Environ. Sci. Technol. 1987, 21, 670-679
American Concrete Institute: Detroit, MI, 1983; pp 1-46. Sersale,R. In Proceedings: Seventh International Congress on the Chemistry of Cement; Paris, France, 1980; paper IV-1/ 3. Karlsson, H. T.; Klingspor, J.; Linne, M.; Bjerle, I. J. Air Pollut. Control Assoc. 1983, 33, 23-28. Ruiz-Alsop, R. N.; Rochelle, G. T. In Fossil Fuels Utilization-Environmental Concerns; Markuszewski, R., Blaustein, B. D., Ed.; ACS Symposium Series 319; American
Chemical Society: Washington, DC, 1986; pp 208-222. Jorgensen, C.; Chang, J. C. S.; Brna, T. G. Environ. Prog. 1987, 6(2),26-32.
Yoon, H.; Stouffer, M. R.; Rosenhoover, W. A.; Statnick, R. M. In Proceedings: Second Annual Pittsburgh Coal Conference; Pittsburgh, PA, 1985; pp 223-242.
(21) Fink, C. E.; McCoy, D. C.; Statnick, R. M. “Flue Gas Humidification with Boiler Limestone Injection for Improved ESP Performance and Increased SO2Removal”;presented at Coal Technology ’85 Conference, Pittsburgh,PA, 1985. (22) Yoon, H.; Ring, P. A.; Burke, F. P. “CoolsideSOzAbatement Technology: 1 MW Field Tests”; presented at Coal Technology ’85 Conference, Pittsburgh, PA, 1985. (23) Bergeson, K. L.; Pitt, J. M.; Demirel, T. Transp. Res. Rec. 1984, 998, 41-46.
Received for review August 8,1986. Accepted February 24,1987. This paper has been reviewed in accordance with the US.Environmental Protection Agency’s peer and administrative review policies and approved for presentation and publication.
Experimental Study of Incremental Hydrocarbon Reactivity Wllllam P. I.. Carter* and Roger Atklnson
Statewide Air Pollution Research Center, Universlty of California, Rlverside, California 9252 1 A series of environmental chamber experiments have been carried out to investigate the incremental reactivities of selected organics with respect to ozone formation in simulated photochemical smog systems. Varying amounts of a test organic were added to or subtracted from a standard four-hydrocarbon “minisurrogate”-NO,-air mixture to determine, as a function of irradiation time, the resulting changes in the amount of ozone formed and NO consumed, relative to the amount of the organic added. The incremental reactivities of toluene, trans-2-butene, and propene decreased significantly with reaction time, with toluene ultimately becoming negatively reactive; n-butane, ethanol, and tert-butyl methyl ether were always positively reactive; and benzaldehyde was always negatively reactive. The results are reasonably consistent with computer model simulations and indicate that the effect of regulating emissions of an organic on ambient ozone will depend not only on the organics reaction rate but also on its reaction mechanism and the conditions under which it is emitted. Introduction When emitted into the atmosphere in the presence of oxides of nitrogen (NO,) and sunlight, most organic compounds contribute to the photochemical formation of ozone, an important air-quality problem in many urban areas. It is well recognized that organic compounds differ significantly in how rapidly they react in the atmosphere and the extent to which their reactions promote or inhibit ozone formation (1-5). However, the definition of the “reactivity” of an organic compound and the most appropriate technique($ by which it is determined have been a subject of debate ( 4 4 , and alternative criteria can yield significantly different rankings of reactivity (1). Many previously proposed reactivity scales (1, S 1 3 ) were based on data from NO,-air irradiations of individual organic compounds in environmental chambers. Thus, for example, the reactivity of an organic has been defined in terms of ozone yields, ozone dosages, NO to NOz conversion rates, consumption rates of the organic, and eye irritation observed in such experiments (1). However, there are problems associated with basing reactivity scales on NO,-air irradiations of single organics, since (a) organics are almost never emitted into the atmosphere in the absence of other reactive compounds (whose reactions affect the overall reaction processes occurring) and (b) chamber 670
Environ. Scl. Technol., Vol. 21, No. 7, 1987
effects are known to have significant effects on the results of such experiments (14),particularly if the organic reacts relatively slowly (14) or if its atmospheric reactions include significant radical removal processes. For such organics, chamber effects (14-16) can significantly affect the overall processes occurring in these irradiations. Several of the problems associated with defining and measuring reactivity on the basis of environmental chamber experiments can be circumvented by defining reactivity in terms of the rates at which the organic reacts with the hydroxyl (OH) radical (3). This criterion, which has the advantages of simplicity and of avoiding problems caused by chamber effects, is based on the recognition that most organics are consumed in the atmosphere primarily by reaction with OH radicals. However, certain classes of organic compounds are consumed in the atmosphere to a significant or even dominant extent by other processes, such as photolysis or reaction with O3 or NO3 radicals (17, 18). In addition, the different classes of organics exhibit many differences in their atmospheric reaction mechanisms (17), and these can significantly influence ozone formation. For example, certain organic compounds actually inhibit ozone formation due to the presence of radical and/or NO, removal processes in their chemical mechanisms (17,19,20), while others enhance ozone formation to a greater extent than expected solely from their reaction rates due to radical sources in their atmospheric reaction mechanisms. An alternative approach examined in this study is to define the reactivity of an organic in terms of the incremental effects caused by additions of small amounts of the organic to an irradiated “base-case” hydrocarbon surrogate-N0,-air mixture, with the base-case mixture being designed to represent current emissions into urban airsheds. This definition of reactivity takes into account the effects of all aspects of the organic reaction mechanism, not just its overall reaction rates with OH radicals and other reactive intermediates. Furthermore, incremental reactivity measurements are based on experiments designed to represent ambient air mixtures more closely than is possible when employing the organic alone, and meaningful reactivity measurements can be made for relatively unreactive compounds that are highly sensitive to chamber effects in single organic-NO,-air irradiations. Although experimental measurements of incremental reactivity also involve conducting environmental chamber experiments, chamber effects are relatively less important in influencing
0013-936X/87/0921-0670$01.50/0
0 1987 American Chemical Society
the results of these experiments since (a) the base-case surrogate-NO,-air mixture can be chosen to be sufficiently reactive so that the influence of chamber effects is minimized and (b) chamber effects tend to cancel out, since reactivity is defined in terms of differences in measured quantities under conditions where chamber effects are expected to be similar. The concept of incremental reactivity is also highly relevant from a regulatory viewpoint, since in most practical cases air-quality regulations and control strategies are concerned with effects of small, or incremental, changes in emissions. In this paper we describe an experimental investigation concerning the incremental reactivities of several organic compounds. The experiments carried out consisted of repeated irradiations of a standard minisurrogate organic-NO,-air mixture designed to represent current emissions of organics and NO, into polluted urban atmospheres (such as the California South Coast air basin), interdispersed with irradiations of the same surrogate-NO,-air mixture to which various levels of selected test organics were added or, for test organics that were also components of the minisurrogate mixture, deleted from the mixture. The organic compounds studied were toluene, benzaldehyde, propene, trans-2-butene, n-butane, ethanol, and tert-butyl methyl ether. These compounds have significant differences in their reaction rates and atmospheric reaction mechanisms (17,21),and the data presented indicate the degree of variability of incremental reactivities for typical classes of organic compounds emitted into urban airsheds. The effects of adding (or subtracting) the test organics to the base-case surrogate-NO,-air irradiations on the sum of the amount of ozone formed and the amount of NO consumed [A([O,] - [NO])] were measured at various times during the irradiations. As discussed below, defining reactivity in terms of changes in both ozone and NO is more directly related to the chemical processes that cause ozone formation in organic-NO,-air irradiations than are changes in ozone alone, and it can be used to give a meaningful assessment of reactivity under conditions when NO is present. The effects on ozone formation and NO consumption are examined as a function of irradiation time since the reactivities observed in the initial stages of the experiment reflect factors that influence the rates at which NO consumption and ozone formation occur, while the reactivities in the final stages of the run reflect factors that influence the ultimate ozone formation potential of the system, provided that the base-case surrogate-NO, experiment is sufficiently reactive so that the maximum ozone formation occurs by the end of the experiment (as was the case with the mixture employed in this study). Indeed, the results of this study show that certain compounds have significantly different effects on the final ozone yields than on the initial rates of ozone formation and NO consumption. To determine whether our current understanding of the chemistry of these irradiated organic-NO,-air mixtures is consistent with our experimentally determined incremental reactivity values and to evaluate the dependence of the incremental reactivity on the amount of test organic added, we also carried out computer model simulations of the experimental reactivity measurements. The results of these simulations are also presented in this paper for comparison with the experimental data, and a brief discussion of the mechanistic factors that account for the incremental reactivity trends observed is presented. Experimental Section Irradiations were carried out in a -6400-L indoor allTeflon chamber (ITC), which consisted of an FEP Teflon
(2-mm thickness) reaction bag fitted inside an aluminum frame. These irradiations were carried out at 50% relative humidity and an average temperature of 303 K. Irradiation was provided by two diametrically opposed banks of 40 Sylvania 40-W BL black lamps, backed by arrays of Alzak-coated reflectors. The light intensity was 70% of the maximum and was periodically monitored by measuring the rate of photolysis of NOz in Nz with the quartz-tube continuous-flow technique of Zafonte et al. (22). The light intensity declined gradually during the course of this study, with the NO2 photolysis rate k1 decreasing from 0.38 min-l for the initial experiments (which were carried out with a new set of lamps) to a constant average value of kl = 0.30 f 0.02 min-’ for irradiations subsequent to ITC-578. The chamber facility and experimental techniques employed and the results of relevant characterization experiments are given in detail elsewhere (23) and are only briefly summarized here. Prior to each experiment, the chamber was filled and emptied at least 3 times with purified air (24) and then filled with purified air at 50% relative humidity. All gaseous reactants were injected by gas-tight, all-glass gas syringes and were flushed into the chamber by a stream of ultrahigh purity N2. For liquid reactants, the desired amount of liquid was injected by a precision micropipet into a 1-L Pyrex bulb, and the contents were flushed into the chamber by a stream of ultrahigh purity Nz. After injection of the reactants, two preirradiation gas chromatographic (GC) analyses were carried out. Providing that no problems were apparent, a further t = 0 sample for GC analysis was taken, and the lights were turned on (t = 0). Samples for GC analyses were then taken throughout the irradiations, usually every hour. The analytical techniques used have been described in detail elsewhere (23,25). NO and NO, were monitored by chemiluminescence analyzers, and O3was monitored by ultraviolet absorption. The organics were quantitatively monitored by gas chromatography with flame ionization detection [or, for peroxyacetyl nitrate (PAN), with electron-capture detection] (23, 25). NO,-air irradiations, with trace ( 10 ppb) amounts of propene and n-butane added to serve as radical tracers, were periodically carried out to monitor the chamber radical source and levels of contamination by reactive organics. In these experiments, the relative rates of consumption of the radical tracers can be used to obtain an estimation of the “radical input rate” (1.9,which is defined as the input rate of hydroxyl radicals from chamber-dependent sources and which is a necessary input parameter to the computer model employed. The radical input rates measured in these runs ranged from 0.04 to 0.08 ppb m i d , in the range normal for this and other chambers constructed of Teflon film (15,261. The NO oxidation rates observed yield information concerning the degree of contamination of the chamber by reactive organics. In most cases the NO oxidation rates in the NO,-air irradiations carried out during this study ranged from 0.1 to 0.15 ppb min-l, showing that the NO oxidation rate caused by background reactive organics was minor compared to the oxidation rates observed in the reactivity assessment runs. The majority of the experiments consisted of alternate standard (base-case) minisurrogate-NO,-air irradiations and minisurrogate-NO,-air irradiations with varying amounts of test organics being added or (when the test organic was also a minisurrogate component) removed. The test organics studied in these experiments were toluene, benzaldehyde, ethanol, tert-butyl methyl ether (TBME), and the minisurrogate components n-butane, N
N
Environ. Sci. Technol., Vol. 21, No. 7, 1987
671
propene, and trans-2-butene. The standard minisurrogate-NO, mixture consisted of nominal initial concentrations of 4 ppm C hydrocarbon surrogate and 0.1 ppm NO,, with an initial NO/N02 ratio of 4. The conditions of the runs with the added (or reduced) test organic were the same as the standard runs, except for the levels of the test organic. Composition of the Base-Case Hydrocarbon Surrogate. The four-component hydrocarbon minisurrogate used in the reactivity experiments consisted of n-butane, propene, m-xylene, and trans-2-butene,with average initial concentrations corresponding to 72.9, 17.2,7.3, and 2.6 mol %, respectively. This mixture was chosen to be a simple representation of a 13-component surrogate developed to be representative, in term$ of the organics present and its photochemical reactivity, of emissions into the California South Coast air basin (27). The minisurrogate was designed to have the same carbon content and OH radical reactivity as the 13-component surrogate and to have approximately the same division, with respect to both carbon content and OH radical reactivity, between the alkanes, alkenes, aromatics, and aldehydes. [The 13-component surrogate also contained acetylene, but its contribution was neglected due to its low OH reactivity (21).] The aldehydes in the surrogate were represented by trans-Bbutene, which reacts rapidly under the conditions of these experiments to form acetaldehyde (17).[trans-2-Butene was used as the representative "aldehyde" since aldehydes are difficult to reproducibly inject into the chamber, and reproducibility of conditions was critical in these incremental reactivity experiments.] Computer model calculations showed that substitution of trans-2-butene for acetaldehyde caused a slight enhancement in the initial NO oxidation rate but did not significantly affect the final ozone yield. Computer Kinetic Model Employed. The chemical mechanism used in the computer simulations is described by Carter et al. (26) and thus is not discussed in any detail here. Briefly, this detailed mechanism consists of an update and an extension of those formulated by Atkinson et al. (28) and Lurmann et al. (29) and, with the exceptions of ethanol (discussed below) and tert-butyl methyl ether, includes the reactions of all of the compounds present in the minisurrogate and of the test organics used in this study. This mechanism has been comprehensively tested (26) against the results of over 400 environmental chamber experiments carried out at our laboratories and at the University of North Carolina. Ethanol is believed to react in the atmosphere primarily with OH radicals, via H-atom abstraction from the CYcarbon atom, to form acetaldehyde and H 0 2 (21): CHaCH2OH
+
OH
-
CH3;HOH
t H20
1
O2
CHsCHO t HO2
This mechanism was used with an OH radical reaction rate cm3molecule-l s-l (21). The constant of 5.27 X 10-12e-176/T atmospheric reaction mechanism of tert-butyl methyl ether (TBME) is not totally understood, and computer calculations were hence not carried out for this test organic. A detailed discussion of the methods used to represent chamber-dependent parameters is given by Carter et al. (26). Briefly, with the absorption coefficients and quantum yields for the photolysis reactions in the chemical mechanism, the rates of the photolysis reactions were calculated from the measured relative spectral distribution of the black lights employed in the ITC, normalized to the measured NO2 photolysis rates. The wall effects were 872
Environ. Sci. Technol., Vol. 21, No. 7, 1987
represented in the model by (26) (a) ozone wall decay at min-', (b) Nz05hydrolysis to HN03 at a rate of 1.3 X a rate of 3.5 X min-l, (c) NO, hydrolysis to 0.2 HONO and 0.8 wall-adsorbed NO, at a rate of 1.4 X min-l, (d) OH radical input from the chamber radical source occurring at a rate given by 0.15kl ppb-l min-l, and (e) NO2 off-gasing occurring at a rate also given by 0.15k, ppb-l min-'. No other chamber-dependent reactions were used in these simulations, nor were any adjustable parameters employed to optimize the fits of the computer simulations to the experimental data. The initial concentrations of NO, NO2, and the minisurrogate components used in the calculations were the averages of the measured initial concentrations of these species. The initial concentrations of the added (or subtracted) test organics were varied to determine the dependence of the incremental reactivity on the amount of test compound added. Results As noted under Introduction, the purpose of this study was to determine, as a function of time, the effects of incremental additions of various test organics on the amounts of ozone formed and NO consumed in irradiations of a standard base-case surrogate hydrocarbon-NO,-air mixture. The following nomenclature is employed. A t a given time t during an experiment in which the concentration of the test organic was changed by an amount A[organic] relative to its level in the base-case surrogate, the reactivity R,(A[organic]) is defined as R,(A[organic]) = A([03] - [NO]), = ([OJt -[03Io) - ([NO], - [Nolo) (1) and the incremental reactivity IR,(A[organic]) is defined as R,(A[organic]) - R,(O) IR,(A[organic]) = (11) A[organic] where R,(O) is the reactivity observed in the base-case experiment (where A[organic] = 0). The limiting incremental reactivity IR,(O) is defined as
IRt(0)
lim
IR,(A[organic]) =
1
d[organic] (111) The latter quantity is considered to be the most appropriate as a measure of the reactivity of a compound since it removes the dependence of the incremental reactivity on the amount of test organic added. Thus, the limiting incremental reactivity is used as the ultimate criterion for reactivity of an organic in the context of this study. Results of Reactivity Experiments. Figure 1shows the concentration-time profiles for ozone and other selected species from two base-case minisurrogate-NO,-air irradiations. The results of the other base-case experiments and of the reactivity assessment experiments with relatively low amounts of added test compound were similar. The ozone concentrations in these irradiations either leveled off or increased slowly (Figure 1)after the initially rapid formation period. The final ozone yields observed in these irradiations (-0.32 ppm) were taken to indicate the maximum ozone-forming potential of this mixture under the experimental conditions employed, since carrying out the irradiation for longer periods would have not resulted in significantly higher ozone yields. Thus, the incremental reactivities defined in terms of the final (6-h) ozone yield reflect the effects of the test organic on the maximum ozone-forming potential of the mixture, while A[organic]-O
0.40,
0.12,
OZONE
NO
0.06k
0.08
1 0
0.03
+p, &
,'x 1
2
3
4
5
I 6
0.00
TRANS-2-BUTENE '*12T
0.09-
------___ -Y r S i
1
Y
Y
m ,
m
2
3
4
5
I
6
0.001 0
1
2
3
4
5
__
Y
I6
ELAPSED TIME Chourd Figure 1. Experimental and calculated time-concentration profiles of Os,NO, PAN, propene, trans-2-butene, and m-xylene during irradiations of the base-case minisurrogate-NO,-air mixtures: ITC-479, (*) experimental data and (-) model calculation: ITC-584, (X) experimental data and (- -) model calculation.
the incremental reactivities defined for lesser time periods, when ozone formation and NO consumption were still occurring, reflect the effect of the test organic on the rate at which NO was consumed and ozone formed. A chronological listing of the minisurrogate-NO,-air irradiations carried out is given in Table I, which is available as supplementary material (see paragraph at end of paper regarding supplementary material). The amounts of test organic added and the reactivity measurements, R,, derived for t = 1,2,3, and 6 h of irradiation are also listed. As expected, the addition (or deletion) of the test organics had an effect on these reactivity values, although in some cases the magnitude of the effects were not much larger than the run-to-run variability of the standard minisurrogate irradiations. Because of this varigbility in the standard experiments and since the incremental reactivity values, IR,(A[organic]), involve R,(O) as well as R,(A[organic]) (eq 11), it is important that the appropriate values of R,(O) associated with each experiment employing the test organic be determined. These values were derived by linear regression analyses of the reactivity measurements of the standard experiments against the run number for groups of runs where the scatter was judged to be due only
to random fluctuations (see ref 23 for further details). These R,(O) values are also gummarized in Table I, together with their estimated single standard deviation uncertainties derived from the linear regressions. The incremental reactivity values calculated from the data given in Table I are summarized in Table F€ and are plotted against A[organic] in Figures 2-8 for the seven test organics employed in this study. The uncertainty estimates shown were derived by assuming that the uncertainties associated with the A([O,] - [NO]) values in the test runs were the same as those for the corresponding R,(O) values; uncertainties in the measured amounts of test organic added (or subtracted) and other possible systematic uncertainties were neglected. Table I1 also gives the limiting incremental reactivity values obtained by extrapolating or (in the cases where the test organic was also a component of the minisurrogate) interpolating the measured incremental values to zero added organic. In several cases the high estimated uncertainties in the changes of A([O,] - [NO]) caused by adding the organic (especially when the amount of added organic was relatively low) make extrapolation of these values to A[organic] = 0 highly uncertain; in those cases Envlron. Sci. Technol., Vol. 21,
No. 7, 1987 673
,-
T=60
0.8
1-60
T=120
0. 2
>. -0.2 -0.3
g
0
1 0,3 c 0.2
2
50
100
150
200
0
50
100
150
200
, il g T=l80
T1.360
< 4
0.8
t-
+
I
0'6 0.4
6
H -8.1
0.2
-0.2 0. 0
-0.3
-0a4j
-0.5
0
50
: 100
~
150
11 200
0
~
1
f;
50
100
150
,
,
Figure 2. Plots of experlmental and calculated incremental reactivities
of toluene as a function of the amount of added toluene: (@) experT=60 -0.2
model calculation.
11120
-0.4
200
ADDED TOLUENE CPPB3
imental data; (-)
-0.2
-200 -188
100 288 300
400-200 -188
E 100 200 300
ME
ADDED PROPENE CPPB3 Figure 4. Plots of experimental and calculated incremental reactivities of propene as a function of the change in the initial propene concentration: (@) experimental data; (-) model calculation.
0. 10
O.ls
21-4.1-
0
t
1-1
$
0 100 200 300 400 500 600
0 188 200 300 488 500 600
q
t * ::qx 1 x j 1
g 0.15
u
u 0.10
z H
: l i u -1.2
-1.4
0 100 200 300 400 500 600
0 100 200 300 400 580 600
ADDED BENZALDEHYDE CPPB3
-8.05
-800 -400
0
400
808 1200 -800 -400
0
400
800 1200
Figure 3. Plots of experimental and calculated incremental reactivities of benzaldehyde as a function of the amount of added benzaldehyde: (@) experimental data; (-) model calculation.
ADDED N-BUTANE CPPBI Figure 5. Plots of experimental and calculated incremental reactivities of n-butene as a function of the change in the initial n-butane concentration: (@) experimental data: (-) model calculation.
the limiting incremental reactivity was estimated by the weighted average of the experimentally measured values, regardless of A[organic]. The methods used to estimate the limiting incremental reactivities are indicated in the footnotes to Table 11. Results of Computer Simulations. The calculated concentration-time profiles for ozone and other selected species monitored in two representative standard base-case minisurrogate-NO,-air irradiations are shown in Figure 1, together with the experimental data. A detailed discussion of the performance of this chemical mechanism in simulating the results of these and other environmental chamber experiments is given by Carter et al. (26). As can be seen from Figure 1,the model gives reasonably accurate
predictions of the maximum ozone yields and reactant consumption rates in these experiments but overpredicts the rates of ozone formation and NO to NOz conversion in the initial stages of the experiments. This overprediction of the initial reaction rates for most of these minisurrogate runs also occurs for other multiorganic-NO,-air run iqdiations carried out in this chamber (26), though this is not observed for similar experiments carried out in other chambers nor in simulations of irradiations of single organic-NO,-air mixtures carried out in the ITC (26). Simulations of other multiorganic-NO,-air runs carried out in this chamber, in which the initial concentrations of the organics and NO, were varied, indicated that the discrepancies between the observed and predicted initial
674
Environ. Sci. Technol., Vol. 21, No. 7, 1987
T=120
T=60
2,,
I
1 1
-
-50
0
50
100
150
0
SE
im
150
4
0.4
-0.4 0J/
;
0
-50
:+
:
58
100
f
l .*l 150
-50
ADDED TRANS-2-BUTENE
CPPB)
Figure 6. Plots of experimental and calculated incrementalreactivities of trans-Bbutene as a function of the change in the Initial trans-2butene concentration: (9)experimental data; (-) model calculation. O.O
W E
T=60
T=120
0
500
1000
1500
2000
0
500
1000
1500
2000
0
500
1000
1500
2000
0
508
1000
1500
2008
ADDED ETHANOL CPPBI Figure 7. Plots of experimental and calculated incrementalreactivities of ethanol as a function of the amount of added ethanol: (9)experimental data: (-) model calculation.
reactivity did not have any discernible dependence on the initial reactant concentrations or the organic/NO%ratios (26). These data suggest that the effects of this discrepancy may cancel out in predictions of incremental reactivity. The computer simulations of the incremental reactivities of toluene, benzaldehyde, propene, n-butane, trans-2butene, and ethanol are compared with the experimental incremental reactivity values in Figures 2-7, and the limiting incremental reactivities calculated for these compounds are given in Table I11 together with the experimental data. While the calculated curves are sometimes outside the ranges of uncertainty estimated for the ex-
2 cz
0 500 1000 1500 2000 2500 3000
0 508 1000 1500 2000 2500 3000
Table 11. Experimental Incremental Reactivities and Extrapolated Limiting Incremental Reactivities Derived from the Results of the Minisurrogate-NO,-Air Irradiations, Where Reactivity Rt Is Defined as A([O,] [NO]), after 1,2,3, and 6 h of Irradiation
-
A[organic], PPb 0
89 180 0
115 347 518
(R,(A[organic]) - R,(O))/A[organic] 0-2 h 0-3 h
0-1 h 0.20" 0.17 f 0.10 0.21 f 0.05 -1.5' -1.11 f 0.07
-0.42 f 0.02 -0.30 f 0.02
-0.10' -0.17 f 0.10 -0.08 f 0.05
-0.23' -0.37 f 0.10 -0.19 f 0.05
Benzaldehyde -1.4' -1.16 f 0.07 -0.59 f 0.02 -0.41 f 0.02
-1.3' -1.06 f 0.07 -0.52 f 0.02 -0.40 f 0.02
-1.6' -1.24 f 0.07 -0.56 f 0.02
0.49 f 0.05 0.72 f 0.07
0.20 f 0.05 -0.01 f 0.07 -0.03d -0.27 f 0.26 -0.14 f 0.09 0.01 f 0.10
0.36 f 0.05 0.43 f 0.05 0.40d 0.27 f 0.13 0.42 f 0.07 0.35 f 0.04
Propene 0.54 f 0.05 0.63 f 0.07 0.42d 0.20 f 0.23 0.33 f 0.09 0.26 f 0.09
0.056 f 0.012 0.05e 0.046 f 0.018 0.042 f 0.008 0.036 f 0.008
n-Butane 0.153 f 0.012 0.09e 0.017 f 0.018 0.050 f 0.008 0.032 f 0.008
2.0 f 0.3 1.9' 1.65 f 0.14 1.14 f 0.08
trans-2-Butene 1.6 f 0.3 1.3' 0.65 f 0.14 0.26 f 0.08
0 888 986 1963
0.04" 0.053 f 0.010 0.035 f 0.009 0.035 f 0.004
Ethanol 0.02" 0.026 f 0.013 0.044 & 0.012 0.015 f 0.006
0
0.07e 0.052 f 0.009 0.030 f 0.005 0.018 f 0.003
-166 -162 0
113 126 334 -701 0
483 1009 1129 -25.3 0
58.9 163
965 1840 2685
0-6 h
Toluene 0.06" 0.01 f 0.10 0.07 f 0.05
tert-Butyl Methyl Ether 0.06e 0.048 f 0.012 0.038 i 0.006 0.031 f 0.004
0.35d
-0.07 f 0.25 0.21 f 0.09 0.11 f 0.09 0.100 f 0.012
0.06e 0.004 f 0.018 0.048 f 0.008 0.034 f 0.008 1.0 f 0.3 0.8* 0.20 f 0.14 -0.06 f 0.08
C
0.081 f 0.012 0.05* -0.006 & 0.018 0.042 f 0.008 0.028 f 0.008 0.4 f 0.3 0.3b 0.03 f 0.04 -0.12 f 0.08
0.01" 0.008 f 0.013 0.050 f 0.012 0.004 f 0.006
0.020 f 0.013 0.050 f 0.012 -0.005 f 0.006
0.04" 0.040 f 0.012 0.046 f 0.006 0.036 f 0.004
0.03" 0.019 f 0.012 0.041 f 0.006 0.021 f 0.004
0.01"
'
" Weighted average of experimental data. Linear extrapolation from data a t the two lowest absolute A[organic] levels. No data. dLinear interpolation between weighted average of data at A[propene] = -166 and -162 ppb and those at A[propene] = 113 and 126 ppb. e Obtained by linear least-squares regression. these processes are manifested by the formation of ozone. Thus, the change in the quantity [O,] - [NO] has a more direct relationship to the processes that are responsible for ozone formation than does the change in ozone alone and thus is appropriate for use in defining reactivity with respect to ozone formation. Although the reaction mechanisms of different types of organics in NO,-air irradiations are complex (17,26),for most organics ozone formation and NO consumption are influenced primarily by four major factors. The incremental reactivity parameters measured in this program and their dependence on the length of time of the irradiation reflect to varying degrees these factors, as indicated below: (1)The rate of consumption of the organic in the atmosphere, particularly with the OH radical and, when applicable, by photolysis or reaction with O3or with the NO3 radical, determines to a large extent the rate,at which the organic converts NO to NOz, i.e., the magnitude of R3, and thus how fast it causes O3 formation. (2) The number of molecules of NO oxidized per molecule of organic reacted determines how much 03 will be formed by consumption of a given amount of the organic. 676
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(3) For organics whose major atmospheric sink is via reaction with OH radicals, the OH radical concentrations present also determine how fast the organic causes O3 formation. If the reactions of the organic enhance radical levels, this will cause more rapid reaction of all organics present that react with OH radicals, and thus increase the rate of ozone formation. The converse is true if the organic is a radical inhibitor. (4) 0, formation can occur only when NO, is present, since as indicated above O3 is formed only by photolysis of NO2. Thus, organics whose atmospheric oxidation mechanisms involve significant NO, sinks will reduce the maximum amount of ozone formed in organic-NO,-air irradiations in which NO, is consumed prior to the end of the experiment. While this factor will influence the ultimate ozone-forming potential of a mixture, it will not affect the rate of formation of O3in the initial stages of an irradiation or under conditions when sufficient NO, is present that NO, availability is not a limiting factor. To illustrate the extent to which the differences in atmospheric reaction rates of the organics studied in this program can account for the differences in some of the reactivity parameters observed, Table I11 lists the calcu-
Table 111. Experimental and Calculated Limiting Incremental Reactivities, Calculated Fractions of Organic Reacted, and Calculated Limiting Incremental Reactivities Normalized for Fraction Reacted, for Selected Organic Compounds for 1,2,3, and 6 h of Irradiation in the ITC Minisurrogate-NO,-Air Irradiations calculated compd benzaldehyde
to1uen e
propene
trans-2-butene
n-butane
ethanol
irrad exptl fraction fraction time, h IRt(0) IR,(O) reacted reacted 1 2 3 6 1 2 3 6 1 2 3 6 1 2 3 6 1 2 3 6 1 2 3 6
-1.5 -1.4 -1.3 -1.6 0.20 0.06 -0.10 -0.23 0.40 0.42 0.35 -0.03 1.9 1.3 0.8 0.3 0.05 0.09 0.06 0.05 0.04 0.02 0.01 0.01
-1.12 -0.76 -0.82 -1.22 0.26 0.04 -0.16 -0.23 0.45 0.43 0.21 0.03 1.5 0.41 -0.11 -0.11 0.028 0.041 0.049 0.048 0.012 0.011 0.014 0.014
0.17 0.25 0.29 0.37 0.09 0.14 0.15 0.20 0.33 0.58 0.73 0.92 0.90 1.00 1.00 1.00 0.036 0.057 0.067 0.087 0.036 0.061 0.073 0.094
-6.7 -3.0 -2.9 -3.3 2.9 0.26 -1.07 -1.16 1.4 0.75 0.29 0.03 1.7 0.41 -0.11 -0.11 0.78 0.71 0.74 0.55 0.35 0.18 0.19 0.15
lated fractions of the test organics that reacted during the time periods used to assess reactivity. For the less reactive compounds, the fractions reacted were relatively small, and this was clearly an important factor in accounting for their relatively low observed reactivities. This effect can be factored out, to a first approximation, by dividing the incremental reactivities by the fractions of test organics reacted, and these calculated ”normalized” incremental reactivities are also included in Table 111. It can be seen that normalizing the incremental reactivities to the amounts reacted reduces significantly the range of reactivities observed. However, Table I11 shows that the differences in the reaction rates do not totally account of the differences in reactivity observed in these experiments and predicted by the computer simulations. The most notable difference in this regard is that: the incremental reactivities differ not only in magnitude but also in sign, with some compounds having “negative” reactivity. For example, toluene has a negative reactivity in terms of the ultimate amount of ozone formed and NO oxidized, and benzaldehyde has negative reactivity both in terms of the ultimate ozone yields and in terms of rates of ozone formation and NO oxidation. This effect, which has been noted previously (30),can be explained by the effects of the photooxidation reactions of the organic on radical levels and NO, removal rates. By far the most extreme example of negative reactivity of the compounds studied in this program is benzaldehyde, which not only has the largest magnitude of negative reactivity but is also the only compound studied that had negative reactivity at all reaction times. This is not unexpected since benzaldehyde is known from previous chamber studies to be a photochemical smog inhibitor (19, 201,and its currently accepted photochemical mechanism (17)predicts that its reaction with hydroxyl radicals [its
major atmospheric sink, together with photolysis to form nonradical products (17)] results in no radical regeneration and in the removal of at least one molecule of NO, from the system. This contrasts with the other organics studied in this program, whose NO,-air photooxidation mechanisms involve predominantly radical regeneration and in which NO, sinks mainly occur as a result of secondary reactions of the products. Toluene is interesting in that it is known to be a highly reactive compound in terms of O3 formation when irradiated by itself in NO,-air systems (31). However, when added to the minisurrogate mixture used here, it has a negative incremental reactivity in terms of the maximum O3 yields but a positive effect on reactivity during the initial stages of the irradiation. The observation of negative reactivity can be explained by the presence of NO, sinks in the toluene photooxidation mechanism (17),which reduce the maximum amount of ozone that can be formed. NO, sinks in the toluene photooxidation mechanism (17) include the formation of PAN, the formation of cresols, which react with NO3 radicals, the formation of benzaldehyde, which, as indicated above, reacts to remove NO,, and (probably) other processes. These NO, sinks do not significantly affect the rates of NO oxidation and ozone formation in the initial stages of the experiment when the availability of NO, is not the limiting factor. Indeed, as seen from Table 111,the incremental reactivity of toluene during the initial stages of the experiment when normalized for its reaction rate is relatively high compared to those of the nonaromatic compounds listed. This is consistent with the formation of radical initiators in its photooxidation mechanism (17, 28). The relatively high normalized reactivities of toluene in the initial stages of the experiments and its negative reactivity in terms of final ozone yields observed in these experiments are reasonably well predicted by our model calculations. The alkenes propene and trans-2-butene are analogous to toluene in that they have relatively high, and positive, incremental reactivities with respect to the initial rates of NO oxidation and ozone formation but have low or slightly negative incremental reactivities with respect to the ultimate amounts of O3formed and NO oxidized. The latter observation could be due, in part, to the fact that these alkenes react directly with 03,although many of the intermediate species formed in these reactions convert NO to NO2 (or NOzto NO3) (17), resulting in net O3 formation. The NO, sink caused by the relatively high yields of PAN in the photooxidations of these alkenes is probably the more significant factor accounting for their low or negative ultimate reactivities.’ The relatively high initial reactivities of these alkenes are attributed to radical formation from the reactions of the alkenes with ozone (where the effects of radical initiation more than compensates for the loss of ozone due to this process) and the photolysis of the aldehyde products. Thus, although these alkenes are quite different from the aromatics in the details of their atmospheric photooxidation reactions (17,26),the net effects of these mechanisms on the incremental reactivities of these compounds are similar. n-Butane and ethanol have positive incremental reactivities with regard to both the initial NO consumption and ozone formation rates and the ultimate amounts of NO oxidized and ozone formed. However, the initial normalized reactivities for these compounds are lower than those for the aromatics and the alkenes, and this is expected since the alkane and alcohol photooxidation mechanisms lack the radical sources present in the aromatic and alkene photooxidations. Moreover, NO, sinks Environ. Scl. Technol., Vol. 21, No. 7, 1987
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are not as important in the photooxidation mechanisms of the alkanes and alcohols as for the alkenes and the aromatics, and hence n-butane and ethanol have less of a tendency to become negatively reactive with regard to the final ozone yields. It can also be seen that n-butane has a reactivity that is 2-4 times higher than that of ethanol, despite the fact that both of these organics react with OH radicals at very similar rates (21). The greater normalized incremental reactivity of n-butane arises because more molecules of NO are consumed per molecule of organic reacted for n-butane than for ethanol. As noted above, the photooxidation of ethanol involves the direct formation of H02 radicals, which consume one molecule of NO in regenerating OH radicals, while the reaction mechanism of n-butane and of the other alkanes is more complex, involving the formation of organic peroxy radicals in addition to H02,which convert additional molecules of NO to NO2 (17,26, 32). Conclusions The purpose of this study was to develop an experimental procedure to measure incremental reactivities relative to ozone formation in the atmosphere and to test this procedure with representative organic compounds whose reaction mechanisms are moderately well understood. Although there are difficulties in obtaining precise incremental reactivity values for relatively unreactive compounds, the experimental technique employed can be used successfully to measure the effects of incremental additions of organics on atmospheric ozone formation. The data obtained show that the incremental reactivities of organic compounds vary widely and depend to a significant extent on the reaction mechanism, rate, and conditions under which these compounds are emitted. Thus, reactivity scales based only on how rapidly the organic reacts, such as those based on OH radical rate constants (3),neglect factors that significantly affect not only the magnitude but in some cases also the sign of the incremental reactivities. In addition, the fact that the magnitude and in some cases the sign of the incremental reactivities can depend on, for example, the availability of NO, shows that it is a gross oversimplification to attempt to define a single reactivity scale for all organics that is applicable under all conditions. The fact tl at the reactivities of organics depend on the conditions under which they are emitted clearly presents difficulties in deriving appropriate control strategies for organic compounds. I t is particularly difficult to assess whether a compound that is marginally reactive or that tends to exhibit negative reactivity under some conditions and positive reactivities under others should be considered to be reactive or unreactive under the conditions of a specific airshed. Airshed model calculations are potentially useful in this regard, provided that (a) the model accurately represents the major conditions of the airshed that would affect incremental reactivities of compounds emitted into it, (b) the chemical reaction mechanisms of the compounds whose reactivities are being assessed are reasonably well understood, (c) experimental data such as those obtained in this study are available to test the ability of the model to accurately simulate the incremental reactivities of the compounds under a variety of conditions, and (d) the sensitivities of predicted incremental reactivities to uncertainties in the chemical reaction mechanisms and in the conditions of the airsheds are reasonably well understood. We are currently carrying out a study of the use of model simulations for this purpose, and the results of this study will be the subject of a future publication. 678
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Acknowledgments We thank Margaret C. Dodd and William P. Long for assistance in carrying out the experimental program. Supplementary Material Available Table I listing the experimental conditions and data obtained from the standard minisurrogate-NO,-air irradiations and minisurrogate irradiations with added (or reduced) amounts of the test organics and estimated standard reactivity values associated with the irradiations with the test organics (3 pages) will appear following these pages in the microfilm edition of this volume of the journal. Photocopies of the supplementary material from this paper or microfiche (105 X 148 mm, 24X reduction, negatives) may be obtained from Microforms Office, American Chemical Society, 1155 16th St., N.W., Washington, DC 20036. Full bibliographic citation (journal, title of article, authors’ names, inclusive pagination, volume number, and issue number) and prepayment, check or money order for $10.00 for photocopy ($12.00 foreign) or $10.00 for microfiche ($11.00 foreign), are required. Registry No. TBME, 1634-04-4;C6H5CH,, 108-88-3;CBH5CHO, 100-52-7; CH&H20H, 64-17-5; H&(CH2)2CH3, 106-97-8; H&CH=CH2, 115-07-1; (E)-H&CH=CHCHs, 624-64-6; NO,, 11104-93-1; 03,10028-15-6.
Literature Cited (1) Altshuller, A. P.; Bufalini, J. J. Enuiron. Sei. Technol. 1971, 5, 39-64 and references cited therein. (2) Dimitriades, B. Proceedings of the Solvent Reactiuity Conference;US. Environmental Protection Agency: Research Triangle Park, NC, Nov. 1974; EPA-650/3-74-010. (3) Darnall, K. R.; Lloyd, A. C.; Winer, A. M.; Pitts, J. N., Jr. Enuiron. Sei. Technol. 1976, 10, 692-696. (4) Pitts, J. N., Jr.; Lloyd, A. C.; Winer, A. M.; Darnall, K. R.; Doyle, G. J. “Development and Application of a Hydrocarbon Reactivity Scale Based on Reaction with the Hydroxyl Radical”; 69th Annual Meeting of the Air Pollution Control Association, Portland, OR, June 27-July 1, 1976; Air Pollution Control Association: Pittsburgh, PA, 1976; paper 76-31.1. (5) Pitts, J. N., Jr.; Winer, A. M.; Darnall, K. R.; Lloyd, A. C.; Doyle, G. J. Proceedings, International Conference on Photochemical Oxidant Pollution and Its Control;Dimitriades, B., Ed.; U.S.Environmental Protection Agency: Research Triangle Park, NC, Jan. 1976; Vol. 11,pp 687-704, paper 14-2, EPA-600/3-77-001b. (6) Pitts, J. N., Jr.; Winer, A. M.; Doyle, G. J.; Darnall, K. R. Enuiron. Sei. Technol. 1978, 12, 100-102. (7) Farley, F. F. Environ. Sci. Technol. 1978, 12, 99-100. (8) Cox, R. A.; Derwent, R. G.; Williams, M. R. Enuiron. Sei. Technol. 1980, 14, 57-61. (9) Heuss, J. M.; Glasson, W. A. Environ. Sci. Technol. 1968, 2, 1109-1116. (10) Glasson, W. A,; Tuesday, C. S. J.Air Pollut. Control Assoc. 1970,20, 239-243. (11) Wilson, K. W.; Doyle, G. J. “Investigation of Photochemical Reactivities of Organic Solvents”;final report on SRI Project PSU-8029; Standford Research Institute: Irvine, CA, Sept. 1970. (12) Glasson, W. A,; Tuesday, C. S. Enuiron. Sei. Technol. 1970, 4, 916-924. (13) Laity, J. L.; Burstain, F. G.; Appel, B. R. Adu. Chem. Ser. 1973,124,95-112. (14) Joshi, S. B.; Dodge, M. C.; Bufalini, J. J. Atmos. Environ. 1982,16, 1301-1310. (15) Carter, W. P. L.; Atkinson, R.; Winer, A. M.; Pitts, J. N., Jr. Int. J . Chem. Kinet. 1982,14, 1071-1103. (16) Bufalini, J. J.; Walter, T. A.; Bufalini, M. M. Environ. Sei. Technol. 1977,11, 1181-1185. (17) Atkinson, R.; Lloyd, A. C. J. Phys. Chem. Ref. Data 1984, 13, 315-444. (18) Finlayson-Pitts, B. J.; Pitts, J. N., Jr. Atmospheric Chemistry: Fundamentalsand Experimental Techniques;Wiley: New York. 1986.
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Kuntz, R. L.; Kopczynski, S. L.; Bufalini, J. J. Environ. Sei. Technol. 1973, 7, 1119-1123. Gitchell, A.; Simonaitis, R.; Heicklen, J. J. Air Pollut. Control Assoc. 1974, 24, 357-361. Atkinson, R. Chem. Rev. 1986,86, 69-201. Zafonte, L.; Rieger, P. L.; Holmes, J. R. Environ. Sei. Technol. 1977, 11, 483-487. Atkinson, R.; Carter, W. P. L.; Winer, A. M. “Evaluation of Hydrocarbon Reactivities for Use in Control Strategies”; final report to the California Air Resources Board; California Air Resources Board Sacramento, CA, May 1983; Contract A0-105-32. Doyle, G. J.; Bekowies, P. J.; Winer, A. M.; Pitts, J. N., Jr. Environ. Sei. Technol. 1977, 11, 45-51. Carter, W. P. L.; Dodd, M. C.; Long, W. D.; Atkinson, R.; Dodge, M. C. “Outdoor Chamber Study to Test Multi-Day Effects”; U.S. Environmental Protection Agency: Research Triangle Park, NC, March 1985; Vol. I, Results and Discussion, EPA-BOO/ 3-84-115. Carter, W. P. L.; Lurmann, F. W.; Atkinson, R.; Lloyd, A. C. “Development and Testing of a Surrogate Species Chemical Reaction Mechanism”; U.S. Environmental Protection Agency: Research Triangle Park, NC, 1986; EPA-600/ 3-86-031. Pitts, J. N., Jr.; Winer, A. M.; Carter, W. P. L.; Doyle, G. J.; Graham, R. A.; Tuazon, E. C. “Chemical Consequences of Air Quality Standards and of Control Implementation
(28) (29) (30) (31) (32)
Programs”; final report to the California Air Resources Board; California Air Resources Board: Sacramento, CA, June 1980; Contract A7-175-30. Atkinson, R.; Lloyd, A. C.; Winges, L. Atmos. Environ. 1982, 16, 1341-1355. Lurmann, F. W.; Lloyd, A. C.; Atkinson, R. J. Geophys. Res. 1986, 91, 10905-10936. Bufalini, J. J.; Dodge, M. C. Environ. Sei. Technol. 1983, 17, 308-311. Atkinson, R.; Carter, W. P. L.; Darnall, K. R.; Winer, A. M.; Pitts, J. N., Jr. Int. J. Chem. Kinet. 1980,12, 779-836. Carter, W. P. L.; Atkinson, R. J . Atmos. Chem. 1985, 3, 377-405.
Received for review August 25,1986. Accepted January 27,1987. W e acknowledge the California Air Resources Board (Contract A0-105-32; Project Monitor Jack K. Suder) for financial support for the experimental portion of this study and the U.S. Enuironmental Protection Agency (Cooperative Agreement C R 810214-01; Project Officer Joseph J . Bufalini) for financial support for the reactivity calculations. Although the research described in this article has been funded in part by the Enuironmental Protection Agency, it has not been subject to Agency review and therefore does not necessarily reflect the views of the Agency, and no official endorsement should be inferred.
Acid Rain and Atmospheric Chemistry at Allegheny Mountain William R. Plerson,* Wanda W. Brachaczek, Robert A. Gorse, Jr., Steven M. Japar, and Joseph M. Norbeck Research Staff, Ford Motor Company, Dearborn, Michigan 48 121
Gerald J. Keelert Department of Atmospheric and Oceanic Science, University of Michigan, Ann Arbor, Michigan 48109
Rain chemistry was measured in August 1983 on Allegheny Mouhtain and Laurel Hill in southwestern Pennsylvania. The average compositioh approximated an H2S04/HN03mixture with a volume-weightedaverage pH of 3.5 and an SO:-/NOy mole ratio of 1.8. There was very little undissociated (weak) acidity and very little S/IV). The acidic rains were associated with air masses traversing SO2 source regions west of the sites; stagnation and intervening precipitation were important influences. The geographic scale for a halving of rain S042- concentration downwind of SO2sources was -440 km. Scavenging ratios were inferred for SO2, aerosol SO-,: and HNOB. On average about half of the rain SO-: resulted from scavenging of SO2,the rest from scavenging of aerosol S042-. The rain H+ was attributed about 25% to “OB, 55% to scavenging of SO2, and 20% to scavenging of aerosol acid SO-.: Cumulative deposition totals in rain were compared with deposition in fog and with dry deposition in the same experiment. A crude acid-deposition budget was calculated as follows: 47% , H2S04in rain; 23%, SOzdry deposition without dew; 1 6 % , HN03 in rain; I l k , HN03 dry deposition without dew; 2 % , HN03 and H2S04in fog and dew; 0.5 %, aerosol dry deposition without dew. Introduction
In August 1983, we carried out a field experiment involving various aspecta of the acid-deposition phenomenon at two rural sites on Allegheny Mountain and Laurel Hill Present address: Harvard School of Public Health, 655 Huntington Avenue, Boston, MA 02115. 0013-936X/87/0921-0679$01.50/0
in southwestern Pennsylvania (Figure 1). This paper deals with the wet (rain) deposition during the experiment. Other aspects of the experiment have been discussed elsewhere: the chemistry of dew and its role in acid deposition ( I ) , the dry deposition of “OB and SO2 to surrogate surfaces (2))and the role of elemental carbon in light absorption and of the latter in visibility degradation (3). The Allegheny and Laurel sites, situated only 150 km downward of the highest density of SOz emissions in the U S . (see ref 4 or Figure 4 of ref 5)) lie in the area of the highest rainfall H+ concentrations and highest annual totals of H+ wet deposition in the US.or, indeed, in North America (6-9). The area moreover has a history of ecological damage ascribed to acid rain; trout kills have been recorded in poorly buffered headwater streams on Laurel Hill since 1960 ( 8 ) ) followed by failures of restocking programs and the disappearance of trout and other fish from some streams-all attributed to acid rain and consequent stream acidification during storm runoff periods (8, IO, 11). This paper describes the rain chemistry in the 1983 experiment and explores chemical relationships with the atmospheric trace gases and aerosol. Air mass trajectories to the sites are briefly considered.
-
Experimental Section
The experiment was conducted August 5-28,1983, on abandoned radio towers atop Allegheny Mountain (at 39.959’ N, 78.8525’ W, elevation 838 m) and Laurel Hill (at 40.099’ N, 79.226’ W, elevation 850 m, 35 km northwest of Allegheny Mountain) (Figure 1). Both sites are
0 1987 American Chemical Society
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