Semipermeable Membrane Devices As Integrative Tools for Monitoring Nonpolar Aromatic Compounds in Air

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Semipermeable membrane devices as integrative tools for monitoring nonpolar aromatic compounds in air

Hanna Söderström

Akademisk avhandling

Som med tillstånd av rektorsämbetet vid Umeå universitet för erhållande av Filosofie Doktorsexamen vid Teknisk-naturvetenskapliga fakulteten i Umeå, framlägges till offentlig granskning vid Kemiska Institutionen, hörsal KB3B1 i KBC, fredagen den 10 december, 2004, kl. 10.00.

Fakultetsopponent: Associate Professor Jochen F. Müller, National Research Centre of Environmental Toxicology (EnTox), University of Queensland, Australia.

Cover illustration by Leif Lindgren, Älvsbyn

Environmental Chemistry Department of Chemistry Umeå University SE-901 87 Umeå Sweden

ISBN 91-7305-782-7

Printed in Sweden by VMC, KBC, Umeå University, Umeå 2004.

Semipermeable membrane devices as integrative tools for monitoring nonpolar aromatic compounds in air

Hanna Söderström, Environmental Chemistry, Umeå University, Umeå, Sweden.

Abstract: Air pollutants pose a high risk for humans, and the environment, and this pollution is one of the major environmental problems facing modern society. Active air sampling is the technique that has been traditionally used to monitor nonpolar aromatic air pollutants. However, active high volume samplers (HiVols) require a power supply, maintenance and specialist operators, and the equipment is often expensive. Thus, there is a need to develop new, less complicated sampling techniques that can increase the monitoring frequency, the geographical distribution of the measurements, and the number of sites used in air monitoring programs. In the work underlying this thesis, the use of semipermeable membrane devices (SPMDs) as tools for monitoring gas phase concentrations of nonpolar aromatic compound was evaluated using the compound classes polychlorinated biphenyls (PCBs), polycyclic aromatic hydrocarbons (PAHs), alkylated PAHs (alkyl-PAHs) and nitrated PAHs (nitro-PAHs) as test compounds.

High wind-speeds increased the uptake and release in SPMDs of PAHs and PCBs with log KOA values > 7.9, demonstrating that the uptake of most nonpolar aromatic compounds is controlled by the boundary layer at the membrane-air interface. The use of a metal umbrella to shelter the SPMDs decreased the uptake of PAHs and PCBs by 38 and 55 percent, respectively, at high wind/turbulence, and thus reduced the wind effect. Further, the use of performance reference compounds (PRCs) to assess the site effect of wind on the uptake in SPMDs reduced the between-site differences to less than 50 percent from as much as three times differences in uptake of PCBs and PAHs. However, analytical interferences reduced the precision of some PRCs, showing the importance of using robust analytical quality control.

SPMDs were shown to be efficient samplers of gas phase nonpolar aromatic compounds, and were able to determine local, continental and indoor spatial distributions of PAHs, alkyl- PAHs and nitro-PAHs. In addition, the use of the SPMDs, which do not require electricity, made sampling possible at remote/rural areas where the infrastructure was limited. SPMDs were also used to determine the source of PAH pollution, and different approaches were discussed. Finally, SPMDs were used to estimate the importance of the gas phase exposure route to the uptake of PAHs in plants. The results demonstrate that SPMDs have several advantages compared with HiVols, including integrative capacity over long times, reduced costs, and no need of special operators, maintenance or power supply for sampling. However, calibration data of SPMDs in air are limited, and spatial differences are often only semi-quantitatively determined by comparing amounts and profiles in the SPMDs, which have limited their use in air monitoring programs. In future work, it is therefore important that SPMDs are properly sheltered, PRCs are used in the sampling protocols, and that calibrated sampling rate data, or the SPMD-air partition data, of specific compounds are further developed to make determination of time weighted average (TWA) concentrations possible.

Keywords: air pollution, alkyl-PAHs, atmosphere, bioavailability, boundary layer, diffusive sampling, emissions, integrative, membrane, monitoring, nitro-PAHs, PAHs, particles, passive samplers, PCBs, plants, PRCs, release rate, sampler design, sources, SPMDs, traffic, uptake rate, wind effect, wood burning.

ISBN 91-7305-782-7

Semipermeabla membraner som integrerande mätverktyg vid miljöövervakning av opolära aromatiska föreningar i luft

Hanna Söderström, Miljökemi, Umeå Universitet, Umeå, Sverige.

Sammanfattning: Luftföroreningar utgör en stor risk för människa och miljö, och är ett av de största miljöproblemen i vårt samhälle. Traditionellt har aktiv luftprovtagning används vid miljöövervakning av opolära aromatiska luftföroreningar. Aktiv högvolymprovtagning kräver dock elektricitet, underhåll och operatörer med specialistkompetens, och utrustningen är ofta dyr. Det finns därför ett behov av nya, mindre komplicerade provtagningstekniker som kan öka mätfrekvensen, den geografiska spridningen samt antalet provtagningsplatser som används i miljöövervakningsprogram. I det arbete som presenteras i denna avhandling undersöktes användningen av semipermeabla membraner (SPMDer) som mätverktyg av opolära aromatiska föreningar i gasfas. Ämnesklasserna polyklorerade bifenyler (PCBer), polycykliska aromatiska kolväten (PAHer), alkylerade PAHer (alkyl-PAHer) och nitrerade PAHer (PAHer) användes som testsubstanser.

Höga vindhastigheter ökade upptaget samt avgången av PAHer och PCBer med log KOA värden > 7.9, vilket visar att upptaget i SPMDer av de flesta opolära aromatiska föreningar kontrolleras av diffusionslagret som bildas vid gränsytan mellan SPMD och luft. Då ett metallparaply användes för att skydda de SPMDerna minskades PAH- och PCB-upptaget med 38 samt 55 procent, vilket visar att metallparaplyet reducerade effekten av vind. Användningen av sk performance reference compounds (PRCs) för att bedömma effekten av vind på upptaget i SPMDer reducerade variationen mellan SPMDer som utsatts för olika vindhastigheter till mindre än 50 procent från upp till tre gångers skillnad i PAH- och PCB- upptaget. Analytiska interferenser reducerade dock precisionen hos några PRCs vilket visar vikten av att använda robust analytisk kvalitetskontroll.

SPMDer visade sig vara effektiva provtagare av opolära aromatiska föreningar i gasfas och kunde användas för att bestämma PAHers, alkyl-PAHers samt nitro-PAHers spridning i luft, lokal, kontinental samt inomhus. Eftersom SPMDer ej kräver elektricitet möjliggjordes även provtagning på otillgängliga plaster där infrastrukturen var begränsad. SPMDer kunde även användas för att bestämma källor till PAH-förorening i luft, och olika tillvägagångssätt för detta ändamål diskuterades. Slutligen användes SPMDer för att bedömma betydelsen av gasfasexponeringen för PAH-upptaget i växter. Dessa resultat visar att SPMDer har många fördelar i jämförelse med högvolymprovtagare som integrerande kapacitet under långa tidsperioder, reducerade kostnader, och ej krav på operatörer med specialistkompetens, elektricitet och underhåll vid provtagning. Tillgängliga kalibreringsdata för SPMDer i luft är dock begränsande, och haltskillnader bestäms oftast endast semikvantitativt genom att jämföra halter och profiler funna i SPMDer, vilket har begränsat deras användning i miljöövervakningsprogram. I framtida arbeten är det därför viktigt att de SPMDerna är ordentligt skyddade, att PRCs används i provtagningsprotokollen, och att fler kalibrerade provtagningshastigheter, eller SPMD-luftfördelningsdata för specifika föreningar tas fram så att tidsvägsmedelkoncentrationer kan bestämmas.

Nyckelord: alkyl-PAHer, atmosfär, avgångshastighet, biotillgänglighet, diffusionslager, diffusionsprovtagning, emissioner, integrerande, källor, luftförorening, membran, miljöövervakning, nitro-PAHer, PAHer, partiklar, passiva provtagare, PCBer, PRCs, provtagardesign, SPMDer, trafik, upptagshastigheter, vindeffekt, vedeldning, växter.

ISBN 91-7305-782-7

Det krävs ett helt nytt sätt att tänka för att lösa de problem vi skapat med det gamla sättet att tänka.

Today´s problems cannot be solved with the same thinking that created them.

Albert Einstein

LIST OF PAPERS

This thesis is based on the following Papers, which are referred to in the text by their roman numbers as listed below.

I. Söderström H.S., and Bergqvist P-A., 2003. Polycyclic aromatic hydrocarbons in a semiaquatic plant and semipermeable membrane devices exposed to air in Thailand. Environmental Science and Technology. 37, 47-52.

II. Söderström H.S., and Bergqvist P-A. Wind effects on passive air sampling of PAHs and PCBs. Accepted for publication in Bulletin of Environmental Contamination and Toxicology.

III. Söderström H.S., and Bergqvist P-A., 2004. Passive air sampling using semipermeable membrane devices at different wind-speeds in situ calibrated by performance reference compounds. Environmental Science and Technology. 38, 4828-4834.

IV. Söderström H.S., Hajšlová J., Siegmund B., Kocan A., Obiedzinski M.W., Tysklind M., and Bergqvist P-A. PAHs and nitrated PAHs in air of five European countries determined using SPMDs as passive samplers. Accepted for publication in Atmospheric Environment.

V. Strandberg B., Söderström H.S, Gustafson P., Sällsten G., Bergqvist P-A., and Barregård L., 2004. Semipermeable membrane devices as passive samplers to determine polycyclic hydrocarbons (PAHs) in indoor environments. Organohalogen Compounds. 66, 44-49.

All papers are reproduced with the kind permission of the publishers of the respective journals.

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ABBREVIATIONS alkyl-PAH Alkylated PAH AU Austria 13C Carbon 13-labelled compound CCC Chemical coordinating center CERC Columbia environmental research center CO Carbon monoxide CV Coefficient of variance CZ Czech Republic d Day (24 hours) DDT Dichlorodiphenyltrichloroethane dw Dry weight EAFs Exposure adjustment factors EMEP European monitoring and evaluation program EPA Environmental protection agency FC Field control GC/MS Gas chromatography/mass spectrometry GPC Gel permeation chromatography H Henry’s law constant 2H Deuterated compound HCB Hexachlorobenzene HCH Hexachlorocyclohexane HiVols Active high volume sampler HR High-resolution

KAW Air-water partition coefficient ke Exchange rate constant KOA Octanol-air partition coefficient KOW Octanol-water partition coefficient KSA SPMD-air partition coefficient KSPMD SPMD equilibrium partition coefficient (also, SPMD-water partition coefficient) LB Laboratory blank LDPE Low-density polyethylene LLE Liquid-liquid extraction LR Low-resolution LRTAP Long range transboundary air pollution MeVols Active medium volume sampler nitro-PAH Nitrated PAH 1-NN 1-nitronaphthalene 2-NN 2-nitronaphthalene

NO2 Nitrogen dioxide OC Organochlorine pesticide PAH Polycyclic aromatic hydrocarbon PAS Passive air samplers PBDE Polybrominated diphenyl ether PCA Principal component analysis PCB Polychlorinated biphenyl PCDD Polychlorinated dioxin PCDF Polychlorinated furan PM Particulate matter POGs Polymer-coated glass sampler POP Persistent organic pollutant PRC Performance reference compound PUF Polyurethane foam QA Quality assurance QC Quality control

RS Sampling rate SE Sweden SK Slovakia

SO2 Sulphur dioxide SPE Solid phase extraction S-PL Poland summer 2000 SPMD Semipermeable membrane device SPME Solid phase microextraction SVOC Semi volatile organic compound TWA Time weighted average USGS United States geological survey’s VOC Volatile organic compound ww Wet weight W-PL Poland winter 1999

TABLE OF CONTENTS

1. INTRODUCTION 1 1.1 Aims and scope 3 2. PASSIVE AIR SAMPLING 5 2.1 Historical use of PAS as monitoring tools 7 2.1.1 Passive air sampling of nonpolar aromatic compounds 8 2.2 Passive sampling using semipermeable membrane devices 9 2.2.1 Sampling principles and factors influencing the sampling rate 12 2.2.2 Target compounds 15 2.2.3 Fields of application 16 2.2.4 Theory and mathematical models of SPMD sampling 18 2.2.4.1 The PRC model 21 3. COMPOUNDS SAMPLED AND THEIR PHYSICAL/CHEMICAL PROPERTIES 23 3.1 Polycyclic aromatic hydrocarbons (PAHs) 23 3.2 Polychlorinated biphenyls (PCBs) 24 4. SAMPLING AND CHEMICAL ANALYSIS 27 4.1 Design of the sampler 27 4.2 Sampling procedure and approach 28 4.2.1 The PRC approach 28 4.2.2 Cleaning of the sampling equipment 29 4.3 Extraction, cleanup and analysis 29 4.4 Quality assurance and quality control (QA/QC) 30 4.5 Data interpretation 32 4.5.1 Semi-quantitative and quantitative approach 32 4.5.2 Multivariate data analysis using principal component analysis (PCA) 33 5. FACTORS INFLUENCING THE UPTAKE OF COMPOUNDS IN SPMDs FROM AIR 35 5.1 Physical/chemical properties of the compounds sampled 36 5.1.1 RS of PCBs in air 37 5.2 Environmental variables 38 5.2.1 Wind-speed/turbulence 39 5.2.1.1 The wind effect test method 40 5.2.1.2 Wind effect at high wind-speeds 42 5.2.1.3 Wind effect in field 44 5.2.2 Temperature 44 5.2.3 Particles 45 5.2.4 UV sunlight 47 6. APPROACHES TO REDUCE THE SITE EFFECTS OF ENVIRONMENTAL VARIABLES 49 6.1 The use of shelter 49 6.1.1 Reduction of the wind effect by shelter 50

6.2 The use of performance reference compounds (PRCs) 52 6.2.1 PRCs´ utility for assessing wind effect 52 6.2.2 PRCs´ utility for assessing site effects in air 57 7. FIELDS OF APPLICATION AND SAMPLING STATEGIES 59 7.1 Determination of local/regional atmospheric distribution 59 7.2 Determination of continental atmospheric distribution 61 7.3 Determination of pollution sources 65 7.3.1 The use of individual compound ratios 65 7.3.1.1 Considerations involved with the use of the individual compound ratios 67 7.3.2 The use of total PAH-patterns 68 7.3.3 Determination of residential wood burning as a source of indoor PAH exposure 69 7.4 Estimation of the consequences of gas phase PAH exposure for plants 72 8. CONCLUSIONS AND FUTURE RESEARCH POSSIBILITIES 77 9. ACKNOWLEDGEMENTS 79 REFERENCES 81

1. Introduction

1. INTRODUCTION

Air pollutants pose a high risk for humans, and the environment, and are one of the major environmental problems facing modern society. For instance, relationships between exposure to outdoor air pollutants and daily mortality, morbidity, hospital admissions, and lung function changes have been found in many studies (e.g., Forsberg et al., 1998; Aga et al., 2003; Sunyer et al., 2003; Zanobetti et al., 2003; Chen et al., 2004). Anthropogenic air pollution has occurred since ancient times, but increased enormously in the 20th century following the industrial revolution in Europe, and continued to increase significantly until the 1970s, due to increases in population, industrialization, urbanization and traffic (Figure 1). Today, many countries have reduced their largest local point sources, such as emissions from industry, while the diffusive sources such as traffic continue to increase as a result of our modern living habits. Air pollution has also become a global problem since many of the air pollutants are transported over long ranges with atmospheric and ocean currents. Thus, even if a country reduces its major emissions, air pollution can continue to increase. In order to control the air quality, and thus the human exposure through air, air monitoring is highly important.

Figure 1. Illustration of the continuous pollution of the atmosphere due to our modern living habits which has become one of the most serious environmental problems facing today’s society.

Air monitoring on a national level is often regulated by international conventions. In Europe, one of the major monitoring programs is the European Monitoring and Evaluation Program (EMEP). This monitoring program was established after the convention on Long Range Transboundary Air Pollution (LRTAP) was signed in 1979 (Geneva, November 1979). The main objective of EMEP is to regularly provide Governments and subsidiary

1. Introduction bodies under the LRTAP convention with useful research results for the development and evaluation of international protocols on the reduction of atmospheric emissions that are discussed and detailed within the convention. The EMEP has three main purposes; to collect emission data, to measure air and precipitation quality, and to model atmospheric transport and deposition. The focus are on acidifying and eutrophying compounds, ground level ozone, heavy metals, persistent organic pollutants (POPs) or semi volatile organic compounds (SVOCs) (classed in this thesis as nonpolar aromatic compounds), volatile organic compounds (VOCs) and particulate matter (PM). The Chemical Coordinating Centre (CCC) co-ordinates the EMEP measurements of chemical air quality and precipitation. In 2002, six countries measured POPs in air and aerosols in a total of eight sites (Aas and Breivik, 2004). Both the distribution of the sites across Europe and the number of sites at which POPs were measured in this monitoring program were considered by the CCC as insufficient (Aas and Breivik, 2004). Thus, the real spatial variability in the atmospheric concentrations of POPs across Europe remains rather unknown.

The standard methods for air sampling of nonpolar aromatic compounds like pesticides, polychlorinated biphenyls (PCBs) and polycyclic aromatic hydrocarbons (PAHs) recommended by CCC/EMEP are active high/medium volume samplers (HiVols/MeVols) including a glass fibre filter (which collects particle-bound compounds) followed by two polyurethane foam (PUF) plugs or XAD-2 (which retain gaseous compounds) coupled to a pump (Anonymous, 2001). Active air sampling is the technique that has been traditionally used in air monitoring. The disadvantages with active volume samplers are that they require power and maintenance, the sampling equipment is often both expensive and complicated, and they have to be operated by specially trained personnel. The high cost of the sampling equipment, the power demands and the requirement for maintenance by specialist operators restrict the number and location of the sites that can be monitored. Furthermore, active sampling is seldom integrative for more than 48 h (see, for instance, reported sampling frequencies by Aas and Breivik (2004)), which makes it difficult to measure episodic pollutant releases and time weighted average (TWA) concentrations. Finally, in active sampling the total amounts of air pollutants (of both the gas phase and the particle- associated compounds) are collected although one or the other of these phases may pose little or no primary risk to living organisms and humans. Thus, there is a need to develop new, less complicated sampling techniques that can increase the monitoring frequency, the geographical distributions of

1. Introduction the measurements, and the number of sites used in air monitoring programs of POPs.

1.1 Aims and scope

In the work described in this thesis, the use of semipermeable membrane devices (SPMDs) as tools for monitoring the gas phase concentrations of nonpolar aromatic compounds in air was evaluated. The work focused on the nonpolar aromatic compounds PCBs, PAHs, alkylated PAHs (alkyl-PAHs) and nitrated PAHs (nitro-PAHs). The main objective was to further develop the SPMDs as passive air samplers (PAS) to provide an alternative sampling method to the conventionally used active volume samplers. Briefly, this thesis discusses the uptake and release rates of the gas phase of nonpolar aromatic compounds in SPMDs, the factors influencing the SPMD sampling rate and the performance of SPMDs in different fields of application.

More specifically, the work described in Papers II and III tested the hypothesis that high wind-speeds/turbulence increase the uptake in SPMDs. Further, the use of performance reference compounds (PRCs) to calibrate the SPMD sampling rate in situ and thus assess site effects at different wind- speeds/turbulence was tested (Papers II and III). Paper III also tested the hypothesis that shelters protecting the SPMDs reduced the wind/turbulence exposure, and thus decreased the uptake and release rates of target analytes.

In unpublished work, the RS values of a number of PCBs for SPMDs in air were estimated by comparing SPMD samplers with conventionally used MeVols. Finally, in Papers I, IV and V the scope for using SPMDs to determine the local, continental and indoor spatial distribution, respectively, of PAHs and some of their substituted derivates, was examined. The work described in Papers I, IV and V also aimed to determine potential point sources of PAHs like traffic and residential heating systems, and to investigate different strategies to detect the specific emission sources. The consequence of gas phase exposure, sampled by SPMDs, on the uptake of PAHs in a semiaquatic plant were also investigated (Paper I).

2. Passive air sampling

2. PASSIVE AIR SAMPLING

Passive air sampling is a sampling method that has advantages in spatial surveys because of its low expense, simplicity, long-term operation and the fact that it does not require electricity. The general sampling principles of PAS and their historical use as monitoring tools are outlined below. In addition, the use of SPMDs as passive samplers is discussed in detail.

The definition of a passive or diffusive sampler is a “device which is capable of taking samples of gases or vapours from the atmosphere at a rate controlled by a physical process such as gaseous diffusion through a static air layer or a porous material and/or permeation through a membrane, but which does not involve the active movement of air through the device” (Anonymous, 2002a). Active sampling of air is normally achieved by pumps. In European standards, the mass of an analyte taken up by a diffusive sampler is described by (Anonymous, 2003)

mS=A×D×(ρ1-ρ2)×t Eq. 1 l

where mS is the mass of an analyte which is sorbed by diffusion in pg, A is the cross-sectional area of the diffusion path or the equivalent sorption surface in 2 2 -1 cm , D is the diffusion coefficient of the analyte in cm ⋅ min , ρ1 is the concentration of the analyte at the beginning of the diffusive layer in µg ⋅ m-3,

ρ2 is the concentration of the analyte at the end of the diffusive layer in µg ⋅ m-3, t is the exposure time in min, and l is the length of the static air layer in the sampler or equivalent for permeation types, in cm (Figure 2).

Ideally, ρ1 is equal to the air concentration of the analyte outside the diffusive sampler and ρ2 is reduced to zero due to sorption or chemical reactions in the sorbent. A ρ1 close to or equal to zero (“zero sink”-condition) serves as the driving force of the diffusion across l.

2. Passive air sampling

A l

B ρ ρ2 ρ 1 Figure 2. Schematic diagram of diffusive sampling and parameters influencing the process (modified from Anonymous, 2003).

The diffusive uptake rate, U, in cm2 ⋅ min-1, is obtained from knowledge of either the physical parameters of the diffusion barrier (Anonymous 2002b) according to

U = ms =A×D Eq. 2 ρ×t l

or by comparison with another analyte for which the diffusive uptake rate, U2, and the diffusion coefficient of the analyte, D2, is known (Anonymous, 2002b) according to

U1=D1×U2 Eq. 3 D2

The major advantages of PAS over active samplers are their low expense and simplicity, since neither a power supply nor specialist operators are required for routine monitoring, and they can operate for long time periods. Thus, PAS can facilitate sampling campaigns at multiple sites simultaneously on local, regional or global levels. They may also be the only option when sampling is required in remote locations, such as natural reserves and mountain areas where there is no electricity supply. The mass uptake of many PAS is linear and integrative during the entire exposure period and TWA concentrations over long time periods can be assessed. Integrative diffusive sampling has several advantages. In the integrative sampling approach episodic pollutant releases are detected to a higher degree than during active grab sampling. TWA concentrations are also more relevant in exposure assessments because they reflect the total exposure during a given time period. Passive sampling is also preferable to active low volume sampling

2. Passive air sampling when the ambient air concentrations are low because integrative sampling concentrates compounds present in trace or ultra-trace levels to levels above detection limits. However, POPs bound to ultra-fine particles (diameter < 1 µm), which can have toxic effects on the respiratory organs of humans and wildlife, are not measured with PAS. In addition, the sampling rate of PAS is affected by the sampling conditions, especially factors like wind- speed/turbulence and temperature, and it is important to ensure that the effect of these parameters is reduced as much as possible and calibrated.

2.1 Historical use of PAS as monitoring tools

Passive sampling was employed for the first time in 1927 (Gorecki and Namiesnik, 2002). The cited passive sampler was used to semi-quantitatively determine the atmospheric concentration of carbon monoxide (CO) (Gorecki and Namiesnik, 2002). However, there was a delay of almost 50 years (until 1973) before diffusive sampling was described in mathematical terms, sampling rates were calibrated, and two types of passive samplers could be used to quantitatively determine the air concentration of nitrogen dioxide

(NO2) and sulphur dioxide (SO2) (Gorecki and Namiesnik, 2002).

Since then, these and new types of PAS have been further developed (e.g., Berglund et al., 1992; Cao and Hewitt, 1993; Krochmal and Kalina, 1997; Wideqvist et al., 2003). To date, PAS have been routinely employed in occupational exposure assessments and in ambient, outdoor air monitoring of a variety of volatile organic compounds (VOCs) like NO2, SO2, ozone, and benzene (see, for example, Sällsten et al., 2001; Stevenson et al., 2001; Bytnerowicz et al., 2002; Modig et al., 2002; Kruså et al., 2004). For instance, the Swedish environmental protection agency has assessed the ambient human exposure of NO2, benzene, butadiene, formaldehyde and acetaldehyde by using personal passive sampling (an approach in which the sampler is carried by people judged to be at risk of exposure) (Sällsten et al., 2001; Modig et al., 2002; Kruså et al., 2004). In total, 40-60 persons were randomly selected from the populations of each of three cities Gothenburg (2000), Umeå (2001) and Stockholm (2002 and 2003), and integrative personal passive sampling was performed for a week in each case (Sällsten et al., 2001; Modig et al., 2002; Kruså et al., 2004).

2. Passive air sampling

2.1.1 Passive air sampling of nonpolar aromatic compounds

The local and continental atmospheric distribution of nonpolar aromatic compounds was first monitored by collecting different plant materials such as tree bark (e.g., Simonich and Hites, 1995), leaves and grass (e.g., Müller et al., 2001), pine needles (e.g., Tremolada et al., 1996; Kylin and Sjödin, 2003; Hellström et al., 2004; Lehndorff and Schwark, 2004) and liches (e.g., Ockenden et al., 1998b). In biomonitoring-based air sampling, the spatial pollution distribution in the plant tissues and the atmosphere are assumed to be closely related. Kylin and Sjödin (2003), for instance, analysed pine needle wax collected at various sites from the central to northern parts of Europe, to monitor the spatial distribution of hexachlorocyclohexanes (HCH) and dichlorodiphenyltrichloroethane (DDT). However, the concentration capacity and sampling rate of plant tissues vary with the plant species and age, the sampling location and season (as shown, for example, by Ockenden et al., 1998b; Müller et al., 2001; Kylin and Sjödin, 2003; Hellström et al., 2004; Lehndorff and Schwark, 2004). For instance, Müller et al. (2001) found interspecies differences between leaves and grasses that were attributed due to differences in the size of the lipophilic compartments and the ratio between the surface area and the volume of the plant species. Ockenden et al. (1998b) also found that nonpolar aromatic compounds were accumulated to a higher degree in lichen than in pine needles. Thus, these kinds of plant data were subject to several uncertainties, which complicate their interpretation and limit the potential of plants as monitoring tools.

Thereafter, monitoring of nonpolar aromatic compounds has been achieved by man-made PAS. Even though man-made PAS have to be deployed and the samplers are more costly, they are preferred over plants because their uniform sampler design facilitates both comparisons between sites and control of the sampling time. Most work has focused on the use of integrative (from weeks up to years) PAS with high concentration capacities for POPs such as SPMDs (e.g., Ockenden et al., 1998a; Ockenden et al., 2001; Lohmann et al., 2001; Meijer et al., 2003; Bartkow et al., 2004), polyurethane foam (PUF) disks (e.g., Shoeib and Harner, 2002; Jaward et al., 2004a; 2004b), XAD-2 resin-based passive samplers (e.g., Wania et al., 2003), polymer-coated glass samplers (POGs) (e.g., Harner et al., 2003) and different leave- resembling samplers with lipophilic coated surfaces like, for example tristerin- coated fiberglass sheets (Müller et al., 2000). Monitoring of the continental atmospheric distribution of nonpolar aromatic compounds using PAS has

2. Passive air sampling only recently been achieved. In 2004, for instance, PUF disks were used to monitor the spatial atmospheric distribution of PCBs, polybrominated diphenyl ethers (PBDEs), organochlorine pesticides (OC), polychlorinated naphthalenes and PAHs simultaneously in 22 countries across Europe (Jaward et al., 2004a; 2004b). Furthermore, the spatial variation in PCBs and hexachlorobenzene (HCB) in Norway and the United Kingdom (UK) was measured by the deployment of SPMDs at 10 and 11 sites for two years between 1994 and 1996 (Ockenden et al., 1998c), and 1998 and 2000 (Meijer et al., 2003), respectively.

2.2 Passive sampling using semipermeable membrane devices (SPMDs)

Semipermeable membrane devices (SPMDs) were developed by scientists at the United States Geological Survey’s (USGS), Columbia environmental research center (CERC) as integrative, passive, in situ samplers of the dissolved fraction of moderately to highly hydrophobic organic compounds in water (US Patents, 5,098,573, Huckins et al., 1992; 5,395,426, Huckins et al., 1995). Thereafter, Petty et al. (1993) demonstrated in a laboratory study that SPMDs are also highly efficient passive samplers of gaseous nonpolar organic pollutants in air. In a study by Prest et al. (1995) the application of SPMDs as PAS in field was demonstrated.

The SPMD was developed to mimic the bioconcentration process of hydrophobic organic pollutants including diffusion through a cell membrane and accumulation into the lipid tissues of an organism. Thus, SPMDs can be viewed as an intermediate between conventional sampling techniques and biomonitoring methods (collecting aquatic organisms) which have traditionally been used to assess environmental exposure (Figure 3). In exposure assessments, the identification of new and existing pollutants, and the determination of TWA pollutant concentrations (and thus estimation of the environmental exposure over time), are fundamental steps. The SPMD technique has the advantages over traditionally used grab water sampling methods like liquid-liquid extraction (LLE), solid phase microextraction (SPME) and solid phase extraction (SPE), that these steps can be facilitated without multiple samples having to be collected over time. SPMDs also facilitate the detection of trace and ultra-trace levels of POPs, because the integrative sampling approach can pre-concentrate the pollutants in situ to levels above detection limits. Utvik et al. (1999), for instance, compared the

2. Passive air sampling five sampling methods based on SPMDs, mussels, an in situ large volume SPE sampler, a SPE disk and LLE, respectively, for monitoring PAH concentrations in seawater (in this study the North Sea) and SPMDs and mussels were found to be the most appropriate approaches for analysing PAHs in seawater.

Figure 3. Picture of a standard 1-mL triolein SPMD placed in water.

A standard SPMD, used for sampling, consists of a 91.4 cm long, 2.5 cm wide and 75-90 µm thick (wall thickness) layflat non-porous tube (tube with no fixed pores, only transient cavities with a diameter ∼ 10 Å) of low-density polyethylene (LDPE) filled with 1 mL (0.915 g) of > 95 % pure triolein (1,2,3- tri[cis-9-octadecenoyl] glycerol) (Figure 4). A standard 1-mL SPMD has a total mass of 4.55 g with a 4:1 LDPE membrane to triolein (w/w) ratio and a membrane surface-area to triolein volume ratio of ∼ 450 cm2 ⋅ mL-1. The ends of the LDPE tube have loops to facilitate its deployment during sampling (Figure 4). In early work, the use of low-molecular-weight dialysis cellulose membrane filled with organic solvent (hexane) was evaluated (e.g., Södergren, 1987; Herve et al., 1991; Johnson, 1991). Although these devices had several drawbacks, field studies showed that the in situ passive sampling approach had potential (Södergren, 1987; Herve et al., 1991; Johnson, 1991). In further research, the use of the nonporous polymers LDPE, polypropylene, polyvinyl chloride, polyacetate and silicone or silastic as membranes were evaluated according to a number of criteria including, for instance, analyte uptake rate, dialytic performance and cost (Huckins et al., 2002b). Of the polymers tested only LDPE and polypropylene were found to be acceptable (Huckins et al., 2002b). LDPE tubes were chosen as membranes because their transient cavities have similar sizes (∼ 10 Å) to the size limits for particles that can be transported through biomembranes (9.8 Å) (Opperhuizen et al., 1985). Thus,

2. Passive air sampling only the truly dissolved organic pollutants with molecular masses < 600 approximately Daltons can diffuse into the membrane tube of the SPMDs, as confirmed for SPMDs in water by Ellis et al. (1995). Triolein is used as the lipid compartment of the SPMD because it is the major storage fat found in aquatic organisms, and the triolein-water partition coefficient (KTW) and the octanol-water partition coefficient (KOW) (a property describing the solubility in water compared to octanol (a lipid phase)) of hydrophobic, organic compounds are very similar (Chiou, 1985). However, SPMDs have a much higher lipid content than aquatic organisms (20 percent in SPMD vs 1-10 percent in fish) and the membrane tube has a higher capacity to concentrate hydrophobic compounds than biomembranes. Thus, when comparing SPMDs and tissues of organisms on a mass-basis, (for example, 1 g SPMD versus 1 g fish), SPMD has a higher concentration capacity for hydrophobic pollutants. SPMDs can therefore be deployed for relatively short time periods, from days to weeks, and still ensure that detectable levels are obtained in the SPMDs (Huckins et al., 2002b). Ockenden et al. (2001), for instance recommended minimum sampling times for PCBs in air of 3-20 days, depending on the congeners present, the limit of detection and the air concentrations.

Membrane 75-90 mm thickness Triolein Transient (lipid) caves size < 10 Å Organic pollutant AIR AIR

Low density polyethylene (LDPE) membrane tube LDPE 91.4 cm long 2.5 cm wide loop

Figure 4. Schematic diagram of a standard 1-mL triolein SPMD including a thin film of triolein (neutral lipid found in many aquatic organism) heat sealed inside a lay flat thin-walled tube of non-porous low-density polyethylene (LDPE). The ends of the LDPE tube have loops to facilitate the deployment during sampling.

2. Passive air sampling

2.2.1 Sampling principles and factors influencing the sampling rate

In a SPMD, passive sampling is achieved by physical absorption in the membrane tube, or diffusion through the membrane tube resulting in absorption of the analyte(s) in the triolein. The uptake of a compound in a SPMD occurs in three different phases; the linear, curvilinear and equilibrium phases (Figure 5). The amount taken up (M) by a SPMD depends on (i) the physicochemical properties of the compound sampled, (ii) the ambient air concentration, (iii) the exposure time, (iv) the sampler design and (v) the environmental sampling conditions. The sampling rate (RS), and the time a chemical remains in the linear and the curvilinear phases until it reaches equilibrium, are independent of the ambient concentrations but depend on (i) the design of the sampler, (ii) the pollutants at the sampling site and (iii) the environmental conditions. Shoeib and Harner (2002) for instance, reported that PCB 28 (tri-CBs) and PCB 52 (tetra-CB) reached equilibrium after 50-75 and 200-400 days, respectively, while the uptake of PCBs with octanol-air partition coefficient (KOA) values > 9 like PCBs 101 (penta-CB), 137 and 138 (hexa-CBs) was linear during the entire 450-day sampling.

M at time t depends on air concentration, sampled compound, sampler design (SPMD configuration + protective devices) and environmental conditions. M at equilibrium depends on air concentration, sampled

compound and SPMD configuration (gives KSPMD). Equlibrium r ea lin rvi Cu Amount M in SPMD (ng)

r a e t in each phase is independent of the air concentration in L but depends on sampled compound, sampler design (SPMD configuration + protective devices),

environmental conditions. (gives RS, ku, ke)

Exposure time t (day) Figure 5. The three mass uptake phases that occurs in a SPMD as a function of the exposure time t in days. The time in each phase depends on the physicochemical properties of the compound sampled, the sampler design including the configuration of the SPMD and the devices protecting the SPMDs, and the environmental conditions (modified from Huckins et al. (2002b)).

2. Passive air sampling

Typically, SPMD sampling is performed either in the linear and time- integrative uptake phase or at equilibrium (Figure 5). In the latter sampling approach, equilibrium ambient concentrations are estimated and the method has the advantage that site effects of the environmental conditions are negligible in some cases. However, the time to approach/reach equilibrium and the magnitude of the SPMD equilibrium partition coefficient (KSPMD) also depend on the environmental conditions, and the user has to demonstrate that steady state has been achieved. This sampling approach is mainly recommended when the sampling situation is quite constant, like in indoor environments. However, the in situ integrative sampling approach is recommended in most cases and was used in the work underlying this thesis. As mentioned previously, there are several advantages with linear and integrative sampling. Episodic pollutant releases are more easily detected, TWA concentrations (which are required in exposure assessments for organisms) can be estimated, and pollutants present in trace or ultra-trace levels are concentrated to levels above detection limits. However, this sampling approach is affected by the environmental conditions at the sampling site.

The uptake in the SPMD can be restricted by three possible barriers; a thin boundary or diffusion layer at the membrane-exposure medium interface, the membrane and the lipid phase (triolein) of the SPMD. The diffusive mass transfer in the boundary layer is controlled by diffusion, whereas the mass transfer in the membrane, due to the transient cavity-structure of the membrane, is restricted by both diffusion and permeation. The transient cavities of the membrane restrict the passage of all compounds, but the truly dissolved or gaseous compounds, as mentioned earlier, will be able to permeate through the membrane, and accumulate in the SPMDs. Thus, only the TWA concentrations of the primary bioavailable pollutants will be estimated from the amounts found in the SPMDs. The resistance to diffusive mass transfer in the lipid of the SPMD is small in comparison to the resistance in the membrane and the aqueous boundary layer, and can be neglected. In water, biofouling at the exterior membrane surface can also restrict the uptake in the SPMDs. However, the uptake is mainly controlled by the boundary layer or the membrane.

In water, several studies have indicated that the linear and integrative uptake of hydrophobic compounds (log KOW > 4.4), is mainly controlled by the boundary layer (Booij et al., 1998; Huckins et al., 1999; Vrana and Schuurmann, 2002). Thus, the thickness of the boundary layer, and therefore

2. Passive air sampling the water flow/turbulence, effect the uptake of most POPs in SPMDs from water (Booij et al., 1998; Vrana and Schuurmann, 2002). For compounds with log KOW < 4.4, the SPMD-water partition coefficient (KSPMD) decreases so that predominantly the membrane controls their uptake. The uptake in the membrane depends on the compounds physical/chemical properties, like molecular three-dimensional geometry and functional groups, because of possible polymer-molecular interactions and steric effects in the membrane. Thus, the advantage of boundary layer control compared with membrane control is that the diffusive mass transfer is non-specific for compounds with similar size, and hydrophobicity, and in air, also volatility (KOW vs KOA). For instance, the water diffusion coefficients of compounds with molecular masses up to 500 g ⋅ mol-1 are in the order of 10-5 cm2 ⋅ s-1 (Booij et al., 1998). However, the effect of environmental conditions is important to calibrate and control. In water, the environmental variables temperature (e.g., Huckins et al., 1999; 2002a; 2002b; Booij et al., 2003) and biofouling (e.g., Huckins et al., 1990; 2002a; 2002b) also influence the uptake of compounds in SPMDs. Temperature (Ockenden et al., 1998a) and high wind-speed/turbulence (Papers II, III) have been shown to affect the uptake in SPMD from air as well. However, the effect of biofouling is expected to be negligible in air, whereas chemicals bound to particles or aerosols can be trapped on the membrane surfaces and influence the amounts taken up by SPMDs from air (Lohmann et al., 2001; Bartkow et al., 2004). UV sunlight is another environmental variable that should be considered during air sampling because photodegradation of UV-sensitive compounds can occur in the SPMDs (Orazio et al., 2002; Bartkow et al., 2004).

In air, the temporal and geographical variations over time in temperature, UV-radiation and wind-speed/turbulence can be high. In particular, the ambient temperature can vary significantly during the course of a day and between days. One way to reduce the site effects of environmental conditions during sampling is to protect the SPMDs from direct exposure to sunlight, rain, wind and particle deposition by using shelters. Results in Paper III, and research by Ockenden et al. (2001) have shown that the effect of wind- speed/turbulence can be reduced by shelter the SPMDs with metal devices that allow air to pass freely around the SPMDs. Furthermore, use of performance reference compounds (PRCs) to assess the biofouling effect on the uptake in SPMDs from water was proposed by Huckins et al. (1993). PRCs can be described as analytically non-interfering organic compounds that exhibit moderately to relatively high SPMD affinities and are added to the lipid-phase of the SPMD prior to membrane enclosure. (It is critical for the

2. Passive air sampling amount of PRC to be equally distributed in the triolein in order to use the whole surface area in the release process.) The theory of the PRC approach holds that the release rates of PRCs are related to the uptake rates of the analogous native compounds sampled and thus can be used to assess the site effects of environmental variables (Figure 6). Several studies has shown that PRCs can also be used to assess the effect of temperature and water flow/turbulence on the uptake in SPMDs from water (e.g., Booij et al., 1998; Huckins et al., 2002a; Vrana and Schuurmann, 2002; Booij et al., 2003). Thus, the use of PRCs will improve the precision of the SPMD sampling. The PRC approach has also been tested for SPMD sampling in air. In the work reported in Papers II and III deuterated (2H) PAHs and carbon 13-labelled (13C) PCBs were used, respectively, as PRCs and it was shown that the release of these compounds can be used to assess the effect of wind on the uptake in SPMDs from air. Ockenden et al. (2001) have also used 13C-PCBs and found a correlation between the uptake and release rates in SPMDs from air. In addition, Bartkow et al. (2004) used 2H-anthracene and 2H-pyrene as PRCs.

High wind Low wind PRC amount M amount PRC Native amount M Native amount

Exposure time t (day) Exposure time t (day) Figure 6. Graphs illustrating how the released amount of performance reference compounds (PRCs) can be used to assess the effects of environmental variables on the uptake of the analogous native compounds. (Here assessment of the effect of variable wind-speed that cause changes in the boundary layer thickness, and thus affect the uptake of target analytes)

2.2.2 Target compounds

The SPMDs concentrate trace and ultra trace levels of the dissolved and vapour phases of nearly all nonpolar organic compounds with molecular weights < approximately 600 Daltons and a cross sectional diameter <

2. Passive air sampling approximately 10 Å, to levels that are much higher than the ambient concentrations and hence detectable. Thus, most nonpolar aromatic compounds can be sampled by SPMDs from the environment. Listed below are examples of compound classes that have been taken up in SPMDs from water (Huckins et al., 2002b; *Ikonomou et al., 2002):

• Alkylated phenols (e.g., nonyl phenol) • Chlorinated anisoles and veratroles • Certain heterocyclic aromatics • Chlorinated and brominated benzenes • Organophosphate pesticides • Non-ionic organometallic chemicals • Organochlorine pesticides (OCs) • Polycyclic aromatic hydrocarbons (PAHs) • Polychlorinated biphenyls (PCBs) • Polychlorinated dioxins (PCDDs) and furans (PCDFs) • Polybrominated diphenyl ethers (PBDEs)* • Pyrethroid insecticides

In water, the RS of compounds increases with increasing log KOW and is maximal around five and six (e.g., Huckins et al., 1999; Luellen and Shea,

2002; Booij et al., 2003). However, for compounds with log KOW < 3, which are more hydrophilic, the concentration capacity of SPMDs decreases dramatically as log KOW decreases, and SPMD technique have no advantages over grab sampling and other relatively low volume techniques. For instance, Vrana et al. (2002) suggested that the linear and integrative uptake of compounds under membrane control (log KOW < 4.4) can be as low as two days in water.

2.2.3 Fields of application

The SPMD technique can be used in a number of different applications to assess both the ambient levels (and thus environmental exposure) and the toxicity of organic pollutants. Specific applications in which SPMDs can be used in field studies are described below.

I. Determination of (i) the presence of new and existing organic pollutants, (ii) the sources, and (iii) the local, regional and continental

2. Passive air sampling

spatial distribution (often in relative terms that is semi-quantitatively) of pollutants. II. Estimation of the TWA concentrations of pollutants that are dissolved or in the vapour phase (quantitatively). III. Estimation of organisms’ exposure to organic pollutants that are dissolved or in the vapour phase. IV. Identification of chemical fractions, with in situ sampled concentrations, of primary bioavailable organic pollutants that can be used in bioassays/biomarker tests and immunoassays.

Table 1 gives examples of field studies which have used SPMDs for different purposes. Many of the applications are closely related, and thus, most of the listed studies include more than one of the applications. The sampling environments (water, air and/or soil) of each study are specified. In the Table, several field studies are listed as having quantitatively determined the TWA concentrations. However, in some of these studies, the sampling rates used were estimated and their approach was semi-quantitative rather than truly quantitative.

2. Passive air sampling

Table 1. Examples of field studies which have used SPMDs for different purposes and in different environments. *Purposes of the field studies presented in the work underlying this thesis. Purpose Field studies (sampling environment in brackets) Ii: To determine presence* Lebo et al., 1992 (creek); Prest et al., 1992 (river); Prest et al., 1995 (seawater+air); Strandberg et al., 1997 (compost); Bergqvist et al., 1998 (seawater); Ockenden et al., 1998a (air); Ockenden et al., 1998c (air); Echols et al., 2000 (river); Zimmerman et al., 2000 (rivers); Rantalainen et al., 2000 (soil at lake shore); Sabaliunas et al., 2000 (rivers); Granmo et al., 2000 (seawater); Booij and van Drooge, 2001 (seawater+air); Lohmann et al., 2001 (air); Lindström et al., 2002 (lakes), Paper I, 2003 (air); Isidori et al., 2003 (air); Meijer et al., 2003 (air); Rastall et al., 2004 (lake); Balmer et al., 2004 (lakes); Petty et al., 2004 (river+waste water); Paper V, 2004 (indoor air); Shaw et al., 2004 (river+seawater); Paper IV, 2004 (air).

Iii: To determine sources* Granmo et al., 2000 (seawater); Lohmann et al., 2001 (air); Lindström et al., 2002 (lakes); Meijer et al., 2003 (air); Paper I, 2003 (air); Balmer et al., 2004 (lakes); Petty et al., 2004 (river+waste water); Paper V, 2004 (indoor air); Shaw et al., 2004 (river+seawater); Paper IV, 2004 (air).

Iiii: To determine the spatial distributions Local* Prest et al., 1992 (river); Prest et al., 1995 (seawater+air); Bergqvist et al., 1998 (seawater); Echols et al., 2000 (river); Rantalainen et al., 2000 (soil+lake shore); Granmo et al., 2000 (seawater); Booij and van Drooge, 2001 (seawater+air); Rastall et al., 2004 (lake); Paper V, 2004 (indoor air). Regional* Sabaliunas et al., 2000 (rivers); Zimmerman et al., 2000 (rivers); Lohmann et al., 2001 (air); Lindström et al., 2002 (lakes), Paper I, 2003 (air); Balmer et al., 2004 (lakes); Shaw et al., 2004 (river+seawater). Continental* Ockenden et al., 1998c (air); Meijer et al., 2003 (air); Paper IV, 2004 (air).

II: To estimate time weighted Prest et al., 1995 (seawater+air); Bergqvist et al., 1998 (seawater); Ockenden average (TWA) concentrations et al., 1998a (air); Ockenden et al., 1998c (air); Echols et al., 2000 (river), of pollutants* Rantalainen et al., 2000 (soil at lake shore); Granmo et al., 2000 (seawater); Lindström et al., 2002 (lakes); Paper I, 2003 (air); Balmer et al., 2004 (lakes); Petty et al., 2004 (river+waste water); Paper V, 2004 (indoor air); Shaw et al., 2004 (river+seawater).

III: To estimate organisms’ Echols et al., 2000 (river); Zimmerman et al., 2000 (rivers); Paper I, 2003 (air). exposure*

IV: To identify chemical Sabaliunas et al., 2000 (rivers); Isidori et al., 2003 (air); Rastall et al., 2004 fractions, with in situ sampled (lake). concentrations, of primary bioavailable organic pollutants that can be used in bioassays/biomarker tests and immunoassays

2.2.4 Theory and mathematical models of SPMD sampling

The basic theory of SPMD sampling, and mathematical models used to estimate the water ambient concentrations from the SPMD concentrations, were developed by Huckins et al. (1993). In the early work, triolein (lipid) was considered to be the only compound-accumulating compartment of the

2. Passive air sampling

SPMD. The concentration of a compound in the lipid compartment of the SPMD is given by (Huckins et al., 1999)

-1 -1 CL = Cw · KLw(1-exp[-ko · Kmw· A· t · KLw · VL ]) Eq. 4

where CL is the concentration in the SPMD lipid compartment, Cw is the water concentration, KLw is the lipid-water partition coefficient, ko is the overall mass-transfer coefficient, Kmw is the membrane-water partition coefficient, A is the SPMD surface area, t is the exposure time in days, KLw is the lipid-water partition coefficient and VL is the lipid volume.

However, the membrane tube also accumulates organic pollutants and, thus, contributes to the amount found in the SPMD. To include the concentrations recovered by the membrane and thus, estimate the water concentrations from the amounts found in the SPMDs, Huckins et al. (1999) suggested that the membrane can be expressed as a lipid-equivalent volume and, thus, the SPMD can be considered as a single-compartment model. However, the differences between the triolein and the membrane tube in both total mass (4:1 membrane/triolein ratio), and concentration capacity (triolein has about four times higher capacity (Huckins, 2004)) has to be compensated for in the model. For compounds under boundary layer control, the uptake in amount ⋅ -1 g SPMD , CSPMD, is then described by

-1 -1 -1 -1 CSPMD = Cw · KSPMD(1-exp[-kw · A · t · KLw (VL + KmL · Vm )]) Eq. 5

where KSPMD is the SPMD-water partition coefficient, kw is the mass transfer coefficient in the boundary layer, KmL is the membrane-lipid partition coefficient, and Vm is the volume of the membrane.

When the uptake of compounds is controlled by the membrane, the pollutant concentrations in the SPMD are described by

-1 -1 -1 -1 CSPMD = Cw · KSPMD(1-exp[-km · Kmw · A · t · KLw (VL + KmL · Vm )]) Eq. 6

where km is the mass transfer coefficient in the membrane.

To simplify eqs. 2 and 3, the terms kwA/KLw(VL+KmLVm) and kmKmwA/KLw(VL+KmLVm), respectively, in each equation are expressed as the -1 exchange coefficient, ke (t ), and the concentration in the SPMD is calculated according to

2. Passive air sampling

CSPMD = Cw · KSPMD(1-exp[-ke · t]) Eq. 7

Under linear and integrative sampling, when the term ket is small or CSPMD/CW

<< KSPMD, the eq. 7 can be reduced to

-1 CSPMD = Cw · ku · t = Cw · RS · t · MSPMD Eq. 8

-1 -1 where ku is the uptake rate constant in mL · d · g SPMD for a specific 3 -1 compound, RS is the volume of water sampled per unit time in dm · d at a given temperature, and MSPMD is the mass of the SPMD in g. The uptake and the elimination of the SPMD are related according to

ku = KSPMD · ke Eq. 9

where KSPMD is the SPMD-water partition coefficient.

The basic theory of SPMD sampling in water developed by Huckins et al. (1993, 1999) was used in this thesis to model the SPMD sampling in air, assuming that the uptake and elimination processes of SPMDs in water and air are related. In the work underlying this thesis only standard 1-mL triolein SPMDs and the integrative sampling approach were used. Thus, the integrative model as expressed by eq. 8 was used to describe the uptake in SPMDs. However, some changes to that model were introduced. First, the subscript SPMD was replaced with SA for SPMD in air. Second, the SPMD -1 concentration, CSPMD, in g ⋅ g SPMD was not modelled since standardized -1 SPMDs were used. Instead, the amount, MSA, in ng ⋅ standard SPMD was modelled and thus, the term MSPMD was not included in the equation, giving

MSA= CA · ku · t = CA · RS · t Eq. 10

-3 where CA is the air concentration in ng · m , ku is the uptake rate constant in -1 mL · d for a standard 1-mL SPMD, and RS is the volume of air sampled per unit time in m3 · d-1, at a given temperature, for a standard 1mL SPMD. In cases where standardized SPMDs are not used, the ku and the RS values have to be normalized by the differences between the masses of the SPMD configuration used and a standard 1-mL SPMD.

2. Passive air sampling

2.2.4.1 The PRC model

As mention earlier, PRCs can be used to calibrate the RS of SPMD in situ and thus assess the effect of environmental variables (Huckins et al., 1993). Two types of PRC approaches can be used to assess the differences between the laboratory pre-calibrated RS and the RS values in situ. One approach is to use PRC-derived exposure adjustment factors (EAFs), described in detail by, for instance, Huckins et al. (2002a). The second approach, described in detail by for instance, Booij et al. (1998), was used in the work underlying this thesis, and is based on the use of ke values calculated from the release of PRC according to

0 MPRC-A = M PRC-A exp(-ke · t) Eq. 11

0 where MPRC-A is the amount of the PRC in the SPMD at time t and M PRC-A is the amount of the PRC in the SPMD at the beginning of the sampling. If the amount of PRC in the SPMD is measured at the beginning and end of the sampling, as was done in work underlying this thesis, eq. 11 is solved as a two-point derivation of ke (Huckins et al., 2002b)

0 -1 ke = ln(M PRC-A/MPRC-A) · t Eq. 12

When the uptake in the SPMD is linear and integrative, the estimated ke value can be used in eq. 8 and the air concentration of the sampled compound is given by

-1 -1 -1 CA = MSA · ke · KSA · t Eq. 13

where KSA is the SPMD-air partition coefficient.

When the PRC amount released is ≥ approximately 60 percent, the uptake of compounds with similar or lower SPMD affinity is nonlinear or at equilibrium. Gale (1998), for instance, suggested that at 60 percent of the steady state concentrations the uptake will be in the curvilinear phase. In such cases, use of eq. 13 will underestimate the CA and instead an exponential or equilibrium model should be used.

In summary, for quantitative data interpretation (converting amounts found in the SPMDs to air concentrations) relevant RS (see eq. 10) or KSA (see eq.

2. Passive air sampling

13) values have to be known. However, for the time being no KSA values are available, and the KOA values have to be used as surrogates for the KSA values as were suggested by, for instance, Ockenden et al. (1998a).

3. Compounds sampled and their physical/chemical properties

3. COMPOUNDS SAMPLED AND THEIR PHYSICAL/CHEMICAL PROPERTIES

The work underlying this thesis was focused on the nonpolar aromatic compounds PCBs, PAHs, alkyl-PAHs and nitro-PAHs. Below selected physical/chemical properties of these compounds are presented.

3.1 Polycyclic aromatic hydrocarbons (PAHs)

PAHs are a wide class of compounds which consist of different numbers of fused benzene rings (generally two to six) in linear, angular or cluster arrangements (Figure 7). The main concerns of PAHs are their wide distribution in the environment, their persistence and the fact that some PAHs have reported mutagenic and carcinogenic properties associated with increasing size of the molecule (compounds with four or more benzene rings being especially carcinogenic), and their metabolic transformation to reactive dihydrodiol epoxides (Pickering, 1999). They are present in the atmosphere primarily due to emissions from gasoline- and diesel-powered motor vehicles, municipal and commercial incinerators, residential heating systems that combust fuels like coal, wood, gas and oil, various industrial processes and volatilization from polluted soils. Over 500 PAHs have been detected in air, but most studies focus on the 16 PAHs designated as priority pollutants by the United States Environmental Protection Agency (the 16 EPA PAHs). PAHs can also exist in substituted forms like alkyl- or nitro-PAHs (Figure 7). High concentrations of alkyl-PAHs compared to unsubstituted PAHs have been found in emissions of petrogenic sources like traffic and fossil fuel combustion systems for home heating (see, for instance, Ramdahl, 1983; Rogge et al., 1993; Nielsen, 1996; Lim et al., 1999; McDonald et al., 2000). Nitro-PAHs originate from both primary sources like emissions of diesel motor vehicles (e.g., Nielsen, 1984; Feilberg et al., 1999; Bamford et al., 2003), and reactions of the parent PAHs in the atmosphere (e.g. Nielsen, 1984; Arey et al., 1986; Atkinson et al., 1987; Atkinson and Arey, 1994). The main concerns regarding nitro-PAHs are that many of them are reported to have direct mutagenic activity and carcinogenicity (e.g., Tokiwa et al., 1994; Durant et al., 1996).

3. Compounds sampled and their physical/chemical properties

a) b) c) Figure 7. Molecular structures of a) the EPA PAH phenanthrene, b) 1- methylphenanthrene and c) 2-nitrofluorene.

To our knowledge, the KOA of PAHs have only been measured in one study by Harner and Bidleman (1998), who determined the log KOA values of fluorene, phenanthrene, pyrene and fluoranthene, respectively at different temperatures (examples in Table 2, pp 55). Lohmann et al. (2000) suggested that the log KOA values for 2-ring to 6-ring PAHs ranged between 6 and 12. Furthermore, PAHs display a wide range of gas-particle partitioning characteristics in the atmosphere. PAHs with two and three rings, which have relatively low log KOA values, are mainly associated with the gas phase, 4- ring PAHs occur in both the gas and particle phase, while PAHs with higher log KOA values, such as PAHs with five rings or more, are mainly bound to particles (see, for instance, Lohmann et al., 2000). In comparison, both alkyl- and nitro-PAHs appear to have lower vapour pressure than unsubstituted PAHs with corresponding numbers of benzene rings in their molecular structure (manifested, for example, by longer retention times in GC-analysis). For instance, Dimashki et al. (2000) found that 2-ring nitro-PAHs (nitronaphthalenes) occurred only in the gas phase, nitro-PAHs with three rings (9-nitroanthracene) were associated with both the gas and particle phases, whereas nitro-PAHs with four rings or more (1-nitropyrene, 2- nitrofluoranthene and 7-nitrobenz[a]anthracene) were mainly bound to particles.

3.2 Polychlorinated biphenyls (PCBs)

PCBs are a group of compounds consisting of 209 congeners with varying degrees of chlorination of the biphenyl scaffold: 1 to 10 chlorine atoms in the molecular structure (Figure 8). In all, there are ten homolog groups (mono- to deca-CBs). In the work underlying this thesis, the PCBs were numbered according to the IUPAC rules, a system developed by Ballschmitter and Zeller (1980). PCBs have been widely used, for instance, as plasticizers, fire retardants, sealing liquids and dielectric fluids for capacitors and transformers. They can be found in all environmental compartments including humans, and

3. Compounds sampled and their physical/chemical properties have shown multiple toxic effects on wildlife and humans (e.g., Giesy and Kannan, 1998; van den Berg et al., 1998; Bosveld et al., 2000; Kannan et al., 2000; Letcher et al., 2002). The major source of PCBs in the atmosphere is volatilization from contaminated sites where they have been stored or dumped, and combustion of PCB-containing material (e.g., Simcik et al., 1997).

Cln

Figure 8. Schematic diagram of the general molecular structure, substitution positions, and nomenclature of the PCBs.

Harner and Bidleman (1996) have measured the log KOA values of 15 PCBs (mono, tetra-, penta-, hexa- and hepta-CBs), at different temperatures, and found log KOA values between 7 and 11 at 20 °C, that increased with increasing chlorination (examples in Table 2, pp 55). Within the same homolog group, PCBs with fewer chlorine atoms in ortho-position were found to have higher KOA values. Furthermore, Chen et al. (2002) predicted the log KOA values of the 209 PCBs, at the same temperature, to be between 5.6 and 12.3 (examples in Table 2, pp 55). In the atmosphere, PCBs predominantly exist in the gas phase (e.g., Ockenden et al., 1998a; Yeo et al., 2003; Tasdemir et al., 2004). For instance, Yeo et al. (2003) found that 90 percent of the total PCBs (gas+particle phase) reside in the gas phase, with mono- to hexa-CBs mainly associated with the gas phase, whereas hepta- to deca-CBs occur in both the gas and particle phase.

4. Sampling and chemical analysis

4. SAMPLING AND CHEMICAL ANALYSIS

The design of the sampler and the sampling procedures used in the work underlying this thesis is presented in detail below. In addition, the chemical analysis of the SPMD samples, the quality assurance and quality control of the data, and the data interpretation are described.

4.1 Design of the sampler

The design of the Exposmeter air sampler used in the work described in this thesis included two standard 1-mL SPMDs obtained from Exposmeter AB, Tavelsjö, Sweden. Each SPMD was mounted carefully between five steel rods attached to a 150 × 140 mm steel disc (a steel device called a spider) and the loops at the ends of the membrane were hung over the two outermost steel rods (Figure 9). The two SPMDs, mounted on separate spiders, were placed horizontally on top of each other inside the upper part of a metal umbrella that was hung with the opening, covered by a metal mesh (340 mm i.d.), facing downwards. The metal umbrella had an inner diameter of 320 and 160 mm at the bottom and top, respectively, decreasing progressively from its lower to its upper part. This design of the metal umbrella was chosen to protect the SPMDs from direct wind, sunlight, rain, and particle deposition while allowing air to pass under and around the SPMDs.

Figure 9. Design of the Exposmeter air sampler used in the work underlying this thesis. Two standard 1-mL SPMDs were deployed on separate 150 × 140 mm steel spiders and placed horizontally on top of each other inside the upper part of a metal umbrella (320/160 mm i.d.) that was hung with the opening, covered by a metal mesh (340 mm i.d), facing downwards (Paper III).

4. Sampling and chemical analysis

4.2 Sampling procedure and approach

In the work described in this thesis, unexposed SPMDs were delivered from Exposmeter in gas-tight solvent-cleaned tin cans (five SPMDs per can) and then stored at -18 °C. The SPMDs were transported in these tin cans to the sampling locations where the SPMD samplers were set up according to previous description (4.1). The samplers were hang on metal racks at 1 to 3 m heights for approximately 21 days when the SPMDs were retrieved and placed in separate pre-cleaned gas-tight tin cans at -18 °C. The average sampling temperature at each sampling location was measured in parallel. In the study reported in Paper IV, environmental conditions such as wind-speed and wind direction were also measured, when possible. To provide controls for the purity of the SPMDs used and the contribution of the sampling and analytical procedures to the amounts found in the SPMDs, one additional SPMD (referred to as the field control, FC) was exposed to air only during deployment and retrieval of the SPMDs, and was analyzed using the same procedure as the other SPMDs.

Air sampling for < 30 days ensures that uptake of most nonpolar aromatic compounds in the SPMDs is linear, and thus sampling is time-integrative. Therefore, the integrative model was used to describe the uptake of pollutants in the SPMDs (see 2.2.4). However, the uptake of some of the most volatile PAHs, such as naphthalene, are in the curvilinear phase at the end of a 3- week sampling period and the naphthalene amounts found in the SPMDs, calculated as ng ⋅ SPMD-1 ⋅ d-1, were underestimated. In addition, time- integrative sampling of these compounds does not occur throughout the sampling and the ability of SPMD to detect episodic pollutant releases declines.

4.2.1 The PRC approach

2H-PAHs with molecular weights no higher than that of deuterated chrysene, together with 13C di- and tri-CBs are commonly used as PRCs. The compounds selected as PRCs, should be suitable for the environmental conditions of the sampling sites and the sampling time in order to ensure that the PRC losses during the sampling are in an acceptable range. The changes 0 in the initial PRC amounts should ideally meet the criteria of 1-(CSPMD/C SPMD) × 100 % > 3 × coefficient of variance (CV) when the PRC releases < 50 %, 0 and CSPMD/C SPMD × 100 % > 3 × CV when the PRC releases > 50 % (Huckins

4. Sampling and chemical analysis et al., 2002b). These criteria are generally met when the amount of PRC released is in the range 20 to 80 % of the initial amount in the SPMDs (Huckins et al., 2002b). For example, if the sampling time is about two weeks, 2H-naphthalene will be completely lost, while no loss of 13C-PCB 180 will be detected. To control the amount of PRC in the SPMD at the beginning of the sampling, one SPMD, referred to as FC, was analyzed with the same procedure as the other SPMDs.

4.2.2 Cleaning of the sampling equipment

The sampling devices were carefully cleaned before use. First, the steel spiders were put in 40 %-hydrochloric acid for 5 minutes and then in a cold water bath for 10 minutes. The acid-washed steel spiders were brushed and washed by hand in water for 1-2 minutes, and then heated for 10 hours at 500 °C. The metal umbrellas were washed by hand in clean water and detergents. Before retrieval of air exposed SPMDs, the interior of new tin cans were shaken in 10 mL of acetone for 20-30 s, and then in 10 mL hexane for another 20-30 s

4.3 Extraction, cleanup and analysis

A wide range of nonpolar organic compounds are taken up in the SPMDs. Therefore an analytical method that is non-destructive and allows the analysis of a large number of organic pollutants is often used (see, for instance, Lebo et al., 1992). The general procedure used for cleaning up SPMD samples, including the extraction and analysis of target or unknown analytes, is presented in Figure 10. A detailed description of the quality of the chemicals used, their sources and the analytical method used in the work underlying this thesis is given in Paper I. Briefly, the SPMDs were brushed and washed by hand in water, then shaken in hexane followed by hydrochloric acid and finally washed in water again, to clean the membrane surface from lipids, soot and other particles. The externally cleaned SPMDs were checked for holes and then the organic pollutants were extracted from them by dialysis for 24×2 h in 95:5 (v:v) cyclopentane:dichloromethane, with an exchange of solvent after 24 h. The extracted SPMDs were discharged and the dialysates were cleaned with a gel permeation chromatography (GPC) system that included a high resolution (HR)-GPC column using 65:45 (v:v) hexane:dichloromethane as eluents. To remove residues of plastic films (polyethylene) and lipids, that

4. Sampling and chemical analysis still interfered with the analysis after the GPC-run, a second clean-up step was developed by testing various open liquid chromatography systems with a range of materials, including acid/basic silica, florisil, deactivated silica and a mixed silica including deactivated silica and basic silicate, with hexane and/or hexane:dichloromethane (50:50 or 75:25 (v:v)) as eluents (for all of the systems tested). The mixed silica gel column with 50:50 (v:v) hexane:dichloromethane as eluent was selected for use in further studies since it gave acceptable recoveries and high purities of the target analytes (PAHs). The SPMD samples were analyzed for PAHs, alkyl-PAHs and PCBs by high- resolution gas chromatography/low-resolution mass spectrometry (HRGC/LRMS) operated in electron ionization mode monitoring selected ions. In Paper IV, the SPMD samples were analyzed for the nitro-PAHs by HRGC/LRMS operated in negative ion chemical ionization mode, using an analytical method described in detail by Dusek et al. (2002).

Size exclusion chromatography

Organic solvent dialysis Chemical fraction (mixed silica)*

Exterior cleaning Chemical analysis (GC/MS)*

Figure 10. General procedures used for cleaning up SPMD samples including the recovery and analysis of target and unknown analytes taken up in exposed SPMDs. *The specific methods used in the work underlying this thesis are shown in parentheses.

4.4 Quality assurance and quality control (QA/QC)

Quality assurance (QA) and quality control (OC) are important to ensure the reliability of the data obtained. One way to improve the quality of the analysis is to include the use of labelled standards. In the work underlying this thesis, different labelled standards were used, depending on the target analytes, and were all added to the extracts after the dialysis. In the studies reported in Papers I and IV, six 2H-PAHs (naphthalene, fluorene, anthracene, pyrene,

4. Sampling and chemical analysis chrysene and benzo[k]fluoranthene) were used, in a mixture, as internal standards for the clean-up. In latter study, two 2H-nitro-PAHs (1-nitropyrene and 6-nitrochrysene) were also added to the dialysates. In the studies described in Papers II and III, one mixture of the 13C-PCBs 28, 47, 52, 101, 105, 118, 138, 153, 156, 180, 194 and 209, and another mixture of 2H-PAHs (acenaphthylene, fluoranthene and benzo[g,h,i]perylene were used in Paper II, and in Paper III, also naphthalene and anthracene was added to the mixture), were added as internal standards. In all cases, deuterated dibenzofuran was used as a recovery standard during the GC/MS-analysis. In cases where the samples were re-analyzed including an extra GPC-run, one 13C tetra- or penta- CB was used as recovery standard. Acceptable recoveries of these standards were within 50-120 percent. Obtained data were adjusted using these recovery values. For detection, the signal to noise ratio should be equal to or greater than three (S/N ≤ 3). In addition, laboratory blanks (LBs) and FCs consisting of solvent and SPMDs, respectively, exposed to air during the deployment and retrieval of the samples, were analyzed in the same manner as the SPMD samples, and were checked for possible interferences within the same retention times as the target analytes.

The reproducibility of the SPMDs was evaluated (Paper III) by analysing the two SPMDs of each sampler separately. The uptake of PAHs and PCBs between the replicate SPMD samples (derived from 21-day sampling and chemical analysis) was 17 and 19 percent, respectively. However, differences in the amounts taken up by some replicate SPMDs were found, indicating that the extremely high wind-speeds applied in this study could have caused wind conditions to differ inside the protective metal umbrellas and thus cause differences in the amounts taken up by the two SPMDs in each sampler. However, in the work presented in Paper III, we found that most replicate SPMDs, sequestered and released similar amounts of the gas-phase PAHs and PCBs, respectively. These results agree with other studies that have concluded that SPMDs sequester gas phase PCDD/Fs, PAHs and PCBs from the atmosphere with good reproducibility (e.g., Lohmann et al., 2001).

4. Sampling and chemical analysis

4.5 Data interpretation

4.5.1 Semi-quantitative and quantitative approach

In Papers I-IV, the SPMD data were interpreted by comparing differences in amounts and profiles found in the SPMDs, and were expressed in ng ⋅ SPMD-1 ⋅ day (d)-1. Since the reproducibility (comparability) of the SPMDs was high, but differences in the RS values between different target analytes and sampling sites were not accounted for, this approach to interpreting data has been defined as semi-quantitative. For quantitative data interpretation (converting amounts taken up in the SPMDs to air concentrations) relevant

RS (see eq. 10) or KSA (see eq. 13) values have to be known.

When Paper I was written, available calibration data for SPMDs in air were limited to field-calibrated RS values for PCBs reported by Ockenden et al. (1998a, 2001) and Shoeib and Harner (2002). Therefore, the RS values of PAHs in water calibrated by Huckins et al. (1999) were used for preliminary calculations of the PAH concentrations in air. At that time, no adjustments were made for the differences in density between water (1.0 kg ⋅ L -1) and air -3 -1 (1.2 kg ⋅ m ), and the RS values of the PAHs in water expressed in L ⋅ d were 3 -1 used as RS values in air expressed in m ⋅ d . Thus, this approach to interpret data was not quantitative, although the values were corrected for different RS values for PAHs in water.

In Paper III, the amounts in the SPMDs were converted to air concentrations 2 for some selected PAHs and PCBs using the in situ-calibrated ke values of H- phenanthrene and 13C-PCB 15, respectively, and the eq. 13. For the selected

PAHs and PCBs, no KSA values were available. Ockenden et al. (1998a) have successfully used the KOA values as surrogates for the KSA and, thus, this approach was applied in Paper II. The KOA values can be directly measured as was made by Harner and Bidleman (1996, 1998) (used in Paper III) or estimated from

-1 -1 KOA = KOW ⋅ KAW = KOW ⋅ R ⋅ T ⋅ H Eq. 14

where KAW is the air-water partition coefficient, H is the Henry’s law constant, R the gas constant, and T is the temperature (used in Paper II).

4. Sampling and chemical analysis

Thus, the approach used in Paper III corrected for the site effects of different sampling conditions, but since no KSA values were available for each compound, it was semi-quantitative rather than truly quantitative. To clarify that the data in Paper III were semi-quantitative, the air concentration data were normalized by the highest concentration determined for each compound.

4.5.2 Multivariate data analysis using principal component analysis (PCA)

For interpretation of the data in Paper IV presented in this thesis, principal component analysis (PCA) was used (Jackson, 1991). In a PCA the variations in a multivariate data matrix X with n rows (objects) and k columns (variables) are projected into a few uncorrelated (orthogonal) principal components (PCs). Basically, the systematic variation of the X matrix is extracted into smaller matrices, the score T and loading P´ matrices, extracting the variation between objects and variables respectively. The first PC is oriented to explain as much variation in the data as possible and presents the best linear summary of X. The second PC is orthogonal to the first, and explains the next largest variation in the data, and so forth. In the multivariate data analysis of this thesis, the number of significant PCs in the PCA was determined by cross-validation (Wold, 1978).

The PCs define a plane with dimensionality of the number of PCs. Each object can be projected onto this plane and the coordinates in this plane gives the score for the object (i.e., the distance between the origin of the plot and the object’s projection into the plane) (Figure 11a). Thus, the score (t) gives the importance of the object for the PC. In a score plot, the scores of two PCs are plotted against each other and thus, visualize the relationship between objects. Hence, two objects close to each other in a score plot are similar and objects far apart are not. The direction of each PC is defined by the loadings (p) as given by the cosine of the angles between the original variables and the PC (Figure 11b). Hence the p values show how the original variables “load” into the PC. Thus, by plotting the p values of two PCs against each other, the importance of each variable for each PC can be revealed. Comparing the score and the loading plot of two PCs reveal the importance of each original variable for each object, e.g. which individual PAH concentrations (variables) that are high or low at the sampling site (object).

4. Sampling and chemical analysis

a) b)

Figure 11. Illustration of the projection of multivariate data into a plane with dimensionality of a few uncorrelated principal components (PCs). a) The score (t) (the distance between the origin and the object’s projection into the plane) for each object gives its importance for each PC, while b) the loadings (p) (the cosine of the angles between the original variables and the PC) gives the importance of each variable for each PC (modified from Eriksson et al. (1999)).

PCA was used in Paper IV to reveal whether there were spatial differences in the concentration of gas phase PAHs within and between five European countries (Austria, Poland, Slovakia, Sweden and the Czech Republic) and to discern, if possible, the PAH-patterns of potential sources. PCA has been performed on SPMD data by Echols et al. (2000) to compare the relative PCB patterns of SPMDs, caged fish and sediments. Two different PC analyses were performed on the data set in our study, which consisted of 11 columns (variables) and 79 rows (objects). The object AU2-99 and the two variables dibenz[a,h]antracene and fluorene were excluded from the two models since more than 50 percent of the corresponding values in the dataset were missing. The variables (the individual PAH concentrations in the data matrix) showed a positive skewness for the objects (sampling sites) when plotted in a histogram. The data were therefore log-transformed to obtain an approximately normal distribution. In the first PCA all variables were mean- centred and scaled to unit variance, while the data in the second PCA were mean-centred and normalized by the total PAH amount sampled at each sampling location. The data were normalized to avoid the results being affected by the differences in PAH-concentrations between objects. The software package SIMCA-P 9.0, 2001, obtained from Umetrics AB, SE-907 19 Umeå, Sweden, was used for the multivariate data analysis presented in work of this thesis.

5. Factors influencing the uptake of compounds in SPMDs from air

5. FACTORS INFLUENCING THE UPTAKE OF COMPOUNDS IN SPMDs FROM AIR

When testing, for instance, the sampling rate, RS, or the effect of air velocity for diffusive samplers of gases and vapours, European standards require the tests to be performed under laboratory conditions in an exposure chamber made of material that is inert with respect to the measured substance (Anonymous, 2002b). In addition, a constant concentration of a calibration gas mixture that either can be traceable to national standards or has been determined by a reference method should be generated in the exposure chamber. When determining the wind effects, different sampling conditions, with variations in exposure time, the air velocity and the sampler orientation should be tested, and at least six samplers should be exposed for each test condition. The results from the six (or more) test samplers should be compared with results obtained simultaneously with six samplers of a reference method or one independently calibrated instrument. When determining the RS of the target analytes, 24 test samplers should be used instead of six. The volume of the generated atmosphere (calibration gas), passing through the exposure chamber, should have sufficient capacity to accommodate all these samplers simultaneously. In addition, the samplers in the exposure chamber should be positioned in such a manner ensuring that there is no interference between each sampler. If the outlet concentration, determined by a reference method or an independent measurement, differs by more than 5 percent from the inlet concentration, the generated air volume is too low or the samplers are positioned too close to one another, and the exposure system should be modified.

The European standard includes several requirements that are problematic to fulfill when testing the performance of standard SPMDs as PAS of nonpolar aromatic compounds. Firstly, generating a calibration gas with constant concentrations of POPs like PAHs and PCBs that partition both to the gas phase and to particles involves several difficulties: I. These compounds have quite low gas phase solubility and generating a gas of these compounds is complicated. II. Generating gas concentration that is constant over time is difficult because the amounts of aerosols and particles in the atmosphere of the exposure chamber can vary and that will effect the gas concentrations. Since these compounds display a wide range of gas- particle partitioning characteristics, this effect will also vary between compounds.

5. Factors influencing the uptake of compounds in SPMDs from air

Another requirement that is problematic to fulfill is to generate an air volume (calibration gas) in the exposure chamber which has the capacity to accommodate the required SPMD samplers and reference samplers simultaneously throughout the sampling period while ensuring that the gas concentrations of the analytes change by less than 5 percent. For instance, ensuring that the SPMDs have at most a 5 percent effect on the air concentration during a 21-day continuous air sampling with six SPMD samplers, and six MeVols, which is a standard method, requires 12,600 3 -1 3 -1 3 (assumed RS: 5 m ⋅ d ) and 35,280 (assumed RS: 14 m ⋅ d ) m air, respectively. Due to these difficulties, the European standard conditions for laboratory exposure tests are almost impossible to meet when testing the performance of SPMDs as PAS of nonpolar aromatic compounds.

5.1 Physical/chemical properties of the compounds sampled

The SPMDs concentrate trace and ultra trace levels of the dissolved and vapour phases of nearly all nonpolar aromatic compounds with molecular weights < approximately 600 Daltons and a cross sectional diameter < approximately 10 Å. In water, KSPMD (i.e. the concentration capacity) and the RS of the compounds, depend on the log KOW value of the target analyte (e.g., Huckins et al., 1990; Booij et al., 1998; Meadows et al., 1998; Huckins et al., 1999; Luellen and Shea, 2002; Vrana and Schuurmann, 2002; Booij et al.,

2003). Luellen and Shea (2002), for instance, suggested RS values of PAHs and alkyl-PAHs in the range 2.11 to 6.06 L ⋅ d-1 that increased with increasing log KOW values up to about 5-5.5, where the RS began to decrease, presumably due to steric effects of the large molecular weight fused-ring PAHs. They also found that PAHs with log KOW > 4.5 remained in the linear uptake phase throughout a 30-day sampling period, whereas the uptake of PAHs with log

KOW < 4.0 or between 4.0 and 4.5 was curvilinear after 7 and 13-15 days, respectively. Furthermore, Booij et al. (2003) suggested that the RS of chlorobenzenes, PCBs and PAHs, respectively, were all in the range 20-200 L ⋅ d-1 (which are high, probably due to the effect of high water flow), and increased with increasing log KOW values up to about six when the RS slightly decreased again. For compounds with log KOW values below three, the concentration capacity (the KSPMD) of SPMDs decrease dramatically, and the use of SPMDs as passive water samplers has no advantages over grab water sampling and other relatively low volume sampling techniques.

5. Factors influencing the uptake of compounds in SPMDs from air

5.1.1 RS of PCBs in air

In unpublished work described in this thesis, the RS of a number of PCBs in SPMDs from air were tested by comparing SPMD samplers with conventionally used MeVols. Instead of following the European standard conditions the experimental setup to determine the RS was simplified for the reasons mentioned above. Thus, the RS were determined in field under indoor air conditions using only four SPMD samplers (instead of 24) and two MeVols (instead of six) with pump rates of 7 and 14 m3 ⋅ d-1, respectively, collecting the gas and particular phase separately. The indoor air sampling was performed in a PCB-contaminated part of a laboratory building that were closed for public use, where both the air concentration of PCBs and the ambient temperature were considered to be quite constant. It is important that the temperature is constant during the calibration since this variable can affect the amount taken up in SPMDs in complex ways. At intermediate and warm temperatures, the RS of compounds in SPMDs from water will increase with increasing temperature (Booij et al., 2003; Huckins et al., 1999), while Ockenden et al. (1998a) suggested that the uptake in SPMDs from air are higher at cold compared to intermediate/warm temperatures. The temperature also affects the partitioning between the gas-phase and particles, and thus the amounts taken up in the SPMDs. Increasing temperature increases vaporization and thus the partitioning into the gas-phase, while decreasing temperature increases condensation onto particles which increases the adsorption of compounds like PAHs to particles. Furthermore, the ambient air concentrations vary differently with temperature depending on the source of contamination. For instance, PAH-emissions from traffic are highest during daytime, while emissions from wood combustions probably increasing during the night.

In this calibration test the average air temperature was 22 °C with a standard deviation of 2.3. The average amounts in standard 1-mL SPMDs (n=4), the gas phase concentrations measured with MeVols and the exposure time (11 3 -1 days) were used in eq. 10 to calculate the RS in m ⋅ d for each compound. The RS of the sum of each congener group (di- to octa) for standard 1-mL 3 -1 SPMDs ranged from 2.5 to 4.5 m ⋅ d (Figure 12). Thus, the RS of the PCB congener groups were rather similar, as expected since all PCB compounds

(log KOW > 5) were probably under boundary layer control. These results are in agreement with those of Shoeib and Harner (2002) who reported indoor- calibrated RS of PCBs, at the same temperature, to be usually in the range 3 to 5, and differed somewhat from outdoor-calibrated RS of PCBs, at an average

5. Factors influencing the uptake of compounds in SPMDs from air temperature of 18 °C (range 7-31 °C) reported by Ockenden et al. (1998a). Our results are also in agreement with Bartkow et al. (2004) who found similar RS of 12 EPA PAHs (fluorene to benzo[g,h,i]perylene) in air (range 0.6 3 -1 to 6.1 m ⋅ d ). However, the concentration capacity (the KSA) of the SPMDs, which was not tested in this work, will vary between each congener group. Shoeib and Harner (2002) reported that equilibrium of PCB 28 (tri-CBs) and PCB 52 (tetra-CB) was reached after 50-75 and 200-400 days, respectively, while the uptake of PCBs with KOA values > 9 like PCBs 101 (penta-CB), 137 and 138 (hexa-CBs) was linear throughout the entire 450-day sampling period. These results show that all PCB congeners tested in the work underlying this thesis, were in the linear uptake phase during the entire 11-day calibration.

8,0

6,0 ) -1 d 3 4,0 (m S R

2,0

0,0 DiCBs TriCBs TetraCBs PentaCBs HexaCBs HeptaCBs OctaCBs

3 -1 Figure 12. The RS (m ⋅ d ) of the sum of each PCB congener group at 22 °C calculated from the average amounts (ng) in standard 1-mL SPMDs (n=4), and the gas phase concentrations (ng ⋅ m-3) measured with active medium volume sampler (MeVols) during a 11 days indoor air sampling.

5.2 Environmental variables

According to European standards for diffusive samplers of gases and vapours the effect of the environmental variables relative humidity, temperature and air velocity on the sampler performance (uptake rate) should all be tested (Anonymous, 2002b). Temperature and wind-speed/turbulence should have an effect on the uptake in SPMDs from air. In addition, chemicals bound to particles or aerosols can be trapped on the membrane surfaces and affect the amount taken up in the SPMD, and the UV light can decrease the amount found in the SPMDs via photodegradation in air. The effect of environmental variables on the uptake of compounds in SPMDs under ambient air

5. Factors influencing the uptake of compounds in SPMDs from air conditions (excluding the effects of these factors under indoor air conditions) are discussed in more detail below.

5.2.1 Wind-speed/turbulence

As previously mentioned, compounds with high log KOW values (KOW values

> approximately 4.4) have high KSPMD values and, thus, rapidly partition to the membrane, and research have indicate that the uptake is restricted by a boundary or diffusion layer at the membrane-water interface (Booij et al., 1998; Huckins et al., 1999; Vrana and Schuurmann, 2002). Thus, the thickness of the boundary layer and, consequently, the water flow/turbulence, affect the uptake of these compounds, as were reported by Booij et al. (1998) and Vrana and Schuurmann (2002). Vrana and Schuurmann (2002), for instance, found that at low flow rates (0.06 to 0.28 cm ⋅ s-1), the uptake of hydrophobic compounds (log KOW > 4) increased with increasing flow rate, but at intermediate rates (0.28-1.14 cm ⋅ s-1) no significant increases in uptake of the compounds were observed. In addition, Booij et al. (1998) reported three times higher uptake rates of PAHs and PCBs under conditions of high compared to low water flow/turbulence.

Although the diffusive mass transfer is much faster in air than in water, the boundary layer is probably thicker and thus, wind-speeds/turbulence should have an effect on the uptake in air as well. Figure 13 illustrates how the wind- speeds/turbulence exposure affects the thickness of the boundary layer, and thus the uptake of PAS. At extremely low wind-speeds/turbulence, when the whole air compartment is quite still, the diffusive mass transfer in the thick boundary layer is highly restricted and the turbulent transfer in low. Under these conditions, the mass transfer to the sampler is so slow that a depletion of the air at the boundary layer-air interface occurs. At increasing wind- speeds/turbulence the thickness of the boundary layer decreases, and the turbulent transfer in the air compartment increases. Thus, the uptake in the sampler will increase. Under intermediate wind conditions, the boundary layer and the turbulent transfer are quite constant, resulting in approximately constant uptake in the sampler. At extremely high wind-speeds, as were tested in the work underlying this thesis, the uptake increases with increasing wind- speeds/turbulence since the turbulent transfer is high, and the thickness of the boundary layer decreases significantly, which reduces the restriction to the diffusive mass transfer in the boundary layer, so for SPMDs the relative degree of control imposed by the membrane should progressively increase. In

5. Factors influencing the uptake of compounds in SPMDs from air the most extreme wind-speeds/turbulence exposure cases, the boundary layer can be eliminated and the uptake in SPMDs will then be solely controlled by the membrane. Thus, the resistance to mass transfer in the boundary layer can be tested by comparing the uptake in samplers at low and high air flows with the uptake at intermediate wind-speeds/turbulence. In studies reported in Papers II and III, the hypothesis that high wind-speeds (range 6-50 m ⋅ s-1) increase the uptakes of PAHs and PCBs in SPMDs from air was tested.

Increase Constant Increase (thick (constant (thin Uptake in the sampler (ng) boundary boundary layer) boundary layer) layer)

-1 Wind-speed/turbulence exposure (m ⋅ s ) Figure 13. Correlation between wind-speeds/turbulence exposure, the thickness of the boundary layer, and uptake in the PAS (modified from Anonymous, 2002b).

5.2.1.1 The wind effect test method

Considering the aforementioned difficulties, we concluded that the European standard conditions for testing the wind effects on SPMD sampling of nonpolar aromatic compounds in air could not be met. Instead, we tested the wind effect in field, where the air volume capacity will be unlimited, and since no point sources of PAHs and PCBs were present at the site, the gas concentrations were reasonably constant between sampling points and over time. More specifically, in the wind effect tests four (Paper II) or five (Paper III) SPMD samplers were deployed simultaneously at different positions inside a wind tunnel (Figure 14), housed in a large dark, un-heated building. Each SPMD used was spiked with PRCs, as described in Papers II and III. Two additional samplers, one deployed outside the wind tunnel (C1) and one outside the building (C2), were used as controls of the uptake in SPMDs from the almost still air inside the building and the uptake at ambient air conditions, respectively. The wind tunnel was built with two compartments to produce different wind-speeds in different parts of the tunnel. An electrical

5. Factors influencing the uptake of compounds in SPMDs from air fan, mounted in the outside wall of the building at the inlet of the wind tunnel, produced constant high wind flow during each sampling period. The wind-speeds could not be measured at specific sampling sites due to high turbulence in the wind tunnel, but the wind-speeds in the inlet (1400 × 1300 mm) and the outlet (400 × 500 mm) of the wind tunnel were calculated to be 6 and 50 m ⋅ s-1, respectively.

Figure 14. Experimental setup of the works described in Papers II and III, including the two samplers used as controls of the uptake in SPMDs outside the wind tunnel (C1) and outside the building (C2), the wind tunnel with two compartments (1300 × 7000 mm and 500 × 2500 mm, respectively), the electrical fan (800 mm i.d.) mounted in the outside wall of the building by the inlet of the wind tunnel*, and four (Paper II) or five (Paper III) SPMD samplers inside the wind tunnel. One SPMD sampler was deployed in the smaller (500 × 2500 mm) compartment of the wind-tunnel, and the remaining three (Paper II) or four (Paper III) were distributed at equal distance in the larger compartment of the wind-tunnel.

Since the tests were performed at ambient air conditions, the air temperatures varied during each sampling period. However, measurements of the air temperatures showed that they varied similarly between sites, and at each sampling time, with average temperatures (measured in the unheated building, in the wind tunnel and outdoors) being 15 °C in Paper II and 9 °C in Paper III. Thus, the samplers in each study were exposed to very similar temperature variations, and the air temperature should not have influenced the differences in uptake between SPMDs. In conclusion, the use of this method, which combined laboratory and field test conditions, fulfilled many of the requirements of European standards (Anonymous, 2002b) although the gas concentrations were not verified with any reference sampling methods

5. Factors influencing the uptake of compounds in SPMDs from air or an independently calibrated measurement, and the wind effect was only qualitatively determined.

5.2.1.2 Wind effect at high wind-speeds

After the 18-day (Paper II) or 21-day (Paper III) sampling to test the wind effect on the uptake in SPMDs, the samples were analyzed for the PAHs and PCBs listed in Papers II and III, respectively. The total amount of PAHs taken up in the SPMDs in the wind tunnel were about two times higher than the amount found in the SPMDs outside the wind tunnel (Figure 15). Similar differences were observed for the uptake of PCBs, with about two to three times higher total amounts of PCBs found in the SPMDs placed inside than those outside the wind tunnel (Figure 15). There were differences in the uptake of PAHs and PCBs inside the wind tunnel as well, with highest uptakes observed at sites one (where turbulence was probably highest) and four (Paper II)/five (Paper III) (where the calculated wind-speed was highest). Highest uptakes of PAHs and PCBs were therefore observed in those SPMDs exposed to the highest (assumed) turbulence/wind-speeds, supporting the hypothesis that increasing wind-speeds increase the uptake in SPMDs from air. Thus, the work described in this thesis showed that, in air, the uptake of most nonpolar aromatic compounds (represented by the studied PAHs and PCBs) in SPMDs is controlled by the boundary layer at the membrane-air interface. However, a wind effect on the uptake in SPMDs was not observed for all target compounds in the work of Papers II and III. The uptakes of acenaphthene and fluorene, which had the lowest log KOA values of the compounds studied, were similar both inside and outside the wind- tunnel, while the uptake of phenanthrene, with a log KOA value of 7.9 (Table 2, pp 55) (log KOW value of 4.5 (Mackay, 1992)) was higher in SPMDs inside than outside the wind-tunnel. These results demonstrate that, in air, the uptake of compounds with log KOA > 7.9 is controlled by the boundary layer. To our knowledge, no wind effect on the uptake in SPMDs from air, and thus no boundary layer control of the uptake of nonpolar air pollutants in SPMDs, had been demonstrated prior to this work. However, Harner et al. (2003) found that high wind increased the uptake of PCBs in unsheltered POGs. For instance, the time to reach equilibrium for PCB 28 was five times higher in still conditions than with an air flow of 4 m ⋅ s-1 in the cited study.

5. Factors influencing the uptake of compounds in SPMDs from air

1 1

) ) -1 d -1 0,8 0,8 d

0,6 0,6

0,4 0,4

0,2 0,2

PCB amount (ng SPMD (ng amount PCB PAH amount (ng SPMD (ng amount PAH a c a,b c 0 0 a) C1C212345 b) C1 C2 1 2 3 4 5

Figure 15. Total amounts (ng ⋅ SPMD-1) of a) PAHs and b) PCBs taken up in the replicate SPMDs in Paper III (first and second bars) and mean values of the amounts taken up in the two SPMDs in Paper II (third bar). aThe first SPMD of C1 was excluded due to sample losses during the GPC-analysis and bthe second SPMD was excluded due to analytical interferences (these results can be found in Paper III). cIn Paper II only four samplers were used. The fourth sampler in Paper II was deployed at the same place as the fifth sampler in Paper III and thus, these data are compared. dLow SPMD amounts probably due to damaged membrane.

The results reported in Paper II, showed that the differences in uptake between the SPMDs inside and outside the wind tunnel increased with increasing log KOA of PAHs up to 8.74. A possible explanation for these results is that as hydrophobicity (and thus log KOA values) increases, the restriction to diffusive mass transfer in the often aqueous boundary layer (condensation at surfaces) increases, and thus high wind-speed in the wind tunnel (which considerably reduces the thickness of the boundary layer) had a greater effect on the uptake of more hydrophobic compounds. These correlations were not observed for the PAHs with higher log KOA values, which are mainly bound to particles, or for the PCBs. These results are not in agreement with those found by Ockenden et al. (2001) who reported that the wind effect on the uptake of PCBs (demonstrated by comparing the uptake in protected SPMD samplers and those exposed to wind) decreased with increasing chlorination (from tri- to octa-CBs). Their results indicate that the uptake of more hydrophobic and larger PCBs is less affected by changes in the thickness of the boundary layer, but this seems unlikely.

The SPMDs in the work presented in Papers II and III were spiked with three 2H-PAHs and four 2H-PAHs together with four 13C-PCBs, respectively, which were used to control the release at different wind-speeds. In all SPMD

5. Factors influencing the uptake of compounds in SPMDs from air samples the levels of the PRC 2H-acenaphthene were below detection limit (< 3 × noise level). Thus, further discussion will only concern Paper III due to the limited amount of PRC data in Paper II. In Paper III, the releases of the remaining seven PRCs from the SPMDs inside the wind tunnel were up to four times higher compared to the release from the SPMDs outside the wind tunnel, and the highest amount were released from the SPMDs at site five, where the calculated wind-speeds was highest. Highest loss rates were therefore observed from the SPMDs exposed to the highest calculated wind- speeds, demonstrating that the releases of PRCs were also restricted by the boundary layer at the membrane-air interface.

5.2.1.3 Wind effect in field

A wind effect on the uptake and release in sheltered SPMDs, exposed to extremely high wind-speeds/turbulence (6 to 50 m ⋅ s-1), was found (Papers II and III). However, the wind-speeds/turbulence in the field will be more intermediate, and less variable between locations. For example, the monthly average ambient wind-speeds in Europe generally range between 1 and 10 m ⋅ s-1 (Anonymous, 2003). Thus, the wind effect in the field will be smaller. For instance, in a field test by Ockenden et al. (2001) only small differences were found when the uptake of PCBs in wind-sheltered SPMD samplers and SPMD samplers exposed to wind were compared. Although the wind effect on the SPMD sampling in the field will be generally low, extremely high wind- speeds/turbulence can occur during sampling and, thus, the use of PRCs to predict the site effect of wind is important.

5.2.2 Temperature

In water, Booij et al. (2003) have reported that the average RS of PCBs, chlorobenzenes and PAHs were about three times higher at 30 °C than at 2

°C. Booij et al. (2003) also found a non-significant effect on the KSPMD at these temperature differences, whereas the LPDE-water partition coefficients were two times higher at 2 °C than at 30 °C. Huckins et al. (2002a) found that the temperature effect on the KSPMD appear to be molecular structure specific with a 2.6 fold differences in the KSPMD of phenanthrene measured at 8 and 30

°C, whereas no temperature effect was found on the measured KSPMD values of PCB 52 and p,p´-DDE. In addition, Huckins et al. (1999) reported a 1.5- fold increase in the RS of PAHs when the water temperature increased from

5. Factors influencing the uptake of compounds in SPMDs from air

10 to 26 °C. Thus, the temperature has a limited effect on the RS of compounds in SPMDs from water.

Few studies have tested the temperature effect on the uptake in SPMDs from air. Ockenden et al. (1998a) found that field-calibrated RS of PCBs were about four times higher in winter than in summer. In winter, the ambient temperatures ranged between -6 to 18 °C (mean 4 °C), while the temperature range in summer was 7-31 °C (mean 18 °C). This temperature effect has obviously never been observed on SPMD sampling in water, since this compartment will be frozen at water temperature below ∼ 0 °C. Ockenden et al. (1998a) suggested that the reason for these results was that at cold winter temperatures decreased the solubility in air more than the solubility in the lipid (triolein), and thus the lipid-air partition coefficient will increase. This hypothesis was supported by Harner et al. (2003) who suggested that the passive sampling medium-air partition coefficient (KPSM-A) will increase by a factor of 2.5 to 3 for every 10 °C decrease in temperature. However, an increase in the KSPMD will only increase the RS of compounds under membrane-control, whereas for compounds under boundary layer control, only the capacity of the SPMDs will increase with increasing KSPMD. Thus, the effect of cold temperatures is rather complex and further research is needed to understand the processes involved in the effect. It is clear though, that the cold temperature effect can have as strong influence on summer and winter data comparisons that are uncorrected for the temperature effect. For instance, the highest and lowest temperatures measured in Umeå, Sweden between 1996 and 2004 were 31.1 and -33.0 °C, respectively (Johansson, 2004). Research is also needed to assess the effects of intermediate and warm ambient air temperatures. However, as previously mentioned, the effect on the RS at intermediate and warm water temperatures is probably limited, unless data on large geographical scales are compared, and we assume that air temperatures above 0 °C affect the RS of compounds in air in a similar way as in water.

5.2.3 Particles

In the work underlying this thesis, the amounts found the SPMDs were expected to originate from air pollutants in the gas phase. However, compounds bound to particles can contribute to the amounts found in the SPMDs in several ways:

5. Factors influencing the uptake of compounds in SPMDs from air

I. Direct particle deposition on the exterior membrane surface during sampling can allow compounds bound to particles to desorbs and permeate through the membrane (particle-membrane transfer), and thus increase the amounts found in the SPMDs. II. Compounds bound to the particles at the exterior SPMD surface that were not accumulated in the SPMDs during sampling, can be extracted from the particles during dialysis instead and thus increase the amounts found in the SPMD samples. The degree of sorption to

particles can vary for compounds with similar log KOA values because of differences in their molecular structures. For instance, due to their planar molecular structure PAHs can probably interact with soot particles to a higher degree than PCBs and chlorobenzenes, which have a more spherical structure.

In addition, particles can give poor dialytical recoveries of the target analytes due to sorption to organic carbon (e.g., soot) at the exterior of the membrane surface (Huckins et al., 1999). Thus, the cleaning of the exterior membrane- surface prior to dialytical extraction is highly important since it reduces the uncertainties of the SPMD data. This procedure will not reduce the effect of particle-associated compounds that accumulate in the SPMDs during sampling. However, when the SPMDs are protected by shelters the amounts of particle-associated compounds taken up in the SPMDs should be limited. For instance, Ockenden et al. (1998a) found that in comparison with HiVols, more than 96 percent of the PCBs sampled by the SPMDs were associated with the gas phase.

The potential of SPMDs to sample particle-associated compounds has been investigated by not cleaning the exterior surface of the SPMD prior to dialytical extraction (Bartkow et al., 2004; Lohmann et al., 2001). Bartkow et al. (2004) found a good agreement between the air concentrations predicted from the SPMD data and those measured with HiVols for the PAHs that are mostly bound to particles. Lohmann et al. (2001) also detected PCDD/Fs and PAHs in the SPMDs that are typically at least 85 percent associated with particles, but predominantly gas phase compounds were found in the samples. The reason for these results, i.e. whether the particle-associated compounds in the samples originated from actual SPMD uptake of desorbed compounds or from dialytical extraction, was not clear, and further research is needed to understand the sampling processes of particle-associated compounds.

5. Factors influencing the uptake of compounds in SPMDs from air

5.2.4 UV sunlight

A fourth environmental variable that can affect the amount of analytes found in the SPMD air samples is UV sunlight. Field tests by Orazio et al. (2002) demonstrated that in SPMDs exposed to direct sunlight, some of the 16 EPA PAHs were photolyzed in just 2 minutes, while in SPMDs sheltered inside foil-topped canisters, for a one-week air sampling, the PAHs degraded severely. In addition, Bartkow et al. (2004) found higher losses of the 2H- PAHs anthracene and pyrene than expected, probably due to photodegradation during sampling, although the SPMDs (820 × 28 mm LDPE tubes filled with 1 mL of triolein) were covered by chambers. Thus, photodegradation of the compounds accumulated in the SPMDs may occur during both sampling and retrieval. It is therefore critical to ensure that the SPMDs are not exposed to any direct or reflected sunlight, and that the sampler design used reduces the UV sunlight sufficiently. In addition, the site effects of UV sunlight can vary significantly, both temporally and geographically. In Umeå, 2004, for instance, the UV radiation (W ⋅ m-2) on the 18th of January was measurable only between 10 am and 4 pm with a maximum below 75 W ⋅ m-2, while the UV radiation in the 15th of July was measurable from 1 am to 10 pm with maximum value around 750 W ⋅ m-2 (values above 300 W ⋅ m-2 mean that the sun is shining) (Johansson, 2004). It is therefore also important to choose sampling locations that are in the shade as much as possible.

6. Approaches to reduce the site effects of environmental variables

6. APPROACHES TO REDUCE THE SITE EFFECTS OF ENVIRONMENTAL VARIABLES

As previously discussed, several environmental variables can affect the uptake in SPMDs during in the field deployments and thus cause deviations from laboratory pre-calibrated RS of the target analytes. In order to compare temporally or geographically different samples it is important to reduce and assess the site effects of these environmental variables.

6.1 The use of shelter

In the work underlying this thesis, the Exposmeter air sampler, described in detail previously, was used (Figure 16). Included in the design of the sampler was a metal umbrella that was used to protect the SPMDs from direct exposure to wind, UV light, particle deposition and rain, but allowing air to pass under and around the SPMDs. Examples of other type of shelter designs that have been used to protect the SPMDs from direct exposure are aluminium boxes (dimensions 300 × 300 × 300 mm) called Stevenson screens (Ockenden et al., 2001) and metal boxes that measuring 110 (h) × 220 (l) × 65 (d) mm (Lohmann et al., 2001).

Figure 16. Design of the metal umbrella (320/160 mm i.d. × 220 mm h) used to shelter the SPMDs in the work underlying this thesis.

The work presented in this thesis tested whether the metal umbrella, designed to reduce the wind-speeds/turbulence inside the sampler, decreased the uptake and release rates in SPMDs at high wind-speeds/turbulence (Paper III). The hypothesis was tested by comparing the uptake and release in

6. Approaches to reduce the site effects of environmental variables unsheltered SPMDs and SPMDs sheltered by metal umbrellas at high wind- speeds/turbulence. In all, five samplers, each including two SPMDs inside a metal umbrella and one SPMD on top, outside the umbrella, were deployed simultaneously for 21 days in a wind tunnel with high air flow (Figure 14).

6.1.1 Reduction of the wind effect by shelter

Comparison between unsheltered SPMDs and SPMDs sheltered by metal umbrellas showed that the use of this shelter design reduced the total uptake of PAHs and PCBs by 38 and 55 percent on average, respectively (Figure 17). Furthermore, the release of 13C-PCB 15 and 2H-phenanthrene decreased by 0- 25 and 13-22 percent, respectively, when the SPMDs were protected by the metal umbrella. Thus, the uptake and release in the SPMDs decreased when they were sheltered by metal umbrellas, demonstrating that the use of the metal umbrella reduced the effect of high wind-speeds/turbulence. Furthermore, the difference between the lowest and highest amounts of PAHs and PCBs taken up in the SPMDs was reduced by 10 and 8 percent, respectively, when the SPMDs were covered with the metal umbrella. These results show that the metal umbrella also reduced the variability in wind effect between sites. However, this test was performed under extremely high and variable wind-speeds/turbulence. Under normal wind conditions, the rate and variation in wind inside the metal umbrella will be much lower, and thus the variability in uptake between sites will be even lower. These results are in agreement with Ockenden et al. (2001), who reported wind-speeds inside a so-called Stevenson’s screen that were below the detection limit (< 0.4 m · s- 1), while the wind-speeds outside the devices were 2.6 to 31.3 m · s-1. In addition, Harner et al. (2003) found that the effect of high wind was reduced to insignificant when POGs were sheltered by chambers.

6. Approaches to reduce the site effects of environmental variables

2500 125

) ) -1 -1 2000 100

1500 75

1000 50

500 25 PCB amount (ng SPMD (ng amount PCB PAH amount (ng SPMD (ng amount PAH 0 0 a) 12345 b) 12345

Figure 17. Total amounts (ng ⋅ SPMD-1) of a) PAHs and b) PCBs in SPMDs sheltered by metal umbrellas (average of two SPMD measurements) (left bars) and in unsheltered SPMDs (right bars) at high wind-speeds/turbulence (modified from Paper III).

The metal umbrella reduced the uptake of the native compounds more than corresponding releases of the PRCs. There are two possible reasons for these results and most likely both of them contribute to the differences in uptake and release between sheltered and unsheltered SPMDs. Firstly, the unprotected SPMDs may have been exposed to such high wind-speeds that the boundary layer at the membrane-air interfaces was very thin or non- existent. Under these sampling conditions the mass transfer of the native compounds in the bulk atmosphere was probably not restricted by the boundary layer and the native compounds were actively transferred to the membrane surface by the high air flow/turbulence. In contrast, the PRCs in the bulk lipid of the SPMD were transferred to the membrane surface by molecular diffusion. Since active transport is a faster transfer process than molecular diffusion, the amount of PRCs transferred to the membrane-lipid interface was probably lower than the amount of native compounds transferred to the membrane-air interface, and thus the PRC release rates could not fully reflect the increasing uptake rates observed in the unprotected SPMDs. Secondly, the unsheltered SPMDs may have been affected by particle deposition on the membrane surface to a much higher degree than the sheltered SPMDs, because of their direct exposure to high air flows. Under such conditions, the amount taken up in the unprotected SPMDs is not only affected by the high wind, but also by direct particle-membrane transfer of compounds: a process that cannot be reflected by the release of PRCs. This hypothesis was supported by the results since the contribution of

6. Approaches to reduce the site effects of environmental variables the 4- and 5-ring PAHs to the total uptake of the PAHs increased by 20 percent in unsheltered SPMDs. The effect of increasing direct particle deposition on the uptake of PCBs should be limited since these compounds are predominantly found in the gas phase. However, further research is needed to verify the anisotropy observed between release and uptake rates of unprotected SPMDs exposed to high wind, and to define the uptake and release mechanisms at these sampling conditions. In conclusion, the results of this study demonstrate that the uptake in unsheltered SPMDs was affected by the high wind, and probably also by particle-deposition, and that these effects were reduced by using the metal umbrella. However, the sampler design used in the work underlying this thesis is probably not optimal for reducing the effect of UV sunlight since the shelter consists of metal, which will reflect the sunlight inside the devices. One way to improve the design of the metal umbrella would be to cover the inside of the metal umbrella with a dark material.

6.2 The use of performance reference compounds (PRCs)

In water, several studies have indicated that PRCs can be used to assess the site effect of water flow/turbulence, temperature and biofouling (Huckins et al., 1993; Booij et al., 1998; Huckins et al., 2002a; 2002b; Vrana and Schuurmann, 2002; Booij et al., 2003). The use of PRCs to assess the site effect of wind-speed/turbulence on the uptake in SPMDs from air will be discussed below. The possibility of using PRCs to assess the effect of particles, temperature and photolysis will also be briefly discussed.

6.2.1 PRCs´ utility for assessing wind effect

The work described in Papers II and III tested if the releases of PRCs were related to the uptake of target analytes under variable wind conditions (from still air to an air flow of 50 m ⋅ s-1). The native target compounds were a number of PAHs and PCBs, while the 2H-PAHs acenaphthene, phenanthrene and pyrene (Paper II), and the 2H-PAHs acenaphthene, fluorene, phenanthrene and pyrene together with the 13C-PCBs 3, 15, 37 and 54 (Paper III), respectively, were used as PRCs. These labelled compounds were selected in order to test the use of PRCs covering as a wide range of vapour pressures as possible with releases that were suitable for different detectors and for sampling periods of varying duration, e.g., one week to several

6. Approaches to reduce the site effects of environmental variables months. After 18 days (Paper II) or 21 days (Paper III) sampling, the levels of the PRC 2H-acenaphthene was below its detection limit (< 3 × noise level) in all SPMDs. In Paper II, the precision of the PRC 2H-pyrene data was low due to analytical interferences in the HRGC-LRMS analysis. Thus, further discussion will only concern Paper III due to the limited amount PRC data in Paper II.

As mentioned previously, both the releases and the uptakes were highest by the SPMDs exposed to the highest calculated wind-speeds, demonstrating that both the uptake of the compounds, and the release of selected PRCs, were controlled by the boundary layer. Thus, the release rates of PRCs were related to the uptake rates of native compounds, showing that they could be used to assess the wind effect. These results are in agreement with Vrana and Schuurmann (2002) who reported that in water, the release of 2H-anthracene increased three-fold when the water flow increased five-fold from 0.06 to 0.28 cm ⋅ s-1. Also Booij et al. (1998) found higher releases of PCB 29, 2H- phenanthrene and 2H-chrysene at high water flow (30 cm ⋅ s-1) than under conditions of low flow (0.03 cm ⋅ s-1). However, the releases of PRCs were not able to fully assess the uptakes of compounds in unsheltered SPMDs probably due to the high turbulent transfer of the native compounds and high particle-deposition on the membrane surface that facilitate direct particle- membrane transfer (effects further discussed in 5.2.3). Thus, in order to use PRCs to predict the effects of high wind it is critical that the SPMDs are protected by metal umbrellas. In addition, both Papers II and III revealed problems with the precision of the PRCs due to analytical interferences. Booij et al. (1998) used PRCs and found that they allowed the uptake kinetics of PAHs to be predicted, but with poor precision due to analytical interferences. Thus, when using PRCs to calibrate the SPMD sampling in situ robust analytical quality control must be applied.

The releases of PRCs generally increased with decreasing log KOA of their analogous native compound (Figure 18). (Figure 18 also shows that the amount of the PRC released for each compound (log KOA value) was affected by the wind-speeds, see differences in the y-values in the figure). Hence, the

SPMD affinity, and thus the KSA, increased with the KOA values of the PRCs used. More specifically, acenaphthene, which had the lowest KOA value, was completely released before the sampling was ended, the releases of 2H- fluorene and 13C-PCB 3 were mostly > 80 percent, the releases of 2H- phenanthrene and 13C-PCB 15 were in most cases between 20 and 80 percent, while the releases of the 2H-pyrene and 13C-PCBs 54 and 37 were within the

6. Approaches to reduce the site effects of environmental variables

20-80 percent range or below 20 percent. These results are consistent with the findings of Ockenden et al. (2001) who observed losses of PCBs 28, 52, 101 and 153 that were 85, 50, 25 and 8 percent, respectively, and non-measurable losses of 13C-PCBs 138 and 180, after a four-month sampling campaign, due to their high KOA values. Consequently, acenaphthene or compounds with similar KOA values should be used as PRCs in short-time (days) deployments, for intermediate-time (weeks) deployments compounds like 2H-phenanthrene and 13C-PCBs 15 and 28 should be used, for longer-time deployments (months to years) compounds like 2H-pyrene and the 13C-PCBs 54 and 37 should be used, and for even longer sampling periods (years) compounds with slightly higher log KOA values should be used.

100

Amount released (%) released Amount 20

0 678910

log KOA

Figure 18. Average amount (%) of PRC released from the replicate SPMDs in the wind tunnel (filled circles), outside the wind tunnel [C1, (triangles)] and outside the building [C2 (squares)] as a function of the log octanol-air partition coefficient (log KOA) of their analogous native compound (Table 2) (Paper III).

The ranges of the ke values calculated using eq. 12, and the SPMD data with remaining PRC-amounts that were 20 to 80 percent of the initial PRC- 2 amounts, are presented in Table 2. The ke values of the H-PAHs were lower than those calculated at 22 °C and wind-speeds between 0.8 and 1.7 m ⋅ s-1 for the 2H-PAHs anthracene (0.090 d-1) and pyrene (0.094 d-1) by Bartkow et al. (2004). The cited authors suggested that their loss rates were too high. In addition, they found similar release rates of anthracene and pyrene, although these compounds differ in their log KOA values. As previously mentioned, they suggested that one possible reason for their results was photolysis of the

PRCs used during sampling. Ockenden et al. (2001) calculated the ke values for PCBs 28, 52, 101 and 153, at 15 ºC and wind-speeds in the range 2.6-31.3 m ⋅ s-1, to be 0.0085, 0.0047, 0.002 and 0.0007 d-1, respectively. Thus, our

6. Approaches to reduce the site effects of environmental variables

calculated ke values were about an order of magnitude higher than theirs. These results could be due to the wind-speeds being much higher during our study. Another possible explanation is that sampling devices with different designs were used, causing either the SPMDs in our study to be more directly exposed to the high air flow, or the SPMDs in their study to be exposed to very still air, and thus their loss rates to be too low.

Table 2. ke values of performance reference compounds (PRCs) (excluding acenaphthene) estimated in Paper III, and the log KOA values of their analogous native compounds. log KOA ke values (d-1)d 13C-PCBs PCB 3 7.0a 0.036-0.061 PCB 15 7.9a 0.018-0.060 PCB 37 9.0b 0.011-0.033 PCB 54 7.3b 0.016-0.042 2H-PAHs Fluorene 7.1c 0.036-0.063 Phenanthrene 7.9c 0.012-0.065 Pyrene 9.2c 0.011-0.018 Log octanol-air partition coefficient (log KOA) at 20 °C , ameasured by Harner and Bidleman (1996), cmeasured by Harner and Bidleman (1998), and bpredicted by Chen et al. (2002), respectively. dExchange rate constant at 9 °C and different wind-speeds.

Since we assumed that all SPMDs were exposed to the same air, the ability of the PRCs used to compensate for the differences in wind-speeds in the wind- tunnel was tested by calculating the air concentrations of some selected PAHs 2 13 and PCBs using the in situ calibrated ke values of H-phenanthrene and C- PCB 15, and eq. 13, and then comparing the variability of these values. There were several factors to consider in these calculations. First, the releases of the PRC 2H-phenanthrene were > 60 percent for the SPMDs inside the wind tunnel. Thus, the uptake of phenanthrene was nonlinear, and the use of the linear uptake model described in eq. 13 would cause the phenanthrene concentrations calculated to be underestimated. Second, the KSA values of the selected compounds were not available and their KOA values had to be used as surrogates. The sources of the surrogate PAH and PCB KOA values were Harner and Bidleman (1998), and Chen et al. (2002), respectively. Since the air concentrations were only semi-quantitatively determined (by using KOA values) data were normalized to the highest concentration determined for each compound (Figure 19).

The CV for the calculated PAH- and PCB-concentrations ranged between 28 to 41 and 28 to 46 percent, respectively (Figure 19). Thus, the use of PRCs to calibrate the ke values in situ and thereby correct for the site effect of wind reduced the between-site differences to less than 50 percent from as much as

6. Approaches to reduce the site effects of environmental variables two- and three-fold difference in the total amount of PCBs and PAHs, respectively, found in the SPMDs. If the use of PRCs had fully compensated for the effect of different wind-speeds, the quantified air concentrations would have been equal between sites. However, we consider that the CV (introduced by the 21-day sampling and the chemical analysis) of the estimated air concentrations were good. Thus, the utility of the PRCs used for assessing the site effect of wind was high.

Phenanthrene (CV: 28 %) PCB 52 (CV: 28 %) 1,0 1,0

0,8 0,8 0,6 * 0,6

0,4 0,4 * 0,2 0,2

0,0 0,0 12345 12345 Pyrene (CV: 29 %) PCB 80 (CV: 29 %) 1,0 1,0 0,8 * 0,8 0,6 0,6

0,4 0,4 * 0,2 0,2 0,0 0,0 12345 12345

Fluoranthene (CV: 41 %) PCB 101 (CV:46%) 1,0 1,0 0,8 0,8

0,6 * 0,6

0,4 0,4 0,2 0,2 * 0,0 0,0 12345 12345

Figure 19. Normalized air concentrations of six PAHs and PCBs estimated using the in situ calibrated ke values of 2H-phenanthrene and 13C- PCB 15, respectively, and the KOA values of PAHs and PCBs suggested by Harner and Bidleman (1998) and Chen et al. (2002), respectively, as surrogates for unavailable KSA values (Paper III). The coefficient of variance (CV), derived from 21-day sampling and chemical analysis, of the estimated air concentrations of each compound was shown in brackets. *Low uptake, and high ke values, probably due to membrane damaged.

6. Approaches to reduce the site effects of environmental variables

6.2.2 PRCs´ utility for assessing site effects in air

The work described in this thesis demonstrated the high potential of the PRC approach to assess the site effect of wind. In order to achieve high precision of the PRCs, the SPMDs have to be sheltered from direct wind exposure and robust analytical quality control should be applied. However, the reduced correlation between the releases of PRCs and the uptakes of native compounds in unsheltered SPMDs indicate that the effects of particles may not be predicted by the use of PRCs (Paper III). To our knowledge, the use of the PRC-approach to predict the effect of temperature and photolysis has not been tested. It has been suggested, however, that photosensitive compounds with high SPMD affinity should be added to the SPMD prior to membrane enclosure, to ensure that compounds accumulated in the SPMD do not photodegrade during sampling and retrieval (Bartkow et al., 2004). However, this approach will only prove if photodegradation occurred or not, rather than quantitatively determine its effects on the uptake of specific compounds.

7. Fields of application and sampling strategies

7. FIELDS OF APPLICATION AND SAMPLING STATEGIES

SPMDs have been used in a number of different applications and environments, but primarily in water (Table 1). The sampling strategy varies depending on the purpose of the study. The fields of application for which SPMDs have been used in the work described in this thesis are presented below, and the chosen sampling strategies are discussed.

7.1 Determination of local/regional atmospheric distribution

The atmospheric distribution of the gas phase PAHs in the Bangkok region, Thailand, was investigated in the work presented in Paper I. SPMD samplers were deployed for three weeks at six different locations (A-F). The sites were chosen in order to cover a wide range of pollution situations ranging from remote areas with background air quality to urban areas. The SPMD samples were analyzed for 14 of the 16 EPA PAHs (excluding naphthalene and acenaphthylene) together with benzo[e]pyrene and 1-methylphenanthrene. The total amounts of the PAHs (excluding 1-methylphenanthrene) found in the SPMDs varied between 17 and 134 ng ⋅ SPMD-1 ⋅ d-1, and were correlated with the assumed degree of pollution at the sampling sites (Figure 20). For example, SPMDs at the urban sites E and F sequestered about ten times more of the PAHs than those deployed at site A (rural area). The predominant individual PAH was phenanthrene, for which the amounts in the SPMDs ranged between 9.8 and 64 ng ⋅ SPMD-1 ⋅ d-1.

7. Fields of application and sampling strategies

160 ) -1 d -1 120

40 PAH amounts (ng SPMD 0 ABCDEF

Figure 20. Total amounts of 14 EPA PAHs (excluding naphthalene and acenaphthylene) taken up in SPMDs from the atmosphere of the Bangkok region, Thailand (pollution was assumed to increase from sites A to F) (modified from Paper I).

In the work reported in Paper IV, the distribution of the gas phase concentrations of PAHs and nitro-PAHs were monitored in five European countries (Austria, the Czech Republic, Poland, Slovakia and Sweden). The study is described in more detail below (see 7.2). Briefly, however, SPMD samplers were deployed during two three-week sampling periods at eight locations in each country. The SPMD data collected within each country differed 5- to 20-fold in PAH-levels, and 6- to 30-fold in nitro-PAH-levels. The highest levels were detected at sites defined as urban, while the lowest levels were detected at remote/rural sites, showing that the variation in the PAH- and nitro-PAH levels within each country followed the subjectively- defined pollution situations.

When using SPMDs to determine the distribution of gas phase concentrations at a local scale, the site effects of environmental variables are generally low, and the variations in the SPMD data reflect the spatial distribution. For example, the average sampling temperatures reported in Paper I was in the range 29.1 to 31.0 °C. However, when comparing winter and summer data, as mentioned above, the temperature effect on the uptake in SPMDs should be considered. For instance, the temporal differences in Poland were high, with uptake of PAH- and nitro-PAHs being one and two orders of magnitude higher, respectively, in winter (average 0.7 °C) than in summer (average 16 °C) (Paper IV). One possible reason for these results is

7. Fields of application and sampling strategies

that the RS of PAHs were probably higher in the winter than in the summer. However, Ockenden et al. (1998a) suggested that the RS of PCBs in summer and winter differed only about four times and thus, the temperature effect cannot be the main reason for the high temporal variations found in Poland. Instead, the main reason was probably that the emissions from coal combustion increased in the winter since residential heating was required when the air temperatures were cold. The results of the work underlying this thesis (Papers I and IV) demonstrate that SPMDs are efficient samplers of gas phase pollutants, and are able to determine the spatial distribution of gas phase PAHs in rather restricted areas. The high efficiency of SPMDs to detect differences in atmospheric concentrations at a local scale has also been reported by Lohmann et al. (2001). In addition, the use of the SPMD technique, which does not require electricity, made sampling possible at remote/rural areas where the infrastructure was limited, like the countryside in Thailand. However, due to the lack of calibration data at the time, the atmospheric concentrations could not be quantitatively determined. Instead, the RS values of PAHs in water calibrated by Huckins et al. (1999) were used in Paper I for preliminary calculation of the PAH concentrations in air.

7.2 Determination of continental atmospheric distribution

In the work of Paper IV, the gas phase PAHs and nitro-PAHs in air of five European countries (Austria, the Czech Republic, Poland, Slovakia and Sweden) were measured during two sampling campaigns in order to determine whether there were spatial differences between northern, central and eastern Europe. Prior to the air sampling campaigns, the research groups of each participating country received written sampling instructions and sampling equipment from the Environmental Chemistry, Umeå University, Umeå, Sweden. According to instructions, the SPMD samplers were deployed by each research group for two three-week periods, one in the autumn of 1999 and the other in the summer of 2000, except in Poland where four-week samples were taken (at different locations) in the winter of 1999 and the summer of 2000. In all, SPMD samplers were deployed at 40 locations; eight sites per country, at similar periods of time in two different seasons (Figure 21). The eight sampling sites were chosen by the research group from each country following the guidance to cover a wide range of pollutant situations from remote and rural sites to urban and industrial areas. The average sampling temperature of each location was measured in parallel and other environmental conditions such as wind-speed and wind direction were measured, when possible (data listed in Table 1, Paper IV). The SPMD

7. Fields of application and sampling strategies samples were analyzed for 13 of the 16 EPA PAHs (excluding acenaphthylene, fluoranthene and pyrene) and the six nitro-PAHs 1- nitronaphthalene (1-NN), 2-nitronaphthalene (2-NN), 2-nitrofluorene, 9- nitroanthracene, 3-nitrofluoranthene, and 1-nitropyrene. However, in 2000, all SPMDs were spiked with native fluorene, so this compound was not analyzed in these samples.

N N

W E W E

S S PL1-2 S

W PL1-2, 4-8 S PL3-6 W PL3 Warsaw S PL7-8

SE4, 6

POLAND SE2

SE8 CZ4 SE7 SE3 Prague CZ3 CZ6 CZ2 #S #S CZECH REPUBLIC CZ1 CZ5 SK1 CZ7 CZ8 SK8 SK6 SK2 SE1 SLOVAKIA SK3 SE5 AU5 AU7 AU4 Bratislava SK4-5 SWEDEN Vienna SK7 AU8 AU3 AU2 AUSTRIA AU1 AU6

Stockholm

0 100 200 300 400 500 Kilometers 0 100 200 300 400 500 Kilometers

Figure 21. Locations of the sampling sites in Sweden (SE), Austria (AU), the Czech Republic (CZ), Slovakia (SK) and Poland (winter 1999, W-PL, and summer 2000, S- PL) (Paper IV).

The total amounts of PAHs sequestered by the SPMDs in 1999 and 2000 ranged between 6.3-1200 and 5.0-1000 ng ⋅ SPMD-1 ⋅ d-1, respectively, and the total amounts of nitro-PAHs range between 0.036-4.0 and 0.0011-2.7 ng ⋅ SPMD-1 ⋅ d-1, respectively. Phenanthrene was the most abundant individual PAH found in the SPMDs, while the three nitro-PAHs 1-NN, 2-NN and 2- nitrofluorene dominated the nitro-PAH-patterns. These SPMD air data showed a wider range than those reported by Lohmann et al. (2001) who measured gas phase concentration of PAHs at 11 sites in the northwest of England. To our knowledge, passive sampling of nitro-PAHs using SPMDs has not previously been reported. However, the levels detected were about a tenth of the gas phase concentrations of 1-NN and 2-NN measured with an

7. Fields of application and sampling strategies active high-volume sampler system in the two cities Birmingham, UK, and Damascus, Syria (Dimashki et al., 2000).

The SPMDs deployed at 40 different locations from northern to central and eastern Europe showed that the levels measured in East Europe (Poland, Slovakia, and the Czech Republic) were about ten times higher than those found in the sampled areas of northern Europe (Sweden), whereas the measured levels of gas phase nitro-PAHs were more than ten times higher than the levels measured in northern (Sweden) and central (Austria) Europe. Thus, the spatial differences in the gas phase concentrations of PAHs and nitro-PAHs between the five countries were high, and the continental distribution of nitro-PAH and PAH-levels were similar. The spatial distribution of gas phase PAHs in Europe had not been measured with SPMDs prior to this study. However, the spatial variations in both PAH- and nitro-PAH concentrations found were similar to the variations in atmospheric PAH-concentrations measured across Europe by Jaward et al. (2004b) using PUF disks.

The differences in the individual and total PAH-levels found in the SPMDs deployed at the 40 locations were visualized in a score plot of the first two PCs (explaining 74 and 9.6 percent, respectively, of the variance) of a PCA model presented in the work of Paper IV (Figure 22). On the left-hand side of the score plot, sampling sites in Sweden and Austria were predominantly found, whereas sampling sites of the Czech Republic and Slovakia were found on the right-hand side. The corresponding loading plot revealed that at the sampling sites to the right in the score plot, viz. East Europe (the Czech Republic (CZ), Slovakia (SK) and Poland 1999 (PL-W)), the levels of e.g., phenanthrene, chrysene and the total PAH-s found in the SPMDs were higher than those found in Sweden, Austria and Poland 2000 (PL-S).

7. Fields of application and sampling strategies

AU2-00AU6-00 AU3-00 SE2-00 2

PL5-S SK6-00 CZ5-00CZ1-99 1 PL3-SPL7-S AU4-00SK7-00CZ6-99 SK3-99 SE5-00PL6-SPL4-S SK7-99CZ7-00 CZ2-99 PL1-SSE4-00 PL2-S SK2-99AU3-99 SK3-00 SE3-00 SK8-99 SK8-00SE1-00 PL3-W SE6-00 AU7-00 AU1-00SK5-99 PL2-WSK1-00SK6-99 0 SE7-00AU5-00 PL8-S AU4-99 CZ8-00CZ6-00PL6-WCZ2-00CZ1-00 PC2 AU8-00 CZ5-99CZ4-99CZ3-99SK1-99PL4-WSK4-00CZ4-00 SK5-00 CZ8-99 PL5-WSK4-99 CZ7-99 PL7-W AU6-99 SK2-00 -1 PL8-W SE8-00 AU8-99AU7-99 SE5-99SE1-99 PL1-W SE4-99AU1-99SE3-99 SE2-99AU5-99 CZ3-00 -2 SE7-99SE6-99 SE8-99

-7 -6 -5 -4 -3 -2 -1 0 1 2 3 4 5 6 7 PC1 Figure 22. Score plot of the first two principal components (PCs) of a PCA model visualizing the correlation between the individual and total PAH-levels found in the SPMDs deployed in Sweden (SW), Austria (AU), Poland (PL-W and PL-S), Slovakia (SK) and the Czech Republic (CZ) during the 1999 and 2000 sampling campaigns (Paper IV). The corresponding loading plot shows that at the sampling sites to the right in the score plot the levels of e.g., phenanthrene, chrysene and the total amounts of PAHs were higher than those found at the sites to the left in the score plot.

The measured weather data reported in Paper IV (Table 1, Paper IV) showed that for each sampling campaign the environmental sampling conditions like air temperatures were generally similar between the sampling locations. Thus, the site effect on the uptake in the SPMDs was limited, and the variations in the SPMD data reflected the spatial variations between the five European countries. These results demonstrate that SPMDs are suitable for integrative sampling of the atmospheric distribution of nonpolar air compounds like PAHs at a continental scale. Results of Paper IV also demonstrates the advantage of using SPMDs in monitoring programs compared to conventionally used HiVols. The SPMD´s low cost, and no need of special trained personal or maintenance during sampling made it possible to accomplish a three-week integrative sampling at 40 locations during two sampling campaigns collecting a total of 80 samples. Such numbers of samples can give a much more complete picture of the spatial distribution of POPs in air across Europe than, for instance, the eight sites distributed in six countries that were included in the air monitoring of POPs in air and aerosol co-ordinated by CCC/EMEP in 2002 (Aas and Breivik, 2004). The utility of SPMDs to measure the continental distribution has been shown in two other studies by Ockenden et al. (1998c) and Meijer et al. (2003) where SPMDs were deployed at 10 and 11 remote sites from the south of the UK to the

7. Fields of application and sampling strategies north of Norway for two two-year sampling campaigns in the periods 1994 - 1996 and 1998 - 2000, respectively, in which PCBs and chlorinated pesticides were measured. However, to date, most SPMD air data have been interpreted by comparing the differences in amounts and profiles found in the SPMDs. To further improve the utility of SPMDs it is important that PRCs are used in the sampling protocols, and calibration data are further developed so that the spatial distribution at both local and global scales can be quantitatively determined.

7.3 Determination of pollution sources

SPMDs have been used to determine the source of pollution in several studies (Table 1). Different approaches to determine the source of the PAHs in air measured by SPMDs are discussed below.

7.3.1 The use of individual compound ratios

One of the approaches used to determine the source of pollution in the atmosphere is to compare levels of a few individual compounds. In the work underlying this thesis, phenanthrene, fluoranthene and pyrene, respectively, were the most abundant individual PAH compounds found in the SPMDs. The ratio of fluoranthene to pyrene has been suggested as a useful indicator of traffic as the source to the atmospheric concentrations of PAHs measured with HiVols (e.g., Yunker et al., 2002). However, when testing this approach in Paper I, where the PAH-levels in air at different levels of traffic intensity were sampled, the fluoranthene/pyrene ratios were not correlated with the assumed degree of traffic intensity, probably because this approach was developed for HiVols, which sample both gas and particle concentrations. Other approaches that have been suggested to detect if traffic is the major source of atmospheric PAH-levels are to use ratios of the particle associated PAHs benzo[g,h,i]perylene to benzo[e]pyrene and coronene to benzo[e]pyrene (Lim et al., 1999; Nielsen, 1996), or to use the benzo[e]pyrene levels as indicators (Müller et al., 1998). Thus, many of the PAH-ratios used today to detect if motor vehicles are the major PAH source are most suitable for HiVols, which sample the total atmospheric concentration (both the gas and the particle phase).

7. Fields of application and sampling strategies

PAS predominantly sample compounds in the gas phase and, due to the size limitation in the membrane of the SPMDs, only the primary bioavailable compounds can accumulate in the SPMDs. Thus, new indicative individual compound ratios have to be found that are suitable for SPMD sampling. When selecting suitable compound ratios for SPMD sampling it is important to consider that the RS of compounds is affected by their physical/chemical properties. Phenanthrene accounted for about 30 to 50 percent of the total PAH amounts found in the SPMDs (Papers I-V). High concentrations of alkyl-PAHs compared to unsubstituted PAHs have been suggested to indicate the presence of petrogenic sources (e.g., Ramdahl, 1983; Rogge et al., 1993;

Nielsen, 1996; Lim et al., 1999, McDonald et al., 2000). Thus, since the RS values of phenanthrene and methylphenanthrene are quite similar (4.46 vs 5.14 L water ⋅ d-1 (Luellen and Shea, 2002), the methylphenanthrene/phenanthrene ratio is a possible indicator of the contribution of traffic to PAH concentrations in air sampled with SPMDs. In the work of Paper I it was shown that the 1- methylphenanthrene/phenanthrene ratio increased with increasing total PAH amounts in the SPMDs, and assumed traffic intensities (Figure 23). These results demonstrate that the 1-methylphenanthrene/phenanthrene ratio can be used to determine whether motor vehicles contributed to the atmospheric PAH-levels measured with SPMDs. However, when using this indicator it is important to consider the residential heating systems used in the sampling area, since the use of fossil fuels for home heating can also cause the concentrations of methylated PAHs to increase relative to levels of unsubstituted PAHs. However, the PAH emissions attributable to this source are negligible in Thailand, and the 1-methylphenanthrene/phenanthrene ratio can be used as an indicator of traffic’s contribution to the PAHs found in the SPMDs in such cases.

7. Fields of application and sampling strategies

160 ) -1 d -1 120

PAH amounts (ng SPMD (ng amounts PAH 0 0,00 0,25 0,50 0,75 1,00 1-methylphenanthrene/phenanthrene ratio

Figure 23. 1-methylphenanthrene/phenanthrene ratio versus the sum of 14 EPA PAHs (excluding naphthalene, acenaphthylene) together with benzo[e]pyrene in SPMDs at sampling sites A-F (Paper I).

In the work reported in Paper IV, the atmospheric levels of the gas phase nitro-PAHs were measured. Feilberg et al. (1999) analyzed diesel exhaust for 1-NN and 2-NN, and found only 1-NN in the samples, whereas Atkinson et al. (1987) found that gas phase reactions between naphthalene initiated by OH radicals produced 1-NN and 2-NN in nearly equal abundance. Thus, the 1-NN to 2-NN concentration ratio can be used to assess whether direct emissions or gas phase reactions produced the atmospheric concentrations of nitro-PAHs. The 1-NN and 2-NN ratios found in the SPMD samples reported in Paper IV were generally close to unity, or between one and two, which indicate that these nitro-PAHs mainly originated from atmospheric formation in the gas phase. Thus, this approach can be used to determine whether direct emissions or gas phase reactions contribute to the levels of 1- NN and 2-NN measured with SPMDs.

7.3.1.1 Considerations involved with the use of the individual compound ratios

When comparing the ratios of individual compounds between different samples, the results can be affected by the purity of the samples, even if internal standards and selective analytical methods have been used. The samples in Paper I were re-run on GPC due to interferences with the pyrene and fluoranthene analysis. The 1-methylphenanthrene/phenanthrene ratios after the 1st and 2nd GPC analyses displayed a linear relationship (R2 = 0.996), but they were twice as high (slope = 2.14) after the 2nd GPC analyses,

7. Fields of application and sampling strategies demonstrating that the ratios between individual compounds are not necessarily constant. Photolysis can also affect the detected ratios. The ratios between 1-NN and 2-NN found in the samples stored for a considerably longer time before analysis were slightly lower than in the other samples (Paper IV). Feilberg et al. (1999) have reported that photodegradation proceeds more rapidly for 1-NN than for 2-NN. Thus, the lower ratios found in some of the samples could have been caused by higher degradation of 1- NN during storage. Unfortunately only 1-nitropyrene and 6-nitrochrysene were used as internal standards, but the results demonstrate the importance of using corresponding labelled standards for each of the photosensitive compounds like nitro-PAHs, to assure the quality of the data. Thus, the ratios of individual compounds between samples are not always stable for the reasons mentioned above.

7.3.2 The use of total PAH-patterns

The total PAH-patterns found in the SPMD samples are probably more stable between samples than the ratios of individual PAH compounds. In work presented in Paper IV, sampling locations influenced by different potential sources of PAH pollution were included in order to discern, if possible, the PAH-pattern of specific sources. As previously described, five European countries were included in this study. In each country, eight locations were selected as sampling sites. Participants in each country were recommended to choose sampling locations (i) in the vicinity and upwind of at least three different potential sources of PAHs, (ii) at sites in rural areas and (iii) at one remote site. The pollution situation and the potential sources at each sampling site are described in detail in Table 1, Paper IV. However, the degree of air pollution varied between countries and the definitions of remote, rural and urban areas for each country were somewhat subjective. Thus, when comparing the data between countries, measured levels were not always comparable with the subjectively defined pollutant situations of the respective sampling locations. However, within each country the highest levels were detected at sites defined as urban, while the lowest levels occurred at remote and rural sites. According to the site descriptions in Table 1, Paper IV, the sources likely to be responsible for the high levels of PAHs and nitro- PAHs detected at the urban and industrial sites were traffic, metal-, chemical-, and petrochemical-industries, tar production, coal-fired systems for residential heating and other types of local combustion systems, which are considered as general sources of atmospheric PAH-pollution.

7. Fields of application and sampling strategies

To visualize the PAH-pattern of each sampling site, and thus attempt to find correlations between PAH-patterns and specific sources, a PCA model was fitted to the dataset normalized to the total amount of PAHs found at each site to exclude the effect of the concentration differences on the results. The score plot of the first two PCs showed that there were differences in the PAH-patterns of each site, while there were no differences in the PAH- patterns between countries or laboratories, demonstrating that both the sampling and the clean-up protocols of the participating research groups were comparable (Figure 4, Paper IV). However, no correlation between specific PAH-patterns and the sources described were detected. For example, sites described as highly influenced by traffic were equally distributed in the score plot. There are several possible reasons for these results. Firstly, the SPMD samplers were located about 1 km from the point source and thus, imissions rather than the emissions of the potential sources were probably measured. When imissions are measured it is difficult to avoid sources being confounded, due to the high number of diffusive PAH sources, and the site descriptions (Table 1, Paper IV) showed that the PAH-levels found at many sites could originate from multiple sources. Secondly, it was difficult to control the environmental conditions and the pollution situations during sampling, and although the intention was to take measurements upwind of described sources, additional unknown sources could also have influenced the PAH-levels of each site. Thirdly, volatilization of PAHs from soil could have affected the results since the samplers were deployed at 1.5 m height. Thus, there are many diffusive sources that can affect the atmospheric PAH concentrations, make specific source-patterns of PAHs very difficult to discern. In addition, the SPMDs sample mainly the gas phase PAHs which can lower the correlation between sources and PAH data. For instance, Müller et al. (1998) found that the correlations between traffic and pollutant data were better for PAHs with 4-rings or more, whereas the correlations for the gas phase PAHs were generally low.

7.3.3 Determination of residential wood burning as a source of indoor PAH exposure

Wood burning is considered to be the most important source of PAH pollution in the air in Sweden, accounting for about 60 percent of the total PAH emissions, while the contribution of traffic is about 30 percent (Boström et al., 2002). Thus, wood burning for residential heating is probably a major source of PAHs in indoor air of Swedish households. In the work reported in Paper V, the PAH-levels in the indoor environments of some

7. Fields of application and sampling strategies

Swedish households were measured with the aim to estimate if residential wood burning was a source of gas phase PAHs in indoor air. Briefly, SPMDs (deployed on spiders but not covered by metal umbrellas due to the low air flow indoors) were placed for two weeks in 16 single-family houses, in the same residential area, during the winter period 2003 in the small Swedish town Hagfors, which is situated in mid-Sweden. Nine of the households (w1- w9) used wood burning for residential heating, while seven of the households (nw1-nw7) used other heating systems (electrical or heating pumps). Cigarette smoking and cooking are other PAH sources that can also contribute to the PAH pollution of indoor air (Dermentzoglou et al., 2003; Endo et al., 2000). Thus, households were recommended to avoid smoking indoor during the sampling period. In addition, it was documented by the participating households how long time they fried food. Samples were analyzed for 15 of the 16 EPA PAHs (excluding naphthalene), retene and the seven methyl- PAHs 2-methylnaphthalene, 1-methylnaphthalene, 2,3-dimethylnaphthalene, 2,3,5,-trimethylnaphthalene, 3-methylphenanthrene, 4,9-methylphenanthrene and 1-methylphenanthrene.

The total amounts of the 15 EPA PAHs and the seven methyl-PAHs were in the ranges 22-400 and 16-320 ng ⋅ SPMD-1 ⋅ d-1, respectively (Figure 24). Thus, the PAH-levels detected were similar or slightly higher than those found in the SPMDs (range 4.9-170 ng ⋅ SPMD-1 ⋅ d-1) deployed outdoors at background to urban sites in Sweden during two sampling campaigns, one in 1999 and the other in 2000 (Paper IV). The five highest PAH-levels were found in SPMDs deployed in houses that used wood burning for residential heating. In these houses, the PAH-levels detected were three to six times higher than the lowest levels detected in the other houses. Thus, the PAH- levels in 11 of the 16 houses were similar and rather low and may represent the background PAH-levels of the indoor air in the area. Four of the five households (w2, 3, 7 and 8) with the highest levels of gas phase PAHs were situated in the same block. It is therefore not clear whether the gas-phase PAHs in the indoor air of these households originated from high PAH- emissions indoors, or if PAH-emissions derived from wood burning outdoors were transferred to the indoor environment via ventilation. However, the results showed that the use of wood burning for residential heating can increase the gas phase concentrations of PAHs in indoor environments even though there were households which used wood burning for residential heating that had background indoor PAH-levels. One possible explanation for the differences in measured levels between houses with wood burning could be that different types (and/or age) of wood was burned, or the

7. Fields of application and sampling strategies operational procedures were different between households. Nevertheless, the results demonstrate that the use of SPMDs that are not covered by metal umbrellas, is a highly sensitive sampling approach, and can be used to determine spatial differences in indoor air concentrations, even though the air is quite still in indoor environments.

450 )

-1 360 d -1

270

180

PAH amountsPAH (ng SPMD 90

0 w1 w2 w3 w4 w5 w6 w7 w8 w9 nw1 nw2 nw3 nw4 nw5 nw6 nw7

Figure 24. Total amounts of 15 EPA PAHs (excluding naphthalene) found in SPMDs deployed in single-family houses with (black bars) and without (grey bars) wood burning as residential heating system (modified from Paper V).

Retene (1-methyl-7-propylphenanthrene) has been suggested as an indicator of atmospheric PAH pollution from wood burning (Ramdahl, 1983; Hawthorne et al. 1988). Retene was found in all samples, indicating that it was present in the indoor environments, but the amounts were low and ranged from 0.3-1.4 ng ⋅ SPMD-1 ⋅ d-1. High retene levels were detected in houses which used wood burning as heating system, but the highest concentration of retene was detected in a house which was not using this type of heating. Thus, the correlation between retene data and the use of residential wood burning was unclear. A possible reason for these results could be that retene is more abundant in emissions from burning of softwood (like coniferious wood) as compared to hardwood (McDonald et al., 2000). Unfortunately, the type of wood burned in this study was not documented, and the utility of retene to detect wood burning as the source of gas phase PAHs in indoor air cannot be clarified from the data obtained in the work presented in Paper V.

7. Fields of application and sampling strategies

7.4 Estimation of the consequences of gas phase PAH exposure for plants

Many studies have found that POPs may accumulate from the atmosphere into plants (e.g., Voutsa and Samara, 1998; Ockenden et al., 1998b; Kipopoulou et al., 1999; Müller et al., 2001; Smith et al., 2001; Kylin and Sjödin, 2003; Hellström et al., 2004). The plant uptake from the air occurs mainly through gas deposition or wet/dry particle deposition. Gas phase PAHs are generally adsorbed to the plant surface and/or accumulated via the stomata in the cuticle of the plants, whereas particle-associated PAHs are mainly retained on the plant surface (Smith and Jones, 2000). In the work described in Paper I, the uptake of PAHs into the water-surface living plant water spinach (Ipomoea Aquatica), and the gas phase concentration of PAHs, was sampled in parallel at three different locations in rural (B), semiurban (C) and urban (D) area in the Bangkok region, Thailand. Water spinach is a perennial, semiaquatic plant that occurs, both wild and cultivated, in the southern Asia and other subtropical areas (Figure 25). Its growth rate is fast and the young shoot, about the upper 30 cm of the plant, is a highly-regarded food that is frequently consumed by the inhabitants of the Bangkok region, Thailand. Since water spinach is often cultivated in rivers and canals in polluted areas, the plant is a potential source of PAH-exposure for humans. Thus, this plant species was selected, even though it is both air and water- living, which complicate the data interpretation by adding water as a possible exposure route for the plant.

7. Fields of application and sampling strategies

Figure 25. Illustration of Ipomoea Aquatica provided by IFAS, Centre for aquatic plants, University of Florida, Gainesville, 1990. The plant has a long, thin stem that generally grows floating on the water surface with horizontally oriented leaves, mostly held slightly above the water surface (obtained from http://aquat1.ifas.ufl.edu/ipaqpic.html).

At each of the three locations B, C and D, one SPMD sampler was deployed at the river side for three weeks. At the time of retrieval of the SPMD samplers, 20 – 25 cm of the upper part of each plant (the edible part) was cut off, wrapped in aluminium foil and transported to the Asian Institute of Technology (AIT), Bangkok, Thailand. These plant shoots were three to ten days old, and had mostly been growing above the water surface. In the laboratory, the plant samples were rinsed in tap water followed by distilled water on the day of collection, to remove particles from the plant surface. Shoots from ten plants were pooled per sample, and triplicate samples were obtained at each site. The samples were weighed (wet weight (ww)), wrapped in aluminium foil and stored at -18 °C. The plant and the SPMD samples were analysed for 15 of the 16 EPA PAHs (excluding acenaphthylene) plus benzo[e]pyrene. The PAH amounts found in the SPMDs were converted to gas phase concentrations using the approach discussed in 4.5.1.

The total average PAH concentrations (n = 3) of the plant samples ranged from 4.2 to 9.6 µg ⋅ g-1 ww, with two times higher levels at site D than at sites B and C. In order to compare these results with data obtained in other studies, the dry weight (dw) of the plant was estimated to be 10 percent of the

7. Fields of application and sampling strategies ww. The recalculated average plant concentrations ranged from 42 to 96 µg ⋅ kg-1 dw and were in the same range as concentrations detected for the 16 EPA PAHs in vegetables growing in an industrial area of northern Greece (Voutsa and Samara, 1998; Kipopoulou et al., 1999).

ww) -1 6

Plant concentration (ug kg a) 0 Fl Py Ant Ace Phe Chr BaP BaA IcdP Fluo DahA BghiP

BkF BbF,

12 )

-3 9

m (ng concentration Air

b) 0

Fl Py Ant Ace Phe Chr BaP BaA IcdP Fluo DahA BghiP

BkF BbF, Figure 26. Individual PAH concentrations in a) the plants and b) the gas phase at sampling sites B (●), C (▲) and D (■) (Paper I).

The major individual PAHs found in the gas phase and the plants were fluorene, phenanthrene, fluoranthene and pyrene (Figure 26). However, PAHs with three rings (MW < 202) were predominant in the gas phase, while the PAH-profiles found in the plants were slightly different with levels of 4- ring PAHs (MW < 202) that were higher or similar to the levels of 3-ring PAHs. Thus, gas phase and particle-bound PAHs were accumulated in plants to a similar degree, indicating that particle phase exposure contributed to the levels found in the plants. Another explanation for the differences in the PAH-profiles between the gas phase sampled with SPMDs, and the plants is

7. Fields of application and sampling strategies that the plants could have taken up additional amounts of the PAHs from the water. However, possible PAH exposure from the water to the plants was not investigated in this study. Furthermore, the concentration of pyrene in the plants (higher levels at site B than C) was not correlated with the traffic intensity at the sampling site. Thus, sources other than traffic as well as other exposure routes than the gas phase like particle and water exposure could have contributed to the higher concentration of pyrene (a 4-ring PAH) in the plants compared to the SPMDs. However, the similarities in the SPMD and plant profiles of 3-ring PAHs demonstrate that the gas phase exposure contributed to the amounts of PAHs accumulated in the plants.

8. Conclusions and future research possibilities

8. CONCLUSIONS AND FUTURE RESEARCH POSSIBILITIES

The work described in this thesis tested the use of SPMDs as integrative tools for monitoring gas phase concentrations of the nonpolar aromatic pollutants PCBs, PAHs, alkyl-PAHs and nitro-PAHs. Conclusions drawn from this work are listed below, and further research possibilities are suggested.

The uptake and release in SPMDs of PAHs and PCBs with log KOA values > 7.9 increase at high wind-speeds (Papers II and III). These results demonstrate that the uptake of most nonpolar aromatic compounds is controlled by the boundary layer at the membrane-air interface. However, the wind-speeds/turbulence in the field will be more moderate and less variable between locations, and thus, the wind effect will be generally low. In future research, the effect of temperature, UV light and particle deposition should be examined. The processes involved in the effect of cold temperature are especially important to understand.

The use of a metal umbrella to shelter the SPMDs decreases the uptake and release of PAHs and PCBs at high wind-speeds and/or turbulence, and thus reduces the wind effect (Paper III). In addition, the uptake of particle- associated PAHs increases in unprotected SPMDs, indicating that the metal umbrellas reduce the effect of particle deposition (Paper III). However, the sampler design used in the work underlying this thesis is probably not optimal for reducing the effect of reflected UV light. In future research the design of the metal umbrella should be improved by testing, for instance, different dark materials to coat the inside of the metal umbrella.

The PRC approach has a high ability to assess the effect of high wind (Papers II and III). However, analytical interferences might reduce the precision of the PRCs (Papers II and III), and it is critical that robust analytical quality control is applied. The effects of particles will probably not be assessed by the use of PRCs (Paper III). The use of the PRC-approach to predict the effect of temperature and photolysis has not been tested. The air temperatures can be highly variable and its effect is quite complex and rather unknown. Thus, the use of the PRC-approach to predict the temperature effect is a very important task to study in future research. The use of photosensitive compounds to predict the effect of photolysis is another future research issue.

8. Conclusions and future research possibilities

SPMDs are efficient samplers of the gas phase concentrations of nonpolar aromatic compounds, and are able to determine the spatial distribution at local scales (Paper I and IV). If uncovered SPMDs are used, they can be highly sensitive samplers, and can be used to determine spatial differences under quite still air conditions, such as those in indoor environments (Paper V). In addition, the use of SPMDs, which do not require electricity, make sampling possible at remote and rural areas where the infrastructure is limited, like the countryside in Thailand (Paper I). SPMDs are also suitable for integrative sampling of the atmospheric distribution of nonpolar aromatic compounds on a continental scale (Paper IV).

The sources of pollution can also be determined with SPMDs (Papers I, IV and V). The 1-methylphenanthrene/phenanthrene ratio is a potential indicator of the contribution of traffic to the gas phase concentrations measured with SPMDs (Paper I). Another approach to determine the pollution source of PAHs is to discern the total PAH-patterns of specific sources. However, SPMDs cannot be used to discern the total PAH-pattern of specific sources, probably because of the high number of diffusive PAH sources influencing the imission measurements (Paper IV). In addition, SPMDs can be used to detect wood burning for residential heating as a possible source of indoor PAH exposure (Paper V), and to determine the importance of the gas phase exposure route to the uptake of PAHs in plants (Paper I).

SPMDs have several advantages in comparison with conventionally used HiVols, including their integrative capacity over long times, reduced costs, and no need of specialist operators, maintenance or power supply during sampling. Thus, the use of SPMDs can increase the monitoring frequency, the geographical distribution of the measurements, and the number of sampling sites, used in air monitoring programs of nonpolar aromatic compounds. However, calibration data of SPMDs are limited, and spatial differences are often only semi-quantitatively determined by comparing amounts and profiles in the SPMDs, which have limited their use in air monitoring programs. In future work, it is important to ensure that SPMDs are properly sheltered,

PRCs are used in the sampling protocols, and that calibrated RS data, or the SPMD-air partition data, of specific compounds are further developed to make determination of TWA concentrations possible.

9. Acknowledgements

9. ACKNOWLEDGEMENTS

Så har jag kommit fram till den del av avhandlingen då jag vill tacka alla er som hjälpt mig under min tid i Umeå som doktorand och peppat mig att simma ända in i kaklet!!

I am grateful for financial support of this work by the European Commission INCO program and the Austrian ministry of education, research and culture. Tack också Helge Ax:son Johnsons stiftelse, Knut och Alice Wallenberg-stiftelsen, the Swedish International Development Cooperation Agency (SIDA), samt Stiftelsen JC Kempes minnes-stipendiefond för ekonomiskt stöd av projekt, resor och konferenser. Tack alla härliga människor på miljökemi som skapar en bra och avslappnad stämmning på jobbet. Speciellt nöjd är jag med superfreda´ som innehåller både innebandy (även om det var ett tag sen jag var med) och gofika, tack för det allihopa! Sen finns det ju alltid några som man vill tacka lite extra. Tack min handledare Pekka för ditt lugn, stöd, dina roliga projektideér, och att du alltid tar dig tid att diskutera forskning med mig. Tack också för alla resor du har tagit med mig på, och för att du har presenterat mig för många duktiga och inspirerande forskare världen över. Bilresan genom klippigabergen, Utah, är ett av många fina minnen från våra jobbresor! Tack också Mats för din handledning, att du stöttat mig i tuffa tider och din konstruktiva kritik av mitt arbete och min avhandling. Tack Patrik för att du lärde mig massor under mitt ex-arbete, och Lena, för att du alltid bryr dig om andra och hjälper och stöttar när det behövs. Tack Fia för att du är så go´ och glad, och att du får mig att skratta! Tack Eva för att du drog med mig till Thailand, och för alla pratstunder om forsket och livet, ditt stöd har varit guld värt! Tack Lijana för din stora omtanke, glada humör och för att du är super att resa med. ”Din hjälp” innan min muntliga presentation i Salt Lake City gjorde susen ;- )! Thanks also to James Huckins, Columbia, USA, and Bo Strandberg, Gothenburg, for sharing your superb SPMD knowledge with me. Keep up the good work! Tack också Leif Lindgren, Älvsbyn, för dina fina illustrationer. Tack alla ni som nu har flyttat men förgyllde min studietid i Umeå med roliga fester, middagar och annat skoj, tack för den härliga tiden! Tack alla vänner här i Umeå för att ni finns och gör var dag kul och gjort doktorerandet lättare. Vad vore en dag utan en TV-kväll med sport, en torsdag utan 24, en kväll utan en runda på IKSU eller en helg utan en skogstur med cykeln?! Tack Älvsbybrudarna för att vi hållit ihop i alla år, allt kul vi har haft tillsammans och för att ni alltid ger mig nya perspektiv på livet, vilket behövs som doktorand! Tack tjejerna i mailgruppen, vad vore en måndag utan en brevlåda full med härliga historia och skvaller. Hoppas jag inte har tråkat ut er på slutet med ändlösa klago-mail. Tjejhelgen i höstmörkret gjorde iallafall super, ser fram emot nästa! För att använda ett slitet uttryck vill jag sist men inte minst tacka min underbara familj. Men hur ska jag kunna skriva vad ni har betytt för mig på bara några rader?! Det går inte helt enkelt... Mamma och Pappa, ni har alltid stöttat och uppmuntrat mig något helt enormt. Utan er hade jag inte varit den jag är, eller nått dit jag är idag! Som ni har peppat mig, ända fram till tryckdag! Tack också mormor för att du alltid brydda dig om ”små-flickorna i Ume”, du finns alltid i mina tankar. Och sen Syrris, hur mycket kul och fint har inte vi haft tillsammans?! Tack för allt, du är världens bästa syster! Och så Anton, min älskade sambo, med dig vid min sida är ingenting omöjligt! Du har peppat, hjälpt och stöttat mig något helt fantastiskt. Tack för att du är så fin!

Det går inte bromsa sig ur en uppförsbacke. Illustration by Leif Lindgren, Älvsbyn

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