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Urban ozone air quality impact of emissions from vehicles using reformulated gasolines and M85
Pergamon
Atmospheric Environment Vol. 28, No. 17, pp. 2777-2787, 1994 Copyright © 1994 Elsevier Science Ltd Printed in Great Britain. All rights reserved 1352-2310/94 $7.00+0.00
1352-2310(94) E0100-X
URBAN OZONE AIR QUALITY IMPACT OF EMISSIONS FROM VEHICLES USING REFORMULATED GASOLINES AND M85 D. P. CHOCK, S. L. WINKLER, T. Y. CHANG, S. J. RUDY a n d Z. K. SHEN Ford Research Laboratory, Ford Motor Company, P.O. Box 2053, MD3083, Dearborn, MI 48121, U.S.A. (First received 1 November 1993 and in final form 3 April 1994) Abstract--The urban ozone air quality impact of exhaust emissions from vehicles using reformulated gasolines and flexible/variable-fuel vehicles using M85 has been studied using emissions data from the Auto/Oil Air Quality Improvement Research Program and a single-cell trajectory air quality model with two different chemical mechanisms (the updated version of Carbon-Bond-IV (CB4) and the LCC mechanism). Peak ozone concentrations are predicted for each fuel for all combinations of the following ambient conditions: low and high atmospheric dilution or mixing height, four NMOG/NOx ratios, two each of the initial NMOG concentration, the vehicular contribution to the ambient air, and the NMOG composition of the initial ambient mixture. The ozone impact of a fuel depends strongly on the atmospheric dilution and NMOG/NO~ ratio of an area. The differences in ozone impact among fuels are limited under the condition of high atmospheric dilution and a high NMOG/NOx ratio. The ozone-forming potentials (OFPs) for the exhaust emissions based on the maximum incremental reactivities (MIRs) for various fuels are generally well correlated with model-calculated peak ozone levels at a low NMOG/NOx ratio. These OFPs can serve to separate out fuels with rather different reactivities, but not fuels with comparable reactivities. Model-calculated ozone levels for various fuels based on CB4 and LCC mechanisms are relatively well correlated at low NMOG/NOx ratios, but much less so at higher ratios. Fuels with a high aromatic content, including high-toluene fuels, tend to be ranked more favorably by CB4 than by LCC. On the other hand, M85 is ranked more favorably by LCC than by CB4. Fuels with a low 90% boiling point and a low content of aromatics and olefins are generally less reactive. M85 would be an attractive fuel if the formaldehyde emissions could be curtailed significantly. Key word index: Reformulated gasolines, alternative fuels, air quality model, trajectory box model, chemical mechanisms.
1. INTRODUCTION In an attempt to improve the ozone air quality in urban areas, the U.S. Federal and some State governments are mandating the use of reformulated gasolines by gasoline vehicles and encouraging the use of alternatively fueled vehicles. A wealth of data has been collected by the A u t o / O i l Air Quality Improvement Research Program (AQIRP) to establish a link between vehicle emissions and properties of reformulated gasolines and alternative fuels. It has long been recognized that individual or mixtures of volatile organic compounds (VOCs) may differ significantly in their effectiveness in forming ozone in the atmosphere. Accordingly, fuels of different compositions may lead to emissions that have different reactivities. In 1991, the California Air Resources Board (CARB, 1991) adopted regulations for vehicular V O C exhaust emissions that allow the use of reactivity-adjustment factors (RAF) determined under a fixed set of environmental conditions and based on a maximum-incremental-reactivity (MIR) scale deAE 28:17-B
veloped by Carter (1990a). Reactivity scales and ozone-forming potentials (OFPs) of emissions, defined as the amount of additional ozone formed due to emissions from unit distance of travel, are generally sensitive to environmental conditions (Carter and Atkinson, 1989; Chang and Rudy, 1991). Therefore, it is important to evaluate the relationship between the reactivity-based O F P s for the different fuels and the corresponding measured or model-predicted ozone levels under a wide range of environmental conditions. It is also important to see how the ozone air quality impact of the fuels changes with respect to these environmental conditions. Even under identical environmental conditions, the model-predicted ozone levels depend on the chemical mechanisms used. Therefore, it is worthwhile to see how uncertainty in the chemistry influences the prediction of the ozone air quality impact of the fuels. This paper studies the impact of ozone air quality, in terms of the reactivity-based O F P s and the predicted ozone levels, of exhaust emissions from vehicles using reformulated gasolines and a methanol fuel. The
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model predictions are m a d e u n d e r various emission a n d a m b i e n t conditions a n d are based o n two different chemical mechanisms. The air quality model used is a simple trajectory model. A l t h o u g h the trajectory model m a y n o t be a d e q u a t e for describing the ozone air quality in a given u r b a n area, it is useful for studying the relative impact of usage of different fuelvehicle systems on ozone f o r m a t i o n u n d e r a wide range of conditions. The emissions data used are from P h a s e I of the A u t o / O i l A Q I R P . The A Q I R P also calculates reactivities of emissions ( H o c h h a u s e r et al., 1992; Schleyer et al., 1992; A Q I R P , 1993) a n d conducts modeling studies of the air quality impact of reformulated gasolines a n d m e t h a n o l fuels in three u r b a n areas (AQIRP, 1991, 1992b; D u n k e r et al., 1992; C h o c k et al., 1993). The present study, o n the other h a n d , uses simplified a s s u m p t i o n s a n d a simple model for the purpose of uncovering the general response of the fuel-vehicle impact on ozone to changes in emissions, e n v i r o n m e n t a l conditions a n d chemical mechanism.
2. M E T H O D O L O G Y
2.1. Fuels and vehicular emissions
The emissions data are from the current-fleet gasoline vehicles (1989 models) using the AQIRP Phase I reformulated gasolines, and prototype flexible-fuel or variable-fuel vehicles (designated FFV/VFVs) using methanol fuels. The
Phase-I gasoline fuels are characterized by four fuel parameters: aromatic content, methyl-tertiary-butyl ether (MTBE) content, olefin content, and the 90% boiling point (T90). Target values (high and low) for each parameter are given in Table 1. (Actual values for each parameter of fuels deviate somewhat from target values.) The fuels are identified by the alphabetic code given in Table 2. Two additional gasoline fuels are also included in the fuel matrix: fuel A or the industry-average (Ind Avg) fuel (32 vol% aromatics, no MTBE, 9 vol% olefins, a T90 of 165°C (330°F), and fuel B, the certification (CERT) fuel, or indolene (30 vol% aromatics, no MTBE, 4.6 vol% olefins, a T90 of 154°C (309°F). The methanol fuel used is M85 (fuel Z), which is a mixture of 85 vol% methanol and 15 vol% lnd Avg fuel. The target Reid vapor pressure (RVP) of all fuels is 60 kPa (8.7 psi). In order to satisfy all the constraints of the RVP and target values of the fuel parameters, the fuel compositions must be altered as necessary. For example, n-butane was added to the gasoline component to adjust the RVP of M85 to 60 kPa. The AQIRP shows that reducing the RVP of the nonoxygenated fuels reduces the diurnal evaporative emissions significantly (AQIRP, 1992a). All gasolines except fuel B have a sulfur content of about 300 ppm by weight. Fuel B has 120 ppm of sulfur. The AQIRP shows that reducing the fuel sulfur content reduces the (post-catalyst) exhaust NMOG, CO and NO x emissions significantly (AQIRP, 1992c). The AQIRP current-fleet gasoline vehicles consist of 16 cars and four light-duty trucks with an odometer range of 16,000-47,000km (10,000-29,000miles). The FFV/VFV fleet consists of 19 (17 cars and 2 trucks) pre-1990 test prototypes with an odometer range of 14,000-53,000 km (9000-33,000 miles). These are not production vehicles. So, comparing the relative ozone impact due to gasoline and methanol fuels based on the present AQIRP data may not be appropriate. However, it is included here to see how the ranking of M85, in terms of ozone air quality benefit, is affected by the use of different chemical mechanisms.
Table 1. Fuel design parameters for the auto/oil gasolines Target value Variable Aromatics (vol%) MTBE (vol%) Olefin (vol%) T90 (°C)
Low
High
Industryaverage
Cert. fuel
20 0 5 138-149
45 15 20 177-182
32 0 12 168
30 0 5 154
Table 2. Auto/oil Phase I reformulated gasolines and methanol fuels and their average measured NMOG exhaust emissions for cars Code A B C D E F G H I J
ID
ExhEM (gkm -1)
Code
0.088 0.079 0.082 0.086 0.106 0.086 0.082 0.075 0.123 0.096
K L M N O P Q R
Amot AmOT aMOT aMot AMOt amOt amoT AMoT
0.086 0.116 0.083 0.076 0.078 0.080 0.102 0.119
Z
M85
0.182
Ind Avg CERT AMot amOT AMOT amot AmOt aMOt AmoT aMoT A/a--high/low aromatics. M/m--high/low MTBE. O/o--high/low olefins. T/t--high/low T90.
ID
ExhEm(gkm ~)
Urban ozone air quality impact The exhaust emissions are based on the U.S. Federal Test Procedure (FTP) which contains three driving phases: cold start or bag 1; hot stabilized or bag 2; and hot start or bag 3. Evaporative and running-loss emission measurements are available for 13 of the 18 gasoline fuels; and the measurements have much greater uncertainties than those for exhaust emissions. In this study, differences in the ozone impact among the gasoline fuels will be based on the differences in exhaust emissions. This assumption was made so that the calculated ozone impact of the fuels can be compared directly with the OFPs which are defined for exhaust emissions. It also allows the inclusion of all gasoline fuels in the comparison, not just those with evaporative and running-loss emissions. The implication of ignoring the differences in evaporative and running-loss emissions among the gasoline fuels will be discussed later. All fuels except M85 are assumed to have evaporative and running-loss emissions identical to those for fuel A. For M85, the measured amounts and compositions of evaporative and running-loss emissions will be used. Use of separate evaporative and running-loss emissions for M85 in the model implies that the model result should not be compared directly with the exhaust-based OFP for M85. Based on the U.S. Environmental Protection Agency's (EPA) MOBILE4 estimates, it is further assumed that the running-loss emissions are 1.8 and 1 times the evaporative emissions for cars and trucks, respectively. For M85, the measured NMOG mass exhaust emissions from the prototype FFV/VFVs are about twice those for fuel A. Since the molecular weight of methanol (32), a major component of the M85 exhaust, is more than twice the molecular weight per carbon (13.88) of the gasoline exhaust, we shall assume that the exhaust mole-carbon (mole-C) emissions for M85 are the same as those for fuel A. The corresponding NMOG exhaust emission is slightly greater than measured. For simplicity, it is also assumed that the NOx and CO emissions for fuel A are applicable to those for other fuels in our model simulations since NOx emissions do not vary as greatly across the fuels as NMOG emissions and the sensitivity of ozone impact to the variation of CO emissions across the different fuels is small. Approximately 20% of all tests with the gasoline fuels and more than 50% of all tests with the methanol fuels have speciated exhaust emissions data which are given in weight percents. In the speciation tests, up to 151 organic species were identified. All tests, however, have data for total nonmethane hydrocarbon mass exhaust emission rates (g km - 1) for the three FTP bags. All evaporative and running-loss tests contain speciated measurements. Their composition measurements for fuel A and for M85 are used in the present study. To estimate the NMOG emission composition, the unidentified species (generally less than 5% by weight) are assumed to have the same composition as the identified nonmethane species (excluding formaldehyde and methanol in the case of methanol fuels). An averaging procedure was used for both the weight percent of each species and the total NMOG mass emission rate for each of the emission phases (bags 1, 2, 3, hot-soak evaporative, diurnal evaporative and running loss). The averaging procedure for the gasoline fuels had three steps: first, averaging over all the tests on a vehicle for each fuel; next, averaging over vehicles of the same model; finally averaging over all models for cars and all models for trucks. For the methanol fuels, straight averaging was used for cars and trucks separately. The averaging procedure above is used to determine the speciated average weight percents from the composition data base and the total NMOG mass emission rates from the full data base for each emission phase and each fuel. Then the average speciated mass emission rates are calculated by multiplying the average weight percents by the average total NMOG mass emission rates for each phase and each fuel. Two of the gasoline cars used the carbureted technology. Their emissions data are not included in the averaging procedure. For each fuel, the average total NMOG mass exhaust
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emissions and the average weight percents of the calibrated nonmethane organic species for the three FTP bags are combined according to the FTP weightings for the three bags. The evaporative and running-loss emissions are also converted to the same units as the exhaust emissions following the MOBILE4 prescription (USEPA, 1989). The average measured NMOG exhaust emissions for cars for all fuels are shown in Table 2. These numbers are similar but not identical to those used in the AQIRP because of differences in assumptions. The converted evaporative emissions for cars are 0.019 gkm I for fuel A and 0.051 gkm -1 for M85. The average measured NOx emission for fuel A is 0.342 g k m - 1; the variation from fuel to fuel is typically 10% or less. It is known that the cold-start (bag 1) exhaust NMOG emissions from current gasoline vehicles dominate the total NMOG emissions and contain a significant amount of unburned fuel components and lower alkenes generated by incomplete combustion. Consequently, the cold-start mode carries the fuel signature into the exhaust emissions and dominates the exhaust mass and reactivity. For M85, the NMOG reactivity is influenced strongly by the formaldehyde emission rate. 2.2. Air quality modelino and simulation conditions The Ozone Isopleth Plotting Package with Optional Mechanism (OZIPM) (Hogo and Gery, 1988) was used to study the air quality impact of the various fuels under various emission and ambient conditions. The two chemical mechanisms used are the Lurmann-Carter-Coyner (LCC) (Lurmann et al., 1987) and the Carbon-Bond IV (CB4) (Gery et al., 1989) mechanisms. The LCC mechanism is an older version of the Statewide Air Pollution Research Center mechanism, the most recent version of which is the SAPRC90 mechanism (Carter, 1990b). The latter contains some rate constant updates and is more detailed; but the overall characteristics of both versions are quite similar. The CB4 mechanism used in the present paper is the recently updated (USEPA, 1993) version which is occasionally referred to as CBM4.1. The simulation conditions are summarized in Table 3. The simulations were started with a given initial concentration of NMOG in the central part of an urban area at 0800 LDT. Two initial NMOG concentrations were assumed: 0.4 and 0.7 ppmC (referred to as "c" and "C", not to be confused with fuel C). They represent the typical range of 6 9 a.m. NMOG concentrations observed in an urban area. The initial NOx concentrations were varied depending on the NMOG/NOx ratio used. The initial NMOG/NOx ratios considered are 6, 8, 10 and 12 ppmC ppm-~ (denoted "R6," "R8," etc.), a range covering most urban areas. NMOG and NOx are emitted to the air parcel as a fraction of the initial concentrations, as shown in Table 3. Pollutants are entrained from aloft as the mixing height changes through the day. Two dilution scenarios were chosen: "low dilution" (the mixing height rising from 300 m at 0800 LDT to 600 m by mid-afternoon; denoted "LD') and "high dilution" (the mixing height rising from 300 to 1500 m, denoted "HD'). The former tends to occur in an areal band downwind of and parallel to a large body of water while the latter is quite typical of inland regions. The ambient NMOG compositions used in the present paper are shown in Table 3: U.S. "urban average" based on 41 cities (denoted "EPA41"), given by Jeffries et al. (1989), and that measured during the Southern California Air Quality Study (SCAQS) in 1987 (Lurmann et al., 1992). The "urban average" NMOG composition database did not have measured aldehydes and the assumed formaldehyde concentrations are much higher than the measured concentrations from SCAQS (see Table 3). To represent the base case for fuel-vehicle emission impact on ozone formation, a certain mole-C fraction (0.15 or 0.30, denoted "v" and "V", respectively) of the initial NMOG mixture was replaced by the composition of the emissions for
2780
D. P. CHOCKet al. Table 3. Simulation conditions Latitude: Temperature: Simulation period:
34.1°N 303 K 0800 to 2100 LDT
Initial concentrations: NMOG NOx (NOz/NO x = 0.25) CO CH,~
Either 0.4 or 0.7 ppmC Variable depending on NMOG/NOx ratio 1.2 ppm 2.0 ppm
Emission schedule
Aloft concentrations: NMOG 03 Mixing height: t Low dilution case High dilution case NMOG composition:$ Mechanism CB4 Mech.
LCC Mech.
Time (LDT) 0800-1300 1300-1800 0.04 ppmC 0.06 ppm
NMOG 0.15 0.05 CO CH 4
Initial at 0800 L D T 300 m 300 m Species PAR ETH OLE TOL XYL HCHO ALD2 ISOP NR
Date: Relative humidity:
21 June 50%
NOx 0.25 0.08 0.5 ppm 1.6 ppm Final at 1500 L D T 600 m 1500 m
Ambient carbon fractions EPA41 SCAQS 0.564 0.550 0.037 0.042 0.035 0.023 0.089 0.096 0.117 0.104 0.021 0.010 0.052 0.035 0.000 0.003 0.084 0.137
ALK4 ALK7 ETHE PRPE TBUT TOLU XYLE TMBZ HCHO ALD2 ISOP ACET MEK NR
CO 0.10 0.03
0.214 0.280 0.037 0.048 0.029 0.080 0.079 0.042 0.020 0.030 0.000 0.000 0.000 0.141
0.235 0.272 0.042 0.033 0.012 0.096 0.088 0.018 0.010 0.033 0.003 0.024 0.010 0.124
Aloft carbon fractions 0.498 0.034 0.020 0.042 0.026 0.070 0.037 0.000 0.273 0.363 0.151 0.034 0.010 0.014 0.042 0.024 0.000 0.070 0.037 0.000 0.000 0.000 0.255
* Fraction of initial concentration emitted per hour (Hogo and Gery, 1988). t Mixing height varies with time according to the characteristic curve recommended by Hogo and Gery (1988). ~ Aloft and ambient EPA41 compositions are from Jeffries et al. (1989). Ambient SCAQS composition comes from Lurmann et al. (1992). Abbreviations for species are defined in Jeffries et al. (1989).
fuel A. For a different gasoline fuel, the exhaust portion of emissions for fuel A was replaced by the mole-C/kin ratio of that fuel to fuel A; the resulting fraction accommodates the exhaust NMOG composition of that fuel. Consequently, the sum of vehicular and ambient mole-C fractions can differ from unity. As stated earlier, the exhaust mole-C fraction of M85 is assumed to be the same as fuel A, but the evaporative and running-loss emission fractions were modified by the mole-C/km ratios relative to the respective emissions for fuel A, as described above. The mole-C percents for NMOG emissions from cars and trucks are taken to be 80/20. For fuel A emissions, this split corresponds roughly to a mileagecontribution split of 90/10. One should note that the composition of the ambient mixture itself already reflects a significant contribution from vehicular emissions. Therefore, our partial replacement of the mixture with the composition and relative amount for a fuel
does not adequately describe the true vehicular contribution to the ambient mixture. However, the relative impact of the fuel will be manifested. Furthermore, the use of two vehicular mole-C fractions allows one to appreciate the sensitivity of this impact to the change in the vehicular contribution to the ambient NMOG. On the whole, we have two alternatives for each of the following conditions: atmospheric dilution, initial NMOG concentrations, vehicular fractions in these concentrations, ambient mixture compositions, and chemical mechanisms. In addition, we have four NMOG/NO x ratios. So we have a total of 128 test cases per fuel. 2.3. Ozone forming potential The OFP of the exhaust NMOG emissions for each fuel is calculated based on the M1R and the maximum ozone
Urban ozone air quality impact incremental reactivity (MOR) scales. MIRs have been determined at relatively low NMOG/NOx ratios where the incremental reactivities of atmospheric NMOG mixtures are at their maximum values, and MORs have been determined at relatively high NMOG/NO= ratios where the daily peak (1 h average) ozone levels are at their maximum values. The OFP (in terms of g 0 3 per kilometer of travel) for each fuel is determined by the sum of products of each NMOG species emission rate (g km ~) and its incremental reactivity (gO3/gNMOG, for either MIR or MOR). Because the emission differences among the gasoline fuels in this study is characterized by the differences in exhaust emissions alone, the calculated OFPs for gasoline fuels can be compared directly with the ozone impact of the fuels estimated by the trajectory model.
3. RESULTS 3.1. Comparison of peak ozone concentrations The ambient conditions, including the atmospheric dilution, the concentration and composition of N M O G in the urban mixture, the N M O G / N O ~ ratio,
0.17-
a: LD
2781
all play a critical role in determining the ozone level. These ambient conditions influence not only the ozone air quality impact from the different fuels, but also the relative ranking of these fuels based on the impact. Figure 1 shows the ozone concentrations for the different fuels under high and low dilution (HD and LD), N M O G / N O x ratios of 6 and 10 p p m C / p p m (R6 and R10, respectively), the two ambient N M O G compositions for urban-area mixtures (SCAQS by upper-case letters and EPA4! by lower-case letters), and the two chemical mechanisms (LCC on the ordinate and CB4 on the abscissa). The initial N M O G concentration and vehicle fraction are held constant at 0.7 ppmC and 0.30, respectively. The ozone levels based on the EPA41 mixture are considerably higher than those based on the SCAQS mixture. This is because EPA41 is a more reactive mixture than SCAQS. It has twice the mole-C fraction of formaldehyde (0.02) as the SCAQS mixture (0.01). The EPA41 mixture also contains higher mole-C fractions of
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Fig. 1. Comparison of peak ozone concentrations of all fuels based on two mechanisms and two urban-mixture compositions (upper-case letters for SCAQS and lower-case letters for EPA41) using an initial NMOG concentration of 0.7 ppmC and a vehicular mole-C fraction of 0.30. (a) LD and R6; (b) HD and R6; (c) LD and R10; (d) HD and R10.
2782
D.P. CHOCKet al.
higher olefins and polyalkylated aromatics which are highly reactive, and a lower fraction of toluene which is not very reactive. Despite the significant differences in the N M O G compositions of the two ambient mixtures, the patterns of the fuels' ozone levels are quite similar. Henceforth, we shall discuss primarily the results based on the SCAQS mixture unless otherwise indicated. Figure 1 reveals the following. First, high dilution does not necessarily imply a lower ozone concentration. In fact, the concentration can be higher than the low-dilution case under a low but identical N M O G / NO~ ratio. This is because the NOx-inhibition effect is most effective at a low N M O G / N O x ratio, and diluting NOx reduces its effectiveness in reducing ozone. Second, the range of the ozone concentrations for the fuels considered is wider for the lower N M O G / N O x ratio than for the higher ratio. This is consistent with the fact that the lower ratio is closer to the region where the MIR applies, so that the sensitivity of the ozone level to the fuel is at or near a maximum. The characteristics of the ozone concentration range will be further discussed below. Third, the correlation between the ozone concentrations predicted by LCC and CB4 is very good for the low NMOG/NOx ratio bt, t deteriorates for the high ratio. At a high NMOG/NOx ratio, the "At" fuels (fuels C, K, O and G, all with a high aromatic content ("A") and a low T90 Ct'). See Table 2) all yield essentially the lowest ozone concentrations compared to other fuels according to CB4, but not the lowest according to LCC. This is because these four fuels have very high toluene contents (36%, 35, 31 and 29% by weight for fuels C, K, O and G, respectively) and toluene is not efficient in ozone production at a high NMOG/NO~ ratio, but is even less efficient according to CB4 (by scavenging NOx, leading to a'negative MOR) than according to LCC. The "AT" fuels (fuels I, R, L and E) are the most reactive fuels. They all have a high aromatic content but with a large proportion of the highly reactive polyalkylated aromatics. Both mechanisms rank them as the most reactive fuels at a low NMOG/NOx ratio. But at a high N M O G / N O x ratio, CB4 again finds these fuels to be not as reactive as some other fuels with a low aromatic content while LCC continues to rank them as the most reactive. The difference in the treatment of the "At" and "AT" fuels results from a difference in aromatic chemistry assumed in both mechanisms and contribute to a significant reduction in the correlation between the predicted ozone concentrations from both mechanisms. It should also be noted that fuel Z is considered more reactive in CB4 than in LCC because the key reactive emission from fuel Z is formaldehyde, and CB4 has a more reactive formaldehyde chemistry (partly due to a higher photolysis rate) than LCC. Figure 2 shows the ozone concentrations for selected fuels as a function of the NMOG/NO~ ratio, the dilution, the initial N M O G concentration, the mole-C fraction of N M O G attributed to vehicles, and the
chemical mechanism. The fuels selected are C, E and Z. Fuel C is one of the least reactive fuels. Fuel E is one of the most reactive, and fuel Z or M85 is the only nongasoline fuel considered here. Under low dilution, the ozone levels tend to peak at an N M O G / N O x ratio of 10 ppmC p p m - ~ or higher. As the ratio decreases, the ozone levels drop rather sharply. The ratio where the ozone concentrations spread the most is around 8 ppmCppm -1. Also, increasing the initial N M O G concentration while keeping the NMOG/NOx ratio the same (by increasing the initial NOx concentration as well) does not change the ozone concentration significantly at a low ratio, but increases the ozone concentration significantly at a higher ratio (10 ppmC p p m - ~or more). This agrees with the expectation that NO x is more effective in inhibiting ozone at a low ratio than at a high ratio. Increasing the fraction of vehicle contribution in the ambient mixture tends to lower the ozone concentration for the less reactive fuels, but increases the ozone concentration, especially at a ratio of around 8 ppmC ppm - 1, for the more reactive fuels. This indicates that the reactive fuels such as fuel E is more reactive than the SCAQS urban mixture while fuel C, and to a lesser degree, fuel Z, are less reactive than the mixture. Under high dilution, the variation of ozone concentrations over an NMOG/NOx ratio of 6 to 10 ppmC p p m - I is more restricted. In contrast to the lowdilution case, the ozone concentrations for the different fuels have a rather limited spread, the greatest spread being at a ratio of about 6 ppmC p p m - 1. This spread is still substantially less than the largest spread in the low dilution case. Thus, in high dilution, the difference in the ozone impact of different fuels is not large except at a low N M O G / N O x ratio. The highest ozone levels tend to occur at an N M O G / N O x ratio of around 8 ppmC p p m - 1. Thus, in principle, a different atmospheric dilution rate requires a different N M O G / N O x ratio on which the MIR or MOR should be defined. For a given initial N M O G concentration, increasing the vehicle contribution from 0.15 mole-C fraction to 0.30 mole-C fraction does not change the ozone concentration substantially. Increasing the initial N M O G concentration at a fixed NMOG/NOx ratio lowers the ozone concentration at a ratio of 6 ppmC ppm- 1 but increases it at a higher ratio. It appears then that the NOx-inhibition effect works better at high dilution than at low dilution. Figure 3 shows the predicted ozone concentrations for fuels C and E for the two dilution rates (HD and LD), two NMOG/NO~ ratios (R6 and R10), the two initial N M O G concentrations (c and C) and vehicular mole-C fractions (v and V), and the two chemical mechanisms (LCC and CB4), based on the SCAQS ambient N M O G composition. The discussions provided earlier are demonstrated vividly in the figure. It is also seen that under LD, the ozone concentration predicted by CB4 has a greater sensitivity to the NMOG/NOx ratio than that predicted by LCC.
Urban ozone air quality impact 0.25
2783
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12
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8
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NMOG/NOx (ppmC/ppm) Fuel C
..... Fu.e! _ E
Fuel Z
• CB4
~7 LCC
Fig. 2. Peak ozone concentrations for fuels C, E and Z as a function of the NMOG/NOx ratio and chemical mechanism (solid lines for CB4, dashed lines for LCC). The eight plots correspond to all combinations of low and high dilution (LD and HD), initial NMOG concentration of 0.4 and 0.7 ppmC ("c" and "C"), and vehicular mole-C fraction in the ambient mixture of 0.15 and 0.30 ("v" and "V"). 3.2. Comparison of ozone-forming potential and peak ozone concentrations Figure 4 shows the relation between the O F P s for the various fuels and the LCC-predicted peak ozone
concentrations at an N M O G / N O x ratio of 6 and 10 p p m C p p m - 1 (R6 and R10). The M I R - and M O R based O F P s are used for R6 and R10, respectively. Both low and high dilution cases are considered, as are the SCAQS and EPA41 base mixture compositions.
2784
D . P . CHOCK et al.
Fuel C HD
0.4(~ 0.35
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R6
R6
RIO C
. v
¢ ¥
V
V
v
RIO
v=V vCv =v/vCv
C Y
0.30 0.25 E
liill
0.20 0.15 0.10 0.05 0.00
3.3. Relative ranking of the fuels
Fuel E HD
o 401
io
0.35 v
R6 Y
v
LD
o °tcLo ° ot°l ] I
V
RIO
v V
I
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V
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0 0.15
.
.
.
ozone concentrations and the OFPs is expected to be somewhat worse, especially for R 10. In particular, the predicted ozone concentrations for fuel Z will be even higher than what might be expected based on its O F P while those for fuels I, R, L and E will be lower than expected, From the above, we see that O F P is useful in providing a rule of thumb in estimating the relative perturbations on ozone air quality of fuels that have rather different reactivities. But it may not be reliable when applied to fuels that have comparable reactivities. Moreover, the range of the OFPs has little to do with the range of the actual ozone concentrations which depend on ambient conditions.
.
0.00
LCC [] CB4 Fig. 3. Peak ozone concentrations for fuels C and E as a function of dilution rate (HD and LD), NMOG/NO x ratio (R6 and R10), initial NMOG concentration ("c" and "C") and vehicular mole-C fraction ("v" and "V"). The initial N M O G concentration is assumed to be 0.7 ppmC and the vehicle contribution is 0.30 in mole-C fraction. The correlation between the predicted ozone concentrations and the MIR-based OFPs is generally good for both the SCAQS and the EPA41 mixtures for R6 (and, actually, for R10 as well) and for both LD and HD. But if we are to expand the OFP scale that covers the low-T90 fuels alone, then the correlation is poor. Here, the impact of ambient N M O G composition also becomes evident. The correlation between the predicted ozone concentrations for R10 and the MOR-based OFPs is only fair at best. Note that since the OFPs are defined for exhaust emissions alone, drawing conclusions for fuel Z may not be appropriate since its evaporative and running-loss emissions have been modified for model calculation. Since the OFPs are derived using a detailed SAPRC mechanism which is more closely related to LCC than to CB4, the correlation between the CB4-predicted
The modeling results show that the "T" (high T90) fuels are more reactive than the "t" fuels, while the relative ozone impacts of the "t" fuels are less than that of fuel A and are less distinct from each other. The "AT" fuels are the most reactive of all gasoline fuels tested, primarily because these fuels lead to the highest exhaust N M O G emission rate and the emissions tend to be highly reactive due to the significant amount of polyalkylated aromatics present. These fuels are also more reactive than the urban mixture so that with a constant N M O G concentration, the greater the vehicular contribution, the higher the ozone level. The ranking by the two mechanisms of these four fuels relative to other "T" fuels can be different at a high NMOG/NOx ratio because of the difference in the aromatic chemistry assumed in both mechanisms. The "t" or low-T90 fuels and fuel B are much less reactive. Among them, the "At" fuels are considered the least reactive by CB4 but not necessarily by LCC especially at a high N M O G / N O x ratio. This is the result of the difference in the toluene chemistry assumed in the mechanisms. According to LCC, fuels H and P are also among the least reactive fuels. Note that fuels G, O, P and H have a high olefin content. They have a somewhat lower exhaust N M O G emission rate, but the emission tends to have higher reactivity. High-olefin fuels also have higher and more reactive evaporative emissions. If evaporative emissions of the individual fuels were included in the simulation, these fuels would lead to higher ozone concentrations as was shown by AQIRP (1991) in the three-city study. The role of MTBE on ozone air quality is small in comparison to the other fuel factors. Fuel Z (M85) produces less ozone than does fuel A at a low N M O G / N O x ratio. However, as the ratio increases, fuel Z produces a comparable amount of ozone as fuel A according to CB4 but still less ozone than does fuel A according to LCC. The difference in the formaldehyde chemistry of the two mechanisms must be reconciled before a more definitive statement on fuel Z can be made. Fuel N and Z have the lowest MIR-based OFPs. But they do not yield the lowest ozone concentrations
Urban ozone air quality impact 0 . 1 7 a" L D
R6
0.17
r
e
2785
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. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . I m
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. . . . . . . . . . . . . . . . . . d a
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.........
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DA 0.11
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c: LD R10
J . . . . . . . . . . . . . . . . d 8 z
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q
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5
E
e
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.
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AD . . . . . . . . . .
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.
.
.
.
.
.
.
.
.
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MOR-based OFP (g On/kin)
0.1g 0.40
N%
0.45
0.50
0.55
0.60
0.65
MOR-based OFP (g Oa/km)
Fig. 4. Comparison of LCC-derived peak ozone concentrations based on the two urban mixtures (upper-case letters for SCAQS and lower-case letters for EPA41) for all fuels and the MIR-(for R6 only) and MOR-(for R10 only) based OFPs. The initial NMOG concentration is 0.7 ppmC and the vehicularmole-C fraction is 0.30. (a) LD and R6; (b) HD and R6; (c) LD and RI0; (d) HD and R10.
in the simulations. In general, however, the MIRbased OFPs of the "t" fuels are quite similar as are their predicted ozone concentrations at a low N M O G / N O x ratio. If we exclude high-olefin fuels that have reactive evaporative emissions, then fuels C, K, B, F and N yield the lowest ozone concentrations. Ranking among these fuels is difficult not only because the differences among their ozone impacts are small to begin with, but also because the different assumptions in the two chemical mechanisms produces a different ranking, and the ranking changes as the atmospheric dilution and N M O G / N O x ratio varies. Fuel Z yields a relatively high ozone impact primarily because of its relatively high formaldehyde emissions (0.012 g km - 1) from the prototype FFV/VFVs. Reduction of the formaldehyde emission would significantly lower the ozone air quality impact of this fuel. 4. C O N C L U S I O N S
The present study shows that the extent of atmospheric dilution and the NMOG/NOx ratio play a AE 28:17-C
critical role in determining the ozone impact of fuels. Varying the dilution leads to very different behavior of peak ozone concentrations. In particular, the relation between peak ozone and the N M O G / N O x ratio is altered significantly. For example, at an NMOG/NOx ratio of 6 ppmC ppm- 1, the peak ozone concentration under high dilution can actually be higher than under low dilution and decreases with increasing initial N M O G (and NOx) concentration due to a more efficient NOx-inhibition effect. Added to this complexity is uncertainty in the chemistry so that different mechanisms predict the peak ozone concentrations somewhat differently. Consequently, both consistent patterns and unresolved issues emerge in the present study. Here are the consistent patterns. The MIR-based OFPs for various fuels are generally well correlated with model-calculated peak ozone levels at a low NMOG/NOx ratio. The correlation is even reasonable for a ratio of 10 ppmC ppm- ~ if the LCC mechanism is used to calculate the peak ozone concentrations. However, while all fuels with a low T90 have relatively low OFPs for exhaust emissions and yield
2786
D.P. CHOCKet al.
relatively low peak ozone levels, there is little correlation between these fuels' O F P s and their corresponding ozone concentrations. Thus, the MIR-based O F P s can serve to separate out fuels with rather different reactivities, but not fuels with comparable reactivities. The O F P s cannot be used to estimate the change in ozone concentration because such change depends strongly on the atmospheric dilution, the N M O G / N O x ratio, and the N M O G concentration. Modelcalculated ozone levels for various fuels based on CB4 and LCC mechanisms are relatively well correlated at low N M O G / N O ~ ratios, while the correlation becomes poorer at higher ratios. The unresolved issues in this study concern the uncertainty in the chemistry of toluene and aromatics in general, and the free-radical initiation process in the chemical mechanisms. Thus, CB4 has a tendency to rank fuels with a high aromatic content less harshly and high-toluene fuels more favorably than LCC at high N M O G / N O x ratios. It also tends to rank fuel Z less favorably than LCC. If we exclude fuels with a high olefin content because of their tendency to have high and reactive evaporative e m i s s i o n s - - a factor not considered in the present study, then fuels with a minimal ozone impact are fuels C, B, N, K and F. If we further exclude the nonoxygenated fuels, then only fuels C and N are left. If we do not want to take advantage of the low reactivity of toluene and simply exclude fuels with a high aromatic content, then fuel N is the only reasonable choice in the gasoline fuel matrix for contributing the least to ozone formation from vehicle emissions. Fuel Z (M85) may still be a very attractive fuel provided that the formaldehyde emissions can be reduced significantly. In terms of actual ozone impact, the advantage of using a fuel with low reactivity depends on, among many things, the typical atmospheric dilution and N M O G / N O x ratio for the area. For example, the impact would be quite limited if the prevalent conditions entail high atmospheric dilution and a high N M O G / N O x ratio. REFERENCES
AQIRP (1991) Air quality modeling results for reformulated gasolines in year 2005/2010. Technical Bulletin No. 3, Auto/Oil Air Quality Improvement Research Program. Available from the Coordinating Research Council, Inc., 219 Perimeter Parkway, Atlanta, GA, 30346. AQIRP (1992a) Emissions results of oxygenated gasolines and changes in RVP. Technical Bulletin No. 6, Auto/Oil Air Quality Improvement Research Program. AQIRP (1992b) Emissions and air quality modeling results from methanol/gasoline blends in prototype flexible/variable fuel vehicles. Technical Bulletin No. 7, Auto/Oil Air Quality Improvement Research Program. AQIRP (1992c) Effects of fuel sulfur on mass exhaust emissions, air toxics, and reactivity. Technical Bulletin No. 8, Auto/Oil Air Quality Improvement Research Program. AQIRP (1993) Reactivity estimates for reformulated gaso-
lines and methanol/gasoline mixtures. Technical Bulletin No. 12, Auto/Oil Air Quality Improvement Research Program, June. CARB (1991) Initial statement of proposed rulemaking reactivity adjustment factors for transitional low-emission vehicles and staff's suggested changes to the proposal. California Air Resources Board, Sacramento, CA. Carter W. P. L. (1990a) Development of ozone reactivity scales for volatile organic compounds. Statewide Air Pollution Research Center, Riverside, CA, US EPA Contract CR 814396-01-1. Carter W. P. L. (1990b) A detailed mechanism for the gasphase atmospheric reactions of organic compounds. Atmospheric Environment 24A, 481-518. Carter W. P. L. and Atkinson R. (1989) Computer modeling study of incremental hydrocarbon reactivity. Envir. Sci. Technol. 23, 864-879. Chang T. Y. and Rudy S. J. (1991) Impact of organic emissions from alternatively-fueled vehicles on urban ozone air quality. In Tropospheric Ozone and the Environment, Transaction No. 19, Air and Waste Management Association, pp. 371 390. Chock D. P., Yarwood G., Dunker A. M., Morris R. E., Pollack A. K. and Schleyer C. H. (1993) Sensitivity of urban airshed model results for test fuels to uncertainties in light-duty vehicle and biogenic emissions and alternative chemical mechanisms auto/oil air quality improvement research program. Presented at Air and Waste Management Association Regional Photochemical Measurement and Modeling Studies Conference, San Diego, CA, 8-12 November. Dunker A. M., Morris R. E., Pollack A. K., Cohen J. P., Schleyer C. H., Chock D. P., Hertz M. and Metcalfe J. E. (1992) Effects of aromatics, MTBE, olefins, and T90 on urban air quality in year 2005/2010 auto/oil air quality improvement research program. Paper 92-119.03, Air and Waste Management Association 85th Annual Meeting, Kansas City, MO, 21-26 June. Gery M. W., Whitten G. Z., Killus J. P. and Dodge M. C. (1989) A photochemical kinetics mechanism for urban and regional scale computer modeling. J. geophys. Res. 94, 12,925-12,956. Hochhauser A. M., Benson J. D., Burns V. R., Gorse R. A., Koehl W. J., Painter L. J., Reuter R. M. and Rutherford J. A. (1992) Speciation and calculated reactivity of Automotive exhaust emissions and their relation to fuel propert i e s - auto/oil air quality improvement research program. Society of Automotive Engineers Paper No. 920325. Hogo H. and Gery M. W. (1988) User's Guide for Executing OZIPM-4 with CBM-IV or Optional Mechanism. EPA/600/8-88/073b, U.S. Environmental Protection Agency, Research Triangle Park, NC. Jeffries H. E., Sexton K. G. and Arnold J. R. (1989) Validation testing for new mechanisms with outdoor chamber data, Vol. 2: analysis of VOC Data for the CB4 and CAL Photochemical mechanisms. U.S. Environmental Protection Agency, Research Triangle Park, NC, Cooperative Agreement CR-813107. Lurmann F. W., Carter W. P. L. and Coyner L. A. (1987) A surrogate species chemical reaction mechanism for urbanscale air quality simulation models. EPA/600/3-87/014a and b, U.S. Environmental Protection Agency, Research Triangle Park, NC. Lurmann F. W., Main H. H., Knapp K. T., Stockburger L., Rasmussen R. A. and Fong K. (1992) Analysis of the ambient VOC data collected in the southern California air quality study. Final Report. STI-99120-1161-FR, CARB Contract No. A832-130, Sonoma Technology Inc., Santa Rosa, CA. Schleyer C. H., Dunker A. M., Whitten G. Z. and Pollack A. K. (1992) Reactivity of total organic gas emissions from
Urban ozone air quality impact reformulated gasolines--auto/oil air quality improvement research program. A W M A Paper No. 92-119.02, Air and Waste Management Association 85th Annual Meeting, Kansas City, MO, 21-26 June. USEPA (1989) User's Guide to MOBILE4 (Mobile Source
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Emission Model). Office of Mobile Sources, U.S. Environmental Protection Agency, Ann Arbor, MI. USEPA (1993) Urban Airshed Model Version 6.20, Dated 920825. U.S. Environmental Protection Agency, Research Triangle Park, NC.
Atmospheric Environment Vol. 28, No. 17, pp. 2777-2787, 1994 Copyright © 1994 Elsevier Science Ltd Printed in Great Britain. All rights reserved 1352-2310/94 $7.00+0.00
1352-2310(94) E0100-X
URBAN OZONE AIR QUALITY IMPACT OF EMISSIONS FROM VEHICLES USING REFORMULATED GASOLINES AND M85 D. P. CHOCK, S. L. WINKLER, T. Y. CHANG, S. J. RUDY a n d Z. K. SHEN Ford Research Laboratory, Ford Motor Company, P.O. Box 2053, MD3083, Dearborn, MI 48121, U.S.A. (First received 1 November 1993 and in final form 3 April 1994) Abstract--The urban ozone air quality impact of exhaust emissions from vehicles using reformulated gasolines and flexible/variable-fuel vehicles using M85 has been studied using emissions data from the Auto/Oil Air Quality Improvement Research Program and a single-cell trajectory air quality model with two different chemical mechanisms (the updated version of Carbon-Bond-IV (CB4) and the LCC mechanism). Peak ozone concentrations are predicted for each fuel for all combinations of the following ambient conditions: low and high atmospheric dilution or mixing height, four NMOG/NOx ratios, two each of the initial NMOG concentration, the vehicular contribution to the ambient air, and the NMOG composition of the initial ambient mixture. The ozone impact of a fuel depends strongly on the atmospheric dilution and NMOG/NO~ ratio of an area. The differences in ozone impact among fuels are limited under the condition of high atmospheric dilution and a high NMOG/NOx ratio. The ozone-forming potentials (OFPs) for the exhaust emissions based on the maximum incremental reactivities (MIRs) for various fuels are generally well correlated with model-calculated peak ozone levels at a low NMOG/NOx ratio. These OFPs can serve to separate out fuels with rather different reactivities, but not fuels with comparable reactivities. Model-calculated ozone levels for various fuels based on CB4 and LCC mechanisms are relatively well correlated at low NMOG/NOx ratios, but much less so at higher ratios. Fuels with a high aromatic content, including high-toluene fuels, tend to be ranked more favorably by CB4 than by LCC. On the other hand, M85 is ranked more favorably by LCC than by CB4. Fuels with a low 90% boiling point and a low content of aromatics and olefins are generally less reactive. M85 would be an attractive fuel if the formaldehyde emissions could be curtailed significantly. Key word index: Reformulated gasolines, alternative fuels, air quality model, trajectory box model, chemical mechanisms.
1. INTRODUCTION In an attempt to improve the ozone air quality in urban areas, the U.S. Federal and some State governments are mandating the use of reformulated gasolines by gasoline vehicles and encouraging the use of alternatively fueled vehicles. A wealth of data has been collected by the A u t o / O i l Air Quality Improvement Research Program (AQIRP) to establish a link between vehicle emissions and properties of reformulated gasolines and alternative fuels. It has long been recognized that individual or mixtures of volatile organic compounds (VOCs) may differ significantly in their effectiveness in forming ozone in the atmosphere. Accordingly, fuels of different compositions may lead to emissions that have different reactivities. In 1991, the California Air Resources Board (CARB, 1991) adopted regulations for vehicular V O C exhaust emissions that allow the use of reactivity-adjustment factors (RAF) determined under a fixed set of environmental conditions and based on a maximum-incremental-reactivity (MIR) scale deAE 28:17-B
veloped by Carter (1990a). Reactivity scales and ozone-forming potentials (OFPs) of emissions, defined as the amount of additional ozone formed due to emissions from unit distance of travel, are generally sensitive to environmental conditions (Carter and Atkinson, 1989; Chang and Rudy, 1991). Therefore, it is important to evaluate the relationship between the reactivity-based O F P s for the different fuels and the corresponding measured or model-predicted ozone levels under a wide range of environmental conditions. It is also important to see how the ozone air quality impact of the fuels changes with respect to these environmental conditions. Even under identical environmental conditions, the model-predicted ozone levels depend on the chemical mechanisms used. Therefore, it is worthwhile to see how uncertainty in the chemistry influences the prediction of the ozone air quality impact of the fuels. This paper studies the impact of ozone air quality, in terms of the reactivity-based O F P s and the predicted ozone levels, of exhaust emissions from vehicles using reformulated gasolines and a methanol fuel. The
2777
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D.P. CHOCK et al.
model predictions are m a d e u n d e r various emission a n d a m b i e n t conditions a n d are based o n two different chemical mechanisms. The air quality model used is a simple trajectory model. A l t h o u g h the trajectory model m a y n o t be a d e q u a t e for describing the ozone air quality in a given u r b a n area, it is useful for studying the relative impact of usage of different fuelvehicle systems on ozone f o r m a t i o n u n d e r a wide range of conditions. The emissions data used are from P h a s e I of the A u t o / O i l A Q I R P . The A Q I R P also calculates reactivities of emissions ( H o c h h a u s e r et al., 1992; Schleyer et al., 1992; A Q I R P , 1993) a n d conducts modeling studies of the air quality impact of reformulated gasolines a n d m e t h a n o l fuels in three u r b a n areas (AQIRP, 1991, 1992b; D u n k e r et al., 1992; C h o c k et al., 1993). The present study, o n the other h a n d , uses simplified a s s u m p t i o n s a n d a simple model for the purpose of uncovering the general response of the fuel-vehicle impact on ozone to changes in emissions, e n v i r o n m e n t a l conditions a n d chemical mechanism.
2. M E T H O D O L O G Y
2.1. Fuels and vehicular emissions
The emissions data are from the current-fleet gasoline vehicles (1989 models) using the AQIRP Phase I reformulated gasolines, and prototype flexible-fuel or variable-fuel vehicles (designated FFV/VFVs) using methanol fuels. The
Phase-I gasoline fuels are characterized by four fuel parameters: aromatic content, methyl-tertiary-butyl ether (MTBE) content, olefin content, and the 90% boiling point (T90). Target values (high and low) for each parameter are given in Table 1. (Actual values for each parameter of fuels deviate somewhat from target values.) The fuels are identified by the alphabetic code given in Table 2. Two additional gasoline fuels are also included in the fuel matrix: fuel A or the industry-average (Ind Avg) fuel (32 vol% aromatics, no MTBE, 9 vol% olefins, a T90 of 165°C (330°F), and fuel B, the certification (CERT) fuel, or indolene (30 vol% aromatics, no MTBE, 4.6 vol% olefins, a T90 of 154°C (309°F). The methanol fuel used is M85 (fuel Z), which is a mixture of 85 vol% methanol and 15 vol% lnd Avg fuel. The target Reid vapor pressure (RVP) of all fuels is 60 kPa (8.7 psi). In order to satisfy all the constraints of the RVP and target values of the fuel parameters, the fuel compositions must be altered as necessary. For example, n-butane was added to the gasoline component to adjust the RVP of M85 to 60 kPa. The AQIRP shows that reducing the RVP of the nonoxygenated fuels reduces the diurnal evaporative emissions significantly (AQIRP, 1992a). All gasolines except fuel B have a sulfur content of about 300 ppm by weight. Fuel B has 120 ppm of sulfur. The AQIRP shows that reducing the fuel sulfur content reduces the (post-catalyst) exhaust NMOG, CO and NO x emissions significantly (AQIRP, 1992c). The AQIRP current-fleet gasoline vehicles consist of 16 cars and four light-duty trucks with an odometer range of 16,000-47,000km (10,000-29,000miles). The FFV/VFV fleet consists of 19 (17 cars and 2 trucks) pre-1990 test prototypes with an odometer range of 14,000-53,000 km (9000-33,000 miles). These are not production vehicles. So, comparing the relative ozone impact due to gasoline and methanol fuels based on the present AQIRP data may not be appropriate. However, it is included here to see how the ranking of M85, in terms of ozone air quality benefit, is affected by the use of different chemical mechanisms.
Table 1. Fuel design parameters for the auto/oil gasolines Target value Variable Aromatics (vol%) MTBE (vol%) Olefin (vol%) T90 (°C)
Low
High
Industryaverage
Cert. fuel
20 0 5 138-149
45 15 20 177-182
32 0 12 168
30 0 5 154
Table 2. Auto/oil Phase I reformulated gasolines and methanol fuels and their average measured NMOG exhaust emissions for cars Code A B C D E F G H I J
ID
ExhEM (gkm -1)
Code
0.088 0.079 0.082 0.086 0.106 0.086 0.082 0.075 0.123 0.096
K L M N O P Q R
Amot AmOT aMOT aMot AMOt amOt amoT AMoT
0.086 0.116 0.083 0.076 0.078 0.080 0.102 0.119
Z
M85
0.182
Ind Avg CERT AMot amOT AMOT amot AmOt aMOt AmoT aMoT A/a--high/low aromatics. M/m--high/low MTBE. O/o--high/low olefins. T/t--high/low T90.
ID
ExhEm(gkm ~)
Urban ozone air quality impact The exhaust emissions are based on the U.S. Federal Test Procedure (FTP) which contains three driving phases: cold start or bag 1; hot stabilized or bag 2; and hot start or bag 3. Evaporative and running-loss emission measurements are available for 13 of the 18 gasoline fuels; and the measurements have much greater uncertainties than those for exhaust emissions. In this study, differences in the ozone impact among the gasoline fuels will be based on the differences in exhaust emissions. This assumption was made so that the calculated ozone impact of the fuels can be compared directly with the OFPs which are defined for exhaust emissions. It also allows the inclusion of all gasoline fuels in the comparison, not just those with evaporative and running-loss emissions. The implication of ignoring the differences in evaporative and running-loss emissions among the gasoline fuels will be discussed later. All fuels except M85 are assumed to have evaporative and running-loss emissions identical to those for fuel A. For M85, the measured amounts and compositions of evaporative and running-loss emissions will be used. Use of separate evaporative and running-loss emissions for M85 in the model implies that the model result should not be compared directly with the exhaust-based OFP for M85. Based on the U.S. Environmental Protection Agency's (EPA) MOBILE4 estimates, it is further assumed that the running-loss emissions are 1.8 and 1 times the evaporative emissions for cars and trucks, respectively. For M85, the measured NMOG mass exhaust emissions from the prototype FFV/VFVs are about twice those for fuel A. Since the molecular weight of methanol (32), a major component of the M85 exhaust, is more than twice the molecular weight per carbon (13.88) of the gasoline exhaust, we shall assume that the exhaust mole-carbon (mole-C) emissions for M85 are the same as those for fuel A. The corresponding NMOG exhaust emission is slightly greater than measured. For simplicity, it is also assumed that the NOx and CO emissions for fuel A are applicable to those for other fuels in our model simulations since NOx emissions do not vary as greatly across the fuels as NMOG emissions and the sensitivity of ozone impact to the variation of CO emissions across the different fuels is small. Approximately 20% of all tests with the gasoline fuels and more than 50% of all tests with the methanol fuels have speciated exhaust emissions data which are given in weight percents. In the speciation tests, up to 151 organic species were identified. All tests, however, have data for total nonmethane hydrocarbon mass exhaust emission rates (g km - 1) for the three FTP bags. All evaporative and running-loss tests contain speciated measurements. Their composition measurements for fuel A and for M85 are used in the present study. To estimate the NMOG emission composition, the unidentified species (generally less than 5% by weight) are assumed to have the same composition as the identified nonmethane species (excluding formaldehyde and methanol in the case of methanol fuels). An averaging procedure was used for both the weight percent of each species and the total NMOG mass emission rate for each of the emission phases (bags 1, 2, 3, hot-soak evaporative, diurnal evaporative and running loss). The averaging procedure for the gasoline fuels had three steps: first, averaging over all the tests on a vehicle for each fuel; next, averaging over vehicles of the same model; finally averaging over all models for cars and all models for trucks. For the methanol fuels, straight averaging was used for cars and trucks separately. The averaging procedure above is used to determine the speciated average weight percents from the composition data base and the total NMOG mass emission rates from the full data base for each emission phase and each fuel. Then the average speciated mass emission rates are calculated by multiplying the average weight percents by the average total NMOG mass emission rates for each phase and each fuel. Two of the gasoline cars used the carbureted technology. Their emissions data are not included in the averaging procedure. For each fuel, the average total NMOG mass exhaust
2779
emissions and the average weight percents of the calibrated nonmethane organic species for the three FTP bags are combined according to the FTP weightings for the three bags. The evaporative and running-loss emissions are also converted to the same units as the exhaust emissions following the MOBILE4 prescription (USEPA, 1989). The average measured NMOG exhaust emissions for cars for all fuels are shown in Table 2. These numbers are similar but not identical to those used in the AQIRP because of differences in assumptions. The converted evaporative emissions for cars are 0.019 gkm I for fuel A and 0.051 gkm -1 for M85. The average measured NOx emission for fuel A is 0.342 g k m - 1; the variation from fuel to fuel is typically 10% or less. It is known that the cold-start (bag 1) exhaust NMOG emissions from current gasoline vehicles dominate the total NMOG emissions and contain a significant amount of unburned fuel components and lower alkenes generated by incomplete combustion. Consequently, the cold-start mode carries the fuel signature into the exhaust emissions and dominates the exhaust mass and reactivity. For M85, the NMOG reactivity is influenced strongly by the formaldehyde emission rate. 2.2. Air quality modelino and simulation conditions The Ozone Isopleth Plotting Package with Optional Mechanism (OZIPM) (Hogo and Gery, 1988) was used to study the air quality impact of the various fuels under various emission and ambient conditions. The two chemical mechanisms used are the Lurmann-Carter-Coyner (LCC) (Lurmann et al., 1987) and the Carbon-Bond IV (CB4) (Gery et al., 1989) mechanisms. The LCC mechanism is an older version of the Statewide Air Pollution Research Center mechanism, the most recent version of which is the SAPRC90 mechanism (Carter, 1990b). The latter contains some rate constant updates and is more detailed; but the overall characteristics of both versions are quite similar. The CB4 mechanism used in the present paper is the recently updated (USEPA, 1993) version which is occasionally referred to as CBM4.1. The simulation conditions are summarized in Table 3. The simulations were started with a given initial concentration of NMOG in the central part of an urban area at 0800 LDT. Two initial NMOG concentrations were assumed: 0.4 and 0.7 ppmC (referred to as "c" and "C", not to be confused with fuel C). They represent the typical range of 6 9 a.m. NMOG concentrations observed in an urban area. The initial NOx concentrations were varied depending on the NMOG/NOx ratio used. The initial NMOG/NOx ratios considered are 6, 8, 10 and 12 ppmC ppm-~ (denoted "R6," "R8," etc.), a range covering most urban areas. NMOG and NOx are emitted to the air parcel as a fraction of the initial concentrations, as shown in Table 3. Pollutants are entrained from aloft as the mixing height changes through the day. Two dilution scenarios were chosen: "low dilution" (the mixing height rising from 300 m at 0800 LDT to 600 m by mid-afternoon; denoted "LD') and "high dilution" (the mixing height rising from 300 to 1500 m, denoted "HD'). The former tends to occur in an areal band downwind of and parallel to a large body of water while the latter is quite typical of inland regions. The ambient NMOG compositions used in the present paper are shown in Table 3: U.S. "urban average" based on 41 cities (denoted "EPA41"), given by Jeffries et al. (1989), and that measured during the Southern California Air Quality Study (SCAQS) in 1987 (Lurmann et al., 1992). The "urban average" NMOG composition database did not have measured aldehydes and the assumed formaldehyde concentrations are much higher than the measured concentrations from SCAQS (see Table 3). To represent the base case for fuel-vehicle emission impact on ozone formation, a certain mole-C fraction (0.15 or 0.30, denoted "v" and "V", respectively) of the initial NMOG mixture was replaced by the composition of the emissions for
2780
D. P. CHOCKet al. Table 3. Simulation conditions Latitude: Temperature: Simulation period:
34.1°N 303 K 0800 to 2100 LDT
Initial concentrations: NMOG NOx (NOz/NO x = 0.25) CO CH,~
Either 0.4 or 0.7 ppmC Variable depending on NMOG/NOx ratio 1.2 ppm 2.0 ppm
Emission schedule
Aloft concentrations: NMOG 03 Mixing height: t Low dilution case High dilution case NMOG composition:$ Mechanism CB4 Mech.
LCC Mech.
Time (LDT) 0800-1300 1300-1800 0.04 ppmC 0.06 ppm
NMOG 0.15 0.05 CO CH 4
Initial at 0800 L D T 300 m 300 m Species PAR ETH OLE TOL XYL HCHO ALD2 ISOP NR
Date: Relative humidity:
21 June 50%
NOx 0.25 0.08 0.5 ppm 1.6 ppm Final at 1500 L D T 600 m 1500 m
Ambient carbon fractions EPA41 SCAQS 0.564 0.550 0.037 0.042 0.035 0.023 0.089 0.096 0.117 0.104 0.021 0.010 0.052 0.035 0.000 0.003 0.084 0.137
ALK4 ALK7 ETHE PRPE TBUT TOLU XYLE TMBZ HCHO ALD2 ISOP ACET MEK NR
CO 0.10 0.03
0.214 0.280 0.037 0.048 0.029 0.080 0.079 0.042 0.020 0.030 0.000 0.000 0.000 0.141
0.235 0.272 0.042 0.033 0.012 0.096 0.088 0.018 0.010 0.033 0.003 0.024 0.010 0.124
Aloft carbon fractions 0.498 0.034 0.020 0.042 0.026 0.070 0.037 0.000 0.273 0.363 0.151 0.034 0.010 0.014 0.042 0.024 0.000 0.070 0.037 0.000 0.000 0.000 0.255
* Fraction of initial concentration emitted per hour (Hogo and Gery, 1988). t Mixing height varies with time according to the characteristic curve recommended by Hogo and Gery (1988). ~ Aloft and ambient EPA41 compositions are from Jeffries et al. (1989). Ambient SCAQS composition comes from Lurmann et al. (1992). Abbreviations for species are defined in Jeffries et al. (1989).
fuel A. For a different gasoline fuel, the exhaust portion of emissions for fuel A was replaced by the mole-C/kin ratio of that fuel to fuel A; the resulting fraction accommodates the exhaust NMOG composition of that fuel. Consequently, the sum of vehicular and ambient mole-C fractions can differ from unity. As stated earlier, the exhaust mole-C fraction of M85 is assumed to be the same as fuel A, but the evaporative and running-loss emission fractions were modified by the mole-C/km ratios relative to the respective emissions for fuel A, as described above. The mole-C percents for NMOG emissions from cars and trucks are taken to be 80/20. For fuel A emissions, this split corresponds roughly to a mileagecontribution split of 90/10. One should note that the composition of the ambient mixture itself already reflects a significant contribution from vehicular emissions. Therefore, our partial replacement of the mixture with the composition and relative amount for a fuel
does not adequately describe the true vehicular contribution to the ambient mixture. However, the relative impact of the fuel will be manifested. Furthermore, the use of two vehicular mole-C fractions allows one to appreciate the sensitivity of this impact to the change in the vehicular contribution to the ambient NMOG. On the whole, we have two alternatives for each of the following conditions: atmospheric dilution, initial NMOG concentrations, vehicular fractions in these concentrations, ambient mixture compositions, and chemical mechanisms. In addition, we have four NMOG/NO x ratios. So we have a total of 128 test cases per fuel. 2.3. Ozone forming potential The OFP of the exhaust NMOG emissions for each fuel is calculated based on the M1R and the maximum ozone
Urban ozone air quality impact incremental reactivity (MOR) scales. MIRs have been determined at relatively low NMOG/NOx ratios where the incremental reactivities of atmospheric NMOG mixtures are at their maximum values, and MORs have been determined at relatively high NMOG/NO= ratios where the daily peak (1 h average) ozone levels are at their maximum values. The OFP (in terms of g 0 3 per kilometer of travel) for each fuel is determined by the sum of products of each NMOG species emission rate (g km ~) and its incremental reactivity (gO3/gNMOG, for either MIR or MOR). Because the emission differences among the gasoline fuels in this study is characterized by the differences in exhaust emissions alone, the calculated OFPs for gasoline fuels can be compared directly with the ozone impact of the fuels estimated by the trajectory model.
3. RESULTS 3.1. Comparison of peak ozone concentrations The ambient conditions, including the atmospheric dilution, the concentration and composition of N M O G in the urban mixture, the N M O G / N O ~ ratio,
0.17-
a: LD
2781
all play a critical role in determining the ozone level. These ambient conditions influence not only the ozone air quality impact from the different fuels, but also the relative ranking of these fuels based on the impact. Figure 1 shows the ozone concentrations for the different fuels under high and low dilution (HD and LD), N M O G / N O x ratios of 6 and 10 p p m C / p p m (R6 and R10, respectively), the two ambient N M O G compositions for urban-area mixtures (SCAQS by upper-case letters and EPA4! by lower-case letters), and the two chemical mechanisms (LCC on the ordinate and CB4 on the abscissa). The initial N M O G concentration and vehicle fraction are held constant at 0.7 ppmC and 0.30, respectively. The ozone levels based on the EPA41 mixture are considerably higher than those based on the SCAQS mixture. This is because EPA41 is a more reactive mixture than SCAQS. It has twice the mole-C fraction of formaldehyde (0.02) as the SCAQS mixture (0.01). The EPA41 mixture also contains higher mole-C fractions of
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Fig. 1. Comparison of peak ozone concentrations of all fuels based on two mechanisms and two urban-mixture compositions (upper-case letters for SCAQS and lower-case letters for EPA41) using an initial NMOG concentration of 0.7 ppmC and a vehicular mole-C fraction of 0.30. (a) LD and R6; (b) HD and R6; (c) LD and R10; (d) HD and R10.
2782
D.P. CHOCKet al.
higher olefins and polyalkylated aromatics which are highly reactive, and a lower fraction of toluene which is not very reactive. Despite the significant differences in the N M O G compositions of the two ambient mixtures, the patterns of the fuels' ozone levels are quite similar. Henceforth, we shall discuss primarily the results based on the SCAQS mixture unless otherwise indicated. Figure 1 reveals the following. First, high dilution does not necessarily imply a lower ozone concentration. In fact, the concentration can be higher than the low-dilution case under a low but identical N M O G / NO~ ratio. This is because the NOx-inhibition effect is most effective at a low N M O G / N O x ratio, and diluting NOx reduces its effectiveness in reducing ozone. Second, the range of the ozone concentrations for the fuels considered is wider for the lower N M O G / N O x ratio than for the higher ratio. This is consistent with the fact that the lower ratio is closer to the region where the MIR applies, so that the sensitivity of the ozone level to the fuel is at or near a maximum. The characteristics of the ozone concentration range will be further discussed below. Third, the correlation between the ozone concentrations predicted by LCC and CB4 is very good for the low NMOG/NOx ratio bt, t deteriorates for the high ratio. At a high NMOG/NOx ratio, the "At" fuels (fuels C, K, O and G, all with a high aromatic content ("A") and a low T90 Ct'). See Table 2) all yield essentially the lowest ozone concentrations compared to other fuels according to CB4, but not the lowest according to LCC. This is because these four fuels have very high toluene contents (36%, 35, 31 and 29% by weight for fuels C, K, O and G, respectively) and toluene is not efficient in ozone production at a high NMOG/NO~ ratio, but is even less efficient according to CB4 (by scavenging NOx, leading to a'negative MOR) than according to LCC. The "AT" fuels (fuels I, R, L and E) are the most reactive fuels. They all have a high aromatic content but with a large proportion of the highly reactive polyalkylated aromatics. Both mechanisms rank them as the most reactive fuels at a low NMOG/NOx ratio. But at a high N M O G / N O x ratio, CB4 again finds these fuels to be not as reactive as some other fuels with a low aromatic content while LCC continues to rank them as the most reactive. The difference in the treatment of the "At" and "AT" fuels results from a difference in aromatic chemistry assumed in both mechanisms and contribute to a significant reduction in the correlation between the predicted ozone concentrations from both mechanisms. It should also be noted that fuel Z is considered more reactive in CB4 than in LCC because the key reactive emission from fuel Z is formaldehyde, and CB4 has a more reactive formaldehyde chemistry (partly due to a higher photolysis rate) than LCC. Figure 2 shows the ozone concentrations for selected fuels as a function of the NMOG/NO~ ratio, the dilution, the initial N M O G concentration, the mole-C fraction of N M O G attributed to vehicles, and the
chemical mechanism. The fuels selected are C, E and Z. Fuel C is one of the least reactive fuels. Fuel E is one of the most reactive, and fuel Z or M85 is the only nongasoline fuel considered here. Under low dilution, the ozone levels tend to peak at an N M O G / N O x ratio of 10 ppmC p p m - ~ or higher. As the ratio decreases, the ozone levels drop rather sharply. The ratio where the ozone concentrations spread the most is around 8 ppmCppm -1. Also, increasing the initial N M O G concentration while keeping the NMOG/NOx ratio the same (by increasing the initial NOx concentration as well) does not change the ozone concentration significantly at a low ratio, but increases the ozone concentration significantly at a higher ratio (10 ppmC p p m - ~or more). This agrees with the expectation that NO x is more effective in inhibiting ozone at a low ratio than at a high ratio. Increasing the fraction of vehicle contribution in the ambient mixture tends to lower the ozone concentration for the less reactive fuels, but increases the ozone concentration, especially at a ratio of around 8 ppmC ppm - 1, for the more reactive fuels. This indicates that the reactive fuels such as fuel E is more reactive than the SCAQS urban mixture while fuel C, and to a lesser degree, fuel Z, are less reactive than the mixture. Under high dilution, the variation of ozone concentrations over an NMOG/NOx ratio of 6 to 10 ppmC p p m - I is more restricted. In contrast to the lowdilution case, the ozone concentrations for the different fuels have a rather limited spread, the greatest spread being at a ratio of about 6 ppmC p p m - 1. This spread is still substantially less than the largest spread in the low dilution case. Thus, in high dilution, the difference in the ozone impact of different fuels is not large except at a low N M O G / N O x ratio. The highest ozone levels tend to occur at an N M O G / N O x ratio of around 8 ppmC p p m - 1. Thus, in principle, a different atmospheric dilution rate requires a different N M O G / N O x ratio on which the MIR or MOR should be defined. For a given initial N M O G concentration, increasing the vehicle contribution from 0.15 mole-C fraction to 0.30 mole-C fraction does not change the ozone concentration substantially. Increasing the initial N M O G concentration at a fixed NMOG/NOx ratio lowers the ozone concentration at a ratio of 6 ppmC ppm- 1 but increases it at a higher ratio. It appears then that the NOx-inhibition effect works better at high dilution than at low dilution. Figure 3 shows the predicted ozone concentrations for fuels C and E for the two dilution rates (HD and LD), two NMOG/NO~ ratios (R6 and R10), the two initial N M O G concentrations (c and C) and vehicular mole-C fractions (v and V), and the two chemical mechanisms (LCC and CB4), based on the SCAQS ambient N M O G composition. The discussions provided earlier are demonstrated vividly in the figure. It is also seen that under LD, the ozone concentration predicted by CB4 has a greater sensitivity to the NMOG/NOx ratio than that predicted by LCC.
Urban ozone air quality impact 0.25
2783
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concentrations at an N M O G / N O x ratio of 6 and 10 p p m C p p m - 1 (R6 and R10). The M I R - and M O R based O F P s are used for R6 and R10, respectively. Both low and high dilution cases are considered, as are the SCAQS and EPA41 base mixture compositions.
2784
D . P . CHOCK et al.
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.
0.00
LCC [] CB4 Fig. 3. Peak ozone concentrations for fuels C and E as a function of dilution rate (HD and LD), NMOG/NO x ratio (R6 and R10), initial NMOG concentration ("c" and "C") and vehicular mole-C fraction ("v" and "V"). The initial N M O G concentration is assumed to be 0.7 ppmC and the vehicle contribution is 0.30 in mole-C fraction. The correlation between the predicted ozone concentrations and the MIR-based OFPs is generally good for both the SCAQS and the EPA41 mixtures for R6 (and, actually, for R10 as well) and for both LD and HD. But if we are to expand the OFP scale that covers the low-T90 fuels alone, then the correlation is poor. Here, the impact of ambient N M O G composition also becomes evident. The correlation between the predicted ozone concentrations for R10 and the MOR-based OFPs is only fair at best. Note that since the OFPs are defined for exhaust emissions alone, drawing conclusions for fuel Z may not be appropriate since its evaporative and running-loss emissions have been modified for model calculation. Since the OFPs are derived using a detailed SAPRC mechanism which is more closely related to LCC than to CB4, the correlation between the CB4-predicted
The modeling results show that the "T" (high T90) fuels are more reactive than the "t" fuels, while the relative ozone impacts of the "t" fuels are less than that of fuel A and are less distinct from each other. The "AT" fuels are the most reactive of all gasoline fuels tested, primarily because these fuels lead to the highest exhaust N M O G emission rate and the emissions tend to be highly reactive due to the significant amount of polyalkylated aromatics present. These fuels are also more reactive than the urban mixture so that with a constant N M O G concentration, the greater the vehicular contribution, the higher the ozone level. The ranking by the two mechanisms of these four fuels relative to other "T" fuels can be different at a high NMOG/NOx ratio because of the difference in the aromatic chemistry assumed in both mechanisms. The "t" or low-T90 fuels and fuel B are much less reactive. Among them, the "At" fuels are considered the least reactive by CB4 but not necessarily by LCC especially at a high N M O G / N O x ratio. This is the result of the difference in the toluene chemistry assumed in the mechanisms. According to LCC, fuels H and P are also among the least reactive fuels. Note that fuels G, O, P and H have a high olefin content. They have a somewhat lower exhaust N M O G emission rate, but the emission tends to have higher reactivity. High-olefin fuels also have higher and more reactive evaporative emissions. If evaporative emissions of the individual fuels were included in the simulation, these fuels would lead to higher ozone concentrations as was shown by AQIRP (1991) in the three-city study. The role of MTBE on ozone air quality is small in comparison to the other fuel factors. Fuel Z (M85) produces less ozone than does fuel A at a low N M O G / N O x ratio. However, as the ratio increases, fuel Z produces a comparable amount of ozone as fuel A according to CB4 but still less ozone than does fuel A according to LCC. The difference in the formaldehyde chemistry of the two mechanisms must be reconciled before a more definitive statement on fuel Z can be made. Fuel N and Z have the lowest MIR-based OFPs. But they do not yield the lowest ozone concentrations
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0.60
0.65
MOR-based OFP (g Oa/km)
Fig. 4. Comparison of LCC-derived peak ozone concentrations based on the two urban mixtures (upper-case letters for SCAQS and lower-case letters for EPA41) for all fuels and the MIR-(for R6 only) and MOR-(for R10 only) based OFPs. The initial NMOG concentration is 0.7 ppmC and the vehicularmole-C fraction is 0.30. (a) LD and R6; (b) HD and R6; (c) LD and RI0; (d) HD and R10.
in the simulations. In general, however, the MIRbased OFPs of the "t" fuels are quite similar as are their predicted ozone concentrations at a low N M O G / N O x ratio. If we exclude high-olefin fuels that have reactive evaporative emissions, then fuels C, K, B, F and N yield the lowest ozone concentrations. Ranking among these fuels is difficult not only because the differences among their ozone impacts are small to begin with, but also because the different assumptions in the two chemical mechanisms produces a different ranking, and the ranking changes as the atmospheric dilution and N M O G / N O x ratio varies. Fuel Z yields a relatively high ozone impact primarily because of its relatively high formaldehyde emissions (0.012 g km - 1) from the prototype FFV/VFVs. Reduction of the formaldehyde emission would significantly lower the ozone air quality impact of this fuel. 4. C O N C L U S I O N S
The present study shows that the extent of atmospheric dilution and the NMOG/NOx ratio play a AE 28:17-C
critical role in determining the ozone impact of fuels. Varying the dilution leads to very different behavior of peak ozone concentrations. In particular, the relation between peak ozone and the N M O G / N O x ratio is altered significantly. For example, at an NMOG/NOx ratio of 6 ppmC ppm- 1, the peak ozone concentration under high dilution can actually be higher than under low dilution and decreases with increasing initial N M O G (and NOx) concentration due to a more efficient NOx-inhibition effect. Added to this complexity is uncertainty in the chemistry so that different mechanisms predict the peak ozone concentrations somewhat differently. Consequently, both consistent patterns and unresolved issues emerge in the present study. Here are the consistent patterns. The MIR-based OFPs for various fuels are generally well correlated with model-calculated peak ozone levels at a low NMOG/NOx ratio. The correlation is even reasonable for a ratio of 10 ppmC ppm- ~ if the LCC mechanism is used to calculate the peak ozone concentrations. However, while all fuels with a low T90 have relatively low OFPs for exhaust emissions and yield
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relatively low peak ozone levels, there is little correlation between these fuels' O F P s and their corresponding ozone concentrations. Thus, the MIR-based O F P s can serve to separate out fuels with rather different reactivities, but not fuels with comparable reactivities. The O F P s cannot be used to estimate the change in ozone concentration because such change depends strongly on the atmospheric dilution, the N M O G / N O x ratio, and the N M O G concentration. Modelcalculated ozone levels for various fuels based on CB4 and LCC mechanisms are relatively well correlated at low N M O G / N O ~ ratios, while the correlation becomes poorer at higher ratios. The unresolved issues in this study concern the uncertainty in the chemistry of toluene and aromatics in general, and the free-radical initiation process in the chemical mechanisms. Thus, CB4 has a tendency to rank fuels with a high aromatic content less harshly and high-toluene fuels more favorably than LCC at high N M O G / N O x ratios. It also tends to rank fuel Z less favorably than LCC. If we exclude fuels with a high olefin content because of their tendency to have high and reactive evaporative e m i s s i o n s - - a factor not considered in the present study, then fuels with a minimal ozone impact are fuels C, B, N, K and F. If we further exclude the nonoxygenated fuels, then only fuels C and N are left. If we do not want to take advantage of the low reactivity of toluene and simply exclude fuels with a high aromatic content, then fuel N is the only reasonable choice in the gasoline fuel matrix for contributing the least to ozone formation from vehicle emissions. Fuel Z (M85) may still be a very attractive fuel provided that the formaldehyde emissions can be reduced significantly. In terms of actual ozone impact, the advantage of using a fuel with low reactivity depends on, among many things, the typical atmospheric dilution and N M O G / N O x ratio for the area. For example, the impact would be quite limited if the prevalent conditions entail high atmospheric dilution and a high N M O G / N O x ratio. REFERENCES
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