Traceability of pollutant measurements for ambient air monitoring

Traceability of pollutant measurements for ambient air monitoring

Trends Trends in Analytical Chemistry, Vol. 23, No. 3, 2004 Traceability of pollutant measurements for ambient air monitoring V. Desauziers Air-qual...

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Trends in Analytical Chemistry, Vol. 23, No. 3, 2004

Traceability of pollutant measurements for ambient air monitoring V. Desauziers Air-quality management is implemented by national monitoring networks. This implies quality assurance of analysis to ensure good traceability of the data obtained that is of great importance for data comparison on temporal and spatial scales and at national and international levels. In this aim, several tools have been established through regulation: definition of the priority pollutants to be monitored; determination of limit values; and, documented standards. These different points are presented in this article and the particular case of measurement of volatile organic compounds (VOCs) in ambient air highlighted, through discrepancies observed between reference methods, the further efforts that need to be made to achieve a better understanding of the variability of results and the need for developments to enhance air monitoring. # 2003 Published by Elsevier B.V. Keywords: Air monitoring; Analysis; Pollutants; Sampling; Traceability

V. Desauziers* Laboratoire Ge´nie de l’Environnement Industriel, Ecole des Mines d’Ale`s, He´lioparc, 2 avenue Pierre Angot, F-64053 Pau Cedex, France

*Tel.: +33 (0)5.59.30.54.25; Fax: +33 (0) 5.59.30.63.68; E-mail: Valerie.desauziers@ ema.fr

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1. Introduction In recent decades, air quality has become a very important concern as more and more studies have shown the great impact of atmospheric pollution on environment and health. With di¡erent processes, such as precipitation and in¢ltration, pollutants can deposit onto soils or into natural waters, reach groundwater and, hence, damage ecosystems. We can highlight the main phenomena of atmospheric pollution as acidi¢cation, photo-oxidation, the greenhouse e¡ect and the impoverishment of stratospheric ozone. Acidi¢cation, through acid rains (HNO3 and H2SO4) are formed from nitrogen oxides and SO2, has been shown to have an impact on forest and natural water ecosystems and on building erosion. Nitrogen oxides (NOx) are of particular concern, as are volatile organic compounds (VOCs), precursors of

photochemical pollutants, such as ozone, aldehydes and peroxyl-acetyl-nitrate (PAN). Compared with tropospheric ozone, which directly impacts human health, stratospheric ozone is necessary as UV ¢lter, and emissions of its main destructors, chloro-£uoro-carbons (CFCs), should be reduced. Depending on their concentrations, some of these pollutants may cause toxic e¡ects to human beings, from nausea and breathing di⁄culties to cancer [1]. This is particularly demonstrated in urban areas where most of the population of industrialized countries is concentrated (e.g., in Europe, 278 million people live in cities with more than 50,000 inhabitants, while 378 million live in rural areas [2]). To improve air quality, regulation has been implemented on an international scale. It generally implies emission inventories, air monitoring and atmospheric dispersion modeling. To estimate emissions and elaborate management strategies, several methodologies have been developed. In Europe, the CORINAIR (CORe INventory of AIR emissions in Europe) helps quantify annual national releases into the atmosphere [3]. At the international level, because of the di⁄culties of data comparison, an attempt to harmonize the methodologies has been going on between various international bodies (e.g., European Commission (EC), the Intergovernmental Panel on Climate Change (IPCC), EUROSTAT, and the International Atomic Energy Agency) [3]. The traditional approach to evaluating air quality involves concentration

0165-9936/03/$ - see front matter # 2003 Published by Elsevier B.V. doi:10.1016/S0165-9936(04)00310-3

Trends in Analytical Chemistry, Vol. 23, No. 3, 2004 measurements in air and comparison with limit values. This is performed through monitoring sites (4300 in North America) which belong to national networks (e.g., State and Local Air Monitoring Stations (SLAMS) in the USA, and ATMO in France). Monitoring air quality through measurement networks implies quality assurance (QA) of analysis to ensure good traceability of the data obtained. Traceability is particularly important in this case because of the necessity for data comparison on temporal and spatial scales, at national and international levels. In this aim, several tools have been established through regulation:  de¢nition of the priority pollutants to be monitored, in agreement with the ability of the techniques to measure them;  determination of limit values that will condition the analytical steps and the method performance (detection limits, repeatability. . .);  establishment of documented standards that describe analytical procedures with minimum quality requirements; and, calibration procedures. These di¡erent points will be discussed in the following sections. Recently, in addition to the classical approach of data comparison with limit values described above, modeling has been included. New regulations, such as the EC’s Framework Directive (96/62/CE) allows the use of atmospheric dispersion modeling systems. They involve spatial and temporal scales, are applied to assess the future air quality against the air-quality objective [4] and can be consistent tools for air-quality management in case of poor spatial covering of monitoring data [5]. However, these models should be carefully validated and scienti¢c uncertainties should be assessed so as to obtain good agreement with the airmanagement policy [4].

Trends (96/62/CE). The monitoring domain was enlarged to 13 pollution indicators. They include pollutants already controlled by the Directives cited above and new compounds: benzene; polyaromatic hydrocarbons (PAHs); carbon monoxide; cadmium; arsenic; and, mercury [2]. One of the objectives of this new Directive is the uniform evaluation of air quality, so it involves improving traceability. To achieve this aim, for each pollutant a set of new Directives (‘‘Daughter Directives’’) are being prepared through working groups that have responsibility for collect information about pollutant levels, health e¡ects, and measurement methods and location. In North America, the same pollutants are measured through 4300 monitoring sites [6]. Among them, particulate matter generally corresponds to PM10, particles 10 m in diameter or smaller. In 1997, the US Environmental Protection Agency (EPA) announced a new PM2.5 standard as being necessary for health and environmental protection [6]. This new standard has initiated a debate in the scienti¢c community regarding the appropriate measurement to be considered and new rules are in progress [6]. Through the 1990 Clean Air Act Amendments and in order to improve monitoring of ozone and its precursors (mainly VOCs), the US EPA initiated the Photochemical Assessment Monitoring Stations (PAMS) program in 1993, and routine VOC measurements have been applied [6]. As a result, individual measurements and therefore separative techniques are required, but long term on-site monitoring imposes a lot of technical constraints in this case. The analytical challenge is particularly strong. Other factors to be considered in selecting or developing analytical procedures are the concentration levels to be quanti¢ed and the frequency of measurements. These are described in the following section for each priority pollutant.

3. Limit values and concentration units 2. De¢nition of pollutants to be monitored Priority pollutants were selected according to criteria that included both their known deleterious e¡ects on health and the feasibility of their measurement, which is of the utmost importance for traceability. For example, in Europe, air quality has been regulated since the early 1980 through EC Directives concerning SO2 and Suspended Particle Matter (80/779/ EEC), lead (82/884/EEC), NO2 (85/203/EEC) [2]. In 1992, ozone was added (92/72/EEC). After 15 years of air quality measurements, it was demonstrated that these Directives were not as e¡ective as expected. As a result, the EC published the new Framework Directive on Ambient Air Quality Assessment and Management

The new proposal in the EC Directive includes requirements in terms of limit values and margin of tolerance. Both in Europe and North America, measurement frequencies and limit values have been de¢ned with respect to the impact of the pollutant on human health (1h, 8h, 24h and annual averaging periods) or on vegetation (annual averaging period) (see Table 1). The limit values de¢ned by the Framework Directive should not be exceeded from 2005 for SO2, CO and PM10, and from 2010 for NO2 and benzene. Until these dates, margins of tolerance are allowed: 43% for SO2, 50% for NO2, PM10 and CO, and 100% for benzene. These margins will progressively reduce every 12 months by equal annual percentages to reach 0% by January 2005 or 2010 according to the pollutant. http://www.elsevier.com/locate/trac

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Table 1. Limit value and averaging time for the priority pollutants in Europe and USA (adapted from Demerjian [6] and Santos-Alvez and Patier [2]) Pollutant

Directive (96/62/CE) Averaging time

US NAAQS*

Limit value 3

CO

8h

10 mg/m

NO2

Annual (heath) Annual (vegetation) 1h 8 h (heath) 1 h (vegetation) 24 h (vegetation) Annual Annual 24 h 1h Annual 24 h

40 g/m3 30 g/m3 (NO+NO2) 200 g/m3 110 g/m3 200 g/m3 65 g/m3 350 g/m3 30 ppb 125 g/m3 20 g/m3 40 g/m3 50 g/m3

Annual Annual

5 g/m3 0.5 g/m3

O3 SO2

PM10 PM2.5 Benzene Lead (on PM10)

Averaging time

Limit value

8h 1h

9 ppm 35 ppm

Annual

53 ppb

8h 1h

80 ppb 120 ppb

24 h

150 ppb

Annual 24 h Annual 24 h

50 mg/m3 150 mg/m3 15 mg/m3 65 mg/m3

3 months

1.5 mg/m3

*NAAQS: National Ambient Air Quality Standards

3.1. Comment on concentration units Table 1 shows that, whether considering EC Directives or US NAAQS, the concentration units of pollutants in air are di¡erent (g/m3 or ppb (or ppm)). By contrast, for water pollutant the concentration units are the same (g/L and ppb, as water density is 1). For air, ppb or ppm (also noted ppbv and ppmv) corresponds to the ratio ‘‘pollutant volume/air volume’’, which is therefore temperature- and pressure-dependent and is not equivalent to g/m3 or mg/m3. For example, 40 g/m3 NO2 corresponds to 44.6 ppbv at 1 atm and 273 K. In this case, the result can be also expressed as 40 g/Nm3 where ‘‘N’’ is for normal conditions of pressure and temperature. This di¡erence in concentration units can cause mistakes and di⁄culties for data inter-comparisons and hence can a¡ect traceability.

4. Reference methods (RMs) For most of the criteria pollutants listed in Table 1, the limit values to be reached are low (ppb or g/m3 levels). As a result, analytical instruments need to be sensitive enough and to be able to provide quantitative results within the recommended averaging times; that generally requires routine measurements on site. To ensure traceability of the measurements performed through monitoring networks, standard operating procedures (SOPs), equivalent measurement methods and QA/QC (quality control) procedures should be implemented. The typical instrumentation and/or measurement methods described in inter-

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national, European and US standards are summarized in Table 2. These RMs have been elaborated and improved in the past 10 years. Development of RMs for suspended particles (0.010^20 m in diameter), heavy metals (Pb, Cd, As, Ni), PAHs and ozone precursors are in progress in Europe and measurement of PM2.5 is under development in USA. The methods cited in Table 2 operate under established QA requirements. For SO2, NO2 and O3, these requirements consist of routine zero, span and precision checks, and periodic audits. Considering individual identi¢cation by gas chromatography (GC) analysis (for benzene and VOCs determination), routine calibration with single or multi-component standards is required. Inter-comparison exercises are also organized to validate RMs (e.g., the HAMAQ (Harmonization of Air Quality Measurements for Important Atmospheric Pollutants in Europe) program that was held from 1996 to 1999, has allowed comparison of the RMs of seven European laboratories for ¢ve pollutants (SO2, NO, NO2, benzene, and CO) [7]. However, despite these procedures being implemented to reach the quality objectives de¢ned in regulations, some monitoring issues remain, as illustrated in the following examples. 4.1. Performance of commercial instrumentation Some standard commercial analyzers are not sensitive or speci¢c enough when they operate in routine monitoring. This is the case for the measurement of nitrogen oxides by chemiluminescence [6].

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Table 2. Reference methods for ambient air monitoring Pollutant

Principle of analysis

DL1/noise [6]

Standard or project

CO

Non-dispersive IR spectrometry

1 ppm/0.5 ppm

NO/NO2

Chemiluminescence

10 ppb/5 ppb

O3

Chemiluminescence UV spectrophotometry

10 ppb/5 ppb

SO2

UV fluorescence

10 ppb/5 ppb

TSP2 PM10 PM2.5 Benzene

Gravimetry Gravimetry Under development in USA Dynamic sampling/thermal desorption/GC Dynamic sampling/solvent desorption/GC Diffusive sampling/thermal desorption/GC Diffusive sampling/solvent desorption/GC Dynamic sampling/GC in situ Dynamic sampling/thermal desorption/GC Diffusive sampling/thermal desorption/GC GC/flame ionisation X-ray fluorescence Atomic absorption

1 mg/m3 1 mg/m3 -

ISO 4224:2000 PR NF EN 14626 (2003) US EPA [8] PR NF EN 14211 (2003) US EPA [8] ISO 10313 :1993 ; US EPA [8] ISO13964:1998; US EPA [8]; PR NF EN 14625 (2003) PR NF EN 14212 (2003) US EPA [8] NF X43-023 (1991); US EPA [8] EPA [8] PR NF EN 14662-1 (2003) PR NF EN 14662-2 (2003) PR NF EN 14662-4 (2003) PR NF EN 14662-5 (2003) PR NF EN 14662-3 (2003) NF EN ISO 16017-1 (2001) NF EN ISO 16017-2 (2003) EPA [8] EPA [8] ISO 9855:1993

VOCs Lead (on PM10) 1

0.01 g/m3

DL: Detection limit; 2TSP: Total Suspended Particulate

The principle is based on a reaction between NO and ozone to form excited NO2, which relaxes to its fundamental energetic level by emitting a speci¢c radiation in the near infrared (NIR) (1200 nm). In order to measure NO2, NO2 is ¢rst reduced in NO with a catalytic converter. Both NO and NO2 (NOx) are then measured and, by di¡erence with NO, the NO2 concentration can be obtained. Two shortcomings a¡ect the use of these analyzers [6]: the converter is not speci¢c (nitric acid and organic nitrates can react); and, overestimation of the NO2 concentration can occur. Moreover, the detection limit for NO is not low enough to evaluate correctly the NO2/NO ratio during periods of high ozone production. Alternatives have been developed, but they should be validated in routine monitoring [6]. 4.2. Representative sampling Because of the complexity of the sample, no completely satisfactory way of measurement can be found, even if RMs exist. This is the case for the measurement of suspended particles. RMs with 10% precision are well described [9,10] for the measurement of particulate matter remaining on a ¢lter after sampling. However, these RMs are subject to positive and negative artifacts because of the complexity of the mixture of constituents in aerosols (e.g., losses of semi-volatile organic com-

pounds and ammonium nitrate may occur, while, by contrast, particle-bound water may be retained and vary according to the seasons and sampling location [9]). As no standard suspended aerosol exists, intercomparison of the di¡erent RMs on the same atmospheric sample would be the only way to evaluate the comparability of the di¡erent devices. A critical review on the measurement of suspended particles can be consulted for more details [10]. 4.3. Need for technology development Quality objectives ¢xed by regulation are sometimes too ambitious with respect to the state of knowledge and technical development (e.g., Demerjian [6] noticed that for PM2.5 in USA, for which quality objectives were stated in US EPA 40 CFR Parts 53 and 58, 1997, considerable development work remains to be done before routine monitoring can be achieved). 4.4. Summary These di¡erent examples show that, even if RMs and QA/QC procedures are well de¢ned, ensuring traceability remains a continuous challenging task involving the improvement of the existing instrumentation for a better routine operation, the continuous development of reliable technologies to obtain representative measurements and to respond to new requirements in quality objectives. http://www.elsevier.com/locate/trac

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5. Calibration Calibration is a key step in the analytical process, and is therefore an important element in determining the accuracy, the precision and the robustness of analyzers. For air analysis, calibration is generally performed by using gas conditioned in pressure cylinders at various concentrations covering the range of interest. For routine use, working standards should be traceable to an NBS (National Bureau of Standards) SRM (Standard Reference Material). Certi¢ed gases are available from suppliers with NBS traceability documentation. An international standard (ISO 6141:2000) describes the requirements for certi¢cates for calibration gases and gas mixtures. For zero checks, zero-grade nitrogen can be used. Generally, a standard gas contains only one compound at a ¢xed concentration. For VOC measurements, there is usually only one calibration curve for all compounds (e.g., when a £ame ionization detector is used and linear and cyclic alcanes are being calibrated). Indeed, in this case, the detector response is proportional to the e¡ective carbon number in the molecule [11]. It is thus possible to calculate the response coe⁄cient for each alcane by using a reference propane standard, as is recommended for total VOC determination in industrial emissions [12]. For other VOC measurements, standard gas mixtures are proposed by the suppliers. Since 1990, these standard mixtures are certi¢ed by inter-comparison exercises between di¡erent international organizations [13]. However, some stability issues are sometimes observed with regard to the compounds and their concentrations in the mixture [14]. Moreover, pressure cylinder mixtures are not relevant for checking the method performance, as it is not possible to vary the relative ratios between compound concentrations. Di¡erent alternative normalized methods for generating standard gases are available. They are based on the static or dynamic generation modes described below.

methods, they are particularly convenient when high sample volumes are required and when the in£uence of air velocity should be checked on a sampling device. International standards (ISO 6145/1 to ISO 6145/8) describe the numerous techniques that can be involved for the preparation of gas mixtures. Among them, permeation is one of the most used. 5.2.1. Permeation. The principle involves the permeation of the target compound through a convenient membrane into a complementary gas £ow. The target compound is placed in a tube which is maintained at constant temperature in a thermostatted chamber. The methodology is described in international standards ISO/DIS 6145/10:1999 and ISO 6349:1979 (e.g., it is recommended for SO2, NO2, and benzene for concentrations ranging from 10 6 to 10 9 in molar fraction with a relative uncertainty of 2.5% (ISO/DIS 6145/ 10:1999)). The permeation rate through the membrane depends on: the compound itself; the membrane structure; the temperature; and, the partial pressure of the compound in and out the tube. These parameters can be maintained constant when the system is used correctly. A source of errors can be the experimental determination of the permeation rates, which requires precise weighing. To overcome this drawback, it is possible to buy permeation tubes certi¢ed by suppliers for permeation rate, uncertainty and lifetime. This method is therefore one of the most convenient for generating gas mixtures because of its robustness and precision. However, the system requires high equilibration time, may not be cost-e¡ective for complex mixtures (as 1 tube is required per compound) and moreover, it is not possible to modify the relative concentrations of compounds, which depend on their permeation rates. An example of home-made system using the permeation method for generation of trace sulfur compounds is shown in Fig. 1 [17]. Successive dilutions of the

5.1. Gravimetric methods These methods are convenient when small volumes of standard gas are required. They are based on the introduction of a known mass of the target compounds (liquid or gas) in an known volume of dilution gas (air for example). Through successive dilutions, it is possible to reach ppmv [15] or ppbv [16]. This procedure is identical to that applied to pressure cylinders, and the international standard ISO/DIS 6142: 1999 describes the criteria for ensuring the validity of the gas mixture. 5.2. Dynamic volumetric methods These methods consist of introducing a £ow-rate of gas A (pure compound or mixture) in a constant £ow-rate of complementary gas B. Compared with gravimetric

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Figure 1. Gas-generation device for trace sulfur compounds based on permeation (adapted from [17]).

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standard gas generated were carried out by using mass£ow meters calibrated for air by the supplier [17,18]. These £ow meters were preferred to rotameters, as they are more precise: 1% of the maximal £ow rate. It should be noted that this precision can be obtained only if a pressure drop of 1 bar applies between the inlet and the outlet of the mass-£ow meter [19]. To improve homogeneity, a mixing chamber was introduced. It is also possible to check the in£uence of air humidity on the analytical method by introducing an humidi¢cation system [17]. The presented set-up allowed each of the ¢ve sulfur compounds studied to be generated in the concentration range 0.01^1000 g/m3 [17].

Three major sources of errors can be identi¢ed with the syringe-injection technique:

5.2.2. Syringe injection. This method is particularly recommended when calibration for numerous compounds in a mixture is required. By contrast with pressure cylinders and permeation tubes, it is possible to modify easily the composition of the standard mixture qualitatively and quantitatively [19]. However, this method is suitable only for the generation of atmospheres of gases or low-to-medium boiling liquids (up to 140 C) [20]. The principle involves loading organic gases or liquids into a gas-tight syringe connected to an atomization chamber. Injection in a complementary gas (air) £ow stream is carried out by means of a motor-drive on the syringe. This method is recommended for concentrations ranging from ppmv to % with dilution gas £ow rate in the range 1^100 L/min (ISO 6145/4: 1986). If desired, the atmosphere generated can be further diluted by using a set-up similar to that presented in Fig. 1. A home-made device based on syringe injection is shown in Fig. 2. This system allows the generation of gas mixtures containing more than 10 VOCs [21]. The concentration range obtained at the ¢rst dilution step is 1^  100 mg/m3 with a precision of 1% [19]. By means of the second dilution step, it is possible to generate sub-g/m3 concentrations with a precision of 10% [19], which is particularly useful for checking the performance of new pre-concentration techniques for VOCs, such as SPME (solid-phase microextraction) [21^ 23].

5.3. Summary To summarize, a lot of techniques are available for generating standard gases. In most cases, certi¢ed pressure cylinders will be preferred, but some speci¢c applications need more £exibility (e.g., complex mixtures and low concentrations), so the analyst should make his own selection to determine the best ¢t between his analytical objectives and the performance of the system chosen. As noted above, numerous methods and principles are described in international standards, but few have been commercialized, so home-made devices are needed. To ensure good accuracy, precision and traceability from one laboratory to another, validation should be carried out through comparison with a reference system by following the procedure described in ISO/DIS 6143:1998.

Figure 2. Standard gas-generation device for trace VOCs based on the

 atomization may not be quantitative for relatively high boiling compounds;  the motor-drive is expected to have a precision of 1%; and,  the £ow meters are expected to have a precision of at least 2%. Moreover, the stability of the delivered concentrations should be carefully examined because of slight irregularities in the syringe drive [20].

6. VOCs as a case study VOCs include non-methane hydrocarbons (NMHCs) and oxygenated organic compounds that play an important role in the photochemistry of the troposphere. Some of these compounds, such as benzene, are carcinogenic, leading to a limit value ¢xed to protect human health of 5 g/m3(see Section 3 above). Di¡erent analytical methodologies can be applied for VOC monitoring in ambient air, and standards describing RMs are listed in Table 2. These methodologies are based on gas chromatography (GC) equipped with FID (£ame ionization detector), PID (photo ionization detector) or mass spectrometer detector (MSD). 6.1. RMs On-site air sampling can be performed on-line with GC for automated ¢eld measurements or o¡-line with GC analysis at laboratory. For on-line analysis, sampling generally involves preconcentration of VOCs by adsorption on solid materials. Di¡erent devices are used, depending on the di¡erent analytical objectives: http://www.elsevier.com/locate/trac

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 adsorbent tubes for dynamic sampling (air is pumped through the adsorbent bed) are recommended for several minutes to several hours average sampling times;  whereas di¡usive sampling cartridges, based on the uptake of VOCs following Fick’s First Law, are suitable for long-term exposure evaluation (several days to several weeks). For both methods, many di¡erent adsorbents with di¡erent adsorption e⁄ciencies are available (e.g., activated charcoal presents the highest adsorption capacity but desorption for analysis is carried out only by solvent extraction, generally by using harmful carbon disul¢de [24,25]). The solvent is then injected in a standard GC con¢guration. However, several mL are needed to complete quantitative extraction but only 1 L is injected, so there is a loss of sensitivity. Moreover, this desorption method is not easy to automate, so it requires a manual procedure in the laboratory. More recently, thermal desorption has been developed to provide better sensitivity and more reliability for routine operation in situ. The main disadvantage of this technique against solvent extraction is the instrument cost, as a speci¢c automated thermal desorber (ATD) is needed. For this technique, weaker adsorbents than activated charcoal are required [25]. Most used for VOCs are organic polymers (e.g., Tenax TA) and carbon-based materials, such as graphitized carbons (e.g., Carbopack B) or micro-porous solids (e.g., Carboxen and Carbotrap). Detailed reviews and method comparisons are presented by Harper [26], and Skov et al. [25]. Air sampling without pre-concentration is also applied by using canisters that are particularly adapted to the most volatile and non-reactive VOCs [25]. For trace analysis, the inner surface of canister should be inert and covered by deactivated fused silica named Silcosteel [27]. Under these conditions, it is possible to store various VOCs at ppb level up to 30 days after sampling [28]. However, this material is expensive and its cleaning is not easy (a speci¢c device being required), leading to possible cross-contamination from one sample to another. By using this sampling mode, pre-concentration and analysis are carried out in the laboratory [25] and only grab sampling can be carried out. Fig. 3 resumes the di¡erent standardized sampling methods for VOC monitoring and shows the complexity to make the correct choice from among the di¡erent RMs proposed. 6.2. Comparison of methods The choice of method is of course also related to the information required. For example, because health e¡ects from benzene are based on cumulative exposure,

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Figure 3. Different principles of reference methods used for monitoring VOCs. (Bold lettering shows operations possible on site).

long sampling times seem to give the most pertinent information. For other VOCs, short sampling times would be preferred to identify pollution peaks and correlate them to meteorological parameters and ozone concentration. Whatever the sampling technique used, results should be traceable to be compared, so comparative studies have been carried out. For benzene, Skov et al. [25] have shown that a di¡usive sampler measures 20% higher values than a BTEX (benzene, toluene, ethylbenzene, xylene) monitor based on continuous measurement by automated GC. This systematic deviation may be because of the uptake rates used for di¡usive sampler [25]. Another source of error can be related to the di¡erence in the concentration evaluated: a calculated weekly average for BTEX monitor; and, a measured average for di¡usive sampling. By contrast, comparison of a BTEX monitor and dynamic sampling on adsorbent tube for 1 hour showed no signi¢cant di¡erence [25]. For VOC analysis, Czaplicka et al. [29] have compared sampling on charcoal tube followed by CS2 extraction and sampling on Carbotrap followed by automated thermodesorption. Standard deviations for the ¢rst technique are more than 10 times above those of thermal desorption for BTEX, and detection limits are also high (around 1.2 g/m3 against 0.4 g/m3 for thermal desorption) [29]. Moreover, ¢nal results obtained with solvent extraction can be easily a¡ected by errors produced by the operator and also by CS2, which can be very easily contaminated by VOC vapors from air. The extract should be therefore stored under inert gas [29]. By contrast, automated thermodesorption avoids contamination, losses because of solvent evaporation and errors because of manual operation. For these reasons and also because on-site methodologies are now preferred for ambient air monitoring, methods involving solvent extraction tend to have been progressively abandoned.

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Generally, further e¡orts should be made to explain the discrepancies found between methods. With this aim, inter-comparison programs, such as NOMHICE (non-methane hydrocarbon inter-comparison experiment) [30] and AMOHA (accurate measurements of hydrocarbons in air) [25,31] were implemented to assess the quality of measurements.

By contrast, for low concentrations ( < 0.5 ppb), monitoring uncertainties are predominant [33]. In this particular case, low concentrations probably represent the background pollution and should not be signi¢cantly in£uenced by sampling location, meteorological parameters or anthropogenic sources.

6.3. Sampling and analytical variability For any given methodology, di¡erent sources of result variability can be determined. These causes of uncertainties are taken into account in QA/QC procedures. These uncertainties may arise from [31]:

7. Conclusion

 Unreliable standards: for NMHCs, daily calibration is achieved through the e¡ective carbon number concept described in Section 5, and zero gas measurements are also carried out regularly for blank determination.  Loss, contamination and reactions: to avoid loss or reactions in the analytical system, inactivated Silcosteel valves and tubings are used. However, another source of sample modi¢cation has been attributed to the Na¢on membrane that is used before the adsorbent trap to remove water vapor. Even if this technique is more selective than other common dryers, such as magnesium chlorate [31], some polar VOCs (e.g., aldehydes and ketones) are signi¢cantly retained and artifacts because of alkene rearrangement during thermal regeneration of Na¢on are observed [32]. Therefore, this dryer can be used only when nonpolar VOCs are investigated, such as NMHCs, and a speci¢c test should be implemented in the QA/QC procedures.  Misidenti¢cation and erroneous quanti¢cation: these occur because of chromatographic peak overlap and bad baseline. Peak overlap is often because of the complexity of ambient air matrices. To simplify the sample, Na¢on dryer is used to remove polar compounds that may be coeluted with non-polar VOCs. In some cases, manual peak integration should replace automatic mode. Bortnick et al. have studied the relative contribution of sampling, analysis and environmental parameters to monitoring variability [33]. The relative importance of sampling and analytical errors varies by compounds: for carbonyls, sampling appears a more important source of error, whereas for several toxic VOCs (benzene, carbon tetrachloride, methylene chloride), analytical error is predominant. Generally, they also stated that errors caused by environmental variability (spatial and temporal) are more important than monitoring errors [33].

Air-monitoring networks have been implemented since the early 1980s. Since then, considerable e¡orts have been made to improve assessment of air quality through new regulations and development of SOPs and QA programs, including inter-comparison exercises. For most of the target pollutants, RMs are well de¢ned and commercial monitoring instrumentation performs satisfactorily. With su⁄cient QA/QC, these RMs are expected to satisfy the primary objectives of the monitoring network. However, signi¢cant measurement issues remain to achieve the objective of quality data and to enhance the performance of the monitoring network. For example, new criteria for pollutants, such as VOCs or PM2.5, require considerable technological and analytical developments because of the lack of sensitivity, signi¢cant variability and the non-representative nature of the sample, especially at trace concentrations. For high concentration levels, as stated by Bortnick et al. [33], it would be more interesting to increase the number of sampling locations and frequency rather than use sophisticated analytical procedures to arrive at a better understanding of spatial and temporal variabilities of data. In this respect, atmospheric modeling is also envisaged as a reliable tool for overcoming the poor spatial coverage of monitoring data and contributing to their traceability. References [1] US Environmental Protection Agency, Regional Approaches to Improving Air Quality, EPA/451-K-97-001, US EPA, Washington, DC, USA, 1997. [2] J.S.G.D. Santos-Alves, R.F. Patier, Sens. Actuators B 59 (1999) 69. [3] CITEPA, Methodology for calculating emissions into the air, 2001 (www.citepa.org). [4] R.N. Colvile, N.K. Wood¢eld, D.J. Carruthers, B.E.A. Fisher, A. Rickards, S. Neville, A. Hugues, Environ. Sci. Policy 5 (2002) 207. [5] C. Borrego, A.I. Miranda, M. Coutinho, J. Ferreira, A.C. Carvahlo, Environ. Poll. 120 (2002) 115. [6] K.L. Demerjian, Atmos. Environ. 34 (2000) 1861. [7] http://www.gazettelabo.tm.fr/2002archives/publics/2000/ 52LNE.html [8] US Environmental Protection Agency, List of Designated Reference and Equivalent Methods, October 1, 1996. NERL, Air

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