Temporal Trends of Polycyclic Aromatic Hydrocarbons in the U.K.

Mar 28, 2008 - Fifteen years of air monitoring data for PAHs at six U.K. sites show statistically ... In the United Kingdom, the Toxic Organic Micropo...
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Environ. Sci. Technol. 2008, 42, 3213–3218

Temporal Trends of Polycyclic Aromatic Hydrocarbons in the U.K. Atmosphere: 1991–2005 SANDRA N. MEIJER,* ANDREW J. SWEETMAN, CRISPIN J. HALSALL, AND KEVIN C. JONES Centre for Chemicals Management and Environmental Science Department, Lancaster Environment Centre, Lancaster University, Lancaster LA1 4YQ, U.K.

Received December 3, 2007. Revised manuscript received February 4, 2008. Accepted February 14, 2008.

Internationally there are few long-term air monitoring programs, which are necessary to assess the effectiveness of source abatement measures as required under the UNECE POPs protocol. In the United Kingdom, the Toxic Organic Micropollutants (TOMPS) program, funded by the Department for Environment, Food, and Rural Affairs (Defra), was started in 1991 and includes regular monitoring of a range of compounds including polycyclic aromatic hydrocarbons (PAHs). In this study, the time series (1991–2005) of atmospheric concentrations of 15 PAHs at six U.K. monitoring sites were investigated. Most show a statistically significant decrease in PAH levels over time, broadly consistent with the reported decline in emissions. Higher levels of heavier PAHs were noted in winter than in summer at most sites. At one coastal site, higher levels of lighter PAHs were noted in summer, possibly due to temperature-driven outgassing of these compounds from seawater. Current annual averages of benzo[a]pyrene are below the recently introduced annual air quality standard of 0.25 ng m-3 at all sites, although quarterly averages have exceeded 0.25 ng m-3 in recent years but only at the urban sites in winter. The atmospheric signature of total PAHs closely mirrors the emission signature, which lends strength to the idea that levels of PAHs in air are still mostly influenced by direct/local sources.

Introduction Polycyclic aromatic hydrocarbons (PAHs) are formed through incomplete combustion. They are emitted to the atmosphere, both in gaseous form and bound to particles, with the major U.K. sources currently being road transport combustion and domestic combustion (1). Because of their toxic, persistent, and semivolatile properties, they were included in the UNECE Convention on Long-Range Transboundary Air Pollution (LRTAP) 1998 Protocol on Persistent Organic Pollutants (POPs) which entered into force in 2003 (2). The POPs protocol identifies specific measures to be taken by parties to cut their emissions of air pollutants. The ultimate objective is to eliminate any discharges, emissions, and losses of POPs. The protocol also obliges parties to reduce their emissions of dioxins, furans, PAHs, and hexachlorobenzene (HCB) below their levels in 1990 (or an alternative year between 1985 and 1995, as often there is not sufficient data available for the * Corresponding author phone: +44 1524 593300; fax: +44 1524 593985; e-mail: [email protected]. 10.1021/es702979d CCC: $40.75

Published on Web 03/28/2008

 2008 American Chemical Society

reference year). However, one practical difficulty is that, internationally, there are very few air monitoring programs where long-term data, extending from these key dates, are available to assess the effectiveness of source abatement measures (e.g., IADN (3, 4) and the EMEP POPs network 5, 6). The United Kingdom is one of the few countries in Europe where these data are available. The U.K. Department for Environment, Food, and Rural Affairs (Defra) funds the regular monitoring of a wide range of atmospheric pollutants. The TOMPS (toxic organic micropollutants) program was started in 1991 and includes regular monitoring of polychlorinated dibenzo-p-dioxins, polychlorinated dibenzofurans (PCDD/Fs), PAHs, and polychlorinated biphenyls (PCBs). In addition to the United Kingdom’s decision to sign the UNECE 1998 POPs protocol, other policy drivers for this program are the European Community’s fourth Air Quality Daughter Directive (2005/107/EC), which includes a target value for benzo[a]pyrene of 1 ng m-3 (annual average), and the establishment of a U.K. Air Quality Objective for PAHs which consists of an annual air quality standard for benzo[a]pyrene of 0.25 ng m-3. All these policy imperatives require sound data on ambient concentrations, trends, and distributions of POPs in the environment. With up to 15 years of monitoring data now available from the TOMPS program for six rural and urban sites, we are in a unique position to investigate temporal trends of PAHs in the U.K. atmosphere. The aim of this paper is to investigate these temporal trends in relation to emission time trends, compound physical chemical properties, location, and seasonal effects.

Monitoring Sites In this paper we analyze PAH time trend data from the six TOMPS sites which have the longest time trend records. They consist of three urban sites in London (LON), Manchester (MAN), and Middlesbrough (MB), two rural sites at High Muffles (HM) and Stoke Ferry (SF), and one semirural site at Hazelrigg (HR) (see Figure 1). At the rural and semirural sites, samplers are located away from major roads, whereas at the urban sites, samplers are located in the city center on the roof of a building. Previously, results from the TOMPS program were reported for London and Manchester only (7) for the first five years of the monitoring program (1991–1995). We now have data for these two longest-running sites for a time period of 15 years, and data for other sites for time periods of between 9 and 14 years (see Figure 1). It is rare to get long time-series like this, especially where the sampling and analysis methods have been consistent throughout the study.

Methods Samples were collected every two weeks with high-volume air samplers, using a glass fiber filter to collect the particlebound compounds and using two in-line polyurethane foam plugs to collect the gas-phase compounds. Typical air volumes collected over a two-week period are 500–700 m3. Extracts were pooled before analysis to obtain quarterly data (Jan-March (Q1), April-June (Q2), July-Sept (Q3), and Oct-Dec (Q4)). Sample collection and analysis, including quality control, has been described in detail in previous papers (7–10). The samples were routinely analyzed for the following 15 PAHs: benzo(a)pyrene (BaP), acenaphthylene (ACL), acenaphthene (AC), fluorene (FL), phenanthrene (PHE), anthracene (AN), fluoranthene (FA), pyrene (PY), benz(a)anthracene (BaA), chrysene (CHR), benzo(b)fluoranthene (BbFA), benzoVOL. 42, NO. 9, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 1. Longest-running TOMPS sites discussed in this paper. Urban sites: London (1991–2005), Manchester (1991–2005), and Middlesbrough (1993–2005). Rural sites: High Muffles (1997–2005) and Stoke Ferry (1997–2005). Semirural site: Hazelrigg (1992–2005). (k)fluoranthene (BkFA), indeno(1,2,3-c,d)pyrene (IP), dibenz(a,h)anthracene (DBAhA), benzo(g,h,i)perylene (BghiP). These are the “EPA16” without naphthalene. The latter was excluded because of problems with quality control due to the volatility of the compound. The concentrations of PAHs presented in this paper refer to the sum of vapor and particle phases. The total data set comprises approximately 300 samples. Of the 15 PAHs screened for, 14 were regularly detected. DBAhA was only detected in 5% of all samples, and although included in the sum of 15 PAHs, the individual compound will not be included in the data analyses. Certain compounds, i.e., ACL, FA, BkFA, and IP were mainly detected at the LON and MB sites; at the other sites, insufficient data were available for these compounds, and trends could not be analyzed. Initial Comments on the Measured Time Trends. Time series for the six sites, showing summer (Q2, Q3) and winter (Q1, Q4) quarters separately, are available for all individual PAHs and for the sum of 15 PAHs in the Supporting Information (Figure S1). A first visual inspection of the time series confirms that levels of PAHs are generally decreasing at all sites for all individual compounds, consistent with observations in North America (11, 12), Europe (6), and the Arctic (13, 14) and broadly consistent with the reported decline in U.K. emissions (1, 15). A suspected outlier is noted for Q2 of 2004 at sites HR, MAN, MB, and SF for compounds AC, FL, PHE, and AN (see circled data points in Figure S1). Therefore, this outlier, although shown in the time series, is not included in any of the statistical analysis of the data. Benzo(a)pyrene. BaP is the most toxic of all PAHs (16). The European Commission working group on polycyclic aromatic hydrocarbons recommended that, from a health perspective, ambient air concentrations of BaP should be below 1 ng m-3 annual mean (16). The United Kingdom recently introduced an air quality standard of 0.25 ng m-3 3214

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(annual mean) for this compound. Time trends for benzo[a]pyrene (BaP) are shown in Figure 2. BaP levels have been falling at all sites. In the early 1990s, quarterly averages were typically 1–4 ng m-3 at urban sites, with the 1991 and 1992 annual averages at LON and MAN exceeding the target value of 1 ng m-3 set by the European Community’s fourth Air Quality Daughter Directive (2005/ 107/EC). By comparison, levels at rural sites were an order of magnitude lower, at around 0.3 ng m-3. Hazelrigg as the only semirural site fell somewhere in between, with levels in the 1990s of below 1 ng m-3. In the most recent year for which data are available (i.e., 2005), levels were significantly lower than in the 1990s. At the two rural sites, annual averages have been well below the standard of 0.25 ng m-3 since the start of monitoring, and at Hazelrigg they have been well below the standard since 1995. At the three urban sites, annual averages exceeded the standard of 0.25 ng m-3 most years until 2001, but in recent years quarterly averages have only exceeded 0.25 ng m-3 in winter. This trend with higher BaP levels in winter than in summer is apparent at all sites. Currently the main source of BaP in the United Kingdom is believed to be fuel combustion for domestic space heating (1) which is more important in winter, thus explaining the higher levels in winter months. Statistical Analysis of Time Trends for All PAHs. Statistical analysis of the time trends for all individual PAHs and for the sum of 15 PAHs was carried out to investigate the statistical significance of the decreasing trends and to calculate apparent half-lives. Assuming first-order kinetics (11, 17), the concentration of the chemical in air is given as C ) C0 e-kt or ln C ) ln C0 - kt where C is the concentration of the compound in air (ng m-3), k is the rate constant, and t is time (years). Therefore, regressing the natural logarithm of the concentration against time should yield a straight line with intercept ln C0 and slope -k. Half-lives can then be calculated as t1/2 ) ln 2/k. Calculated half-lives, together with regression statistics (p-value, 95% CI), are available in Table S1 for all individual PAHs and for the sum of 15 PAHs. The first-order decay curves are shown in Figure S2. Only statistically significant halflives (p < 0.05 for the slope of the regression) are reported. Out of the 74 regressions, 63 were statistically significant. Out of the 11 regressions that were not significant, eight were at the rural sites High Muffles and Stoke Ferry. This indicates that at these sites, which are more remote from sources, the PAH signal is more weathered and influenced by meteorology and less dependent on source patterns. The time series for these two sites is shorter, and therefore the likelihood of detecting significant trends is decreased given the variability in the data (17). The compound with the fewest significant decreasing trends is AC, for which only Hazelrigg and London showed a significant trend. Lighter PAHs such as AC are more prone to degradation in the atmosphere (18) and are therefore less likely to directly reflect source patterns and associated time trends. Half-lives range from 2.2 years (AN at Middlesbrough) to 13.4 years (PHE at Manchester), with an average half-life of 5.9 years for all compounds at all sites. These half-lives are comparable to those obtained in other longterm monitoring campaigns. For example, Cortes et al. (11) reported atmospheric half-lives of gas-phase PAHs ranging from 2.2 to 9.3 years for three IADN sites in the Great Lakes region. Another study looking at temporal trends of PAHs at seven U.S. and Canadian IADN sites, with data spanning a longer time period, found half-lives of total PAHs in Chicago were 8.7 in the vapor phase and 8.9 in the particle phase (12). This study found total PAH concentrations at the rural sites declined very slowly or showed no significant decline. This agrees with the findings in the current study, where fewer

FIGURE 2. BaP concentration (ng m-3) at the six sites, showing winter and summer quarters. The dashed line indicates the recently introduced air quality standard. significant trends were found for the rural sites HM and SF. In Europe, long-term monitoring of a range of POPs is carried out by individual countries as part of the EMEP POPs network (5). The longest time series available are from Czech Republic (10 years) and Sweden (12 years). Apart from the Czech data (6) which shows concentrations of total PAH decreasing with time, none of this data is published in the open literature. However, the data is reported every year by NILU (5) and can be freely downloaded from their Web site (19). We applied our statistical analysis to the Swedish data and found that PAHs at the Rorvik site (SE2/14) showed no significant trend, except for AN (half-life 9.5 years). At the other site (SE12), significant decreasing trends giving half-lives ranging between 4.5 and 8.7 years were found for a number of PAHs, again in the same range as found in our study and the other studies mentioned previously. Seasonality. Previous studies have identified seasonality in PAH levels using earlier TOMPS data (7, 20), based on annual averages for only a few years. With the longer time series available now, a more conclusive assessment of summer-winter differences can be carried out. However, because of the declining trends it is not possible to compare summer and winter levels averaged over the whole time series. Therefore, statistical analysis (ANOVA) was carried out on the natural logarithm of the concentration versus time to determine whether there was a significant difference in intercept between the summer and winter regression lines, using the same slope for both summer and winter regression. This analysis was therefore carried out only on data sets where it was previously confirmed that there is no significant difference in the slope of both regression lines when regressed separately. ANOVA results of natural log transformed data are summarized in Table 1. The compounds showing the clearest summer-winter differences are BaP, CHR, BbFA, and BghiP,

with higher levels in winter. These compounds are the heavier, mainly combustion-related PAHs. The Lancaster site differs from the others in that although higher levels are found in winter for BaP, BbFA, and BghiP in line with the general trend, higher summer than winter levels were observed for PHE, AN, and PY. This was noted previously (7, 20) by other authors reporting earlier results from the TOMPS program. Possible causes suggested by Prevedouros et al. (20) were enhanced volatilization from vegetation/soils or greater traffic volume in summer. However, these factors are not exclusive to the Hazelrigg site and they are therefore unlikely to be the cause of the observed patterns. Another possible cause is enhanced outgassing of lighter PAHs from the Irish Sea in summer. Hazelrigg is only 8 km from the coast and is likely to be influenced by air–water exchange over Morecambe Bay under prevailing (i.e., westerly) winds. One might argue that the High Muffles and Middlesbrough sites are also near the sea (10 and 20 km, respectively). However, they are upwind of the sea under prevailing conditions. Other studies have also found higher atmospheric concentrations of the lighter PAHs in summer than in winter (21–23). These were also at coastal sites, suggesting a local influence of seawater outgassing of light PAHs on atmospheric concentrations. Although the increased levels of PHE and AN at Hazelrigg are most noticeable in summer, levels of these compounds are generally more abundant (% of total PAH) at Hazelrigg than at the other sites, suggesting that the influence of seawater outgassing on local PAH levels in air may be a yearround phenomenon. Discussion of the Variability in the Data. Although over 85% of the regressions are statistically significant at the p < 0.05 level (see Table S1), the 95% confidence intervals show that the uncertainty in the calculated half-lives is quite substantial, due to the large variability in the data. The detection of significant trends relies on various factors (17), VOL. 42, NO. 9, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 1. Summary of Summer-Winter Differences, Including the p-Value in Parenthesesa

BaP ACL AC FL PHE AN FA PY BaA CHR BbFA BkFA IP BghiP EPA15

Hazelrigg

High M.

London

Manchester

Middlesbr.

Stoke Ferry

W (0.000) NA ns (0.115) ns (0.203) S (0.000) S (0.000) NA S (0.000) ns (0.830) ns (0.772) W (0.000) NA NA W (0.000) S (0.000)

W (0.000) NA W (0.002) W (0.000) ns (0.762) ns (0.116) NA W (0.007) ns (0.057) W (0.008) W (0.000) NA NA W (0.000) ns (0.055)

W (0.000) W (0.000) ns (0.064) W (0.000) ns (0.458) W (0.001) NA NA W (0.000) W (0.000) W (0.000) W (0.000) W (0.000) W (0.000) ns (0.058)

W (0.000) NA W (0.042) W (0.000) ns (0.386) W (0.000) NA W (0.037) W (0.000) W (0.000) W (0.000) NA NA W (0.000) W (0.014)

W (0.002) ns (0.610) ns (0.602) ns (0.449) ns (0.322) ns (0.491) ns (0.584) ns (0.824) ns (0.205) W (0.028) W (0.010) W (0.003) W (0.002) W (0.001) ns (0.138)

W (0.000) NA ns (0.086) W (0.001) ns (0.121) ns (0.245) NA W (0.000) W (0.000) NA W (0.000) NA NA W (0.000) W (0.000)

a W ) winter > summer (significant at p < 0.05); S ) summer > winter (significant at p < 0.05); ns ) no significant difference at p < 0.05; NA ) not analyzed (either because of insufficient data points or because the two regression lines had significantly different slopes and using the same slope for both regressions was not warranted).

FIGURE 3. U.K. atmospheric emissions of the EPA16 PAHs between 1990 and 2005 (kg). most importantly the variability in the data and the length of the study and also, to a lesser extent, the sampling frequency. Because of the length of the current–time series, significant trends can be detected despite moderately high data variability. Various possible causes can be identified to explain the variability in the data. These will be discussed below. 1. The seasonality of PAH concentrations in air. The TOMPS data presented here showed mostly higher levels in winter but also higher levels of the more volatile PAHs in summer at the HR site. Various other studies have found an increase of PAH levels in winter (6, 12, 14, 24, 25). This increase was usually found for both gas-phase and particle-bound PAH, and it has been attributed to increases in fuel consumption for domestic heating and seasonal variation in the atmospheric boundary layer height. For example, Dimaski et al. (25) found a significant negative correlation between PAH concentrations (total of gas and particle) and temperature. However, correction of the PAH concentrations for the boundary layer effect resulted in a nonsignificant correlation with temperature. The increase of both gas-phase and particle-bound PAHs in winter is therefore not directly related to temperature. On the other hand, it was found that gas-phase PAHs were positively correlated with temperature (11, 12), and Cortes et al. (11) showed that including temperature in the regression equation for gas-phase PAHs reduced variability in the data. The fact that the data presented here are a combination of gas and particle phase concentrations and the fact that they represent a quarterly 3216

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average concentration make correcting for temperature unjustified here. 2. Visual inspection of the data suggests a higher variability in the data in the early years of the time series, i.e., in the early 1990s. Possible explanations for a decreasing variability over time include a change in analytical methods and a decrease in the related analytical error, or a change (i.e., decrease) in variability of emissions over time. As the analytical methods have been consistent throughout the study, analytical error and a possible decrease in it over time can be ruled out. In terms of changes in variability of emissions, the emission estimates show that the relative importance of the different sources of PAHs changes in 1996, with diffuse sources such as traffic and domestic combustion taking over from industrial (i.e., point) sources as major emission sources (see Figure 3). The greater variability in air concentrations before 1996 may have been caused by the fact that emissions from point sources are potentially more variable than those from diffuse sources such as traffic. The data was statistically assessed for changes in variability over time by calculating the residuals around the linear regression line of natural-log transformed concentrations versus time (i.e., the same regression as used to calculate the half-lives in Table S1). This was done by assessing changes in the variance of the residuals with time by splitting up the data set into two parts: an ‘early’ part (before 1996) and a ‘late’ part (after 1996). The variance of the residuals in the early years was compared to the variance in the later years with an f-test to see if they are significantly different (p < 0.05).

FIGURE 4. Relative contribution of individual PAHs to the total (expressed as a percentage of EPA15), both in the emitted mixture and measured in air. Results of the f-test are shown in Table S2. The f-test was carried out only on data sets which had already shown a significant downward trend and for the HR, LON, MAN, and MB sites only, as at the rural sites HM and SF, no data were available before 1996. At the HR, LON, and MAN sites, there was a significantly greater variability in the early data for some of the lighter PAHs but not for the heavier PAHs. This difference between the lighter and the heavier PAHs could be related to the difference in physical-chemical properties, with lighter PAHs being more prone to processes which affect their atmospheric levels, such as atmospheric degradation and air-surface exchange processes. It is interesting to note that at the Middlesbrough site the variance of the residuals is actually higher in the later years. This is due to an increase in levels in 1996 and 1997 at this site, which will be discussed below. However, generally there is a decrease in variability in atmospheric levels of PAHs, which could be attributed to tighter regulation on emissions from large industrial (point) sources. 3. The Middlesbrough site shows temporary increases in levels of a number of PAHs, especially AN, PY, and BaA, in 1996 and 1997. Although these anomalous values do not affect the significance of the first-order fit, they do reduce the confidence in the fit. It is unclear why the levels of these compounds are consistently higher than the general trend over the course of two years. A sudden temporary change in diffuse sources such as domestic combustion and traffic, which are major sources of most PAHs, would not be able to cause such a big change in levels of PAHs because they are made up of a large number of small emitters and it is unlikely that they would all change in the same direction at the same time. The Middlesbrough site is in a very industrial area, and it is nearer to large point sources than any of the other sites, so another explanation might be a temporary change in emissions from a point source nearby. However, the main point sources in the vicinity of the MB site are downwind of the monitoring site. Additionally, these point sources emit pollutants from tall stacks which are designed to disperse the pollutant plume at large height above ground level. Increased emissions locally from tall stacks are usually associated with certain short-term meteorological conditions which “knock down” the plume, and therefore could not cause consistent increased levels over a 2-year period. The most plausible explanation for the anomalous levels in 1996 and 1997 is that they are due to a temporary change in activity related to a fairly local and probably stationary source, such as the undertaking of building work or road works nearby. Changes in Emitted and Measured Signature over Time. U.K. emission estimates are available through the U.K.’s National Atmospheric Emissions Inventory (26) (NAEI) for 1990–2005 for 16 PAHs (the “EPA16”), broken down by source type according to the 1996 IPCC source categories (27). The

emission inventories were compiled through the use of emission factors (i.e., emission per unit of activity) which have been reported in the literature. The emission factor is multiplied by a sector activity, i.e., the annual production or consumption to give the estimated annual emission (1). These emissions are therefore estimates and the absolute amounts are subject to considerable error (1), but they are a valuable tool for identifying overall emission time trends as well as changes in the relative importance of different source categories and changes in composition of the PAH mixture emitted over time. Investigation of emissions suggests that the relative importance of different types of sources has changed greatly during the period 1990–2005 (see Figure 3 and Figure S3). In 1990, industrial processes (in particular anode baking for the primary aluminum industry) were believed to be the main source of PAHs to the U.K. atmosphere, accounting for 46% of all emissions. Emissions from this source have decreased significantly since 1995 because of improved abatement measures following the implementation of the 1990 Environmental Protection Act (1, 15). By 2005, industrial processes made up less than 1% of total emissions. In 1990, agricultural burning was also an important source, but this practice was banned in the United Kingdom in 1993. By 2005, transport was the main source of PAHs, accounting for 65% of all emissions, followed by other fuel combustion activities (22%) not related to power generation, manufacturing/construction, or transport, e.g. mainly fuel consumption for domestic space heating. As the dominant source types have changed over time, the relative amounts of PAHs emitted have presumably also changed during this period. Naphthalene makes up most of the sum of 16 PAHs emitted by all source types. In 1990 it accounted for approximately 50% of total PAH versus 70% in 2005. Note that although naphthalene is the main PAH making up total emissions, this compound is not usually reported in air because of difficulties in getting reproducible data for this compound, and it is not screened for in the TOMPS program. To compare the PAH signature in air with the emitted signature, relative amounts of PAHs emitted have therefore been recalculated using the sum of 15 PAHs, i.e., without naphthalene. In 1990, emissions of the EPA15 were dominated by phenanthrene (34%), followed by fluoranthene (16%), pyrene (9.4%), fluorene (8.6%), and anthracene (8.2%). In 2005, however, emissions were dominated by acenapthylene (29%) followed by phenanthrene (23%), fluorene (11%), acenapthene (7.9%), fluoranthene (7.5%), and pyrene (7.1%). Figure 4 shows the relative amounts of individual compounds (as a percentage of total PAH) for the years 1991 (first year of monitoring) and 2005, both for estimated emissions and for concentrations measured in air. The measured concentrations, expressed as a percentage of EPA15, are averaged per year and include data from the six VOL. 42, NO. 9, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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monitoring sites. A reduced data set was used for this, as a significant portion of the samples did not have a data entry for FA (i.e., compound is “not analyzed” instead of “not detected”), and as this compound normally makes up about 10% of the total, including those samples would significantly skew the percentages. These data points were therefore left out. This was not necessary for other compounds which lacked data entries, as these compounds normally make up less than 1% of the total. The 2005 measured signature in air does not show the increased prominence of acenapthylene in the 2005 emission pattern. The lighter three-ring PAHs such as ACL have much shorter theoretical half-lives in air than the four–six-ring PAHs (18). This might explain the low relative amount in air of this compound. For the remaining PAHs, the atmospheric signature seems to closely mirror emissions, with phenanthrene, fluorene, fluoranthene, and pyrene (in that order) being the most abundant compounds in the atmosphere in 2005. This lends strength to the idea that levels of PAHs in air are still mostly influenced by direct/local sources.

Acknowledgments The authors thank AEA Energy & Environment (acting on behalf of Defra) for providing the speciated PAH emission data. The authors also thank Dr. Robert Lee, Vicky Burnett, Danielle Lock, Susan Hodson, and David Hughes for help with the collection and analysis of the air samples.

Supporting Information Available Two figures showing time trends for summer and winter quarters, and fitted first-order curves, for all PAHs at all sites; one figure showing the breakdown of EPA16 emissions by sector for 1990 and 2005; one table showing calculated halflives for all PAHs at all sites; and one table showing p-values of the f-test carried out to assess changes in data variability over time. This material is available free of charge via the Internet at http://pubs.acs.org.

Literature Cited (1) Wenborn, M. J.; Coleman, P. J.; Passant, N. R.; Lymberidi, E.; Sully, J.; Weir, R. A. Speciated PAH inventory for the UK; AEA Technology Environment: Harwell, Oxfordshire, 1999. Available at http://www.naei.org.uk/reports.php. (2) UN-ECE Convention on Long-Range Transboundary Air Pollution. 1998 Protocol on Persistent Organic Pollutants; http:// www.unece.org/env/lrtap/welcome.html. (3) Simcik, M. F.; Basu, I.; Sweet, C. W.; Hites, R. A. Temperature dependence and temporal trends of polychlorinated biphenyl congeners in the Great Lakes atmosphere. Environ. Sci. Technol. 1999, 33, 1991–1995. (4) Hillery, B. R.; Simcik, M. F.; Basu, I.; Hoff, R. M.; Strachan, W. M. J.; Burniston, D.; Chan, C. H.; Brice, K.; Sweet, C. W.; Hites, R. A. Atmospheric deposition of toxic pollutants to the Great Lakes as measured by the Integrated Atmospheric Deposition Network. Environ. Sci. Technol. 1998, 32, 2216–2221. (5) Aas, W.; Breivik, K. Heavy metals and POP measurements 2005; Norwegian Institute for Air Research, EMEP/CCC-Report 6/2007. Available at http://www.nilu.no/projects/ccc/reports.html. (6) Holoubek, I.; Klanova, J.; Jarkovsky, J.; Kohoutek, J. Trends in background levels of persistent organic pollutants at Kosetice observatory, Czech Republic. Part I. Ambient air and wet deposition 1996–2005. J. Environ. Monit. 2007, 9, 557–563. (7) Coleman, P. J.; Lee, R. G. M.; Alcock, R. E.; Jones, K. C. Observations on PAH, PCB, and PCDD/F trends in U.K. urban air, 1991–1995. Environ. Sci. Technol. 1997, 31, 2120–2124.

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