Health benefits from ozone reduction in the United States in Chicago and Detroit. Estimates for 2005 suggested major reductions in mortality would have occurred in the central U.S. cities of Chicago, Dallas, and Houston, whereas smaller impacts were estimated for Atlanta, Seattle, and Washington, DC. Discussion The goal of our analysis was to estimate the health benefits that would be achieved by meeting the current ozone NAAQS, and additional benefits that might accrue from meeting more stringent standards. The HIA approach is based on research linking ozone exposures to adverse health effects, such as studies that indicate associations of a 10-ppb increase in daily ozone with a 0.87% increase in total mortality (Bell et al. 2004) and a 4.0% increase in respiratory mortality (Jerrett et al. 2009). Our results quantify the immediate health impacts of achieving ozone standards, instead of projecting benefits into the future. These results are of particular public health significance in light of the executive deci- sion to withdraw proposed ozone NAAQS in the range of 70–60 ppb (Office of the Press Secretary 2011). Hubbell et al. (2005) estimated that an 80-ppb ozone regulatory standard would result in 840 prevented mortalities for the United States based on 2000–2002 air quality data. Using more recent baseline incidence rates, population sizes, and estimated C-R func- tions, we estimated that a further reduction of 5 ppb from the current standard would result in nearly twice the number (approximately 1,900/year) of avoided all-cause mortalities. A 2008 U.S. EPA regulatory impact assess- ment for a proposed 65-ppb ozone standard suggested smaller overall health impact esti- mates than ours based on a 60-ppb standard (U.S. EPA 2008b). For example, the U.S. EPA estimated 450 avoided premature mor- talities and 1.1 million avoided school-loss days based on a 65-ppb standard, whereas we estimated 1,700 avoided premature mortali- ties and 3.7 million avoided school-loss days under a 60-ppb standard. Geographic patterns of ozone concentrations estimated by Fann et al. (2011) were similar to patterns estimated for the present study, including areas of high ozone in southern California, the industrial Midwest, and portions of the South, including St. Louis and Atlanta. In contrast with the present study, the U.S. EPA used CMAQ (Community Multiscale Air Quality) program modeling to project ozone levels to the year 2020, the expected date of full NAAQS compliance, assuming a series of increasingly stringent emissions reductions from future rules (U.S. EPA 2008b). In addition, the U.S. EPA assess- ment excluded the San Joaquin and South Coast Air Basins, significant population areas, because these regions are not expected to meet the 2020 projection (U.S. EPA 2008b). We used more recent data to estimate immediate health effects that would result from achiev- ing standards, without assuming changes according to a regulatory compliance timeline. We also elected to use air monitor data to avoid the uncertainty associated with air qual- ity modeling assumptions such as emissions inventories and chemical mixing (U.S. EPA 2008b). However, the use of monitor data has some disadvantages. For example, AQS distributions are heterogeneous and several regions are sparsely monitored, notably the Rocky Mountains [see Supplemental Material, Figure S1 (http://dx.doi.org/10.1289/ ehp.1104851)]. Interpolated ozone concen- trations in these areas have greater variability than locations proximate to other monitors, which may reduce the accuracy of estimated effects in these locations. The use of 3 years of air quality data fol- lows the U.S. EPA ozone NAAQS (U.S. EPA 2008c), while highlighting the tempo- ral variability observed in the annual ozone exceedance. For example, we estimated sub- stantial reductions in ozone levels for Houston in 2005 and 2006, but the reduction would have been much smaller in 2007. Atypical temperature and precipitation during the sam- pling season may partially explain this, high- lighting the regional impact of weather on air pollutant concentrations. Also, we estimated reductions in premature mortality per MSA using a national risk function estimate. A pre- ferred approach would be to apply appro- priate city-specific estimates (Hubbell et al. 2009), but this would have been difficult to do given the data and analytical platform used for our analysis. Avoided respiratory-related ED visits were estimated using C-R functions with varying population age ranges. A potential concern is that pooling estimates where some age groups are more heavily represented than others may generate biased estimates; however, pooling improves geographic coverage and provides more robust results at the national level. The MSA assessment might be improved by using population weighted results because it is dif- ficult to separate effects due to the number of residents from effects due to ozone reduction. However, because our goal was to demonstrate overall health benefits, absolute numbers of prevented outcomes are a more relevant metric for assessing public health impacts compared to normalized estimates. Based on our analysis, a 60-ppb ozone standard would result in the greatest numbers of avoided health effects. However, attain- ment of a 60-ppb standard, which is close to the non-anthropogenic background level of 40 ppb, would likely present significant challenges for regulatory interventions, and a Environmental Health Perspectives • volume 120 | number 10 | October 2012 less stringent alternative NAAQS standard of 70 ppb would also yield significant improve- ments to human health. On the other hand, implementation of tighter emissions regula- tion is important because ambient ozone levels are predicted to rise with changes in global climate (Chang et al. 2010) and preliminary studies suggest that the adverse health effects of ozone may be increased in warmer tem- peratures (Foster et al. 2000). Reductions in ambient ozone and related health benefits would also result in substantial economic ben- efits that were not considered in our analysis. Though the statutory requirements of the CAA preclude consideration of economic fac- tors in setting the NAAQS, attainment costs are considered during the implementation phase of air pollution control policy decisions. Although there is substantial consistency across epidemiological and clinical studies regarding the evidence of adverse health con- sequences associated with exposure to ambi- ent ozone, uncertainty remains regarding the dose–response relationship, especially at low ozone concentrations. It is not clear from the published literature whether the biological processes driving short-term and long-term induced mortality are the same. Therefore, it is currently unknown whether premature deaths observed in long-term cohort studies are clinically unique from those observed in short-term time-series studies. This dispar- ity may lead to a potential overestimation of ozone-linked mortality. Additionally, because the U.S. EPA’s current science assessment (U.S. EPA 2008c) identifies no threshold for the relationship between ozone exposure and premature mortality, our assessment underestimates the overall avoided premature mortality in regions where ozone concentra- tions fall below the assumed background level of 40 ppb. By restricting our analysis to the ozone season, 1 May–30 September, we may have underestimated the magnitude of ozone-related health effects in regions with longer peak ozone seasons, including southern California and Texas. However, national estimates that were not restricted to the ozone season were compa- rable to those reported. Geographic underesti- mation may also have occurred in rural regions that lack sufficient monitors but are exposed to biogenic volatile organic compounds from ter- restrial plant sources (Fiore et al. 2005). There were also statistical and methodo- logical limitations related to the software platform used in our analysis. Each step of risk assessment presents uncertainty, and it is difficult or impossible to comprehensively quantify the propagated effect of uncertainty introduced at multiple points in the analysis (Briggs et al. 2009; Levy et al. 2001). Our model did not account for uncertainty related to the ozone interpolation, the incidence rate 1409