Review of passive accumulation devices for monitoring organic micropollutants in the aquatic environment

Environmental Pollution 136 (2005) 503e524 www.elsevier.com/locate/envpol

Review of passive accumulation devices for monitoring organic micropollutants in the aquatic environment Frank Stuer-Lauridsen* COWI Consulting Engineers and Planners, Parallelvej 2, DK 2800 Kongens Lyngby, Denmark Received 29 January 2004; accepted 10 December 2004

Major developments in the passive sampling of organic contaminants in aquatic environments will support future monitoring, compliance and research. Abstract Over the past 15 years passive sampling devices have been developed that accumulate organic micropollutants and allow detection at ambient sub ng/l concentrations. Most passive accumulation devices (PADs) are designed for 1e4 weeks field deployment, where uptake is governed by linear first order kinetics providing a time weighted average of the exposure concentration. Semipermeable membrane devices (SPMDs) are the most comprehensively studied PADs, but other samplers may also be considered for aquatic monitoring purposes. The applicability of the PADs is reviewed with respect to commonly monitored aqueous matrices and compounds, the detection limits, and for use in quantitative monitoring related to requirements embedded in the EU Water Framework Directive, the US and EU Water Quality Criteria, and the Danish monitoring aquatic programme. The PADs may monitor O75% of the organic micropollutants of the programmes. Research is warranted regarding the uptake in PADs in low flow environments and for the development of samplers for polar organic compounds. Ó 2005 Elsevier Ltd. All rights reserved. Keywords: Passive sampling devices; SPMD; Hydrophobic compounds; Monitoring; Contaminants; PAH; PCB; Pesticides

1. Introduction Aquatic monitoring programmes are generally based on collection of discrete samples of the water phase. The subsequent chemical analysis will provide a snapshot of the concentration in the environment at the particular time. In environments where the contaminant concentrations may vary over time, it is often desirable to expand the time window and increase the resolution by taking more samples. Such pseudo time-integrated sampling of water, be it automatic or manual, is both

* Tel.: C45 4597 2211; fax: C45 4597 2212. E-mail address: [email protected] 0269-7491/$ - see front matter Ó 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2004.12.004

costly and cumbersome, and rarely if at all used in large scale monitoring programmes. In this type of monitoring programme, samples of sediment or biota are the usual choice to represent time-integrated waterborne contamination on the long term (years) and short term (weeks to months), respectively. There are, however, drawbacks attached to the evaluation of aquatic quality based on sediment and biota data for organic micropollutants. It can be very difficult or impossible to assess the influence of for example sediment bioturbation and resuspension events, sediment sorbent quality, degradation and elimination rates, or the actual and recent condition of the sampled biota, all of which may have substantial influence on the observed concentrations. Passive sampling methods may allow for a combination

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of the control and reproducibility offered by the conventional grab sampling and the time-integration offered by sediment and biota sampling without the problems associated with sample history and degradation. Such methods directed at providing a feasible and reproducible way of monitoring the aquatic environment have emerged since the mid-1980s. These methods are candidate as versatile tools in aquatic monitoring programmes, allowing a time-integrated monitoring of organic micropollutants directly in the aqueous phase as an alternative to conventional sampling techniques. A passive sampling device for chemicals is an object that collects chemical compounds without provision of energy from an external source. This review deals with devices that have the potential to be used practically in monitoring organic micropollutants in the aquatic environment, i.e. it is restricted to devices accumulating the chemicals during an in situ deployment and allows this chemical to be quantified after retrieving the sampler. The present generation of simple and reproducible passive sampling devices allow for both screening or estimating low concentrations of organic micropollutants. The methods, the monitored compounds and the matrices are covered and discussed with respect to commonly monitored organic micropollutants. The achieved detection limits, precision and possibilities of quality control are included since these issues are to be addressed in relation to monitoring purposes. However, interesting applications are continuously published regarding passive sampler applications for ecotoxicology testing, for porewater measurements in sediment, and for comparing accumulation in passive sampling devices to the accumulation in biota and sediments, these issues are not a topic for this review. 1.1. Passive sampling devices For a number of years the monitoring of indoor airborne pollutants for worker’s health reasons has included passive dosimetry of noxious gasses, since the diffusive collection by such devices was demonstrated already in the early 1970s (Palmes and Gunnison, 1973). In environmental monitoring the analysis of collected pine needles is a well established method to monitor the occurrence of organic chemicals in air (Kylin et al., 1994), a methodology quite similar to indoor dosimetry, but equally passive since it does require the pine needle to live through the exposure phase. In the aquatic environment similar ‘‘dosimetry’’ is applied when indigenous or transplanted organisms are collected and analysed for their contamination. Sampling of for example fish, bivalves, macro algae is common practise in many national and international programmes directed at monitoring metals and organic pollutants in the aquatic environment. These programmes and their

biological ‘‘dosimeters’’ are, however, not aimed at providing quantitative information on the concentrations of the chosen compounds in the water phase, but rather indicate the level of contamination in the aquatic biosphere. There are exceptions, however; e.g. the transplantation method employing ‘‘moss bags’’ (Roy et al., 1996) does include a correlation of the transplants’ concentration to the aqueous contamination. The truly passive sampling devices for chemicals in the aquatic environment have only recently gained a wider reputation. 1.2. The first passive sampling devices for the aquatic environment The first passive sampling methods were aimed at monitoring the concentrations of dissolved inorganic compounds in surface water (Benes and Steinnes, 1974) and at the sedimentewater interface (Mayer, 1976; Hesslein, 1976) by measuring equilibrium concentrations in water enclosed in dialysis membranes. Similar ‘‘peepers’’ with a 0.45 mm filter are still used frequently in sediment for porewater measurements of selected metals (Bufflap and Allen, 1995). The drawback of these devices is that the few enclosed millilitres of water in a passive sampling device equilibrating at a 1:1 ratio with its milieu is not enough for analytical determination of either the concentrations of most environmentally relevant metals or the typical organic micropollutants of concern. Several methods capable of accumulating metals from the ambient aqueous phase have been published since the early 1980s (Srikameswaran et al., 1984; Morrison, 1987; Davison and Zhang, 1994) and simultaneously a quest for suitable methods for organic micropollutants has taken place. A method for accumulating organic compounds in a passive sampler in a soil environment was described by Coutant et al. (1985), and soon after several passive sampling devices for organic micropollutants in the hydrosphere were developed in a short period of time: the solvent-filled dialysis membrane bag (So¨dergren, 1987), the carbon filled passive dosimeter (DiGiano et al., 1989), the permeation sampler (Zhang and Hardy, 1989), and the SPMD (Huckins et al., 1990). Some of the general features of the PADs for the aquatic environment have previously been reviewed, e.g. in Palowitch (1996); Kot et al. (2000).

2. Description of PAD methods The desire to isolate the relevant compounds from their aqueous matrix or co-occurring compounds interfering with analytical chemical procedures typically calls for a relative selective collecting phase which may be an organic solvent, a resin or a polymer coating,

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and sometimes a diffusion barrier; a semipermeable membrane. A schematic presentation of an accumulating sampler is given in Fig. 1. Briefly, the process of accumulating a compound in the sampler is the following: a chemical compound in the water is carried to the sampler by convection; it diffuses across the boundary phase surrounding the sampler, and passes through the membrane pores by conduction. It is finally solubilised in the solvent or sorbed to a bonded receiving phase. The final phase is chosen to act as a sink for the chemical, thus ensuring an effective gradient across the sampler’s interface to the ambient water. Obviously, the effectiveness of the sampler is related to the surface area, and to increase this several of the samplers allow the membrane to completely enclose the receiving phase forming a bag, tube or sandwich, thus forming a permeable housing. The accumulating sampling devices will be referred to as passive accumulation devices (PADs), thus forming a linkage to the original pads developed for sampling in air (Palmes and Gunnison, 1973). Some of the devices reach equilibrium very fast and these are usually referred to as equilibrium sampling devices (ESDs). In Table 1 the range of identified PAD methods for the aquatic environment is shown. There is a variety of membrane and solvent/resin characteristics among the published PADs, but many of them have been tested with only one or a few compounds, and the performance characteristics toward a broader range of chemicals have not been investigated. It is clear from Table 1 that the PAD methods have been applied for a variety of purposes ranging from estimating the concentration compounds in true dissolution in the water phase to the collection of compounds for subsequent testing for

Fig. 1. Schematic representation of a passive accumulation device. The membrane is typically a hydrophobic material and the receiving solvent or resin acts as a sink for the compounds of interest. To increase the surface area several of the samplers allow the membrane to completely enclose the receiving phase, thus forming a permeable housing.

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toxicity. One particular method, the semipermeable membrane device (SPMD) has received considerable attention in the literature with more than 100 scientific papers published by 2003 (USGS, 2004). The present review does not attempt to cover all published studies on SPMDs, but rather presents examples on relevant applications for passive sampling methods, including SPMDs, that accumulate organic micropollutants from the aqueous environment. With the exception of SPMDs, the PADs have not been applied under conditions that resemble environmental monitoring programmes where many sites are sampled more or less simultaneously often by different laboratories. A few methods are included in Table 1 that may have potential for in situ aquatic monitoring purposes (EmporeÒdisk and solid phase micro extraction, SPME), although they are presently used as equilibrium sampling devices in rapid exposures in laboratory applications (Mayer et al., 2003). Other PADs are included even though they have only been used in soil studies, but because they are considered applicable in aquatic studies. The following four PADs have been used more widely.

2.1. Solvent filled dialyses tubing The first PAD method for the aquatic environment was developed by So¨dergren (1987) at the University of Lund, Sweden. It is a simple device consisting of a tube of dialysis membrane (regenerated cellulose, rayon) with 3 ml of organic solvent inside, typically hexane. The membrane excludes molecules larger than approximately 1000 Da, a size which is close to the cut off of the biological membrane. Hydrophobic organic micropollutants diffuse through the membrane from the water and accumulate in the organic phase. The method is based on the partitioning process similar to bioconcentration of hydrophobic organic contaminants in fish and invertebrates and was therefore believed to mimic bioconcentration. It is an extraordinary cheap and simple PAD that does not need a housing, sophisticated solvents, specialised tools or similar. However, the hydrophilic membrane may tend to impede the uptake of hydrophobic compounds. Herve et al. (1991) detected the less hydrophobic compoundsdhexachlorocyclohexane and chlorophenolsdbut not the more hydrophobic PCBs known to be present near the paper mill effluent. The solvent filled dialyses membrane tubing has a surface area to volume ratio of three in the reported configuration, and it can only be used with difficulty for quantitative monitoring of organic micropollutants in water (Johnson, 1991). It has not been used very much in recent years, and must be now be characterised as the prototype of the generation.

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Table 1 Overview of published PAD methods that have been used or may be used in situ PAD method Solvent-filled dialysis membranes Passive dosimetry

Content/membrane

Hexane or heptane/ dialysis membrane Activated carbon/ perforated acrylic Permeation sampling XAD-7, Tenax Ta/polycarbonate membrane Semipermeable membrane Triolein/polyethylene devices (SPMD) Passive sampling device C-18, XAD-4, 2,2,4trimethylpentane/ polyethylene Passive in situ Hexane/polyethylene concentration membrane extraction sampler (PISCES) Passive sampler (TMP) Trimethylpentane/ polyethylene EmporeÒ discs Filter with C18 resin/no membrane Groundwater monitoring Various adsorbents/Goretex membrane In situ monitoring Polyurethane/fibreglass Groundwater monitoring C18, XAD, Amberlite and polyurethane/ceramic membrane Passive sampler Porapak Q, Tenax-TA/ for volatile aromatics silicone polycarbonate in water Polar organic Various chromatographic chemical integrated materials/various sampler (POCIS) membranes EmporeÒ discs in situ Filter with C18 resin/ polysulfonene and polyethylene LDPE and silicone No solvent or resin/LDPE passive sampling or silicone Groundwater monitoring Dowex Optipore L-493/ ceramic membrane Blue cotton Phthalocyanine trisulphonate/no membrane TLC plate Thin-layer chromatography plate/no membrane Solid phase Various materials/no micro extraction membrane (SPME)

Target compounds

Sampling purposea

Reference

Rankb

Lipophilic

Screening

So¨dergren (1987)

M

p-Xylene and atrazin

Quantitative, not tested in situ Quantitative, not tested in situ

DiGiano et al. (1989)

L

Zhang and Hardy (1989)

L

Huckins et al. (1990)

H

PCBs and pesticides

Quantitative, screening, toxicity tests Screening in soil

Zabik et al. (1992)

L

Organochlorines

Screening

Litten et al. (1993)

L

Chlordane and dieldrin

Quantitative

Peterson et al. (1995)

M

Lipophilic

Equilibrium sampling

Verhaar et al. (1995)

L

VOCs and SVOCs

Screening

L

Aromatic compounds PAHs, PCBs and pesticides

Screening, toxicity tests Quantitative

Mehltretter and Sorge (1995) Madsen et al. (1996) Grathwohl and Schiedek (1997)

Monocyclic aromatics

Quantitative (via gas phase)

Phenolic compounds

Lipophilic

L L

Lee and Hardy (1998)

L

Polar organic compounds Quantitative

Alvarez et al. (1999)

L

Organic compounds

Quantitative

Kingston et al. (2000)

Lc

PAHs and PCBs

Booij et al. (2000)

Booij et al. (2000)

M

BTEX and naphthalenes

Quantitative

Martin et al. (2003)

L

Phthalates and aromatics

Toxicity tests

Sayato et al. (1990)

L

Organophosphates

Screening

LeBlanc et al. (2003)

L

Various compounds

Screening, equilibrium sampling

Verbruggen et al. (2000)

L

a Several sampling purposes are identified: determination of contaminant concentration in the water (quantitative,), qualitative or semiquantitative determination of occurrence (screening), collection of material for toxicity tests (toxicity tests), determination of concentration by equilibrium sampling (equilibrium sampling). b Applicability ranking for monitoring based on annual average of studies published with the method since first published (as of August 2004): H (high) O1; M (medium) O0.5; L (low)
0.5. c If combined with the studies involving EmporeÒ disc equilibrium sampling a medium ranking is achieved.

2.2. Semipermeable membrane device (SPMD) The SPMD was first published in 1990 (Huckins et al., 1990) and a substantial number of studies have followed (>100). The SPMD is a low density polyethylene tube (lay flat style) filled with approximately

1 ml of triolein and sealed at both ends. Due to the hydrophobic membrane material the hydrophobic compounds pass easily and accumulate in the triolein and in the membrane of SPMD. Ionised compounds or metals will not pass the membrane. The capacity in the membrane for compounds with an octanolewater

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coefficient less than 200 is little and SPMD is not an optimal collector for compounds with a LogPow less than three. It has a high surface-to-volume ratio of 400. SPMD has been reviewed for environmental applications (Huckins et al., 1996), analytical chemical issues (Petty et al., 2000b) and quality control (Lu et al., 2002). Recently, a guidance document has been developed (Huckins et al., 2002) detailing the procedures in the field and laboratory. The SPMD has also been used for sampling in air (e.g. Petty et al., 1993). It is available commercially and special deployment canisters have been developed. Several research groups work with related designs, choosing other materials and dimensions, e.g. Booij et al. (1998); Hofelt and Shea (1997). 2.3. EmporeÒ disks EmporeÒ disks are filters coated with a standard absorption material often a C-18 resin (octadecanyl), which are commercially available and conventionally used for extraction of hydrophobic organic compounds in water samples. It has a high surface-to-volume ratio, which prompts its use as an equilibrium sampling device for extraction directly in the sample; it has been tested in complex aqueous samples (e.g. Verhaar et al., 1995), and these applications with rapid equilibrium are used primarily in laboratory studies. However, a passive in situ sampler with a C18 EmporeÒ disk as the receiving phase and a long linear uptake phase due to a rate limiting membrane (polysulfon or polyethylene) has been tested by Kingston et al. (2000). The use of a well-known standard disk in terms of size, capacity, specificity of the sorbent, etc. adds to the applicability of the method as an ESD. Also, the surface area-to-volume ratio is advantageous compared to many PADs, but the freedom to choose from many receiving phases also fuels a need to determine uptake rates for a compound in each configuration, when used as a PAD. 2.4. Tri-methyl pentane passive sampler (TMP) Zabik et al. (1992) first tested polymeric tubing with tri-methyl pentane (iso-octane) in soil soaked with run-off from a pesticide rinse pad. They tested several receiving phases and membranes and found the combination of polyethylene and tri-methyl-pentane superior for pesticides. Peterson et al. (1995) tested a similar polyethylenetri-methyl pentane device for the aquatic environment in laboratory and in situ experiments. It was shown that the sampling of pesticides was linear and proportional to the concentration in the surroundings for at least 3e7 weeks. The method has been further developed and applied in several field experiments in river water (e.g. Leonard et al., 1999). Although not completely avoided, the use of TMP as the solvent receiving phase ameliorates the problem of prohibitive loss of solvents observed with, e.g.

507

hexane, when used with LDPE membranes. In the reported configuration it has a surface-to-volume ratio of three, and for the pesticides tested it has an accumulation factor of 200e300. The sampler is simple to make from standard materials, but has yet to gain a wider audience outside of Australia.

3. Other PADs In addition to the PADs, which have been used more widely, a number of different methods have been published which all may be applied in the aquatic environment. However, they have in common that they have only been used in one or a few studies (Table 1). In fairness, some PADs such as the LDPE sampler and the POCIS have yet only been on the ‘‘market’’ for a very short time.  Active carbon with a lid of acrylic polymer with holes creating a non-turbulent diffusion zone (DiGiano et al., 1989); the method has been tested in the laboratory on p-xylene and atrazine.  Polymer absorbent in a silicone polar carbon membrane; tested for a range of phenolic compound, although not in situ (Zhang and Hardy, 1989).  The passive in situ concentration extraction sampler (PISCES) used for PCB is a brass tubing filled with hexane and enclosed by polyethylene membranes (Litten et al., 1993). It is simple and cheap to produce, but has a poor surface area-to-volume ratio (!0.1) and is prone to loss of solvent.  Groundwater well samplers: various adsorption materials enclosed in a Gore-Tex membrane (Mehltretter and Sorge, 1995) or a ceramic membrane (Grathwohl and Schiedek, 1997); recently, also a Teflon housing with ceramic membranes and Dowex optipore 1-493 adsorption material (Martin et al., 2003). The samplers have been used for monitoring of BTEX, PAH, PCB and pesticides.  Polyurethane plugs in fibre glass used for toxicity testing (Madsen et al., 1996). This device was produced to meet a specific research purpose rather than as a new passive sampler, but the use of polyurethane plugs, well known from sampling of contaminants in air, presents a new configuration of a PAD.  A passive sampling device for volatile aromatics in water has been tested under laboratory conditions (Lee and Hardy, 1998). The principle is to collect compounds in the gas phase of the sampler during the exposure in the water. This approach resembles that of the groundwater samplers mentioned above.  Devices made from silicone tubing or LDPE tubing without solvent or resin has been applied by Booij et al. (2000); Luellen and Shea (2002) for PCBs and

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Fig. 2. Time line for the development of passive samplers for the environment.

PAHs. Some simplicity in preparation and postdeployment procedures and in interpretation is offered by the omission of the lipid receiving phase of the SPMD.  A PAD with the acronym POCIS was presented by Alvarez et al. (1999) and a patent application on this sampler has been submitted (Petty et al., 2001). The POCIS is directed towards polar compounds, and in a field study in Denmark 29 of 46 monitored polar pesticides were detected (Stuer-Lauridsen et al., 2004). The POCIS sampler is comparable to the EmporeÒ disk sampler in the way that it can be loaded with different resins. Several passive devices are used without a diffusion limiting membrane and employ more specific resins: The

‘‘blue cotton’’ method is used often in Japan for collecting large amounts of environmental contaminants for later testing for mutagenicity (Sayato et al., 1990). Blue cotton is a rayon backbone material impregnated with the densely blue copper phthalate cyanine trisulphonate, and when deployed in situ it samples aromatics with three or more rings and phthalates. The rapid laboratory extraction method SPME which is available for a variety of compounds in water and air (Eisert and Levsen, 1996) is emerging as an equilibrium sampling device (Mayer et al., 2003). It has not yet been used for in situ monitoring of the water phase, but is already popular for estimation of bioconcentration potential in laboratory experiments (Verbruggen et al., 2000). Recently, LeBlanc et al. (2003) presented a PAD based on thin-layer chromatographic plates which has been tested in situ with two organophosphate pesticides. When it comes to deciding procedures of a monitoring programme, one of the key properties of a sampling method and analytical chemical procedure is its robustness. In the case of PADs exposed for extended periods of time the physical robustness is of primary concern, but also the sampling method’s precision when applied by the many different people involved in a large scale monitoring programme should be considered. Among the PADs, the SPMDs are presently used by several laboratories and deployment procedures and clean up methods are standardised. The PADs based on standard materials have an inherent advantage in this respect, e.g. the EmporeÒ disk method; a future SPME method or the simple LDPE method should also be robust (Fig. 2).

4. PADs in aqueous matrices The majority of the reported deployments of PADs have been in surface waters, both limnic and marine waters. However, PADs have been used in a diversity of aqueous matrices and examples will be drawn from studies of sewage water, groundwater wells, storm water overflows and other matrices of the hydrosphere. Table 2 presents an overview of the matrices where selected PAD methods have been used.

Table 2 Matrices where PAD methods have been applied with an indication of the typical levels of contamination measured Matrix

SPMD

Solvent-filled dialysis membrane bags

EmporeÒ discs

TMP

Other PAD methods

Wastewater and storm water Aquifer, well water Rivers and streams Lakes Marine waters Soil sediment

H H H, L L L H

H H H H H H

H, L e L e L e

e e L L e H

e H H H H H

The approx. criteria for high (H) and low (L) concentrations estimated in the matrix is 100 ng/l.

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An important feature of the PAD is that is does not sample the ‘‘third’’ phase of sorbed compounds in natural water. In several studies it has been shown that PADs sample only the truly dissolved phase and that contaminants bound to dissolved organic matter or particles are not transported into SPMDs (Ellis et al., 1995; Axelman et al., 1999), or into solvent-filled dialysis membrane bags (Knulst, 1992; Stuer-Lauridsen, 1996). 4.1. Wastewater, storm water and groundwater In the harsh wastewater environment SPMDs have been deployed in the sewerage line in a screening survey to identify unknown sources of several organic pollutants (Stuer-Lauridsen and Kjølholt, 2000), but SPMDs have also been used directly in the purified effluent of a WWTP by Petty et al. (2000a). In a study of aromatic and chlorinated compounds, Wang et al. (2001) deployed SPMDs at several sampling sites within the main WWTP in Beijing. In WWTPs EmporeÒ disks (Verbruggen et al., 1999; van Loon et al., 1996), solvent filled dialysis membrane bags (So¨dergren, 1987), and recently the chromatography plate method (LeBlanc et al., 2003) have also been used. None of the studies report damage to the SPMDs or to the other samplers from the deployment conditions preventing the collection of compounds. SPMDs were also used to monitor PAHs in a creek, which receives storm water overflow from urban areas (Lebo et al., 1996). Passive conventional sampling is already an established methodology in monitoring of landfill percolate and contaminated groundwater, but several PAD methods have been developed specifically for this matrix (DiGiano et al., 1989; Grathwohl and Schiedek, 1997; Martin et al., 2003). Gustavson and Harkin (2000) measured 16 PAHs in contaminated groundwater wells with SPMDs and found a good agreement with conventional sampling. Finally, Zabik et al. (1992) tested several prototype PADs in the laboratory and exposed one in situ in soil affected by run-off from a pesticide rinsing area. Exposure and calculation of aqueous concentrations in these matrixes from PADs differ from working in the aquatic environment in general, because the volume of water in a groundwater well, typically in a contaminated site, or in the interstitial space in soil or sediment is limited and often less than the clearance volume of the sampler. In that case modified equations for calculation of aqueous concentrations of contaminants are used. 4.2. Freshwater Through the work of Huckins, Petty and others SPMDs have been deployed and studied primarily in the freshwater environment, which is not surprising as the

509

developers’ laboratory is situated in Columbia, Missouri, in the central part of the USA. The deployments range from tiny creeks (Lebo et al., 1995) to the Missouri River (Petty et al., 1995b) in the USA, but it is beyond the scope of the present review to present the extensive SPMD research. Most of the applications are focussed on a recurring suite of priority pollutants: organochlorine pesticides, PCBs, polychlorinated dioxins/furans and PAHs. A survey of SPMD applications in freshwater and marine waters can be found in Prest et al. (1995) and in Huckins et al. (2002). A number of researchers have taken SPMDs to rivers in other parts of the world and frequently monitored a broader range of compounds: e.g. Wang et al. (1999b) to the Yanghe River in China, Norrgren et al. (2000) to the Kafue River in Zambia, and Sabaliunas and So¨dergren (1997) to the Neris and Obelis Rivers in Lithuania. PAD monitoring of contamination of lakes has received less interest, but in several local monitoring programmes of paper mill effluent in Finland, chlorinated compounds in lakes have been studied with solvent filled dialysis membrane bags (Herve et al., 1991) and SPMDs (Herve et al., 1995). The two methods have also been used for collection of PCBs (Gale et al., 1997; Johnson, 1991) and for PAHs (Lebo et al., 1995) in lakes. Solvent filled dialysis membrane bags (Herve et al., 1991; Knulst, 1992), the EmporeÒ disk (Verbruggen et al., 1999), the PISCES (Litten et al., 1993), and the TMP-sampler of Leonard et al. (1999) have also been used in the freshwater environment. All samplers appear to be well suited for freshwater studies. 4.3. Marine waters In the marine environment, particularly SPMDs have been used. The locations have often been harbours (Hofelt and Shea, 1997), bays and other near-coastal environments in the proximity of local emission sources (Peven et al., 1996; Granmo et al., 2000). SPMDs have also been used in less sheltered coastal areas (Prest et al., 1995; Booij and van Drooge, 2001) and in a few occasions under the conditions of open sea: near an offshore oil production platform in a side-by-side experiment with mussels (Røe and Johnsen, 1999), and in a study off the coast of Sweden of the contaminants in the polluted plume of more brackish water resulting from the 1995 floodings in Europe (Bergqvist et al., 1998). In a harbour and coastal environment solvent-filled dialysis membrane bags have also been used to study the concentration gradient of tributyltin in the water column towards the sediment surface (Stuer-Lauridsen and Dahl, 1995). None of the samplers found in the literature are quoted to have properties that would render them incompatible with the aqueous matrices of interest. However, the somewhat fragile solvent-filled dialysis membrane bags are not well suited for high energy and

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a number of parameters related to the partitioning process (receiving phase, ionisation, Kow) and the permeability of the membrane, and to the analytical procedures applied after retrieving the sampler from the environment. Since the sampler indiscriminately collects a wide range of compounds with properties similar to the interesting ones, these procedures are crucial to limit interference and may include extraction of the sampler material, clean up, derivatisation, reextraction, evaporation and finally the applied instrumental methods. A thorough review of the consideration of these parameters in relation to SPMDs can be found in Petty et al. (2000b). In Table 3 a comparison is shown of the compounds reported from in situ and laboratory studies with PADs and the requirements for monitoring under a number of international agreements and with the Danish monitoring programme as an example of national monitoring requirements.

abrasive environments such as the coastal zone where the SPMDs or the polyethylene samplers (Booij et al., 2003) appear to be more robust.

5. Compounds detected in PADs The PAD will slowly extract compounds from the water when deployed and the accumulation of the compounds in the less than 10 ml receiving phase typically ranges between 50 and 1000 times depending on the compound and the PAD. In the SPMD this means that between 0.5 and 5 l of water is extracted, and it is therefore possible to detect low nanogram per litre concentrations of organic micropollutants with one sampler. It is important to note that one single sampler does not collect all the organic micropollutants relevant to monitoring programmes. The integrated selectivity and specificity of a given sampler is both a function of

Table 3 Compounds identified under the EU Water Framework Directive (EU, 2001), the priority substances of regional conventions for the North East Atlantic (OSPAR, 2002,), the Baltic Sea (HELCOM, 1998), and the organic micropollutants listed for surface water monitoring in the Danish NOVANA programme (NERI, 2002) compared to the compounds detected with PADs in the aquatic environment (both in situ and laboratory data) Compound group

EU WFD

OSPAR

HELCOM 19/5

Danish Aquatic Monitoring Programmea

PADsb

Pesticides Organochlorine pesticides Organophosphates Pyrethroids Phenoxy acids Sulphonyl urea and urea derivatives Triazines Dithiocarbamates Carbamates Benzonitrile Halogenated carboxy acids Aromatic hydrocarbons (BTEX) Alkyl phenol compounds Halogenated aliphatic hydrocarbons Halogenated aromatic hydrocarbons Brominated diphenylethers Polychlorinated phenyls Chlorophenols Polycyclic aromatic hydrocarbons Phosphate triesters Softeners (phthalates) Anionic detergents (LAS) Organotin compounds Dioxins and furans Steroid hormones Pharmaceuticals Otherd

e X X e e X

e X e e e e

e X e e e e

e Fc F e F F

e SPMD, SF, TMP, PSD SPMD, TIMP SPMD, TIMP e PSD

X e e e e X X X X X e e X

X e e e e e e e e e X e X e e e X X X e e

e e e e e e X X e e X e X e X e X X X X X

F, F F F F e F F, e e F F, F e F F, F F

a

X e X e e e e

e e

M

M

M

M

SPMD, e e e e SMPD, SPMD SPMD SPMD, SPMD SPMD, SPMD, SPMD, SPMD SPMD SPMD SPMD, SPMD SPMD e e

ED, PSD

SF

ED SF, ED, PSD SF ED, SF

SF

F designates freshwater and M marine water. The passive samplers are abbreviated as follows: SPMD, semipermeable membrane device; SF, solvent-filled dialysis membrane bags; ED, EmporeÒ discs; TMP, tri-methyl pentane sampler; PSD, other passive sampling devices mentioned in Table 1 except SPMD, SF, ED and TMP. c Not in aqueous samples. d In situ detection of petroleum hydrocarbons, triclosan, chloro compounds, trialkyl phosphates, fragrances are reported from SPMD, ED and PSD. HELCOM 19/5 mentions algae metabolites. b

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5.1. Organochlorines and organohalogens Organochlorine pesticides and other organochlorines, such as PCBs, have been amongst the most frequent compounds used for testing of SPMDs and other PADs. The pesticides are still widely used and the pesticides and PCBs are part of many monitoring programmes and listed among the ‘‘dirty dozen’’ of the Convention on Persistent Organic Pollutants (United Nations, 1998). The pesticides have most often been hexachlorocyclohexane (lindane, g-HCH) or dichlorodiphenyl trichlorethane (DDT), but many other organochlorine compounds have been detected. The SPMDs, with their strong hydrophobic properties, have been applied frequently in studies of these compounds and besides lindane and DDT the list includes hexachlorobenzene, aldrin, dieldrin, heptachlor, chlordan, nonachlor, endrin, toxaphene and mirex (Huckins et al., 1990; Prest et al., 1992; Petty et al., 1995a; Ellis et al., 1995; Hofelt and Shea, 1997), polychlorinated dioxins and furans (Lebo et al., 1995) and PCBs (Bergqvist et al., 1998). Solvent filled dialysis membrane bags have also been used for monitoring organochlorines (So¨dergren and Okla, 1988; Herve et al., 1991; Burmaster et al., 1991), and other PADs have been calibrated and tested for dieldrin (Kingston et al., 2000) or monitored chlordane and dieldrin in situ (Peterson et al., 1995). Polybrominated diphenyl ethers have recently appeared as organohalogens of environmental concern and they have been detected with SPMDs (Booij et al., 2001; Rayne and Ikonomou, 2002). Studies of the short chain chlorinated paraffins have not been found, although these compound are frequently listed as relevant for monitoring. 5.2. Various pesticides and biocides Chlorpyriphos and malathion are examples of organophosphates that have been tested in SPMDs (Sabaliunas and So¨dergren, 1997; De la Torre et al., 1995), and recently LeBlanc et al. (2003) also tested their thin layer chromatography plate sampler on diazinon and chlorpyrifos. Several pyrethroids have been measured with SPMDs: deltamethrin and fenvalerate (Sabaliunas and So¨dergren, 1997), fenvalerate (Huckins et al., 1990), esfenvalerate (Stuer-Lauridsen and Nielsen, 1999; Bu¨gel Mogensen et al., 1998), and allethrin (Wang et al., 1999a), but not with other PADs. Atrazine has been tested with SPMD (Wang et al., 1999b), but triazines are relatively water soluble and have otherwise only been collected with the emerging polar samplers: atrazine in POCIS (Alvarez et al., 1999) and the antifouling agent Irgarol 1059 with the in situ EmporeÒ sampler (Kingston et al., 2000). A range of cotton farming pesticides were tested by Leonard et al. (2002) with a PAD based on Peterson et al. (1995), and Zabik et al. (1992) tested endosulfan and trifluralin in soil soaked with rinse pad

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run-off. Many modern pesticides are more polar than most PADs are designed for. However, in a recent 5 month study of three streams in Denmark, 29 of 46 monitored polar pesticides were detected with the POCIS sampler (Stuer-Lauridsen et al., 2004). 5.3. Aromatic and alkylated aromatic compounds The volatile BTEXs (benzene, toluene, ethylbenzene, xylene) and two-ringed aromatics are relatively water soluble compounds and have therefore received less attention with PADs than the polyaromatic compounds. SPMDs have been used for BTEX studies in contaminated river and urban stream water (Wang et al., 1999b; Moring and Rose, 1996), but the capacity of the SPMD for these less hydrophobic compounds is little and during standard 28 days of exposure the sampler may reach equilibrium. Polyurethane in fibreglass bags has been used in interstitial water and soil for the collection of methylnaphthalenes and subsequent test of toxicity (Madsen et al., 1996). When it comes to the ubiquitous environmental contaminants PAHs, most PADs are very well suited for these compounds and PAHs have been investigated in a number of studies. The full range of PAHs, alkylated PAHs and heterocyclic sister compounds have formed core compounds for much of the SPMD research. A few examples of studies are given here: PAHs (Bennett et al., 1996; Lebo et al., 1996; Stuer-Lauridsen and Kjølholt, 2000; Stuer-Lauridsen et al., 1998), PAHs and methylated PAHs (Peven et al., 1996; Sabaliunas and So¨dergren, 1997), and PAHs including the heterocyclic thiophenes and carbazoles (Moring and Rose, 1996; Lebo et al., 1992). Blue cotton has also been used for in situ collection of PAHs (Sayato et al., 1993), although primarily for testing of carcinogenic potential. In one case the solventfilled dialysis membrane bags was (partly) unsuccessfully applied to detect PAHs (Burmaster et al., 1991). Finally, in a laboratory test of the in situ EmporeÒ disk the three ring PAH phenanthrene, was used as the test substance (Kingston et al., 2000) and in many studies not related to PAD the EmporeÒ disk has also been shown to sample PAHs. Phenols and phenolic compounds have been studied in laboratory investigations with the permeation sampling device developed by Zhang and Hardy (1989), and Granmo et al. (2000) determined chlorophenols in situ in seawater with SPMDs. Also, the pesticide degradation product and traffic emission component 4-nitrophenol has been monitored with SPMD (Wang et al., 1999a). Finally, the hormone disrupting nonylphenol and nonylphenol ethoxylates have been monitored in situ with SPMDs (Stuer-Lauridsen and Kjølholt, 2000). No record of Bisphenol A detection has been found, and remarkably little information on phthalates or other plasticisers is available from PADs. This latter may

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be due partly to the occurrence of DEHP as a manufacturing component in the SPMD, thus requiring a special cleaning of the SPMD if DEHP is of concern in the environment. 5.4. Other compounds A range of compounds other than those mentioned above are often included in monitoring programmes or in special surveys, such as antifouling agents, pharmaceuticals or surfactants. The latter are always problematic in any partitioning process and PADs which share this feature with standard chemical extractions, are not a good collector of surfactants except for the non-ionic surfactants, such as alkyl phenol ethoxylates. Pharmaceuticals have recently emerged as environmental contaminants (Halling-Sørensen et al., 1998), but as yet only steroid estrogens have been tested in POCIS (Alvarez et al., 1999). Tributyltin (Stuer-Lauridsen and Dahl, 1995) and triphenyltin (Stuer-Lauridsen, 1996) has been measured in situ with solvent-filled dialysis membrane bags and several organotins have been found with SPMDs (Følsvik et al., 2000). These compounds present analytical challenges in conventional water samples because they need to be determined at sub ng/l concentrations, but due to their positive charge they are also difficult to determine at these levels with the PADs used until now. As previously mentioned the antifouling agent Irgarol 1059 has also been found with the in situ EmporeÒ sampler (Kingston et al., 2000). The bactericide triclosan and its derivative methyl-triclosan have been found in SPMDs exposed in Swiss waste water (Lindstro¨m et al., 2002). Chlorinated aliphatic hydrocarbons, for example trichlorethylene, have been measured by SPMDs in the water phase (Stuer-Lauridsen and Kjølholt, 2000), but these compounds are typically too volatile for measurements in water. Volatiles in sewage systems have been measured by a PAD exposed in the overlying air space (Vorkamp et al., 1998). Other compounds such as fragrances have been reported in an EmporeÒ disk study (Verbruggen et al., 1999) and tributylphosphate was detected with SPMD in wastewater (Stuer-Lauridsen and Kjølholt, 2000). In association with tracking and identification of crude oil it has been shown by Luellen and Shea (2003) that the biomarkers for petroleum hydrocarbons, hopane and sterane, and their distinct and recognisable pattern are conserved in SPMDs. It is obvious that not one single PAD has a suitable capacity for all the organic compounds of interest, in particular the more water soluble compounds. This is hardly surprising since a sequestering phase and membrane of low polarity by nature do not form a good media for hydrophilic compounds. However, the hydrophobic properties of the SPMD allow for the detection of extremely hydrophobic compounds in the aqueous

phase, including polychlorinated dioxins and furans, PCBs, four ring and larger PAHs, brominated biphenyl ethers and pyrethroid insecticides. Although the uptake constant of compounds above logKow 5e6 decrease, these compounds can be detected by SPMDs at ambient concentrations, and this is the type of compound that is otherwise not included in a monitoring programme for water because of analytical difficulties. The more hydrophilic compounds must be sampled using PADs with polar membranes. Only recently have such PADs emerged that will accumulate charged or polar compounds, such as many modern pesticides and pharmaceuticals. 6. Application for monitoring programmes 6.1. Quantitative use/measurement of concentration in water phase Most published PAD methods aim at calculating the concentration of the compounds in the aqueous phase based on the accumulated concentration in the PAD. The equations used for estimating the aqueous concentration based on PADs are all variations of the first order relationship (a selection is shown in Appendix A). The simplest way of estimating the concentration in the ambient environment is at or near the equilibrium between the PAD and the water phase (Fig. 3). Under this regime a first-order equation reduces to Eq. (1) and to calculate the concentration in the water it is only required that a distribution coefficient KPAD and the concentration of the substance in the PAD (CPAD) are known: KPAD Z

CPAD CPAD 0 CWater Z CWater KPAD

ð1Þ

Fig. 3. Time dependent concentration profile of organic micropollutant in passive accumulating sampling device. Most in situ samplers, such as SPMD, TMP and Empore disk (in situ) have a slow uptake and are exposed and collected during a linear phase regime, whereas equilibrium samplers are exposed until they have entered the nearequilibrium phase.

F. Stuer-Lauridsen / Environmental Pollution 136 (2005) 503e524

For the PAD based on hydrophobic interaction it should be possible to estimate the KPAD from inherent properties of the substance as expressed in values for fugacity or Kow. Methods with a rapid equilibration time of hoursedays such as the SPME and EmporeÒ disk without a diffusion limiting boundary may employ such an approach (Mayer et al., 2003). However, this approach is not suited for PADs with more than month long equilibration times such as SPMDs (Huckins et al., 1996) or solvent filled dialysis membrane bags (Johnson, 1991). For the SPMDs Huckins et al. (2002) present the following equation solved for the concentration of compound in water and with diffusion in the membrane as the rate controlling step: CWater Z

CSPMD !VSPMD Rs !t

ð2Þ

where VSMPD is the volume of SPMD and Rs is the SPMD sampling rate (the volume of water cleared per time unit, 1/day). Rs is conceptually linked to the clearance rate for, e.g. bivalves, and it is specific for the compound and the PAD, and needs to be determined in a laboratory experiment. This is the equation typically used for estimating the aqueous concentration of a compound, and the necessary Rs values for standard SPMDs can be found in Huckins et al. (1996, 2002) for PAHs, PCBs and certain pesticides. The Rs typically fall in the range of 0.5 to 5 l/day with the most hydrophobic compounds having the higher values. A more comprehensive model, a three layer model, has been proposed (Gale, 1998) and also a simpler equation has been proposed (Bu¨gel Mogensen et al., 1998). In groundwater, the uptake of PAHs into a PAD with air-filled adsorbent, has been estimated with a double film diffusion model (Martin et al., 1999). In several studies it has been shown that SPMDs absorb from the truly dissolved concentration of the aqueous phase (Ellis et al., 1995; Axelman et al., 1999). Therefore, the aqueous concentration estimated with PADs does not necessarily reflect the concentration measured in unfiltered conventional water samples, particularly in the case of very hydrophobic compounds and/or high concentrations of dissolved organic matter. Provided the concentration of dissolved organic carbon is known, a simple recalculation of the concentration of an organic micropollutant accumulated in the SPMD to the total concentration in the water can be made (Meadows et al., 1998), thus facilitating comparison to a back catalogue of values from traditional water sampling. Since SPMDs do not sample ionised compounds, but do accumulate the protolysed uncharged species, a recalculation is employed based on pKa, see e.g. the study on chlorophenols (Granmo et al., 2000). In other attempts to estimate aqueous concentrations of micropollutants based on partitioning, several studies

513

have been carried out on the use of SPMDs and fish or sediment. Here, most often simple estimates of steady state concentrations are given. In one study of PCBs in Saginaw River the aqueous concentration was calculated from the PCB congener concentrations in sediments, SPMDs and fish (Echols et al., 2000). The estimated aqueous concentrations of PCBs (approximately 2e20 ng/l) based on accumulation in SPMDs or fish were overall similar to published water concentrations (1.4e 14 ng/l), but for individual congeners large differences occurred related to the properties of the three receiving phases of fish, sediment and PAD. An inherent problem with direct comparison of the aqueous concentration derived from SPMDs, sediment and biota is that different equilibration times may lead to comparison of results from different levels of non-equilibration (Gale et al., 1997). Other studies instead calculate concentrations based on uptake rates (e.g. Rantalainen et al., 2000). Rantalainen et al. (1998) also compared aqueous concentrations of PCDDs, PCDFs and PCBs in the Lower Fraser River collected by SPMDs and Infiltrex sampler, i.e. extraction of 100 l of water by XAD-2 resin. They demonstrated that at low concentrations of the compoundsddown to 1015 g/ldthe ‘‘conventional’’ and passive sampling yielded comparable results (same order of magnitude) were found for those compounds not affected by high blank levels. The calculation and interpretation of time weighted average (TWA) based on linear uptake rates has been applied for air sampling in the work environment for SPME (Martos and Pawliszyn, 1999). The aqueous concentrations calculated by the equations mentioned above will also derive TWAs of the exposure concentrations as described by Huckins et al. (1996). During deployment of a PAD the aqueous concentration of a compound may not be constant, but increase and decrease under the local conditions. Although, theoretical considerations are given in Huckins et al. (1996); Gale (1998), experimental studies with pulsed aqueous concentrations of the contaminants and the consequence for the TWA concentration in the SPMDs or other PADs have not yet been reported. Hickie et al. (1995) used a simple first-order toxicokinetics model approach to describe bioconcentration in fish exposed in pulses and this may also prove valuable for PADs. For the rapid equilibrium samplers Mayer et al. (2003) points out that a basic requirement is that the device response time must be shorter that the fluctuations being measured.

6.2. Flow conditions The importance of the flow for the mass transport over the surface of SPMDs has been debated intensely since an important assumption for recalculation of

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accumulated amount to aqueous concentration is that the rate limiting step is the crossing of the polymeric membrane rather than the aqueous boundary layer. Under flow velocities of 30 cm/s or less the exchange kinetics over the membrane for compounds with Log Kow>4 are governed by the unstirred layer of water (Booij et al., 1998). The change in mass transfer coefficient was 3e4 for a 1000-fold change in flow velocity from 0.03 to 30 cm/s. In flow experiments by Huckins et al. (2002) only a 50% increase in SPMD sampling rates was found for flow regimes from 0.004 to 0.2 cm/s. In many environments flow is considerably slower than the 30 cm/s and the exposure regime may influence the observed uptake. Vrana and Schu¨u¨rmann (2002) also concluded that flow under environmental conditions affected the sequestration in the SPMD. Louch et al. (2003) observed that on a three site transect across a river, the site with the lowest flow had the highest concentrations and they argue that since this is contrary to theory, it must be a result of actual differences in concentrations across the river. Information on flow, temperature and the measurement of a permeability reference compound (PRC) may provide valuable information to correct for flow conditions. The use of permeability reference compounds in quantitative analysis is treated below. The effect of an environmental range of flow on the boundary layer appeared to be up to a factor of 3e4 (Booij et al., 1998); however, this was measured without the protective cages typically used in field deployments, which will reduce the effect of turbulent flow. Under the very low flow regime in a groundwater well Gustavson and Harkin (2000) used an encounter volume model to describe contaminant accumulation, which is also the method suggested by DiGiano et al. (1989) for their passive dosimetry device. The model is applicable when the PAD clearance volume is greater than the water turnover volume as in groundwater wells and interstitial spaces where the PAD may clear the volume of contaminants encountered.

6.3. Biofouling and permeability Any unprotected surface submersed in an aqueous ecosystem will eventually be the substrata for bacteria, flora and fauna. However, several PADs, i.e. the solvent-filled dialysis membrane bags, the PISCES and the TMPs have been exposed for many weeks in marine and freshwater without significant biofouling. This impregnating effect has been attributed to the slow seeping of solvent from the sampler through the membranes. Indeed, these samplers do suffer from loss of their low molecular weight solvents (Stuer-Lauridsen, 1996; Litten et al., 1993; Leonard et al., 1999), and the authors typically state that only samplers with a given residual volume are used for the study. In contrast, the

SPMD with its non-toxic high molecular weight solvent tends to acquire a biofilm and in some cases an epiphyton growth during the exposure and the sampler must be cleaned when retrieved. More important, however, is that the biofouling lowers the exchange rate between the SPMD and the ambient environment by increasing the thickness of the diffusion layer surrounding the sampler. Huckins et al. (2002) reported 20e70% impedance in PAH uptake in severe cases. Richardson et al. (2002) concluded that biofouling, and the measures applied to remediate it, complicates the use of SPMDs beyond practical applications, but others routinely use a PRC to correct for biofouling and other environmental conditions during deployment.

6.4. PAD methods and detection limits It has already been showed by DeVita and Crunkilton (1998) that detection limits of SPMDs for PAHs are well below aqueous detection limits required by US EPA. In Table 4 the detection limits published for PAD methods are compared to the quality objectives for surface waters in the EU and the required detection limits of the Danish surface water monitoring programme. The majority of published detection limits for PADs are from studies with SPMDs, and here it is typically seen that the detection limits required for monitoring purposes and the level of quality criteria relevant for organochlorine pesticides, PCBs and PAHs and other hydrophobic compounds are met by the detection limits of SPMDs. Most other passive sampling devices do not meet the detection limits required for hydrophobic compounds in aquatic ecosystems not affected by point sources. Calculations of reliable quantitative estimates of aqueous concentrations of organic micropollutants have not been possible for solvent-filled dialysis membrane bags (Johnson, 1991). However, with the TMP sampler Peterson et al. (1995) detected 4 ng/l chlordane and 18 ng/l for dieldrin as the lowest estimated concentrations, and 10 ng/l was estimated for endosulfan compounds by Leonard et al. (2000). Detection at this lower range would suffice for most aquatic monitoring of hydrophobic organochlorine and organophosphate pesticides. For some of the less hydrophobic compounds such as diuron and triazine pesticides, methods such as the EmporeÒ disk and POCIS are available. Although not a typical monitoring object, fragrances have been estimated in surface water samples down to 1 ng/l by the EmporeÒ disk (Verbruggen et al., 1999). No limits of detection are available for SPMDs regarding naphthalene, but for PAHs it can be seen that SPMDs meet the required detection limits. In a comparison between conventional sampling and SPMDs in groundwater wells, the PAD exhibited 3e70 times lower detection limits for 16 PAHs

Table 4 Detection limits required in Danish monitoring programme, EU quality criteria and the lowest estimated aqueous concentrations or aqueous detection limit for SPMD or other PADs (in ng/l) Compounds

Danish monitoring requirements (NERI, 2002)

EU quality criteria for the aquatic environment (Bro-Rasmussen et al., 1994)

SPMD based detection limit or lowest measurement

e 10 10 10 10 10 10 10 e e 10 e e 50a 100a 100a 100a 50a e

1e10 2 2 10 5 10 10e200 1000 e e 10e10,000 e e 10,000 10,000 10,000 10,000 1000 e

e 0.02 0.01 0.25 28 0.1 300b e 100 e e e e e 100c e 100c e e

10 10 10 10 1 10 10 10 10 10 10 10 10 10 e 30 50 50 10

e e e e 10 e e e e e e e e e e e e e e

10.2 9.4 4.9 3.0 51 2.8 2.3 2.1 2.8 3.3 4.5 9.2 10.3 8.6 e 23,500 e 3300 2

Ref.

Other PAD-based detection limit or measurement

Petty et al. (1995a) Petty et al. (1995a) Petty et al. (1995a) Petty et al. (1995b) Ellis et al. (1995) Sabaliunas and So¨dergren (1997) Wang et al. (1999b)

Wang et al. (1999b) Wang et al. (1999b)

DeVita DeVita DeVita DeVita DeVita DeVita DeVita DeVita DeVita DeVita DeVita DeVita DeVita DeVita

and and and and and and and and and and and and and and

Crunkilton Crunkilton Crunkilton Crunkilton Crunkilton Crunkilton Crunkilton Crunkilton Crunkilton Crunkilton Crunkilton Crunkilton Crunkilton Crunkilton

Wang et al. (1998) Wang et al. (1998) Petty et al. (2000a)

(1998) (1998) (1998) (1998) (1998) (1998) (1998) (1998) (1998) (1998) (1998) (1998) (1998) (1998)

e e e 4 e e 10e40 e e 8.5 e 42 e e e e e e

e e e e e e e e e e e e e e e 5000 5000 5000 e

Ref.

Peterson et al. (1995)

Leonard et al. (2002)

Kingston et al. (2000) Kingston et al. (2000)

F. Stuer-Lauridsen / Environmental Pollution 136 (2005) 503e524

Organochlorine pesticides 4,4#-DDT DDT metabolites Dieldrin Endrin Lindane Organophosphates Triazines Atrazin Irgarol Various pesticides Diuron Aromatic hydrocarbons Benzene Toluene Ethylbenzene o,m,p-Xylene Naphthalene Polycyclic aromatic hydrocarbons (PAHs) Acenaphthylene Acenaphthene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benzo[a]anthracene Benzo[b]fluoranthene Benzo[k]fluoranthene Benzo[a]pyrene Indeno[1,2,3-cd]pyrene Dibenzo[a,h]anthracene Benzo[g,h]perylene Chlorophenols 2,4-Dichlorophenol 4-Chloro-3-methylphenol 2,4,6-Trichlorophenol Polychlorinated biphenylsd

Criteria and detection limits

Zhang and Hardy (1989) Zhang and Hardy (1989) Zhang and Hardy (1989) 515

(continued on next page)

a b c d

Criteria and detection limits Danish monitoring requirements (NERI, 2002)

EU quality criteria for the aquatic environment (Bro-Rasmussen et al., 1994)

SPMD based detection limit or lowest measurement

Ref.

Other PAD-based detection limit or measurement

e e e e e e e e 10 e

e e e e e e e e e e

0.85 0.01 0.009 0.001 0.0058 0.77 0.022 e 0.004 0.0001

Huckins et al. (1993) Rantalainen et al. (1998) Rantalainen et al. (1998) Rantalainen et al. (1998) Rantalainen et al. (1998) Rantalainen et al. (1998) Rantalainen et al. (1998)

e e e e e e e e e e

50e100

10e100

e

5

10

120

Rantalainen et al. (1998) Booij et al. (2001)

Petty et al. (2000a)

Not measured in surface water in Danish monitoring programme. Chlorpyriphos, not a field study. Quantitative correlation not determined. In Danish monitoring programme analyses are for PCB congeners 28, 31, 52, 101, 105, 118, 138, 153, 156 and 180.

Ref.

700e800

Lee and Hardy (1998)

1200

Lee and Hardy (1998)

F. Stuer-Lauridsen / Environmental Pollution 136 (2005) 503e524

PCB-52 PCB-80 PCB-79 PCB-78 PCB-81 PCB-77 PCB-126 PCB-153 PCB-169 Polybrominated diphenyl ethers Halogenated aromatic hydrocarbons Hexachlorobenzene

516

Table 4 (continued ) Compounds

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(Gustavson and Harkin, 2000). For some of the recently emerged and least water soluble environmental contaminants, the polybrominated diphenyl ethers, Booij et al. (2001) estimated SPMD based aqueous concentrations of 0.1e5 pg/l in marine water. The many applications of SPMDs and the low detection limits obtained for hydrophobic compounds with log Kow>3 points to this PAD as superior. However, the TMP and EmporeÒ disk samplers also display comparable detection limits for this type of compounds, but remains to be tested for a wider range of compounds. 6.5. PADs and quality assurance, precision and accuracy Monitoring programmes are used to evaluate the state of the environment, to judge the spatial and temporal impact of man’s activities and sometimes in response regulate the emissions of industry and households. Therefore, careful quality assurance of the data generated is an integral part of a monitoring programme. If PADs are to be part of a monitoring programme there must be the possibility to establish a quality assurance programme. The primary quality control of a PAD is carried out, as with any sampler, upon the point of sampling. When the PAD is retrieved it must be meticulously inspected for signs of punctures, holes, discolouring, or any malfunction, similar to the condition check of a manual or automated device or a collected organism. However, a more sophisticated control of the integrity and performance of a PAD can

be exercised by the addition of a ‘‘performance’’ tracer to the PAD. Such permeability reference compounds (PRCs) are used for correcting the uptake rates of the sequestered pollutants by comparing the rate of dissipation of the PRC during deployment with a known dissipation rate from SPMD determined in laboratory tests. Huckins et al. (1996); Booij et al. (1998) have added PRCs, e.g. deuterated phenanthrene suited for the analytical method, to the SPMD type sampler before deployment and estimated the effect of on-site current velocities, temperature and the impeding effect of biofouling on the device’s sampling characteristics. PRCs have also been developed for LDPE and silicone samplers (Booij et al., 2002) and the procedure can and should be used for all PADs, although different compounds may be more appropriate depending on the sampler properties and analytical considerations. The precision of PADs expressed by the relative standard deviation on measured concentrations is shown for a number of studies in Table 5. The relative standard deviation between replicates (typically nZ3e4) ranges from a few percent for hydrophobic compounds up to 50e60% for less hydrophobic compounds. For 25 organochlorine pesticides measured by SPMD in the Lower Missouri River the average coefficients of variation between five sites ranged from 19 to 26, with DDE and the isomers of hexachlorocyclohexane always contributing with high coefficients of variation (Petty et al., 1995b). After a 3e4 weeks exposure period in field studies, a typical deviation of 20e25% is found between replicates. In laboratory studies the relative standard

Table 5 Typical relative standard variations for passive accumulating devices when applied in quantitative studies Method SPMD

Parameter Concentration Concentration Concentration

Compound PAHs Dioxins and furans Organochlorine pesticides Pesticides

LPDE

Estimated concentration in water Uptake rate Uptake rate constant Uptake rate constant in field Uptake rate constant Partition coefficientd

Silicone

Partition coefficientd

PCBs and PAHs

Solvent-filled dialysis membrane bags Passive collection via gas phase TMP (estuary)

Concentration

Organotins

Permeation coefficient

Chlorinated monoaromates and nitro compounds Chlordane and dieldrin

a b c d

Concentration

RSD %

Average CV Range

PAHs Chlorophenols PCBs Pesticides PCBs and PAHs

Relative percent difference. Values from uncorrected values. Based on 95% confidence intervals. Partition to polymer in methanol/water mixture.

Reference a

Between batches Within batches Between batches Within batches

14e56 5 19e26 1e72 1.9e4.7

DeVita and Crunkilton (1998) Lebo et al. (1995) Petty et al. (1995b) Sabaliunas and So¨dergren (1997)

3e-11 10.2e28.9 0e45b

Huckins et al. (2002) Wang et al. (1998) Meadows et al. (1998)

31.4e44.8 15e25 2e6 1.9e36.7 1e19 30e50

Sabaliunas and So¨dergren (1997) Booij et al. (2002) Booij et al. (2002) Stuer-Lauridsen and Dahl (1995)

1.6e8.8

Lee and Hardy (1998)

23e49c

Peterson et al. (1995)

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deviation between replicate SPMDs is typically less than 5%. The accuracy of a method is often taken as the value found in round robin trials. Since such calibration exercises are not available in the open literature for PADs, one may instead compare results obtained with PADs with the standard spot sampling in the field. The former is only valid when constant concentrations are found in the surroundings, as in the groundwater well study by Gustavson and Harkin (2000). The authors conclude that the SPMDs detect aqueous concentrations of PAHs consistent with those found by conventional sampling although some underprediction of the two ring PAHs and overprediction of four ring PAHs are seen with the models for calculation of aqueous concentrations of PAH. Another indication of accuracy may be based on the recovery of spikes and internal standards in SPMDs. For PAHs the recovery ranged from 44% to 120% in Lebo et al. (1992). For exposures up to 64 days the lowest recovery of pyrene and benzanthracene was 48 and 46%, respectively, and the highest 76 and 74% after deployment in a stream (Lebo et al., 1996). Average recoveries for spiked PAHs in SPMDs ranged from 63 to 82% (DeVita and Crunkilton, 1998). 6.6. PADs for monitoring purposes The range of compounds detected with PADs which are shown in Table 3, the detection limits obtained or the lowest detected concentrations seen in Table 4 and the standard variations found (Table 5) suggests that PADs may find applications in monitoring programmes. To evaluate the applicability of the methods in monitoring, a number of regulatory instruments are used as references points in Tables 3 and 4: the EU water framework directive (EU, 2001), the regional conventions: the joint monitoring programme of the North-East Atlantic (OSPAR, 2002), the Helsinki Commission recommendation 19/5 on the Baltic Sea (HELCOM, 1998), and finally

as an example of national monitoring, the Danish aquatic monitoring programme (NERI, 2002). In Table 6 an assessment is given for the applicability of the PADs based on their properties relative to the above mentioned monitoring obligations and to widely acknowledged quality criteria for organic compounds in the aquatic environment: those proposed by the scientific committee of the European Commission (Bro-Rasmussen et al., 1994) and the priority toxic pollutants of the US (US EPA, 1999). Most PADs can be assessed regarding their use in a semi-quantitative approach, but at present it is the SPMDs that appear to have the widest applicability and it is the only method for which information is available on in situ quantitative estimates of water concentrations for a broader range of compounds than the two to five often structurally related compounds used in many studies. Polar compounds with LogKow values of less than 2e3 are poorly accumulated in most of the PADs, and this is a drawback when comparing to monitoring requirements and quality criteria. In these lists a range of pesticides appear with such properties because in particular the modern pesticides are designed to have semi-hydrophilic properties. However, PADs for polar compounds such as the POCIS are emerging. A special case is the very hydrophobic compounds, which share properties with very persistent and very bioaccumulative compounds. For example, the ecotoxicological assessment criteria for water given by the OSPAR (1996) for PAHs and PCBs are among the most protective quality criteria for water stipulating safe concentrations in the low ng/l range. Actual sampling and measuring of these compounds in seawater by conventional methods rarely occurs and when it does it typically involves specialised equipment and column extraction of hundreds of litres of water. However, SPMDs and PADs are well suited for collection of these and other very hydrophobic compounds in water. PADs may in fact provide an opportunity to gain information of a number of hydrophobic compounds that are

Table 6 Tentative PAD method ranked according to the percentage of organic compounds on the list, for which the method has potential for semiquantitative sampling List of priority pollutants

SPMD

Solvent-filled dialysis membrane tubing

TMPa

EmporeÒ discb

Combined for other PADsc

Water Framework Directive EU Water Quality Criteria US Water Quality Criteria OSPAR, HELCOM recommendation Danish Monitoring Programme

Good Medium

Poor Poor

Good Medium

Good Medium

Medium Poor

Good

Medium

Good

Good

Medium

Good

Poor

Good

Good

Poor

Medium

Poor

Medium

Medium

Poor

Rank: good (O75%); medium (25e75%); poor (!25%). For quantitative monitoring only SPMD is yet available. a Combined for Zabik et al. (1992); Peterson et al. (1995); Leonard et al. (2002). b Extrapolated from few laboratory and field data (van Loon et al., 1996; Kingston et al., 2000). c Including data from exposures in soil.

F. Stuer-Lauridsen / Environmental Pollution 136 (2005) 503e524

normally omitted from determination in the aqueous phase due to analytical difficulties, such as many dioxins/furans, pyrethroids, brominated flame retardants, heavier PAHs and short chained chlorinated paraffins. Criteria used for identifying compounds of environmental concern are related to properties of persistence and accumulation often associated with high octanolewater coefficients, e.g. the Persistent, Bioaccumulative and Toxic (PBT) criteria (OSPAR, 1998) or the criteria for the persistent organic pollutants of the Stockholm (POP) Convention (United Nations, 1998). PADs such as SPMDs and others based on simple accumulation in hydrophobic phases are particularly well suited for such compounds. There is an important distinction between the concentration of a hydrophobic compound measured in a conventional water sample and the result from a PAD, since in the grab sample the concentration of a compound is comprised of the truly dissolved compound, the fraction adsorbed to dissolved organic matter and, if the sample is not filtered, a third fraction bound to particles. Thus if simply extracted and analytically determined the concentration is a total of several fractions. In contrast, the concentration in a PAD is derived from the truly dissolved fraction only, which is considered to be the primary concentration available for toxicity, bioaccumulation and degradation. When the truly dissolved concentration is known from the PAD sample it is, however, straightforward to calculate the fractions sorbed, if the dissolved and particulate organic carbon concentrations are available. The usefulness of the PAD method in monitoring obviously depends on its ability to sequester pollutants of priority, but it must also be versatile and practical. In environmental air monitoring of contaminants, a passive sampler with polyurethane foam disks has already formed the backbone of a monitoring exercise on organochlorines across Europe (Jaward et al., 2004). In terms of being practical, reproducible and accurate SPMDs have been shown to fulfil the criteria for successful application in an aquatic monitoring programme. Emerging now are EmporeÒ disk methods and SPMEs. Both methods are considered as possible ESDs for aqueous samples and due to a relatively simple handling, clean up and analysis holds promise for in situ applications. The SPME may be applied as a linear uptake PAD for environmental samples in situ with simple modifications to prolong the linear uptake. Such an SPME device for exposure in air has recently been developed by Martos and Pawliszyn (1999); Gorlo et al. (1999). If one wants to avoid conventional sampling completely a sampler for polar compounds and one for metals will also be needed, but the PADs should be seen as complimentary to existing approaches, mostly equally good and sometimes even better. The use of PADs in a monitoring context is a feasible option and the choice between PADs and

519

conventional sampling should be governed by the requirements of the programme.

7. Conclusions The last 15 years has seen the development of a number of PADs, including some that allow for the detection and screening of a broad range of organic micropollutants in the aquatic environment. All the published PADs can be applied for qualitative screening of compounds in the dissolved phase of the hydrosphere, but for a few PADs it is also possible to calculate the concentration of the compound in water from the amount of compound detected in the PAD. The SPMD, the EmporeÒ disk method and the TMP sampler are the PADs that may qualify for application in aquatic monitoring programmes. The most thoroughly investigated design, the SPMD, has been applied to a broad range of organic micropollutants and has been shown to quantitatively estimate aqueous concentrations for compounds with Log Kow values of 3e6. The accuracy and precision of SPMDs estimates of Cw is adequate for monitoring of the majority of organic micropollutants in monitoring programmes and often better than typical chemical analysis for very strongly hydrophobic compounds. It is worth noticing that in the development of monitoring programmes such compounds (dioxins/ furans, PCBs, brominated diphenyl ethers and pyrethroid pesticides) are typically omitted from determination in the aqueous phase due to considerations of the detection limits achievable by conventional chemical analysis. The polar compounds, such as many pesticides and pharmaceuticals are presently poorly covered by PAD applications, but some recently developed PADs (POCIS, EmporeÒ disk) holds promise in improving the coverage of compounds with polar properties. The environmental conditions of flow, temperature, and biofouling do affect the accumulation of a compound in the sampler, but the effects can be monitored and corrected by monitoring the recovery of a permeability reference compound. The use of PRCs should be included in all field exposures of PADs. In many aquatic systems the concentrations of the compounds in question are not constant, but may fluctuate or occur in more or less pronounced pulses. The resulting concentrations in the PADs are time weighted averages of the exposure period, and little research has yet been directed to verify the theoretical quantitative expression of pulsed or intermittent exposure and the uptake pattern in aqueous PADs. Such information will allow users of PADs to present regulators with realistic scenarios for the environmental concentration of a certain compound when PADs are applied in aquatic monitoring programmes or used to measure aqueous contaminant concentrations against

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quality criteria. The PADs, in particular the SPMD, can be used to monitor the O75% of the organic micropollutants that are listed in water quality criteria of the EU and the US, the EU water framework directive, the recommendations of OSPAR and the Danish national monitoring programme.

study the applicability of PADs in the Danish Aquatic Monitoring Programme. Thanks are due to Mr Morten Birkved for tracking down and organising many PAD papers, and to two anonymous reviewers for constructive criticism of a previous version.

Acknowledgements

Appendix A.

The work that initiated this review was originally contracted from Dr Alf Aagaard of the Danish EPA to

All equations in Table A.1 apply to conditions of linear uptake.

Table A.1 Equations used for calculation of ambient concentrations of organic micropollutants based on concentration in PADs PAD method

Target compounds

Equation for prediction of concentration in aqueous matrix

Explanation

Reference

Semipermeable membrane devices (SPMD)

Lipophilic

CSPMD !VSPMD CW Z Rs !t

Huckins et al. (1996)

Solvent-filled dialysis membranes (SF)

PCBs

CL ZCW !a!ð1  ebt Þ

EmporeÒ disks (ED)

Lipophilic

k1 Cd;t ZCa ! !ð1  ek2 t Þ k2

Passive sampler, TMMPS

Chlordane and dieldrin

CPAD ðtÞZKPAD ðtÞ!CWater

Permeation sampling

Phenolic compounds

MZK!C!t

CSPMD, concentration in the SPMD at time t; CW, concentration in the aqueous phase; VSPMD, volume of SPMD; Rs, SPMD sampling rate (value for analyte determined in laboratory experiment); t, time CL, concentration in the lipid; CW, concentration in the water; a, statistical estimate of k1 and k2 ratio, determined using octanole water partition coefficient as seed value; b, statistical estimate of k2 as above, DAm/Im/VL Ca, concentration in the aqueous phase; k1 and k2, uptake and elimination rates; Kd, k1/k2, partition coefficient disk to water, calculated using empirical equation based on the octanolewater partition coefficient CPAD(t), concentration of analyte in PAD at time t; KPAD(t), partition coefficient PAD to water at time t (linear relationship) for analyte determined in laboratory experiment); CWater, concentration of analyte in water M, mass collected; C, external concentration; K, permeation constant (determined); t, exposure time

Johnson (1991)

Verbruggen et al. (1999)

Peterson et al. (1995)

Zhang and Hardy (1989)

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F. Stuer-Lauridsen / Environmental Pollution 136 (2005) 503e524 Table A.1 (continued) PAD method

Target compounds

Groundwater monitoring

PAHs

Equation for prediction of concentration in aqueous matrix   1 Cg ! C  FZZ w Z g w H CD !H D e

g

ðm=EÞ!L D!A!t

Passive dosimetry

Altrazine and p-xylene

CZ

Passive sampler

Organic compounds

MD ZCW !RWD !t

Passive sampler for volatile aromatics in water

Monocyclic aromatics

CZ

m K!t

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Explanation

Reference

F, steady state flux; Zwandg, film thickness of water saturated membrane and air film, resp.; Deandg, effective diffusion coefficients in membrane and air, resp.; Cwandg, concentration of analyte in water and air; H, Henry’s law constant C, contaminant concentration; m, mass collected; E, extraction efficiency; D, molecular diffusivity (calculated with standard techniques); L, length of diffusion pathway; A, area of diffusion pathway; t, exposure time MD, mass of analyte in receiving phase (at time t); CW, concentration of analyte in water; RWD, sampling rate (determined as accumulation factor at time t); t, time C, external concentration; m, mass collected; K, permeation constant (determined); t, exposure time

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