3.1 Elevated Mixing Ratios of Selected VOCs
 Throughout the campaign, individual VOC mixing ratios varied by 1 to 4 orders of magnitude in the hourly samples (Table 1), indicating significant variability in the sampled air masses and local air quality conditions. This is demonstrated by the numerous periods of both relatively clean (e.g., portions of 20 and 26–27 February and 9–11 March) and polluted (e.g., portions of 19, 24–25, and 28 February and 2–3 and 5 March) air masses observed at BAO (Figure 2).
Figure 2. Time series of selected VOC mixing ratios measured during NACHTT. Samples were collected at the top of each hour at 22 m on the BAO tower. Missing data points indicate that mixing ratios were below the limit of detection. Concurrent peaks in mixing ratios of (a) the natural gas associated alkanes ethane and propane, and (b) dimethylsulfide (DMS) and (c) acetonitrile occurred throughout the campaign. Mixing ratios of (d) the combustion associated compounds ethyne and ethene tracked those of (e) aromatic compounds benzene and toluene, and to a lesser extent, (f) the oxygenated VOCs methanol and ethanol. (g) Mixing ratios of the alkyl nitrates 2-butyl nitrate and 2-pentyl nitrate and total alkyl nitrates showed a different temporal distribution indicating their primarily photochemical source.
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 Of the measured VOCs, the C2-C5 alkanes (excluding cyclopentane and 2,2-dimethyl propane), methanol, ethanol, acetone, and acetaldehyde were the 10 most abundant (Table 1). High mixing ratios of alkanes were observed throughout the entire monthlong campaign. At BAO, the average total measured VOC (C2-C10 alkane + C2-C6 alkene + C6-C10 aromatics + ethyne + biogenics + OVOCs + alkyl nitrates + DMS + COS) distribution was dominated by alkanes (mean ± standard deviation: 69 ± 17%). The C2-C5 alkanes represented approximately 66% of the total VOC mixing ratio, and the maximum total C2-C5 alkane mixing ratio was 324 ppbv. The prevalence of this high alkane signal over the monthlong campaign indicates the presence of a continuous source of alkane emissions.
 Alkane mixing ratios generally decreased with increasing carbon number. Of the measured VOCs, ethane (mean ± standard deviation: 22 ± 22 ppbv) and propane (16 ± 19 ppbv) were the two most abundant, with maximum mixing ratios of each exceeding 100 ppbv, representing 31% and 20% of total measured VOCs, respectively (Table 1). The C4 alkanes comprised 11% of the total measured VOCs with mean mixing ratios of 7.1 and 2.9 ppbv for n-butane and i-butane, respectively. Iso-pentane and n-pentane mixing ratios were nearly identical with means of 2.0 ppbv and medians of 1.05 ppbv.
 Relatively clean air masses were also sampled throughout the campaign (e.g., portions of 20 and 26 February and 10–12 March; see Figure 2). Mean mixing ratios from these periods were used to define the background VOC mixing ratios (see Table 1). Mean C2-C7 n-alkane mixing ratios for the entire campaign period (including the background observations) were between 7 times higher for ethane and 11 times higher for n-butane than the background levels. Mean C2-C6 n-alkane mixing ratios were approximately an order of magnitude higher than those observed at other semirural and remote sites, including Thompson Farm in Durham, New Hampshire, USA [Russo et al., 2010a]; Alert, Canada [Gautrois et al., 2003]; and Whiteface Mountain, New York [Gong and Demerjian, 1997]. Other alkanes exhibiting a similar enhancement above the background mixing ratio included n-heptane, n-octane, i-butane, i-pentane, 2-methylpentane, 3-methylpentane, cyclopentane, methylcyclopentane, cyclohexane, and methylcyclohexane.
 In contrast to the alkanes discussed above, mixing ratios of other VOCs were similar to measurements from other semirural sites. Mean mixing ratios of commonly used tracers of urban emissions, ethyne (633 ± 382 pptv) and tetrachloroethylene (C2Cl4) (10 ± 7 pptv), were only double the background mixing ratio and were similar to previously reported wintertime measurements at semirural sites [Russo et al., 2010a; Gautrois et al., 2003; Yokouchi et al., 1996; Gong and Demerjian, 1997]. Mean mixing ratios of the dominant alkenes, ethene (434 ± 465 pptv) and propene (104 ± 142 pptv), and aromatics, benzene (186 ± 126 pptv) and toluene (190 ± 190 pptv), were approximately 2 times higher than the background mixing ratio (Table 1); however, all were within one standard deviation of the wintertime mean reported at Thompson Farm [Russo et al., 2010a]. Mean OVOC mixing ratios exhibited smaller enhancements or no enhancement above background (factor of 1 to 2 times the background value). Mean mixing ratios of C3-C5 alkyl nitrates, excluding 1-propyl nitrate, were elevated above background by a factor of 2–4. Biogenic NMHCs were observed at low mixing ratios during the campaign as would be expected outside of the growing season.
 A comparison of NMHC mixing ratios observed at BAO with those reported in two major North American polluted areas (Galveston Bay near Houston, TX and Mexico City, Mexico), several U.S. cities, and a natural gas-producing region in the SW U.S. (Figure 3) demonstrates the magnitude of the observed mixing ratios. It should be noted that the results presented here were measured in late winter and include samples collected throughout the entire day. The mean and median C2-C4 alkane mixing ratios at BAO were higher than those observed in 28 different U.S. cities [Baker et al., 2008], while the range of pentane, ethene, ethyne, benzene, and toluene mixing ratios were comparable to major U.S. cities (Figure 3). The mean and median ethane and propane mixing ratios and the mean n-butane mixing ratio were higher than those observed in Galveston Bay and the Houston ship channel. The 75th percentile i-butane, n-pentane, and i-pentane mixing ratios at BAO were similar to the mean levels reported for Galveston Bay and the Houston ship channel [Gilman et al., 2009]. The C2-C4 alkane levels were higher at BAO than over the Anadarko Basin (Texas, Oklahoma, Kansas, eastern New Mexico) in 2001–2002 where Katzenstein et al.  attributed the observed high C1-C4 alkane mixing ratios to emissions from O&NG production. The mean ethane mixing ratio at BAO was higher than in Mexico City, and the mean C3-C5 alkane mixing ratios were within the range of the 24 h averages observed in and downwind of Mexico City [Apel et al., 2010]. The ethene, ethyne, benzene, and toluene levels at BAO were lower than in Mexico City. This comparison illustrates that the observed alkane mixing ratios at BAO represent a relatively large, unrecognized source of VOCs.
Figure 3. Box and whisker plot of several NMHCs measured at BAO during 18 February to 13 March 2011 in comparison to measurements from other polluted sites. In the box and whisker plot, the solid black line = median, dashed black line = mean, top and bottom of box = 75th and 25th percentiles, respectively, top and bottom whisker = 90th and 10th percentiles, respectively, and top and bottom black circle = 95th and 5th percentiles, respectively. Measurements from other polluted sites are represented by different symbols: red triangle = range of mean mixing ratios reported for 28 U.S. cities measured during summer between 1999 and 2005 [Baker et al., 2008]; blue square = mean mixing ratio in Houston ship channel/Galveston Bay, August–September 2006 [Gilman et al., 2009]; pink diamond = 24 h mean mixing ratio in Mexico City (TO, top pink diamond) and at site T1 30 km northeast of Mexico City (bottom pink diamond) during March 2006 [Apel et al., 2010]; green triangle = mean mixing ratio in SW U.S. during September 2001 (top green triangle) and April 2002 (bottom green triangle) [Katzenstein et al., 2003].
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3.4 Alkyl Nitrates
 Alkyl nitrates (RONO2) are produced from VOC precursors through reactions involving OH radical and NO [Roberts, 1990; Bertman et al., 1995]. Observed total C1-C5 alkyl nitrates (total RONO2 = sum of individual C1-C5 alkyl nitrates) were dominated by 2-butyl nitrate (27 ± 18 pptv), 2-propyl nitrate (16 ± 8 pptv), and 2-pentyl nitrate (12 ± 10 pptv) (Table 1 and Figure 2g). The mean total RONO2 mixing ratio (73 ± 44 pptv) was more than double that observed at a semirural site during winter in New Hampshire [Russo et al., 2010b] and similar to summer values reported for the polluted Pearl River Delta in China [Simpson et al., 2006]. However, comparisons with previous studies are complicated because the mixing ratios of individual alkyl nitrates in benchmark standards used by multiple research groups have changed [Simpson et al., 2011]. Mixing ratios of C3-C5 alkyl nitrates were highly correlated with 2-butyl nitrate (r2 > 0.95) throughout the campaign. Ethyl nitrate showed a weaker correlation with 2-butyl nitrate (r2 = 0.83), and methyl nitrate was completely uncorrelated (r2 ≈ 0), suggesting the importance of other primary or secondary sources.
 Alkyl nitrate mixing ratios were not correlated with alkane mixing ratios (Table 2a); however wind direction dependence, vertical profiles, diurnal variations in mixing ratios, and photochemical age estimations indicate strong local photochemical production from natural gas-associated precursors as the major source at BAO. Similar to the C2-C5 alkanes, the mean total RONO2 was also highest in the northeast wind sector (Figure 5e). The decrease in mixing ratio with height in vertical profiles of individual C2-C5 RONO2 (see 2-butyl nitrate in Figure 4e) and total RONO2 suggests production from a local ground level source. Mixing ratios of total RONO2 were also highest during the afternoon (12:00–16:00 MST) supporting a primarily local photochemical source.
 The photochemical age of the alkyl nitrates in sampled air masses was estimated by comparing observed ratios of individual RONO2 to their parent hydrocarbons to values predicted using a simplified sequential reaction scheme [Bertman et al., 1995] (reaction (1):
This reaction scheme assumes only OH-initiated photochemical production of a given alkyl nitrate (RONO2) from its parent hydrocarbon (RH) at the rate kA and photochemical removal at the rate kB as the only removal mechanism. The evolution of alkyl nitrates in a given air mass is then described by equation (1):
Where the alkyl nitrate to parent hydrocarbon ratio , at a given time t, is calculated assuming a starting ratio of 0. Predicted and observed alkyl nitrate to parent hydrocarbon ratios are plotted against the ratio for the most abundant alkyl nitrate, 2-butyl nitrate, and its precursor, n-butane, in Figure 8. The analysis could not be performed for methyl nitrate because methane was not measured. For C3-C5 alkyl nitrates, the observed values were similar to the expected values and indicated that most observations represented photochemical production of RONO2 from alkane emissions less than 1 day old (Figure 8). Observed ethyl nitrate to ethane ratios still indicated photochemical production from fresh emissions but were enhanced in ethyl nitrate relative to the expected values (Figure 8). This deviation suggests another source of ethyl nitrate at BAO, potentially production from oxidation of longer chain alkanes [Russo et al., 2010b; Bertman et al., 1995]. The high alkyl nitrate mixing ratios observed at BAO suggest that transport of alkyl nitrates produced in this region, and subsequent photolysis to yield NO, could be an important source of NOx in downwind locations with limited primary NOx emissions.
Figure 8. Photochemical age of alkyl nitrates. Comparison of observed alkyl nitrate:parent hydrocarbon ratios (green circles) to predicted values (gray squares) provide an estimate of photochemical age. Most observed alkyl nitrate:parent hydrocarbon ratios indicated that alkyl nitrates were less than 1 day old (red dashed line) and all were less than 3 days old (blue dashed line).
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3.5 OH Reactivity
 Tropospheric O3 is a criteria pollutant which can have detrimental effects on human health and vegetation [Lippmann, 1989] and is an important greenhouse gas with an estimated positive radiative forcing of 0.35 W/m2 [Solomon et al., 2007]. Ozone is produced in the troposphere from the OH-initiated oxidation of hydrocarbons in the presence of nitrogen oxides. Thus, the mixing ratios of O3 depend on the specific composition of hydrocarbons present in a region or an air mass. The OH reactivity is a measure of the initial rate of peroxy radical formation and can be interpreted as the potential of a specific compound to ultimately produce O3. Here we calculate the OH reactivity of the trace gases observed at BAO as a measure of their O3 production potential and estimate the OH reactivity attributable to natural gas associated emissions.
 A compound's OH reactivity is calculated as the product of its concentration [X] in molecules cm−3 and its OH rate constant kOH, X in cm3 molecule−1 s−1 [Atkinson, 2003; Atkinson et al., 1997, 2006; Atkinson and Arey, 2003] (equation (2)):
Reactivity was calculated for CO, CH4, NO2, and the VOCs (C2-C10 alkanes, C2-C6 alkenes, C6-C10 aromatics, ethyne, biogenic hydrocarbons, acetonitrile, COS, DMS, and OVOCs) in all of the hourly samples (Figure 9). In order to better estimate total OH reactivity of VOCs at BAO, constant values for CH4 and formaldehyde were used because these compounds were not measured during the NACHTT campaign. The mean CH4 mixing ratio (1867 ppbv) reported by Petron et al.  was used. For formaldehyde, the average winter 24 h median values reported by Eisele et al. [2009, Table 4.1] at two Colorado Front Range sites was used. Lastly, to avoid biasing the analysis towards samples with a greater number of compounds above the limit of detection (LOD), we assumed that the mixing ratio for any observation below the LOD was a random value between 0 and the LOD. This assumption adds a small amount of random noise to the calculation (<0.01% of the calculated total OH reactivity) but results in more representative total OH reactivity values. The propagated uncertainty of the total OH reactivity estimates presented here was 55% and included the uncertainty of each compound's rate constant for reaction with OH and the uncertainty in the mixing ratio of the compound.
Figure 9. Percentage contributions of VOCs to OH reactivity for the entire campaign and during background periods for all measured VOCs plus NO2, CO and using constant values (a and b) for CH4 and formaldehyde which were not measured, (c and d) for NMHCs and OVOCs using a constant value for formaldehyde, and (e and f) for C2-C10 alkanes only. Numbers above each pie chart represent the total reactivity, and numbers within each slice are the percentage contribution of the given compound or class of compounds.
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 The overall mean calculated OH reactivity was 7.0 ± 5.0 s−1 with a maximum of 31.3 s−1 (Table 1 and Figure 9a). Nitrogen dioxide made the largest contribution to OH reactivity (over 40%) followed by the alkanes (21%), the OVOCs (13%), CO (12%), CH4 (6%), alkenes (3%), and the aromatics (1%) (Figure 9). The mean NO2, CO, and CH4 contributions to OH reactivity were 3.1, 0.64, and 0.29 s−1, respectively. The mean calculated OH reactivity at BAO was on the lower end of total OH reactivity values reported in the literature which typically range from between 5 s−1 and over 100 s−1 [see Lou et al., 2010]. Similar to the mixing ratios compared in Figure 3, the mean calculated OH reactivity at BAO was within the range of total summertime OH reactivities measured in moderately sized cities including Houston, TX (range of 7–25 s−1 [Mao et al., 2010]), Nashville, TN (mean ± standard deviation of 11.3 ± 4.8 s−1 [Kovacs et al., 2003]), and Mainze, Germany (range of 6–18 s−1 [Sinha et al., 2008]). Calculated wintertime OH reactivity at BAO was on the low end of ranges reported for megacities including Paris, France (range of 10–120 s−1 [Dolgorouky et al., 2012]), Mexico City, Mexico (range of 10–200 s−1 [Shirley et al., 2006]), New York City, NY (range of 10–100 s−1 [Ren et al., 2003, 2006]), and Tokyo, Japan (range of 10–100 s−1 [Yoshino et al., 2006]). However, these comparisons are complicated by the fact that the direct total OH reactivity measurements from the studies mentioned above include contributions from species not measured during NACHTT, and total OH reactivity measurements at the BAO site would likely be higher than the results of the summation method discussed here. However, the similar magnitudes of OH reactivity at BAO and in these cities again suggests that regional VOC emissions were more representative of urban air quality and that there is the potential for enhanced O3 formation downwind of this site.
 The mean total VOC contribution to OH reactivity, which excludes NO2, CH4, and CO, was 3.0 ± 2.7 s−1, which is effectively equivalent to the OH reactivity reported during this campaign by Gilman et al. , with a maximum of 14.6 s−1. The OH reactivity of the VOCs was dominated by the alkanes (1.8 ± 2.2 s−1, 49%), OVOCs (0.78 ± 0.37 s−1, 37%), and alkenes (0.24 ± 0.28 s−1, 9%) (Table 1 and Figure 9c). This result is consistent with the dominance of alkanes and OVOCs in the VOC mixing ratio signature discussed above. Propane and n-butane constituted the largest percentage of the alkane reactivity at 23% and 20%, respectively (Table 1 and Figure 9e). Ethane, i-butane, i-pentane, and n-pentane made similar contributions (8–10%) to the alkane reactivity. The dominance of propane in the alkane contribution to OH reactivity is caused by its greater rate constant for reaction with OH compared to ethane which had a higher mean mixing ratio than propane (section 3.1). The C6 alkanes (n-hexane, 2-methylpentane, 3-methylpentane, methylcyclopentane) each contributed 2–4% to the total OH reactivity of the VOCs (Table 1).
 Similar signatures dominated by alkane reactivity were observed in the megacities of Tokyo, Japan [Yoshino et al., 2006] and Mexico City, Mexico [Shirley et al., 2006]. Alkane emissions from petrochemical processing facilities in Houston, TX also resulted in a similar contribution from alkanes to total OH reactivity [Mao et al., 2010]. Comparisons to other rural or semirural sites are difficult because measurements in these areas are typically conducted during the summer to quantify the contribution of biogenic VOCs to OH reactivity. For example, the mean VOC reactivity at BAO was within one standard deviation of the mean reactivity (4.15 ± 2.64 s−1) reported during the summer at Thompson Farm, NH during the 2004 International Consortium for Atmospheric Research on Transport and Transformation (ICARTT) campaign [White et al., 2008]. However, biogenic hydrocarbons accounted for the bulk of summertime VOC OH reactivity during ICARTT but accounted for less than 1% of the OH reactivity at BAO.
 Following the alkanes, OH reactivity of the OVOCs was the second most important contributor to total OH reactivity of the VOCs during NACHTT. The estimated formaldehyde (0.20 s−1) and the measured acetaldehyde (0.24 ± 0.22 s−1) and ethanol (0.20 ± 0.16 s−1) accounted for the bulk of OVOC reactivity accounting for 31%, 26%, and 24% of the OVOC reactivity, respectively. While methanol was the most abundant OVOC, its slower rate of reaction with OH resulted in it contributing only 14% to the total OVOC reactivity.
 Other classes of compounds made minor contributions to the VOC reactivity at BAO. Alkenes accounted for 9% of the calculated reactivity (Figure 9c), with ethene (0.10 ± 0.10 s−1) and propene (0.08 ± 0.10 s−1) contributing 38% and 28%, respectively, to the total alkene reactivity (Table 1). The mean aromatic OH reactivity at BAO was only 4% of the total OH reactivity from VOCs (Figure 9c) and was dominated by the xylenes (0.04 ± 0.04 s−1), toluene (0.03 ± 0.03 s−1), and the trimethylbenzenes (0.02 ± 0.03 s−1) (Table 1).
 To estimate the effect of natural gas emissions on OH reactivity, we compare the mean alkane contribution to OH reactivity observed for the whole campaign to the mean alkane contribution to OH reactivity calculated during the background periods defined in section 3.1 above. Only the OH reactivity of alkanes was considered as they were the dominant VOCs associated with natural gas production as described above in section 3.2. Methane was also excluded from this analysis as it was not measured during NACHTT. The mean alkane contribution to OH reactivity under background conditions was 0.11 ± 0.04 s−1, 94% less than the mean alkane contribution to OH reactivity for the whole campaign. Percent contributions of individual C2-C10 alkanes to the total OH reactivity of the alkanes were similar under background conditions as compared to the whole campaign period underscoring the strong correlation between these compounds and their common source. The difference in mean OH reactivity from alkanes between the entire campaign and the background periods was 1.70 s−1, amounting to 24% of the mean total OH reactivity or 57% of the mean OH reactivity of the VOCs (excluding CH4, CO, and NO2) observed during the campaign. This value is similar to that reported by Gilman et al.  calculated using the results of their multivariate regression source apportionment. This additional OH reactivity attributable to natural gas emissions could have an important impact on local to regional tropospheric O3 formation. The longer photochemical lifetimes of alkanes during the winter result in transport of these compounds and subsequent increased O3 formation in downwind regions [e.g., Kemball-Cook et al., 2010]. Any additional input of reactive VOCs could be important for air quality in the Colorado Front Range as regional municipalities continue efforts to comply with federal air quality standards.
 The major VOCs emitted from natural gas processing, specifically the alkanes, have lifetimes ranging from days to months during the winter. Consequently, year-round natural gas VOC emissions will likely impact the air quality of downwind areas [i.e., Kemball-Cook et al., 2010]. In order to forecast and predict air quality, accurate local/regional emission rate estimates of photochemical smog precursors (i.e., VOCs) are necessary. Here we use the hourly samples from 22 m to calculate fluxes of alkanes and benzene at BAO and estimate regional emissions associated with natural gas production.
 Boundary layer development throughout the day and night was monitored using the meteorological parameters measured on the moving carriage on the BAO tower between the surface and 250 m. Vertical profiles of potential temperature were examined for each night to determine when the nocturnal boundary layer (NBL) was fully developed and to estimate the height of the NBL. A shallow nocturnal inversion layer with a depth of 10 to 80 m developed each night. The 22 m inlet height of the canister samples was below the top of the inversion layer on five of the 23 nights (18–19 February, 22–23 February, 27–28 February, 2–3 March, and 3–4 March). The variation of trace gas mixing ratios throughout the day can be calculated using the mass balance equation (equation (3)) [Talbot et al., 2005; Zhou et al., 2005; Sive et al., 2007; White et al., 2008]:
Where [XBL] is the concentration of compound X in the boundary layer, t is time, ER is the emission rate of compound X, H is the boundary layer depth, P is the chemical production rate of compound X, kOH is the rate constant for the reaction of compound X with OH, [OH] is the concentration of OH, Ve is the vertical transfer coefficient, X is the concentration of compound X in the mixed layer above the boundary layer, and advection is a term representing horizontal transport of compound X. The NMHCs investigated here are not produced chemically in the atmosphere; thus, chemical production P reduces to 0. The only important removal mechanism for alkanes is reaction with OH. If only measurements made at night are used, then chemical removal by OH, kOH[OH][XBL], can be neglected. The nitrate radical (NO3) is a potential nocturnal oxidant; however, given the low reaction rate constant for the reaction between NO3 and the alkanes considered here, this chemical removal term can be neglected. Under the stable NBL with low wind speeds and relatively homogeneous regional emissions, it can be assumed that advection and vertical mixing are negligible, and changes in VOC mixing ratios can be attributed to local emissions. Based on our assumptions, equation (3) reduces to the following (equation (4)):
Thus, the flux or emission rate ER can be calculated when the change in trace gas concentration per unit time under the nocturnal boundary layer and the boundary layer height H are known. We used a nocturnal boundary layer height of 40 m for these calculations based on observed potential temperature profiles (not shown) for the five nights used in the calculation.
 Hourly average mixing ratios of natural gas associated VOCs varied diurnally. Mixing ratios of these compounds (ethane, propane, i-butane, n-butane, i-pentane, n-pentane, n-hexane, 2-methylpentane, 3-methylpentane, cyclopentane, methylcyclopentane, cyclohexane, methylcyclohexane, and benzene) generally increased under the stable NBL and reached their maximum hourly average mixing ratios at 08:00–09:00 h MST (Figure 10). This nocturnal trend was followed by a steady decrease after sunrise as wind speeds, vertical mixing, and photochemical processing increased until mixing ratios reached a minimum at 14:00 h MST and remained at minimum levels through approximately 22:00 h MST (Figure 10). Emission fluxes were calculated for the compounds which exhibited a strong correlation (r2 > 0.7) between the change in mixing ratio and time between 01:00 and 05:00 h (Table 3). It is worth noting that only alkane and benzene mixing ratios increased (r2 = 0.71–0.98) consistently under the NBL suggesting that a strong local emission source, which was persistent throughout the night, was responsible for the increases observed.
Figure 10. Hourly mean mixing ratios of selected VOCs. Symbols represent the mean mixing ratio for all hourly samples collected at 22 m on the BAO tower. Error bars represent the standard error of the mean. (a–c) Most VOCs displayed a diurnal cycle with higher mixing ratios during the night and lower mixing ratios during the late afternoon. (d) The alkyl nitrates showed an opposite diurnal pattern indicating their photochemical production during the day. The gray box highlights the 01:00 to 05:00 MST time period used for calculating VOC fluxes.
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Table 3. Emission Rates of Selected VOCs at BAO and Extrapolated Regional Emission Rates
| ||BAO Emission Ratea||Wattenberg Fieldb||Weld Countyc||Previous Estimates for Weld Countyd|
|Compound||109 molecules cm−2 s−1||r2||Gg yr−1||Gg yr−1||Gg yr−1|
|Ethane||181 (81)||0.79||9.4 (2.1)||29 (9)|| |
|Propane||168 (80)||0.74||13 (3)||40 (14)||15–65|
|i-Butane||39 (17)||0.82||3.9 (0.8)||12 (3)|| |
|n-Butane||97 (41)||0.84||9.8 (1.9)||31 (8)||5–42|
|i-Pentane||30 (12)||0.86||3.7 (0.6)||12 (3)||1–20|
|n-Pentane||28 (11)||0.88||3.4 (0.6)||11 (2)||1–20|
|n-Hexane||6 (2)||0.86||0.89 (0.16)||2.8 (0.6)|| |
|n-Heptane||1.4 (0.6)||0.86||0.24 (0.04)||0.74 (0.17)|| |
|n-Octane||0.4 (0.2)||0.86||0.08 (0.01)||0.25 (0.06)|| |
|n-Nonane||0.2 (0.1)||0.78||0.03 (0.01)||0.11 (0.03)|| |
|Neopentane||0.2 (0.1)||0.94||0.03 (0.003)||0.09 (0.01)|| |
|2,3-Dimethylbutane||0.3 (0.1)||0.75||0.04 (0.01)||0.14 (0.05)|| |
|2,2-Dimethylbutane||0.4 (0.1)||0.86||0.05 (0.01)||0.17 (0.04)|| |
|2-Methylpentane||5.3 (2.2)||0.86||0.79 (0.14)||2.5 (0.6)|| |
|3-Methylpentane||2.8 (1.2)||0.83||0.41 (0.08)||1.3 (0.3)|| |
|2,4-Dimethylpentane||0.2 (0.1)||0.80||0.04 (0.01)||0.12 (0.04)|| |
|2,3-Dimethylpentane||0.4 (0.2)||0.86||0.07 (0.01)||0.21 (0.05)|| |
|2-Methylhexane||0.4 (0.2)||0.74||0.07 (0.02)||0.22 (0.07)|| |
|2,2,4-Trimethylpentane||0.5 (0.2)||0.83||0.09 (0.02)||0.29 (0.08)|| |
|2-Methylheptane||0.3 (0.1)||0.91||0.06 (0.01)||0.19 (0.03)|| |
|3-Methylheptane||0.2 (0.1)||0.94||0.05 (0.01)||0.15 (0.02)|| |
|Cyclopentane||1.5 (0.6)||0.90||0.18 (0.03)||0.57 (0.11)|| |
|Methylcyclopentane||3.0 (1.0)||0.98||0.44 (0.03)||1.4 (0.1)|| |
|Cyclohexane||1.7 (0.6)||0.97||0.25 (0.02)||0.8 (0.08)|| |
|Methylcyclohexane||1.5 (0.5)||0.98||0.25 (0.01)||0.77 (0.06)|| |
|Benzene||1.4 (0.7)||0.71||0.18 (0.05)||0.57 (0.21)||0.05–2|
 Calculated emission fluxes generally followed the trend in mixing ratios at BAO with the highest fluxes calculated for ethane (181 ± 81 × 109 molecules cm−2 s−1) and propane (168 ± 80 × 109 molecules cm−2 s−1) followed by n-butane (97 ± 41 × 109 molecules cm−2 s−1) and fluxes of other compounds following a decreasing trend with increasing number of carbon atoms (Table 3). The total flux of the NMHCs listed in Table 3 (570 × 109 molecules cm−2 s−1) was higher than the O&NG associated total VOC flux of 370 × 109 molecules cm−2 s−1 reported in 2007 for Garfield County, CO, the county with the second highest number of natural gas wells in Colorado [Colorado Department of Public Health and Environment (CDPHE), 2009]. This value was calculated from total VOC annual emissions reported in CDPHE , using a VOC mixing ratio-weighted molecular mass of 53.5 g mole−1 and an estimated area of natural gas operations in Garfield County of 2000 km2. In addition, the ethane emission rate for a portion of the Anadarko Basin in the southwest U.S. (study area of 590,400 km2) has been estimated at 0.3–0.5 Tg yr−1 [Katzenstein et al., 2003]. This emission rate corresponds to a flux of 32–54 × 109 molecules cm−2 s−1, which is a factor of 3–5 lower than that estimated during NACHTT. While there is considerable uncertainty associated with the calculation of emission fluxes in all of these studies and the comparison is complicated by differences in the year of the measurement, it is likely that differences in the spatial density of wells or natural gas production volume between the Anadarko Basin, Garfield County, and Weld County contributed to the differences in estimated emission rates, and future studies should examine quantitative relationships between these factors and trace gas emissions.
 In order to compare emissions observed during NACHTT to previous emission rate estimates for this region, emission fluxes were extrapolated to regional emission rates for the Wattenberg Field and for Weld County, CO (Table 3). Regional emissions for the Wattenberg Field were calculated using an area of 2530 km2 corresponding to the area of highest spatial density of natural gas wells (see Figure 1), while the regional emission rates estimated using the area of Weld County, 10,341 km2, includes regions outside of the Wattenberg Field where the spatial density of natural gas wells is lower. Our estimates assume that NMHC emission rates at BAO were representative of regional emissions regardless of the number of wells in a given area; therefore, the regional emission rate estimates for the Wattenberg Field and Weld County represent lower bound and upper bound estimates, respectively.
 Petron et al.  reported both bottom-up and top-down estimates of emissions of methane, propane, n-butane, i-pentane, n-pentane, and benzene from venting and flaring of natural gas in all of Weld County. Their three top-down emission scenarios were constrained by the observed methane to propane ratio but utilized different inventory-based molar ratio estimates for vented natural gas resulting in the range of emission rates shown in Table 3. A comparison of the top-down and bottom-up estimates by Petron et al.  suggests that O&NG industry emissions estimates of the measured compounds are underreported by a factor of approximately 1.5–2, with the exception of benzene, which was lower by a factor of approximately 5. Propane, n-butane, pentane, and benzene emission rates extrapolated to the area of the Wattenberg Field during NACHTT were on the low end of the range reported by Petron et al.  for Weld County (Table 3), while those extrapolated to the area of Weld County were in the middle of the range of previous estimates. The average of the two emission rate estimates for the Wattenberg Field and for Weld County calculated for NACHTT agrees most closely with the Petron et al.  top-down emission estimate scenario that used the median methane to propane molar ratio from a survey of 77 Wattenberg natural gas samples (median = 15.43; [COGCC, 2007]) and resulted in their lowest top-down emission estimate (28.5 Gg yr−1 propane, 13 Gg yr−1 n-butane, 6 Gg yr−1 i-pentane, 6 Gg yr−1 n-pentane, and 400 Mg yr−1 benzene). Our results support the findings of Petron et al.  but suggest that the discrepancy between emission inventories and actual emissions of these compounds is closer to a factor of 1.5 for the alkanes and a factor of 3 for benzene.
 Several factors could have contributed to the lower estimated emissions reported in this study compared to those of Petron et al. . BAO is located within the Wattenberg Field, but the highest spatial density of natural gas wells was northeast of BAO, and emission rates may be higher in areas with a greater well density. Similarly, emission rates are likely lower throughout the remainder of the county where the spatial density of natural gas wells is lower. Thus, there is considerable uncertainty in the assumption that the emission rate at BAO is representative of all of Weld County or of the Wattenberg Field. Also, NACHTT was conducted during the winter when emissions of these volatile compounds are likely different than during the warmer summer months included in the Petron et al.  study. Our estimates are based on observations of emission rates on five nights; longer-term observations through multiple seasons may prove more representative. Lastly, differences in well drilling and completion practices and natural gas storage, transport, and processing practices may also contribute to the differences in emission rates discussed above.
 Annual emissions estimates for Weld County from NACHTT can be compared to previous estimates to examine the effect of the regulations instituted by the CDPHE in 2008. Table 4 compares our measurement-based emissions estimates for the combined total of the natural gas associated VOCs listed in Table 3, excluding ethane, to inventory-based estimates from 2006 (Western Regional Air Partnership Phase III inventory [Bar-Ilan et al., 2008b]). Ethane was excluded because we assumed that the 2006 inventory employed the EPA definition of VOCs, which excludes both methane and ethane, as Bar-Ilan et al. [2008a, 2008b] described their VOC estimates as the total of all state or EPA permitted sources. While the number of active wells increased by 61% and gas production increased by 27% from 2006 to 2011, the change in natural gas associated VOC emissions ranged from a 40% decrease to an 86% increase based on our lower and upper bound estimates, respectively. Using the average of our lower and upper bound estimates of VOC emissions yields a 23% increase relative to 2006 estimates, which is similar to the percent increase in natural gas production from Weld County. Normalizing VOC emissions to the number of gas wells and to the amount of natural gas produced shows that VOC emissions from each individual well likely decreased between 2006 and 2011 but that the large increase in the number of active wells and the decrease in the amount of gas produced per well over this time period resulted in nearly equivalent VOC emissions per unit of gas produced. These comparisons indicate that the 2008 regulations enacted by the CDPHE were most likely effective in controlling VOC emissions but that the continued development of natural gas resources in the region has offset the gains achieved and resulted in greater overall VOC emissions from the region.
Table 4. Comparison of VOC Emissions Estimates From Weld County in 2006 and 2011
| ||Natural Gas Production||VOC Emissions|
|Year||Active Wellsa||Gas Produced (Bcf)b||Tons yr−1c||Tons yr−1 well−1||Tons yr−1 Bcf gas−1|
 The EPA recently instituted new requirements for emission control and green completion of natural gas wells [EPA, 2012]. State and municipal governments across the country are also reevaluating their regulations in response to public concerns about human health and environmental quality. The baseline values for the wide suite of compounds reported here will be useful for determining the efficacy of future regulatory and policy changes on controlling emissions.