Andrew P.
Rutter
a,
James J.
Schauer
*a,
Glynis C.
Lough
a,
David C.
Snyder
a,
Catherine J.
Kolb
a,
Sara
Von Klooster
a,
Todd
Rudolf
a,
Helen
Manolopoulos
a and
Mark L.
Olson
b
aEnvironmental Chemistry and Technology Program, 660 N. Park Street, University of Wisconsin-Madison, Madison, WI 53706, USA. E-mail: aprutter@wisc.edu, jjschauer@wisc.edu; Fax: +1 (608) 262 0454; Tel: +1 (608) 262 4495
bUnited States Geological Survey, 8505 Research Way Middleton, WI 53562, USA
First published on 31st October 2007
Gaseous elemental mercury (GEM), particulate mercury (PHg) and reactive gaseous mercury (RGM) were measured every other hour at a rural location in south central Wisconsin (Devil’s Lake State Park, WI, USA) between April 2003 and March 2004, and at a predominantly downwind urban site in southeastern Wisconsin (Milwaukee, WI, USA) between June 2004 and May 2005. Annual averages of GEM, PHg, and RGM at the urban site were statistically higher than those measured at the rural site. Pollution roses of GEM and reactive mercury (RM; sum of PHg and RGM) at the rural and urban sites revealed the influences of point source emissions in surrounding counties that were consistent with the US EPA 1999 National Emission Inventory and the 2003–2005 US EPA Toxics Release Inventory. Source-receptor relationships at both sites were studied by quantifying the impacts of point sources on mercury concentrations. Time series of GEM, PHg, and RGM concentrations were sorted into two categories; time periods dominated by impacts from point sources, and time periods dominated by mercury from non-point sources. The analysis revealed average point source contributions to GEM, PHg, and RGM concentration measurements to be significant over the year long studies. At the rural site, contributions to annual average concentrations were: GEM (2%; 0.04 ng m–3); and, RM (48%; 5.7 pg m–3). At the urban site, contributions to annual average concentrations were: GEM (33%; 0.81 ng m–3); and, RM (64%; 13.8 pg m–3).
Mercury exists as three different species in the atmosphere, each with very different chemical and physical properties, and therefore different source–receptor relationships. Gaseous elemental mercury (GEM) is resistant to chemical oxidation, which means that it has an atmospheric residence time between 0.5–2 years,7–9 allowing it to become intrahemispherically well mixed. Oxidized mercury (II) compounds are operationally defined as particulate mercury (PHg) and reactive gaseous mercury (RGM), which can be considered sub-classes of reactive mercury (RM; sum of RGM and PHg).10,11RM compounds are more water soluble and less volatile than GEM, resulting in more rapid removal from the atmosphere.7,8 The difference in time scales of removal between GEM and RM means that RM will be deposited within a few tens to hundreds of kilometres from the source, whereas GEM will be mixed into the global pool before it is oxidized to RM and subsequently deposited,7,8,12 a process which is very slow and spatially diffuse. The contrast between the GEM and RM source–receptor relationships means that reductions in RM emissions have the potential to yield large reductions in regions of mercury deposition at short and medium distances from point source emissions, whereas reductions in GEM emissions from point sources would yield much smaller decreases at short and medium distances from the emission sources.
The ability to quantify the impact of point sources on local mercury deposition is essential in formulating mercury emission control policies. Several studies have been conducted to assess the origins and magnitudes of mercury point source impacts at US receptors. The studies fall into two categories: (i) modeling studies;9,12–14 and (ii) field measurements,15–23 often combined with source apportionment and back trajectory models.24–35 Modeling studies typically provide estimates of average source impacts over large areas, but impacts of point sources on individual receptors short distances downwind are not as well represented, due to the size of the model grid scales and uncertainties in the mercury emission inventory.9,12,36 These limitations reduce the usefulness of modeling studies to air quality managers and regulators interested in protecting specific water bodies. Field studies can provide a higher spatial and temporal resolution that is very specific to a receptor site, and therefore are potentially powerful tools for defining source–receptor relationships.
In this study, we present a methodology for identifying source regions and calculating an estimate of source impacts on atmospheric concentrations of GEM and RM at a receptor site. The methodology used semi-continuous real time GEM, PHg and RGM measurements made at a rural site for one year and at a predominantly downwind urban site during the following year. The objectives of the study were achieved by: (i) understanding the differences between rural and urban GEM, PHg and RGM concentration measurements; (ii) quantifying the local and regional impact of GEM, PHg and RGM point sources; and (iii) understanding the seasonality of GEM, PHg and RGM concentrations at both sites.
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Fig. 1 Monthly averages of gaseous elemental mercury (GEM) concentrations at Devil’s Lake State Park, WI, between April 2003 and March 2004 (a) and Milwaukee, WI, between July 2004 and May 2005 (b). The dashed lines represent the annual GEM averages of 1.61 ± 0.01 ng m–3 (a) and 2.48 ± 0.02 ng m–3 (b). All uncertainties are represented with standard errors of the averages. |
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Fig. 2 Monthly averages of particulate mercury (PHg) and reactive gaseous mercury (RGM) concentrations at Devil’s Lake State Park, WI, between April 2003 and March 2004 (a), and Milwaukee, WI, between July 2004 and May 2005 (b). The dashed line represents the PHg and RGM annual averages of: (i) 9.1 ± 0.1 pg m–3, and 5.3 ± 0.2 pg m–3 (a); and (ii) 11.8 ± 0.3 pg m–3, and 10.3 ± 0.2 pg m–3 (b). All uncertainties are represented with standard errors of the averages. |
Fig. 2a and b exhibited seasonal variations in the relative distribution of RM between RGM and PHg. This was caused by the effect of ambient temperature on the vapor pressure of RM, which is semi-volatile.8,45,46 More reactive mercury is associated with atmospheric particles during the winter months (80–90%), when the temperature is colder and the vapor pressure of RM is lower, than in the warmer summer months (40–60%). This phenomenon was observed at both the rural and the urban sites.
Fig. 3a and b show time series excerpts from RM concentration measurements made at DLSP between April 2003 and March 2004 (Fig. 3a), and Milwaukee between June 2004 and May 2005 (Fig. 3b). Fig. S2a–h† in the ESI show complete time series of GEM, PHg, RGM, and RM measurements made at both sites. Short-lived concentration increases in all of the three mercury species were interpreted as the impacts from point sources because they could not be explained by atmospheric oxidation of GEM to RM,7,32,33,47,48 or the re-emission of previously deposited mercury.49 The data presented in Fig. 3a and b, and Fig. S2a–h† revealed that increases in all of the mercury species at DLSP and Milwaukee did not fit the diurnal pattern that would be expected with re-emission of previously deposited RM and GEM.49–54 Furthermore, correlations between reactive mercury and ozone were very weak at both DLSP (r2 < 0.02) and Milwaukee (r2 < 0.05) before and after point source impacts were removed, indicating that at these locations, rapid GEM oxidation by ozone was not a strong contributor to short-lived increases in RM concentrations. Therefore, the assertion that short-lived increases in RM at DLSP and Milwaukee were predominantly caused by point sources was considered to be valid. Fig. S3–S5† in the ESI contain illustrative examples of individual short-lived increases linked through back trajectories to point sources listed in the EPA NEI.43 Forty-eight hour back trajectories were calculated for these point source impacts using the NOAA web based HYSPLIT model (v.4.8)55 and are also presented in Fig. S3–S5.† Absences in short-lived concentration increases at DLSP were defined as periods during which point sources did not impact the measurement site. These periods are shown in Fig. 3a as the sections of data represented with bold solid lines.
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Fig. 3 Excerpts of reactive mercury concentration measurements made at (a) Devil’s Lake State Park, WI, USA (DLSP) and (b) Milwaukee, WI, USA. The bold solid lines in (a) depict sections of the DSLP data sets during which no point source impacts were observed, that were used to calculate a reference mean and a point source threshold (shown in (a) and (b)). The reference mean and point source threshold values were then used to estimate point source impacts at DLSP and Milwaukee. |
In order to estimate the contributions of point sources to atmospheric mercury concentrations measured at DLSP and Milwaukee, a simple data analysis method was developed around this approach to identify point source impacts. Mean GEM and RM concentrations during periods without point source impacts were calculated for southern Wisconsin using the bold sections of the DLSP data presented in Fig. 3a and S2a–d† of the ESI: (i) GEM = 1.5 ± 0.2 ng m–3 (1 std. dev.; n = 1474); and, (ii) RM = 6.5 ± 4.0 pg m–3 (1 std. dev.; n = 1203). These mean concentrations will be hereafter referred to as the reference means. A threshold three times the standard deviation above the reference mean was used to sort the GEM and RM concentrations from both DLSP (Fig. 3a) and Milwaukee (Fig. 3b) into two classes: (i) periods when the receptor site was impacted by emissions from a point source; and (ii) periods of time when no impacts from point sources were observed. The concentration data were sorted using the histogram analysis tool available in Microsoft Excel®. Any concentrations higher than the point source thresholds (ST; i.e. for GEM, ST = 1.5 + (3 × 0.2) ng m–3 = 2.1 ng m–3) were considered to be impacts from point source emissions.
Once the concentrations had been sorted into the point source and non-point source classes, the following calculations were performed:
![]() | (1) |
![]() | (2) |
The results of the point source impact calculations are presented in Table 1 as the contributions of point sources to annual average GEM and RM concentrations measured at DLSP and Milwaukee. Table 1 shows that point sources have a lower impact on GEM concentrations than non-point sources at both DLSP and Milwaukee. At DLSP, the impact on the annual average GEM concentrations was 2.4%, while at Milwaukee the average contribution was 33%. Non-point sources contributed the majority of GEM at both sites: 98% at DLSP and 67% at Milwaukee. In contrast, the contributions of point sources to RM concentrations measured at DLSP and Milwaukee were much higher than contributions of point sources to GEM concentrations. Contributions of point sources to annual average RM concentrations were 48% at DLSP and 64% at Milwaukee, while non-point sources contributed 52% and 36%, respectively.
Devil’s Lake State Park, WI, USA | Milwaukee, WI, USA | |||||||||
---|---|---|---|---|---|---|---|---|---|---|
Annual average conc. | Contribution to annual average conc. | Annual average conc. | Contribution to annual average conc. | |||||||
Non-point source | % | Point source | % | Non-point source | % | Point source | % | |||
GEM/ng m–3 | 1.62 | 1.58 | 98 | 0.04 | 2 | 2.48 | 1.67 | 67 | 0.81 | 33 |
RM/pg m–3 | 11.8 | 6.0 | 52 | 5.7 | 48 | 21.7 | 7.9 | 36 | 13.8 | 64 |
In order to determine the source–receptor relationships at DLSP and Milwaukee, a knowledge of the types of nearby points sources and their locations was necessary. Fig. 4 is a wind rose of all of the large mercury emitting point sources (defined in this study as a source that emits more than 50 lbs yr–1) listed by either the NEI43 or the TRI.44 The Milwaukee and DLSP measurement sites are shown as a black circle at the origin, and a white circle to the west, respectively. The sources listed in the inventories were sorted into ten categories: coal fired power plants; industrial inorganic chemical manufacturers; chlor-alkali production facilities; metallurgical coke production facilities; steel mills/blast furnaces; asphalt and paving manufacturers; secondary non-ferrous metals refineries; petroleum refineries; paper mills; and miscellaneous, which included a crematory, a battery processing plant, a corn processing plant and a waste-to-energy facility. The mercury emissions (not speciated into GEM, PHg and RGM) from each of these categories are presented in Table 2. Coal fired power plants contributed 54% of the mercury emissions in the area, making them the highest emitting source category of mercury in the region. Industrial inorganic chemical manufacturers are the next largest emitting category (17%), followed by the remaining categories, each of which emit less than 8% of the total. Neither the NEI nor the TRI provided mercury speciation information for any of the facilities in the inventories. The 2002 NEI provided a table of nominal speciation information for a wide variety of process categories, but gave no assurance of the accuracy of these values for any particular facility. The speciation of mercury point source emissions controls how close to the emission source deposition occurs. For example, it is clear from Table 1 and Fig. S2a–h† that RM concentrations in both DLSP and Milwaukee were impacted by point sources emitting substantial amounts of reactive mercury. In order for environmental impact assessments to be successful, it is important that accurate speciation information be provided to air quality professionals for all sources emitting mercury.
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Fig. 4 A diagram that shows the locations of mercury point sources (>50 lbs yr–1) listed in the US EPA National Emission Inventory and Toxics Release Inventory. Also shown are the rural and urban measurement sites at Devil’s Lake State Park and Milwaukee, WI. |
Type of facility | Emissions/lbs yr–1 | % |
---|---|---|
Coal fired power plants | 6918 | 54 |
Industrial inorganic chemical manufacturers | 2122 | 17 |
Chlor-alkali plants | 1082 | 8 |
Coke production facilities | 650 | 5 |
Steel mills/blast furnaces | 565 | 4 |
Asphalt and paving | 443 | 3 |
Secondary non-ferrous metal refineries | 215 | 2 |
Petroleum refineries | 159 | 1 |
Paper mills | 154 | 1 |
Miscellaneous | 438 | 3 |
Total | 12![]() |
Fig. 5a and b show pollution roses for GEM and RM at DLSP. The data presented were greater than the point source threshold (ST), so that only point source impacts are shown. Fig. 5a and b revealed that elevated concentrations of GEM and RM came predominantly from the east and southeast, southwest, and northwest. The data in Fig. 5a and b corresponded well with the NEI and TRI source locations presented in Fig. 4. DLSP was clearly impacted by point sources located in counties to the east and southeast (Columbia, WI; Milwaukee, WI; Cook, IL; DuPage, IL; Will, IL; Lake, IL), southwest (Grant, WI; Linn, IA), and northwest (Vernon, WI; Allamakee, IA).
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Fig. 5 Pollution roses for atmospheric gaseous elemental mercury (a) and reactive mercury (b) for Devil’s Lake State Park, WI, between April 2003 and March 2004. Both pollution roses only show data above the point source threshold. |
Fig. 6a and b are pollution roses of GEM and RM concentrations measured in Milwaukee that were greater than the ST. The figures show that elevated concentrations of GEM and RM came predominantly from 110° to 240° and 0° to 45°. A comparison of the source locations in Fig. 4, with the data presented in Fig. 6a and b, shows that the elevated concentrations arriving at the measurement site from the south are likely due to emissions from point sources in the following counties: Milwaukee (WI), Kenosha (WI), Cook (IL), Lake (IL), DuPage (IL), Will (IL), Lake (IN), Jasper (IN) and Porter (IN).43 The elevated concentrations that approached the measurement site from the east (the direction of Lake Michigan) and west were probably due to emissions from point sources to the north and south of Milwaukee (Fig. S4)† combined with air recirculation during lake breeze meteorological effects.56–58 Elevated concentrations from the east were also caused by point sources in Michigan counties Muskegon and Ottawa on the opposite shore of Lake Michigan impacting the site during easterly wind patterns (Fig. S5).† It is possible that some of the mercury observed from non-point source directions could have been emitted by small point sources (less than 50 lbs yr–1) or unregistered point sources distributed throughout the city. The recirculation of air during lake breeze events made it difficult to distinguish between recirculated emissions from known point sources, and emissions from unregistered point sources. In order for effective air quality management decisions to be made for Milwaukee, the effect of the lake breeze system on the recirculation of known RM and GEM point source emissions will need to be determined.
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Fig. 6 Pollution roses for atmospheric gaseous elemental mercury (a) and reactive mercury (b) at Milwaukee, WI between July 2004 and May 2005. Both pollution roses only show data above the point source threshold. |
This study demonstrates the value of high time resolved mercury concentration and speciation measurements made over 1 year at receptor sites deemed sensitive to mercury impacts from point sources. From such measurements, it is possible to estimate the point source impacts on atmospheric mercury concentrations, and therefore dry and wet deposition, using the data analysis method described. These impact estimates can then be used to assess the potential success of proposed air quality management schemes on mitigating the deposition of reactive mercury to sensitive receptors.
Footnote |
† Electronic supplementary information (ESI) available: Supplementary information, Fig. S1–S5. See DOI: 10.1039/b710247j |
This journal is © The Royal Society of Chemistry 2008 |