General Features Of Photochemical Smog Diurnal and Seasonal Cycle

Ozone and other secondary reaction products show a pronounced diurnal cycle with peak concentrations typically occurring in late afternoon. The diurnal cycle of 03 shows a sharp contrast with the diurnal cycle of primary species, including NOx, HC, and CO (see Fig. 1). The primary species typically have peak concentrations in early morning and much lower concentrations during the daytime as concentrations are diluted through the process convection-driven vertical mixing. Because production of 03 requires sunlight, peak concentrations often occur at the time of maximum vertical mixing and often coincide with diurnal minima in precursor concentrations. The diurnal cycle of 03 is also influenced by nighttime removal of 03 near the ground (through surface deposition or through reaction with directly emitted NO). Especially in urban areas, 03 concentrations near the surface are often very low at night. The characteristic increase in 03 during the morning hours (6 to 10 a.m.) is usually driven by convective mixing that breaks up the nighttime inversion at the

*The family consisting of hydrocarbons and oxygenates such as formaldehyde, hcho, is properly referred to by acronyms such as volatile organic compounds (voc) or reactive organic gases (rog). In this chapter they will be referred to collectively as hydrocarbons (hc).

Diurnal Ozone Concentraties Cities

surface and mixes down air (from 100 to 1000 m above the surface) with higher 03. The subsequent rise in 03 after 10 A.M. is often associated with photochemical production.

The city of Los Angeles is especially susceptible to high-ozone events because it frequently sees a combination of high sunlight, warm temperatures, and a low-level thermal inversion (typically 500 to 1000 m above the surface) during the daytime. In most other cities the conditions that favor ozone formation (sunlight and warm temperatures) coincide with vigorous vertical mixing (up to 2000 m), which has a moderating effect on ozone concentrations. Thermal inversions, which trap pollutants near the ground, are more commonly associated with cold temperatures and often with fog. These conditions would produce high concentrations of primary pollutants but not ozone.

Unless stated otherwise, the discussion of ozone concentrations below refers to the diurnal peak or near-peak concentrations that occur during the afternoon.

Concentrations and Regional Transport

The global background concentration of 03 near the surface is 20 to 40 ppb parts per billion (ppb), although these values probably represent an increase in comparison with preindustrial concentrations. There have also been episodes in which high concentrations of 03 originating in the upper troposphere, ultimately of stratospheric origin, may have been transported to the surface.

During air pollution events in the United States and Europe, peak 03 frequently exceeds 125 ppb, which is the current government health standard in the United States.* In Los Angeles during the 1970s and 1980s air quality violations (i.e., 03 < 125 ppb) were reported on approximately 180 days per year. In the 1990s the frequency of violation has been lowered to 90 days per year. Most other major cities in the United States record violations on 5 to 10 days per year. In Europe, ozone exceeds 125 ppb on just a few days per year, while in Mexico City at present, ozone exceeds 125 ppb on 200 days per year. Concentrations above 200 ppb are found only during the most severe events, and concentrations as high as 490 ppb have been observed in Los Angeles and in Mexico City.

In addition, 80 to 100 ppb ozone is frequently observed in rural areas of the eastern United States and Europe during regional events. In these events, air with ozone concentrations above 80 ppb frequently extends over a 1000 x 1000 km region and extends vertically to 1000 to 2000 m above the surface (e.g., Clarke and Ching, 1983). These events are often associated with stagnant high pressure systems in which air may be trapped under a subsidence inversion at —2000 m. An example in eastern North America is shown in Figure 2. In the example, ozone above 120 ppb

* As of 1997, a 1 -h average concentration in excess of 125 ppb constitutes an air quality violation in the U.S. metropolitan areas that record violations of the 1 -h standard on more than 3 days over a 3-year period are held in violation of clean air laws and are asked to submit a plan for pollution reduction. Since 1977, most U.S. cities have been continually in violation. It was recently proposed that the 1-h air quality standard be replaced with a standard based on 8-hour average concentrations, in which an 8-h average concentration in excess of 85 ppb is counted as an air quality violation.


OZONE (ppb)

June 15 1988

Figure 2 Peak ozone concentrations in the eastern United States during a severe air pollution event (June 15, 1988) based on surface observations at 350 EPA monitoring sites. The shadings represent values of 30 to 60ppb (lightest shading) to 180 to 210ppb (darkest shading) with 30-ppb intervals in between. Values reported for Canada and the Atlantic Ocean are inaccurate since no observations were available for these locations. First printed in Sillman (1993).

Figure 2 Peak ozone concentrations in the eastern United States during a severe air pollution event (June 15, 1988) based on surface observations at 350 EPA monitoring sites. The shadings represent values of 30 to 60ppb (lightest shading) to 180 to 210ppb (darkest shading) with 30-ppb intervals in between. Values reported for Canada and the Atlantic Ocean are inaccurate since no observations were available for these locations. First printed in Sillman (1993).

was found in many metropolitan areas, especially in the corridor extending from Washington to New York and Boston. However, ozone above 90 ppb covered a much larger area extending from Kentucky to Maine. Although ozone concentrations above 120 ppb were generally associated with plumes from specific urban areas (or from coal-fired power plants), concentrations of 90 to 100 ppb were found at rural sites throughout the region. In addition, unusually high ozone (> 200 ppb) was found in Acadia National Park in Maine. The high ozone in Maine is most likely due to transport from Boston (300 km distant) and the New York area (700 km distant).

Environmental and Health Impacts

The impact of ozone and acid aerosols on human health has been the subject of intense scrutiny. Ozone and aerosols have been associated with a variety of lung ailments. Short-term symptoms (including lung inflammation, asthmatic responses, and measured impairment of lung functions) have been found in experiments in response to ozone concentrations as low as 120 ppb. High-ozone events have been correlated with increased admissions to hospitals for respiratory diseases and with increased mortality rates. For a summary of findings, see Lippman (1993) and Bascomb et al. (1996).

In addition, ozone concentrations of 80ppb have been found to cause damage both to forests and to agricultural crops. Crop damage from ozone in the United States has been estimated to cause monetary losses of $ 1 to 2 billion per year. For a summary of findings, see U.S. Congress (1989) and National Research Council (NRC, 1991).

Dependence on Temperature

As stated above, ozone in polluted regions shows a strong dependence on temperature. This dependence on temperature is important as a basis for understanding variations in ozone concentrations from year to year or between cities. As shown in Figure 3, elevated ozone is always associated with temperatures in excess of 20° C and is often with temperatures above 30°C. In the eastern United States and Europe, year-to-year variations in ozone concentrations are often the result of variations in temperature and cloud cover, rather than in changes in emission of pollutants.

The reason for the dependence on temperature is due largely to the chemistry of ozone formation. Cardelino and Chameides, (1990) and Sillman and Samson (1995) found that the temperature dependence was associated with the temperature-dependent decomposition rate of PAN. PAN becomes longer lived at lower temperatures, and formation of PAN results in the removal of NO, , hydrocarbons, and odd hydrogen radicals (described below), all of which suppress ozone formation. PAN, also a component of photochemical smog, tends to reach maximum values at intermediate temperatures (5 to 10°C). Jacob et al. (1993) proposed that ozone correlates with temperature partly because the meteorological conditions that favor ozone formation (high solar radiation and light winds) tend to be associated with warm temperatures. In addition, the emission rate of biogenic hydrocarbons (a major ozone precursor, discussed below) increase sharply with increasing temperature. Ozone is affected by temperature only in polluted regions. Temperature apparently has little impact on ozone production at the global scale (Sillman and Samson, 1995).

300 250 200 150 100 50 0

New York

New York

300 250 200 150 100 50 0

: : : :


; ; : ; -

- >

! J

: - F




280 285 290 295 300 305 310 315 Max. Temperature (K)

280 285 290 295 300 305 310 315 Max. Temperature (K)

Figure 3 Diumal peak 03 (ppb) vs. maximum surface temperature observed in the New York-New Jersey-Connecticut metropolitan area for April 1 through September 30, 1988. From Sillman and Samson, 1995.


Role of Biogenic Hydrocarbons

The main precursors of photochemical smog, NO, and hydrocarbons, are emitted into the atmosphere by a variety of human activities—transport (chiefly automobiles), coal-fired industry (especially electric power plants), and biomass burning. However, significant amounts of hydrocarbons occur naturally and are emitted by vegetation, primarily from trees. The most important of these biogenic hydrocarbons are isoprenes (C5H8), emitted by oaks and other deciduous trees, and a- and /?-pinenes (Ci0H16), which are emitted from conifers. These species react chemically in the same way as anthropogenic hydrocarbons and can function as precursors to photochemical smog. In the United States it is estimated that emission of biogenic hydrocarbons equals or exceeds emission of anthropogenic hydrocarbons (Geron et al., 1994). Even in urban areas biogenic hydrocarbons can account for a significant fraction of total hydrocarbon emissions and can have a large impact on the formation of smog (Chameides et al., 1988). The impact of isoprene is especially large because it reacts rapidly, with a chemical lifetime of one hour or less. Consequently even small amounts of isoprene (0.5 ppb) can have a large impact on ozone.

It should be emphasized that naturally occurring hydrocarbons will not lead to the formation of photochemical smog in the absence of human activities because smog formation requires NO, in addition to hydrocarbons. Although some NO, is emitted naturally through biological activity, naturally occurring NO, emissions are too small to allow significant formation of 03 and other components of smog. Biogenic NO, is estimated to be 7% of total NO, emissions in the United States (Williams et al., 1992) and most of this is associated with agriculture (especially with the use of nitrate fertilizer).

Ozone Production Efficiency

The ozone production efficiency represents the rate of production of ozone divided by the loss rate for NO, [P(03)/L(N0J]. Liu et al. (1987) first introduced the concept of ozone production efficiency and used it as a basis for estimating global production of ozone as a function of estimated NO, emissions. A central feature of the ozone production efficiency is the tendency toward lower values in more polluted environments. Recent estimates suggest that ozone production efficiency is 10 to 30 in the remote troposphere but just 3 to 5 in urban areas.


The relation between ozone, NOv and hydrocarbons can be illustrated by an isopleth plot (Fig. 4), which shows instantaneous rates of ozone production as a function of NO, and hydrocarbon concentrations. It can be seen that ozone production is a highly nonlinear process, especially with regard to NO,. Ozone production as a function of NO^ shows well-defined local maxima, usually at a specific HC/NO, ratio. This region of maximum ozone (the "ridge line") can be thought of as a divide

Mexico City Voc Nox

Figure 4 Isopleths giving net rate of ozone production (ppb per hour, daytime average, solid line) as a function of ROG (ppbC) and NOx (ppb). The dashed lines and arrows show the calculated evolution of ROG and NOx concentrations in a series of air parcels over an 8-h period (9 a.m. to 5p.m ), each with initial ROG/NOA = 6 and speciation typical of urban centers in the United States, based on calculations shown in Milford et al. (1994).

10 20 50 100 200

Hydrocarbons (ppbC)

Figure 4 Isopleths giving net rate of ozone production (ppb per hour, daytime average, solid line) as a function of ROG (ppbC) and NOx (ppb). The dashed lines and arrows show the calculated evolution of ROG and NOx concentrations in a series of air parcels over an 8-h period (9 a.m. to 5p.m ), each with initial ROG/NOA = 6 and speciation typical of urban centers in the United States, based on calculations shown in Milford et al. (1994).

between two regimes with different photochemical behavior. Above the ridge line (with low HC/NOx ratios), ozone production rates increase with increasing HC but decrease with increasing NOx (hydrocarbon-limited regime). Below the ridge line (with high HC/NOx ratios) ozone production rates increase with increasing NOx and will be largely unaffected by changes in hydrocarbons (NOx-limited regime). The existence of these two regimes has an enormous impact on public policy because it affects the choice of control strategies for reducing high ozone levels. If ozone production is dominated by NOx-limited chemistry, then reductions in NOx emissions would be necessary to reduce ozone concentrations. If production is dominated by HC-limited chemistry, then reductions in hydrocarbons would be needed. There is also a complex relation between NOx, HC, and particulates, which also affects policy choices (Meng et al., 1997).

An important feature of HC-NOx chemistry is the tendency for polluted air to evolve toward the NOx-limited regime as the air mass ages and moves downwind. Air is most likely to show HC-limited chemistry when it is close to emission sources, especially in large cities. As the air mass ages, the HC/NOx ratio increases and the chemistry shifts to the NOx-limited regime. This shift from HC-limited to N(Delimited chemistry as polluted air ages is illustrated by the air parcel trajectories in Figure 4.

In terms of photochemical mechanisms, the split into NOx-limited and HC-limited regimes is closely related to the chemistry of odd hydrogen, defined as the sum of OH, H02, and R02 radicals (where R represents a hydrocarbon chain). The sequence of reactions that lead to photochemical smog (including production of both


ozone and acid aerosols) is usually initiated by reactions that involve the OH radical, and availability of OH (which is produced from reactions involving sunlight, 03 and HzO) often controls the rate of ozone production. The split into NOt- and HC-limited regimes is determined by the loss mechanism for odd hydrogen (Sillman, 1995; Sillman et al., 1990; Kleinman 1991, 1994; summarized in Sillman, 1999). The reaction sequences are shown in the postscript to this chapter.

Kleinman (1991, 1994) has shown that the split between NOA- and HC-limited regimes can be explained simply in terms of the supply of NO^ relative to the source of odd hydrogen. NOA-limited chemistry occurs when the supply of odd hydrogen exceeds the supply of NO, , while HC-limited chemistry (also referred to as light-limited chemistry) occurs when the supply of NOA exceeds that of odd hydrogen. This explanation is useful because it provides a conceptual basis for understanding how HC-NOA chemistry varies from location to location and from event to event. For example, HC-limited chemistry is most likely to occur in large cities and during severe events, when the supply of NOA is largest, and also during periods of low sunlight, which limits the source of odd hydrogen. NO (.-sensitive chemistry is more likely in smaller cities and during more moderate events (i.e., with lower NO^) and in far downwind locations (Milford et al., 1994). These trends in HC-NOA chemistry are all consistent with Kleinman's description.

The following is a summary of factors that affect the variation between HC-limited and NOv-limited chemistry.

As illustrated in Figure 4, HC-sensitive chemistry is associated with low HC/NOA ratios and NO,-sensitive chemistry is associated with high HC/NOA. The importance of HC/NOA ratios was identified in the early research into the causes of photochemical smog in the 1950s (Haagen-Smit and Fox, 1954).

For many years, the U.S. Environmental Protection Agency (EPA) used a rule of thumb that HC-NOA chemistry could be deduced based on the HC/NOt ratio at 6 to 9 a.m., where a ratio of 10* or less was presumed to correspond with HC-sensitive chemistry and a ratio of 20 or more would correspond with NOA-sensitive chemistry. This approach has been discredited (e.g., NRC, 1991) because it failed to take into account many of the other factors that affect HC NO, chemistry, described below. The chemical impact of hydrocarbons also depends on the reactivity of the hydrocarbon species, so that NOA-sensitive chemistry is more likely when HC reactivity is higher. It is often useful to think of hydrocarbons and HC/NOt ratios in terms of reactivity-weighted sums rather than just total concentration.

*Hydrocarbon concentrations are often expressed in parts per billion carbon (ppbC), which represents a sum of species concentrations in ppb weighted by the number of carbon atoms contained. HC/NO^ ratios are expressed in pppC/ppb.

Biogenic Hydrocarbons

The inclusion of biogenic hydrocarbons in analyses of photochemical smog often has a large impact on HC NOr chemistry. Biogenic hydrocarbons cause an increase in HC/NOr ratios and therefore cause a shift toward NOr-sensitive chemistry. The impact of biogenic hydrocarbons is often overlooked because (i) biogenic hydrocarbons are extremely reactive, and consequently have an impact on chemistry out of proportion to their ambient concentrations; and (ii) biogenic emissions are zero at night and low during the morning hours, and are therefore underrepresented in the traditional morning HC/NOr ratio.

Historically, the role of biogenic hydrocarbons on urban ozone formation was not recognized until 1988. More recently, it has been shown that emission estimates used by the U.S. EPA underestimated biogenic emissions by factors of 3 or more (Geron et al., 1994). Unpublished results from model calculations suggest that use of the higher biogenic emission estimates would cause a shift from primarily HC-sensitive chemistry to primarily NOr-sensitive chemistry in many cities of the eastern United States. The impact of biogenic hydrocarbons is smaller in Europe (Simpson, 1995).

Geographical Variation

The pattern of geographical variation in HC-NOr chemistry is largely associated with the photochemical aging process. As stated above, fresh emissions are often in an HC-limited state but evolve toward NOr-limited chemistry as the air mass ages. In addition, the total accumulation of ozone in a fully aged air mass appears to be controlled entirely by NOr rather than by HC. In other words, a reduction in hydrocarbons in an HC-limited region has the effect of deferring ozone production until an air mass moves downwind and disperses, but may have little effect on the number of ozone molecules that is produced once the chemistry has run to completion.

The contrast between HC-limited chemistry in an urban center and NOr-limited chemistry in downwind suburban regions has been dominated most extensively for Los Angeles (Milford et al., 1989, 1994). Extensive measurements and model calculations have supported the view that downtown Los Angeles has HC-limited chemistry. By contrast, ozone in rural locations in the eastern United States (representing photochemically aged air) is usually sensitive to NOr rather than HC, although there are exceptions. The highest ozone concentrations typically occur in suburban locations approximately 6h downwind of major urban centers. These locations represent an intermediate situation between the HC-limited chemistry of urban centers and NOr-limited chemistry in far downwind locations. It is often uncertain whether peak ozone concentrations are associated with HC-limited or NOr-limited chemistry, and predictions (derived from model calculations) are frequently dependent on model assumptions, e.g., about emission rates, winds, and vertical mixing.

It should be emphasized that predictions for HC-NOx sensitivity for individual locations are all highly uncertain at this time. HC-NOr predictions are often based on model calculations with little supporting evidence from ambient measurements.

The most detailed analyses, including both model calculations and analyses of ambient measurements, have been done for rural sites in the eastern United States (e.g., Buhr et al., 1995; Roselle and Schere, 1995; Jacob et al., 1995; Sillman, 1995; Trainer et al., 1993) and for Los Angeles (e.g., Milford et al., 1989; Jacobson et al., 1996) Other area evaluations have been done for Atlanta (Sillman et al., 1995), Mexico City (Sosa et al., 2000), and Europe (Simpson, 1995). For a more complete summary, see Sillman, (1999) and NRC (1991).

Evaluation of Ozone Production through Measurements

Two types of ambient measurements are especially important for evaluating the impact of photochemistry as a source of 03. One is the correlation between 03 and CO (e.g., Parrish et al., 1993). Because CO is primarily a product of human activities (either industry or biomass burning), a positive correlation between these species is interpreted as a signal for photochemical smog, especially in the remote troposphere. A second measurement is the correlation between 03 and the sum of total reactive nitrogen (NO^, including N0V, PAN, HN03, and other organic nitrates) and between 03 and the sum of N0V reaction products (NO^-NO^, or NOz). Because 03 and NOz are both produced by similar photochemical processes, there is a strong correlation between these species in polluted environments at times of photochemical activity.

The correlation between 03 and NOz is also interpreted as a measure of ozone production efficiency, defined as the ratio of net production of ozone to the loss rate for NOx [P(03)/L(N0J]. The slope between 03 and NOz is determined partly by the ozone production efficiency but is also influenced by atmospheric removal processes, especially for HN03. The ozone production efficiency is often used as a basis for interpreting ozone chemistry (e.g., Liu et al., 1987). In addition, Sillman (1995, 1998, 1999) has proposed that the value of the ratio 03/N0z can be used as an "indicator" for NOv-sensitive versus HC-sensitive ozone chemistry.


Ozone is produced by a reaction sequence that is initiated by reaction of hydrocarbons or CO with the OH radical. Although individual hydrocarbons follow complex reaction pathways, they often conform to the following pattern:

R02 represents a hydrocarbon chain with 02 attached. The group of R02 radicals and H02 all react rapidly with NO:

resulting in an intermediate hydrocarbon by-product (R'CHO) and N02. This is followed by photolysis of N02:

The resulting oxygen atom rapidly combines with 02 to form ozone (O + 02 + M 03 + M).

The rate of formation of ozone and other smog elements, including sulfate and nitrate aerosols, depends critically on the OH radical, which initiates the reaction sequence. The complex dependence of ozone on NOt and hydrocarbons is closely linked to the chemistry of OH and associated radical species, including H02 and R02 radicals. Because the reaction sequence (1) through (4) operate on the radicals OH, H02, and R02 without changing the sum OH + H02 + R02, it is useful to regard the latter sum as a family of species (odd hydrogen). Much of the complexity of ozone chemistry can be understood by analyzing sources and sinks for this family. Odd hydrogen sources are almost all photolytic reactions and include the following:

Odd hydrogen is removed by reactions that produce hydrogen peroxides and nitric acid:

NO, hno

Formation of peroxyacetyl nitrate (PAN) is also a significant sink for odd hydrogen.

It is possible to derive an analytic solution for OH and for the rate of production of ozone as a function of NOx and HC based on the above reactions (Sillman et al., 1990, 1995). The solution has the form of a fourth degree polynomial for OH, NOx, and HC (or for ozone production, NOx, and HC) and reproduces many of the qualitative features of OH and ozone production as a function of NOx and HC (Fig. 4). HC-limited chemistry occurs when formation of nitric acid (10) represents the dominant loss mechanism for odd hydrogen. In this situation reactions (6), (7), and (10) form an approximate steady state that determines OH. Increased NOt



Cycle Photochemical Smog

1131 1138 1224 1231 1557 1606

Figure 5 Stages in the chemical development of a power plant plume. The three sets of profiles show measurements of S02 (surrogate for NOr, heavy solid line), ozone (dotted line), particulate sulfur (S;J, line-dot-line), all in ppb; and the light-scattering coefficient (5scat, 10"4/m, light solid line) made during crosswind aircraft traverses through the plume of the Cumberland power plant in NW Tennessee on August 23, 1978. The traverses at 80, 110, and 160 km downwind distances illustrate the "early," the "intermediate," and the "mature"stages of chemical development of the plume, respectively. From Gillani et al., 1996.

1131 1138 1224 1231 1557 1606

Figure 5 Stages in the chemical development of a power plant plume. The three sets of profiles show measurements of S02 (surrogate for NOr, heavy solid line), ozone (dotted line), particulate sulfur (S;J, line-dot-line), all in ppb; and the light-scattering coefficient (5scat, 10"4/m, light solid line) made during crosswind aircraft traverses through the plume of the Cumberland power plant in NW Tennessee on August 23, 1978. The traverses at 80, 110, and 160 km downwind distances illustrate the "early," the "intermediate," and the "mature"stages of chemical development of the plume, respectively. From Gillani et al., 1996.

causes a decrease in OH and a consequent decrease in the rate of ozone production [approximately equal to the rate of (1)]. Increased HC causes a modest increase in OH [due to (7)] and a larger increase in the rate of (1), which leads to increased ozone production. NOx-limited chemistry occurs when formation of peroxides [(8) and (9)] represents the dominant sink for odd hydrogen. In this situation the sum H02 + R02 is determined by the steady state between (6), (8), and (9) and is relatively insensitive to changes in NOr or HC. The rate of ozone formation, approximately equal to the rate of reactions (3) and (4), increases with increasing NOt and is largely unaffected by HC.

At nighttime 03 is removed by reaction with NO, as follows:

During the daytime reactions (5) and (11) both occur rapidly, but the combination has little effect on ozone concentrations. However, at nighttime, (5) does not occur and (12) results in removal of 03. Reaction (11) also causes a decrease in ozone during the daytime in the vicinity of a large emission source of NO, e.g., coal-fired power plants. Power plant plumes typically show a decrease in 03 immediately downwind of the plant, followed by recovery and subsequent increase in 03 as the ozone-forming reactions (1) to (4) occur (see Fig. 5) (White et al., 1983; Gillani and Pleim, 1996). This pattern of reduced 03 near the plume source followed by enhanced 03 downwind is similar to the pattern of evolution of urban plumes with HC-limited chemistry near emission sources and NO^-limited chemistry further downwind.


Support for this work was provided by the U.S. National Science Foundation (grant #ATM-9713567).


Bascomb, R., P. A. Bromberg, D. L. Costa, R. Devlin, D. W. Dockery, M. W. Frampton, W. Lambert, J. M. Samet, F. E. Speizer, and M. Utell. Health effects of outdoor air pollution. Am. J. Resp. Crit. Care Med., 153, 477^198, 1996.

Brimblecombe, P., The Big Smoke: A History of Air Pollution in London since Medieval Times, Methuen, London, 1987.

Buhr, M., D Parrish, J. Elliot, J. Holloway, J. Carpenter, P. Goldan, W. Kuster, M. Trainer, S. Montzka, S. McKeen, and F. C. Fehsenfeld, Evaluation of ozone precursor source types using principal component analysis of ambient air measurements in rural Alabama. J. Geophys. Res., 100, 22853-22860, 1995.

Cardelino, C. A., and W. L. Chameides, Natural hydrocarbons, urbanization, and urban ozone, J. Geophys. Res., 95, 13971-13979, 1990.

Chameides, W. L., R. W. Lindsay, J. Richardson, and C. S. Kiang, The role of biogenic hydrocarbons in urban photochemical smog: Atlanta as a case study, Science, 241, 14731474, 1988.

Clarke, J. F., and J. K. S. Ching, Aircraft observations of regional transport of ozone in the northeastern United States, Atmos. Environ., 17, 1703-1712, 1983.

Geron, C. D., A. B. Guenther, and T. E. Pierce, An improved model for estimating emissions of volatile organic compounds from forests in the eastern United States, J. Geophys. Res., 99, 12773-12791, 1994.

Gillani, N. V, and J. E. Pleim, Sub-grid-scale features of anthropogenic emissions of NOx and VOC in the context of regional Eulerian models, Atmos. Environ., 30, 2043—2059, 1996.

Haagen-Smit, A. J., and M. M. Fox, Photochemical ozone formation with hydrocarbons and automobile exhaust, J. Air Pollut. Control Assoc. 4, 105-109, 1954.

Jacob, D. J., B. G. Heikes, R. R. Dickerson, R. S. Artz, and W. C. Keene, Evidence for a seasonal transition from NO,.- to hydrocarbon-limited ozone production at Shenandoah National Park, Virginia, J. Geophys. Res., 100, 9315-9324, 1995.

Jacob, D. J., J. A. Logan, G. M. Gardner, R. M. Yevich, C. M. Spivakowsky, S. C. Wofsy, S. Sillman, and M. J. Prather, Factors regulating ozone over the United States and its export to the global atmosphere, J. Geophys. Res., 98, 14817-14827, 1993.

Jacobson, M. Z., R. Lu, R. P. Turco, and O. P. Toon, Development and application of a new air pollution modeling system—Part I: Gas-phase simulations. Atmos. Environ., 30, 19391963, 1996.

Kleinman, L. I., Seasonal dependence of boundary layer peroxide concentration: The low and high NOx regimes, J. Geophys. Res., 96, 20721-20734, 1991.

Kleinman, L. I., Low and high-NO, tropospheric photochemistry,./ Geophys. Res., 99, 1683116838, 1994.

Lippman, M., Health effects of tropospheric ozone: Review of recent research findings and their implications to ambient air quality standards, J. Expos. Anal. Environ. Epidemiol., 3, 103-128, 1993.

Liu, S. C„ M. Trainer, F. C. Fehsenfeld, D. D. Parrish, E. J. Williams, D. W. Fahey, G. Hubler, and P. C. Murphy, Ozone production in the rural troposphere and the implications for regional and global ozone distributions, J. Geophys. Res., 92, 4191^207, 1987.

Meng, Z., D. Dabdub, and J. H. Seinfeld, Chemical coupling between atmospheric ozone and particulate matter, Science, 277, 116-119, 1997.

Milford, J., D. Gao, S. Sillman, P. Blossey, and A. G. Russell, Total reactive nitrogen (NOv) as an indicator for the sensitivity of ozone to NOY and hydrocarbons, J. Geophys. Res., 99, 3533-3542, 1994.

Milford, J., A. G. Russell, and G. J. McRae, A new approach to photochemical pollution control: Implications of spatial patterns in pollutant responses to reductions in nitrogen oxides and reactive organic gas emissions, Environ. Sci. Technol., 23, 1290-1301, 1989.

Miller, D. F., A. J. Alkezweeny, J. M. Hales, and R. N. Lee, Ozone formation related to power plant emissions, Science, 202, 1186-1188, 1978.

National Research Council (NRC), Committee on Tropospheric Ozone Formation and Measurement, Rethinking the Ozone Problem in Urban and Regional Air Pollution, National Academy Press, Washington, DC, 1991.

Parrish, D. D., J. S. Holloway, M. Trainer, P. C. Murphy, G. L. Forbes, and F. C. Fehsenfeld, Export of North American ozone pollution to the North Atlantic Ocean, Science, 259, 1436-1439, 1993.

Roselle, S. J., and K. L. Schere. Modeled response of photochemical oxidants to systematic reductions in anthropogenic volatile organic compound and NOr emissions, J. Geophys. Res., 100, 22929-22941, 1995.

Ryerson, T. B., M. Trainer, J. S. Holloway, D. D. Parrish, L. G. Huey, D. T. Sueper, G. J. Frost, S. G. Donnelly, S. Schauffler, E. L. Atlas, W. C. Kustler, P. D. Goldman, G. Hubler, J. F. Meagher, and F. C Frehsenfeld, Observations of ozone formation in power plant plumes and implications for ozone control stategies, Science, 292, 719-723, 2001.

Sillman, S. Tropospheric ozone: The debate over control strategies, Annu. Rev. Energy Environ., 18, 31-56, 1993.

Sillman, S., The use of NOv, H202 and HN03 as indicators for 03-N0Y-R0G sensitivity in urban locations, J. Geophys. Res., 100, 14175-14188, 1995.

Sillman, S., The relation between ozone, NOt and hydrocarbons in urban and polluted rural environments. Millenial review series, Atmos. Environ., JJ(12), 1821-1845, 1999.

Sillman, S., K. Al-Wali, F. J. Marsik, P. Nowatski, P. J. Samson, M. O. Rodgers, L. J. Garland, J. E. Martinez, C. Stoneking, R. E. Imhoff, J-H. Lee, J. B. Weinstein-Lloyd, L. Newman, and V Aneja, Photochemistry of ozone formation in Atlanta, GA: Models and measurements, Atmos. Environ., 29, 3055-3066, 1995.

Sillman, S., D. He, M. Pippin, P. Daum, L. Kleinman, J. H. Lee and J. Weinstein-Lloyd, Model correlations for ozone, reactive nitrogen and peroxides for Nashville in comparison with measurements: Implications for VOC-NO^ sensitivity,./ Geophys. Res., 103,22629-22644, 1998.

Sillman, S., J. A. Logan, and S. C. Wofsy, The sensitivity of ozone to nitrogen oxides and hydrocarbons in regional ozone episodes, J. Geophys. Res., 95, 1837-1851, 1990. Sillman, S. and P. J. Samson, The impact of temperature on oxidant formation in urban, polluted rural and remote environments, J. Geophys. Res., 100, 11497-11508, 1995. Simpson, D., Biogenic emissions in Europe, 2, Implications for ozone control strategies, J.

Geophys. Res., 100, 22891-22906, 1995. Sosa, G., J. West, F. San Martini, L. T. Molina and M. J. Molina, "Air Quality Modeling and Data Analysis for Ozone and Particulates in Mexico City." MIT Integrated Program on Urban, Regional and Global Air Pollution Report No. 15, 76 pages, October 2000, available from Trainer, M., D. D. Parrish, M. P. Buhr, R. B. Norton, F. C. Fehsenfeld, K. G. Anlauf, J. W. Bottenheim, Y. Z. Tang, H. A. Wiebe, J. M. Roberts, R. L. Tanner, L. Newman, V C. Bowersox, J. M. Maugher, K. J. Olszyna, M. O. Rodgers, T. Wang, H. Berresheim, and K. Demeijian, Correlation of ozone with NO, in photochemically aged air, J. Geophys. Res., 98, 2917-2926, 1993.

U.S. Congress, Office of Technology Assessment, Catching Our Breath: Next Steps for Reducing Urban Ozone, OTA-O-412, U.S. Government Printing Office, Washington, DC, 1989.

White, W. H., D. e. Patterson, and W. e. Wilson, Jr., Urban exports to the nonurban troposphere: Results from project MISTT, J. Geophys. Res., 88, 10745-10752, 1983. Williams, E. J., A. Guenther, and F. C. Fehsenfeld, An inventory of nitric oxide emissions from soils in the United States, J. Geophys. Res. 97, 7511-7519, 1992.

Was this article helpful?

0 0


  • david sankt
    What is the characteristics of photochemical smog?
    3 years ago

Post a comment