban watershed data (i.e., 37% artificial areas, 44% agriculture 19% forestry). The calculated rate of export of Ag-NPs is therefore equal to 5.5 ± 3.0 g km -2 y -1 or 3mg.capita -1 .y -1 for artificial areas.

Considering the large uncertainties due to the high variability of Ag-NPs concentrations along the year for artificial land and agricultural practices we can conclude that both activities generate the same amounts of Ag-NPs expressed in mg.capita -1 .y -1 , the highest of all being the forested watershed (i.e. 80 mg.capita -1 .y -1 vs 5 mg.capita -1 .y -1 ). The major characteristic of forested watershed is that its lithology corresponds to a sandy soil vs. carbonated soils for the other watersheds and that it is richer in carbon in the top-soil [START_REF] Bonnot | L origine des m taux et la d namique du inc dans le bassin de la Seine[END_REF]. Recent studies have shown that soils organic matter could significantly contribute to the formation of silver nanoparticles in soils [START_REF] Huang | Significant contribution of metastable particulate organic matter to natural formation of silver nanoparticles in soils[END_REF] . The low number of inhabitants and the favored production of Ag-NPs by soils organic matter would account for this higher number.

The origin of the Ag-NPs for the agricultural and urban watersheds is probably different from the geochemical background and is from anthropogenic origin since in the Seine river watershed, biosolids were and are used on agricultural land [START_REF] Buzier | Inputs of total and labile trace metals from wastewater treatment plants effluents to the Seine River[END_REF][START_REF] Thévenot | Critical budget of metal sources and pathways in the Seine River basin (1994-2003) for Cd, Cr, Cu, Hg, Ni, Pb and Zn[END_REF]. Artificial areas will generate the highest flux of Ag-NPs by more than one order of magnitude compared to the two other land uses when expressed in g km -2 y -1 . The CORINE Land Cover Map (updated in 2006) can be used to calculate the total Ag-NPs annual export for the entire catchment of the Seine River (i.e., 78650 km 2 ) using two end-members. An urban one corresponding to 13% of the total catchment (Ag-NPs = 5.5 ± 3.0 g km -2 y -1 ) and an average value corresponding to the agricultural and forested end-members, which represents the rest of the land use in the catchment (Ag-NPs = 0.4 ± 0.3 g km -2 y -1 ). The calculated flux is equal to 80 ± 51 kg y -1 corresponding to a release per capita of 4 mg.capita -1 .y -1 . The urban source representing 56 ± 30 kgy -1 . The same calculation can be done for the dissolved fraction of Ag with the same hypothesis. The dissolved flux of Ag is 94 ± 50 kg y -1 with the urban source representing 64 ± 30 kg y -1 . In total a flux of 174 kg y -1 of silver would be exported from the catchment dissolved and nano. This flux is much smaller than the particulate flux calculated by Ayrault equal to 1700 kgy -1 in 2003 [START_REF] Ayrault | Silver and thallium historical trends in the Seine River basin[END_REF] corresponding to a flux of 85 mg.capita -1 .y -1 . Table S4 data show a huge input to soils of bulk silver in France compared to the nano Silver (whatever the units used) and suggest that the difference between our seine river flux and the one from [START_REF] Ayrault | Silver and thallium historical trends in the Seine River basin[END_REF] could be due to a missing source input not captured by our 3 watersheds but existing at the scale of the larger seine river watershed. For instance the cause could be higher inputs of biosolids from WWTPs on agricultural land (Thevenot et al., 2007;[START_REF] Buzier | Inputs of total and labile trace metals from wastewater treatment plants effluents to the Seine River[END_REF] since we do not have the mass of biosolids that are actually spread over the lands in our three sites. In addition the data from [START_REF] Ayrault | Silver and thallium historical trends in the Seine River basin[END_REF] was calculated with data from sediments of two cores about 100 km downstream of Paris draining 96% of the Seine River basin and are located within the last major meander of the Seine River before it reaches its estuary. This location has been flooded at least once each year until 2004 [START_REF] Ayrault | Silver and thallium historical trends in the Seine River basin[END_REF]. Flood events were never recorded in our small watersheds during Electronic Supplementary Material (for online publication only) Click here to download Electronic Supplementary Material (for online publication only): Supplementary materials V3.0.docx


L M Lechuga

J Tamayo

M Alvarez

L G Carrascosa

A Yufera

R Doldan

Zinoviev

M Álvarez

R Doldán

E Peralías

A Rueda

J A Plaza

K Zinoviev

C Domínguez

A Zaballos

M Moreno

C Martínez-A

D Wenn

N Harris

C Bringer

V Bardinal

T Camps

C Vergnenègre

C Fontaine

V Díaz

A highly

Keywords: biosensor, microsystem, Si technology, microcantilevers, DNA biochip

HAL is

Introduction

There is an increasing interest in systems based on micro/nanotechnologies for ultrasensitive and miniaturised biosensors and biochips. The application of this new microand nanotechnology techniques to microarray production will result in both miniaturization of biochip and an increase in sensitivity of the assays. One of the most promising approaches in micro/nanobiosensors is in the area of cantilever based sensors, which are a new class of highly sensitivity devices able of performing local, high resolution and label-free molecular recognition measurements [START_REF] Mckendry | Multiple label-free biodetection and quantitative DNA-binding on a nanomechanical cantilever array[END_REF][START_REF] Arntz | Label-free protein assay based on a nanomechanical cantilever array[END_REF]. These sensors derive from the microfabricated cantilevers used in atomic force microscopy [START_REF] Giesbers | The Electrical Double Layer on Gold Probed by Electrokinetic and Surface Force Measurements[END_REF] and are based on the bending induced in the cantilever when, for example, a biomolecular interaction takes place on one of its surfaces. The microcantilevers translate the molecular recognition of biomolecules into a nanomechanical motion [START_REF] Fritz | Translating biomolecular interactions into nanomechanics[END_REF] (from a few to hundreds of nanometres), that is commonly coupled to an optical or piezoresistive read-out system [START_REF] Lavrik | Cantilever transducers as a platform for chemical and biological sensors[END_REF][START_REF] Raiteri | Micromechanical cantilever-based biosensors[END_REF][START_REF] Datskos | Micro and Nanocantilever Sensors[END_REF] (see Fig. 1 for details). These devices can specifically detect single-base mismatches in oligonucleotide hybridisation [START_REF] Fritz | Translating biomolecular interactions into nanomechanics[END_REF] among other interesting applications. In these assays [START_REF] Lavrik | Cantilever transducers as a platform for chemical and biological sensors[END_REF], nucleic acids are immobilized only on one side of a micromachined cantilever [START_REF] Duguid | How do DNA repair proteins locate potential base lesions? A chemical crosslinking method to investigate O-6-alkylguanine-DNA alkyltransferases[END_REF]. Exposure of the cantilever to a sample containing complementary nucleic acid gives rise to a cantilever bending [START_REF] Appiah-Kubi | The effect of varying salt concentrations on the response of microcantilevers[END_REF] of a few nanometers. The nanomechanical response is due to the surface stress change of the active side with respect to the other side, in which DNA is not immobilized [START_REF] Alvarez | Development of nanomechanical biosensors for detection of the pesticide DDT[END_REF]. Surface stress arises from the electrostatic, van der Waals and steric intermolecular interactions on the surface [START_REF] Alvarez | Nanomechanics on the formation of DNA self-assembled monolayers and hybridisation on microcantilevers[END_REF]. The deflection is measured with sub-nanometer resolution by an optical system in which a laser beam reflects off the surface of the cantilever and is detected by a custom position sensitive photodetector.

In this paper we show the development of a portable multibiosensor microsystem based on microcantilever sensors able to detect hybridization of nucleic acids partly complementary to human genes with high sensitivity. The biosensor microsystem consists of: i) an array of twenty microcantilevers as biosensing tranducers, ii) an optical detection system (lasers and photodetectors) capable of measuring the cantilever movement with subnanometer resolution and iii) a polymer microfluidic system for reagents delivery to the sensor surface [12].

In order to obtain the biosensor microsystem the main steps to be accomplished are: the fabrication of arrays of 20 microcantilevers, the development of a self-assembly chemistry for the DNA probe immobilization, the fabrication of a polymer microfluidic header integrated with the cantilever array, the development of an optical detection subsystem with a monolithic array of 20 single mode oxide-confined VCSELs and a coupling element, the integration of an array of 20 segmented photodetectors and CMOS circuits for the detection and processing of the laser beam displacement, and finally, the assembly of the different components with an accurate alignment. To our knowledge, this device will be the first biosensor microsystem based on nanomechanics, in which sensors, optical detection system and microfluidics will be integrated making available this device for clinical testing. The microsystem is intended to be used initially for the detection of mutations and polymorphisms of the BRCA1 human gene involved in the development of breast cancer, but has many other applications in the biodiagnostic and environmental areas.

Fabrication of the microcantilever array transducer

Silicon, silicon nitride and silicon oxide cantilevers are available commercially with different shapes and sizes in analogy to AFM cantilevers. The reported technologies allow the fabrication of ultra thin cantilevers of up to 12 nm thick [START_REF] Kawakatsu | Millions of cantilevers for atomic force microscopy[END_REF] and millions of them on one wafer [START_REF] Li | Ultrathin single-crystalline-silicon cantilever resonators: Fabrication technology and significant specimen size effect on Young's modulus[END_REF] with typical lengths between 10 and 500 µm. However, for a specific biosensor application the cantilever geometry, dimensions and materials have been modelled in order to obtain a highly sensitive nanomechanical response with a small sensor area. The thickness, length, and width of the cantilever to be fabricated are derived from the modeling using Finite element simulations (ANSYS). Reducing the thickness and increasing the length results in increment of sensitivity of the device but also leads to complexity in fabrication technology.

A fabrication technology of single-crystalline-silicon thin cantilevers using commercially available Silicon-On-Insulator (SOI) wafers has been developed and optimized with the objective to achieve 100% yield of cantilevers per wafer. It has been possible to fabricate arrays of thousands of identical cantilevers on one wafer. The technology allows to fabricate arrays of 20 cantilevers with discrete and common windows, each being 200 or 100 μm long, 40 μm wide, and 340 nm thick (see Figure 2a).

The sequence of the fabrication technological steps is shown in Fig. 2B. The front side contained a structural silicon layer (1) and a sacrificial silicon oxide layer [START_REF] Arntz | Label-free protein assay based on a nanomechanical cantilever array[END_REF]. The back side of the silicon substrate (3) has a silicon oxide layer (4). As a first step, the back side was covered with an aluminium layer (5) as a mask material for DRIE. The initial multi-layer structure used for the fabrication of cantilevers is presented in Fig. 2a. Afterwards, cantilevers on the front side were formed by dry etching of silicon layer through the pattern obtained by photolithography. Finally, the silicon oxide was removed from the areas where the silicon substrate must be etched. Before the next step the components side was covered with a photoresist layer [START_REF] Raiteri | Micromechanical cantilever-based biosensors[END_REF]. Etching the holes in 450 µm thick silicon substrate with DRIE takes about one hour. The etching gives almost vertical walls in the cavities with small, about 20 µm, lateral undercut at the top. At this stage, clean silicon oxide membranes with cantilevers and photoresist layer on top were obtained (see Fig. 2B). The photoresist layer 6 prevented the membranes of cracking and allows to obtain 100 % yield of the production (among 2500 cantilevers per wafer). The membranes were etched in vapours of HF (49%) to release the cantilevers. Afterwards wafers were shortly rinsed with DI water. Finally, the photoresist film was removed by oxygen plasma etching.

The mechanical properties of the cantilevers have been experimentally determined (see Table 1). The spring constant of cantilevers were evaluated using a calibrated AFM probe. In addition, the harmonic characterisation of the cantilevers was also done. The experimental results match with the simulated ones when the overetching of the cantilevers is considered.

The measured resonance frequencies are also in good agreement with the calculated ones. The Q factors are a little smaller than the Q factor of commercial ones, however this is in good agreement with the results obtained in published works for very thin cantilevers.

Immobilization and regeneration protocols.

The oligonucleotides receptors must be bound to the cantilever surface by an extremity and separated from the surface to facilitate the duplex formation of the nucleotides closer to the surface. This has been achieved through the development of a self-assembly chemistry which allows the addressing of DNA immobilization at the nanometer-scale level only on one side of the cantilever array. The density will be tuned by using spacer molecules with a terminal group that inhibits non-specific adsorption. The optimised immobilisation procedure is based on thiolated self-assembled monolayers. Figure 3 shows the bending of the cantilevers observed during the immobilisation of the oligonucleotide receptors. Due to the attractive interaction between the nucleotide chain and the gold, the most favourable structure for adsorbed thiolated DNA monolayer is the DNA strongly anchored at the surface by the thiol linker and with the nucleotide chain in contact with the surface. To enhance the accessibility of the DNA probes to the target and assures that the DNA probes are only attached to the gold surface through the terminal sulphur atom of the thiol linker, the surface was exposed to 6mercapto-1-hexanol (MCH) after DNA immobilization (see Figure 3). MCH adsorption produces a large compressive surface stress, giving a downwards bending of about 30-35% of the cantilever deflection resulting from the thiolated DNA immobilisation [START_REF] Alvarez | Nanomechanics on the formation of DNA self-assembled monolayers and hybridisation on microcantilevers[END_REF].

To understand the effects of the MCH treatment, we performed radiolabelling experiments, (see Figure 4) using the thiol modified DNA radioactively labelled. The quantification of the immobilized DNA before and after the MCH treatment was carried out with a scintillation counting. The results indicate that the MCH molecules displace about 60% of the immobilized DNA, suggesting that only 40% of the immobilized chains are covalently attached to the gold surface of the cantilever [START_REF] Alvarez | Nanomechanics on the formation of DNA self-assembled monolayers and hybridisation on microcantilevers[END_REF].

Microfluidics system.

A polymer microfluidic flow cell has been developed which integrates with the cantilever array. The flow cell incorporates a single flow channel to allow the delivery of reagents to the array of cantilevers, and subsequent removal to waste. The flow cell is required to be a lowcost disposable item. It consists of three acrylic layers and two PDMS gaskets. The acrylic layers are directly machined on a Datron CAT 3D M6 micromill and include recesses for the gaskets and the cantilever chip. In order to produce the PDMS gaskets, a mould tool was machined directly into brass. The PDMS is mixed and then forced into the assembled mould using a syringe. The mould is then placed into an oven at 90°C for one hour to cure the PDMS. The cantilever chip is mounted into a recess in one of the acrylic layers using UVcuring adhesive. After this, the oligonucletide receptors are immobilised on the individual cantilevers using traditional micro-spotting or ink-jet printing techniques.

The assembled flow cell is shown in figure 5. This flow cell plugs onto a non-disposable base unit, which consists of three fluid ports and two metal guide pins, as shown in figure y.

One of the fluid ports is the inlet, one is the outlet, and the third acts as a priming port in order to remove any air bubbles from under the cantilever chip. When the flow cell is pressed onto the base unit, the vertical tubes on the base unit that transport the fluids locate into holes in the PDMS gasket and form a seal (Figure 5B). In this way, the flow cell may be removed from the base unit and a new flow cell attached very simply and quickly. The liquid volume in the flow cell is about 25μl.

This flow cell has a single common fluid path that addresses all of the cantilevers in the array. A variant on this flow cell is currently being developed which has a discrete fluid path to each cantilever. This presents a much greater challenge but will allow the immobilisation to be performed on individual cantilevers by flowing solutions through the appropriate channel.

The immobilisation process can also potentially be monitored in real-time if the flow cell is connected to the optical system.

The following challenges for the discrete path flow cell have been overcome: fabrication of discrete fluid channels with 250µm pitch; alignment of the cantilever chip to the fluid channels in an acrylic layer; gluing and sealing of the cantilever chip to the acrylic layer ensuring that there is no leakage path between adjacent cantilevers; making multiple reversible fluid connections to the header in a rapid and reliable way; and finally ensuring a good optical path to and from the cantilever surface.

Optical subsystem for excitation.

The development of an optical detection subsystem must integrate all of the required elements with the necessary alignment accuracy of the VCSEL, cantilever and photodetector arrays. The design and fabrication of the optical subsystem takes into account several requirements: size specifications, expected beam deflection on photodetector, efficient light injection and collection, the requirement for a coupling element, as well as the control of the output beam which tends to diverge along propagation. For that reason, an original optical subsystem has been implemented, which enables to target accurately each cantilever following the parallelism of the arrays and the microfluidics, amplifying the cantilever deflection without any resonant measurement.

The optical subsystem consists of a 20 single-mode oxide-confined amplifying VCSEL (Vertical Cavity Surface Emission Lasers) array emitting at 850 nm through a 5.5μm AlOx diameter. The divergence of these mono-mode sources is around 22º FWHM, which means that their divergence has to be decreased in order to enlarge the cross talk free propagation distance that have been fabricated (see Figure 6A). In addition, a microlens array must be used to collimate the optical beams. The best lens design has been calculated to have, for a PMMA lens, a focal length of 257μm F/#: 1.71. The beam divergence after such a customized lens, located 263μm far from the VCSEL source, has been measured to be about 1.5°.

The DNA hybridisation induces a weak surface stress of the cantilever used as a transducer.

During the system operation, cantilevers are lighted by the VCSEL laser sources described above. The beam deflection originated from this stress is very small and need to be amplified in order to be detected. The deflection angle of the cantilever is a function of its length: the longer the cantilever is, the bigger the reflection angle. For the minimum level to be detected, the cantilever end moves in the nanometer range, which results in a deflection angle around 0.28 mdeg for a 200μm long cantilever. The simplest way to detect the cantilever movement is to place a photodetector far enough so that the spot displacement on the sensitive surface exceeds the position resolution of the detection device, as routinely used in AFM equipments.

However, this solution hinders the miniaturization of the whole biosensor system so in order to ensure the compactness of the whole system, a coupling element has been fabricated, in which the deflected beam propagates through the air, and whose structure is at the same time the mechanical support frame for the global system. Its consists of two parallel one-side metallized and SiO2 protected silicon wafers separated by air, mounted on a high speed machined metallic support. The minimum propagating beam path should be equal to 10cm to get reach to required sensitivity. For an incident angle of 10° and for a mirror separation of 1cm, such a path is obtained for a coupler length of 2cm. The principle of the coupling element is shown in Figure 6B.

Electro-Optical subsystem for detection and data acquisition.

For the detection of the displacement of the laser beam reflected off the cantilever backside, an array of 20 photodetectors is used. The Position Sensing Detector (PSD) is employed to measure changes on the position of the light spot emitted by a VCSEL cell. The process is illustrated in Fig. 7 where a gaussian distribution of the laser power on the photodetector array is assumed with an elliptic spot with dimensions R x = 3σ x and R y = 3σ y (σ x and σ y are the standard deviation of the gaussian). When biomolecular interaction takes place, the cantilever deflection (d) goes from the initial value (d0) to other (d1), provoking a displacement of the Spot Centre (SC) from its initial location (SC0) to the new one (SC1).

In general, an area image sensor consists of an array of pixels, each containing a photodetector that converts incident light into photocurrent and some of the readout circuits needed to convert the photocurrent into electric charge or voltage and to read it off the array.

PSD systems [START_REF] Gonnason | A Low Cost High Resolution Optical Position Sensor[END_REF][16] [START_REF] Massari | A Fast CMOS Optical Position Sensor With High Subpixel Resolution[END_REF] track spot paths along a photodetection surface. This tracking is characteristic of a sensing process, that in our case it must be in the range from submicrometers up to few tens of micrometer. Thus, the photodetectors and the proposed detection algorithm have been designed and optimized to obtain this resolution. Since the form of the spot is assumed invariant the spot centre (SC) is considered as a good candidate to study the spot movement. However, for some standard pixel sizes, the SC could be always inside a pixel and the required resolution in our application is below the pixel size. With this in mind, an algorithm for the PDS is proposed in this paper divided into the two following steps.

• Step1: Final result for step 1 is that a two-dimensional array of voltages (frame) is generated. With this information, the pixel where the spot is will be known, having a resolution of the SC position equal to the pixel size, or equivalently, a pixel resolution system.

• Step2: To increase the accuracy of the measurement, the centroid of the spot is calculated using a subset of the acquired frame according to step 1, which is an approximation of the SC.

A 128 x 16 pixel array system for PSD was designed. It is implemented in a CMOS technology and is composed of a pixel array, control digital circuitry, analog processing and output buffers. In CMOS there are several classical structures for optical power to current conversion, based mainly on diodes and bipolar transistors as photodetection devices. A pdiffnwell photodiode structure is chosen in this case to isolate as much as possible the photocurrent in each pixel. For the signal pixel array, the well known Active Pixel Sensor (APS) [START_REF] Fossum | CMOS Image Sensors: Electronic Camera-On-A-Chip[END_REF] structure, shown in Figure 8A

6.

Final integration and packaging.The final integration of the whole system is being performed by incorporating the microfluidic header into an optical sub-system which includes the optical sources, coupling element, position sensing photodetectors and mechanical alignment stages. This sub-assembly, shown in Figure 11, is mounted in an instrument that incorporates fluid components, power supplies, temperature control, data acquisition and CMOS control electronics.

Optical Waveguide Cantilevers.

In order to achieve further integration an optical cantilever sensor has been implemented.

The detection method is based on monitoring the light exiting the microcantilever. This integrated waveguide cantilever sensor can be fabricated as array of independent waveguide channels designed for monitoring of bio-specific reactions. The sensor can work in static or in dynamic modes, either by monitoring the deflection or by monitoring the changes in the resonance frequency of the cantilever. The advantage of the device is that the transducer is integrated with the receptor on one chip, the complete optical system is in the plane of the cantilever chip, and only a simple external photodetector is required for measuring the optical output power. No preliminary alignment or adjustment is needed, except for light coupling into the chip, which does not affect seriously the performance of the device if the coupler is well designed. The sensitivity of the device is comparable to the one of the AFM principle based instruments.

The principle of operation is based on the sensitivity of energy transfer between two butt coupled waveguides to their misalignment with respect to each other (see Figure 12 for further details). The technique can be considered as an alternative to the known methods used for read-out of nanomechanical response of microcantilevers sensors. The cantilever displacement can be detected with resolution of 18 fm/√Hz. The limit is defined generally by the shot noise of a conventional photodetector used for the readout of the output signal. Realtime parallel monitoring of several channels can be realised. In contrast to other devices based on standard optical readout, no preliminary alignment or adjustment is required except for light coupling, which does not affect seriously to the performance of the device if the coupler is well designed. The sensitivity of the device is comparable to the one based on standard microcantilevers.

The fabricated device, shown in the photograph in Figure 13, contains an array of 20 waveguide channels. Samples with 100 µm and 200 µm long cantilever beams were fabricated. The cantilevers were 500 nm thick and in air supported two guiding modes, according with the simulations. Silicon nitride input and output waveguides were 120 nm thick. The cantilevers on chip were located over a common cavity located in the centre of the chip. The width of the cavity assigned by the cantilever length was correspondingly 100 µm or 200 µm. Both the cantilevers and the waveguides were 40 µm wide. The external facets of the input and output waveguides were made right at the edge of the chip. This avoids the polishing of the facets and provides relatively efficient coupling using direct focusing into the chip and to collect light exiting the receptor using the full numerical aperture.

The theoretical analysis and experiments demonstrated that the device possesses sufficient sensitivity for detection of sub-angstrom scale cantilever deflections. The sensitivity of the technique is comparable to optical read-out standard microcantilevers, but is more attractive for the applications where the parallel real-time monitoring of several specific reactions on one chip is needed. The device does not need preliminary adjustment or alignment procedures except for light coupling, it can work in a static or in a dynamic modes, either by monitoring the deflection or by monitoring changes in the resonance frequency of the cantilever.

CONCLUSIONS

A new biosensor microsystem is being developed based on the integration of microcantilever sensors, optical excitation and detection subsystem and microfluidic headers. The device is intended to be portable to allow decentralised analysis testing. The microsystem comprises an array of 20 micromechanical cantilevers produced in silicon technology, a polymer microfluidic system for delivery of the samples, an array of 20 lasers of vertical emission and chips with the photodetectors and the CMOS circuitry for signal acquisition and conditioning, capable of measuring the cantilever deflection with sub-nanometer resolution. In a further innovation, an optical waveguide cantilever transducer has been also developed in order to improve the final performance and integration of the microsystem.

  photocurrent is integrated in the parasitic capacitor in node Vpixel (see Figure8.(b)), which is approximately the photodiode capacitor. Transistor Mseg works as a voltage follower and Msw is used to connect/disconnect the pixel to the vertical bus (for columns) which connects all pixels in a column.To correct the influence of the main sources of noise in APS based system, the Correlated Double Sampling (CDS)[START_REF] Mendis | CMOS Active Pixel Image Sensors for Highly Integrated Imaging Systems[END_REF][20] technique is used. It is based on sampling two times the APS output: the first, when Mres is switched off, and the second, at the end of the integration interval. The difference will cancel the noise and also will eliminate the non-zero substrate effect of Mres threshold voltage. The schematic in Figure 8.(c) shows the CDS circuits used. The branch controlled by SHR takes samples of the signals immediately after the reset signal is activated, and the branch controlled by SHS does it at the end of the integration interval. The voltage Voutj = VoutSj -VoutRj is obtained at the CDS output. Since there is a CDS circuit for each column, the outputs are multiplexed to obtain the frames. The system has been designed in a 0.35 μm CMOS technology, 2 polys and 4 metals. The layout of the 10 μm x 10 μm pixel is shown in Figure 9.(a), the size of the photodiode is 4 μm x 6 μm, therefore the fill factor is 24 %. A microphotograph of the fabricaded chip is in Figure 9B, where only 5 of the 20 PSDs are shown. Preliminary experimental results are shown in Figure 10 giving resolutions under 2 μm, which is agreement with the precission of the used micrometer.

Figures

  Figures

Fig. 1

 1 Fig.1 Principle of working of a nanomechanical biosensor

Fig. 2 .

 2 Fig.1 Principle of working of a nanomechanical biosensor

Figure 3 .

 3 Figure 3. Real time monitoring of the oligonucleotide receptors immobilisation. A) Deflection signals of the cantilever sensor for the immobilization of a 2 µM ss-DNA (27 mer) 5'thiol modified and the corresponding unmodified ss-DNA. It is shown that no deflection signal is observed for the unmodified ss-DNA case. B) Simultaneously 2 µM thiolated DNA (27 mer)immobilization detection in the cantilever array platform showed in Fig.2. The differences in the cantilever response are due to the dispersion values in the cantilever spring constant and the initial conditions of the gold surface on each cantilever. These differences can be corrected by previous calibration.

Figure 4 .Figure 6 A 1 sFigure 8 Fig. 9 (Fig. 10 (Figure 11 .Figure 12 .Figure 13 A

 4618910111213 Figure 4. Radiolabelling data of the thiolated DNA coverage on gold before and after the MCH treatment

Table 1 .

 1 Comparative values of the mechanical properties for fabricated and commercial cantilevers.

	L = 200 µm	Materials properties E (10	11	N/m)	ν	Thickness x with (µm)	K (N/m)	Z [18]∆σ = 1mN/m	F (kHz) (air)	Q
	Veeco (SiN) (V-shaped)	1.75			0.25	0.6 x 40	0.12	1.4	20.9	30
	Olympus (SiN)	1.75			0.25	0.8 x 40	0.10	0.8	23.9	35
	Micromasch (silicon)	1.69			0.20	1 x 50	0.26	0.6	38.4	45
	CNM (silicon)	1.69			0.20	0.33 x 40 0.009	5.1	7.8	6
	For the silicon nitride cantilever a young's modulus of E =1.75 N/m2 has been used,
	and for the silicon, E =1.49 N/m2

ACKNOWLEDGEMENTS

This work is being supported by the EU Contract IST-2001-37239. Spanish authors are also supported by national project TIC2002-10473-E.


V Bardinal

R Legros

C Fontaine

In situ measurement of AlAs and GaAs refractive index dispersion at epitaxial growth temperature

In situ molecular beam epitaxy control of III-V optoelectronic device growth has been achieved by dynamic optical reflectometry with tunable excitation wavelength, through the use of a titanium:sapphire laser light source. This new multi-wavelength reflectometry method was used to determine the values of the AlAs and GaAs refractive indices at growth temperature (600 °C). Index dispersion between 760 and 960 nm is presented and found to be in good agreement with the existing models.

The increasingly common use of epitaxially grown Bragg reflectors or other multilayer structures in optoelec-tronic devices such as vertical cavity surface emitting lasers 1 or reflection modulators 2 has significantly increased the need for high precision layer thicknesses. The epitaxial techniques used for the fabrication of these devices, chemical vapor deposition (CVD), and molecular beam epitaxy (MBE), are more likely to meet this stringent requirement if real-time thickness control can be performed. Over the last years, certain in situ techniques have been developed for that purpose, based on reflection high energy electron diffraction (RHEED), 3 optical based flux monitoring, 4 spectroscopic ellipsometry, 5 interferometric pyrometry, 6 and dynamic optical reflectometry (DOR). 7 While RHEED -which can only be used for MBE-leads to an accurate control of tiny thicknesses, i.e., for quantum wells, the others are better suited for the measurement of higher range thicknesses,

as those used in reflectors. In the latter case, in situ DOR has commonly been used with single-wavelength laser light sources. 7 However, a predetermined probe wavelength makes it difficult to control multilayer thicknesses. This limitation has recently been overcome by the use of broadband sources, 8 which allow those wavelengths that are most sensitive to a particular process to be selected. In this letter, an alternative multi-wavelength technique is reported. This new method takes advantage of a titanium:sapphire cw tunable laser source which combines wavelength tuning capability in a wide spectral range with laser source characteristics: stability of the intensity and precision of the probe wavelength.

In this letter, GaAs and AlAs refractive indices at growth temperature are measured using this technique between 1.29 and 1.63 eV (760-960 nm), a range of great interest for III-V optoelectronic devices. So far, this wavelength range has received little attention. Indeed, only optical index dispersion for both binaries between 1.6 and 3 eV obtained by means of spectral reflectance have been recently published. 9 In the case of GaAs, the 1.5-5 eV range has also been investigated by spectroscopic ellipsometry, leading for this material to refractive index values which were then modelized 10 or fitted with a harmonic oscillator approximation (HOA) model. 11,12 In this work, index measurement was also performed at 633 nm in addition to the range investigated, in order to allow comparison with the published values, generally obtained at this wavelength. Finally, the index values measured and their dispersion are shown to be in good agreement with the existing models. (100) GaAs substrates were prepared by a UV-ozone oxidation 13 prior to being loaded into a 2300 Riber system. Substrate temperature was measured with a pyrometer. To ensure temperature reproducibility, a systematic procedure was used, leading to oxide desorption at 610 °C. Moreover, the temperature range involved was observed to be very sensitive to the As contamination of the pyrometer viewport and care was taken to keep it clean throughout the study. Growth was performed at 600 °C for both III-V binaries. Growth rates, in the range of 1 µm/h, were adjusted by RHEED oscillations so as to optimize the growth of both binaries for the same As pressure. Temperature ramps were applied to III element cells to avoid flux transients at shutter openings. The experimental apparatus for in situ selective DOR measurements is shown in Fig. 1. External light source is provided by a titanium:sapphire cw tunable laser via an optical fiber. The laser beam illuminates the growing layer surface at an angle of 2° to the normal and is then reflected and detected by a Si photodiode. The resulting signal is demodulated by a lock-in amplifier and acquired as a function of time by a computer. Tuning of the wavelength within the range investigated from 760 to 960 nm is achieved by means of a computer-controlled step-by-step motorization of the Lyot filter. This spectral range can be analyzed as a result of the use of two sets of laser mirrors.

Dynamic optical reflectometry consists of subjecting at nearly normal incidence a light beam on the surface of a growing sample. 7 The reflected beam exhibits oscillations due to multiple reflection interferences as the layer thickness increases. The oscillation period is equal to /2n, where n and  are the real part of the layer refractive index and the incident beam wavelength, respectively. Assuming a constant growth rate G, the refractive index n of the growing material can easily be inferred from these reflectivity oscillations by using the following relation: n=/(2G), where  is the oscillation period. Whatever the underlying multilayer structure this procedure can nonetheless be applied by taking care not to use the initial oscillation for this calculation, which is systematically shorter than the other oscillations due to phase shift. When the controlled layer is of the same type as the substrate, GaAs in our case, a decrease in contrast, or even cancellation of the oscillations can sometimes be observed. This phenomenon has already been described 8 elsewhere and can be ac counted for by a phase effect. This arises when the optical thickness of the already grown multilayer is about a multiple of /2, which makes the uppermost GaAs layer behave as if it were directly grown onto the GaAs substrate. In order to avoid this detrimental effect, a reflective multilayer, com-posed of some ~/4n-thick AlAs-GaAs alternating layers, was systematically grown prior to the GaAs period measurement. This procedure was found to greatly improve the oscillation contrast. However, GaAs measurement accuracy was observed not to be as good as for AlAs where absorption is not significant (see Fig. 3). First, index calibration was carried out. To do this, two samples were grown in the same run, under the same conditions of growth rate and temperature (600 °C). DOR was achieved at 0=920 nm on sample (a) using the aforementioned procedure. For ex situ calibration of the growth rate by double crystal x-ray diffraction, sample (b) was used. It consisted of 16 periods of GaAs and AlAs alternating layers, with selected thicknesses of about a quarter-wave for 0. These GaAs and AlAs thicknesses were measured to be 61.45 (±0.5) and 73.85 (±0.5) nm, respectively. These data were later confirmed by spectroscopic ellipsometry. GaAs and AlAs growth rates at 600 °C were then derived from these values measured at room temperature and from the thermal expansion coefficients of both semiconductors, 14 allowing the values of AlAs and GaAs refractive indices to be thoroughly obtained for sample (a): n(AlAs)=3.11

and n(GaAs)=3.72. Given our experimental conditions (=±0.5 s, =±0.05 nm, thickness=±0.5 nm), the preceding values were determined within 0.8%. These index values, which correspond to 600 °C and 920 nm, were then used as a reference in the remaining part of the study.

In a second step, index dispersion was determined. During the spectral study, attention was paid to maintaining the substrate temperature constant and equal to 600 °C. Reflectivity oscillation periods were measured at 920 nm, for reference, and at 760, 780, 810, 835, 860, 900, 940, and 960 nm (see an example given in Fig. 2). GaAs and AlAs refractive index variations were then inferred from their 920 nm reference value and from the experimental dispersion of the oscillation period. The results are shown in Fig. 3 for both materials. In addition, DOR measurements were performed by replacing the tunable laser with a He-Ne laser at =632.8 nm to provide comparison with the index values published and mainly measured at this wavelength. Our results were first compared in the case of GaAs with the HOA model dispersion at 611 °C. 10 A good agreement was obtained, as shown in Fig. 3, although the model yields slightly higher values than the experiment which can be accounted for by the temperature difference. The high temperature index dispersion was also calculated for both binaries from the Moss semi-empirical model, [START_REF] Moss | Optical Properties of Semiconductors[END_REF] which relates index and band gap dependencies on temperature according to the following relation To do this, the index dispersion at room temperature n() was taken from Afromovitz's model, [START_REF] Afromovitz | [END_REF] while the temperature variation of the band gap was derived from Varshni's empirical relation 17,18 and experimental values 18 for GaAs and AlAs, respectively. This leads to a dispersion law of the refractive index at 600 °C plotted in Fig. 3 for both materials. Our experimental values turn out to be in good agreement with this law, which does not include any adjustable parameters. We now use this model for first-order modeling of multilayer optical behavior at high temperature.

In order to compare our results with those published in the literature, mainly at 633 nm, the Moss model was also used to bring these values to 600 °C (see Table I). They were found to differ greatly. These discrepancies could be accounted for by the difference in the models involved in the experimental techniques of characterization and by the lack of precision of the indicated temperatures. Indeed, achieving absolute calibration of surface temperatures in a reduced pressure (MOCVD) or high vacuum (MBE) environment is highly difficult.

In our case, the values obtained for the GaAs and AlAs refractive indices within the wavelength range investigated will be used to control the growth of these semi-conductors at the temperature we consider to be 600 °C and calibrated by a systematic procedure.

To conclude, a new technique of in situ dynamic optical reflectometry taking advantage of a tunable laser light source has been presented. In a first step, this technique was successfully used to measure AlAs and GaAs refractive in-dices at a growth temperature between 760 and 960 nm, i.e., in a range in which they are seldom reported. These measured values were shown to be in good agreement with a model based on the Moss semiempirical expression, which can then be used for the design of the optical behavior of III-V materials at growth temperature. This tunable DOR technique turns out to be highly flexible and particularly precise throughout this study. These advantages are now being used for precise real-time monitoring of the MBE growth of GaAs/AlAs Bragg reflectors and light modulators, as will be reported shortly.

.

FIG. 1 . Experimental setup 8 FIG. 2 . 9 FIG. 3 .

 18293 FIG. 1. Experimental setup

Acknowledgements

The authors would like to thank Dr. A. Mun ˜oz-Yagu ¨e for fruitful discussions and Dr. G. Rolland and Dr. F.

Bertin (LETI-CEA) for double-crystal x-ray diffraction and spectroscopic ellipsometry measurements.