Passive sampling (PS) is a very useful approach for the monitoring of emerging contaminants in environmental matrices, showing greater sensitivity than can be achieved by current best practice – grab sampling – and is applicable to a wide variety of compounds. An EU Directive (2013/39/EC) has added substances to the existing Water Framework Directive (WFD) Priority Substance list. Investigation into PS in the monitoring of these compounds is necessary to show the potential of this technique in supporting monitoring requirements under the WFD. A catchment-based approach evaluated the occurrence of these compounds in Irish surface waters. This work deals with the challenges associated with the use of PS in a legislative context, and for routine monitoring of emerging contaminants. Looking at a number of sites across Ireland, upstream and downstream of wastewater treatment plants, the focus was on polar analytes and polar PS (POCIS). With method limits of detection (LODs) of 0.001 mg L−1 pharmaceuticals and endocrine disrupting chemicals (EDCs) were found in water and passive samples alike, whereas the polar pesticides were not often detected or were below the annual average environmental quality standard levels. The results of this study show the potential for PS as a monitoring technique for emerging and watch-list chemicals.

Passive sampling (PS) techniques are rapidly developing as very cost-effective state of the art tools to identify and measure ultra-trace micro-pollutants in water. It is possible to obtain improved compound detectability and sensitivity relative to those obtained using more ‘traditional’ spot water sampling and analysis techniques. On an operational front, several questions remain regarding the merits of PS in providing representative sampling of chemicals in environmental waters, the quantification of water column pollutant loadings and its suitability for adoption for checking legislative compliance.

PS is also a valuable tool owing to the ability to sample large volumes of water, the ease of deployment and processing compared with water or biota sampling and the non-mechanical operation meaning no external power input is required (Harman et al. 2012). PS have the ability to measure the freely dissolved fraction, or bioavailable fraction of a compound in water (Vrana et al. 2005; Smedes et al. 2010). This is a measure of the availability of the specific compound to organisms in the environment and directly applies to risk assessment as the freely dissolved fraction is proportional to the chemical activity of the compound (Harman et al. 2009; Booij et al. 2016). One important consideration is that passive samplers can provide more representative information than infrequent spot sampling on the occurrence of pollutants in water bodies, particularly where concentrations fluctuate markedly in time.

It is theoretically possible to use passive sampling to quantify and/or screen for the vast majority of organic compounds. However, a number of conditions must be met before PS techniques can be considered as fully suitable for the purposes of routine monitoring. Their application for Water Framework Directive (WFD) (2000/60/EC) (European Parliament 2000) purposes requires additional performance criteria. These include the calculation of accurate uptake rates in order to calculate time weighted average (TWA) contaminant concentrations in water, and strict protocols for in situ deployment. The authors have previously provided a full summary of a role for passive sampling in supporting WFD compliance and monitoring (Jones et al. 2015).

Passive sampling is a diffusion-based technique, with samplers consisting of a material with a high affinity for the analytes of interest. In some samplers (e.g. low density polyethylene strips and silicone rubber sheets), this receiving phase can be bare, while in other samplers (e.g. polar organic chemical integrative sampler (POCIS)), it is separated from the sampling medium (water) by a diffusion limiting membrane. Different combinations of membrane and receiving phase allow the sampling device to be tailored for various purposes (Booij & Smedes 2010; Smedes et al. 2010; Jones et al. 2015). As a high affinity for pollutants of interest is selected for the receiving phase, it achieves a higher concentration within the sampler compared with that in the water phase and can provide significant improvements in analytical detection limits. There are the two fundamentally different modes of uptake by receiving phases, absorption (in partition samplers such as polydimethylsiloxane (PDMS)), and adsorption when target compounds form strong bonds to the sampling material (e.g. POCIS) (Booij & Smedes 2010; Smedes et al. 2010; Jones et al. 2015). Uptake of pollutants by adsorption-based samplers is more complex, and the useful sampling period is limited by the capacity of the receiving phase for a particular analyte. In these samplers, factors such as competitive binding may be important (Booij & Smedes 2010; Smedes et al. 2010; Jones et al. 2015). Passive samplers such as the POCIS are often used as alternative sampling methods for EDCs as they can provide time-integrated trace levels of pollutants in water and have been suggested as complementary methods for WFD surveillance, operational and investigative monitoring (European Commission 2009). Sampling rates for these in situ devices which can consider all environmental factors are not yet viable, thus POCIS is generally applicable for use as a qualitative/screening device (Figure 1).

The primary aim of this study was to gather knowledge (national and international) on the current status of monitoring emerging chemicals in aquatic environments, and to assess the value of PS-based techniques for this purpose, particularly within the context of the WFD and Marine Strategy Framework Directive (MSFD) (European Parliament and Council of the European Union 2008). In Dublin and Cork catchments, studies were carried out over two years to assess based variations in chemical contamination. Analytical methods for pesticides, pharmaceutical and non-polar organics were adapted and developed for application to water, and passive sampler extracts.

The study includes oestrogens, oestrone (E1) and 17β oestradiol (E2), the synthetic oestrogen 17α-ethynylestradiol (EE2) and diclofenac. This is because an increasing volume of scientific data indicates potential negative effects of steroid oestrogens at European Union (EU) level, to or via the aquatic environment. There is also a threat posed to the environment by the occurrence of pharmaceuticals and personal care products (PPCPs) (Snyder et al. 2009; Wilkinson et al. 2017). The main source of PPCPs is through the wastewater management systems present in all major urbanised areas (Zhang et al. 2008; Cabeza et al. 2012; Bu et al. 2013). The use of wastewater treatment plants (WWTPs) serves as a means for the removal of organic matter and other substances prior to discharge into the environment, although not all the compounds contained within the effluent can be removed or degraded to a harmless product. The increased construction of WWTPs, under EU regulations, has partly removed the discharge of raw sewerage into the environment. The ingestion of pharmaceuticals and the metabolism (or lack of) and their subsequent excretion (generally via urine and faeces) serves as a major route to the environment. The objective of this part of the study was to screen a number of samples associated with WWTP runoff for pharmaceuticals and steroids, and to assess the value of passive sampling in monitoring these compounds.

These concerns have resulted in the inclusion of E1, E2 and EE2 in the EU watch list of substances for Union-wide monitoring (European Commission 2015), whereupon it is anticipated this will generate high-quality data on their concentrations in the aquatic environment for the purpose of supporting future prioritisation exercises. Under this legislation maximum method detection limits of 0.4 ng L−1 for E1 and E2 and 0.035 ng L−1 for EE2 on a total water basis have been set. It is generally recognised that there are a number of difficulties in reaching such low detection limits using traditional spot water sampling and marine strategy (MS) methods due to the low instrument sensitivities required, volumes of water required, increased dilution effects and representative sampling in what is generally a dynamic environment.

Pesticides and biocidal products comprise the majority of the emerging compounds (nine) most recently added to EU water legislation (2013/39/EU). These compounds are the focus of the second part of the study. This case study focussed on aclonifen, bifenox, dicofol, quinoxyfen, cybutryne, dichlorvos and terbutryn.

Aclonifen (2-chloro-6-nitro-3-phenoxyaniline) is a diphenyl-ether (DPE) herbicide used in the pre-emergence control of broad-leaved and grass weeds (Directive 2011a; Loos 2012a). Bifenox is a selective herbicide used in control of annual broad-leaved weeds and some grasses in numerous crops, e.g. in cereals, maize, soya beans, rice and other crops (Directive 2011b; Loos 2012a, 2012b). Cybutryne (Irgarol) is an effective triazine herbicidal biocide (or algicide) mainly used as an antifouling agent in paints for boats and vessels. It is applied at marine as well as at inland freshwater sites (Manzo et al. 2006; Loos 2012a, 2012b). Dicofol is an organochlorine pesticide (arcaricide; miticide) that is chemically related to dichlorodiphenyltrichloroethane (DDT), and used for controlling mites that damage cotton, fruit trees and vegetables (Hoekstra et al. 2006; Fujii et al. 2011; Loos 2012a, 2012b). Heptachlor is an insecticide which is banned and is no longer used in the EU. Heptachlor-epoxide is its degradation product (Directive 2011c; Loos 2012a, 2012b). Quinoxyfen is a fungicide often used to control powdery mildew infections on grapes and hops (Cabras et al. 2000; California Department of Pesticide Regulation 2004; Loos 2012a, 2012b). Until now, limited work has been published which included all of these compounds in relation to surface water occurrence. Analysis of these compounds is problematic, requiring a range of specific methods and strict protocols. This study involved collection of passive and grab samples in a catchment and monitoring these compounds for two periods (summers 2014 and 2015).

Site selection

A number of specific requirements had to be met by sites to be used in the study. As far as possible the choice of deployment locations took into account current WFD monitoring to maximise synergies (e.g. with spot sampling data). Practical considerations for site selection for PS deployment included access, security of deployments, options for use of structures for deployment, restrictions on deployment of moorings and tidal range in transitional and coastal waters. Locations of the sites selected, and types of sample collected at each site are outlined in Table 1.

Table 1

Overview of sampling and rationale from this study (Figure 2)

CountySiteRationalePOCISWater
Cork Inchigeelagh Upstream river ✓ ✓ 
Inniscarra Downstream river ✓ ✓ 
Shandon Riverine/transitional ✓ ✓ 
Lough Mahon Riverine/transitional ✓ ✓ 
Outer bay Riverine/transitional ✓ ✓ 
Dublin Poolbeg High pressure coast ✓ ✓ 
Osberstown Riverine/transitional ✓ ✓ 
Lucan Bridge Downstream river ✓ ✓ 
Kilcullen Bridge Upstream river ✓ ✓ 
CountySiteRationalePOCISWater
Cork Inchigeelagh Upstream river ✓ ✓ 
Inniscarra Downstream river ✓ ✓ 
Shandon Riverine/transitional ✓ ✓ 
Lough Mahon Riverine/transitional ✓ ✓ 
Outer bay Riverine/transitional ✓ ✓ 
Dublin Poolbeg High pressure coast ✓ ✓ 
Osberstown Riverine/transitional ✓ ✓ 
Lucan Bridge Downstream river ✓ ✓ 
Kilcullen Bridge Upstream river ✓ ✓ 

In order to evaluate the benefit of the catchment approach to identify point sources and pathways of pollution, the River Lee catchment in Cork was selected for the study of potential impact from the natural oestrogens, oestrone (E1) and 17β oestradiol (E2) and the synthetic oestrogen 17α ethynylestradiol (EE2). The river Lee flows from an area of low anthropogenic activity, through agricultural land to Cork City entering the Celtic Sea via the industrialised Cork Harbour (100 km2 surface). Within the study area, Lough Mahon and the River Lee in Co. Cork received a poor WFD classification status for fish. Evidence of pollutant stress in mussels has also been revealed by scope for growth measurements, potentially due to untreated wastewater from population centres such as Cork City. Secondary treated waste enters Cork Harbour from Cork City (333,000PE) and from a number of other potential inputs ranging including industrial waste and riverine inputs of agricultural run-off and untreated waste from the River Lee catchment.

In Dublin, the river Liffey is one of the most important rivers in Ireland and is a major water source for the Greater Dublin Area. The Liffey has been a controlled river for more than 50 years. It is dammed at Pollaphuca and Leixlip and there are also three hydro-electric power stations along its course. Due to this artificially controlled hydraulic regime, the Liffey is defined as a ‘Heavily Modified Water Body’ under the Water Framework Directive (Parliament 2000). The water quality Q-rating in the Liffey varies along its length between 5 (close to source) and 2–3 in the lower reaches (close to sea). The Eastern River Basin Management Plan categorises the Liffey as of ‘moderate’ quality throughout the majority of its course (Dublin City Council 2015). The river is designated in parts under the Urban Waste Water Treatment Regulations 2001, as a ‘sensitive’ river (downstream of the Osberstown sewage treatment works to Leixlip Reservoir, County Kildare) (European Parliament 2004). The WFD characterisation process has identified WWTP as primary pressures on the Liffey system.

Water samples

A 1 L grab sample (5 L for pharmaceuticals and EDCs) was collected in shatterproof glass bottles at each sampling point (n = 3). These samples were extracted using solid phase extraction (SPE), and analysis was carried out by liquid chromatography-mass spectrometry (LC-MS) or gas chromatography-mass spectrometry (GC-MS) depending on compounds analysed.

Water extraction for pesticides

The sample was passed through a styrene di-vinyl benzene polymer solid phase extraction (SPE). The cartridge was washed, dried and the steroids desorbed with dichloromethane. This extract was evaporated to incipient dryness and dissolved in a 90:10 v/v% iso-hexane:propan-2-ol mixture and cleaned-up using normal phase chromatography with an Amino LC column fractionation. The resulting extract was again evaporated to incipient dryness before being dissolved in 50% aqueous CH3OH. This extract was then fractioned using reverse phase chromatography with a standard C18 phase column. The resulting extract was evaporated to incipient dryness and dissolved in 10% aqueous CH3OH. 100 μL of this was analysed using high performance liquid chromatography with negative ion atmospheric photo-ionisation interface and time of flight mass spectrometric detection.

Water extraction for pharmaceuticals and EDCs

Water samples (5 L) were extracted as per Ronan & McHugh 2013). Briefly, samples were filtered with 0.45 μm glass fibre filters prior to SPE. 5 L water samples were filtered with 0.45 μm glass fibre filters (Whatman GF-F, General Electric company, USA). Sample clean-up was conducted using Oasis hydrophilic-lipophilic-balanced (HLB) 3 mL, 60 mg SPE cartridges (Waters, Milford, MA, USA). Bakerbond SPE™ 75 mL reservoirs (JT Baker, Avantor Performance Materials, Centre Valley, PA, USA) were attached to the SPE cartridges to enable application of the sample. Cartridges were rinsed and conditioned with 3 mL pestiscan grade CH3OH and deionised water, prior to addition of the sample extract. Cartridges were rinsed with 3 mL 5% CH3OH in deionised H2O, and samples were eluted in 2 mL CH3OH under a gentle vacuum. The sample was then reduced to dryness under nitrogen at 35 °C and re-suspended in 130 μL of 1:4 CH3OH: water. 10 ng of external standard warfarin (Sigma Aldrich, St Louis, MO, USA) was added. 20 μL of this solution was injected into the high performance liquid chromatography (HPLC) system.

Passive samplers

POCIS devices used in this study were as developed by Alvarez et al. 2004 (Alvarez et al. 2004) and were supplied by the National Laboratory Service (NLS), Environment Agency, United Kingdom. They consisted of a layer of Oasis HLB sorbent (230 mg approximately, Waters, Milford, MA, USA) sandwiched between two membrane layers of polyethersulphone polymer, fixed in place by two circular steel rings secured with stainless steel nuts and bolts. Each POCIS disc was mounted on holders (three per sampler) and placed inside a protective perforated stainless-steel housing. A POCIS device (three POCIS discs) and one single POCIS disc for use as a field blank were stored in an airtight metal canister at −30 °C and transported to the site in cold conditions. POCIS devices were deployed at 1 m depth at selected sites for a period of one month. The field blank was exposed to the environment during deployment and retrieval. Storage and transportation of the devices was conducted in a manner that would not expose them to contamination. Samplers and field blanks were stored at −30 °C prior to transport (in cold conditions) to the NLS, UK for extraction and analysis.

Passive sampler extracts were analysed as per the Environment Agency (UK) Blue Book 220, method B. On arrival to the laboratory the POCIS discs were cleaned to remove any debris or stubborn deposits. Any excess water was removed where necessary with absorbent tissue and the discs allowed to air-dry overnight. The loaded POCIS was carefully dismantled and the sorbent transferred with CH3OH into a glass chromatography column fitted with a glass wool plug and a stopcock. The CH3OH used to rinse the sorbent into the column was collected to be pooled with the SPE eluent. The sample was eluted with 40 mL CH3OH (adjusted to near dropwise flow). The collected eluate and rinse was evaporated using rotary evaporation to approximately 1 mL.

A schematic illustrating this approach is shown in Figure 3.

Figure 1

Schematic illustrating components of the POCIS sampler, the mechanism of analyte interactions and the deployment apparatus used for multiple samplers. The samplers are used for polar analytes, the Kow are shown in the table.

Figure 1

Schematic illustrating components of the POCIS sampler, the mechanism of analyte interactions and the deployment apparatus used for multiple samplers. The samplers are used for polar analytes, the Kow are shown in the table.

Close modal
Figure 2

Map of Ireland with two inserts highlighting Dublin catchment (top), and Cork catchment (bottom). Sampling sites are marked in stars. WWTPs are marked with circles.

Figure 2

Map of Ireland with two inserts highlighting Dublin catchment (top), and Cork catchment (bottom). Sampling sites are marked in stars. WWTPs are marked with circles.

Close modal
Figure 3

Basic schematic showing the steps involved in the monitoring approach using passive sampling.

Figure 3

Basic schematic showing the steps involved in the monitoring approach using passive sampling.

Close modal

Analysis for pharmaceuticals and EDCs

The HPLC system consisted of an Agilent 1200 Series Game Pad, Agilent 1200 Binary Pumps, an Agilent 1200 High Performance Autosampler and an Agilent 1200 thermostatted column compartment (Agilent Technologies Deutschland GmbH, Böblingen, Germany). A Kinetex 4.6 × 50 mm, 2.1 mm ID C18 2.5 μm particle size column (Phenomenex, Macclesfield, Cheshire, UK) was used for chromatographic separation. For LC-MS/MS experiments, an Applied Biosystems 3200 Q-TRAP was used (Applied Biosystems, Foster City, CA, USA). The mobile phases were 0.025% TEA (triethylamine) in deionised water (A) and 0.025% TEA in 5:95 water: acetonitrile (B) flowing at 300 μL per minute with a gradient as follows: 0 to 2 min (5 to 31.5% B), 2 to 7 min (31.5 to 34% B), 7 to 9 min (100% B) and 9 to 12 min (5% B). The ESI source (Turbo-Ionspray, Applied Biosystems) was operated in negative mode at 550 °C and −4,400 V. Quantification was performed using matrix-matched standards for mussel tissue, and aqueous standards for water samples.

Analysis for pesticides

The HPLC system consisted of an Agilent 1100 Series Game Pad, Agilent 1100 Binary Pumps, an Agilent 1100 High Performance Autosampler and an Agilent 1100 thermostatted column compartment (Agilent Technologies Deutschland GmbH, Böblingen, Germany). A Luna phenyl hexyl 2 × 150 mm, 3 μm particle size column (Phenomenex, Macclesfield, Cheshire, UK) fitted with a Luna phenyl propyl 4 × 2 mm guard column was thermostatted at 60 °C was used for chromatographic separation. For HPLC-MS/MS experiments, an Applied Biosystems 3200 Q-TRAP was used (Applied Biosystems, Foster City, CA, USA). The mobile phases were deionised water (A) and 0.025% TEA in 95:5 CH3OH: acetone (B) flowing at 300 μL per minute with a gradient as follows: 0 to 0.5 min (5 to 20% B), 0.5 to 1 min (20 to 40% B), 1 to 12 min (40 to 80% B), 12 to 14 min (80% B) and 14 to 14.5 min (80 to 5% B) with analysis as per the Environment Agency Blue Book 220.

This is the first time that pharmaceuticals are being proposed for addition to existing European water legislation and very low environmental quality standards (EQS) values (and often analytically challenging) have been set for many of the proposed priority substances, EDCs and pharmaceuticals. Diclofenac is a non-steroidal anti-inflammatory drug. EE2 is a synthetic estradiol used in contraceptive pills and for the treatment of menopausal and post-menopausal symptoms. E2 is the predominant natural female sex hormone and is the most active of the naturally occurring estrogenic hormones and is a key intermediate in industrial synthesis of other estrogens and of various hormonal 19-norsteroids.

From the results presented in Table 2, it can be seen that pharmaceutical compounds were detected in both water and passive samplers collected in 2014 and 2015. As accurate sampling rates for POCIS were not available, levels are presented as ng/device for the particle size distributions (PSDs) for diclofenac and as estimated concentrations (ng L−1) for the EDCs, with the relative percentage loading of each compound in relation to others being sampled in parenthesis. Concentrations and determinations as measured in both the passive sampling devices and grab samples are indicative of the potential influence of the presence of large WWTPs along the Dublin catchment. Samples collected at the Lucan sampling point (downstream from the Leixlip WWTP) show the highest levels of compounds of interest in the water phase. This shows that E1 was consistently the compound with the highest loading in the PSDs compared with E2 and EE2. This is consistent when compared with the concentrations detected in water samples, where E2 was often detected when E2 and EE2 were not. The 5 L (n = 3) grab samples collected at each sampling point allowed the achievement of a low limit of quantitation (LOQ) (method LOQ of 0.001 mg L−1).

Table 2

Concentrations and relative percentage occurrence (brackets) of pharmaceuticals and EDCs along the liffey catchment in water and passive samples collected in 2014 and 2015

MatrixKilcullenOsberstownLucanPoolbeg
Analyte  2014    
Dcl* POCIS 0.66 0.93 17.48 3.32 
E1 <0.23 (65.7) 0.29 (70.7) 0.37 (75.5) 0.42 (82.4) 
E2 <0.06 (17.1) <0.06 (14.6) <0.06 (12.2) 0.05 (9.8) 
EE2 <0.06 (17.1) <0.06 (14.6) <0.06 (12.2) <0.04 (7.8) 
Dcl Water 4.05 4.17 29.26 3.92 
E1 <0.07 (100) nd 0.33 (43.4) 1.92 (89.3) 
E2 nd 0.33 (100) 0.43 (56.6) 0.23 (10.7) 
EE2 nd nd nd nd 
Analyte  2015    
Dcl* POCIS 0.93 6.2 5.4 1.7 
E1 <0.23 (65.7) 0.31 (72.1) 0.42 (77.8) 0.41 (75.9) 
E2 <0.06 (17.1) <0.06 (14.0) 0.06 (11.1) 0.07 (13.0) 
EE2 <0.06 (17.1) <0.06 (14.0) <0.06 (11.1) <0.06 (11.1) 
Dcl Water 5.25 3.03 26.28 4.1 
E1 <0.23 (65.7) 0.31 (72.1) 0.42 (77.8) 0.41 (75.9) 
E2 <0.06 (17.1) <0.06 (14.0) 0.06 (11.1) 0.07 (13.0) 
EE2 <0.06 (17.1) <0.06 (14.0) <0.06 (11.1) <0.06 (11.1) 
MatrixKilcullenOsberstownLucanPoolbeg
Analyte  2014    
Dcl* POCIS 0.66 0.93 17.48 3.32 
E1 <0.23 (65.7) 0.29 (70.7) 0.37 (75.5) 0.42 (82.4) 
E2 <0.06 (17.1) <0.06 (14.6) <0.06 (12.2) 0.05 (9.8) 
EE2 <0.06 (17.1) <0.06 (14.6) <0.06 (12.2) <0.04 (7.8) 
Dcl Water 4.05 4.17 29.26 3.92 
E1 <0.07 (100) nd 0.33 (43.4) 1.92 (89.3) 
E2 nd 0.33 (100) 0.43 (56.6) 0.23 (10.7) 
EE2 nd nd nd nd 
Analyte  2015    
Dcl* POCIS 0.93 6.2 5.4 1.7 
E1 <0.23 (65.7) 0.31 (72.1) 0.42 (77.8) 0.41 (75.9) 
E2 <0.06 (17.1) <0.06 (14.0) 0.06 (11.1) 0.07 (13.0) 
EE2 <0.06 (17.1) <0.06 (14.0) <0.06 (11.1) <0.06 (11.1) 
Dcl Water 5.25 3.03 26.28 4.1 
E1 <0.23 (65.7) 0.31 (72.1) 0.42 (77.8) 0.41 (75.9) 
E2 <0.06 (17.1) <0.06 (14.0) 0.06 (11.1) 0.07 (13.0) 
EE2 <0.06 (17.1) <0.06 (14.0) <0.06 (11.1) <0.06 (11.1) 

Concentrations of diclofenac (Dcl) in POCIS presented as amount (ng/device) while E1, E2 and EE2 in POCIS is estimated based on sampling rates. Other results in ng L−1. nd = not detected.

Steroid estrogen data from spot water sample and POCIS data available for samples collected between 2013 and 2014 from the River Lee catchment are presented in Table 3. Sites were assessed on a catchment basis and include freshwater and estuarine/marine samples. Five sampling points in the Lee catchment in County Cork were sampled once each in 2013 and 2014. E2 and EE2 were not detected in water samples in 2013 and 2014. E1 was added to the suite of compounds in 2014 and was detected in Lough Mahon and the Outer Harbour at similar concentrations as those calculated in POCIS. E1 was also detected in water at Iniscarra and in POCIS in Iniscarra and Shandon. Concentrations of E2 in POCIS were higher in 2013, with a surprising spike of EE2 detected in Iniscarra, a remote upstream site (1.39 ng L−1).

Table 3

Concentrations of pharmaceuticals and EDCs along the Lee catchment in water and passive samples collected in 2013 and 2014

MatrixInchigeelaghIniscarraShandonLough MahonOuter Harbour
Analyte  2013     
Dcl* POCIS na na na na na 
E2 <0.5 <0.5 <0.5 2.36 1.98 
EE2 <0.2 1.39 <0.2 <0.2 <0.2 
Dcl Water 7.2 3.5 4.5 7.8 2.8 
E2 nd nd nd nd nd 
EE2 nd nd nd nd nd 
Analyte  2014     
Dcl* POCIS <0.1 0.14 0.92 0.75 0.27 
E1 <0.51 0.24 0.37 0.48 0.37 
E2 <0.13 <0.04 <0.04 0.06 0.09 
EE2 <0.12 <0.04 <0.04 <0.04 0.07 
Dcl Water 9.7 12.7 6.2 8.7 4.3 
E1 nd 0.41 nd 0.41 0.54 
E2 nd nd nd nd nd 
EE2 nd nd nd nd nd 
MatrixInchigeelaghIniscarraShandonLough MahonOuter Harbour
Analyte  2013     
Dcl* POCIS na na na na na 
E2 <0.5 <0.5 <0.5 2.36 1.98 
EE2 <0.2 1.39 <0.2 <0.2 <0.2 
Dcl Water 7.2 3.5 4.5 7.8 2.8 
E2 nd nd nd nd nd 
EE2 nd nd nd nd nd 
Analyte  2014     
Dcl* POCIS <0.1 0.14 0.92 0.75 0.27 
E1 <0.51 0.24 0.37 0.48 0.37 
E2 <0.13 <0.04 <0.04 0.06 0.09 
EE2 <0.12 <0.04 <0.04 <0.04 0.07 
Dcl Water 9.7 12.7 6.2 8.7 4.3 
E1 nd 0.41 nd 0.41 0.54 
E2 nd nd nd nd nd 
EE2 nd nd nd nd nd 

Results in ng L−1. nd = not detected, na = not analysed. Sites are listed from upstream to downstream (left to right).

From the results of this work, the currently unreliable nature of polar passive sampling was explored, as grab samples often detected estrogenic events that were missed by the passive samplers. A 2016 interlaboratory study (ILS) found that for most polar compounds both the analytical variability and the variability of applied calibration data contributed similarly to the overall variability of water concentration estimates. The same study also found the exchange of polar compounds between sampler and the aqueous phase was often observed to be anisotropic and therefore generally not possible to use release of PRCs to calibrate the uptake rate for calculation of TWA water concentrations (Vrana et al. 2016). This result could be combatted through the deployment of more PSDs, multiple PSDs at each site and flow data collected to allow for more accurate sampling rates to be calculated. The current study does not demonstrate any advantages of polar passive sampling for pharmaceuticals and EDCs outside of screening potential. This is reflected in a recent position paper arising from a NORMAN network-supported workshop in Lyon (France) in November 2014 (Miège et al. 2015).

Results of this work are presented in Tables 25, showing data for 2014 and 2015 in Dublin and for 2013 and 2014 in Cork catchments. Targeted pesticides were not detected in many cases and were otherwise present in concentrations below the annual average environmental quality standard (AA EQS). The analytes were detected using passive samplers in 2014 in Dublin at Kilcullen, but these were absent in 2015. The results from the passive samplers allowed for the assessment of pharmaceutical concentrations where large enough grab samples were not feasible. The benefits of this technique are counted as the ease of use and deployment, the selectivity and the ability to sample large volumes of water.

Table 4

Concentrations of pesticides in passive samplers (ng/device) and in water (ng mL-1) in the Dublin catchment sites.

MatrixAA EQSKilcullenOsberstownLucanPoolbeg
Analyte   2014 
Aclonifen POCIS (ng/device)  nd nd nd nd 
Bifenox  nd nd nd nd 
Cybutryn  nd 0.017 nd nd 
Dichlorvos  nd nd nd nd 
Dicofol  nd nd nd nd 
Heptachlor/heptachlor epoxide  nd nd nd nd 
Quinoxyfen  nd nd nd nd 
Terbutryn  0.007 0.023 nd Nd 
Aclonifen Water (ng mL−1120 0.001 0.0009 0.001 nd 
Bifenox 12 0.001 nd 0.001 nd 
Cybutryn 2.5 0.0004 0.0006 0.0005 0.0003 
Dichlorvos 0.6 nd nd nd nd 
Dicofol 1.3 nd nd nd nd 
Heptachlor/heptachlor-epoxide 0.002 nd nd nd nd 
Quinoxyfen 150 0.001 0.0007 0.001 0.0003 
Terbutryn 65 0.0004 0.0006 0.0006 0.0004 
Analyte   2015 
Aclonifen POCIS (ng/device)  nd nd nd nd 
Bifenox  nd nd nd nd 
Cybutryn  nd 0.006 nd 0.007 
Dichlorvos  nd nd nd nd 
Dicofol  nd nd nd nd 
Heptachlor/heptachlor epoxide  nd nd nd nd 
Quinoxyfen  nd nd nd nd 
Terbutryn  0.007 0.031 0.054 0.055 
Aclonifen Water (ng mL−1120 0.002 nd nd nd 
Bifenox 12 0.004 nd nd 0.003 
Cybutryn 2.5 0.0007 nd nd 0.0002 
Dichlorvos 0.6 0.0004 0.0005 0.0003 0.0002 
Dicofol 1.3 nd nd nd nd 
Heptachlor/heptachlor epoxide 0.002 nd nd nd nd 
Quinoxyfen 150 0.0065 0.0003 nd nd 
Terbutryn 65 0.0002 nd 0.0009 0.0005 
MatrixAA EQSKilcullenOsberstownLucanPoolbeg
Analyte   2014 
Aclonifen POCIS (ng/device)  nd nd nd nd 
Bifenox  nd nd nd nd 
Cybutryn  nd 0.017 nd nd 
Dichlorvos  nd nd nd nd 
Dicofol  nd nd nd nd 
Heptachlor/heptachlor epoxide  nd nd nd nd 
Quinoxyfen  nd nd nd nd 
Terbutryn  0.007 0.023 nd Nd 
Aclonifen Water (ng mL−1120 0.001 0.0009 0.001 nd 
Bifenox 12 0.001 nd 0.001 nd 
Cybutryn 2.5 0.0004 0.0006 0.0005 0.0003 
Dichlorvos 0.6 nd nd nd nd 
Dicofol 1.3 nd nd nd nd 
Heptachlor/heptachlor-epoxide 0.002 nd nd nd nd 
Quinoxyfen 150 0.001 0.0007 0.001 0.0003 
Terbutryn 65 0.0004 0.0006 0.0006 0.0004 
Analyte   2015 
Aclonifen POCIS (ng/device)  nd nd nd nd 
Bifenox  nd nd nd nd 
Cybutryn  nd 0.006 nd 0.007 
Dichlorvos  nd nd nd nd 
Dicofol  nd nd nd nd 
Heptachlor/heptachlor epoxide  nd nd nd nd 
Quinoxyfen  nd nd nd nd 
Terbutryn  0.007 0.031 0.054 0.055 
Aclonifen Water (ng mL−1120 0.002 nd nd nd 
Bifenox 12 0.004 nd nd 0.003 
Cybutryn 2.5 0.0007 nd nd 0.0002 
Dichlorvos 0.6 0.0004 0.0005 0.0003 0.0002 
Dicofol 1.3 nd nd nd nd 
Heptachlor/heptachlor epoxide 0.002 nd nd nd nd 
Quinoxyfen 150 0.0065 0.0003 nd nd 
Terbutryn 65 0.0002 nd 0.0009 0.0005 

nd = not detected. Detection limits for aclonifen, bifenox, cybutryn, dichlorvos, dicofol, heptachlor/heptachlor epoxide, quinoxyfen, terbutryn = 0.5–5.0 ng mL-1. AA EQS – annual average environmental quality standard.

Table 5

Concentrations of pesticides in passive samplers (ng/device) and in water (ng L−1) in the Cork catchment sites

MatrixInchigeelaghIniscarraShandonOuter HarbourLough Mahon
Analyte   2013 
Aclonifen Water (ng L−1120 nd nd nd nd nd 
Bifenox 12 nd nd nd nd nd 
Cybutryn 2.5 nd nd nd nd nd 
Dichlorvos 0.6 nd 0.003 0.0004 0.002 0.003 
Dicofol 1.3 nd nd nd 0.0002 0.0001 
Heptachlor/heptachlor epoxide 0.002 nd nd nd nd nd 
Quinoxyfen 150 1.00 1.50 6.00 0.27 0.87 
Terbutryn 65 nd 0.0006 nd nd nd 
Analyte   2014 
Aclonifen POCIS (ng/device)  nd nd nd nd nd 
Bifenox  nd nd nd nd nd 
Cybutryn  nd nd nd nd nd 
Dichlorvos  nd nd nd nd nd 
Dicofol  nd nd nd nd nd 
Heptachlor/heptachlor epoxide  nd nd nd nd nd 
Quinoxyfen  1.00 1.00 0.61 0.93 0.67 
Terbutryn  nd nd nd nd nd 
Aclonifen Water (ng L−1120 nd nd nd nd nd 
Bifenox 12 nd nd nd nd nd 
Cybutryn 2.5 nd 0.0055 nd nd nd 
Dichlorvos 0.6 0.0013 0.0007 0.0019 0.0005 0.0007 
Dicofol 1.3 nd nd nd nd nd 
Heptachlor/heptachlor epoxide 0.002 nd nd nd nd nd 
Quinoxyfen 150 1.00 0.30 5.73 8.00 6.40 
Terbutryn 65 0.0001 0.0001 0.0002 0.0001 0.0002 
MatrixInchigeelaghIniscarraShandonOuter HarbourLough Mahon
Analyte   2013 
Aclonifen Water (ng L−1120 nd nd nd nd nd 
Bifenox 12 nd nd nd nd nd 
Cybutryn 2.5 nd nd nd nd nd 
Dichlorvos 0.6 nd 0.003 0.0004 0.002 0.003 
Dicofol 1.3 nd nd nd 0.0002 0.0001 
Heptachlor/heptachlor epoxide 0.002 nd nd nd nd nd 
Quinoxyfen 150 1.00 1.50 6.00 0.27 0.87 
Terbutryn 65 nd 0.0006 nd nd nd 
Analyte   2014 
Aclonifen POCIS (ng/device)  nd nd nd nd nd 
Bifenox  nd nd nd nd nd 
Cybutryn  nd nd nd nd nd 
Dichlorvos  nd nd nd nd nd 
Dicofol  nd nd nd nd nd 
Heptachlor/heptachlor epoxide  nd nd nd nd nd 
Quinoxyfen  1.00 1.00 0.61 0.93 0.67 
Terbutryn  nd nd nd nd nd 
Aclonifen Water (ng L−1120 nd nd nd nd nd 
Bifenox 12 nd nd nd nd nd 
Cybutryn 2.5 nd 0.0055 nd nd nd 
Dichlorvos 0.6 0.0013 0.0007 0.0019 0.0005 0.0007 
Dicofol 1.3 nd nd nd nd nd 
Heptachlor/heptachlor epoxide 0.002 nd nd nd nd nd 
Quinoxyfen 150 1.00 0.30 5.73 8.00 6.40 
Terbutryn 65 0.0001 0.0001 0.0002 0.0001 0.0002 

nd = not detected. Detection limits for aclonifen, bifenox, cybutryn, dichlorvos, dicofol, heptachlor/heptachlor epoxide, quinoxyfen, terbutryn = 0.5–5.0 ng mL−1.

From the results of this work the currently unreliable nature of polar passive sampling was explored as grab samples often detected estrogenic events that were missed by the passive samplers. A 2016 ILS found that for most polar compounds both the analytical variability and the variability of applied calibration data contributed similarly to the overall variability of water concentration estimates. The same study also found the exchange of polar compounds between sampler and the aqueous phase was often observed to be anisotropic and therefore generally not possible to use release of PRCs to calibrate the uptake rate for calculation of TWA water concentrations (Vrana et al. 2016). This result could be combatted through the deployment of more PSDs, multiple PSDs at each site and flow data collected to allow for more accurate sampling rates to be calculated. The current study does not demonstrate any advantages of polar passive sampling for pharmaceuticals and EDCs outside of screening potential. This is reflected in a recent position paper arising from a NORMAN network-supported workshop in Lyon (France) in November 2014 (Miège et al. 2015).

As these compounds have only recently been included in the proposed legislation, there is little information on widespread occurrence and background levels across Europe. However, some studies have reported the determination of these compounds in relation to WWTP inputs. Diclofenac is a widely used non-steroidal anti-inflammatory and antipyretic drug mainly administered via the oral route. It is rapidly absorbed in the body. Un-metabolised or partially excreted diclofenac is known to be mainly resistant to biological wastewater treatment and reaches the water cycle (Xiao et al. 2015). In 1998, diclofenac was detected in rivers and lakes in Switzerland with the data strongly suggesting input from human medical use via WWTPs. Levels detected were in the high ng L−1 level, much higher than the results from the River Liffey (Buser et al. 1998). A study carried out in Berlin in 2000 found levels of diclofenac in concentrations from n.d – 1,030 ng L−1 in surface water samples downstream from WWTP and sewage treatment plants (STPs) (Heberer et al. 2002). Looking at pharmaceutical residues in STP effluents from France, Greece, Italy, Sweden, Germany, the UK and Canada (n = 65), a review in 2005 reported concentrations of diclofenac from n.d. to 0.81 μg L−1 (Petrović et al. 2005). A further study in France reported the following levels of diclofenac in various matrices: 0.9 ng L−1 (tap water), 0.7 ng L−1 (surface water), 2.6 ng L−1 (marine water) and 9 ng/500 mL (wastewater). This is comparable with the results from this study. They also collected samples in the Hérault watershed and detected diclofenac in levels of 1.36–33.2 ng L−1, higher than levels found in the Liffey (Togola & Budzinski 2008). In 2008, a review reported that diclofenac is ineffectively removed by WWTPs, with removal efficiency varying from 0% to 80% depending on the operation conditions of the WWTPs (Zhang et al. 2008). A study from 2014 found levels of diclofenac in WWTPs in Greece (n = 32) to range from n.d. to 5,164 ng L−1 (Kosma et al. 2014). Another recent study looking at seawater concentrations of NSAIDs in Portugal found diclofenac to occur from n.d. to 30 ng L−1 (n = 5) (Paíga et al. 2015). A study reported diclofenac levels of 46–1,171 ng L−1 (WWTP effluent; n = 4) and <LOD (river water; n = 1). The same study also reported POCIS TWA levels of diclofenac to be 0.39 ng L−1 in river water and 0.25 ng L−1 in tap water (Tanwar et al. 2015).

Discharge from STP or WWTP, as well as direct introduction via excretion by livestock, represent major sources of estrogenic compounds to surface water. This is particularly relevant as the upstream sampling sites in this study are located in agricultural areas. Information on steroid hormones in the environment is comparatively less available than diclofenac. Occurrence and concentration of hormones in effluent, freshwater and groundwater varies in different countries. A recent review reported on the occurrence and concentration of hormones found in effluent, freshwater and groundwater in different countries. This report showed that in the effluent of WWTP and STP, the concentrations of estrogenic compounds usually are below 50 ng L−1, but there are unexpected high concentrations of estriol and 17α-estradiol about 590 ng L−1 and 180 ng L−1, respectively, found in America (Li et al. 2014). A review in 2011 summarised that the concentrations of hormones in influent and effluent are generally quite similar, indicating that the elimination of hormones is incomplete or that the compounds are resistant to typical wastewater treatment processes (Pereira et al. 2011). A study in Turkey found the following levels of hormones in surface waters (n = 6): E1 (n.d.–6.04 ng L−1) with a percentage occurrence of 0–40%; E2 (n.d.–10.2 ng L−1) with a percentage occurrence of 20–80%; EE2 (n.d.–14 ng L−1) with a percentage occurrence of 0–40% (Aydin & Talinli 2013).

This is the first report of a catchment-based assessment of pharmaceuticals in Ireland. As a result of their growing use, pharmaceuticals have been found in aquatic systems, in STP effluents, surface waters and even drinking waters. Although the quantities of pharmaceuticals and their bioactive metabolites being introduced into the environment is likely low, their continuous input may lead to a high long-term concentration and promote continual, but unnoticed effects on aquatic and terrestrial organisms. As such, it is important to screen and monitor these compounds in our waterways and to limit any possible sources of point pollution such as WWTPs and STPs. Monitoring these compounds is necessary to provide wider knowledge about their occurrence in the environment, to understand their fate, and organism exposure levels.

Oestrone (E1) and 17β oestradiol (E2) have been detected in Irish waters in both spot water samples and passive samplers. 17α ethynylestradiol (EE2) was detected in one passive sampler from Cork in 2013. While EE2 was not detected in spot water samples, the method LOD (limit of detection) does not meet the specified LOD for watch list; hence further development of such methods may potentially result in the detection of EE2 in some water bodies. Results from this study indicates that, for the coastal and freshwater sites studied, the presence of steroid oestrogens is low. Higher E1 and E2 levels were found in samples originating from higher population centres, e.g. Dublin and Cork, particularly at sites in close proximity to WWTP effluent discharges.

Significant progress has been made in recent years (and throughout the lifetime of this project) in respect of the development of mechanistic sampler-water exchange models and improved in situ calibration. The results presented in this study are comparable to other European studies. One such example is a large-scale inter-laboratory comparison carried out in 2010 across 24 laboratories (Miege et al. 2012). This study included a number of polar pesticides from the WFD and compared different types of polar passive sampling, including POCIS. Similar μg L−1 levels were detected and, as with this study, the passive sampling devices were found to allow determination of analytes that were otherwise rendered undetectable in the grab water samples. Similar recommendations were also made.

Overall, it is recognised that passive sampling undoubtedly provides a means by which low concentrations of hydrophobic pollutants (down to pg L−1 levels) can be measured with at least the same level of accuracy as conventional and ‘accepted’ spot sampling methods. There is a key additional benefit that concentration information is subject to less biological influences such as size and metabolism; however, it is also clear that passive sampling still faces considerable challenges in order for its applicability to be demonstrated beyond doubt.

There is a role for passive sampling in both screening and trend monitoring and in feeding into risk-based approaches to operational monitoring with future derived passive sampling threshold values installed in an assessment hierarchy/framework where exceedance of the threshold value could be used to flag potential contamination issues.

Alvarez
D. A.
,
Petty
J. D.
,
Huckins
J. N.
,
Jones-Lepp
T. L.
,
Getting
D. T.
,
Goddard
J. P.
&
Manahan
S. E.
2004
Development of a passive, in situ, integrative sampler for hydrophilic organic contaminants in aquatic environments
.
Environmental Toxicology and Chemistry
23
(
7
),
1640
1648
.
Booij
K.
,
Robinson
C. D.
,
Burgess
R. M.
,
Mayer
P.
,
Roberts
C. A.
,
Ahrens
L.
,
Allan
I. J.
,
Brant
J.
,
Jones
L.
,
Kraus
U. R.
,
Larsen
M. M.
,
Lepom
P.
,
Petersen
J.
,
Pröfrock
D.
,
Roose
P.
,
Schäfer
S.
,
Smedes
F.
,
Tixier
C.
,
Vorkamp
K.
&
Whitehouse
P.
2016
Passive sampling in regulatory chemical monitoring of nonpolar organic compounds in the aquatic environment
.
Environmental Science & Technology
50
(
1
),
3
17
.
Bu
Q.
,
Wang
B.
,
Huang
J.
,
Deng
S.
&
Yu
G.
2013
Pharmaceuticals and personal care products in the aquatic environment in China: a review
.
Journal of Hazardous Materials
262
,
189
211
.
Buser
H.-R.
,
Poiger
T.
&
Müller
M. D.
1998
Occurrence and fate of the pharmaceutical drug diclofenac in surface waters: rapid photodegradation in a lake
.
Environmental Science & Technology
32
(
22
),
3449
3456
.
Cabras
P.
,
Angioni
A.
,
Garau
V. L.
,
Pirisi
F. M.
,
Cabitza
F.
,
Pala
M.
&
Farris
G. A.
2000
Fate of quinoxyfen residues in grapes, wine, and their processing products
.
Journal of Agricultural and Food Chemistry
48
(
12
),
6128
6131
.
California Department of Pesticide Regulation
2004
PUBLIC REPORT 2004-01 Quinoxyfen, California Department of Pesticide Regulation, Sacramento, CA, USA.
Directive, S.-G. on R. of the P. S. L. (under W. G. E. of the C. I. S. for the W. F.
2011a
Aclonifen EQS Dossier, EU, European Commission, Brussels, Belgium.
Directive, S.-G. on R. of the P. S. L. (under W. G. E. of the C. I. S. for the W. F.
2011b
Bifenox EQS Dossier, EU, European Commission, Brussels, Belgium.
Directive, S.-G. on R. of the P. S. L. (under W. G. E. of the C. I. S. for the W. F.
2011c
Heptachlor/Heptachlor epoxide EQS Dossier, EU, European Commission, Brussels, Belgium.
Dublin City Council
2015
Eastern River Basin District. 2015 (February). [online] http://www.erbd.ie (accessed 2016).
European Commission
2009
Common Implementation Strategy for the WFD, Guidance Document No. 19, Guidance on Surface Water Chemical Monitoring
.
Brussels, Belgium
.
European Commission
2015
Commission Implementing Regulation (EU) 2015/495 of 20 March 2015 establishing a watch list of substances for Union-wide monitoring in the field of water policy pursuant to Directive 2008/105/EC of the European Parliament and of the Council. Official Journal of the European Union, L78/40(C(2015) 1756), 20–30.
European Parliament
2000
Directive 2000/60/EC of the European Parliament and of the Council of 23 October 2000 Establishing A Framework for Community Action in the Field of Water Policy
.
European Union, European Parliament, Brussels, Belgium
.
European Parliament
2004
S. I. No. 254/2001 — Urban Waste Water Treatment Regulations, Brussels, Belgium.
European Parliament and Council of the European Union
2008
Directive 2008/56/EC of the European Parliament and of the Council
.
Official Journal of the European Union
164
,
19
40
.
Fujii
Y.
,
Haraguchi
K.
,
Harada
K. H.
,
Hitomi
T.
,
Inoue
K.
,
Itoh
Y.
,
Watanabe
T.
,
Takenaka
K.
,
Uehara
S.
,
Yang
H.-R.
,
Kim
M.-Y.
,
Moon
C.-S.
,
Kim
H.-S.
,
Wang
P.
,
Liu
A.
,
Hung
N. N.
&
Koizumi
A.
2011
Detection of dicofol and related pesticides in human breast milk from China, Korea and Japan
.
Chemosphere
82
(
1
),
25
31
.
Harman
C.
,
Allan
I. J.
&
Vermeirssen
E. L. M.
2012
Calibration and use of the polar organic chemical integrative sampler – a critical review
.
Environmental Toxicology and Chemistry/SETAC
31
(
12
),
2724
2738
.
Hoekstra
P. F.
,
Burnison
B. K.
,
Garrison
A. W.
,
Neheli
T.
&
Muir
D. C. G.
2006
Estrogenic activity of dicofol with the human estrogen receptor: isomer- and enantiomer-specific implications
.
Chemosphere
64
(
1
),
174
177
.
Kosma
C. I.
,
Lambropoulou
D. A.
&
Albanis
T. A.
2014
Investigation of PPCPs in wastewater treatment plants in Greece: occurrence, removal and environmental risk assessment
.
Science of the Total Environment
466–467
,
421
438
.
Loos
R.
2012a
Analytical Methods for the new Proposed Priority Substances of the European Water Framework Directive (WFD)
.
European Commission – DG Joint Research Centre (JRC)
,
Ispra
,
Italy
.
Loos
R.
2012b
Chemical Monitoring under the Water Framework Directive (WFD) – Current Challenges NORMAN Inter-Laboratory Study (ILS) on Passive Sampling of Emerging Pollutants Chemical Monitoring on site in support of WFD implementation. Dissemination Workshop/JRC Innovation Transfer Event, Ispra, Italy
.
Manzo
S.
,
Buono
S.
&
Cremisini
C.
2006
Toxic effects of irgarol and diuron on sea urchin Paracentrotus lividus early development, fertilization, and offspring quality
.
Archives of Environmental Contamination and Toxicology
51
(
1
),
61
68
.
Miege
C.
,
Schiavone
S.
,
Dabrin
A.
,
Coquery
M.
,
Mazzella
N.
,
Berho
C.
,
Ghestem
J.-P.
,
Togola
A.
,
Gonzalez
C.
,
Gonzalez
J.-L.
,
Lalere
B.
,
Lardy-Fontan
S.
,
Lepot
B.
,
Munaron
D.
,
Tixier
C.
,
Miège
C.
,
Schiavone
S.
,
Dabrin
A.
,
Coquery
M.
,
Mazzella
N.
,
Berho
C.
,
Ghestem
J.-P.
,
Togola
A.
,
Gonzalez
C.
,
Gonzalez
J.-L.
,
Lalere
B.
,
Lardy-Fontan
S.
,
Lepot
B.
,
Munaron
D.
,
Tixier
C.
,
Miège
C.
,
Schiavone
S.
,
Dabrin
A.
,
Coquery
M.
,
Mazzella
N.
,
Berho
C.
,
Ghestem
J.-P.
,
Togola
A.
,
Gonzalez
C.
,
Gonzalez
J.-L.
,
Lalere
B.
,
Lardy-Fontan
S.
,
Lepot
B.
,
Munaron
D.
&
Tixier
C.
2012
An in situ intercomparison exercise on passive samplers for monitoring metals, polycyclic aromatic hydrocarbons and pesticides in surface waters
.
Trends in Analytical Chemistry
36
(
0
),
128
143
.
Miège
C.
,
Mazzella
N.
,
Allan
I.
,
Dulio
V.
,
Smedes
F.
,
Tixier
C.
,
Vermeirssen
E.
,
Brant
J.
,
O'Toole
S.
,
Budzinski
H.
,
Ghestem
J. P.
,
Staub
P. F.
,
Lardy-Fontan
S.
,
Gonzalez
J. L.
,
Coquery
M.
&
Vrana
B.
2015
Position paper on passive sampling techniques for the monitoring of contaminants in the aquatic environment – achievements to date and perspectives
.
Trends in Environmental Analytical Chemistry
8
(
November
),
20
26
.
Paíga
P.
,
Lolić
A.
,
Hellebuyck
F.
,
Santos
L. H. M. L. M.
,
Correia
M.
&
Delerue-Matos
C.
2015
Development of a SPE-UHPLC-MS/MS methodology for the determination of non-steroidal anti-inflammatory and analgesic pharmaceuticals in seawater
.
Journal of Pharmaceutical and Biomedical Analysis
106
,
61
70
.
Pereira
R. O.
,
Postigo
C.
,
de Alda
M. L.
,
Daniel
L. A.
&
Barceló
D.
2011
Removal of estrogens through water disinfection processes and formation of byproducts
.
Chemosphere
82
(
6
),
789
799
.
Petrović
M.
,
Hernando
M. D.
,
Díaz-Cruz
M. S.
&
Barceló
D.
2005
Liquid chromatography-tandem mass spectrometry for the analysis of pharmaceutical residues in environmental samples: a review
.
Journal of Chromatography A
1067
(
1–2
),
1
14
.
Smedes
F.
,
Bakker
D.
&
de Weert
J.
2010
The Use of Passive Sampling in WFD Monitoring
.
Deltares
,
The Netherlands
.
Snyder
S.
,
Lue-Hing
C.
,
Cotruvo
J.
,
Drewes
J. E.
,
Eaton
A.
,
Pleus
R. C.
&
Schlenk
D.
2009
Pharmaceuticals in the Water Environment. National Association of Clean Water Environment (NACWA) and Association of Metropolitan Water Agencies (AMWA), Washington, DC, USA, 38.
Tanwar
S.
,
Di Carro
M.
&
Magi
E.
2015
Innovative sampling and extraction methods for the determination of nonsteroidal anti-inflammatory drugs in water
.
Journal of Pharmaceutical and Biomedical Analysis
106
,
100
106
.
Togola
A.
&
Budzinski
H.
2008
Multi-residue analysis of pharmaceutical compounds in aqueous samples
.
Journal of Chromatography A
1177
(
1
),
150
158
.
Vrana
B.
,
Allan
I. J.
,
Greenwood
R.
,
Mills
G. A.
,
Dominiak
E.
,
Svensson
K.
,
Knutsson
J.
&
Morrison
G.
2005
Passive sampling techniques for monitoring pollutants in water
.
TrAC Trends in Analytical Chemistry
24
(
10
),
845
868
.
Vrana
B.
,
Smedes
F.
,
Prokeš
R.
,
Loos
R.
,
Mazzella
N.
,
Miege
C.
,
Budzinski
H.
,
Vermeirssen
E.
,
Ocelka
T.
,
Gravell
A.
&
Kaserzon
S.
2016
NORMAN Interlaboratory Study (ILS) on Passive Sampling of Emerging Pollutants
.
European Union, Brussels, Belgium
.
Wilkinson
J.
,
Hooda
P. S.
,
Barker
J.
,
Barton
S.
&
Swinden
J.
2017
Occurrence, fate and transformation of emerging contaminants in water: an overarching review of the field
.
Environmental Pollution
231
,
954
970
.