Longterm ozone trends at rural ozone monitoring sites across the ...

3 downloads 6598 Views 2MB Size Report
monitoring sites across the United States, 1990–2010, J. Geophys. Res., 117 .... ozone plumes from the free troposphere to the surface of the western USA with ...
JOURNAL OF GEOPHYSICAL RESEARCH, VOL. 117, D22307, doi:10.1029/2012JD018261, 2012

Long-term ozone trends at rural ozone monitoring sites across the United States, 1990–2010 Owen R. Cooper,1,2 Ru-Shan Gao,2 David Tarasick,3 Thierry Leblanc,4 and Colm Sweeney1,2 Received 7 June 2012; revised 9 October 2012; accepted 10 October 2012; published 28 November 2012.

[1] This analysis provides an up-to-date assessment of long-term (1990–2010) rural ozone trends using all available data in the western (12 sites) and eastern (41 sites) USA. Rather than focus solely on average ozone values or air quality standard violations, we consider the full range of ozone values, reporting trends for the 5th, 50th and 95th percentiles. Domestic ozone precursor emissions decreased strongly during 1990–2010. Accordingly 83%, 66% and 20% of summertime eastern U.S. sites experienced statistically significant ozone decreases in the 95th, 50th and 5th percentiles, respectively. During spring 43% of the eastern sites have statistically significant ozone decreases for the 95th percentile with no sites showing a significant increase. At the 50th percentile there is little overall change in the eastern U.S. In contrast, only 17% (2 sites) and 8% (1 site) of summertime western U.S. sites have statistically significant ozone decreases in the 95th and 50th percentiles, respectively. During spring no western site has a significant decrease, while 50% have a significant median increase. This dichotomy in U.S. ozone trends is discussed in terms of changing anthropogenic and biomass burning emissions. Consideration is given to the concept that increasing baseline ozone flowing into the western U.S. is counteracting ozone reductions due to domestic emission reductions. An update to the springtime free tropospheric ozone trend above western North America shows that ozone has increased significantly from 1995 to 2011 at the rate of 0.41  0.27 ppbv yr 1. Finally, the ozone changes are examined in relation to regional temperature trends. Citation: Cooper, O. R., R.-S. Gao, D. Tarasick, T. Leblanc, and C. Sweeney (2012), Long-term ozone trends at rural ozone monitoring sites across the United States, 1990–2010, J. Geophys. Res., 117, D22307, doi:10.1029/2012JD018261.

1. Introduction [2] Within the United States, ozone has been recognized since the 1940s and 1950s as an air pollutant detrimental to human health and vegetation [National Research Council, 1991], and today ozone pollution is a widespread problem in many regions around the world resulting from both local and distant sources [Dentener et al., 2011]. The U.S. Clean Air Act of 1970 created a National Ambient Air Quality Standard (NAAQS) for ozone requiring regions that exceed the standard to implement ozone precursor emission reductions such that the NAAQS can be attained. Two decades later, a National Research Council [1991] review of ozone 1 Cooperative Institute for Research in Environmental Sciences, University of Colorado Boulder, Boulder, Colorado, USA. 2 NOAA Earth System Research Laboratory, Boulder, Colorado, USA. 3 Experimental Studies Research Division, Meteorological Service of Canada, Environment Canada, Ottawa, Ontario, Canada. 4 Table Mountain Facility, Jet Propulsion Laboratory, California Institute of Technology, Pasadena, California, USA.

Corresponding author: O. R. Cooper, NOAA Earth System Research Laboratory, CSD04, 325 Broadway, Boulder, CO 80305, USA. ([email protected]) ©2012. American Geophysical Union. All Rights Reserved. 0148-0227/12/2012JD018261

trends across the country concluded that “Despite the major regulatory and pollution-control programs of the past 20 years, efforts to attain the National Ambient Air Quality Standard for ozone largely have failed.” In those days the U. S. Environmental Protection Agency’s (EPA) principal statistical measure of ozone concentrations was the composite nationwide average of the second highest 1-h daily maximum concentrations in a given year, which decreased by 14% between 1980 and 1989 but with large interannual variability. The National Research Council noted that it was difficult to ascertain progress in reducing ozone pollution when this primary metric was highly susceptible to meteorological variation and when only 10 years of data were available. [3] As time went by, additional years of data allowed for more robust trend analyses, and different metrics were devised to give a better idea of ozone changes relevant for human and vegetative exposure [Lefohn and Shadwick, 1991; Fiore et al., 1998, and references therein]. Using summertime afternoon ozone data from 549 ozone monitoring sites across the U.S., Fiore et al. [1998] found that no large region of the U.S. experienced a significant increase in ozone during 1980–1995. Significant ozone decrease were mainly confined to the New York City, Los Angeles and

D22307

1 of 24

D22307

COOPER ET AL.: RURAL U.S. OZONE TRENDS, 1990-2010

Chicago metropolitan regions, with decreases of the 90th percentile being more pronounced than for the medians. These urban ozone decreases appeared to be related to decreases in volatile organic compound (VOC) emissions as national NOx emissions were constant over the 15 year period. [4] Up-to-date national ozone trend information is provided by the U.S. EPA on its website (http://www.epa.gov/ airtrends/ozone.html). Using the annual 4th highest daily maximum 8-h average ozone mixing ratio (the metric used to determine exceedance of the NAAQS for ozone, currently set at 75 ppbv) from 247 monitoring sites across the country the EPA shows that ozone decreased from 97 ppbv during 1980–1984 to 75 ppbv during 2006–2010. The overall national trend is a 28% decrease from 1980 to 2010 and a 17% decrease from 1990 to 2010. These trends are based on data that were primarily collected in urban areas with most sites concentrated in the eastern USA or along the U.S. west coast. Using May– September average ozone values at 180 rural and urban sites across the country the EPA shows that 2001–2010 ozone decreases were similar for rural and urban sites. However, ozone decreases were greater in the east than in the west. While the average of all sites shows a nationwide decrease, not all sites have decreasing ozone. From 1980 to 2008 71% of U.S. sites showed statistically significant decreases in the 4th highest daily maximum 8-h average ozone value, and 2% had statistically significant increases; for 1994–2008 these values were 51% and 1%, respectively [Lefohn et al., 2010]. As described below, these results show that U.S. emission controls are reducing the frequency and magnitude of extreme ozone episodes. [5] Because ozone is a secondary pollutant its production is closely related to the primary emissions of precursor trace gases. The EPA National Emissions Inventory estimates that for the period 1990–2010, total U.S. anthropogenic emissions declined by 49%, 58% and 44% for the ozone precursors NOx (=NO + NO2), CO and VOC, respectively (U.S. Environmental Protection Agency, National Emissions Inventory (NEI) air pollutant emissions trends data, http://www. epa.gov/ttnchie1/trends/, 2012, hereinafter referred to as EPA, online report, 2012). According to several recent photochemical modeling studies, these reductions in ozone precursors are responsible for the downward trend in ozone observed in the eastern U.S. and in southern California, at least for the median and extreme ozone values during the ozone season that lasts from May through September [Frost et al., 2006; Kim et al., 2006; Gilliland et al., 2008; U.S. Environmental Protection Agency, 2009; Butler et al., 2011; Hogrefe et al., 2011; Pozzoli et al., 2011]. [6] While emission reductions appear to be reducing the frequency of high ozone events, several studies have shown that mixing ratios of the lower ozone percentiles, such as the 5th, 10th or 20th percentiles are increasing across the country [Lin et al., 2000; Lefohn et al., 2010; Hogrefe et al., 2011]. Lin et al. [2000] suggested that the increase in the lower ozone percentiles was due to an increase in the baseline ozone flowing into the U.S. (baseline ozone is defined as the observed ozone at a site when it is not influenced by recent, locally emitted or produced anthropogenic pollution [Dentener et al., 2011]). A modeling study by Jacob et al. [1999] lent support to this hypothesis, which calculated that rising Asian anthropogenic emissions can increase the quantity of baseline ozone impacting the U.S., even offsetting ozone reductions caused

D22307

by decreased domestic emissions. Many studies concurrent with, and subsequent to the findings of Jacob et al. [1999] have provided ample in situ evidence that Asian pollution plumes reach the lower free troposphere and the surface of western North America [Jaffe et al., 1999; McKendry et al., 2001; Parrish et al., 2004a; Dentener et al., 2011 ]. [7] The possibility of significant changes in baseline ozone due to rising Asian emissions has the strongest implications for the western United States, the region of the country most likely to be impacted by Asian outflow [Reidmiller et al., 2009; Brown-Steiner and Hess, 2011; Lin et al., 2012a]. Several recent studies have shown that mean and median ozone values are increasing at some rural sites in the western USA, especially during spring. Analysis of springtime ozone measurements from rural sites on the U.S. west coast and Lassen Volcanic National Monument in elevated northeast California, as well as hydrocarbon measurements from field campaigns [Parrish et al., 2004b], showed that ozone increased from the 1980s through 2002, and that the photochemical environment of the eastern North Pacific Ocean marine boundary layer had also changed [Jaffe et al., 2003; Parrish et al., 2004b]. These studies suggested that the changes in the baseline air entering North America were related to increased anthropogenic emissions in Asia. Focusing on elevated National Park ozone monitoring sites across western North America, Jaffe and Ray [2007] found that deseasonalized mean daytime ozone values increased at 7 of the 9 sites analyzed. Similarly, Oltmans et al. [2008] reported that the positive trend at elevated Lassen Volcanic National Monument continued through 2006, the most recently available data at the time of their analysis. Returning to the west coast, Parrish et al. [2009] combined surface ozone measurements from several U.S. coastal sites and filtered the data to characterize baseline air masses flowing onshore from the North Pacific Ocean marine boundary layer. The analysis showed statistically significant increases in mean ozone for winter, spring and summer from the 1980s through 2007. An increasing ozone trend has also been measured in the free troposphere above western North America during springtime (1984–2008), with the strongest rate of increase associated with the air masses with greater transport from the atmospheric boundary layer of south and east Asia [Cooper et al., 2010]. In response to these findings, several recent studies have focused on the transport mechanisms that bring enhanced ozone plumes from the free troposphere to the surface of the western USA with an emphasis on identifying plumes heavily influenced by Asian emissions [Weiss-Penzias et al., 2006; Parrish et al., 2010; Huang et al., 2010; Cooper et al., 2011; Pfister et al., 2011; Lin et al., 2012a]. [8] The studies described above have made great progress toward understanding and quantifying the impact of local and upwind emissions on U.S. surface ozone trends. In addition many recent studies report rural and urban ozone trends across the United States and provide insight into the reasons for the trends, or lack thereof [Chan, 2009; Chan and Vet, 2010; Lefohn et al., 2010; National Park Service, 2010; Hogrefe et al., 2011; Pozzoli et al., 2011; Sather and Cavender, 2012]. Individually, these studies have various foci, such as exceedances of the NAAQS for ozone, and cover a range of time periods through 2005 or 2008. However, the literature is lacking a current overview of U.S. ozone trends that is sensitive to the seasonal variability in

2 of 24

D22307

COOPER ET AL.: RURAL U.S. OZONE TRENDS, 1990-2010

ozone chemistry and transport processes, and that is aimed at understanding long-term (20-year) trends for low, mid- and high mixing ratios. The goals of this study are to provide an up-to-date nationwide survey of surface ozone trends, discuss the observed trends in light of regional and global changes in ozone precursor emissions and temperature, and to serve as a benchmark for modeling studies of U.S. ozone trends. Specifically, this study is aimed at understanding regional trends in ozone and therefore only analyzes rural measurements, which are less susceptible to ozone titration by fresh NO emissions, and only examines daytime measurements (11:00– 16:00 local time) when the atmospheric boundary layer is well mixed and when nighttime surface deposition is not an issue. Rather than focusing on the broad “ozone season” of May–September as is generally done by studies interested in extreme ozone events, this study examines ozone trends separately for spring (March, April and May), and summer (June, July, August) which have contrasting transport patterns [Moody et al., 1998; Cooper et al., 2002] and photochemistry [Emmons et al., 2003]. While most of the emphasis is placed on spring and summer when exceedances of the NAAQS for ozone are most likely, winter ozone trends are also briefly reported to allow the contrast of ozone trends between seasons with strong and weak domestic ozone production. As noted by the studies above, ozone trends can differ for high, mid- and low ozone events, therefore trends are calculated for the 95th, 50th and 5th ozone percentiles.

2. Method [9] Due to the interannual variability of ozone [National Research Council, 1991; Parrish et al., 2012], determining robust ozone trends requires many years of data and this study focuses on the 21-year period of 1990–2010. The rural ozone measurements in this study were obtained from the U.S National Park Service (NPS) and The Clean Air Status and Trends Network (CASTNET). Data from one additional rural site, Whiteface Mountain Summit, New York were obtained from the University of Albany. Information on the locations of the 53 sites retained for this study can be found in Table 1 and Figure 1. [10] The NPS Air Resources Division administers an extensive air monitoring program throughout the park system, designed to establish current air quality conditions and to assess long-term trends of air pollutants that affect park resources [National Park Service, 2010]. The data are also used to determine compliance with the NAAQS and to assess national and regional air pollution control policies. The NPS data used in this study were collected by the National Park Service Gaseous Pollutant Monitoring Program and downloaded from the NPS Gaseous Pollutant and Meteorological Data Access Page: http://ard-request.airresource.com. Additional NPS data from the Brevard Rd., Saratoga, and Mt. Greylock sites were downloaded from the U.S. EPA Air Quality System Datamart website: http:// www.epa.gov/ttn/airs/aqsdatamart/access/interface.htm. [11] CASTNET is a national air quality monitoring network (administered and operated by EPA’s Clean Air Markets Division) that collects data to assess trends in air quality, atmospheric deposition, and ecological effects due to changes in air pollutant emissions. The network began in 1991 with the incorporation of 50 sites from the National Dry

D22307

Deposition Network, which had been in operation since 1987. CASTNET operates more than 80 regional sites throughout the contiguous United States, Alaska, and Canada, located in rural areas where urban influences are minimal. The CASTNET data used in this study were downloaded from: http://epa.gov/castnet/javaweb/index.html. [12] A total of 53 sites were available for this analysis, all spanning the 1990–2010 period. While all of the sites are in rural areas, none can be considered to be truly remote from all anthropogenic emission sources. To minimize any local effects and to ensure that the measurements are as representative as possible of regional ozone, only midday measurements were analyzed, when the daytime atmospheric boundary layer is well-mixed. Twelve sites are located in the western U.S. and 41 sites in the eastern U.S. No sites were available in the Great Plains. [13] Two of the sites experienced monitor relocations. From 1990 to 1996 measurements from Yellowstone National Park were made at the Lake Yellowstone site on the northwest shore of Lake Yellowstone. In 1996 the monitor was moved 1.5 km to the northwest to the Water Tank site, which is located 40 m above Lake Yellowstone. Because the monitor was moved simultaneous measurements at both sites were not available for comparison. While the local transport patterns are slightly different for the two sites due to the impact of the lake breeze at the lower elevation site, using data from the well-mixed midday period minimizes the differences as discussed by Jaffe and Ray [2007]. Measurements from Joshua Tree National Park were made at the Lost Horse Ranger Station site from 1990 until September 1993, when the monitor was moved to the Black Rock site, 20 km to the northwest. The terrain is similar for the two sites but the new location is closer to urban areas. While the relocation of the monitor covers a substantial distance, air quality in Joshua Tree National Park is overwhelmingly dominated by air pollution transport from the Los Angeles Basin which regularly places the rural park in exceedance of the NAAQS for ozone [Sullivan et al., 2001; Rosenthal et al., 2003; Langford et al., 2010]. Therefore the ozone measured at Joshua Tree National Park is largely a reflection of emissions in the Los Angeles Basin and the variation in local emissions is minor in comparison. [14] Jaffe and Ray [2007] report ozone trends through 2004 at several National Park Sites. They noted that the inlet height for many National Park monitoring sites was increased from 3.5 m to 10 m above ground level in the mid-1990s. These sites included several used in the present study: Lassen Volcanic National Park, Rocky Mountain National Park, Yellowstone National Park and Glacier National Park. Jaffe and Ray [2007] determined that the increase of the inlet height resulted in a statistically significant increase in ozone at night when the sites are typically influenced by temperature inversions, but they found no significant change in ozone during daytime conditions when surface warming has eroded the nighttime inversion. [15] Ozone trends were calculated as follows. In order to minimize local effects and to ensure that the ozone measurements are representative of the well-mixed daytime atmospheric boundary layer, only hourly average data reported at 11, 12, 13, 14, 15 and 16 local time were considered. Trends were calculated separately for spring (March, April, May), summer (June, July, August), and winter (December, January,

3 of 24

AL AR FL GA IL IL IN IN MA MA MD ME MI MI MS NC NC NC NH NJ NY NY NY OH OH OH PA PA PA

Chiricahua NM Chiricahua NM Grand Canyon NP Grand Canyon NP Sequoia/Kings Canyon NPs Lassen Volcanic NP Pinnacles NM Joshua Tree NP Joshua Tree NP Rocky Mountain NP

AZ AZ AZ AZ CA CA CA CA CA CO CO WY WY WY WY MT

4 of 24

Saratoga

Blue Ridge Parkway

Cape Cod NS Appalachian Trail

Yellowstone NP Yellowstone NP Glacier NP

Park Unit

State

Sand Mountain Caddo Valley Sumatra Georgia Station Alhambra Bondville Vincennes Salamonie Reservoir Fox Bottom Area Mount Greylock Summit Beltsville Ashland Ann Arbor Unionville Coffeeville Coweeta Cranberry Route 191 (S Brevard Rd) Woodstock Washington’s Crossing Connecticut Hill Stillwater Whiteface Mountain Summit Oxford Deer Creek State Park Lykens Laurel Hill State Park Penn State University Arendtsville

Entrance Station Entrance Station The Abyss The Abyss Lower Kaweah Manzanita Lake Fire Station SW of East entrance Black Rock Lost Horse Ranger Station Long’s Peak Gothic Pinedale Centennial Lake Yellowstone Water Tank West Glacier Horse Stables

Site

34.29 34.18 30.11 33.18 38.87 40.05 38.74 40.82 41.98 42.64 39.03 46.60 42.42 43.61 34.00 35.06 36.10 35.5 43.95 40.31 42.40 43.01 44.37 39.53 39.64 40.92 39.99 40.72 39.92

32.01 32.01 36.06 36.06 36.57 40.54 36.49 34.07 34.09 40.28 38.96 42.93 41.36 44.56 44.56 48.51

North Latitude

Elevation (m)

Western United States 109.39 1570 109.39 1570 112.18 2073 112.18 2073 118.78 1890 121.58 1756 121.16 335 116.39 1244 116.19 1265 105.55 2743 106.99 2926 109.79 2388 106.24 3178 110.39 2361 110.40 2400 113.00 976 Eastern United States 85.97 352 93.10 71 84.99 14 84.41 270 89.62 164 88.37 212 87.49 134 85.66 250 70.02 41 73.17 1140 76.82 46 68.41 235 83.90 267 83.36 201 89.80 134 83.43 686 82.05 1219 82.6 675 71.70 258 74.87 61 76.65 501 73.65 120 73.90 1,483 84.73 284 83.26 267 82.00 303 79.25 615 77.93 378 77.31 269

West Longitude

12/27/88 10/4/88 12/28/88 6/28/88 6/28/88 2/9/88 8/4/87 6/28/88 4/1/87 5/1/89 11/1/88 12/20/88 6/28/88 6/28/88 12/27/88 11/4/87 12/27/88 4/1/89 12/27/88 12/27/88 9/28/87 7/1/88 1/1/73 8/18/87 9/28/88 1/10/89 12/15/87 1/6/87 6/28/88

1/1/92 4/1/89 1/1/93 7/21/81 6/1/84 10/1/87 4/1/87 10/1/93 5/1/87 7/1/87 5/16/89 12/27/88 8/19/91 6/1/87 6/26/96 4/1/89

Start Date

6/25/96

9/22/93

12/31/92

9/1/92

End Date

CN CN CN CN CN CN CN CN NPS NPS CN CN CN CN CN CN CN NPS CN CN CN NPS U. of Albany/ NY DEC CN CN CN CN CN CN

CN CN CN NPS NPS CN CN CN NPS CN CN CN CN NPS CN CN

Network

SND152 CAD150 SUM156 GAS153 ALH157 BVL130 VIN140 SAL133 CACO-XX (CC) APTR-MG BEL116 ASH135 (AL) ANA115 UVL124 CVL151 COW137 PNF126 BLRI-RO WST109 WSP144 CTH110 SARA-ST WFMS (WF) OXF122 DCP114 LYK123 LRL117 PSU106 ARE128

CHA467 CHA267 GRC474 (GC) GRCA-AB (GC) SEKI-LK (SK) LAV410 (LV) PIN414 (PN) JOT403 (JT) JOTR-LH (JT) ROM406 (RM) GTH161 (GO) PND165 (PD) CNT169 (CN) YELL-LV (YS) YEL408 (YS) GLR468

Network Site Abbreviation

Table 1. Names and Locations of the 53 CASTNET (CN) and National Park Service Gaseous Pollutant Monitoring Program (NPS) Rural Ozone Monitoring Sites Used in This Studya

D22307 COOPER ET AL.: RURAL U.S. OZONE TRENDS, 1990-2010 D22307

CN CN CN NPS CN CN CN CN CN CN CN CN 1/12/88 1/3/89 3/22/88 7/1/88 7/23/88 6/12/89 5/1/83 6/2/87 11/3/87 9/27/88 11/10/87 1/19/88 384 622 302 1243 793 361 1073 920 150 472 210 510 80.15 78.77 85.73 83.61 83.94 83.83 78.44 80.56 78.31 90.60 80.85 79.66 41.43 41.60 36.04 35.70 35.63 36.47 38.52 37.33 37.17 45.21 38.88 39.09 Shenandoah NP

Great Smoky Mountains NP Great Smoky Mountains NP

M.K. Goddard Kane Experimental Forest Edgar Evins State Park Cove Mountain Look Rock Speedwell Big Meadows Horton Station Prince Edward Perkinstown Cedar Creek Park Parsons M.K. Goddard State Park

PA PA TN TN TN TN VA VA VA WI WV WV

a

MKG113 KEF112 ESP127 GRSM-CM (CM) GRS420 SPD111 SHN418 (BM) VPI120 PED108 PRK134 CDR119 PAR107

Network Site Abbreviation Network End Date Start Date Elevation (m) West Longitude North Latitude Site Park Unit State

Table 1. (continued)

NP, NM, and NS denote National Park, National Monument and National Seashore, respectively. Site abbreviations used in Figure 1 are listed in parentheses in the Network Site Abbreviation column.

COOPER ET AL.: RURAL U.S. OZONE TRENDS, 1990-2010

D22307

D22307

February). If a site had less than 50% data availability in any month in any season then that particular season was discarded. For each site and each season, data must be present for at least 18 out of 21 years during 1990–2010. Accordingly, any site with a reported trend in this analysis has greater than 85% data completeness over the 21 year study period, based on seasonal availability. All available daytime hourly measurements were used to compute the seasonal 5th, 50th and 95th ozone percentiles for each year. From this point forward, the 5th, 50th and 95th ozone percentiles for spring will be referred to as SpO305, SpO350 and SpO395, respectively. Likewise the summer percentiles will be referred to as SuO305, SuO350 and SuO395, and winter as WiO305, WiO350 and WiO395. The trend, or ozone rate of change (ppbv per year), over 1990– 2010 was calculated separately for the 5th, 50th and 95th ozone percentiles with a straight line fit through the data using the least squares method of simple linear regression. The trends were calculated with a reference year of 2000. The p value indicates the statistical significance of the linear relationship, determined by first calculating R, the correlation coefficient between ozone and time. The null hypothesis that R2 = 0 (no linear relationship) was tested using the standard F-statistic (ratio of the mean square regression to the mean square residual). If the probability p associated with the F statistic was small (p ≤ 0.05), the null hypothesis was rejected with a confidence level ≥ 95%. [16] To examine the impact of the Denver metropolitan region on ozone in the nearby Rocky Mountains, surface hourly ozone, CO and NO2 measurements from the Denver region were obtained from the U.S. EPA Air Quality System Data Mart website (http://www.epa.gov/ttn/airs/aqsdatamart/). Surface CO measurements from Niwot Ridge (40.05 N, 105.63 W, 3526 m a.s.l.) in the Rocky Mountains west of the Denver metropolitan region were measured by the NOAA Earth System Research Laboratory Carbon Cycle Group using weekly flask samples [Novelli et al., 2003]. [17] In addition to the surface ozone analysis, this study also provides an update to the free tropospheric springtime ozone trend above western North America that was originally calculated for 1984–2008 and reported by Cooper et al. [2010]. This update follows the same general methodology as Cooper et al. [2010] and adds all available ozone measurement from 2009 to 2011, made between 3.0 and 8.0 km above sea level (a.s.l.) in the region 25 –55 N, 130 –90 W. The only difference in methodology between the present study and Cooper et al. [2010] is that this analysis uses all available data between 3.0 and 8.0 km a.s.l. from all years during 1984–2011, whereas Cooper et al. [2010] used a particle dispersion model to identify and remove measurements made in the lowermost stratosphere that occasionally extended below 8.0 km a.s.l. Using the original 1984–2008 data set of Cooper et al. [2010] we compared the ozone rate of change with and without measurements made in the lowermost stratosphere. Due to the small percentage of data points in the lowermost stratosphere, the stratospheric measurements made little impact on the 50th ozone percentile values (seasonal medians were no more than 1 ppbv greater in the data set with stratospheric measurements) and made little difference in the rate of increase of the ozone 50th percentile. [18] Ozone data for 2009–2011 were collected by (1) electrochemical concentration cell ozonesondes, accuracy: 10% [Smit et al., 2007]; (2) an ozone lidar, accuracy: 5 of 24

D22307

COOPER ET AL.: RURAL U.S. OZONE TRENDS, 1990-2010

D22307

Figure 1. (a) Topographic map of the United States showing the locations of the 53 rural ozone monitoring sites used in this study. Sites discussed in the text are labeled, and are also listed in Table 1. (b) EDGARv4.1 global NOx emissions inventory for the year 2005 at 0.1 degree resolution. 5–25% [McDermid et al., 2002]; (3) MOZAIC commercial aircraft, accuracy: (2 ppbv + 2%) [Thouret et al., 1998]; and (4) a variety of research aircraft flights with instrument accuracies that are generally better than 5% or 5 ppbv. The ozone measurements were made at the following locations: 1) NOAA ozonesondes from the monitoring sites of Boulder, CO and Trinidad Head, CA, as well as the IONS-2010 California sites of Shasta, Pt. Reyes, San Nicolas Island and Joshua Tree National Park [Cooper et al., 2011]; 2) Environment Canada ozonesondes from the monitoring sites of Kelowna, Bratt’s Lake and Edmonton; 3) MOZAIC commercial aircraft profiles at Portland, OR and Calgary, Alberta; 4) lidar profiles from the NASA JPL Table Mountain facility, CA; 5) the NASA WB-57 aircraft profiles from Houston, TX during the spring 2011 MACPEX experiment; and 6) NOAA Global Monitoring Division aircraft profiles from Briggsdale, CO, Estevan Point, BC, Trinidad head, CA, Beaver Crossing, NE, West Brook, IA, and Southern Great Plains, OK. The total number of profiles from all platforms for 2009, 2010 and 2011 was 99, 186 and 100, respectively. [19] Several gridded data sets were used in the analysis: [20] 1) Column NO2 trends were calculated for several regions within midlatitude North America using monthly gridded data from the polar-orbiting GOME and SCIAMACHY sensors, produced and made freely available

by the Tropospheric Emission Monitoring Internet Service in The Netherlands (www.temis.nl). These products are based on the methodology of Boersma et al. [2004] and Richter et al. [2005]. Data were available for 1996–2011 at 0.25  0.25 horizontal resolution. [21] 2) The EDGARv4.1 global anthropogenic NOx emission inventory for 2005 at 0.1  0.1 resolution was provided by European Commission, Joint Research Centre (JRC)/ Netherlands Environmental Assessment Agency (PBL): Emission Database for Global Atmospheric Research (EDGAR), release version 4.1 http://edgar.jrc.ec.europa.eu, 2010. [22] 3) Gridded Population of the World, Version 3 (GPWv3) data at 2.5 min horizontal resolution for 1990 and 2010 were produced and made available by the Center for International Earth Science Information Network (CIESIN), Columbia University; and Centro Internacional de Agricultura Tropical (CIAT) (Gridded Population of the World, Version 3 (GPWv3). Palisades, NY: Socioeconomic Data and Applications Center (SEDAC), Columbia University, http://sedac. ciesin.columbia.edu/gpw, 2005, hereinafter referred to as CIESIN, online report, 2005). [23] 4) Interannual variability of NOx emissions from wildfires was calculated for several regions within midlatitude North America using monthly gridded data from the Global

6 of 24

COOPER ET AL.: RURAL U.S. OZONE TRENDS, 1990-2010

D22307

D22307

Figure 2. Change in population for various county and municipal regions across midlatitude North America from 1990 to 2010. Blue lines demarcate the regions used in the satellite column NO2 analysis shown in Figure 3. The vertical line through Texas is the meridian at 102 degrees West and separates the eastern U.S. from the western U.S., as described in the text. Numbers indicate the percent increase in human population in each region during 1990–2010. Light blue dots indicate the locations of the ozone monitors in Yellowstone (YS), Rocky Mountain (RM) and Grand Canyon (GC) National Parks. The population data were made available by CIESIN (online report, 2005). Fire Emissions Database (GFED3) [van der Werf et al., 2010]. Data were available for 1997–2010 at 0.5  0.5 horizontal resolution.

3. Results 3.1. Trends in Population, Emissions, and Free Tropospheric Baseline Ozone [24] Before presenting the ozone trend analysis, and to facilitate the discussion of the results, this section describes the changes in population, ozone precursor emissions, and free tropospheric baseline ozone across the United States, all of which affect surface ozone levels. During 1990–2010 the U.S. experienced a broad increase and geographical shift in its human population. Overall, the population of the contiguous U.S., southern Canada and northern Mexico increased from 287 to 352 million, a 22% increase (Figure 2) that also shifted southwards and westward, with the greatest rate of increase in the west at 37%. Even faster rates of increase occurred in several western sub-regions, with the population of central Colorado and southeast Wyoming increasing by 44%, and the population of northwest Arizona and Las Vegas increasing by 95% (Figure 2). [25] Despite the increase in population, ozone precursor emissions in the eastern and western U.S. have decreased. As mentioned in the Introduction, the EPA estimates that total U.S. anthropogenic emissions declined by 49%, 58% and 44% for NOx, CO and VOC, respectively, during 1990– 2010. However, the changes have not been smooth with most of the reductions occurring after 1996, and there can be strong seasonal variations, for example power plant NOx emissions in the eastern U.S. are much less during May–September in an effort to reduce ozone during the typical ozone season [Butler et al., 2011]. An independent and up-to-date assessment of regional ozone precursor changes can be gathered from the satellite tropospheric column NO2 retrievals made freely

available by the Tropospheric Emission Monitoring Internet Service in The Netherlands. This gridded product is ideal for the purposes of this paper because the retrievals are available (1996-present) for much of the study’s time period, and because ozone production in rural areas of the U.S. is typically NOx limited in spring and summer [McKeen et al., 1991; Fiore et al., 2009]. Even in the western USA where biogenic VOC emissions are weakest, ozone is much more sensitive to changes in anthropogenic emissions than biogenic [Fiore et al., 2011]. Therefore, for the remainder of this study, NO2 (or NOx) will be used as the primary indicator of ozone precursor emission distributions and trends. [26] Figure 3a shows the change in tropospheric column NO2 above the contiguous U.S., southern Canada and northern Mexico during spring and summer (column amounts in spring are greater than summer due to the longer NO2 lifetime in spring). Changes in column NO2 do not exactly match changes in emissions due to the change in NO2 lifetime that occurs with changes in ambient NO2 concentrations [Lamsal et al., 2011]. From 1996 to 2011 the total mass of NO2 above the U.S. decreased by 41% in spring and 33% in summer, in general agreement with the 49% emissions decrease estimated by the EPA for the longer period of 1990–2010. Column NO2 declines are also observed within the large regions of the northeastern, southeastern and western U.S. in spring and summer, all statistically significant based on linear regression analysis at the 95% confidence level (Figure 3a). Small western regions that cover the Colorado/Wyoming Front Range, northwest Arizona/Las Vegas, and Yellowstone National Park and surroundings (Figure 3b) also show declining column NO2. However, trends are only statistically significant (based on linear regression analysis at the 95% confidence level) for the northwest Arizona/Las Vegas region. [27] The column NO2 retrievals detect NO2 from all sources including wildfires. Due to the sporadic occurrence of wildfires, which may be missed by the polar orbiting satellites that

7 of 24

D22307

COOPER ET AL.: RURAL U.S. OZONE TRENDS, 1990-2010

D22307

Figure 3. (a) Trends in tropospheric column NO2, expressed as the total mass of NO2 detected per region by the GOME and SCIAMACHY instruments for spring (circles and thick lines) and summer (squares and thin lines) across the continental USA, and three sub-regions. The boundaries of these regions are shown in Figure 2. (b) Tropospheric column NO2 trends for three small areas in the western USA. Also shown is the interannual variability in the total quantity of NO2 emitted per year by biomass burning (GFED v3) in the same regions as shown in Figure 3a for (c) spring and (d) summer. measure column NO2, trends in wildfires across the U.S. may not be well represented by the monthly gridded column NO2 product. Because wildfires have the potential to produce large amounts of ozone in the U.S. [Jaffe and Wigder, 2012], monthly NO2 emissions from fires are also examined across the U.S., based on monthly estimates in the Global Fire Emissions Database (GFED). Figures 3c and 3d shows GFED biomass burning NO2 emissions for the same regions as

Figure 3a for spring and summer, 1997–2010. Emissions are greater in summer than spring, with both seasons showing large interannual variability. Linear regression analysis indicates that none of the examined regions has a statistically significant trend. The strongest fire season in the western USA from 1997 until 2010 was summer (JJA) 2007 which emitted 94 Gg NO2. According to the EDGAR v4.1 inventory for 2005, the anthropogenic NO2 emitted from the

8 of 24

D22307

COOPER ET AL.: RURAL U.S. OZONE TRENDS, 1990-2010

D22307

Figure 4. April–May ozone trends above western North America between 3 and 8 km above sea level for 1984–2011 and 1995–2011. All available data from ozonesondes, research aircraft and commercial aircraft were utilized, with no filtering by air mass origin. Numbers at the bottom of the graph indicate the sample size for each year. western USA in this season was 784 Gg. Therefore, the strongest fire season in the western U.S. from 1997 to 2010 was equal to 12% of the anthropogenic NO2 emissions. Given the lack of a trend in wildfire emissions across the U.S. and their relatively low rates compared to anthropogenic sources, U.S. fires are not expected to influence long-term ozone trends during 1997–2010. [28] An important consideration for rural surface ozone trends across the U.S. is the change in baseline ozone that has been observed during springtime. Cooper et al. [2010] used all available free tropospheric ozone measurements above midlatitude North America from April–May, 1995–2008 to show that median ozone values in the 3.0–8.0 km range were increasing at the rate of 0.63  0.34 ppbv yr 1. The rate of increase was stronger when the analysis was limited to air masses with a greater likelihood of transport from south and east Asia. Adequate data coverage was not available for a summertime analysis. Because free tropospheric ozone can be transported to the surface of the U.S. [Cooper et al., 2011; Lin et al., 2012a], it is likely that a trend in free tropospheric ozone could influence ozone trends at the surface. [29] Analysis of satellite column NO2 retrievals (not shown) indicates that NO2 emissions in east Asia have doubled from 1996 to 2011, increasing at a relative rate similar to CO2 emissions (as first reported by Richter et al. [2005]), while emissions in midlatitude North America and Europe have decreased. In response, surface ozone has decreased in the eastern U.S. (this study) and Europe [Logan et al., 2012], and increased in east Asia [Beig and Singh, 2007; Ding et al., 2008; Wang et al., 2009; Lin et al., 2010; Li et al., 2010; Parrish et al., 2012]. Apparently, the sphere of influence of the

increase of emissions and ozone within Asia extends downwind as far as the free troposphere above western North America [Cooper et al., 2010]. Globally there has been little change in total anthropogenic NOx emissions from 1990 to 2010, while emissions in China have doubled [Granier et al., 2011]. If global NOx emissions remain constant but continue to shift to east Asia it can by hypothesized that the ozone production efficiency within east Asia will decrease with increasing emissions [Wu et al., 2009], while the exported Asian pollution plumes will become diluted with air masses from regions with decreasing ozone production. As a result the high ozone rate of increase observed downwind of Asia (0.63  0.34 ppbv yr 1 for 1995–2008 [Cooper et al., 2010]) may not be sustained and could decrease. [30] Figure 4 is an update to the results of Cooper et al. [2010] showing the April–May free tropospheric ozone trend above western North America for 1995–2001 and 1984–2011. For 1995–2011 the median ozone rate of increase is 0.41  0.27 ppbv yr 1, 30% lower than, but still within the statistical uncertainty of the rate of increase determined for 1995–2008 by Cooper et al. [2010]. Several more years of data are required to determine if the ozone rate of increase has in fact declined. Regardless, from 1995 to 2011, free tropospheric ozone above western North America has increased significantly by 6.5 ppbv, and from 1984 to 2011 ozone increased by 14 ppbv. 3.2. Present-Day Rural Ozone Distribution [31] Figure 1 shows the locations of the 53 U.S. rural ozone monitoring sites used in this analysis. Twelve sites had data available for 1990–2010 in the western U.S., all

9 of 24

D22307

COOPER ET AL.: RURAL U.S. OZONE TRENDS, 1990-2010

D22307

Figure 5. Spring (MAM) daytime ozone mixing ratios at 50 rural sites shown as (a) 95th, (b) 50th, and (c) 5th percentiles for the years 2006–2010. Red and blue numbers indicate the average ozone values for each percentile in the western and eastern U.S., respectively. located in mountainous or elevated regions, with 10 sites over 1 km a.s.l. and 6 sites over 2 km a.s.l. While all sites are in rural areas, ozone monitors at Pinnacles National Monument (PN), Sequoia/Kings Canyon National Parks (SK), Joshua Tree National Park (JT) and Rocky Mountain National Park (RM) are less than 100 km from large urban areas (Figure 1b). Ozone monitor density is greater in the eastern U.S. with 41 available sites from northern Maine to northern Florida. Due to the lower terrain of the eastern U.S. the monitor elevations are much lower than in the west with only 5 sites having elevations greater than 1 km a.s.l. Eastern U.S. rural ozone monitors also have a closer proximity to large urban areas due to the greater population, which exceeded the

western population by a factor of 3 in 2010 (CIESIN, online report, 2005). In addition, 2005 anthropogenic NOx emissions in the eastern USA exceeded western emissions by a factor of 3.8 (EDGAR v4.1). These distributions are based on an east-west dividing meridian of 102 W and include the regions of southern Canada and northern Mexico, as shown in Figure 2. [32] Present-day (2006–2010) rural ozone mixing ratios are shown across the USA for spring and summer in Figures 5 and 6, respectively. Figure 5 and Table 2 show that the range of SpO395 values in the west are similar to those in the east with Joshua Tree National Park (JT) being the only site with notably high ozone due to its location downwind of the Los Angeles Basin. SpO350 and SpO305 values are

10 of 24

D22307

COOPER ET AL.: RURAL U.S. OZONE TRENDS, 1990-2010

D22307

Figure 6. Summer (JJA) daytime ozone mixing ratios at 53 rural sites shown as (a) 95th, (b) 50th and (c) 5th percentiles for the years 2006–2010. Red and blue numbers indicate the average ozone values for each percentile in the western and eastern U.S., respectively. typically greater in the west than in the east. Given the greater emissions of ozone precursors in the eastern U.S. it might be expected that greater ozone values would be found in the east than in the west. However, previous studies have shown that ozone in the free troposphere (>2 km a.s.l) is similar on the east and west coasts of the U.S. in springtime due to the rapid eastward transport and lack of stagnation events during this season [Cooper et al., 2005], and that free tropospheric ozone is much greater than at the surface [Cooper et al., 2011]. The elevated terrain of the western USA is much more likely to be influenced by descending free tropospheric air masses than the eastern USA, air masses that can contain enhanced ozone due to transport from upwind emission regions such as Asia or from the stratosphere [Brown-Steiner and Hess, 2011; Cooper et al., 2011;

Table 2. Average of the Ozone Percentiles at Each Site in the Western and Eastern U.S. for Spring and Summer During 1990–1994 and 2006–2010 Spring

Summer

1990–1994 2006–2010 1990–1994 2006–2010 Western U.S. 95th % Western U.S. 50th % Western U.S. 5th %

64 50 38

66 53 40

70 53 36

69 54 39

Eastern U.S. 95th % Eastern U.S. 50th % Eastern U.S. 5th %

72 47 25

66 47 28

80 52 29

66 46 28

11 of 24

D22307

COOPER ET AL.: RURAL U.S. OZONE TRENDS, 1990-2010

D22307

Figure 7. Spring (MAM or AM) daytime ozone trends (1990–2010) at 52 rural sites for the (a) 95th, (b) 50th and (c) 5th percentiles. Indicated are sites with statistically significant (red dots) and insignificant (pink dots) positive trends, and sites with statistically significant (dark blue dots) and insignificant (light blue) negative trends. Vectors indicate the ozone rate of change in ppbv yr-1 as shown in Figure 7a. Zhang et al., 2008; Lin et al., 2012a, 2012b]. While U.S. springtime surface ozone mixing ratios are sensitive to local emissions [Reidmiller et al., 2009; Lin et al., 2012a], terrain and the impact of baseline air masses appear to have a strong influence on the nationwide rural ozone distribution. These impacts are most apparent for the greater SpO305 and SpO350 values in the west, while for SpO395, ozone production in the eastern USA appears to match the effects of enhanced baseline ozone on the western USA. [33] During summer, eastward transport is weaker than in spring and emissions from upwind regions have less of an impact on U.S. surface ozone relative to domestic emissions

[Reidmiller et al., 2009]; ozone transport from the stratosphere is also less in summer than in spring [Stohl et al., 2003; James et al., 2003a, 2003b]. The diminished role of long range transport in summer might be expected to result in a smaller contrast between western and eastern ozone. However, average western surface ozone is still greater than eastern ozone for the 5th, 50th and 95th percentiles, and the difference between east and west in summer is actually greater than in spring (based on the 50th percentile; Table 2) indicating that terrain is still an important factor in summer for the nationwide ozone distribution. This comparison includes all 12 sites in the western USA, however, two sites,

12 of 24

D22307

COOPER ET AL.: RURAL U.S. OZONE TRENDS, 1990-2010

Joshua Tree and Sequoia/Kings Canyon are heavily influenced by anthropogenic emissions in southern California. When these two sites are omitted, the western summertime 95th, 50th and 5th ozone percentiles decrease to 65, 51 and 37 ppbv, respectively. This omission makes the western SuO395 1 ppbv less than the eastern USA value, but SuO350 and SuO305 are still greater than the eastern USA values by 5 and 9 ppbv, respectively. 3.3. 1990–2010 Rural Ozone Trends Across the U.S. [34] The 1990–2010 rural ozone trends are shown in Figures 7 and 8 and statistically significant trends are summarized for the eastern and western U.S. in Table 3. In spring 41% of the eastern U.S. sites have statistically significant ozone decreases for the 95th percentile with no sites showing a significant increase. In contrast, no site in the western U.S. has a significant decrease, while 25% have a significant increase. At the 50th percentile there is little overall change in the eastern U.S. while 50% of western sites have a significant increase, and no western site has a significant decrease. At the 5th percentile increases outweigh decreases in both the east and west. Two sites in the west, Lassen National Volcanic Monument (LV) and Yellowstone National Park (YS) have significant increases for all three percentiles. The most likely explanation for the contrast in eastern and western ozone trends appears to be emission reductions causing decreases of extreme ozone events in the eastern U.S. [Hogrefe et al., 2011], while increased baseline ozone is likely causing ozone to rise at many sites in the western U.S. [Parrish et al., 2004b, 2009, 2012; Cooper et al., 2010]. We hypothesize that the increase in baseline ozone may also contribute to the rise in the eastern USA 5th percentile. However, decreasing NOx emissions may also contribute to the rise in the eastern USA 5th percentile as urban areas will experience less ozone titration, and under weak photochemical conditions the result is the export of plumes to rural areas with greater ozone concentrations. [35] In summer the impact of domestic emission controls appears to have strongly affected the eastern U.S. across the entire ozone range, with 83%, 66% and 20% of these sites experiencing statistically significant ozone decreases in the 95th, 50th and 5th percentiles, respectively. No eastern site has a significant increase, in any percentile. Despite emission reductions in the west only 17% and 8% of sites have statistically significant ozone decrease in the 95th and 50th percentiles, respectively, all occurring in the polluted region of central California (Pinnacles National Monument (PN) and Sequoia/Kings Canyon National Parks (SK)). Surprisingly, Joshua Tree National Park (JT) has no significant downward trend in summer with the 5th percentile actually increasing significantly, despite improvements in ozone air quality in the Los Angeles Basin (http://www.arb.ca.gov/ adam/index.html). Overall, western U.S. sites show more of an increase than a decrease in summer (Table 3). [36] Table 2 summarizes very interesting seasonal changes in the timing of peak ozone values. In the western U.S., median ozone in summer was 3 ppbv greater than in spring during 1990–1994. But for the present-day (2006–2010) summertime ozone is only 1 ppbv greater than spring. This decreased seasonal contrast is primarily caused by greater ozone increases in spring than summer. Spring/summer differences for the 5th and 95th percentiles have also diminished

D22307

over the past 20 years. The contrast is even stronger for the eastern U.S. In 1990–1994 eastern U.S. median ozone in summer exceeded spring by 5 ppbv, but today summer ozone is actually 1 ppbv less than spring, entirely due to decreasing ozone in summer. Spring/summer differences for the 5th and 95th percentiles have also disappeared. While ozone across the USA used to peak in summer (1990–1994), today there is a very broad spring/summer maximum. Seasonal ozone shifts have been observed at other long-term monitoring sites in the U.S. since the 1980s [Bloomer et al., 2010; Lefohn et al., 2010]. [37] While this paper focuses on spring and summer, for completeness, winter ozone trends are reported in Figure 9 and summarized in Table 3. This comparison shows how ozone trends are different in winter when domestic ozone production is at a seasonal minimum. Significant ozone decreases only occur at 9% of sites in the east or west and only for the 95th percentile. In contrast, ozone increases are much more common. In the west roughly a quarter to a third of all sites has significant increases across the range of ozone values. In the east just one site has a significant increase in WiO395 while 44% and 58% of sites have significant increases in WiO350 and WiO305, respectively. 3.4. Impact of Temperature Trends [38] As discussed in Section 3.1 changes in emissions and baseline ozone can impact surface ozone trends across the USA, but another potential impact is regional climate change. Of all the meteorological variables affecting ozone concentrations in polluted regions, temperature is the most important [Jacob and Winner, 2009]. Surface ozone and temperature are highly correlated in the eastern United States with observed ozone increases of 2–3 ppbv per  C [Bloomer et al., 2009]. In California maximum daily ozone increases with maximum daily temperature with a slope of 2–4 ppbv per  C averaged across the 1990s and 2000s, although the slope becomes very small or even negative at temperatures above 39 C [Steiner et al., 2010]. The increase in ozone with temperature is linked to 1) the temperature-dependent lifetime of peroxyacetylnitrate (PAN), 2) the temperature dependent biogenic emission of isoprene, and 3) air mass stagnation and sunny skies that accompany high surface temperatures [Jacob and Winner, 2009]. Many modeling studies have calculated the response of ozone to a warmer future climate due to rising concentrations of anthropogenically produced greenhouse gases. The robust findings indicate increased surface ozone in polluted regions, but decreased surface ozone in remote regions due to higher water vapor concentrations [Jacob and Winner, 2009]. Murazaki and Hess [2006] examined the potential impact of climate change on U.S. surface ozone from the period 1990–2000 to 2090–2100. They calculated that a 2–4 C increase in surface temperature would result in surface ozone enhancements of up to 6 ppbv in polluted regions. Baseline ozone in the free troposphere above the U.S. west coast would decrease by 2–6%, due to the shorter lifetime of ozone in a warmer climate, resulting in a decreased impact at the surface of the western U.S., up to 2 ppbv. Wu et al. [2008] found similar results for the shorter time period of 2000–2050 (U.S. summertime temperature increases as much as 2–3 C), except they found little change in ozone in the mid-troposphere of northern midlatitudes. Wu et al. [2008] also found that in April, shifting synoptic scale transport patterns could result in

13 of 24

D22307

COOPER ET AL.: RURAL U.S. OZONE TRENDS, 1990-2010

D22307

Figure 8. As in Figure 7 but for 53 sites in summer. increased baseline ozone impacting some regions of the western U.S. [39] Global average surface temperature increased by 0.074 C  0.18 C per decade when estimated by a linear trend for the 100 year period,1906–2005 [Intergovernmental Panel on Climate Change (IPCC), 2007], while the rate of increase since the late 1970s is 0.15 C–0.20 C per decade [Hansen et al., 2010]. As global temperatures have increased since the 19th century so too has the global tropospheric ozone burden, primarily due to rising anthropogenic emissions of ozone precursors [Lamarque et al., 2005]. Based on the model studies of ozone response to future climate change, one might assume that past ozone changes have also been influenced by climate change since the 19th century. A recent intercomparison of 10 atmospheric chemistry models

run with year 2000 emissions but with 2000s and 1850s climate shows a range of responses of tropospheric ozone to the observed temperature increase [Stevenson et al., 2012]. Six out of 10 models indicate ozone decreases at the surface of the northern hemisphere midlatitudes due to observed climate change, but the decreases are small (