Modelling the atmospheric mercury cycle-chemistry in fog droplets

Pergamon

Atmospheric Environment Vol. 29, No. 12, pp. 1441-1457, 1995 Copyright © 1995 Elsevier S¢i¢nc¢ Ltd Printed in Great Britain. All rights reserved 1352-2310/95 $9.50 + 0.00

1352-2310 (94)00323-8

MODELLING THE ATMOSPHERIC MERCURY CYCLE--CHEMISTRY IN FOG DROPLETS KARIN PLEIJEL and JOHN MUNTHE Swedish Environmental Research Institute (IVL), P.O. Box 470 86, S-402 58 G&eborg, Sweden (First received 22 March 1994 and in final forra 8 October 1994)

Abstraet--A raodel has been developed describing the mass transport and chemistry of different forms of mercury in the atmosphere (the CAM model). 48-hour simulations of an air parcel containing a fog have been used to .examine the influence of a number of chemical parameters on dissolved divalent mercury, Hg(II), in fog droplets. Representation of chlorine chemistry was found to be very important for modelling of mercury species, as mereury-chloride complexes dominate the dissolved Hg(II) fraction in competition with the reactive Hg(II)-S(IV) complexes. If the pH is increased, the importance of HgCI2 will decrease in favour of Hg(l I)--S(IV)complexes which, in turn, will lead to lowered concentrations of dissolved Hg(II), due to an enhanced production of volatile Hg° via reduction of HgSO v At low SO 2 concentration (0.5 < SO 2 < 10 ppb) dissolved mercury is strongly inversely dependent on the gas phase SO 2 concentration. The ozone concentration is almost linearly related to the dissolved Hg(II) content. Total mercury content (dissolved plus adsorbed Hg(II)) is strongly correlated to soot concentration. At high soot concentrations all Hg(II) is expected to be found in the adsbrbed form. Key word index: Mercury, atmosphere, cloud, fog, aerosol, oxidation, reduction, sensitivity analysis, model.

1. INTRODUCTION Since the 1960s the problems caused by mercury in the environment has attracted considerable attention. Today, in the 1990s, this problem remains a hot issue, Some forms of mercury (methyl mercury) are extremely toxic, capable cf causing serious damage to the central nervous system. Mercury is a natural component of coal, peat and natural gas, and thus combustion of these fuels will lead to mercury emissions. In addition to this, waste incineration, chlorine-alkali production, crematory incineration and the use of certain pesticides lead to mercury emissions (Lindqvist et al., 1991). Mercury is also emitted from natural surfaces such as forest soil and freshwater lakes (Xiao et al., 1991a) as well as the open ocean (Fitzgerald, 1986). A major fraction of these emissions are probably re-emissions of previously deposited mercury originating from anthropogenic emissions. Mercury has for a long period of time accumulated in soil due to atmospheric deposition. A continuous leaching out of this soil mercury occurs and will remain a threat to aquatic ecosystems in the future. Although the mcrcury problem has received attention for a long tittle, the behaviour of mercury is so complex that many questions remain unsolved. Mercury is the only metal that occurs in the gaseous phase in nature, and is capable of long distance transport in

its atomic elemental form (Hg°). In air, Hg ° is the dominant form, and background concentrations usually range from 2 to 4 ng m - 3 in the northern hemisphere. A smaller fraction, a few per cent, is associated to particles (Brosset, 1982; Slemr et al., 1985; Iverfeldt, 1991; Xiao et al., 1991b) and traces of methyl mercury forms are also present in air (Fitzgerald et al., 1991; Brosset and Lord, 1995). The atmospheric cycle of mercury is closely linked to changes in its chemical form. In precipitation, mercury is present in both particulate and dissolved forms. In industrialised regions, the fraction associated with particulate matter is predominant and the dissolved forms are usually less than 50% of the total concentration of 5 to 50 ng ( - 1 (2.5 x 10-11_2. 5 x 10- t0 M ) . Methyl mercury forms are present at concentrations around 0.05 to 0.5 ng d - 1 (2.5 x 10-13-2.5 x 10-12 M) (Munthe and Iverfeldt, 1993). The major removal mechanisms for atmospheric mercury are wash-out of airborne particulate and oxidised forms, aqueous oxidation of elemental mercury to more water-soluble forms and dry deposition. Oxidation of Hg ° by 0 3 to Hg(II) is known to occur in the aqueous phase (Iverfeldt and Lindqvist, 1986; Munthe, 1992) and is probably the major removal mechanism for Hg °. However, a reverse process where dissolved Hg(II) is reduced back to Hg ° is also known to occur (Munthe et al., 1991) and may lead to the transfer of mercury from the aqueous droplets to the

1441

1442

K. PLEIJEL and J. MUNTHE

gas phase. Earlier studies have shown that the total amount of Hg ° removed from the atmosphere is a function of the rates of oxidation and reduction of the mercury and will depend not only on the amount of mercury present in the atmosphere but also the presence of O a and SO 2 (Munthe, 1991, 1992). More realistic model calculations, combined with laboratory experiments and field studies, are needed to explain the relation between emissions, atmospheric transformations and deposition of mercury. This work focuses on the chemical behaviour of mercury in fog droplets. In order to retain the possibility of a full understanding of the chemical processes in the droplets, the model describes the gas and aqueous phase chemistry in detail, while the description of dynamic parameters are held at a low but acceptable level. This paper represents the first in a series aimed at formulating a model for the transport and transformation of atmospheric mercury. In future papers, the further development of this model will be described with focus on precipitation processes and validation.

2. INCLUSIONOF MERCURYCHEMISTRYINTO AN ATMOSPHERICPHOTOCHEMICALMODEL The model simulates an air mass following a wind trajectory. Initially, the air mass is assumed to contain an unpolluted fog, which is gradually influenced by air pollutants emitted from the underlying surface. The chemical behaviour inside the fog is followed for a period of 48 h. The processes covered in this model, called the CAM model (Chemistry of Atmospheric Mercury) are emissions into and deposition out of the fog air, transport in and out of fog droplets, as well as the chemical reactions occurring in the gas phase and in the fog droplets. The association of mercury onto particles within droplets is also considered. The model used is an extended and modified version of the gas-phase trajectory model developed at Harwell (Derwent and Hov, 1979). 2.1. Gas-phase reactions The gas-phase chemistry of the model comprises around 80 reactions including both thermal and photochemical reactions. The VOC chemistry is simplified in accordance with the CBM-IV (carbon bond mechanism), described by Gery et al. (1989) and Dodge (1990). The CBM-IV model version used here includes in total 31 species, of which 7 represent different carbon bond groups. The gas-phase reactions used in the model are presented in the Appendix, Table A1. 2.1.1. Mercury in the gas phase. In the gas phase, elemental mercury vapour (Hg °) is the predominant form, constituting around 95% of the total mercury in air (Lindqvist and Rodhe, 1985; Slemr et al., 1985). Elemental mercury is not very soluble in water, thus removal of mercury from the air mainly occurs after

chemical conversion to water soluble divalent mercury Hg(II) complexes (Munthe, 1991, 1992,1993). Dry deposition can also be an important process, but the mechanisms have not yet been well enough described to allow modelling. The only known reaction of Hg°(g) of potential importance is believed to be the oxidation by O3(g), shown in reaction (1). Hg°(g) + O3(g)

k' , HgO(g) + 0 2(g).

(1)

In the standard case of the modelling simulations this reaction is excluded. Based on studies performed in the late forties (P'yankov, 1949), an estimated value of the rate constant, k~ = 4.9 x 10 - I s molecules cm -3 s- ~, has been calculated (Yarwood and Niki, 1990) for reaction (1). Using this value, a relatively short atmospheric chemical lifetime can be estimated. However, in a recent experimental study, Hall (1995) concluded that this rate constant was greatly overestimated. Although some uncertainties remain concerning the mechanisms of the reaction, an atmospheric chemical lifetime of 6 to 12 months is indicated with the newer. results. The reaction, to some extent, occurs on surfaces and it is not unlikely that this has influenced the earlier experimental results. It is not likely that this gas-phase reaction represents a major pathway for the removal of Hg ° from air. For this reason, it has not been included in this model, at the present degree of detail. Recent measurements (Stratton and Lindberg, 1995) suggest that significant concentrations of gaseous divalent mercury species may exist in ambient air. However, while it is not yet known if gaseous divalent mercury species are generated from homogeneous oxidation reactions or via direct emissions, interactions of these mercury forms and fog/cloud droplets will certainly be of importance for the overall deposition rate of mercury and will be included in future versions of this model. 2.2. Aqueous-phase reactions 2.2.1. Chemistry of mercury. The only significant redox reactions of mercury considered in this model study, are assumed to take place in the aqueousphase. In Fig. 1 the fate of mercury in fog droplets as modelled in this study is shown. Emissions of mercury are assumed to occur in the elemental form, from sources over which the air parcel passes (see Section 2.4 for further details). The elemental mercury will pass through the droplet surface until an equilibrium concentration is established according to Henry's law (see discussion of mass transfer processes in Section 2.3). Elemental mercury, dissolved in the droplet, Hg°(ac0, will be oxidised by ozone to divalent mercury, Hg(II), which forms different complexes according to the chemical composition of the droplet, Removal of mercury from the droplet, occurs through reduction of Hg(lI) via decomposition of HgSO 3 to Hg°(aq) or through photolysis of Hg(OH)2 to

Chemistry in fog droplets

.

°(aq)

ct-

I _soH"

1443

\ /

I-IsCl*

so.

i

so,

o~IcoH) 0q)

Fig. 1. Fate of mercury in the atmosphere, as modelled in this study. Hg°(aq). Particulate mercury is assumed to form as mercury complexes adsorb onto particles within the droplet. The magnitude of the flux of Hg ° over the air/water interface i,; dependent on the concentrations of Hg ° in air and water relative to the Henry's law constant. The driving forces are the two opposing processes; aqueous oxidation of Hg °, which will lead to the transfer of mercury from the air into the droplet, and the reduction of aqueous Hg(II), which will act in the reverse direction. Elemental merc~Lry, dissolved in cloud droplets, Hg°(aq), will be oxidised mainly by ozone, O3(aq), forming HgO(aq), which reacts further to divalent mercury, Hg(II) or HgZ+(aq), with k 2 = 4 . 5 x 1 0 7 M -1 s - t (Munthe, 1992) and k 3 = 1 x 101° M - t s -1, for reactions (2) and (3). The rate of reaction (3) has not been investigated experimentally, but the accumulation of HgO(aq) is not likely from equilibrium considerations. For simplicity, k 3 is set to a value close to the diffusion controlled limit. HgO(aq) + O3(aq)

h2 , HgO(aq) + O2(aq)

(2)

HgO(aq) + H+(aq)

k3 , Hg2+(aq) + OH-(aq). (3)

Another oxidant of potential significance is OH(aq), which is capable of oxidising Hg°(aq) to Hg 2+ (aq), based on preliminary results from on-going work (Munthe, 1994). Divalent mercury, Hg(II), forms a number of complexes, depending on the ions available in the droplet. Hg2+(aq) and sulphite ions, S O l - , form the complexes HgSO 3 and Hg(SO3) 2- via reactions (4) and (5), and Hg(II) is reduced back to Hg ° via HgSO 3 as

shown in reaction (6) (Munthe et al., 1991). Hg 2 + (aq) + SO 2- (aq) ~

xs

, HgSO 3(aq),

(4)

K 5 = 2 x 10-13 M HgSO3(aq) + SO 2-(aq),

x6

, Hg(SO3) 2-(aq),

K 6 = 4 x 10- t2 M HgSOa(aq )

(5)

k, ,Hg°(aq ) + Prod.,

k7 = 0.6 s - 1.

(6)

Hg2+(aq) and chloride ions, CI-, form the chloride complexes HgCl + and HgCl 2 in reactions (7) and (8). A model experiment has been conducted, which included the effect of partitioning H g e l 2 between the gas and aqueous phase. When the dimensionless distribution constant 2.9 x 10 - s (M,j,/Mw, t,r) was used (Iveffeldt and Lindqvist, 1980), the escape of HgCl 2 from the aqueous to the gas phase reduced total dissolved Hg(II) only by 0.3%. The reverse process where gaseous divalent mercury is adsorbed from the gas phase to the droplets may be important but data on the presence and generation of these species are not readily available. Hg 2 +(aq) + C l - (aq),

xs

~HgCl + (aq),

K s = 1.82 x 10- 7 M HgC1 + (aq) + CI- (aq), K 9 = 3.31x 10 -7 M .

(7) r"

, HgCI 2(aq), (8)

Hg2+(aq) and the hydroxide ion, O H - , form HgOH + and Hg(OH)2, (reactions (9) and (10)), and

1444

K. PLEIJEL and J. MUNTHE

the divalent mercury bound as Hg(OH)2 can be reduced back to Hg ° (aq) by photolysis in reaction (11). In addition to these complexes, a mixed complex, HgOHC1, may be formed by reaction (12). Hg2+(aq)+OH_(aq) '

r~o

~HgOH+(aq),

Klo = 2.51 x 10 -11 M HgOH + (aq) + O H - (aq),

(9) KI1

, Hg(OH)2 (aq),

(lO)

K H = 6.31 x 10 -12 M Hg(OH)2(aq) hv.k~, Hg0(aq ) + Prod., k12 = 3 x 10-7 s - 1 , at midday, 60°N HgOH +(aq) + C l - ( a q ) ,

r~

(11)

,HgOHCl(aq),

Kl3 = 3.72x 10 -a M .

(12)

Equilibrium constants in reactions (4), (5), (7)-(10) and (12) have been taken from Smith and Martell (1976). The rate constant for the photolysis reaction (11) is from Xiao et al. (1994). In the cases where equilibrium constants have been used in the present model simulations, the second-order reaction in which the complex is formed is assumed to be limited by aqueous diffusion, using the rate 1 x 101° M - 1 s - ~ (Eigen and Wilkins, 1965). The rate coefficient for the dissociation is then calculated from this rate and the equilibrium constant. The adsorption rate of all forms of dissolved Hg(II) onto soot particulate within the droplet is assumed to be limited by aqueous diffusion. The upper limit for the adsorption is set by an empirical relation described in Petersen et al. (1995), as in equation (13) below, and the process is modelled assuming reversible transport. 1

Hg(II),d, = Hg(II),q. Csoot---"k

(13)

r

where Csoot is the soot concentration in the aqueous phase in (g ma), r is the soot particulate radius in (m), k is a constant of value 5 x 10 -6 (m4 g - l ) . (If soot concentration in air is used a scavenging ratio of 5 x 105 can be used to convert it to soot concentration in the aqueous phase.). 2.2.2. Other aqueous chemical reactions. The aqueous-phase chemistry, apart from the mercury reactions shown above, is described in Appendix, Table A3. Sixty-two species are included in 98 reactions of which 18 are chemical equilibria and seven are photolysis reactions. Particles serving as fog condensation nuclei may contain water-soluble material (for example inorganic salts) and are thus expected to influence the chemical composition of the aqueous content in the droplet. The influence of this has been neglected in the present study, and instead the influence of different fixed pH values has been examined.

2.3. Transport of species across the #as-droplet interface Fogs are composed of droplets and gaseous material, as well as dry particles (condensation nuclei). The processes involved in the transfer of species across the gas-droplet interface include: the transfer of gas to the air-water interface, gas-water equilibrium at the interface (according to Henry's law), chemical equilibrium within the droplet, aqueous-phase mass transport of dissolved species, aqueous-phase chemical reactions, aqueous-phase mass transport of reaction products, transfer of volatile products into the gasphase and gas-phase transport of these species (Schwartz, 1986). The reversible transport of species into and out of droplets was calculated according to Jacob (1986). Mass transport in fogs is thereby assumed to be controlled by molecular diffusion (Schwartz, 1986). The absorption flux, ~in (in molecules per cm a of air per second) of species i to the fog droplet is given by ~i,=

3rlLDg i a2 ' ( C i )

(14)

and the volatilisation flux ~out (in molecules per cm 3 of air per second) of species j to the gas phase is given by ~out - 3rILD~, ~ 6.023 x 1020 C* a2 KnRT

(15)

where a is the droplet radius (in cm), L the liquid water content (vol./vol.), Dg.i is the diffusion coefficient (in cm 2 s - ~) of species i in air, ( C i ) is the bulk gas-phase concentration (in molecules per cm a of air) far from the droplet surface, C~ is the aqueous-phase concentration (M) at the droplet surface, K n is Henry's law constant (M atm-1), R = 8.206 x 10 -2 atm ~ m o l - ~ K - ~ is the gas constant, T is the temperature (K), and ~/ is a coefficient correcting for free molecular effects (0 < q < 1). q is a function of Kn and ct, where Kn is the Knudsen number, i.e. the ratio of the mean free path of air, 2, to the droplet radius (Kn = 22/a). The accommodation (or sticking) coefficient, ~t, will strongly depend on the nature of the molecule itself as well as on the surface properties. This quantity varies greatly, and measured values satisfy 1.10 -4 < ct ~< 1.0 (Chameides and Davis, 1982; Schwartz, 1986; Jacob, 1986) for different substances. No information is available on the sticking coefficient for Hg ° and a value of 0.1 has been used for all species in this study. A number of simulations were performed where the sticking coefficient for Hg ° was varied from 0.001 to 1.0. No change in the rate of incorporation of mercury into the droplets was found under the conditions employed for the modelling. This suggests that this parameter is of minor importance when modelling the atmospheric cycling of mercury. However, in the future extended version of this model, the physical processes of cloud precipitation will be incorporated and further tests of the sensitivity

Chemistry in fog droplets to varying sticking coefficients will be performed under more realistic conditions. The value of Henry's law constant used in this model (0.112 M a t m - 1; from Clever et al., 1985) suggests an aqueous concentration of Hg ° of about 3 x 10-14 M, which is about a factor of 1000 less than what is expected for total mercury in precipitation in relatively clean areas. Thus a significant amount of l-Ig° has to be oxidised before a "steady-state" concentration with equal rates of oxidation and reduction will be established. In the present model app)[ication, where the aqueous mercury concentration is followed from a "clean" situation, i.e. with no Hg in the droplets, the mass transfer rate (e.g. sticking efficiency) is clearly not limiting to the production of aqueous mercury. In a more realistic situation, where droplets are formed from condensation nuclei, the mercury concentration is expected to increase more rapidly due to the simultaneous oxidation of Hg ° with droplet growth and possibly also dissolution of divalent mercury forms adsorbed on the dry condensation nuclei. Thus the actual mass transfer of Hg ° into the droplets would not be as great as in these model calculations. In Iverfeldt and Lindqvist (1986) the transfer of Hg ° into an aqueous phase in the presence of elevated concentrations of O3 was investigated. The aqueous solution used was found to be an efficient trap for the mercury which was interpreted as a rapid mass transfer of Hg ° into the aqueous phase followed by oxidation to Hg(II). 2.4. Emissions Emissions of air pollutants along a simulated trajectory are assumedL to correspond to the emissions in the late 1980s from southern Sweden. The emissions in kg per person and year used are 30 kg of SO2, 48 kg of NO x (as NO2), 200 kg of CO and 58 kg of VOCs. The emissions are assumed to be proportional to the population density of on average 60 inhabitants k m - 2 (SCB, 1989). The diurnal variation as well as the distribution betwecn different VOC species are described in more detail in Pleijel and Andersson-Sk61d (1992). Emissions of me~rcury are held constant and these have been set to 42 g k m - 2 yr, a value representative for southern Sweden (Pacyna, 1989). All emitted mercury is assumed to be in the form of gaseous elemental mercury, Hg°(g). If mercury emissions are excluded, the concentration of total Hg(II) in fog droplets will decrease by 3% a,; an average over 48 h while the average decrease in Hg°(g) and Hg°(ac0 will be 6%. Thus, the contribution to the atmospheric mercury from emissions in Sweden is almost neglible in comparison to the amount of mercury already present in the air mass. 2.5. Boundary concentrations As the model studies cover a sequence of 48 h over a Swedish land area, the results are dependent on the initial conditions chosen. A sensitivity test on some of

1445

the boundary conditions has been carried out and is analysed in Section 5. The calculations start just before an air mass is assumed to arrive at the Swedish west coast. To obtain realistic initial concentrations, data from actual measurements at the Swedish coastline were used. Concentrations of species not measured at the Swedish coast, were set to modelled values of an air mass assumed to pass over the British Isles and transported across the North Sea, while some initial concentrations have been set to values that correspond to measured data at other locations in Europe. The initial concentrations used are shown in Appendix, Table A4. 2.6. Deposition Dry deposition was treated using mean values of deposition rates for 0 3, H N O 3, SO2, NO2, PAN and species that act similar to PAN. The deposition rates, shown in Table A5 of Appendix, were chosen to correspond to a representative Swedish terrain being covered with 50% forest. Only a non-precipitating fog was simulated, and wet deposition was excluded. The sensitivity analysis indicates that the inclusion of wet deposition of the fog, does not greatly influence the concentration of mercury in the aqueous phase during the 48 h studied. 2.7. Meteorology and other physical data The chemical reactions are assumed to take place in a well-mixed box, including a fog (or a low-level cloud) of 600 m thickness. The solar radiation intensity and diurnal variation on the 10th of May was chosen at the latitude 58°N (G6teborg, Sweden). Wind speed was set to 2.5 m s -x, relative humidity inside the fog 100%, and temperature 20°C. The radius of a fog droplet usually varies from 0.5 to 30 #m, with the most common radius at 0.8 #m (Warneck, 1988). Liquid water content (LWC) reaches a maximum in the upper half of the clouds, and typical mean values for a non-precipitating cloud are 0.1-1 g m -3, and for a fog 0.1 to 0.5 g m -3. In the standard case a uniform droplet distribution using fog droplets of radius 10 gm and a LWC of 1 x 10 -6 (1 g m -3) was assumed. Photolytic processes occurring inside a fog are expected to be enhanced in the upper part of the fog due to increased scattering. In the lower part however, the photolysis is expected to be reduced due to the solar radiation reduction by the fog. These variations in solar flux have not been included in the standard version of the CAM model. Results from a sensitivity analysis suggest that an increase in solar flux by 20% will lead to an increase in dissolved Hg(II) with 5%, whereas a 20% solar flux decrease will lower the concentration of dissolved Hg(II) by 11%. 2.8. Mathematical description of the model In mathematical terms, the chemical part of the model consists of ca 190 differential equations, repres-

1446

K. PLEIJEL and J. MUNTHE

enting the reactions in the gas and aqueous phase for ca 90 chemical species in the boundary layer. The calculation program FACSIMILE/CHEKMAT (Curtis and Sweetenham, 1987), employing Gears method for solving the differential equations, has been run on a personal computer. For a species, i, the differential equation that represents the concentration development in time, Ci(t), is expressed as dCi(t) = P~ - LiC~(t) + Ei dt h

Vi, o Ci(t )

(16)

h

where P~ is the chemical production rate (in molecules cm-3 s-1), Li is the chemical loss (in s-1), E~ is the emission rate (in molecules crn-3 s-1), h is the height of the mixing layer (in cm) and V~.0 is the dry deposition rate (in cm s-1).

3. SIMULATIONRESULTS The simulations were carried out using the combination of CBM-IV gas-phase chemistry, the aqueousphase chemistry and the aqueous-phase mercury chemistry described in the previous section. The development of the divalent mercury content inside fog droplets, during the 48 h simulation, is shown in Fig. 2. The simulation was carried out without setting the pH to a fixed value, which means that pH will decrease from 4.5 in the beginning to 3.3 at the end of the simulation. Most of the divalent mercury is found in the adsorbed phase---a natural consequence of the empirical relation that these calculations are based upon. A considerable amount of the dissolved Hg(II) is found as the HgCI 2 complex. Smaller fractions are found as HgOHC1 and HgCI +. The dominance of the chloride complexes is due to the relatively high concentrations of chloride that is assumed to be available

in the droplets (see Section 5), the stability of the mercury-chloride bond and the pH chosen (Fig. 3.). In Fig. 3 the average dissolved Hg(II) concentration as well as the relative mercury complex distribution for pH's ranging between 3 and 5.2 is shown. The pH dependence is strong, and a higher pH correlates to lower concentrations of Hg(II). An increase in pH will favour production of S O l - and HgSOa, thus production of Hg ° through reaction (7) will be enhanced, leading to less dissolved Hg(II). The increase of O H - and S O l - with increasing pH will lead to a decrease in the relative importance of HgCI 2 in favour of HgOHC1, Hg(OH)2 and Hg(SO3) ~-. The simulations were conducted using the assumption that mercury complex equilibria are established at a bimolecular rate of 1 x 101° M - 1 s- 1 (see Section 2.2.1). If a bimolecular rate of 1 x 10s M - i s - 1 is used instead (i.e. equilibria are established at a rate 100 times slower), the total Hg(II) is reduced by 24%, and the distribution between different complexes is displaced towards more HgOHCI and less HgCI 2. At a slower rate of establishing equilibria the reduction of HgSO 3 to Hg ° will be favoured, leading to less Hg(II). The decrease of Hg(II) through photolysis of Hg(OH) 2 to Hg ° is of minor importance, compared to the decrease through the reduction of HgSO 3. In the "no-set-pH" case, Hg(II) increased by 24% when reduction of HgSO3 to Hg ° was excluded, while the influence on Hg(II) through reduction via photolysis of Hg(OH) 2 (reaction (11)), was only parts per thousand. If no transport of gaseous species into the droplet, nor any aqueous phase chemistry is included in the model simulations, the concentration of ozone, OH and SO2 is illustrated in parallel to fog simulation concentrations in Figs 4-6. The "no-fog" ozone concentration follows a diurnal pattern with a maximum

5.5E-11 5E-11 4.5E-11 4E-11 3.5E-11 3E-11 :S 2.5E-11 2E-11 1.5E-11 1E-11 5E-12 0

I"1Hg (11) ads nnHGOHCL n HGOH2 i HGOH1P tHGCL2 ! HGCL1P • HGSO322M IMHGSO3 •HGO 7

13

19

1

7

13

19

1

7

Time (hours)

Fig. 2. Development of divalent mercury content, Hg(II), in the simulated fog droplets, during 48 h, starting at 7 a.m. "Hg(II) ads" means all Hg(II) complexes that are adsorbed onto particles within the droplet, HGOHCL is the dissolved complex HgOHCI, HGOH2 is Hg(OH)2, HGOH1P is HgOH +, HGCL2 is Hg(CI)2, HGCL1P is HgCI +, HGSO322M is Hg(SO3)~-, HGSO3 is HgSO3, HGO is HgO.

Chemistry in fog droplets

1447

13.OOE-12

[ ] HGOHCL

7.00E-12

[ ] HGOH2

~5.00E-12

M

!5.00E-12

[ ] HGOHIP

,$.00E-12

[ ] HGCL2

3.00E-12

[ ] HGCL1P

2.OOE-12

•

1.00E-12 0.00E + O0

HGSO322M HGSO3

NosetpH

pH 3.0

pH 4.5

pH 5.2

[ ] HGO

Fig. 3. Average concentration and complex distribution of dissolved Hg(II), for different pH values. The ' "No-set-pH" case, means pH is not set to a fixed value, and will decrease from 4.5 at the beginning to 3.3 at the end. For explanation of symbols, see Fig. 2.

o Fog O3(aq) o Fog O3(g)

60t O , o"

50

~

~

"

-0.40

O3(g)

4o A

D.30 ~

o

e~

i "'13._

30

i

8

-

°'°'°-a~ o ~

,

~

-0.25 "°'°"° ° o-o,o.o,o .o.o.~

~

2O -I!

- 0.50 - 0.45

_

-0.2o

- 0.15

O,o,X~o, °

,,

"°'°'°'oo<>-<>o.<>oo ~

- 0.I0

.4

!

0.05

A

I

13

I

I

19

I

I

I

7

I

13

I

19

0.00

I

Time (hours) Fig. 4. Development of gaseous ozone outside the fog (O3(g)), inside the fog (Fog O3(g)), and of aqueous ozone inside the fog droplets (Fog O3(aq)), over a 48 h period.

1.60E-H)7

• OH(g)

q2.00E-03

1.40E.~O

a Fog OH(g) o Fog OH(aq)

-It'SOE'03 11.60E_03

',,,

1.20E+07

r

1.00E4-07 ~ ~

~

/ L20E.03

11.oo,,-o |,.oo,,.o4

$.00E-~

It,

'

2.00E+06 O.OOE+OG

|6.00~.-04

: o\ \

"-" 4.00D~-~

13

_

H H HI~'H H H= • 13 19 I 7 Time (hours)

|2.ooE-o4

-'o ~

o 19

0

t ,,.oo,,-o,, .....~ . . . .

I

~n OOE+OO

r-"

Fig. 5. Dewdopment of gaseous O H outside the fog (OH(g)), inside the fog (Fog OH(g)), and of aqueous O H inside the fog droplets (Fog OH(aq)), over a 48 h period.

1448

K. PLEIJEL and J. MUNTHE

4.00 r-3.50 ,

}

3.00

• SO2(g) 13 Fog SO2(g)

c~ ~ 2.00

0 Fog S 0 2 ( a q )

~

~

-

. ~ v ~ t 4 t 4 t ' P ~ ~-~

1.60

I

2.50 2.00 -

i\

_..--f-

/

1.50

oo,

o,,

1.00 "~

1.00 0.50 i - ~ Q 0.00 7

_O F ~ O

0'00o'O'DtJ

0 13

19

I

7 13 Time (hours)

19

I I

0.20 0.00 7

Fig. 6. Development of gaseous sulphur dioxide outside the fog (SO2(g)), inside the fog (Fog SO2(g)), and of aqueous sulphur dioxide inside the fog droplets (Fog SO2(aq)), over a 48 h period. in the late afternoon/early evening and a minimum in the early morning, before the sun has risen. When a fog is simulated, the concentration of gaseous ozone is greatly decreased due to the initial transport into fog droplets. In this case the ozone inside the droplets will influence the behaviour in the gas phase as well, and thus play the role as an important aqueous oxidant. A rapid ozone consumption within the droplet leads to further gas-to-droplet transfer. This effect dominates over the effect of ozone production in the gas phase. For the OH radical, a similar diurnal pattern is seen, but the maximum is more pronounced, and occurs at midday. At night the production of OH is inhibited, and the remaining OH is rapidly consumed. The gaseous OH radical concentration in the fog is markedly reduced due to the initial transport of OH into the droplet. The gas-phase production during daytime will however still cause an OH peak at midday. The SO 2 concentration in the gas phase outside the fog is, due to the emissions applied in the model, continuously increasing. In contrast to 03 and OH, no diurnal pattern is observed for SO 2. The behaviour of SO2 in the fog strongly differs from outside the fog. The initial transport into the droplet will immediately reduce the SOz(g), after which the behaviour of SO2 is strongly dependent on the H202 concentration, via the aqueous three-body reaction: HSO~ + H 2 0 z + H ÷ --* SO 2- + 2H + + H 2 0 . (17) At the end of the simulation there is only a very small amount of H202 left, and the consequence is an increasing trend of SO2.

4. COMPARISONOF MODELLED AND MEASUREDDATA The modelled total concentrations of mercury presented in Fig. 2 range from about 5 × 1 0 -12 to 5× 10-11M. Most of this is mercury adsorbed to particles and the dissolved divalent complexes reach

a maximum of about 5 x 10-12M during the simulations. No data are available on measured concentrations of mercury in fog or cloud water. In precipitation, however, total mercury is routinely monitored within the Swedish environmental monitoring program. Typical monthly averages of mercury in precipitation are in the range 4 x 10-11 to 2 x 10-10 M at monitoring sites on the Swedish west coast (Munthe, 1994). The weighted annual average is around 5 x 10-11 M, which is in good agreement with the model results. It is not possible to measure the concentrations of the individual divalent complexes due to the low concentrations. However, some analytical data exist of measurements of an operationally defined mercury fraction, "reactive Hg" or "HglIa", in precipitation believed to mainly consist of dissolved mercury complexes. Iverfeldt (1991) reported an average value of about 1.93 ng ~ - ~ (9.5 x 10-12 M) for the Swedish west coast during 1985-1989. Similar values (1-2.5 x 1 0 - 1 1 M ) have also been reported from measurements in the U.S.A. (Bloom and Watras, 1989). These measured concentrations are somewhat higher than the modelled values but the difference is not great considering uncertainties of the nature of the defined mercury fractions.

5. SENSITIVITYTESTSOF DISSOLVEDHg(ll) TO SULPHUR DIOXIDE, OZONE, CHLORIDE AND SOOT CONCENTRATION The CAM model, described in the previous section, has been used for a sensitivity test covering initial sulphur dioxide, ozone and chloride concentrations as well as soot concentration. 5.1. Sulphur dioxide concentration---sensitivity test The initial value of gaseous sulphur dioxide has been changed from 0.5 to 50 ppb, and the resulting dissolved Hg(II) formed is illustrated in Fig. 7. A higher start value of SO2 leads to lower concentrations of dissolved Hg(II) as shown in Fig. 7. The

Chemistry in fog droplets

.....

I

I

t

I

I

t

I

t

20

40

60

80

100

120

140

160

0.9 -t 0.8 ]

1449

Ozone

180

2OO

(Pl~t)

•

I"1

0.7

~

O.S 0.4

t

0

•

[]

0.3 -t

0 0.2 4

E] []

0.1 O~ 0

0

0 I

I

I

I

I

I

I

t

I

5

10

15

20

25

30

35

40

45

50

so= ( ~ )

Fig. 7. Relation between dissolved Hg(II), initial SOz and O3 concentration. Filled squares (ll) mark the Hg(II) concentration when the SO2 concentration, read on the lower x-axis, is changed. The relative unit 1 correspond,'; to 6.2' 10-12 M in this case. Open squares ([]) mark the Hg(II) concentration when the O3 concentration, read on the upper x-axis, is changed, and unit 1 corresponds to 3.8.10-11 M in this case.

reduction is caused by the increase in SO~ -, favouring the complex HgSO3, which will spontaneously reduce back to Hg °. This chain of processes leads to a total decrease of dissolved Hg(II) in the droplet. In the lower concentration range (SO2 = 0.5 to 10ppb), a change of the SO~ level leads to a greater change in Hg(II) concentration, compared to in the higher concentration range (SO 2 = 20 to 50 ppb). Higher SO 2 levels are seen to cause a higher percentage of HgCI 2 and a lower percentage of HgOHCI. This can be explained by the lower pH resulting from oxidation of SO2, as described above. A lower pH gives less O H - , thus leading to less HgOH + and HgOHCL This pH effect dominates over the increased concentration of SO~- which would otherwise lead to more HgSO 3 and Hg(SO3)~-. The opposite reasoning is true for the lower SO2 values, where a higher pH gives more O H - , and thus more formation of HgOHC1. 5.2. Ozone concentration---sensitivity test For ozone, the initial values were changed in the range 10-200 ppb in the sensitivity test, The relation between ozone concentration and dissolved Hg(II) is almost linear, and is illustrated in Fig. 7. A higher initial value of ozone thus causes a higher concentration of dissolved Hg(II), due to an increased oxidation rate of Hg °. Higher 0 3 gives a higher percentage of HgCI 2 and a lower percentage of HgOHCI. This can, as in the experiment of changing the SO 2 above, be explained by a lower pH caus.ed by a higher 03 concentration. The influence of 03 on pH, goes via an increased H 2 0 2 level at a higher Oa concentration. The pH will then decrease through the aqueous three-body reacAE 29*1Z-H

tion: HSO 3 + H 2 0 2 -k H + --, SO 2- + 2H + + H 2 0

(17)

which is more important than the direct reaction of 0 3 and S(IV). 5.3. Chloride concentration--sensitivity test The initial chloride concentration was changed between 1.10 -7 and 1.10 -3 M. Due to enhanced production of the complex HgC12, the total Hg(II) concentration will increase rapidly, especially in the range 1"10-6 < CI- < l ' 1 0 - 4 M , as seen from Fig. 8. 5.4. Soot concentration--sensitivity test The influence of different soot concentration levels on Hg(II) in the dissolved and adsorbed form in the droplet has been studied over a soot concentration range 0.5-100/zgm 3. As seen from Fig. 9, the total amount of Hg(II), (dissolved Hg(II) plus adsorbed Hg(II)), will increase with increasing soot concentration. The adsorbed part will increase with increasing soot concentration, and at the highest soot concentration studied, almost all of the Hg(II) in the droplet will be found in the adsorbed phase. At the lowest soot concentration, most of the Hg(II) will be dissolved, while the partitioning between dissolved and adsorbed Hg(II) will be 50/50 at soot concentration 1.0/lg m -3 .

6. D I S C U S S I O N

The aim of this work has not been to accurately simulate a specific measurement event, but rather to advance the understanding of the chemistry of mercury in fog droplets. The box-model used in this study

1450

K. PLEIJEL and J. MUNTHE

•

•

•

I

I 0.9 0.8 - 0.7 "~ 0.6 d O ¢.)

0.5 o -0.4

I 1.00E-06

I 1.00E-05

I I .OOE-04

0.3

-

0.2

-0.1 0

n

I 1.00E-07

-

I 1.00E-03

CI- (M) Fig. 8. Relation between dissolved Hg(II) and initial CI- concentration. Unit 1 corresponds to 7.2-10-12 M.

4.50E-11 •

Diss Hg(ll)

n ads Hg(II)

4.00E- 1I

~,~.,,,-"~/

3.50E- I I -

3.00E-I I

- 2.50E-11 -

2.00E-I 1

-

1.50E-I I

-

1.00E-I I

- 5.00E-I 2 I 1.00E-07

1.00E-06

1.00E-05

1.001]-04

0.00E+00 1.00E-03

Soot concentration (g/m3) Fig. 9. Relation between dissolved, adsorbed and total Hg(II) with various soot concentrations in the range 0.5 to 100 #gm -3.

serves as a c o m p l e m e n t to larger-scale models which include m u c h more d y n a m i c processes. The results can be used as a base resource for the d e v e l o p m e n t of condensed chemical schemes t h a t are d e m a n d e d from Eulerian modellers. T h e variations of predicted mercury c o n c e n t r a t i o n s with varying chemical composition was reported in this paper. O f equal i m p o r t a n c e is to investigate the variations of predicted mercury c o n c e n t r a t i o n s with varying physical parameters such as rate of c o n d e s a t i o n / e v a p o r a t i o n , t e m p e r a t u r e a n d sticking coefficients. These questions were not addressed in the present study b u t is currently being investigated a n d will be reported at a later stage. Acknowledgements--This work has been financially supported by the Swedish Environmental Protection Board (SNV) and ELFORSK AB. The authors would also like to

thank Prof. Oystein Hov, Dr. JLke Iv¢ffeldt, Dr. Evert Ljungstr6m and Prof. Oliver Lindqvist for their valuable comments and recommendations.

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Chemistry in fog droplets Chameides W. L. (1986) Possible role of NO a in the nighttime chemistry of a cloud. J. geophys. Res. 91, 5331-5337. Chameides W. L. and Davis D. D. (1982) The free radical chemistry of cloud droplets and its impact upon the composition of rain. J. ffeophys. Res. 87, 4863-4878. Clark A. I., Intyre A. E. M., Lester J. N. and Perry R. (1984) Ambient air measurements of aromatic and halogenated hydrocarbons at urban, rural and motorway locations. Sci. tot. Envir. 39, 265-279. Clever H. L., Johnson S. A. and Derrick M. E. (1985) The solubility of mercury and some sparingly soluble mercury salts in water and ~Lqueous electrolyte solutions. J. phys. chem. Ref Data 14, 631-680. Cotton F. A. and Wilkinson G. (1980) Advanced Inorganic Chemistry. A Comprehensive Text, 4th Edn. Wiley, New York. Curtis A. R. and Sweetenham W. P. (1987) FACSIMILE/ CHEKMAT User's manual. AERE R 12805, Harwell Laboratory, Oxfordshixe, U.K. Deister U., Neeb R., Helas G. and Warneck P. (1986) Temperature dependence of the equilibrium CHz(OH) e + HSO~ = CH2(OH) SO~- + H20 in aqueous solution. J. phys. Chem. 90, 3213-3217. Derwent R. G. and Hov O. (1979) Computer modelling studies of photochemical air pollution formation in North West Europe. AERE R-9434, Harwell Laboratory, Oxfordshire, U.K. Dodge M. C. (1990) Formaldehyde production in photochemical smog as predicted by three state-of-the-science chemical oxidant mechanisms. J. geophys. Res. 95, 3635-3648. Eigen M. and Wilkins R. G. (1965) The kinetics and mechanism of formation of metal complexes. ACS Series 49, American Chemical Society, Washington D.C., U.S.A. Faust B. C. and Hoign6 J. (1990) Photolysis of Fe (III)hydroxy complexes as sources of OH radicals in clouds, fog and rain. Atmospheric Environment 24A, 79-89. Ferm M. (1991) IVL, Box 470 86, 402 58 G6teborg, Sweden, personal communication. Ferm M., Samuelsson U., Sj6din A. and Grennfelt P. (1984) Long-range transpc,rt of gaseous and particulate oxidised nitrogen compounds. Atmospheric Environment 18, 1731-1735. Finlayson-Pitts B. J..and Pitts J. N. Jr (1986) Atmospheric Chemistry. Fundamentals and Experimental Techniques. Wiley, New York. Fitzgerald W. F. (19:86) Cycling of mercury between the atmosphere and oceans. In The Role of Air-Sea Exchange in Geochemical Cycling (edited by Buat-Menard P.), NATO Advanced Institute Series, pp. 363-408. Reidel, Dordrecht, The Netherlands Fitzgerald W. F., Mason R. P. and Vandal G. M. (1991) Atmospheric cycling and air-water exchange of mercury over mid-continenlal lacustrine regions. War. Air Soil Pollut. 56, 745-767. Gery M. W., Whitten G. Z., Killus J. P. and Dodge M. C. (1989) A photochemical kinetic mechanism for urban and regional scale computer modelling, d. geophys. Res. 94, 12,925-12,956. Graedel T. E. and Goldberg K. I. (1983) Kinetic studies of raindrop chemistry. 1. Inorganic and organic processes. J. geophys. Res. 88, 10,865-10,882. Graedel T. E., Mandich M. L. and Weschler C. J. (1986) Kinetic model studies of atmospheric droplet chemistry. 2. Homogeneous transition metal chemistry in raindrops. J. oeophys. Res. 91, 5205--5221. Hall B. (1995)The gas phase oxidation of elemental mercury by ozone. War. Air Soil Pollut. (in press). Hoffmann M. R. (1986) On the kinetics and mechanism of oxidation of aquated sulphur dioxide by ozone. Atmospheric Environmenl~ 20, 1145-1154. Huie R. E. and Neta P. (1987) Rate constants for some oxidations of S(IV) by radicals in aqueous solutions. Atmospheric Environment 21, 1743-1747.

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Iverfeldt A. (1991) Occurrence and turnover of atmospheric mercury over the Nordic Countries. Wat. Air Soil Pollut. 56, 251-265. Iveffeldt A. and Lindqvist O. (1980) Determination of distribution equilibria between water and air for certain mercury compounds. Report No. 415, The Coal Health Environment Project, The Swedish State Power Board, S-162 78 V/illingby, Sweden. Iverfeldt A. and Lindqvist O. (1986) Atmospheric oxidation of elemental mercury by ozone in the aqueous phase. Atmospheric Environment 20, 1567-1573. Jacob J. D. (1986) Chemistry of OH in remote clouds and its role in the production of formic acid and peroxymonosulfate. J. oeophys. Res. 91, 9807-9826. Lee Y. -N., Klotz P. J., Schwartz S. E. and Newman L. (1986) Kinetics of hydrogen peroxide-sulphur(IV) reaction in rainwater collected at a Northeastern U.S. site. J. oeophys. Res. 91, 13,264-13,274. Lind J. A., Lazrus A. L. and Kok G. L. (1987) Aqueous phase oxidation of sulphur(IV) by hydrogen peroxide, methylhydroperoxide, and peroxyacetic acid. J. oeophys. Res. 92, 4171-4177. Lindqvist O. and Rodhe H. (1985) Atmospheric mercury--a review. Tellus 37B, 136-159. Lindqvist O., Johansson K., Aastrup M., Andersson A,, Bringmark L., Hovsenius G., H~tkanson L., Iveffeldt A., Meili M. and Timm B. (1991) Mercury in the Swedish environment--Recent research on causes, consequences and corrective methods. War. Air Soil Pollut. 55, 261. Lindskog A. and Moldanova J. (1994) The influence of origin, season, and time of the day on the distribution of individual NMHC measured at R6rvik, Sweden. Atmospheric Environment 28, 2383-2398. L6vblad G. and Sj6berg K. (1989) Monitoring of air and precipitation at the IVL PMK-stations. Yearly report 1988 (in Swedish, English summary). SNV Report 3645, 171 85 Solna, Sweden. L6vblad G., Sj6berg K., Ehrencrona C. H. and Peterson K. (1991) Monitoring of air and precipitation at the IVL PMK-stations. Yearly report 1990 (in Swedish, English summary). SNV Rapport 3940, 171 85 Solna, Sweden. L6vblad G., Sj6berg K., H~tllinder C. and Peterson K. (1990) Monitoring of air and precipitation at the IVL PMKstations. Yearly report 1989 (in Swedish, English summary). SNV Report 3787, 171 85 Solna, Sweden. Munthe J. (1991) The redox cycling of mercury in the atmospheJe. Thesis, University of G6teborg and Chalmers University of Technology, G6teborg, Sweden. Munthe J. (1992) The aqueous oxidation of elemental mercury by ozone. Atmospheric Environment 26A, 1461-1468. Munthe J. (1993) Mercury in the atmosphere. Emissions, transformations, deposition and effects. IVL-report B 1110, G6teborg, Sweden. Munthe J. (1994) unpublished data at IVL, G6teborg, Sweden. Munthe J. and Iverfeldt/~. (1993) Wet deposition of methylmercury in Sweden. Proc. from the 1993 EPA/A& W M A international symposium, Measurement of Toxic and Related Air Pollutants., Durham, NC, May 3-7. Munthe J., Xiao Z. F. and Lindqvist O. (1991) The aqueous reduction of divalent mercury by sulphite. War. Air Soil Pollut. 56, 621-630. M611er L. and Jonsson A. (1985) Oxygenated organic species in urban air (in Swedish). SNV-report 3005, 171 85, Solna, Sweden. M611er D. and Mauersberger G. (1992) Cloud chemistry effects on tropospheric photooxidants in polluted atmosphere-model results. J. atmos. Chem. 14, 153-165. Nahir T. M. and Dawson G. A. (1987) Oxidation of sulphur dioxide by ozone in highly dispersed water droplets. J. atmos. Chem. 5, 373-383. Neta P., Huie R. E. and Ross A. B. (1988) Rate constants for reactions of inorganic radicals in aqueous solution. J. phys. Chem. 17, 1027-1284.

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Pacyna J. M. (1989) Mercury emission inventories for Europe. Paper presented at the Workshop on Modelling and Atmospheric Transport and Deposition of Mercury, GKSS, Geestacht, Germany, Sept. 18-19, 1989. Petersen G., Iverfeldt/k. and M u n t h e J. (1995) Atmospheric mercury species over central and northern Europe. Model calculations and comparison with observations from the Nordic air and precipitation network for 1987 and 1988. Atmospheric Environment 29, 47-67. Pleijel K. and Andersson-Sk61d Y. (1992) A comparison of different chemical models for use in regional simulations of air pollution (in swedish). IVL-report B 1048, IVL, Box 470 86, 402 58 G6teborg, Sweden. P~ankov V. A. (1949) Kinetics of the reaction between mercury vapor and ozone. Zhur. Obshchev., Khim. J. Gen. Chem. 15, 224. SCB (1989). Statistical book of the year 1989 (in swedish). Statistiska Centralbyr~m, Stockholm, Sweden. Schwartz S. E. (1986) Mass-transport considerations pertinent to aqueous phase reactions of gases in liquid-water clouds. In Chemistry of Multiphase Atmospheric Systems (edited by Jaeschle W.). N A T O ASI Series G6. Springer Berlin Heidelberg, Germany. Seigneur C. and Saxena P. (1984) A study of atmospheric acid formation in different environments. Atmospheric Environment 18, 2109-2124. Shah J. J. and Singh H. B. (1988) Distribution of volatile organic chemicals in outdoor and indoor air. Envir. Sci. Technol. 22, 1381-1388. Singh H. B., Salas L. J., Cantrell B. K. and Redmond R. M. (1985) Distribution of aromatic hydrocarbons in the ambient air. Atmospheric Environment 19, 1911-1919. Slemr F., Schuster G. and Seiler W. (1985) Distribution, speciation and budget of atmospheric mercury. J. atmos. Chem. 3, 407--434.

Smith R. M. and Martell A. E. (1976) Critical Stability Constants. Vol. 4: Inorganic Complexes. Plenum, New York. Stratton W. J. and Lindberg S. E. (1995) Use of a refluxing mist chamber for measurements of gas-phase watersoluble mercury(II) species in the atmosphere. Wat. Air Soil Pollut. (in press). Sturges W. T. and Harrison R. M. (1989) The use of nylon filters to collect HCI: Efficiencies, interferences and ambient concentrations. Atmospheric Environment 23, 1987-1996. Warneck P. (1988) Chemistry of the Natural Atmosphere. Academic Press, 24-28 Oval Road, London N W l 7DX, U.K. Weschler C. J., Mandich M. L. and Graedel T. E. (1986) Speciation, photosensitivity, and reactions of transition metal ions in atmospheric droplets. J. geophys. Res. 91, 5189-5204. Xiao Z. F., M u n t h e J. and Lindqvist O. (1991b) Sampling and determination of gaseous and particulate mercury in the atmosphere using gold coated denuders. Wat. Air. Soil Pollut. 56, 141-151. Xiao Z. F., M u n t h e J., Schroeder W. H. and Lindqvist O. (1991a) Vertical fluxes of volatile mercury over forest soil and lake surfaces in Sweden. Tellus 43B, 267-279. Xiao Z. F., M u n t h e J., Str6mberg D. and Lindqvist O. (1994) Photochemical behaviour of inorganic mercury compounds in aqueous solution. In Mercury as a Global Pollutant-toward Integration and Synthesis (edited by Watras C. J. and Huckabee J. W.). Lewis, in press. Yarwood G. and Niki H. (1990) A critical review of available information on transformation pathways for the mercury species in the atmospheric environment. Report prepared for Atmospheric Environment Service, Environment Canada, Downsview, Ontario, Canada.

APPENDIX

Table A1. Chemical gas-phase reactions used in all calculations (Gery et al., 1989; Dodge, 1990). kf=J<0> kf = 5.8 D - 3 4 . M * (TEMP/300.0) @ - 2.6 kf = 1.8 D - 1 2 * E X P ( - 1370./TEMP) kf = 9.3 D - 12 kf = 9.0D - 3 2 . M k f = 1.1 D - 3 1 * M * ( T E M P / 3 0 0 . 0 ) @ - 1.8 kf = 1 . 2 D - 1 3 * E X P ( - 2450./TEMP) kf=J<7> kf=J kf = 2.1 D - 11*M*EXP(100.0/TEMP) kf = 2.2 D - 10 k f = 1 . 9 D - 1 2 * E X P ( - 1000./TEMP) k f = l A D - 14*EXP( - 600./TEMP) kf = 33.9.J < 0 > kf = 1.7 D - 11*EXP( - 150/TEMP) kf = 7.2 D - 14*EXP( - 1414./TEMP) k f = 1 . 1 6 D - 12 kf= 1.3D-21 kf = 2.439 D - 2 kf = 1.7 D - 20* EXP (530./TEMP) kf = 8.2 D - 3 9 . G H 2 0 kf = 4 . 0 D - 12 kf=J<8> kf = 6.6D - 12 k f = 1.6D - 24 k f = 1 . 5 D - 11 kf = 9.4D - 15*EXP(778./TEMP) kf = 3.7D - 12*EXP(240./TEMP)

: GNO2 = GNO + GO : GO + GO2 = GO3 : GO3 + GNO = GNO2 : GO + GNO2 = GNO : GO + GNO2 = GNO3 : GO + GNO = GNO2 : GO3 + G N O 2 = G N O 3 : GO3 = GO : GO3 = GOD : GOD = GO : GOD + GH20 = GOH + GOH : GO3 + G O H = G H O 2 : GO3 + G H O 2 = G O H : G N O 3 = , *0.89:GNO2, *0.89:GO, *0.11 : G N O : GNO + GNO3 = GNO2 + GNO2 : GNO2 + GNO3 = GNO + GNO2 : GNO2 + GNO3 = GN205 : GN205 + GH20 = GHNO3 + GHNO3 : GN205 = GNO2 + GNO3 : GNO + GNO = GNO2 + GNO2 : GNO + GNO2 = GHONO + GHONO : GNO + GOH = GHONO : GHONO = GNO + GOH : GOH + GHONO = GNO2 : GHONO + GHONO = GNO + GNO2 : GNO2 + GOH = GHNO3 : GOH + GHNO3 = GNO3 : GNO + GHO2 = GNO2 + GOH

C h e m i s t r y in f o g d r o p l e t s

1453

T a b l e A I . (Continued) k f = 1 . 7 D - 12 kf = 6.8D + 13*EXP(10200./TEMP) k f = 1.3 D - 1 2 * E X P ( 3 8 0 . / T E M P ) k f = 5.9 D - 1 4 * E X P ( l l 5 0 . / T E M P ) k f = 2.1 D - 3 8 , G H 2 0 * E X P ( 5 8 0 0 . / T E M P ) kf-J<9> kf= 2.9D12*EXP(160./TEMP) k f = 2 . 2 D - 13 k f = 1 . 1 D - 11 kf=J<5> kf=J<6> kf= 2.9D- II*EXP(1550./TEMP) k f --- 5.4 D - 16 kf--- 1 . 2 D - l l * E X P ( - 9 8 6 . / T E M P ) kf = 6.9D - 12*EXP)250./TEMP) k f = 2 . 5 D - 15 kf=J<10> k f --- 4.2 D - 1 2 * E X P ( 1 8 0 . 0 / T E M P ) k f = 4.7 D - 12 k f = 1.95 D + 1 6 * E X P ( 13543./TEMP) kf-- 2.5D12 k f = 6 . 4 E - 12 kf = 1.1D + 2*EXP( - 1710./TEMP) kf= 8.0D13 kf---- 1 . 0 D + 1 5 * E X P ( - 8 0 0 0 . / T E M P ) kf = 1.6D + 3 k f = 1 . 5 D - 11 k f = 1.2 D - 1 1 * E X P ( - 3 2 4 . / T E M P ) k f = 5.2 D - 1 2 * E X P ( 5 0 4 . / T E M P ) k f = 1.4 D - 1 4 . EX1a ( - 2 1 0 5 . / T E M P ) kf= 7.6D-

15

kf = 1.0D - II*EXP(

- 792./TEMP)

kf= 1.66D- 12*EXP(474./TEMP) kf = 1.2D- 14*EXF(- 2633./TEMP) k f = 2.1 D - 1 2 * E X P ( 3 2 2 . / T E M P ) k f = 8 . 0 D - 12 k f = 4.2 kf=4.1D11 kf= kf= kf = kf=

2.2D11 1 . 3 D - 11 3.75,J < 5 > 2.9D11

kf = 5.4D - 17*EXP( - 500./TEMP)

kf= 1.6D-

II*EXP(II6./TEMP)

kf= kf = kf = kf= kf = kf =

11 5 > 12 14,EXP(1300./TEMP) 13 12

1.7D4.0*J < 8.1 D 1.7D6.7 D 6.0D -

: GNO2 + GHO2 = GHO2NO2 : GHO2NO2 = GNO2 + GHO2 : GOH + GHO2NO2 = GNO2 : GHO2 + GHO2 = GH202 : GHO2 + GHO2 = GH202 : GH202 = GOH + GOH : GOH + GH202 = GHO2 : GOH + GCO = GHO2 : GOH + GHCHO = GHO2 + GCO : GHCHO = GHO2 + GCO + GHO2 : GHCHO = GCO : GHCHO + GO = GOH + GHO2 + GCO : GNO3 + GHCHO = GHNO3 + GCO + GHO2 : GALD2 + GO = GCH3COO2 + GOH : GALD2 + GOH = GCH3COO2 : GALD2 + GNO3 = GCH3COO2 + GHNO3 : GALD2 = GHCHO + GXO2 + GCO + GHO2 + GHO2 : GCH3COO2 + GNO = GNO2 + GXO2 + GHCHO + GHO2 : GCH3COO2 + GNO2 = GPAN : GPAN = GCH3COO2 + GNO2 : G C H 3 C O O 2 + G C H 3 C O O 2 --- G H C H O + G H C H O + G X O 2 + GXO2 + GHO2 + GHO2 : G C H 3 C O O 2 + G H O 2 = , *0.79: G H C H O , . 0 . 7 9 : G X O 2 , *0.79: G H O 2 , *0.79 : G O H : GOH = GXO2 + GHCHO + GHO2 : G P A R + G O H = ,*0.87: G X O 2 , * 0 . 1 3 : G X O 2 N , *0.11 : G H O 2 , *0.11 : G A L D 2 , * 0 . 7 6 : G R O R , * - 0.11 : G P A R : G R O R = ,*1.1 : G A L D 2 , * 0 . 9 6 : G X O 2 , * 0 . 9 4 : G H O 2 , * 0 . 0 4 : G X O 2 N , *0.02: G R O R , * - 2.1 : G P A R : GROR = GHO2 : GROR + GNO2 = : G O + G O L E = ,*0.63: G A L D 2 , *0.38: G H O 2 , *0.28: G X O 2 , *0.3: G C O , *0.2: G H C H O , *0.02: * G X O 2 N , *0.22 : G P A R , *0.2: G O H : GOH + GOLE = GHCHO + GALD2 + GXO2 + GHO2, * 1.0: G P A R : GO3 + GOLE =, *0.5:GALD2, *0.74:GHCHO, *0.33:GCO, * 0 . 4 4 : G H O 2 , * 0 . 2 2 : G X O 2 , *0.1 : G O H , * - 1 . 0 : G P A R : GNO3 + GOLE = GHCHO + GALD2 + GNO2, .0.91 :GXO2, *0.09: G X O 2 N , * - 1.0: G P A R : GO + GC2H4 = GHCHO + GCO, *0.7:GXO2,. 1.7:GHO2, *0.3: G O H : G C 2 H 4 + G O H = G X O 2 + G H O 2 , * 1.56: G H C H O , *0.22: G A L D 2 : G C 2 H 4 + G O 3 = G H C H O , *0.42: G C O , * 0 . 1 2 : G H O 2 : G O H + G T O L = , *0.08: G X O 2 , *0.36: G C R E S , *0.44: G H O 2 , *0.56:GTO2 : G T O 2 + G N O = , *0.9: G N O 2 , *0.9: G H O 2 , F*0.9 : G O P E N : GTO2 = GCRES + GHO2 : G O H + G C R E S = , , 0 . 4 : G C R O , *0.6: G X O 2 , *0.6: G H O 2 , .0,3 :GOPEN : GCRES + GNO3 = GCRO + GHNO3 : GCRO + GNO2 = : GOPEN = GCH3COO2 + GHO2 + GCO : GOPEN + GOH = GXO2 + GCO + GCO + GHO2 + GHO2 + GCH3COO2 + GHCHO : G O P E N + G O 3 = , *0.03: G A L D 2 , *0.62: G C H 3 C O O 2 , *0.7: G H C H O , *0.03: G X O 2 , *0.69: G C O , *0.08 : G O H , *0.76: G H O 2 , *0.2: G M G L Y O X : G O H + G X Y L = , *0.7 : G H O 2 , *0.5: G X O 2 , *0.2 : G C R E S , *0.8: G M G L Y O X , * 1.1 : G P A R , *0.3: G T O 2 : GMGLYOX + GOH = GCH3COO2 + GXO2 : GMGLYOX = GCH3COO2 + GCO + GHO2 : GXO2 + GNO = GNO2 : GXO2 + GXO2 = : GXO2N + GNO = :'GXO2 + GH02 =

T h e r e a c t i o n c o n s t a n t s h a v e b e e n u p d a t e d a c c o r d i n g to Pleijel a n d A n d e r s s o n - S k 6 1 d (1992). P h o t o l y s i s rates, J < 0 > to J < 18 > , a r e s h o w n in T a b l e A2. All species s t a r t w i t h a G to i n d i c a t e the g a s p h a s e , a n d the sign = is u s e d f o r a r i g h t h e a d i n g a r r o w ( - - , ) . T h e n o t a t i o n o t h e r w i s e ' f o l l o w s G e r y et al. (1989). @ r e p r e s e n t s a n e x p o n e n t i a l f u n c t i o n i.e. X @ Y = X y.

1454

K. P L E I J E L a n d J. M U N T H E T a b l e A2. C a l c u l a t i o n of p h o t o l y s i s r a t e s a n d p h o t o l y s i s r a t e c o n s t a n t s J

=C*A*EXP(B*SEC)

where J is the p h o t o l y s i s r a t e in u n i t (s - 1) 0 < N R < 18 a c c o r d i n g to the r e a c t i o n in q u e s t i o n C is a f a c t o r r e d u c i n g (or e n h a n c i n g ) the s o l a r r a d i a t i o n w i t h i n a fog S E C = I / S I N ( 0 ) , w h e r e 0 is the s o l a r angle, defined b e l o w SIN(0) = sin(lat)'sin(decl) + c o s ( l a t ) ' c o s ( d e c l ) . c o s ( L H A ) , w h e r e lat is latitude, decl is the d e c l i n a t i o n a n g l e for a given d a y , L H A is the local t i m e a n g l e . A a n d B are c o n s t a n t s defined for r e a c t i o n 0 to 18 below: A<0> A A<2> A<3> A<4> A<5> A<6> A<7> A<8> A<9> A<10> A < 11 > A< 12> A<13> A<14> A< 15> A<16> A< 17> A< 18>

= 1.4500E-02 =2.0000E-04 =3.3237E-05 =3.5270E-02 =8.9387E-02 =5.4000E-05 =6.6500E-05 =1.2300E-03 =2.8000E-03 =2.2000E-05 =1.3500E-05 =2.6200E-04 =6.3300E-07 =7.2600E-07 =2.1100E-04 = 1.5800E-04 =2.5600E-05 =5.0200E-07 =3.3500E-02

B<0> B< 1 > B<2>= B<3>= B<4>= B<5>= B<6>= B<7>= B<8>= B<9>= B<10> B < 11 > B<12> B<13> B<14> B< 15> B<16> B< 17> B<18>

= -4.0000E-01 = - 1.40000E + 00 --5.6640E--01 -8.0965E-02 -5.9174E-02 -7.9000E-01 -6.0000E-01 -6.0000E-01 -4.5000E-01 -7.5000E-01 = -9.4000E-01 = -6.0000E-01 = -6.0000E-01 = -8.4900E-01 = -3.4500E-01 = -6.7200E-01 = -7.0300E-01 = -9.2000E-01 = -2.8000E-01

T a b l e A3. A q u e o u s - p h a s e c h e m i s t r y u s e d in the m o d e l . S o m e o f the r e a c t i o n s a r e the s a m e as in M611er a n d M a u e r s b e r g e r (1992). T h e n o t a t i o n 1 D - 4 o r 1E-4 m e a n s 1 x 10 - 4 , N A V is A v o g a d r o ' s c o n s t a n t (6.022045 x 1023 m o l e c u l e s m o l - l ) . @ r e p r e s e n t s a n e x p o n e n t i a l f u n c t i o n i.e. X @ Y = X r R a t e coefficients

Reaction/equilibrium

Ref.

kf kf kf kf kf

: : : :

CO2 = H2CO3 H20 = HIP + OH1M H2CO3 = HCO31M + HIP HCO31M = CO32M + HIP

CW80 J86 CW80 G86

: : : : : :

NH3 = NH41P + OH1M HO2 = O21M + HIP FE3P + OH1M = FEOH2P FEOH2P + OH1M = FEOH21P FEOH21P + OH1M = FEOH3 HCOOH = HCOO1M + HIP

C84 C84 G86 G86 W86 C84

= 0.03, k b = 20 = 1D-4, k b = 1 D 1 0 ( N A V * I D - 3 ) = 4.16D3, k b = 1 D 1 0 ( N A V * I D - 3 ) = 4.84D-1, k b = 1 D 1 0 ( N A V * I D - 3 ) = 1 D I 0 * I . 7 D - 5 * E X P ( - 4325* (1/TEMP-1/298)), kb = 1D10/(NAV*ID-3) kf = 2 D 5 , k b = 1 D 1 0 / ( N A V * I D - 3 ) k f = I D 1 0 / ( N A V * I D - 3 ) , k b = 1,54D-2 kf = 1D10/(NAV*ID-3), kb = 1D10/3.08D10 kf = 1D10/(NAV*ID-3), kb = 1D10/6.29D14 kf = 1D10*l.8D-4, kb = 1D10/(NAV*ID-3) k f = 1 D 1 0 * 1 0 @ ( - 4.7) kf = 1D10/(NAV*ID-3) kf = 6.25D2, k f = 6 . 2 5 D 2 / 1 . 3 1 6 D - 2 / ( N A V * 1D-3) k f = 6.25D2, k b = 1 D 1 0 / ( N A V * I D - 3 ) kf= 1D10/(NAV*ID-3), kb = 1D10/2D5 kf = 1.58E-12*lE10, kb = 1E10/(NAV*IE°3 kf = 5.1E-4*IE10, kb = 1E10/(NAV*IE-3) kf= J(15) kf = J(16> kf = 1 E 9 / ( N A V * I E - 3 ) kf = 1E10/(NAV*IE-3) kf = 4.6E3/(NAV*lE-3)@2 kf = 5E5/(NAV*IE-3) kf = 1.2E9/(NAV*IE-3) k f = 2 . 5 E 8 / ( N A V * 1E-3) kf = 1.3E9/(NAV*IE-3) kf = IE8/(NAV*IE-3)

: CH3COOH

= CH3COO1M

+ HIP

: SO2 = HSO31M + HIP : HSO31M = SO32M + H1P : CL1M + CL = CL21M : H202 = HIP + HO21M : HONO = HIP + NO21M : HONO = NO + OH : NO21M = NO + OH + OH1M : HONO + OH = NO2 + H20 : NO21M + OH = NO2 + OH1M : HONO + H202 + HIP = NO31M + HIP + HIP + H20 : NO21M + 03 = NO31M + 02 : NO21M + NO3 = NO2 + NO31M : NO21M + CL21M = NO2 + CL1M + CL1M : NO2 + OH = NO31M + HIP : NO2 + NO2 = HONO + HIP + NO31M

C84 SM76 SM76 CD82 J86 J86 J86 J86 J86 J86 J86 J86 J86 J86 J86 J86

C h e m i s t r y in fog d r o p l e t s T a b l e A3, kf = kf = kf = kf = kf = kf = kf = kf = kf = kf = kf = kf= kf = kf = kf = kf = kf = kf = kf = kf = kf = kf =

3E8/(NAV*-3) J(17) J(18) 8.8E8/(NAV*IE-3) 8.5D7/(NAV*ID-3) 1.5D6/(NAV*ID-3) 2.8E6/(NAV*IE-3) 1.5E7/(NAV*IE-3) 9 . 1 E 6 / ( N A V * 1E-3) 4E8/(NAV*IE-31 8E5/(NAV*IE-31 J(13) J(14) 6D9/(NAV* ID-3) 7 D 9 / ( N A V * 1D-3) 1D10/(NAV*ID..3) 2D9/(NAV*ID-3) 1.5D9/(NAV*ID-3) 1 D 8 / ( N A V * I D -- 3) 4 . 5 D 7 / ( N A V * 1D-3) 7.2D7/(NAV*lD-3)@2 1.7D7/(NAV*ID-3)@2

kf = + kf = kf = kf = kf = kf=

(3.5D7*(H1P/(NAV*ID-3)) 610)/(NAV* I D - 3 ) 2.4D4/(NAV*113,-3) 3 . 2 D 5 / ( N A V * 113,-3) 1D9/(NAV*ID-3) 1.4D9/(NAV*ID-3) 1 D 8 / ( N A V * 1D-3)

kf kf kf kf kf kf kf kf kf

4.5D9/(NAV*ID-3) 5.2D9/(NAV*ID-3) 1.5D9/(NAV*ID-3) 6 D 8 / ( N A V * 1D-3) 2 . 5 D 4 / ( N A V * 1D-3) 7.5D4/(NAV*ID-3) 2 D 9 / ( N A V * 1D-3) 2 D 8 / ( N A V * 1D-3) 7.5D7/(NAV*lD-3)@2

= = = = = = = = =

k f = 3 / ( N A V * 1D-3) k f = 3 . 6 D 3 / ( N A V * 1D-3) k f = 4 . 2 / ( N A V * 1D-3) kf = 2D9/(NAV*ID-3) kf = 1.5D8/(NAV*ID-3) k f = 3 D 9 / ( N A V * 1D-3) k f = 2 . 1 D 5 / ( N A V * 1D-3) k f = 2 . 3 D 7 / ( N A V * 1D-3) kf = 4.6D4/(NAV*ID-3) kf = 1.2D9/(NAV*ID-3) kf = 2.0D9/(NAV* ID-3) kf= J(ll) kf = 3D8/(NAV*ID-3) k f = 6 0 0 / ( N A V * 1D-3) kf = 1.8D8/(NAV*ID-3) k f = 7 . 2 D 8 / ( N A V * 1D-3) kf = 1.1D8/(NAV*ID-3) kf = 7.3D4/(NAV*ID-3) k f = 8 D 9 / ( N A V * 1D-3) k f = 1 D 7 / ( N A V * 1D-3) k f = 1 . 7 D 4 / ( N A V * 1D-3) k f = v e r y fast kf= 1D8/(NAV*ID-3) kf = 2.33D-4 kf = 4.5D9/(NAV*ID-3) kf = 1D9/(NAV*ID-3) kf = 1.4D5/(NAV*ID-3) kf = 7.3D6/(NAV* ID-3) kf = 1.8D10/(NAV*ID-3), kb = 16/(NAV*ID-3)

1455

(Continued)

: NO2 + NO = H1P + HIP + NO21M + NO21M : NO31M = NO2 + OH + OH1M : NO3 = NO + 02 : SO41M + NO21M = SO42M + NO2 : HO2 + O21M = HO21M + 02 : O21M + HCO31M = HO21M + CO31M : HO21M + 03 = OH + O21M + 02 : OH + HCO31M = CO31M + H20 : SO41M + HCO31M = SO42M + H1P + CO31M : O21M + CO31M = 02 + HCO31M + OH1M : H202 + CO31M = HO2 + HCO31M : H202 = OH + OH : 03 = OH + OH : OH + OH = H202 : HO2 + OH = H20 + 02 : O H + O 2 1 M = O H I M q- 0 2 : OH + 03 = HO2 + 02 : 03 + O21M = OH + 02 + 02 + OH1M : HO2 + O21M = H202 + 02 + OH1M : H202 + OH = HO2 + H20 : HSO31M + H202 + HIP = SO42M + HIP + HIP + H20 : HSO31M + CH3OOH + HIP = SO42M + HIP + HIP + ROH

J86 J86 J86 J86 A84 A84 J86 J86 J86 J86 J86 CD82 CD82 CD82 J86 J86 J86 J86 J86 CD82 L86

: : : : : :

HSO31M + CH3COO2H = SO42M + HIP + CH3COOH SO2 + 03 = SO42M + H1P HSO31M + 03 = SO42M + HIP+ 02 SO32M + 03 = SO42M + 02 CH2OHSO31M + OH = SO31M + HCHO + H20 HSO31M + NO3 = SO41M + SO42M + NO31M + H1P + H1P : H S O 3 1 M + O H = S O 3 1 M q- H 2 0 : SO32M + OH = SO31M + OH1M : SO31M + 02 = SO51M : SO51M + SO51M = SO41M + SOIM + 02 : SO51M + HSO31M = SO31M + HSO51M : SO51M + HSO31M = SO42M + SO41M + HIP : SO41M + HSO31M = SO42M + SO31M : SO41M + CL1M = SO42M + CL : HSO51M + HSO31M +HIP = SO42M + SO42M + H1P ÷ H1P + H1P : HCHO + HSO31M = CH2OHSO31H : CH2OHSO31M + OH1M = SO32M + HCHO + H20 : CH2OH2 + SO32H = CH2OHSO31H + OH1M : CH2OH2 + OH = HCOOH + HO2 : HCOOH + OH = CO2 + HO2 + H20 : H C O O I M + O H = C O 2 + H O 2 4- O H 1 M : HCOOH + NO3 = NO31M +HIP + CO2 + HO2 : CH3COOH + OH = PROD + HO2 : CH3COOH + NO3 = PROD + NO31H : CH3CHO + OH = CH3CO + H20 : CH3CO + 02 = CH3COOH + HO2 : FEOH21P = FEOH1P + OH : FE2P + OH = FEOH1P : H202 + FE3P = HO2 + FE2P + H1P : O21M + FE3P = 02 + FE2P : O21M + FE2P = H202 + FE3P + OHIM + OH1M : O21M + MN2P = H202 + MN3P + OH1M + OHIM : H202 + MN3P = HO2 + MN2P + H1P : O21M - CU2P = CU1P + 02 : FE3P + CUIP = FE2P + CU2P : FE2P + MN3P = FE3P + MN2P : N205 = HNO3 + HNO3 : NO3 + CL1M = NO31M + CL : PAN = PROD + NO31M : CL21M + HO2 = CLIM + CL1M + HIP + 02 : CL21M + O21M = CLIM + CL1M + 02 : C L 2 M 4- H 2 0 2 = C L 1 M + C L I M + H I P + H O 2 : OH1M + CL21M = CL1M + CL1M + OH

L87 H86 ND87 ND87 J86

: CL + OHIM

N88

= CL1M + OH

L87

C86 HN87 HN87 HN87 HN87 HN87 HN87 HN87 HN87 HN87 G86 D86 D86 C84 C84 C84 C86 G86 G86 G86 G86 FH90 G86 G86 G86 G86 G86 G86 G86 G86 G86 J86 SS84 CD82 CD82 J86 J86

K. PLEIJEL and J. MUNTHE

1456

Table A3. (Continued) kf = kf = kf = kf = kf =

1.6D5, kb = 1.45D10/(NAV*lD-3)@2 3.1D9/(NAV*ID-3) 4.5D7/(NAV*ID-3) fast fast

: CL = H1P + CL1M + OH : CL + HO2 = HIP + CL1M + O2 : CL + H202 = HIP + CL1M + HO2 : HNO3 = NO31M + HIP : HCL = CL1M + HIP

N88 GG83 GG83

A84: Adewuyi et al. (1984). C84: Chameides (1984). C86: Chameides (1986). CD82: Chameides and Davis (1982). CW80: Cotton and Wilkinson (1980). D86: Diester et al. (1986). FH90: Faust and Hoign6 (1990). G86: Graedel et al. (1986). GG83: Graedel and Goidberg (1983). H86: Hoffmann (1986). HN87: Huie and Neta (1987). J86: Jacob (1986). L86: Lee et al. (1986). L87: Lind et al. (1987). ND87: Nahir and Dawson (1987). N88: Neta et al. (1988). SM76: Smith and Martell (1976). SS84: Seigneur and Saxena (1984). W86: Weschler et al. (1986). Table A4. Boundary concentrations used in the modelling Specie Gaseous species Hg ° CO2

03 HO 2 OH

H202 CO SO2 HC1 NO 2 NO HNO3 NO3 N20 5 PAN NH3 C2H4 Toluene Xylene HONO HO2NO2 CHaCOO 2 m-glyoxal O O(1D) Aldehydes 2 Parat~nes Olefines HCHO CH3OOH CH3C(O)OOH HCOOH CH3COOH CH3OH Aqueous species Fe 3+ Fe 2 + Mn 3 + Mn 2+ Cu 2+ Cu + pH

Concentration

Unit

Ref.

3 330 35 1.59 x 107 1.6 x 107

ngm -3 ppm ppb molec cm- 3 moleccm -3 ppb ppb ppb ppb ppb ppb ppb ppb ppb ppb ppb ppb ppb ppb ppb ppb molec cm-3 ppb molec cm-a molec cm - 3 ppb ppb ppb ppb ppb ppb ppb ppb ppb

1 2 3 4 4 5 6 3 7 3 3 6 6 4 8 2 9 9, 4 9, 4 6 4 4 4 4 4 9, 4 9, 4 9, 4 6 2 2 2 2

0.2 120 1.2 0.25 2.4 0.3 0.I 0.01 0.015 0.1 0.5 0.21 0.53 0.37 0.03 0.006 3.61 x 106 0.21 2 x 103 3 x 10- s 0.38 16.16 0.4 1 0.1 0.01 0.5 0.15 0.8

0.8 x 10 -6 M 0.1 × 10- 6 M 1 x 10-lo M 0.3 x 10-6 M 1 × 10 - l ° M 0.01 × 10 -6 M 4.52 - l°log[H +]

2 2 2 2 2 2

Comment

Average May-Sept Modelled value Modelled value

Average May-Sept Average May-Sept Average May-Sept

Modelled value Typical european mean

Modelled value Modelled value Modelled value Modelled value Modelled value

Typical Typical Typical Typical

European European European European

mean mean mean mean

Typical Typical Typical Typical Typical Typical

European European European European European European

mean mean mean mean mean mean

1. Iverfeldt (1991). 2. M611er and Mauersberger (1992). 3. L6vblad and Sj6berg (1989). L6vblad et al. (1990, 1991) at R6rvik (Swedish west coast). 4. Pleijel and Andersson-Sk61d (1992), modelled value at the Swedish west coast, for a trajectory crossing the Nordic Sea, from Great Britain to Sweden. 5. Ferm (1991). 6. Finlayson-Pitts and Pitts (1986). 7. Stnrges and Harrison (1989). 8. F e r m et al. (1984) and Finlayson-Pitts and Pitts (1986). 9. Lindskog and Moldanova (1994), Singh et al. (1985), Shah and Singh (1988), Clark et al. (1984), M6ller and Jonsson (1985). Toluene, xylene, aldehydes 2, parafines and olefines represent groups of organic species defined in Gery et al. (1989). The groups include a number of species each, and the concentrations of individual organic compounds are divided into these groups as described in Pleijel and Andersson-Sk61d (1992).

Chemistry in fog droplets Table A5. Dry deposition rates in cm s-1 Species

Deposition rate (cm s- 1)

HNO3 SO2 NO 2 03 H202 PAN*

2.0 0.5 0.15 0.5 0.5 0.2

* This value is used for PAN and species similar to PAN.

1457