Matthieu Petit
email: petit@lasmea.univ-bpclermont.fr

Christine Robert-Goumet

Luc Bideux

Bernard Gruzza

V Matolin

S Arabasz

B Adamowicz

D Wawer

M Bugajski

Passivation of InP(100) substrates: first stages of nitridation by thin InN surface overlayers studied by electron spectroscopies

Keywords: 68.49Uv, 68.55Jk, 79.60Dp, 61.46.+w, 78.55.-m indium nitride, indium phosphide, nitridation, electron spectroscopies, AES, ESCA, photoluminescence, passivation

INTRODUCTION

III-V compound semiconductor surfaces and interfaces have been intensively investigated in the past decades because of both fundamental and technological interest. Among them indium phosphide is known to have a high carrier mobility : this material have a strong potential use in integrated optoelectronic devices [START_REF] Sze | Physics of Semiconductor Devices[END_REF].

Treatments of indium phosphide surfaces will improve temporal and thermal stability of devices developed on the base of this semiconductor. Designed technology of III-V nitridation can have considerable importance for modern electronics, optoelectronics, quantum devices and other devices with nanometer sizes.

However the nitridation of III-V semiconductors is quite difficult because of the small sticking probability of nitrogen on the surface [START_REF] Suzuki | [END_REF]. For InP, several kinds of nitridation process have been used to produce nitridated films: among them, alkali metals promoted nitridation which lead to the formation of mainly InPNx layers [3,4], direct nitrogen ion beam nitridation resulting in nitridated layers showing In-N, In-N-P and P-N bondings [START_REF] Suzuki | [END_REF] and plasma nitridation which entails the creation of In-N and P-N bonds [5].

In this article, we are focused on the first steps of InP (100) nitridation following at a low scale by electron spectroscopies (AES, ESCA). The reactive nitrogen beam was produced by a high voltage glow discharge source (GDS) with two incident angles : normal and grazing to the normal of the surface. Results by photoluminescence ex-situ method obtained on the InN/InP ended structures are also reported. The InP(100) surfaces have been firstly bombarded by argon ions to remove contamination layers but this cleaning produce a preferential phosphorus desorption which entails the creation of metallic indium droplets (coverage rate 25%, average height : 4 monolayers) [6]. This method based on an indium precursor state has been fruitfully used to create InN atomic layers. Clearly passivation of the InP substrates was observed underlining the efficiency of the process.

EXPERIMENTAL

Two ultra vacuum chambers (10 -7 Pa) have been used for this study. One was equipped with ESCA spectroscopy using the MgK  radiation (1253,6 eV) and the analysis were performed on ended InN/InP(100) structures. The other UHV chamber, having Auger electron spectroscopy (AES) and nitrogen source, was devoted to samples elaboration. During the nitridation process, Auger peaks can be recorded in real time.

Loading of the InP(100) substrates (S doped, n = 4,7 10 16 cm -3 ) were done after ex-situ chemically cleaning [7,8]. This cleaning is based on successive baths of H2SO4, a solution of 3% bromine in methanol and deionised water. The sample surface is controlled by Auger electron spectroscopy with a retarding field analyser.

a) Indium based precursor state.

After introduction of the samples in UHV chamber, a low level contamination by carbon and oxygen was detected. These impurities were removed in situ by Ar + ions bombardment using a low energy (E0 = 300 eV, ion current = 2 A cm -2 , t = 15 min). This first treatment induces P depletion and well controlled metallic indium islands formation [9,10]. These nano-droplets on the surface have been extensively studied in the past by authors using mainly electron spectroscopies and simulation for a good understanding of the phenomenon [6,10,11]. It has been shown that roughly each crystallite contains about ten In atoms and the whole quantity of metallic indium is equivalent in coverage to one monolayer [6]. Atoms of V element as antimony deposited on this In rich surface induce 3D 2D transformation and the formation of thin InSb overlayers [12,13].

Under heating (T = 573 K) one also clearly observes the vanishing of In droplets [10,[START_REF] Gruzza | Thesis of Blaise Pascal University[END_REF] which would appear again if the temperature decreases up to 300 K.

In this study we have advisedly used the indium rich InP(100) surfaces and the heating effect to include nitrogen atoms in the overlayer region.

b) Nitridation process.

The reactive nitrogen beam was produced by a high voltage glow discharge source (GDS). This kind of source mainly creates nitrogen atoms, with some N + , N 2+ and N2 species in very low amount. Two directions of the nitrogen flow were possible: a grazing incidence or a normal one to the surface sample. Time was a fundamental parameter of the nitridation.

Another one was the temperature fixed equal to T = 523 K which was optimal for the process because the decomposition temperature of InP is around T = 573 K.

RESULTS AND DISCUSSIONS. a) A.E.S. analysis :

The nitridation of the In rich InP surface was followed by recording the Auger peaks of the three atomic elements: The RFA apparatus gives best results for low electron kinetic energies. For this reason phosphorus Auger peak has been well observed. On the other hand, the evolution of intensities gives important information along the duration of the process. Concerning the knowledge on the environment of indium atoms, better results were obtained with XPS spectroscopy and are presented in the following part of this article.

In the figure 1 one has reported the ratio

N P I I
of the Auger signal intensities of nitrogen and phosphorus atoms respectively. Clearly for a grazing incidence flow the nitridation seems better because of the increase of the N Auger peak and the decrease of the phosphorus signal.

Really the mode of preparation with a liquid indium layer on the surface acting as a precursor plays an important role. In the literature [15,16] we can find opposite results concerning influence of the beam direction on other initial kind of InP surface. In their case, the nitridation process is performed by ion gun method. As regard us, the explanation is simply due to the longer path trajectory of incident N atoms in the indium overlayer. -the region 1: the beginning of the nitridation -the region 2: the optimisation of the nitridation (the maximum for R) -the region 3: a decrease of the factor R From this graph one can deduce the time to create a complete InN overlayer, which must be equal to 40 minutes for both the two possible directions of the N beam. Considering the evolution of these two lines it is more continuous in the case of grazing incidence, one must also have in mind the figure 1 indicating more nitrogen adsorption in this case. Moreover in spite of a smaller maximum the curve is above the other one. For these reasons we can consider a better efficiency for grazing incidence N flux.

Concerning the indium M4N45N45 one remarks an energy shift of the peak which is maximum (E = 2 eV) for t = 40 min and also with the use of a grazing nitrogen beam. This result must be correlated to the optimisation of the nitridation of the sample.

b) ESCA results :

Looking at the photoelectron spectra of In4d level presented in the figure 4, a shift of this peak is observed versus the nitridation time. This shift is equal to 0.8 eV for t = 20 min and maximum (2 eV) for t= 40 min. Prolonged time of nitridation (t = 60 min) implies a diminution of this shift (1 eV). In fact these results must be compared to those from the figure 3 where three steps for the nitridation process have been put in evidence.

More precisely, the XPS analyses are reported in the spectra series of the figures 5.

Remembering that indium 4d peak is due to a double contribution (4d3/2, 4d5/2) it is possible to decompose each experimental peak recorded at different times of nitridation. The contributions of oxides -present on the surface since the samples have been exposed to the atmosphere during the transfer from one UHV chamber to the other to perform XPS analysiswere not included in the figure for a readable question.

At the beginning (t = 0 min) only In-P contribution (18.5 eV and 19.5 eV) and In-In contribution due to the initial droplets (17.5 eV and 18.5 eV) are present (figure 5a). After 20 minutes of nitridation, we observe a decrease of the In-In peaks and the apparition of a new contribution at E = 19.5 eV and 20.5 eV characteristic of the In-N bonds [17] (figure 5b).

The time t = 40 minutes corresponds to the time for a complete nitridation of the substrate, no more metallic indium is present on the surface (figure 5c). When the experiment is prolonged (t = 60 min) one observes again the presence of In-In peaks (figure 5d).

c) Interpretation attempt

From the previous results one can deduce that nitridation experiment follows two steps. In the first times, we observe the nitridation of the surface up to a complete consummation of initial quantity of free indium which is liquid on the surface at T = 523 K. Then the correct time of the process is close to 40 minutes for which a maximum of two InN monolayers have been condensed on the surface. The initial state of In crystallites on InP is well known: mean sizes 4 monolayers and coverage 25%. The two nitrided monolayers are corresponding to the maximum of possible stacking layers with an initial quantity of metallic indium equal to one monolayer.

At this time the nitridation is stopped due to the lack of free indium atoms. An irreversible process is observed coming with the apparition of In-In bonds. A possible satisfying explanation can be a low desorption of InN at this temperature and a damage of the surface. This phenomenon is comparable to the evaporation of InP substrate (congruence temperature 573 K) [18]. The schematic figure 6 summarises the nitridation process in time.

EX-SITU PHOTOLUMINESCENCE EXPERIMENTS :

Room temperature ex-situ photoluminescence measurements upon different excitation light intensities from InP samples after nitridation were carried out at the Institute of Electron Technology (Warsaw, Poland). The green line (514.5 nm) of Ar + laser was used as an excitation source. The photoluminescence signal was collected by an optical fibre and analysed by a spectrometer equipped with a CCD camera.

A set of near-band edge PL spectra for 30, 40 and 50 minute nitrided InP(100) surfaces is shown in Fig. 7a. It is clear that PL intensity (IPL) varies markedly after different nitridation times although no change in spectrum shape was observed. The strongest photoluminescence was found for the sample after 40 min nitridation. Recently, we proved that PL is affected by highly complicated surface recombination processes [19] and is a strongly nonlinear function of both excitation light intensity, and surface state distribution [20]. Thus, in order to assess the recombination properties of InN-InP interface, PL measurements at different excitation light intensities () were performed. The results, in terms of a PL quantum efficiency, i.e. ratio YPL=IPL/ for various nitridation time are summarised in Fig. 7b. The significant increase of YPL (by about half an order of magnitude upon high ) is observed for 40 minutes nitridation in comparison to 30 and 50 minutes processing. We attribute this increase to the decrease of interface recombination rate, probably due to the reduction of interface state density. This hypothesis is additionally supported by the fact that the largest dynamics of YPL changes upon increasing  is also noticed for 40 minutes nitridation. It is well explained in terms of the gradual saturation of interface states as non radiative interface recombination centers [20]. Apparently, the smaller interface state density coincides with the faster saturation, and thus with better-marked YPL changes. So that, nitridation for 40 minutes yields the best passivation effect of InP substrate by InN overlayer.

It should be emphasised that the PL-based characterisation of InN-InP interface is well consistent with the analysis of electron spectroscopy results. From comparison of Fig. 7b and Fig. 3 it is clear that the YPL behaviour follows the amount of nitrogen content at the interface, indicating different mechanism of nitridation. Therefore the correlation between the electronic properties and the chemical composition of the interface is evident.

CONCLUSION

The formation of metallic indium on the InP(100) by Ar + bombardment is a precursor for formal elaboration of InN monolayers on the substrate. Experimental choices of parameters are crucial. The temperature T = 523 K seems the best one and well allows the inclusion of nitrogen atoms. Time around 40 minutes is really optimal, all metallic indium is consumed and the process must be stopped. Better efficiency is obtained when the nitrogen beam has grazing incidence. Electron spectroscopies (AES, ESCA) give quantitative results at the nanoscale, the knowledge of the structure InN/InP(100) is absolutely necessary to well understand ex-situ analysis. Optical measurements are well correlated to the quality of sample nitrided overlayer. Passivation of the semiconductor surface has been really put in evidence by such kind of experimental results. Some complementary experiments (LEED) must be performed to conclude on the crystallinity of the InN overlayers. Such sort of heterostructures is promising substrates for optoelectronics and nanoelectronics. One of the future objectives is the use of an indium evaporation source in order to have also the possibility to increase the thickness of InN overlayer.

-

  P  LMM (E = 117 eV) -N  KLL (E = 383 eV) -In  MNN (E = 403 eV and 411 eV)

  figure 2c (t = 40 min) indicates a more important contribution of P-N in comparison to that of P-In. Following this idea we have reported in the figure3the ratio:

Figure 1 :

 1 Figure 1: Ratios of Auger signals of nitrogen and phosphorus atoms versus nitridation time and for two positions of the N beam.

Figure 2a :

 2a Figure 2a: Phosphorus Auger transition after in-situ cleaning of the InP substrate.

Figure 2b :

 2b Figure 2b: Phosphorus Auger transition after 20 minutes of nitridation. The decomposition in series of triplets indicates the presence of P-N bonds.

Figure 2c :

 2c Figure 2c: Phosphorus Auger transition after 40 minutes of nitridation. The decomposition in series of triplets indicates the presence of P-N bonds.

Figure 3 :Figure 4 :

 34 Figure 3: Ratio of the intensities of P-N and P-In Auger signals, showing the influence of the nitrogen beam direction. Versus time of nitridation, three regions are put in evidence Figure 4: The indium 4d peak for different times of nitridation and the associated energy shift.

Figure 5a :

 5a Figure 5a: The decomposition of the In4d ESCA peak after the Ar + cleaning showing the metallic indium on the surface.

Figure 5b :

 5b Figure 5b: The decomposition of the In4d ESCA peak after 20 minutes of nitridation. In-N bonds are apparent.

Figure 5c :

 5c Figure 5c: The decomposition of the In4d ESCA peak after 40 minutes of nitridation. No more metallic indium detected.

Figure 5d :

 5d Figure 5d: The decomposition of the In4d ESCA peak after 60 minutes of nitridation. A decrease of the In-N signal, a detection of metallic indium.

Figure 6 :

 6 Figure 6:A schematic representation of the surface nitridation along duration of the process.

Figure 7a :

 7a Figure 7a: Room temperature photoluminescence spectra obtained on different nitrided InP

  FIGURE 1

  FIGURE 3

  FIGURE 6

P-InKinetic energy (eV)

ACKNOWLEDGEMENTS

This work was partially performed within the V FPEC Project of Centre of Excellence in Physics and Technology of Semiconductors Interfaces and Sensors (CESIS) under the contract G6MA-CT-2002-04042. One of the author (S.A.) acknowledges the Marie Curie Host Fellowship funded by the European Commission.


Romain Jacquet

Cécile Miège
email: cecile.miege@irstea.fr

Foppe Smedes

Céline Tixier

Jacek Tronczynski

Anne Togola

Catherine Berho

Ignacio Valor

Julio Llorca

Bruno Barillon

P Marchand

Marina Coquery

Comparison of five integrative samplers in laboratory for the monitoring of indicator and dioxin-like polychlorinated biphenyls in water Authors

Keywords: Integrative samplers, polychlorinated biphenyls, water monitoring, time weighted average concentration, modeling

published or not. The documents may come

Comparison of five integrative samplers in laboratory for the monitoring of indicator and dioxin-like polychlorinated biphenyls in water

Introduction

Like many other hydrophobic organic contaminants, polychlorinated biphenyls (PCBs) have toxic effects on living organisms, including human beings [START_REF] Carpenter | Polychlorinated biphenyls (PCBs): Routes of exposure and effects on human health[END_REF].

In aquatic environments, PCBs are principally adsorbed on particulate matter due to their hydrophobicity (logK OW > 4.5); hence, their concentration in the dissolved phase is therefore very low, typically in the ng/L to pg/L range. Monitoring such low concentrations with traditional bottle (or grab) sampling is challenging and requires sophisticated analytical methods such as isotopic dilution mass spectrometry. Furthermore, grab sampling only provides a snapshot of the contaminant concentration at a particular time without taking temporal variations into account.

Since two decades, several integrative sampling devices have been developed for the monitoring of organic contaminants in aquatic environments [START_REF] Greenwood | Potential applications of passive sampling for monitoring non-polar industrial pollutants in the aqueous environment in support of REACH[END_REF][START_REF] Söderström | Strategies for monitoring the emerging polar organic contaminants in water with emphasis on integrative passive sampling[END_REF][START_REF] Lohmann | Use of passive sampling devices for monitoring and compliance checking of POP concentrations in water[END_REF]. These samplers enable the improvement of limits of quantification (LOQ) by accumulation and concentration of contaminants over long-term exposure. Moreover, when they are used in the integrative phase of uptake (i.e. integrative samplers), time-weighted average (TWA) concentrations over the exposure period can be calculated, leading to a better representativeness of measurements.

Several integrative samplers, at different stages of development, are now available for monitoring non-polar organic contaminants. The semi-permeable membrane device (SPMD) is one of the most comprehensively studied integrative sampler; it consists of a low-density polyethylene (LDPE) lay-flat tubing filled with a small quantity of triolein. It was designed to sequester and concentrate freely dissolved organic contaminants with logK OW ranging from three to eight and has already been extensively used for the monitoring of PCBs in water (Huckins et al. 2006). Next to biphasic sampling devices like SPMD, singlephase integrative samplers, such as LDPE strip and silicone rubber (SR), are gaining interest due to simpler modelling of contaminant transport processes and easier sample

Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011

processing. Numerous studies have shown the suitability of LDPE strips for the monitoring of hydrophobic organic contaminants, such as polyaromatic hydrocarbons (PAHs) or PCBs in various water bodies [START_REF] Booij | Temperature dependent uptake rates of non-polar organic compounds by semi-permeable membrane devices and low-density polyethylene membranes[END_REF][START_REF] Carls | Monitoring polynuclear aromatic hydrocarbons in aqueous environments with passive low-density polyethylene membrane devices[END_REF][START_REF] Adams | Polyethylene devices: passive samplers for measuring dissolved hydrophobic organic compounds in aquatic environments[END_REF][START_REF] Anderson | Field trial and modeling of uptake rates of in situ lipid-free polyethylene membrane passive sampler[END_REF]. Silicone rubber was also found to be a suitable alternative to SPMD for the monitoring of hydrophobic contaminants [START_REF] Rusina | Polymer selection for passive sampling: a comparison of critical properties[END_REF]. Indeed, SR sheets have been successfully used for the monitoring of PAHs and PCBs from 2002 in The Netherlands [START_REF] Smedes | Monitoring of chlorinated biphenyls and polycyclic aromatic hydrocarbons by passive sampling in concert with deployed mussels[END_REF]. Chemcatcher can house different combinations of receiving phases and membranes as appropriate for polar or non-polar contaminants monitoring (Greenwood et al. 2007). The first non-polar version of Chemcatcher, made of a C18 Empore disk and a LDPE membrane, aimed at sampling contaminants with logK OW greater than three [START_REF] Kingston | Development of a novel passive sampling system for the time-averaged measurement of a range of organic pollutants in aquatic environments[END_REF]. A recent optimization of the sampler, by adding a small volume of octanol between the receiving phase and the membrane, was proposed to decrease the internal sampler resistance to mass transfer of hydrophobic compounds with logK OW above five [START_REF] Vrana | Performance optimisation of a passive sampler for monitoring hydrophobic organic pollutants in water[END_REF]. Chemcatcher has already been used during field campaigns for the monitoring of PAH and organochlorine pesticides [START_REF] Vrana | Field performance of the Chemcatcher passive sampler for monitoring hydrophobic organic pollutants in surface water[END_REF].

Finally, developed since 2008, CFIS (Continuous Integrative Flow Sampler) is a new active (i.e. using a pump) sampler designed for the determination of TWA concentrations of organic compounds in water [START_REF] Llorca | Constantly stirred sorbent and continuous flow integrative sampler. New integrative samplers for the time weighted average water monitoring[END_REF]. Briefly, CFIS is a fully immersible device consisting of a small peristaltic pump powered by batteries and producing a constant water flow through the glass cell containing a PDMS sorbent. The main advantage of CFIS is that sampling rates are unaffected by water turbulence or velocity and thus, the use of performance reference compounds (PRCs) is not required. It has already been used for the monitoring of PAH and organochlorine pesticides in wastewater treatment plant effluent [START_REF] Llorca | Constantly stirred sorbent and continuous flow integrative sampler. New integrative samplers for the time weighted average water monitoring[END_REF].

Over the past 20 years, a variety of models has been developed to better describe the transfer kinetics of hydrophobic contaminants into integrative samplers (Booij et al. Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011

2007). Whatever the integrative sampler and model considered, the calculation of TWA concentrations of contaminants in water from amounts accumulated in the sampler requires the knowledge of sampling rate (R S ) and sampler-water partition coefficient (K SW )

for each compound. Sampling rates are determined by laboratory calibration under controlled exposure conditions. In situ R s calibration is needed to take into account differences between laboratory vs. in situ exposure conditions (i.e. flow velocity, biofouling or temperature); it is achieved by the use of internal surrogates (performance reference compounds, PRCs), spiked in samplers prior to exposure [START_REF] Huckins | Development of the permeability/performance reference compound approach for in situ calibration of semi-permeable membrane devices[END_REF]. K SW can be determined experimentally (Smedes et al. 2009 for LDPE and SR) or estimated via empirical relationships as a function of logK OW (Huckins et al. 2006for SPMD, Vrana et al. 2006for Chemcatcher, Booij et al. 2003[START_REF] Adams | Polyethylene devices: passive samplers for measuring dissolved hydrophobic organic compounds in aquatic environments[END_REF][START_REF] Lohmann | Global Aquatic Passive Sampling (AQUAGAPS): using passive samplers to monitor POPs in the waters of the world[END_REF] for LDPE). Concerning CFIS, that is an "active" sampler, the use of PRC and K SW is not necessary. Indeed, a pump enables to control the water flow during exposure and the temperature effect is known by previous calibration in laboratory (from 5°C to 35°C). By this way, R s estimated in laboratory for each PCB is corrected according to the average temperature encountered during in situ exposure, to be directly used for the determination of TWA concentrations [START_REF] Llorca | Constantly stirred sorbent and continuous flow integrative sampler. New integrative samplers for the time weighted average water monitoring[END_REF].

Very few intercomparison exercises on integrative samplers have been performed until now. [START_REF] Allan | Field performance of seven passive sampling devices for monitoring of hydrophobic substances[END_REF] or [START_REF] Miège | An in situ intercomparison exercise on passive samplers for monitoring metals, polycyclic aromatic hydrocarbons and pesticides in surface waters[END_REF] [START_REF] Rusina | Calibration of silicone rubber passive samplers: experimental and modeled relations between sampling rate and compound properties[END_REF]. Using these five samplers allows comparing different strategies for integrative sampling: passive vs. active (with pump) sampling, use of PRC or not, use of different models and equations to assess TWA concentration. By exposing these five integrative samplers into the same experimental calibration system, a first objective was to compare their performances in accumulating PCBs (uptake, repeatability and linearity). Moreover, since there is no detailed guideline on integrative sampling, a second objective was to compare different methods of calculation of TWA concentrations (models, partition coefficients values and selected PRCs).

Materials and methods

Integrative samplers

The main characteristics of the studied integrative samplers as well as the PRCs tested and main steps of their processing are summarized in Table 1. Further details on their characteristics, pretreatment and analysis are given in Supplementary data (S1).

Target molecules

The exposure of samplers was performed with 19 PCBs: PCB 18, indicator PCBs (PCB 28,52,101,118,138,153 and 180) and dioxin-like PCBs (PCB 77,81,105,114,118,123,126,156,157,167,169 and 189). All these PCBs were purchased from Cil Cluzeau (Courbevoie, France) and delivered in a custom-made solution used to spike the silicone rubber sheets (referred as "dosing sheets" hereafter).

Exposure device and strategy

The exposure device was custom-made (PIC, Olivet, France) and consisted of a tank (height = 120 cm, diameter = 47 cm), a stirrer and six holders; all these pieces being made of stainless steel to minimize adsorption. A scheme of the exposure device is presented in Supplementary data (S2). The tank was filled with 200 L of tap water agitated with a stirrer (height = 100 cm, width = 18 cm) set in motion by an electronic engine (Heidolph RZR 2102 control Z; VWR, Fontenay-sous-Bois, France). Rotation speed was set at 33 rpm to obtain a water velocity of about 5 cm/sec near the samplers exposed in the tank. To regulate water temperature, the exposure tank was placed in a 300 L polyethylene tank (CVC series, Manutan, Gonesse, France) and water was cooled by an aquarium chiller (Teco TR20, Europrix, Lens, France). The exposure tank contained six holders (height = 100 cm, width = 20 cm) set along the wall. Two holders were used to support the dosing sheets and the four other holders were used to fix four types of integrative samplers: SPMD, SR, LDPE strip and Chemcatcher. Each holder had four

Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011

positions enabling the simultaneous exposure of four samplers at different depths. CFIS were installed outside the tank but were exposed to the tank water by use of glass tubes.

PCBs were dosed to the tank water by immersing a large amount of dosing sheets [START_REF] Rusina | Calibration of silicone rubber passive samplers: experimental and modeled relations between sampling rate and compound properties[END_REF]. This allowed for maintaining a constant concentration of about 1 ng/L of each studied PCB throughout the experiment. Dosing sheets were first mounted in the exposure device and the tank was filled with tap water. Then, the exposure system was allowed to equilibrate under agitation and temperature regulation during two days, after which water was renewed. This step allowed for cleaning the system and eliminating traces of methanol that might have remained in the dosing sheets from spiking. After another two days of equilibration, samplers were deployed in the tank.

Exposures in the water tank lasted up to three months. SPMD, SR and LDPE strips were exposed during 1, 3, 7, 14, 21, 28 (in triplicate), 56 and 91 days. Chemcatcher and CFIS were exposed during 3, 7, 14 (in triplicate), 21, 28 and 56 days. Before and after exposure, samplers were stored at -20°C.

Temperature of the tank water was recorded every 6 h over the whole exposure duration. During the first month, water (1 L) was sampled weekly for determination of pH, conductivity and dissolved organic carbon (DOC) concentration. During the last two months, these measurements were performed every two weeks. The concentrations of PCBs in water were calculated from their concentration in dosing sheets [START_REF] Rusina | Calibration of silicone rubber passive samplers: experimental and modeled relations between sampling rate and compound properties[END_REF] and using their PDMS -water partition coefficients [START_REF] Smedes | Polymer-water partition coefficients of hydrophobic compounds for passive sampling: application of cosolvent models for validation[END_REF]). Dosing sheets were sampled every two weeks by cutting six small pieces at different depths in the tank obtaining a total amount of about 1 g of material. Further details about the preparation and analysis of dosing sheets are given in Supplementary data (S1).

Quality controls

Interlaboratory assay

Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011

In order to assess the interlaboratory variability in PCB analysis, a standard solution was prepared and sent for analysis by each of the five laboratories involved in this study with its own analytical method. This solution contained the 19 studied PCBs at concentrations ranging from 50 to 130 µg/L and was conditioned in amber glass vials prior to shipment. A good agreement was found between laboratories since the relative standard deviations (RSD) on measured concentrations ranged from 3 to 13 % depending on the congeners.

Blank samplers

After their preparation, several samplers were kept as procedural blanks in order to evaluate any possible contamination during fabrication, spiking, storage, processing and analysis. These procedural blanks were stored at -20°C until processing. Other samplers were used as "field" blanks and exposed to the ambient air during the handling of deployed samplers to take account for any possible contamination during deployment and retrieval.

These "field" blanks were stored at -20°C until pro cessing. No contamination by PCB was measured in procedural and field blanks.

At last, for each type of sampler (except CFIS), a blank sampler (not spiked with PRC) was exposed in the tank during the whole exposure duration in order to assess any possible contamination with PRC between samplers. Only low amounts of rapidly releasing PRC were observed representing less than 4 % of the concentration in these exposed spiked samplers.

Calculations

Several models have been developed to describe the transfer of hydrophobic contaminants into the various available integrative samplers and to calculate the TWA concentrations in water from the accumulated amounts in the samplers [START_REF] Booij | Theory, modeling and calibration of passive samplers used in water monitoring[END_REF].

Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011

In this work, for SPMD, SR, LDPE strip and Chemcatcher, TWA concentrations of PCBs in water were calculated from the following equation (Huckins et al. 2006):

                - - = SW S S SW S W K V t R K V N C exp 1 (1)
where C W is the TWA concentration of PCB in water (ng/L), N is the mass of PCB accumulated in sampler (ng), V S is the volume of sampler (L), K SW is the sampler-water partition coefficient of PCB (L/L), R S is the sampling rate of PCB (L/day) and t is the exposure duration (day). For SR and LDPE, V S is replaced by M S , the mass of sampler (kg), and K SW is expressed in L/kg.

For CFIS, TWA concentrations of PCBs were calculated from the following equation, which is a simplification of equation 1 for the sampling during the linear uptake phase (Huckins et al. 2006)

: t R N C S W = (2)
More details on calculations of TWA concentrations of PCBs in water for each sampler are given in Table 2 and Supplementary data (S3).

Results and discussion

Exposure conditions

During the three months exposure, water temperature remained constant (22.6°C ± 0.1°C) and pH only slightly varied (7.5 ± 0.2). In contrast, water conductivity slightly decreased from 380 to 310 µS/cm and DOC concentration showed an increase from around 1.5 to 5 mg/L (Supplementary data, S4 between 3 and 11 %, n=7) and derived water concentrations ranged from 0.37 ng/L (PCB 189) to 3.80 ng/L (PCB 114), with a mean value of 1.29 ng/L. The exposure conditions were therefore considered as constant during the whole experiment.

Comparison of uptake curves and PRC candidates

To compare the uptake of the five studied integrative samplers, this uptake was normalized to a surface area of 100 cm 2 (N A ). Plotting N A versus time showed that the uptake rate (slope) ranged within a factor five as depicted in Figure 1A for PCB 81 (4 Cl atoms), PCB 114 (5 Cl atoms), PCB 138 (6 Cl atoms) and PCB 180 (7 Cl atoms). With exception of PCB 81 (approaching equilibrium for LDPE), the N A for SPMD and LDPE were quite similar, whereas SR was showing considerable higher uptake than SPMD. This could have been caused by the fact that SPMD and LDPE were both similarly fixed on a spider holder perpendicular to the flow, whereas SR was fixed in parallel with the flow (S2). The overall lower N A for Chemcatcher is likely connected to the "beaker" shape configuration creating a longer diffusion path between sampler and bulk water. For CFIS, the N A was not expected to be comparable as this sampler has its own flow regime; but by chance, this N A was at the same level as that of the passive samplers. Remarkable is that CFIS showed the highest N A (PCB 138 and 180) as well as the lowest (PCBs 81 and 114).

It seems that for CFIS, the N A for indicator PCBs were markedly higher than for dioxin-like PCBs, a phenomenon that was not observed for the other samplers.

Although, the N A of most PCBs were linear with time for the whole exposure duration; there were some outliers, indicated by (a) in Figure 1. Indeed, the N A for SPMD exposed for 91 days was similar to that at 56 days, and that even occurred for the most hydrophobic PCBs that could impossibly have obtained equilibrium. The low N A of the SPMD at 91 days exposure however coincides with a reduced release of PRCs, as shown in Figure 1B for PCB 29. It is not clear whether the lower exchange for the SPMD

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exposed during 91 days is caused by observed biofouling or different mounting position of the spider holder giving a different flow regime. Anyway, these observations underline the importance of PRC application. Figure 1B shows the release curves of some PRC candidates spiked in samplers prior to exposure. PRCs were selected according to the criteria reported in Table 2. The release rates of PRCs were used to calculate the TWA concentrations of PCBs in section 3.4.1.

For LDPE strip, PCBs 18, 28 and 52 reached the equilibrium phase of uptake within the 91 days of exposure and PCBs 77, 81 and 101 were in the curvilinear uptake phase, as was PCB 18 in SPMD. With the exception of these less hydrophobic congeners, N A in LDPE strip, as well as in SR, Chemcatcher and CFIS, increased during the three months of exposure with an overall good linearity. Linear uptake phase durations of all PCBs are given in Supplementary data for SPMD, SR and LDPE strip (S5).

Discussion on sampling rates

When N A is divided by (t x C W ), a R SA (L.d -1 .100 cm -2 ) is obtained for the compounds in linear kinetic phase (cf. equation 2). Figure 2 allows comparing the R SA of the different compounds and between the five samplers. It is important to note that R SA for SPMD, SR and LDPE are close with an average RSD of 21 % whereas that value increases to 43 % when data of all samplers are considered. PCBs approaching equilibrium (grey bars) were not included.

Figure 2 is based on the 28 days exposures for SPMD, SR and LDPE, and on the 14 days exposures for Chemcatcher and CFIS, because these exposures were performed in triplicate and allowed an evaluation of the uptake repeatability (error bars in Figure 2).

Overall, repeatability of R SA was very satisfying, with a mean value of RSD (combining all the 19 PCBs) lower than 14 % (9 % for SPMD, 5 % for SR, 7 % for LDPE, 9 % for

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Chemcatcher and 14 % for CFIS). This variability was only slightly higher than that observed in the dosing sheets over time.

Despite the differences in materials (membranes, sequestering phases) and configurations, SPMD, SR, LDPE strip and Chemcatcher exhibited similar patterns of R SA contrary to CFIS. Ignoring the compounds that were approaching equilibrium (PCB 18, PCB 28 and PCB 52), the RSD of the R SA ratios between two samplers (1 ratio per PCB, n=16 to 19) can be used as a measurement for agreement between patterns. This revealed that the pattern ratio between SPMD and SR showed a RSD of 10 % (average ratio of 0.7) and excellent agreement. For both SPMD-LDPE and SR-LDPE, the RSD of the pattern ratio was 16 %, with average ratios of 0.9 and 1.2 respectively. This larger RSD for LDPE was due to relatively higher uptakes of PCBs 77, 81, 126 and 169, i.e. of non-ortho substituted PCBs. For SPMD-Chemcatcher and SPMD-CFIS, the pattern agreement was much lower with RSD of 40 % and 50 % and average ratios of 2.1 and 1.2, respectively. For SPMD-Chemcatcher, leaving PCB 189 out reduced the RSD of the pattern ratio to 23 %. The different level of R SA for the CFIS sampler can be explained by the different flow regime in the cell (outside the tank) compared to that inside the tank, but we cannot explain the much higher R SA of the indicator PCBs compared to those of the dioxin-like PCBs. The RSD of pattern ratios for the membrane samplers were consistent and only about a factor two higher than the repeatability of R SA for each triplicate samplers;

this indicates that uptake processes of the different compounds were similar for SPMD, SR and LDPE.

Between compounds, R SA values were quite scattered (20 to 50 % RSD). To explain this scatter between compounds, the origin of C W should be considered. The calculation of R SA in the evaluation above was actually done according to:

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Dose SW A W A SA C t K N C t N R = = (3)
where C Dose is the concentration in the sheets dosing the water phase. Equation 3 clearly shows that R SA is proportional to K SW and any uncertainty in the K SW is included in C W and, subsequently in the value of R SA . The K SW has a considerable uncertainty (Difilippo and Eganhouse 2010) and can easily be in the same range as the between compound variation of 32, 31 and 41 % observed for SPMD, SR and LDPE respectively. The K SW uncertainty can also explain why the expected decrease of R SA with increasing hydrophobicity (and Mw) [START_REF] Booij | Temperature dependent uptake rates of non-polar organic compounds by semi-permeable membrane devices and low-density polyethylene membranes[END_REF], Huckins et al. 2006[START_REF] Rusina | Calibration of silicone rubber passive samplers: experimental and modeled relations between sampling rate and compound properties[END_REF]) is not visible.

Evaluation of TWA concentrations

Comparison of TWA concentrations

The TWA concentrations of PCBs in water were calculated from PCB uptake and PRC release (except for CFIS) measured for triplicate samplers exposed during 14 days for Chemcatcher and CFIS and during 28 days for SPMD, SR and LDPE strip. The calculations of TWA concentrations were first carried out as indicated by the developer of the sampling system as listed in Table 2 and Supplementary data (S3). For Chemcatcher, the model used for calculations was stated to be only applicable for compounds with logK OW ranging from 3.7 to 6.8 [START_REF] Vrana | Modelling and field application of the Chemcatcher passive sampler calibration data for the monitoring of hydrophobic organic pollutants in water[END_REF]) but was applied also for more hydrophobic PCBs. For SPMD, LDPE strip and Chemcatcher, TWA concentrations of PCBs in water were calculated using only PRCs whose dissipation was between 20 and 80 % in order to prevent quantification problems due to insignificant release or concentrations close to LOQ. In contrast, all PRCs spiked in SR were used (Booij and Smedes, 2010 literature, mostly experimentally determined and modeled for six of the dioxin-like PCBs [START_REF] Smedes | Polymer-water partition coefficients of hydrophobic compounds for passive sampling: application of cosolvent models for validation[END_REF]. For SPMD and Chemcatcher, logK SW were determined from empirical relationships as a function of logK OW (Huckins et al. 2006[START_REF] Vrana | Calibration of the Chemcatcher passive sampler for the monitoring of priority organic pollutants in water[END_REF]. For CFIS, with no PRC used, TWA concentrations of indicator PCBs were computed with R S previously determined in laboratory calibration experiments [START_REF] Llorca | Constantly stirred sorbent and continuous flow integrative sampler. New integrative samplers for the time weighted average water monitoring[END_REF]. Sampling rates of dioxin-like PCBs were extrapolated from R S of indicator PCBs having the same number of chlorine atoms.

For Chemcatcher, repeatability between triplicate samplers was satisfying (12-22 %) but TWA concentrations of PCBs in water were up to 12 times higher than the average of the four other samplers, suggesting that the use of the model [START_REF] Vrana | Calibration of the Chemcatcher passive sampler for the monitoring of priority organic pollutants in water[END_REF][START_REF] Vrana | Modelling and field application of the Chemcatcher passive sampler calibration data for the monitoring of hydrophobic organic pollutants in water[END_REF] for hydrophobic compounds was not applicable for PCBs. Chemcatcher results were therefore not included in the comparison between samplers.

Figure 3 shows the ratios of the TWA concentrations of PCBs calculated from the four samplers and the nominal concentrations of PCBs in water derived from concentrations in dosing sheets. Overall, concentrations computed from the four samplers were reasonably close. The highest difference was observed for PCB 153 with a factor of eight between the lowest calculated concentration (0.5 ng/L for LDPE strip) and the highest (4.0 ng/L for CFIS). For SPMD, calculated TWA concentrations of the 19 PCBs were between 0.8 and 2.1 times (average 1.5) the nominal concentrations in water.

Repeatability between triplicate samplers was between 8 and 26 % (average 15 %). The same tendency was observed for SR, with calculated TWA concentrations between 1.5 and 2.8 times (average of 2.1) higher than the nominal concentrations in water, and with RSD for triplicate samplers between 10 and 27 % (average 18 %). In contrast, for LDPE strip, TWA concentrations were found between 0.4 and 1.2 times (average of 0.7) the nominal concentrations in water with RSD between 7 and 22 % (average 13 %). Finally, for CFIS, TWA concentrations were between 0.5 and 5.0 times (average 1.5) the nominal

Influence of the data treatment method (PRC and K SW ) for SPMD, SR and LDPE strips

The method of calculation (the model used, the selection and the use of PRCs, the choice of the partition coefficients) influences the TWA concentration results for a given sampler. In order to observe the influence of these parameters, TWA concentrations for the 28 days exposures were calculated again for SPMD, SR and LDPE with alternative methods.

For SPMD, initial logK SW used in part 3.4.1., obtained from the empirical model of Huckins et al. (2006), was replaced by new logK SW according to [START_REF] Booij | Correction to An Improved Method for Estimating in Situ Sampling Rates of Nonpolar Passive Samplers[END_REF].

For SPMD and LDPE, we also tested to use all the PRCs spiked (see Table 1), following the method of Booij and Smedes (2010). At the opposite, for SR, instead of using all the PRCs, we considered only PRCs whose release after 28 days were between 20 and 80 %;

three PRCs (i.e. PCBs 2, 3 and 10) were then retained. TWA concentration from SR were then calculated with PCBs 2, 3 and 10 and were found similar (RSD < 7 %), we only present results obtained with PCB 10 because of a smaller RSD.

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The ratios of the TWA concentrations on the concentrations of PCBs in water calculated from dosing sheets are illustrated in Figure 4. For SPMD, in spite of using a different relation for logK SW , the differences on TWA concentrations were relatively small (slight decrease with Ksw from [START_REF] Booij | Correction to An Improved Method for Estimating in Situ Sampling Rates of Nonpolar Passive Samplers[END_REF]. Indeed, the relation mainly affects the more hydrophobic PCBs but not the PRCs (low hydrophobic PCBs).

Consequently, the sampling rate used to calculate C W for the more hydrophobic PCBs is slightly affected. Besides, the use of all the PRCs spiked in SPMD, instead of only one, resulted in a very slight increase of TWA concentrations. For SR, the change in PRC used induced a slight decrease of the TWA concentrations. Indeed, by using only one PRC, these concentrations were between 1.1 and 2.1 times higher than the concentrations in water (with an average of 1.6), instead of 2.1 found with all PRCs. For LDPE, the use of all

PRCs instead of only one resulted in slight increase of TWA concentrations.

Note that the variations in PRC choice above are for illustration. We recommend using all PRCs for R S estimation as no information is lost and uncertainties in logK SW of the PRC may be averaged out. However uncertainty in logK SW remains an issue also for target compounds. Measurement of accurate logK SW is very difficult, experimental logK SW values are scarce and can be considerably scattered [START_REF] Difilippo | Assessment of PDMS-water partition coefficients: Implications for passive environmental sampling of hydrophobic organic compounds[END_REF]. Models predicting the logK SW from logK OW can have typical uncertainties ranging from 0.13-0.36 log unit (factor 1.4-2.4) (Booij and Smedes 2010). Moreover, the selection of other logK OW sources than those used for creating the predictive relations, may contribute to further variability. Considering the above, the results reported here for the three membrane samplers with a general variation of about a factor two are very satisfying, as they were based on logK SW either experimentally determined or from an empirical relationship with logK OW , different calculations models, different selection criteria of PRCs, and obtained from different samplers, processed in different laboratories.

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Conclusions

The designed calibration system for the simultaneous exposure of the five integrative samplers enabled to maintain sufficiently constant exposure conditions up to three months and PCB uptake in samplers showed overall good linearity with time and repeatability. The three membrane samplers (SR, LDPE and SPMD) are efficient to accumulate large amounts of PCBs and have great potential for low LOQ when used in water monitoring programs. TWA concentrations of PCBs in water calculated from the different samplers were in good agreement, except for Chemcatcher whose model for hydrophobic compounds [START_REF] Vrana | Calibration of the Chemcatcher passive sampler for the monitoring of priority organic pollutants in water[END_REF][START_REF] Vrana | Modelling and field application of the Chemcatcher passive sampler calibration data for the monitoring of hydrophobic organic pollutants in water[END_REF] was not proven to be suitable for PCBs in this study. For the four other samplers, despite the variety of materials, geometries and calculation methods, TWA concentrations were generally between 0.5 and 3 times the nominal water concentrations calculated from dosing sheets, which is quite satisfying in the domain of ultra-trace (ng/L level) micropollutants analysis in aquatic environments.

At last, it must be underlined that TWA concentrations in water can be calculated through the use of various models, PRCs and logK SW values. For the transfer of these sampling tools to water basin managers, it is therefore of crucial importance that protocols detail the calculation methods. Moreover, any results on TWA concentration (from the literature or in situ monitoring programs) should be accompanied with detailed information on calculation method used (i.e. model and equations, PRCs and logK SW values).

Intercomparison exercises on sampling and processing, but also the determination of partition coefficients K sw , should enable to progress on the knowledge and harmonization of practices for the use of integrative Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18 3,10,14,29,37,55,78,104,155,166,201,PCB 1,2,3,10,14,21,30,50,55,78,104,145,204 PCB 10,14,29,104,112,204 Bip-d10, Ace-d10, Flu-d10, Phe- a SPMD: semipermeable membrane device, membrane length = 91.4 cm, width = 2.5 cm, thickness = 70-95 µm ; triolein volume = 1 ml, weigth = 0.915 g.

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b SR: silicone rubber, length = 12.5 cm, width = 5.5 cm, thickness = 500 µm.

c LDPE strip: low-density polyethylene strip, length = 20 cm, width = 2.5 cm, thickness = 80 µm. d Chemcatcher: non-polar version, membrane diameter = 4.7 cm, thickness = 40 µm ; C18 Empore disk diameter = 4.7 cm, volume = 0.6 mL (i.e. 0.45 mL octanol + 0.15 mL C18). e CFIS: continuous flow integrative sampler, 90 PDMS pieces with length = 2 mm, width = 2 mm, thickness = 400 µm or 3 Twisters® with length = 2 cm, diameter = 2 mm.

f LDPE: low density polyethylene, PDMS: polydimethylsiloxane.

S1. Additional information about materials and methods

S1.1. SPMD

SPMD were purchased from Exposmeter (Tavelsjö, Sweden) and had the standard configuration, as defined by Huckins et al. (2006): an area-to-volume ratio of about 460 cm 2 /ml of triolein (purity ≥ 95%), an approximate lipid-to-membrane mass ratio of 0.25 and a 70-95 µm wall thickness.

Prior to their exposure, SPMD were spiked with several PRC (Table 1). The membrane was perforated at one end and 25 µl of isooctane containing 4 mg/L of each PRC was injected with a 50 µl syringe. Membrane was then heat-sealed and SPMD were stored in air tight cans at -20°C until e xposure.

After exposure and retrieval, SPMD were stored in air tight cans at -20°C until processing. Prior to extraction, mounting loops were cut and SPMD membranes were cleaned with Milli-Q water, wiped with paper and measured to determine their exact surface area. Recovery of accumulated PCBs was carried out by dialysis in 125 ml of cyclohexane at 15°C during 24h in darkness. This op eration was repeated one time.

Dialysis was performed at 15°C in order to reduce t he amount of co-extracted material [START_REF] Meadows | Large-scale dialysis of sample lipids using a semi-permeable membrane device[END_REF]). Both dialysates were combined and after addition of internal surrogates (100 ng of PCBs 34, 119, 141 and 209), the solvent was evaporated to 1 ml. Then, extracts were diluted (100 µl completed to 1 ml with cyclohexane) and purified on disposable Florisil cartridges (6 ml, 1 g) conditioned with 10 ml of cyclohexane. Extracts were loaded on the cartridges and allowed to soak during 5 min after which elution was performed with 10 ml of cyclohexane/methylene chloride 95/5 (v/v). After evaporation under nitrogen with 10 µl of n-dodecane, used as keeper, extracts were reconstituted in 1 ml of isooctane

Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011 containing 10 µg/L of 2,4,5,6-tetrachloro-m-xylene (TCX) and octachloronaphtalene (OCN), used as internal standards.

Analysis of SPMD extracts were performed with a Varian (Les Ulis, France) 3800 GC-ECD using two chromatographic columns purchased from Restek (Lisses, France), a RTX-PCB column (30 m x 0.25 mm x 0.25 µm) and a RTX-5 column (30 m x 0.25 mm x 0.25 µm). Both columns were equipped with 10 m of uncoated guardcolumns. PCB 157 could not be quantified because of co-eluting peaks on both columns. All results were corrected for recovery of internal surrogates.

S1.2. Silicone rubber (SR)

Silicone rubber sheets of 60 x 60 cm and 0.5 mm thickness were purchased from Altec Products Limited (Cornwall, UK). From this large sheets sampler were prepared by cutting them at a size of 12.5x5.5 cm and were soxhlet extracted with ethylacetate for 100 h prior to use.

To measure of the exchange rate through the release of PRC, 27 exposure sheets were spiked with PRC (Table 1). Spiking was done by soaking the sheets in methanol containing the PRC and followed by adding portions of water to gradually increase the water content [START_REF] Lohmann | Use of passive sampling devices for monitoring and compliance checking of POP concentrations in water[END_REF]. The procedure is similar to that for the dosing sheets, but starting with 300 ml of methanol. Time periods and methanol percentages were equal to the procedure described for dosing sheets (S1.6). After spiking, the sheets were washed with Milli-Q water, individually packed in diffusion closed glass jars numbered from 1 to 27 and stored at -20°C when not exposed in the tank. After thawing, sheets were carefully wiped dry with a tissue before exposure.

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Prior to analysis of the bulk samples, a one day exposure sheet was analyzed to allow estimation of appropriate dilutions ensuring later extracts to fit the calibration curve. Prior to extraction, each sheet was spiked with internal surrogates: 1000 ng of PCB 143 and up to 28 days exposure, 10 ng of 13 C labeled indicator PCB were added. One and three days exposures, that were performed four and two times, respectively, were considered as single exposures and extracted together. Exposed and unexposed sheets were soxhlet extracted by 120 ml of methanol for 16 hours.

Extracts were Kuderna-Danish evaporated until 2 ml and 20 ml of hexane were added. By Kuderna-Danish evaporation, the methanol was azeotropically removed and consequently the extract was transferred to hexane. Hexane extracts were further concentrated to 1 ml using a gentle stream of nitrogen for short exposures and gradually diluted for longer exposures, up to 15 ml of hexane for a 90 d exposure. To each extract TCN was added to obtain a concentration at 100 ng/ml.

Extracts were analysed on an Agilent (Palo Alto, CA, USA) HP 6890 Series GC-MS with an Agilent HP 5973N mass selective detector. The column was an Alltech AT-5MS (30 m x 0.25 mm x 0.25 µm) from Grace (Deerfield, IL, USA).

Selected Ion Monitoring (SIM) mode was applied for quantification using appropriate masses and two masses were monitored for each PCB. All calculations were done based on TCN as an internal standard.

S1.3. LDPE strips

LDPE strips were prepared from additive-free LDPE lay-flat tubing purchased from Brentwood plastics (MO, USA). Single layered strips were obtained by cutting sections of the tubing twice along the side edges. A mounting loop was prepared at each extremity of the strips and removed before the whole analytical treatment.

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Prior to use, LDPE strips were pre-extracted twice by soaking in cyclohexane overnight and then spiked with six PRC (Table 1) following the method described by [START_REF] Booij | Spiking of performance reference compounds in low density polyethylene and silicone passive water samplers[END_REF]. Briefly, LDPE strips were soaked overnight in a PRC solution in methanol/water 80/20 (v/v). Spiked strips were stored at -20°C until exposure.

After exposure, LDPE strips were kept in the dark at -20°C until further treatment. The mounting loops were removed and strips were rinsed with Milli-Q water. Strips were then extracted twice by soaking overnight in cyclohexane.

Surrogate standards (PCBs 30,198,209 and TCN) were spiked into cyclohexane at the beginning of the extraction. After extraction, strips were removed from cyclohexane and allowed to dry for weight determination. The combined cyclohexane extract was concentrated to 4 ml and an aliquot of 200µl was taken for dioxin-like PCB analysis. The exact volume was controlled gravimetrically by weighting both extracts. The analytical protocols for cleanup and analysis of indicator PCBs and PRCs have been described previously [START_REF] Johansson | Polybrominated diphenyl ethers (PBDEs) in mussels from selected French coastal sites: 1981-2003[END_REF]. Briefly, the clean-up and fractionation of all extracts were made by adsorption chromatography on a two layer silica/alumina column. The first fraction eluted with hexane was analysed for PCB by GC-ECD according to the procedure described earlier [START_REF] Johansson | Polybrominated diphenyl ethers (PBDEs) in mussels from selected French coastal sites: 1981-2003[END_REF] reported result was chosen for each non-coeluting congener on the appropriate column. Dioxin-like PCB analysis were conducted at LABERCA laboratory (Nantes, France) according the method of [START_REF] Costera | PCDD/F and PCB transfer to milk in goats exposed to long-term intake of contaminated hay[END_REF]. Separation of coplanar (nonortho) PCBs from non-planar PCBs was achieved on an activated mixture of

S1.4. Chemcatcher

Chemcatchers were purchased from the University of Portsmouth and were prepared according to their own protocol (University of Portsmouth, 2009).

Before use, C18 Empore disks were soaked in methanol overnight in a clean glass beaker. Then, they were placed on a 47 mm diameter disk vacuum manifold platform and 50 ml of methanol were slowly passed through the disks, followed by 150 ml of ultrapure water. Then 250 ml of water spiked with 300 µl of PRC (Table 1) standard solution at 2 µg/ml in methanol was filtered through the disks. The Empore disks were then dried under vacuum during 30 min and put on the sampler PTFE supports. Then, 1 ml of 45% (v/v) n-octanol in methanol was applied evenly to the surface of each C18 Empore disk. The resulting volume of n-octanol was 450 µl. The LDPE membranes (pre-cleaned by soaking in n-hexane during 24h and dried) were put on the top and any air bubbles were smoothed away from between the two layers by gently pressing the top surface of the membrane using a clean paper tissue. The PTFE supporting disks were placed in the sampler bodies and fixed in place to form

Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011 a watertight seal between the membrane and the top section of the sampler [START_REF] Vrana | Performance optimisation of a passive sampler for monitoring hydrophobic organic pollutants in water[END_REF][START_REF] Vrana | Calibration of the Chemcatcher passive sampler for the monitoring of priority organic pollutants in water[END_REF]. Chemcatchers were stored at -20°C until exposure.

After exposure, samplers were rinsed with ultrapure water, carefully disassembled and LDPE membranes were removed and rinsed with acetone. PCBs were extracted from the Empore disks with 5 min of ultrasonic bath in acetone followed by 5 min in ethyl acetate/2,2,4-trimethylpentane 50/50 (v/v). The disks were then removed and the solvent extracts, combined with the LDPE membrane rinsates, were filtered through a drying cartridge containing sodium sulfate. Extracts were reduced under nitrogen at 450 µl and transferred to 2 ml vials prior to analysis with a solution of chrysene-d12 (internal standard) in n-octanol. The final volume was adjusted to 500 µl with n-octanol (University of Portsmouth, 2009;[START_REF] Vrana | Performance optimisation of a passive sampler for monitoring hydrophobic organic pollutants in water[END_REF].

Sampler extracts were analysed with a Varian (Les Ulis, France) 240 GC-MS/MS system using a Varian CP-SIL 8 CB (50 m x 0.25 mm x 0.25 µm) capillary column equipped with a guard-column.

S1.5. CFIS

The CFIS device was developed by Labaqua and was prepared with PDMS pieces or PDMS in Twister ® format from Gerstel (Mülheim an der Ruhr, Germany).

The PDMS pieces were obtained by cutting a PDMS tubing in pieces of 2 x 2 mm.

Every device contains 90 PDMS pieces or 3 Twisters ® . Prior to use, Twisters ® were conditioned in an empty thermodesorption tube at 300ºC for 4 h with an helium flow of 75 ml/min.

After exposure, the Twisters ® or PDMS pieces were removed from the CFIS, gently dried with a paper tissue and finally introduced in glass desorption tubes.

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Analysis were performed by thermodesorption-GC-MS using an Agilent (Palo Alto, CA, USA) 6890 GC -5973 MS system equipped with a Gerstel thermal desorption unit TDS-2 and connected to a Gerstel programmed-temperature vaporization (PTV) injector CIS-4 Plus by a heated transfer line. Analysis were carried out using an TRB-5ms column (30 m x 0.25 mm x 0.25 µm) from Teknokroma (Barcelona, Spain).

S1.6. Dosing sheets

Silicone rubber sheets of 60 x 60 cm purchased from Altec Products Limited (Cornwall, UK) were cut in pieces of 5 x 15 cm and four holes were pinched in the 4-12 cm middle part. A total of 1.1 kg of dosing sheets, corresponding to a surface area of around 3.5 m 2 , were prepared this way. Prior to use, sheets were soxhlet extracted with ethylacetate for one week and subsequently, the ethylacetate was extracted from the sheets by two times 4 h with 2 L of methanol. once with milli-Q water. Two quality control samples were taken just after spiking and another one after the mounting of dosing sheets in the tank.

Before analysis, samples of dosing sheets were stored in glass jars at -20°C.

For each analysis, about 1 g of dosing sheet was extracted with 50 ml of methanol in a glass jar with an aluminum foil lined lid shaked overnight at 100 rpm. Prior to extraction, 1000 ng of PCB 143 was added as an internal surrogate. The extraction was repeated and the combined extract was Kuderna Danish evaporated to about 2 ml followed by addition of 20 ml of hexane and Kuderna Danish evaporation was repeated. The obtained hexane extract was transferred to a 15 ml vial and brought to 10 ml on weight basis. From the extract, 1 ml (on weight basis) was transferred to a vial and 100 ng of TCN (1,2,3,4-tetrachloronaphtalene) were added. Further, a dilution was made by measuring 100 µl into a vial with 100 ng of TCN and adding hexane to 1 ml. Extracts were analysed as described for silicone rubber exposure sheets (S1.2).

Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18

39 . 0 , ,         = i PRC PRC S i S V V R R
Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011 5. Calculation of analyte time-weighted average concentration in water, C W,i (ng/L):

                - - = i SW S i S i SW S i t i W K V t R K V N C , , ,
, , exp 1 6. Calculation of analyte linear uptake phase duration, t 1/2,i (day):

i e i S i SW S i k R K V t , , , , 2
                - - = - PW S i S PW S i t i W K M t F MW A K M N C 47 . 0 , , exp 1
3. Calculation of analyte linear uptake phase duration, t 1/2,i (day): , 2014, vol. 98, pp. 18-27 The 6. Calculation of analyte linear uptake phase duration, t 1/2,i (day):

Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011 Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011

                - - = i DW S i S i DW S i t i W K V t R K V N C , , ,
, , exp 1 6. Calculation of analyte linear uptake phase duration, t 1/2,i (day): Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18-27 The 3.6 ± 0.3 4.2 ± 0.7 10.0 ± 1.2 PCB 28 5.67

3.6 ± 0.3 4.2 ± 0.7 10.0 ± 1.2 PCB 52 5.84

3.5 ± 0.3 4.0 ± 0.6 9.6 ± 1.2 PCB 77 6.36

3.5 ± 0.3 4.0 ± 0.6 9.6 ± 1.2 PCB 81 6.36

3.5 ± 0.3 4.0 ± 0.6 9.6 ± 1.2 PCB 101 6.38

3.4 ± 0.3 3.8 ± 0.6 9.4 ± 1.2 PCB 105 6.65

3.4 ± 0.3 3.8 ± 0.6 9.4 ± 1.2 PCB 114 6.65

3.4 ± 0.3 3.8 ± 0.6 9.4 ± 1.2 PCB 118 6.74

3.4 ± 0.3 3.8 ± 0.6 9.4 ± 1.2 PCB 123 6.74

3.4 ± 0.3 3.8 ± 0.6 9.4 ± 1.2 PCB 126 6.89

3.4 ± 0.3 3.8 ± 0.6 9.4 ± 1.2 PCB 138 6.83

3.3 ± 0.3 3.6 ± 0.6 9.1 ± 1.2 PCB 153 6.92

3.3 ± 0.3 3.6 ± 0.6 9.1 ± 1.2 PCB 156 7.18

3.3 ± 0.3 3.6 ± 0.6 9.1 ± 1.2 PCB 157 7.18

3.3 ± 0.3 3.6 ± 0.6 9.1 ± 1.2 PCB 167 7.27

3.3 ± 0.3 3.6 ± 0.6 9.1 ± 1.2 PCB 169 7.42

3.3 ± 0.3 3.6 ± 0.6 9.1 ± 1.2 PCB 180 7.36

3.2 ± 0.3 3.4 ± 0.6 8.9 ± 1.2 PCB 189 7.71

3.2 ± 0.3 3.4 ± 0.6 8.9 ± 1.2 a values from [START_REF] Hawker | Octanol water partition-coefficients of polychlorinated biphenyl congeners[END_REF] Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011

Florisil/

  Carbopack C/Celite 545. Analyses were performed by GC-HRMS (gas chromatograph (HP-7890) from Hewlett Packard -Palo Alto, CA, USA; mass spectrometer (JMS-800D) from Jeol -Japan) equipped with a a HT8-PCB capillary column (60 m x 0.25 mm x 0.25 µm) from SGE Analytical Science (Australia).

  of F through nonlinear regression by fitting the measured release of PRC with modeled data as a function of K pw and MW -0.47 .[Booij and Smedes, 2010] of analyte time-weighted average concentration in water, C W,i (ng/L):

  version of the article published in Chemosphere

= 4 .

 4 Calculation of analyte i sampling rate, R S,i (L.day -1 ): of analyte time-weighted average concentration in water, C W,i (ng/L):

  of Chemcatcher-water partition coefficient, K DW,i[START_REF] Vrana | Calibration of the Chemcatcher passive sampler for the monitoring of priority organic pollutants in water[END_REF]]: of PRC sampling rate, R S,PRC (L.day -1 ): of analyte sampling rate, R S,i (L.day -1 )[Vrana 2007]: factor taking into account environmental conditions and determined from R s of PRC [Pi = log R s, PRC -22.775 log K ow,PRC +4.061 (log K ow,PRC ) 2 -0.2318 (log K ow,PRC ) 3 ] 5. Calculation of analyte time-weighted average concentration in water, C W,i (ng/L):

  of analyte time-weighted average concentration in water, C W,i (ng/L): of analyte linear uptake phase duration, t 1/2,i (day):

  Author-produced version of the article published inChemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011samplers (non-polar Chemcatcher, SPMD, silicone rod and strip and two modified versions of MESCO), exposed in a flow-through calibration system with Meuse river water spiked with PAHs, PCBs and organochlorine pesticides (concentrations ranging from 20 to 700 ng/L). This laboratory experiment only lasting five days showed that the mass of contaminant absorbed normalized to the sampler surface area was comparable if uptake was controlled by diffusion through the water boundary layer.

	In the context of the ECLIPSE project (2009-2011) 1 , we have studied five integrative
	samplers that well represent the various types used nowadays for PCBs in term of
	receiving phase and configuration (dimensions, holders): SPMD, SR, LDPE strip,

tested in situ the performance of several PSs (including non-polar Chemcatcher, LDPE, membrane enclosed sorptive coating -MESCO, SR and SPMD) for the monitoring of hydrophobic compounds (among PAHs, PCBs or organochlorine pesticides) in the river Meuse (The Netherlands) (Allan et al. 2009) or the river Rhône (France) (Miège et al. 2012) respectively. Although different integrative samplers and methods of calculation were used, relatively consistent TWA concentrations were obtained (variation by a factor up to two). Allan et al. (2010) compared under laboratory conditions the performances of six different integrative Chemcatcher (apolar version) and CFIS. After PRC spiking or not, samplers were exposed under constant agitation and temperature in water contaminated with 19 indicator and dioxin-like PCBs for periods ranging from one day to three months. A constant PCB concentration of about 1 ng/L was achieved by immersing a large amount of spiked silicone rubber sheets

  Author-produced version of the article published inChemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011

). PRC-based sampling rates for SPMD, SR and LDPE strip are given in Supplementary data (S6). Required logK SW for SR and LDPE strip were available in the

  Author-produced version of the article published inChemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011 concentrations in water and RSD between 2 and 80 % (average 21 %). For CFIS, PCBs 180 and 153 were mainly responsible for high mean RSD and TWA concentration, as illustrated in Figure 3. In summary, except for Chemcatcher, TWA concentrations of PCBs computed from the different samplers were in agreement with concentrations in water calculated from dosing sheets and good repeatability was found. These results are very satisfying considering that they were obtained from different samplers, processed in different laboratories and obtained with different calculations methods (i.e. different models, different selection criteria of PRCs and logK SW either experimentally determined or extrapolated from empirical relationship function of logK OW ).

  Author-produced version of the article published inChemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011been found up to date. Further outputs of the ECLIPSE project should follow dealing on the application and comparison of these five integrative samplers in situ.

	Figures
	Fig. 1. Surface area normalized uptake (N A ) plotted versus time (panel A) for PCB 81 (4 Cl
	atoms), PCB 114 (5 Cl atoms), PCB 138 (6 Cl atoms) and PCB 180 (7 Cl atoms) in the five
	integrative samplers. Uptake was normalized to a membrane surface area of 100 cm 2 .
	Panel B shows release curves of some selected PRCs.
	sampling, especially for priority
	chemical monitoring and regulatory programs in compliance with the Water Framework
	Directive and the Marine Strategy Framework Directive. To be noted that the challenge of

PRC strategy is even more crucial for integrative samplers used for hydrophilic compounds (i.e. POCIS, polar Chemcatcher, …), since very few PRC candidates have

from samplers / [PCB] in water SPMD SR LDPE strip CFIS Author

  Surface area normalized sampling rates (R SA , L.d -1 .100 cm -2 ) of PCBs in the five 599 integrative samplers. Error bars represent the standard deviations (n=3). R SA was 600 normalized to a membrane surface area of 100 cm 2 . PCB congeners are on the x-axis. 601 Ratios of the time-weighted average (TWA) concentrations of PCBs calculated from 606 integrative samplers on the PCB water concentrations calculated from dosing sheets. The 607 TWA concentrations were calculated as indicated in Table2. PCB congeners are on the x-608 -produced version of the article published inChemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011

															-27 Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18-27
																							The original publication is available at http://www.sciencedirect.com The original publication is available at http://www.sciencedirect.com
																											doi : 10.1016/j.chemosphere.2013.09.011 doi : 10.1016/j.chemosphere.2013.09.011
	602 609	Fig. 2. PCBs approaching equilibrium are represented with grey bars. Fig. 3. axis.
	603 610	PCB 157 could not be quantified in SPMD because of co-eluting peaks (ND).
	604 611
	605				10										SPMD (28 days)
					9
				Rs in L d -1 100cm -2	3 4 5 6 7 8
					2
					1																						ND
					0
	Rs in L/100cm 2 PCB 28 [PCB] SR PCB 18	PCB 52	18 PCB 28 77	52 PCB 101 81	118 PCB 138 101	SR (28 days) 153 180 77 81 PCB 105 PCB 114 PCB 118	105 PCB 114 123	123 PCB 126 126	156 PCB 157 138	167 PCB 169 153	189 PCB 156	PCB 157	PCB 167	PCB 169	PCB 180	PCB 189
			3
						18		28		52	101		118	138		153		180		77		81		105	114	123	126	156	157	167	169	189
			2												LDPE (28 days)
		[PCB]	1	2 Rs in L/100cm
			0
				PCB	PCB		PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB
				18	28	18	28 52		52 77	101	118 81	138 101	153	180 105	114	77	81 118		105 123	114	123 126	126 138	156	157 153	167 156	169	189 157	167	169	180	189
					10										Chemcatcher (14 days)
					9
				Rs in L/100cm 2	3 4 5 6 7 8
					2
					1
					0
							18	28		52	101	118	138		153	180		77	81		105	114	123	126	156	157	167	169	189
																			CFIS (14 days)
				m 2

Fig. 4. Ratios of the time-weighted average (TWA) concentrations of PCBs calculated from integrative samplers on the concentrations of PCBs in water. The TWA concentrations were calculated as indicated in Table 2 (reference method) and by changing SPMD-water partition coefficient values (for SPMD) and selection criteria of PRCs (for SPMD, SR and LDPE strip). PCB congeners are on the x-axis.

ND: not determined (PCB 157 could not be quantified in SPMD because of co-eluting peaks).

from samplers / [PCB] in water

  Author-produced version of the article published inChemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011

							Reference method: all PRC	One PRC
	LDPE strip
	3
	2
	1
	0
	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB	PCB
	18	28	52	77	81	101	105	114	118	123	126	138	153	156	157	167	169	180	189

[PCB] from samplers / [PCB] in water ND

Table 1 .

 1 Main characteristics and processing steps of the five integrative samplers 620

			SPMD	a	SR	b	LDPE strip c	Chemcatcher	d	CFIS e
	Suppliers		Exposmeter	Altec Products Limited	Brentwood plastics	University of Portsmouth	LABAQUA
	Receiving phase (membrane f , solid or liquid phase)	LDPE + Triolein	PDMS	LDPE	LDPE + C18 Empore disk + 450 µL of n-octanol	PDMS
	Surface area (cm	2 )	457		138	100	17	8
	Weight (g)		4.5		4.1	0.39	0.62	0.144
	Total volume (cm	3 )	4.95		3.4	0.425	0.6	0.147
	PRC spiking method	Syringe injection of isooctane solution	Soaking in water / methanol solutions	Soaking in water / methanol solutions	Percolation of aqueous solution	No PRC
	PRC

g PCB

Table 2 .

 2 Author-produced version of the article published inChemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011 Strategies (models, criteria to select PRCs and methods to evaluate logK SW ) used for the calculation of TWA concentrations of PCBs in water for each integrative sampler.Author-produced version of the article published inChemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011

	TCN: 1,2,3,4-tetrachloronaphthalene

g PRCs: performance reference compounds, Bip: biphenyl, Ace: acenaphthene, Flu: fluorene, Phe: phenanthrene, Pyr: pyrene, B[a]a: benzo[a]anthracene. h

  UK). All PCB congeners were quantified on both columns and theAuthor-produced version of the article published inChemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011

	. For dioxin-like PCB analysis, separation of coplanar (non-ortho) PCB from
	non-planar PCB was achieved on an activated mixture of Florisil/Carbopack C/Celite
	545.
	Analysis of LDPE extracts for indicator PCBs were performed with a Varian
	(Les Ulis, France) 3800 GC fitted with two electron capture detector and two columns
	of different polarities: a DB-5 column (60 m x 0.25 mm x 0.25 µm) from Agilent (Palo
	Alto, CA, USA) and a HT8 column (50 m x 0.25 mm x 0.25 µm) from SGE Europe Ltd
	(Milton Keynes,

  Then the water/methanol mixture was discarded and sheets were washedAuthor-produced version of the article published inChemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011

	Then the dosing sheets were
	immersed in 1.6 L of methanol and the custom-made spiking solution obtained from
	CIL Cluzeau (Courbevoie, France) was added. This solution contained PCB 18 (0.34
	mg), PCB 28 (0.60 mg), PCB 52 (0.9 mg), PCB 77 (2.2 mg), PCB 81 (3.3 mg), PCB
	101 (2.28 mg), PCB 105 (2.6 mg), PCB 114 (9.6 mg), PCB 118 (2.6 mg), PCB 123
	(2.7 mg), PCB 126 (3 mg), PCB 138 (6 mg), PCB 153 (5 mg), PCB 156 (5 mg), PCB
	157 (7 mg), PCB 167 (8 mg), PCB 169 (10 mg), PCB 180 (10 mg), PCB 189 (7 mg)
	in 5 ml of ethylacetate. After 6 h of shaking, water was added to obtain a 90%
	methanol solution that was shaken for 32 h. Dilution with water was continued by
	10% steps as follows: 36 h at 80%, 48 h at 70%, 56 h at 60% and 80 h at 50%
	methanol.

The six holders with passive samplers (from the left to the right : 2 dosing sheets as source of contamination, 1 SPMD, 1 SR, 1 LDPE, 1 Chemcatcher) CFIS outside the tank

  -27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011 Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011 S3. Detailed calculations of TWA concentrations of PCBs in water N t,i : amount of compound i accumulated in the sampler at time t

	V s : volume of sampler
	V i : molecular volume of compound i
	M s : mass of the sampler
	K ow, i : octanol-water partition coefficient of compound i
	A s : surface area of the sampler
	MW i : molecular weight of compound i
	S3.1. SPMD
	1. Calculation of PRC release rate constant, k e,PRC (day -1 ):
	k	e	,	PRC		=	-	ln	(	N	t	,	t PRC			N	0	,	PRC	)
	2. Calculation of SPMD-water partition coefficient, K sw,i [Huckins et al. 2006]:
	log	K	SW	,	i	=	-	2	61 .	+	2	321 .	log	K	OW	,	i	-	0	1618 .	( log	K	OW	,	i	) 2
	3. Calculation of PRC sampling rate, R S,PRC (L.day -1 ):
	S R	,	PRC	=	S V	K	SW	,	PRC	k	e	,	PRC
	4. Calculation of analyte i sampling rate, R S,i (L.day -1 ):

  original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011 Calculation of LDPE-water partition coefficient, K PW,i[START_REF] Smedes | Polymer-water partition coefficients of hydrophobic compounds for passive sampling: application of cosolvent models for validation[END_REF]]:

	S3.3. LDPE
	1. Calculation of PRC release rate constant, k e,PRC (day -1 ):
	k	e	,	PRC	=	-	ln	( N	t	t PRC ,	N	0	,	PRC	)
	2. ( MPF MW K i i i PW + + + = 06 . 1 90 . 0 0141 . 0 log ,	0	.	21	for	tetra	-	ortho	d substitute	PCB	)
	(with	MPF	=	( number meta of total number	) chlorine chlorine atoms atoms para of +	, meta-para chlorine fraction)
	R	S	,	PRC		M	S	K	PW	,	PRC	k	e	,	PRC

3. Calculation of PRC sampling rate, R S,PRC (L.day -1 ):

  original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011S4. Results of the physicochemical analysis of water sampled during exposure 944 945Author-produced version of the article published inChemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011 S6. PRC-based sampling rates of PCBs (L/day/100cm2)

		Day D0	Temperature (°C) 22.61 Log K OW a	pH 7.7 SPMD Conductivity (µS/cm) 380 SR LDPE strip DOC a (mg/L) 1.50
		D3 PCB 18	22.61 5.24	7.3	380	1.20
		D7	22.61	7.8	375	1.90
		D17	22.61	7.5	375	1.35
		D21	22.63	7.5	380	1.80
		D28	22.60	7.7	375	3.25
		D42	22.60	7.2	355	4.50
		D56	22.59	7,0	325	5.15
		D70	22.65	7.5	320	5.15
		D91	22.65	7.6	310	4,70
	946	a DOC: dissolved organic carbon
	947
	948

EChantilLonneurs Intégratifs pour la mesure de PCB dans la phase disSoute de miliEux aqueux,

2009- 2011, coord. Irstea (C. Miège), funded by the French Axelera cluster

Acknowledgments

The ECLIPSE project is part of the French PCB Axelera project, supported by the "Chemistry and environment French competitive" cluster from Lyon and Rhône-Alpes (http://www.axelera.org/en/). We thank Nadège Bely (IFREMER) and Henry Beeltje (TNO -Organisation for Applied Scientific Research, The Netherlands) for technical assistance. The authors thank an unknown reviewer for his constructive comments which helped us to significantly improve the quality of this paper.

Author-produced version of the article published in Chemosphere, 2014, vol. 98, pp. 18-27 The original publication is available at http://www.sciencedirect.com doi : 10.1016/j.chemosphere.2013.09.011

Supplementary data

Supplementary data associated with this article can be found in the online version.


M Petit

Ch Robert-Goumet

L Bideux

B Gruzza

Z Benamara

N Bachir Bouiadjra

V Matolin

Auger electronic spectroscopy and electrical characterisation of InP(100) surfaces passivated by N 2 plasma

Keywords: AES, Indium phosphide(100), Indium nitride, electrical measurements

Auger electron spectroscopy (AES) was used to investigate the processes taking place during the initial stages of InP(100) surfaces nitridation. This AES study combined with electrical measurements (intensity-potential) shows that the processes greatly differ depending on the nitridation angles. Results show that with grazing angle for nitrogen flow, the nitridation process is more efficient. Results obtained with AES spectra are coherent with electrical measurements : Hg/InN/InP(100) Schottky diodes present better electrical characteristics in the case of a grazing flow. That means, the adsorption of nitrogen on the surface is more important for this configuration.

Introduction :

Group III-nitride semiconductors are promising wide band-gap materials for micro and optoelectronic devices since nitride alloys can cover the full visible spectra -from near infrared (Eg= 0,9 eV for InN [1][2]) to the far ultraviolet (Eg=6,2 eV for AlN [3]).

The InP surfaces nitridation has different interests as the realisation of InN buffer layers before the InN films growth, or as passivated layers. Indeed, no protective material exist for III-V semiconductors as SiO2 on Si. On the contrary, the presence of In2O3 or InPO4 on InP surfaces produces important degradation for the electrical properties of the structures. That is why the surface passivation is one of the most important processes in the fabrication of InP-based electronic and optoelectronic devices. Different methods have been employed for the nitridation of InP(100) surfaces :

by N2 + ions beam bombardment [4-5-7] by plasma nitridation [6] It is known that the nitrogen parameters influence the structural and electronic properties of the nitridated surfaces. For instance, previous works [4] have shown that the normal incidence provides maximum efficiency of ion incorporation for N2 + ions beam bombardment (energy range between 0,3 and 1 keV) In this paper we proposed to investigate the influence of the incident angle of the nitrogen flow produces with a plasma source. This nitridation process is monitored by Auger electron spectroscopy. We also compare the different current-tension electrical measurements of samples nitridated by different angles.

Experimental procedures:

S-doped InP(100) samples have been used, they were chemically cleaned ex-situ with successive ultrasonic baths [8] (H2SO4, 3% bromine-methanol desionised water) before introduction in an ultra high vacuum chamber (10 -6 to 10 -7 Pa). InP(100) surfaces have been cleaned by in-situ Ar + ionic bombardment before the nitridation. This ionic cleaning leads to two phenomena : the removal of the contamination layers mainly due to carbon species [9], and the creation of metallic indium droplets in well controlled quantity by preferential phosphorus sputtering [10][11]. These indium droplets will play a key role during the nitridation process. The surface state is monitored by EPES combined with AES [9]. All the in situ experiments have been driven with a retarding field analyzer (OPR 309 Riber).

The nitridation process has been performed with a high voltage plasma discharge cell (glow discharge source (GDS)) to create one or two monolayers of InN on InP(100) substrates through the consumption of indium droplets by nitrogen. In this kind of nitrogen source, a continuous plasma is produced by a high voltage (about 2 kV). A majority of N atomic species are created. However some amount of N2 + is produced but the energy of ions is very low (few eV). Two angles have been studied for the nitrogen flow : 0° (normal incidence) and 80°( grazing incidence) versus the normal of the surface. Low temperature processing is essential in nitridation of InP, indeed the InP decomposition temperature is about 350°C. The substrate is heated to 250°C during the nitridation process, according to previous work investigating the influence of the temperature on the nitridation process [12].

Results and discussion :

1-Auger analysis :

The nitridation of InP (100) substrates was studied using Auger spectroscopy (AES). The measurements have been carried out simultaneously with the nitridation process. To characterise the chemical bonding and composition of the buffer layer growing during the process, the Auger signals of In-MNN, P-LMM and N-KLL transitions at about 403 eV and 411 eV, 117 eV and 383 eV respectively have been used.

 Evolution of the nitrogen Auger peak (N-KLL):

During the nitridation process, the nitrogen Auger signal for the two incident angles of the nitrogen flow has been monitored (figure 1). An increased of this signal is put in evidence. However, we observed a difference between the maximum of Auger intensity between the two angles. For the normal angle, a step is reached from 15 minutes of nitridation time. On the contrary, for the grazing angle, the nitrogen Auger intensity is higher and reaches a maximum at 40 min of nitridation time. After 40 minutes, the N Auger intensity decreases, probably due to a destruction of nitride layers. This first observation can give us the following information : the nitridation process seems to be more efficient with a grazing incident nitrogen flow on the surface.

 Ratio RPIn and RPN :

We have calculated two ratios RPIn and RPN which have been defined in previous work [9,13].

We just remind that the phosphorus Auger peak can be decomposed into three gaussians corresponding to P-In bonds before nitridation and into two groups of three gaussians after nitridation : the first group is related to P-In bonds and the second group to the P-N bonds. RPIn gives information about the amount of P-In bonds and RPN, information about the quantity of P-N bonds produce during the nitridation process. Figure 2 shows the evolution of these two ratios during the nitridation time and for the two incident angles of the nitrogen flow. The graph can be decomposed into three parts : -from 0 to 35 min, RPIn decreases and RPN increases in the both angles with RPIn > RPN : P-N bonds increases but P-In bonds are still the most numerous. -Between 35 and 42 min, RPIn < RPN : the quantity of P-N bonds is higher than the P-In bonds but in the same range for the two angles.

-After 42 min, RPIn becomes greater than RPN once again but the discrepancy is more important in the case of normal angles than in the case of the grazing angle. This phenomena is attributed to the damaging of the surface under the etching effect of the impinging nitrogen flow.

 Observation of the Indium Auger shift :

Figure 3 represents the shift of the indium Auger spectra. The low-kinetic energy shift is more pronounced during the 10-15 first minutes of the nitridation. This shift can be explained by a change in the chemical environment of the indium atoms on the surface. During this short interval, the incoming N atoms substitute for the P atoms and the formation of In-P-N compound is responsible for the observed Auger shift. Indeed, the ionicity of the In-N bonds is higher than that of the In-P bonds due to an increased charge transfer from the In to anion as phosphorus is substituted by nitrogen [14]. We remind the value of Pauling electronegativity of P and N : 2,19 and 3,04 respectively. A shift of the core levels of In to higher binding energies is expected and then a low kinetic energy shift of the In-MNN Auger transition can be observed. We observed a great difference between the two flow angles. The shift is more pronounced for the grazing incident flow than for the normal one. It means that the chemical environment of the indium atoms is higher N content.

To conclude, the Auger spectroscopy demonstrates that the maximum of the nitridation process is around 40 minutes for both incident angles. The P Auger peaks study shows that for the two angles, there is formation of P-N bonds. But N Auger peaks and In shift seem to prove that the nitridation is more efficient in the case of grazing nitrogen flow. After 40 minutes, we observe a decreasing of the P-N bonds and of the quantity of N atoms for the grazing angle. We think it is caused by the destruction of the nitride layers. This damaging may be due to several phenomena which take place simultaneously : dissociation of the substrate surface under temperature [15], evaporation of the layers and sputtering effect of the nitrogen flow. The last phenomenon seems to be weaker for the grazing flow. Further XPS analysis are necessary for a better understanding of this damaging process.

2-Electrical I(V) measurements :

The Hg/InN/InP Schottky diodes -where "InN" is the nitride layers created by the nitridation process-were characterised by electrical measurements (I-V). InP substrates nitridated during 60 min have been used. Parameters like the saturation current Is, the barrier height BN, the serial resistance Rs have been calculated from log(I) = f(V) and I(V) curves and are summarised in table 1. The expression which connects the intensity I to the applied potential V is : [16][START_REF] Sze | Physics of semiconductors devices[END_REF]  

s s B q× V-R ×I I(V)=I ×exp n×k ×T     
 where :

V is the direct applied voltage ; n is the mean ideality factor ; RS is serial resistance of the structure ; T is the temperature ; kB is the Boltzman constant IS is the saturation current.

The barrier height BN is given by :

* 2 BN S A T Φ =kT ln J      
Where :

A * is the Richardson constant ;

JS is the surface density current.

To extract the electrical parameters from the experimental characteristics, we have considered the part of weak polarization (figure 4) from which we determined the saturation current IS which corresponds to a null bias voltage. The shift between the curve of weak bias voltage and that of high biasing gives RS value. The diffusion potential Vd corresponds to the equilibrium of the structure (I = 0) and is quoted to the I(V) curve. The direct and reverse I(V) characteristics are shown in figure 4 for both normal and grazing angles. The direct currents have exponential forms. The direct current for the grazing flow is higher. The electrical properties of the Hg/InN/InP Schottky diodes are influenced by the nitridation process. It means that the passivation effect of the nitride layer is different with the incident nitrogen flow angle.

Conclusions :

The processes taking place during the nitridation of InP(100) surfaces greatly differ depending on the incident angle of the nitrogen flow. Auger spectra demonstrate that the grazing flow is more efficient : the amount of nitrogen is greater, the shift of indium auger peak is more important showing that the chemical environment of the indium atoms is richer in In-N bonds. Then the evolution of RP-In and RP-N ratios shows that the sputtering effect is weaker with the grazing flow, the nitride layers are less damaged. We also observed significant differences between grazing and normal flow with electrical measurements : These observations prove that the passivation effect of the nitride layers depends on the incident nitrogen angle flow.

Figures and tab caption

Tab 1 :Figure 1 :

 11 Electrical parameters calculated from log I(V) and I(V) curves of the Hg/InN/InP Schottky diodes. Evolution of the intensity of the nitrogen Auger Peak during the nitridation time for the two incident angles of the nitrogen flow.

Figure 2 :Figure 3 :

 23 Variation of the ratios RPN and RPIn according to the incident angle of the nitrogen flow. Variation of the shift of the indium Auger peak.

Figure 4 :

 4 log I = f(V) curves of the Hg/InN/InP Schottky diodes for the normal and grazing nitrogen flows. Inset : I(V) curves.

Figure 1 0

 1 Figure 1

Table 1

 1

		Bn (eV)	Is (A)	Vd (V)	Rs ()
	InP (Grazing)	0.66	6 10 -5	2.16	50
	InP (Normal)	0.85	3.16 10 -8	2.43	340