Chemosphere 118 (2015) 268–276

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Low density polyethylene (LDPE) passive samplers for the investigation of polychlorinated biphenyl (PCB) point sources in rivers Nicolas Estoppey a,⇑, Julien Omlin b, Adrien Schopfer a, Pierre Esseiva a, Etiënne L.M. Vermeirssen c, Olivier Delémont a, Luiz F. De Alencastro b a b c

School of Criminal Sciences, University of Lausanne, Batochime, 1015 Lausanne, Switzerland Central Environmental Laboratory, Ecole Polytechnique Fédérale de Lausanne (EPFL), Station 2, 1015 Lausanne, Switzerland Swiss Centre for Applied Ecotoxicology Eawag-EPFL, 8600 Dübendorf, Switzerland

h i g h l i g h t s  LDPE passive samplers were assessed to investigate PCB point sources in rivers.  Influence of velocity on the uptake was evaluated in river-like flow conditions.  Associations between velocity and concentration in samplers (Cs) were determined for indicator PCBs.  Velocity weakly influences Cs and makes LDPE strips an efficient investigative tool.  PCB sources in a Swiss river were rapidly localized by comparing Cs among sites.

a r t i c l e

i n f o

Article history: Received 5 March 2014 Received in revised form 25 August 2014 Accepted 3 September 2014 Available online 13 October 2014 Handling Editor: Andreas Sjodin Keywords: Passive sampling Pollution sources Investigation Water velocity Low density polyethylene (LDPE) Polychlorinated biphenyl (PCB)

a b s t r a c t This study aims to provide a passive sampling approach which can be routinely used to investigate polychlorinated biphenyl (PCB) sources in rivers. The approach consists of deploying low density polyethylene (LDPE) strips downstream and upstream of potential PCB sources as well as in their water discharges. Concentrations of indicator PCBs (iPCBs) absorbed in samplers (Cs) from upstream and downstream sites are compared with each other to reveal increases of PCB levels. Cs measured in water discharges are used to determine if released amounts of PCBs are compatible with increases revealed in the river. As water velocity can greatly vary along a river stretch and influences the uptake at each site in a different way, differences in velocity have to be taken into account to correctly interpret Cs. LDPE strips were exposed to velocities between 1.6 and 37 cm s1 using a channel system built in the field. Relationships between velocity and Cs were established for each iPCB to determine the expected change in Cs due to velocity variations. For PCBs 28 and 52, this change does not exceed a factor 2 for velocity variations in the range from 1.6 to 100 cm s1 (extrapolated data above 37 cm s1). For PCBs 101, 138, 153 and 180, this change only exceeds a factor 2 in the case of large velocity variations. The approach was applied in the Swiss river Venoge to first conduct a primary investigation of potential PCB sources and then conduct thorough investigations of two suspected sources. Ó 2014 Elsevier Ltd. All rights reserved.

1. Introduction 1.1. Scope and aim As a result of the increasing environmental awareness since the 1960s, international, regional and national environmental law has substantially developed over the last decades (Carpi and Schweighardt, 2012). Numerous pieces of legislations aim to ⇑ Corresponding author. Tel.: +41 79 353 79 84; fax: +41 21 692 46 05. E-mail address: [email protected] (N. Estoppey). http://dx.doi.org/10.1016/j.chemosphere.2014.09.032 0045-6535/Ó 2014 Elsevier Ltd. All rights reserved.

reduce (or even cease) emissions of pollutants into the environment (e.g. the Stockholm Convention on persistent organics compounds) and some intend to apply the polluter pays principle (e.g. the European environmental liability directive 2004/35/CE; Drumbl, 2010). In order to reach those objectives, the identification of pollution sources is a key issue and low cost investigative tools are urgently needed (Allan et al., 2006; Mudge, 2008). Because of their persistence and toxicity, polychlorinated biphenyls (PCBs), like other persistent organic compounds, are of great concern. In Switzerland, despite their complete ban since 1986, levels of dioxin-like PCBs in fishes from some rivers have been reported to

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be higher than maximal levels permitted by the European Union (Schmid et al., 2010). According to these authors, these high levels of PCBs measured in some river stretches are mainly due to point sources. The localization and treatment of the most problematic PCB sources are thus realistically achievable. The aim of this study is to develop a methodology which can be routinely used to investigate PCB point sources in rivers. The approach is required to reveal increases of PCB pollution between sites upstream and downstream of potential PCB sources (characteristics of ‘‘potential PCB sources’’ are given in Section 3.2). When potential PCB sources have water discharges to the river, the approach is also required to show if they contain PCBs and if the amounts of discharged PCBs are compatible with the increases revealed in the river. As the pollution at the moment of sampling needs to be determined, PCB measurements have to be carried out in water rather than in sediments (which reflect historical pollution and are mobile) or in fishes (which move). However, due to their hydrophobicity, PCBs are present at very low concentrations in the dissolved phase, typically in the ng L1 to pg L1 range (Jacquet et al., 2014). Therefore, the tool selected to develop the methodology must be sensitive enough to capture low PCB concentrations in river water. As PCB concentrations at the different sites can fluctuate, it is also necessary that the tool takes into account those variations. Finally, the tool has to be flexible enough to be deployed at sites that are difficult to access. 1.2. Passive sampling as an investigative tool For more than two decades, passive sampling has become a powerful tool to monitor organic contaminants in aquatic environments (Huckins et al., 1993). The in-situ enrichment of contaminants in samplers over long-term exposure enables to achieve limits of quantitation (LOQ) low enough to measure hydrophobic contaminants (such as PCBs) in water (Vrana et al., 2005a; Zabiegała et al., 2010; Jacquet et al., 2014). In addition, the use of integrative passive samplers enables to sequester pollutants from episodic pollution. Sampling materials are simple, small and do not require a power supply in the field (Vrana et al., 2005a). These advantages make passive sampling a method of choice to investigate PCB sources in rivers. The use of traditional grab sampling would require sophisticated analytical methods to be sensitive enough and would not take into account variations in concentrations (Jacquet et al., 2014). The use of automated sampling devices would not be practicable since it necessitates secured sites with a power supply. In addition, grab and automated sampling require the transport of water to the laboratory with the connected risks of contamination and losses through wall adsorption (Lohmann et al., 2012). The use of passive sampling to investigative pollution sources was reported by Allan et al. (2006) but there is very little literature on this specific application. The strategy described here deploys passive samplers upstream and downstream of potential PCB sources as well as in their water discharges. Concentrations of PCBs accumulated in the samplers (Cs) depend on concentrations of PCBs in water (Cw) at the different sites. Cs measured at the downstream and upstream sites are compared to reveal increases of PCB levels in the river. Cs measured in water discharges is then used to determine if PCBs are indeed released into the river and if the released amount is compatible with the increase observed in the river. Passive sampling has however a disadvantage which complicates data interpretation: uptake by samplers depends on exposure conditions (water velocity, temperature and biofouling). Water velocity can greatly vary among river sites and is the most problematic exposure parameter in source investigation (see Section 2.3). Therefore, the influence of water velocity has to be taken into account to correctly compare Cs among sites. To compensate for

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variations in velocity, performance reference compounds (PRCs) can be used (Booij and Smedes, 2010). These compounds desorb from samplers and are indicative of the exposure conditions that affect not only their release but also the uptake of similar compounds. However, it implies to load PRCs in samplers before exposure, to extract additional samplers to determine the initial quantity of PRCs, to release pollutants in rivers and to use models to interpret the obtained data. Consequently, a requirement to use PRCs may dissuade investigators from using passive sampling as a routine investigative tool. In this study we propose an alternative method which uses low density polyethylene (LDPE) passive samplers and accounts for variation in water velocity without using PRCs. In a first part, influence of velocity on the uptake was evaluated using a channel system supplied with river water. In a second part, the efficiency of the investigative tool was evaluated by an operational deployment in the Swiss river Venoge. 2. Theoretical considerations relevant for PCB source investigations 2.1. Exchanges between water and samplers Principles governing passive sampling have already been well established for hydrophobic pollutants (Vrana et al., 2005a; Huckins et al., 2006; Booij et al., 2007; Lohmann et al., 2012). The uptake of such pollutants by passive samplers is characterized by an initial linear step, followed by a curvilinear step and an equilibrium step. At a given time (t), the pollutant concentration in a sampler (Cs) is given by:

Cs ¼

   Ns Rs ¼ C w K sw 1  exp  t ms K sw ms

ð1Þ

where Ns is the mass of pollutant in the sampler, ms is the mass of the sampler, Cw is the concentration of the pollutant in water, Ksw is the sampler-water partition coefficient and Rs is the (initial) sampling rate. Rs corresponds to the volume of water that the sampler clears of pollutant per unit of exposure time (t) immediately after deployment. It is proportional to the sampler surface area (A) and inversely proportional to the mass transport resistance (Io) of pollutants from the bulk water to the receiving phase. In the case of single-phase samplers, like LDPE, Io is the sum of the resistances in three phases: the water boundary layer (w), the biofilm layer (b) and the receiving phase (s). In defining each of these resistances by the thickness of the particular phase (di), the diffusion coefficient in the particular phase (Di) and the partition coefficient between the particular phase and water (Kiw), Rs is given by Eq. (2) (Vrana et al., 2005b):

Rs ¼

A dw Dw

þ

db Db K bw

þ DsdKssw

ð2Þ

2.2. Sampler characteristics The selected sampler is required to absorb a quantifiable amount of pollutant (Ns) in an acceptable sampling time (e.g. 6 weeks). According to Eq. (1), the maximal Ns that samplers can absorb (regardless of time constraints) depends on Ksw and ms, or in other words, on their affinity with the pollutant and their storage capacity. Among the samplers reviewed by Vrana et al. (2005a), semipermeable membrane devices (SPMDs), low density polyethylene (LDPE) and silicone strips are the ones which offer a good affinity for all PCBs and the best sensitivity (detection at sub pg-level). LDPE and silicone tend to become more popular because they are single-phase (low cost, simpler construction and extraction) (Rusina et al., 2007; Smedes et al., 2010). In order

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to reach a quantifiable Ns in a routinely acceptable sampling time, samplers have to possess a sufficiently high Rs. According to Eq. (2) (Vrana et al., 2005b), samplers designed to have a high surface area (A) and a thin receiving phase (ds) offer a high Rs. Thus, the use of strips is optimal. Finally, Rs is also increased with the use of a receiving phase which has high diffusion coefficients (Ds) for the investigated pollutants and a high Ksw. Ds of PCBs are higher in silicone than in LDPE (about 2.3–3.2 units of log) but LDPE is thinner (2 units of log if 0.05 mm thick LDPE from this study is compared with the 5 mm thick AlteSil silicone) (Rusina et al., 2010a). According to Smedes et al. (2009), sampler-water partition coefficients of PCBs are of the same order of magnitude for LDPE and silicone (maximal difference of 0.5 units of log). Thus, for a same surface area, Rs is slightly higher in silicone than in LDPE. Regarding practical aspects, LDPE is easier to handle in the laboratory as it swells less in solvents compared to silicone. In addition, the high release of oligomers from silicone can severely disrupt analysis if they are not properly removed (Rusina et al., 2007; Smedes et al., 2010). 2.3. Exposure conditions Rs depends on the exposure conditions such as water velocity, water temperature and biofouling (Lohmann et al., 2012). An increase of water velocity intensifies turbulences at the sampler surface and decreases the water boundary layer thickness (dw) (Mayer et al., 2003). The resulting decrease of transfer resistance in this phase leads to an increase of Rs (Eq. (2)). An increase of the amount of biofouling causes an increase of the biofilm layer thickness (db) and thus a decrease of Rs (Vrana et al., 2005a). Finally, an increase of temperature leads to an increase of diffusion coefficients (Ds, Db and Dw) and an decrease of partition coefficients (Ksw, Kbw), therefore causing opposite effects on Rs. As exposure conditions can vary among sites, they can cause different concentrations in samplers for equal contamination levels (Huckins et al., 2002; Smedes, 2007). Fig. 1 illustrates the situation in which samplers are deployed upstream and downstream of a potential source of pollution. The concentration in samplers (Cs) at each site is given by Eq. (1). An increase between Cs,upstream and Cs,downstream can be due to an increase of aqueous concentration (Cw,downstream > Cw,upstream) but also to an increase of sampling rate (Rs,downstream > Rs,upstream). In order to correctly determine the existence (or the absence) of PCB emissions between two sites, the influence of exposure conditions has to be evaluated. This requirement also applies to the situation in which sampling is carried out in water discharges of potential PCB sources. In addition, to evaluate if PCBs are released into the river (Cs,discharge), it is assessed if the released amount matches the increase observed in the river, that is, if the following mass balance equation is satisfied:

C s;discharge Q discharge ¼ C s;downstream Q downstream  C s;upstream Q upstream

ð3Þ

where Qdischarge, Qupstream and Qdownstream are the volumetric flow rates of the water discharge and of the river at the upstream and the downstream sites respectively. To determine whether the

relation is satisfied, the influence of exposure conditions on Cs,discharge, Cs,upstream and Cs,downstream must be taken into account. Temperature and biofouling are not problematic in source investigation in rivers since only very limited variations are expected between two adjacent river sites. In addition, according to literature, the impact of temperature and biofouling on the uptake is small. In SPMDs, an increase of temperature of 20 °C leads to an increase of the uptake rate of about a factor 2 (Booij et al., 2003). Ksw (which appears not only in the definition of Rs) is largely independent of temperature for PCBs (Lohmann et al., 2012). In fouling conditions the decrease of the uptake was reported to be of a factor 2 by Richardson et al. (2002) or even to be not significant by Harman et al. (2009). On the contrary, water velocity is a problematic parameter in source investigation in rivers. Indeed, high variability is expected among sites due to changing riverbed morphology, for example, slope and cross-section variations. In addition, according to literature, the impact of velocity on the uptake can be very high depending on the range of encountered velocities. Using SPMDs, Vrana and Schüürmann (2002) reported increases of penta and hexachlorobenzene uptakes respectively of factors about 3 and 9 for an increase of velocity from 0.06 to 0.28 cm s1. Booij et al. (1998) reported an increase of PCB uptake by SPMDs of a factor of about 3 for an increase from 1 cm s1 to 30 cm s1. Using silicone, Rusina et al. (2010b) measured an increase of polyaromatic hydrocarbon (PAH) uptake of about a factor 10 for an increase from 0.14 to 9 cm s1. As shown by Booij et al. (1998), the uptakes by SPMD and LDPE strips are very similar since both of them have a common phase: the LDPE membrane. While uptakes by LDPE strips and SPMD are expected to be influenced by exposure conditions in a similar way, the uptake by silicone is expected to be more influenced by water velocity. Indeed, Eq. (2) reveals that (in the absence of biofouling) the uptake is rate limited by the membrane when dw/Dw  ds/DsKsw and by the water boundary layer when dw/Dw  ds/DsKsw (Huckins et al., 2006; Allan et al., 2010). Silicone has a ds/Ds ratio slightly smaller than LDPE and tends to be more water boundary layer controlled. Thus it is expected to be more affected by variation of velocity. Based on this consideration and the advantages described in the previous chapter, LDPE was selected in this study. 3. Experiments Experiments to evaluate the influence of velocity on the uptake (part A) and to investigate pollution sources (part B) were conducted in the Venoge River (Switzerland). This river was selected because levels of dioxin-like PCBs in fishes were reported to be higher than maximal levels permitted by the EU (Schmid et al., 2010). The Venoge River flows 36 km from the foot of the Jura Mountains to Lake Geneva. It has a mean flow of 4.13 m3 s1 (FOEN, 2012) and a watershed of 241 km2 with 9% of urban area in 2003 (Trevisan et al., 2012). The six indicator PCBs (iPCBs, IUPAC nos. 28, 52, 101, 138, 153 and 180) were selected because they are present at higher levels than other PCBs in the environment. They are thus useful for routine investigation. Since PCBs were

Fig. 1. Samplers deployed upstream and downstream of a potential PCB source. Concentration in samplers (Cs) is given by Eq. (1).

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commercialized in the form of mixtures, a source which releases iPCBs also releases other PCBs such as dioxin-like PCBs. 3.1. Part A: Influence of velocity using a channel system A channel system adapted from Vermeirssen et al. (2008) was built at a site situated 1.5 km from the Venoge River mouth (see map and photographs in Supplementary data S1). The system was mounted on a formwork panel (2.5 m long, 0.5 m wide) supported by six table legs with adjustable height. The formwork panel was situated about 1.5 m below the river’s water level. On one of the ends of the formwork panel, a Plexiglass dispenser box (0.70  0.40  0.45 m) was raised of 0.23 m using two plastic crates. Three PVC tubes (51 mm diameter) delivered water to the box by siphoning. At the bottom of the box, four steel vents (51 mm diameter) supplied water to four Plexiglass channels (2 m long, 0.09 m wide, 0.15 m high) placed directly on the formwork panel. Each vent was fitted with a PVC screw cap with a hole in its centre. The hole in each of the vents had a different diameter to obtain four desired volumetric flow rates. Flow was maintained constant by using an overflow in the dispenser box (large slit to release surplus water) that lead to a constant hydraulic head of 0.30 m in the box. The ends of the channels near the box were closed, at the other end the channels were fitted with knife gate valves (made of Plexiglass) enabling to adjust the water level in each channel. The valves apertures were set to obtain a minimum water level of 0.09 m and the desired water velocities. A first experiment (A.1) was conducted with hole diameters of 24, 30, 34 and 42 mm, leading to volumetric flow rates of 0.66, 1.41, 2.28 and 3.13 L s1 and velocities of 6.3, 13.1, 28.4 and 37.0 cm s1. A second experiment (A.2) was carried out with a reduced lowest velocity (hole diameter of 10 mm instead of 24 mm) and slightly adapted water levels in the channels. Accordingly, water velocities in this second experiment were: 1.6, 12.9, 25.7 and 35.8 cm s1. As a verification, velocities were also measured using two other methods: an anemometer (Flowatch, JDC Electronic SA, Switzerland) and a measure of time required for particles to flow a distance of 1 m. 3.2. Part B: Investigation of pollution sources A primary investigation of potential sources was carried out in autumn 2011 (part B.1). According to Schmid et al. (2010), main PCB point sources are landfills and other sites contaminated by old material containing PCBs (e.g. capacitors or joint sealing). Wastewater treatment plants and channelized runoffs from urban surface are considered as secondary PCB point sources. In the Venoge, an inventory was made of outlet pipes and tributaries coming from (or flowing near) sites with these characteristics. Ten potential sources were selected. One sampler was deployed upstream and downstream of each of those potential sources as well as in the water discharges (see map in Supplementary data S2). Several sources were detected and thorough investigations were conducted in autumn 2012 (part B.2). Two samplers were deployed upstream and downstream of these sources. They were immersed close to the suspected sources (30–100 m) in order to exclude other visible water discharges between the two sites. Two samplers were also deployed in these water discharges. At each sampling site, water velocity was weekly measured in front of the samplers using the same anemometer as in part A.1. 3.3. Material LDPE was obtained from Semadeni AG (Switzerland) in the form of a roll of tubular sheet of 0.30 m width and 0.05 mm thickness. Indicator PCBs (nos. 28, 52, 101, 138, 153 and 180) as well as

271

PCB 189 were purchased from Dr. Ehrenstorfer GmbH (Germany). Labelled iPCBs (13C, 99%) were obtained from Cambridge Isotop Laboratories (USA). All solvents used for the extraction, purification and analysis steps (dichloromethane, hexane, methanol and isooctane) were of pesticide grade and were purchased from Romil LTD (United Kingdom) and Carlo Erba (France). Florisil adsorbent (60– 100 mesh) was purchased from Sigma–Aldrich (USA). Sulphuric acid was obtained from Merck KGaA (Germany). 3.4. Preparation LDPE sheet was cut into 0.30  0.09 m (part A) and 0.30  0.03 m (part B) strips. The required number of strips for each experiment was cleaned by Soxhlet extraction (1 L extractor) using dichloromethane (24 h) and methanol (24 h). Then, the strips were transferred into a 1 L sealed amber glass bottle and kept in the fridge until deployment. 3.5. Deployment and recovery In part A, a sampler consisted of two 0.30  0.09 m strips. Each of them was fixed to an iron rod (diameter: 5 mm) using staples. The two iron rods were maintained side by side in the channel by a Plexiglass block. In each channel, three samplers were immersed (one behind the other) for 6 weeks. Every week velocity was measured and the three samplers were rotated in order that each sampler spent a total of 2 weeks in each position. In part B, samplers consisted of six 0.30  0.03 m strips. They were knotted to an iron bar (diameter: 10 mm) and tied with wire. The iron bars were fixed into the bed of the river (or secured with a cord in source discharges). At each location one sampler was deployed for the primary investigation and two samplers were deployed for the thorough investigation. Every week, any branches and leaves which were caught on the samplers were removed. In addition, the velocity and the temperature were measured. At the end of the 6-week sampling period, the strips of each sampler were placed in separate closed aluminium containers. In the lab, they were briefly rinsed with Milli-Q water and patted dry with a tissue paper (part A) or dried on aluminium sheets for 24 h (part B). Then, the strips were cut to 20 cm length, weighted and stored in the freezer (20 °C) until extraction. In each experiment, three samplers were used as field controls. They were not immersed in river water, but exposed to air during deployment and recovery of deployed samplers. The field controls were kept in the freezer during the sampling period. 3.6. Extraction and clean-up The strips of each sampler (including field controls) were weighed and cut into pieces (about 0.01  0.03 m). They were placed into separate 100 mL glass Soxhlet thimbles. Previously 1 cm of anhydrous sodium sulphate was added on the sintered discs of the thimbles. One mL of a solution containing each 13Clabelled indicator PCB at 25 ng mL1 was added as a surrogate standard to the strips used in parts A and B.2. A PCB 189 solution (0.5 mL of 25 ng mL1) was added to the strips used in part B.1. The extraction was carried out for 16 h with 150 mL of dichloromethane (70 °C). The extracts were solvent exchanged to hexane and reduced to 1 mL in a rotary evaporator. Extracts underwent clean-up on Florisil columns (5 g previously deactivated with water (4% by wt.) and protected with 4 g of sodium sulphate) and were eluted using 50 mL of hexane. The extracts were reduced to 1 mL and transferred to centrifuge tubes in where a final clean-up was performed with 1 mL of sulphuric acid (H2SO4). After centrifugation (3000 rpm for 10 min), the supernatants were transferred to 2 mL GC vials and reduced to 0.5 mL under a nitrogen

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stream. They were solvent exchanged to isooctane and reduced to 0.3 mL. Finally, extracts were transferred to GC vials with inserts and stored at 20 °C until analysis. 3.7. Analysis Extracts from part A were quantified by GC–MS/MS (Thermo Scientific: Trace 1310 coupled with TSQ Quantum XLS Ultra). Two lL of the extracts were injected at 280 °C in a splitless mode. Helium was used as carrier gas and Argon as collision gas. All extracts were first injected on a ZB-5MS column (60 m  0.25 mm, id 0.25 lm) with a constant flow rate of 1.5 mL min1. The temperature program started at 80 °C (0.5 min), increased to 160 °C (20 °C min1), and then to 300 °C (4 °C min1) and was held for 10 min. The mass spectrometer was operated in electron impact at 70 eV in the selected reactive monitoring (SRM) mode. For each PCB congener, two SRM transitions were used for quantitation and confirmation. The transfer line and ion source temperatures were set at 290 °C and 250 °C respectively. As a confirmation, all extracts were injected on a ZB-XLB column (20 m  0.18 mm, id 0.18 lm) with a flow rate of 1.2 mL min1. The temperature program started at 80 °C (0.5 min), increased to 170 °C (60 °C min1), and then to 300 °C (7 °C min1) and was held for 9.5 min. All other parameters were identical to the ones set for the injections with the ZB-5MS column. Extracts from part B were quantified by GC– MS (Agilent Technologies: 7890A GC System coupled with 5975C inert XL MSD). Two lL of the sample was injected at 250 °C in a splitless mode. Helium was used as carrier gas with a constant flow mode (1 mL min1) with a HP-5MS column (30 m  0.25 mm, id 0.25 lm). The temperature program started at 50 °C (2 min), increased to 100 °C (15 °C min1), and then to 290 °C (10 °C min1) and was held for 15 min. The mass spectrometer was operated in SIM mode with one quantifier and two qualifiers ions per chlorinated class. The temperature of the transfer line, the ion source and the quadrupole were set up at 300 °C, 230 °C and 150 °C respectively. In GC/MS sequences, a calibration curve (3–500 ng mL1 of each iPCB, 125 ng mL1 of iPCB 13C as internal standard) was analyzed after every six samples. The surrogate standard was used to compensate for losses during extraction, clean-up and measurement. The final result was expressed in mass of PCB (ng) per mass of LDPE (g).

4. Results and discussion 4.1. Part A: Influence of velocity using a channel system The concentration of each iPCB absorbed in LDPE strips (Cs) after the 6-week sampling period is given in Fig. 2. Each bar indicates the mean Cs calculated from three samplers deployed in the same channel. Cs were overall higher in experiment A.2 than in experiment A.1. This can be explained by the fact that the average river flow was of 8.2 m3 s1 during experiment A.1 and of 2.0 m3 s1 during experiment A.2 (FOEN, 2013). Assuming that PCB sources emitted similar amounts of PCBs during both experiments, the difference of river flow led to an aqueous concentration more than four times higher in A.2 than in A.1. ANOVA (p = 0.05) and Tukey (a = 0.05) tests were used to compare mean Cs from each channel. In experiment A.1, Cs of PCBs 28 and 52 did not increase significantly with increasing velocity. It appears that after 6 weeks in the studied range of velocities (6.3– 37.0 cm s1), equilibrium between LDPE and water must have been reached (or almost reached) for PCBs 28 and 52. For PCBs 101, 138, 153 and 180, a velocity increase from 6.3 to 37.0 cm s1 led to significant Cs increases of factors of about 1.3. However, there was no statistically significant Cs increase between the lowest velocities (6.3 and 13.1 cm s1) and between the highest velocities (28.4 and 37.0 cm s1). In experiment A.2, an increase of velocity from 1.6 to 35.8 cm s1 led to significant increases of Cs for all iPCBs. Cs increased of factors of about 1.4, 1.5, 2.2, 2.3, 2.4 and 2.7 for PCBs 28, 52, 101, 138, 153 and 180, respectively. Surprisingly there was a small but significant decrease of Cs with a velocity increase from 25.7 to 35.8 cm s1. This could probably be induced by heavy rainfall during the last day of the sampling period. On that day the average river flow rate reached 12.4 m3 s1 whereas it was only 1.7 m3 s1 for the 41 other days of the sampling period. The resulting reduction of aqueous concentrations probably caused a slight desorption of iPCBs from LDPE. Since this release increased with velocity, it could explain part of the small decrease of Cs from 25.7 to 35.8 cm s1. In order to obtain a relationship between Cs and velocity, data of experiments A.1 and A.2 were fitted with the model given by Vermeirssen et al. (2008): c

C s ¼ að1  exp½bv Þ

ð4Þ

Fig. 2. Concentration of iPCBs absorbed in LDPE strips (Cs) deployed in the channel system for 6 weeks in experiment A.1 (left) and A.2 (right). Error bars represent the standard deviation (n = 3). Limits of quantification (LOQ) were measured from three field controls.

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The fitting was done using Matlab R2012b. Fig. 3 shows the fitted data of PCBs 28 and 101 in experiments A.1 and A.2 The fitted data of PCBs 52, 138, 153 and 180 are given in Supplementary data (S3). Curve fitting in experiment A.1 suffers from the absence of data at very low velocities. It is particularly problematic for PCBs 28 and 52 which reached equilibrium and to which the model assumptions may not be applicable (Vermeirssen et al., 2008). The velocity of 1.6 cm s1 in experiment A.2 enabled to improve the precision of fitting at low velocities. Thus, only the results from experiment A.2 were used to take into account the influence of velocity in part B (source investigation). The results from experiment A.1 were used to ensure that results of experiments A1 and A.2 were in the same order of magnitude. Using the fitted equations, Cs was calculated at different velocities (ranging from the lowest tested velocity and extrapolated to 100 cm s1). Double entry tables were constructed to offer the investigator the possibility to determine the expected factor of increase of Cs due to velocity increases between two sites. These tables are given in Supplementary data (S4) and examples of tables for PCBs 28 and 101 (constructed from experiment A.2) are given in Table 1. The maximal velocity shown in the double entry tables is 100 cm s1 because it corresponds to maximal velocities measured in the Venoge. If investigations are to be carried out in rivers with higher flow rates, the fitted equations can be used to extrapolate increases of Cs for higher velocities. Both experiments A.1 and A.2 revealed that the influence of velocity on the uptake by LDPE

strips is low in the studied velocity range. For PCBs 28 and 52, the increase of Cs does not exceed a factor 2 (grey cells in double entry tables) whatever the velocity variation. For PCBs 101, 138, 153 and 180, the increase of Cs exceed a factor 2 (blank cells in double entry tables) only in the case of large differences in velocity between two sites. The factor by which Cs changes due to velocity increases between two sites is called Fhighest/lowest. In order for Fhighest/lowest to always exceed 1, it was chosen to always refer to velocity increases between two sites (not to velocity decreases). Thus, when velocity is higher at the downstream site than at the upstream site (vdownstream > vupstream), Fhighest/lowest is written as Fdownstream/upstream and in the opposite case (vupstream > vdownstream), Fhighest/lowest is written as Fupstream/downstream. Table 2 illustrates the situations an investigator can encounter and the way the investigator should use the factor Fhighest/lowest. 4.2. Part B: Investigation of pollution sources 4.2.1. Primary investigation of pollution sources The upper graphs of Fig. 4 give the concentrations of PCBs 28 and 101 absorbed in LDPE strips deployed downstream and upstream of potential PCB sources in the Venoge as well as in the discharges of potential sources. The lower graphs give the average velocity measured at each site. The results of PCBs 52, 138, 153 and 180 are given in Supplementary data (S5). Comparisons between Cs,upstream and Cs,downstream were done only if both concentrations

Fig. 3. Data obtained from the channel system experiments for PCBs 28 and 101 fitted with Eq. (3). Experiment A.1 (left) and experiment A.2 (right). Error bars represent the standard deviation (n = 3). The fit obtained for PCB 28 in experiment A.1 (dotted line) was not satisfactory (i.e. low R-square value).

Table 1 Expected factor of increase of Cs due to velocity increases between two sites. This factor is referred to as Fhighest/lowest in the text and was obtained by calculating the ratio of Cs measured (or extrapolated) at the two velocities of interest in experiment A.2 (Cs measured at the highest of the two velocities divided by Cs measured at the lowest of the two velocities).

1.6 1.6 1.0 5 1.2 10 1.3 15 1.3 20 1.4 30 1.5 35.8 1.5 40 1.5 50 1.6 100 1.7

5 1.0 1.1 1.2 1.2 1.3 1.3 1.3 1.4 1.5

Site with lowest velocity[cm s-1] 10 15 20 30 35.8 40 50 100

1.0 1.1 1.1 1.2 1.2 1.2 1.2 1.3

1.0 1.0 1.1 1.1 1.1 1.2 1.3

1.0 1.1 1.1 1.1 1.1 1.2

1.0 1.0 1.0 1.1 1.2

1.0 1.0 1.0 1.1

1.0 1.0 1.0 1.1 1.1 1.0

PCB 101 Site with highest velocity [cm s-1]

Site with highest velocity -1 [cm s ]

PCB 28

1.6 1.6 1.0 5 1.4 10 1.6 15 1.8 20 2.0 30 2.2 35.8 2.3 40 2.4 50 2.5 100 3.1

5 1.0 1.2 1.3 1.5 1.6 1.7 1.8 1.9 2.2

Site with lowest velocity[cm s-1] 10 15 20 30 35.8 40 50 100

1.0 1.1 1.2 1.3 1.4 1.5 1.5 1.9

1.0 1.1 1.2 1.3 1.3 1.4 1.7

Dotted lines separate interpolated and extrapolated results obtained with the fitting equations of experiment A.2.

1.0 1.1 1.2 1.2 1.3 1.5

1.0 1.0 1.1 1.1 1.4

1.0 1.0 1.0 1.1 1.1 1.0 1.3 1.3 1.2 1.0

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Table 2 Situations that can occur when investigating PCB sources and resulting conclusions.

a

The absence of PCB sources theoretically leads to ‘‘Cs,downstream/Cs,upstream = Fdownstream/upstream’’. Potential removal processes (e.g. sorption to particles followed by settling) occurring between the upstream and downstream sites could however lead to ‘‘Cs,downstream/Cs,upstream < Fdownstream/upstream’’. The investigator must be aware that such removal processes could hide the impacts of PCB sources on the river. b The absence of PCB sources theoretically leads to ‘‘Cs,upstream/Cs,downstream = Fupstream/downstream’’. Potential removal processes occurring between the two sites could however lead to ‘‘Cs,upstream/Cs,downstream > Fupstream/downstream’’. Such removal processes could hide the impacts of PCB sources on the river. c Even if Cs,downstream 6 Cs,upstream, the fact that Cs,upstream/Cs,downstream is lower than Fupstream/downstream means that an increase of PCB concentration occurs between the two sites, i.e. that one or several PCB sources are present.

Fig. 4. Upper graph: Concentration of PCBs 28 (left) and 101 (right) absorbed in LDPE (Cs) strips deployed downstream and upstream of potential PCB sources in the Venoge (solid line) as well as in discharges of potential sources (dashed lines). Lower graph: Average velocity at each sampling site.

were above LOQ (dotted line). If only one of the two concentrations was above LOQ, that one was compared with the LOQ value. Cs,discharge was used to determine if water discharges of potential PCB sources effectively contain PCBs. Four problematic sites were localized in this primary investigation. In each of these cases, vdownstream was higher than vupstream and it had to be verified that the ratio Cs,downstream/Cs,upstream was higher than the factor Fdownstream/upstream. The first case was the increase between Cs,site10 and Cs,site12 which corresponded to sites upstream and downstream of a wastewater treatment plant (WWTP). The ratios Cs,site12/Cs,site10 of PCBs 28, 101, 138 and 153 (respectively 2.7, 1.7, 2.0, 2.0) were higher than the factors Fsite12/site10 (1.2, 1.4, 1.5, 1.5). Cdischarge showed that the discharges of the WWTP contained PCBs. The second case was the increase between Cs,site12 and Cs,site14 which corresponds to sites upstream and downstream of a settling basin collecting water from industrial sites. Only the ratio Cs,site14/Cs,site12 of PCB 28 (1.3) was higher

than Fsite14/site12 (1.1). High Cdischarge of the six iPCBs were measured in the discharge of the basin. The third case was the increase between Cs,site14 and Cs,site15 which corresponds to sites upstream and downstream of several pipes whose origins were not known (probably industrial sites). The ratio Cs,site15/Cs,site14 of the six indicator PCBs (1.4, 1.5, 1.8, 2.1, 2.0, 1.8) was higher than Fsite15/site14 (1.0, 1.0, 1.1, 1.1, 1.1, 1.1). The fourth case is the increase between Cs,site16 and Cs,site17 which corresponds to sites upstream and downstream of a landfill. The ratio Cs,site17/Cs,site16 of PCBs 28 and 101 (1.8, 1.6) was higher than Fsite17/site16 (1.2, 1.5).

4.2.2. Thorough investigation of two PCB sources The WWTP and the settling basin were investigated in more detail. Fig. 5 give the concentration of the six iPCBs absorbed in LDPE deployed upstream and downstream of these two suspected PCB sources as well as in their discharges.

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Fig. 5. Concentration of the six iPCBs absorbed in LDPE strips (Cs) deployed upstream and downstream of two suspected PCB sources: a wastewater treatment plant (left; WWTP) and a settling basin (right). Error bars represent the standard deviation (n = 2).

Comparisons between Cs,upstream and Cs,downstream were done using a t-test (p < 0.05). In the case of the WWTP, Cs,downstream was significantly higher than Cs,upstream for all indicator PCBs. As vupstream was higher than vdownstream, the increase of Cs between the two sites could not be due to velocity but to the presence of the WWTP. The average flow rate of the WWTP (Qdischarge) was 0.03 m3 s1, the flow of the Venoge (downstream of the WWTP) was 1.90 m3 s1 (Qdownstream) (FOEN, 2012). To verify the mass balance (Eq. (3)), Cs,upstream, Cs,downstream and Cs,discharges where normalized to the downstream site which presented the lowest velocity. They were thus divided respectively by Fupstream/downstream, 1 and Fdischarge/downstream. The product ‘‘Qdischarge Cdischarge’’ is about 7.5 times lower than the term ‘‘Qdownstream Cdownstream  Qusptream Cupstream’’. The mass of discharged PCBs explained thus only 14% of the increase between the upstream and downstream sites. The most likely reason is that the downstream sampling site was situated only 30 m from the water discharges. The mixing of waters from the WWTP and from the river was probably not complete. The measurement of conductivity (carried out after the investigation) showed that it was not homogenous across the whole width of the river. Thus, if practicable, the measurement of conductivity should be conducted by the investigator to optimize the selection of sampling sites. However, the larger the distance between the suspected source and the downstream site is, the more likely it becomes that there are removal process (e.g. sorption to particles followed by settling) and other potential PCB sources between the two sites. Consequently, the investigator has to accept that in many cases a complete mixing of waters at the downstream site (and thus a satisfied mass balance) is not possible. In the case of the WWTP, it cannot be totally excluded that biofouling did not have any influence on the uptake. The mass of biofouling (estimated from the weight before extraction) on the samplers from the discharge was twice the mass on samplers from the river. It could have slightly decreased the uptake and possibly also contributed to the fact that the mass balance was not satisfied. In the case of the settling basin, Cs,downstream was significantly higher than Cs,upstream only for PCB 52. The ratio Cs,downstream/ Cs,upstream is higher than the factor Fdownstream/upstream and the increase of Cs between the two sites is due to the presence of the settling basin. The very low volumetric flow rate of the water discharge was not officially measured and could only be estimated (0.001 m3 s1); the flow of the Venoge was about 1.87 m3 s1. To verify the mass balance (Eq. (3)) for PCB 52, Cs,upstream, Cs,downstream and Cs,discharges were normalized to the site in the water discharge which presented the lowest velocity. They were thus divided respectively by Fupstream/discharge, Fdownstream/discharge and 1. The prod-

uct ‘‘Qdischarge Cdischarge’’ is about 2.7 times lower than the term ‘‘Qdownstream Cdownstream  Qusptream Cupstream’’. The mass of discharged PCBs explained thus only 37% of the increase between the upstream and downstream sites. Similarly to the case of the WWTP, the mixing of waters from the river and the settling basin was likely not complete at the downstream site. In addition, the volumetric flow rate of the settling basin water discharge most probably increased during rain events and could have been underestimated since it was not continuously measured. The operational deployment in the Venoge River showed that the proposed approach can be used as an efficient tool to point out PCB sources. For such preliminary investigations, a limited number of measurements per sampling site suits the aim of providing rapid information. Two measurements enable to roughly estimate the variation of Cs at each sampling site while keeping laboratory working time sufficiently low for large-scale investigation. In order to assign the PCB release to a polluter, additional data should however be acquired. The variation of Cs at the different sampling sites and at different distances from the suspected source should be measured by deploying more samplers. In addition, to assess the magnitude of the pollution, aqueous concentration should be determined by using passive samplers loaded with PRCs and, when practicable, by collecting water samples from discharges.

5. Conclusion Using the field based channel system the influence of water velocity on PCB uptake by LDPE strips was determined under river-like flow conditions. Relationships between water velocity and Cs were determined for the six iPCBs and offer the investigator the possibility to determine the expected change of Cs due to velocity variations among sites. This change was shown to be low for velocity variations encountered in rivers and makes LDPE strips an efficient investigative tool. The use of PRCs can be avoided for such an application of passive sampling. Levels of dioxin-like PCBs in fishes from the Venoge were reported to exceed maximal levels permitted by the EU without knowing the sources of PCBs emission. According to the strategy described in this paper, LDPE strips were deployed downstream and upstream of potential PCB sources as well as in their water discharges. Within a few months, PCB sources could be pointed out and authorities informed of their presence. Based on the methodology and results from this study, similar investigations can be conducted in other rivers.

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