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Mercury partitioning in a power plant plume and its influence on atmospheric removal mechanisms
000&6981/80/0201-0227
Atmospheric EnvironmentVol. 14, pp. 227-231. Q Pergamon Press Ltd. 1980. Printed in Great Britain.
W2.CW
MERCURY PARTITIONING IN A POWER PLANT PLUME AND ITS INFLUENCE ON ATMOSPHERIC REMOVAL MECHANISMS STEVEN
E.
LINLJBERG
Environmental Sciences Division, Oak Ridge National Laboratory, Oak Ridge, TN 37830, U.S.A. (First received 5 April 1979 and
inJina1 form
1 September 1979)
Abstract - Air samples were. collected isokinetically in the plume of a major coal-fired power plant during helicopter flights. The dominant form of mercury in the plume was elemental Hg vapor, ranging from 92 to 99% of the total Hg concentration in air. There was no evidence of significant gas to particle conversion during plume aging. The predominance of the vapor form is conducive to long-range transport and removal by precipitation scavenging.
VAPOR AND PARTICLE-ASSOCIATED MERCURY IN A POWER PLANT PLUME
exposure to airborne mercury have been insufficiently evaluated and may pose a long-term threat (National Research Council, 1978). Current atmospheric emissions of mercury are estimated to exceed releases to waterways by more than an order of magnitude, yet there are still no legislated ambient air quality standards for mercury (Harriss and Hohenemser, 1978). It is estimated that N 35% of the global atmospheric emission of mercury from man’s activities results from coal combustion, with the ratio of anthropogenic to natural emissions ranging over orders of magnitude on regional scales (Harriss, in press). The release of Hg vapor during coal combustion is well recognized with estimates of the fraction of the feed coal mercury discharged as a vapor ranging from 90 to 97% (Billings and Matson, 1972; Klein et al., 1975; Anderson and Smith, 1977). Although studies have reported mercury vapor emission rates and concentrations in stack gases and particles, to our knowledge only limited data have been published on the concentration of Hg vapor in power plant plumes. These reports have not considered particle-associated mercury, however (e.g., see Jepsen and Langan, 1971). This paper describes an investigation of mercury partitioning between particulate and vapor forms in the plume of a modern coal-fired power plant. The effects of chronic
EXPERIMENTAL
Samples were collected in the plume of the Cumberland power plant during a cooperative study with the Air Quality Branch of the Tennessee Valley Authority. The Cumberland Power Plant is located approx. 80 km northwest ofNashville, Tennessee. The plant includes two horizontally opposed boilers with a total electric power generation capacity of 2600 MW. Each unit is equipped with a 305-m stack and an electrostatic precipitator with a design efficiency of 99%. The units bum pulverized coal from western Kentucky with a
sulfur content of -4%. During the studies reported here (26-27 October 1976) only one of the two units was operating. A Sikorsky S-58 helicopter was used to collect samples during flights across the path of plume flow during early morning hours while the plume was visually well defined. Atmospheric stability during the period of sample collection was determined to be class D (J. Meagher, personal communication). Plume center was determined by continuous readout of SOz, NO, NO,, Aitken, and condensation nuclei levels while samples were drawn through Teflon intake probes. Two 5-cm-o.d. Teflon tubes were positioned at the lower left side of the helicopter and extended _ 1 m in front of the helicopter body. This position was determined to be out of the influence of the downwash of the main rotor at the sampling speed of 25-30 m s-l (Meagher et al., 1978). The forward motion of the aircraft during sampling forced air through these tubes which extended into the helicopter cabin. The air was sampled isokinetically by means of smaller probes positioned 20cm inside the rear end of each Teflon tube. Isokinetic sampling was achieved by adjusting the velocity of sampling to match the velocity of the main stream, by using probes designed to offer minimum disturbance, and by directing the probes into the stream of flow. Samples were collected at altitudes ranging from 250 to 450 m and distances of 0.25,7, and 22 km downwind of the source stack during horizontal passes perpendicular to the wind direction in an attempt to sample similarly aged aerosols at each downwind distance. Cross sectional plume samples were collected at each of these downwind distances by activating the gas and aerosol sampling systems as the aircraft entered the plume boundary on each pass, and deactivating the system as the aircraft excited the plume. Ozone monitoring was used to determine the plume boundary because of the fast response of the instrumentation and because 0, depletion generally is detectable beyond the SO, plume profile. Passes were continued across the entire plume envelope at each distance until a sufficient sample had been collected for analysis. For the experiments reported here, a single set of gas and aerosol samples was collected at each location. In-plume sampling times ranged from 5 to 35 min and required 20-50 passes. Gas samples for elemental Hg vapor (Hg”) were collected using activated charcoal absorption traps as described elsewhere (Van Hook et al., 1979 ; Lindberg and Turner, 1977), while aerosol samples were collected using standard filter methods (0.1 pm Teflon membranes) under isokinetic flow conditions. Mercury analysis was by cold vapor atomic absorption spectroscopy (Van Hook et al., 1979). Analytical
221
228
F
STEVE\:
precision for Hg determinations was on the order of & 5’;,,. considerably higher than those indicated by the (‘NC’ Collection and analysis of replicate ambient air samples for data. The discrepancy likely relates to the difficulty of Hg’ resulted in a reproducibility of - + 15”,, (Lindberg and obtaining truly representative samples at the 0.-75-km Turner, 1976, 1977; Lindberg et ul., 1979b). location. Interestingly, the dispersion factors c;rIcu Further details related to plume identification, local wealated between the 7- and X-km points are III con ther conditions, and airborne sampling methods have been described elsewhere (Meagher et rrl., 1978; Lindberg et al., siderably better agreement. The factor calculated from 1979a). Results of the sulfur oxidation studies have been CNC data is 0.12 while that from the Hg vapor reported by Meagher rc al. (1978) while results of in-plume concentrations is 0.20. The problem at the 0.2%km particle-solubility studies for other trace elements have been sampling location results from a combination ot the reported by Lindberg and Harriss (1979). narrow cross section of the plume at this distance. the response time of the instrumentation. and the resulting RESULTS AND DISC‘USSlOh short sampling time in the plume during each pass (Meagher et al., 1978). Because of this the absolute The relationships between plume age (as indicated concentrations of Hg vapor in the plume should be by distance from source), particulate Hg[Hgr.,] conregarded cautiously, particularly at the point closest to centration, Hg vapor to particulate Hg concentration the smokestack. It will be more meaningful in the ratio [Hg”/Hg,,], and total suspended particle (TSP) remainder of this discussion to consider the ratios of concentration in the plume are illustrated in Fig. 1. As the measured concentrations. expected, Hg in the plume sample collected 0.25 km If significant adsorption or condensation oi Hg from the stack was dominated by the vapor phase. vapor were occurring as the plume cooled and mixed Approximately 927; of the total mercury in the plume with ambient air, as suggested elsewhere (Williston, was present as Hg” at this point at a measured 1968; Staff, 1971; Billings and Matson, 1972; Locconcentration of 1700ng m-3. During plume travel keretz, 1974) the vapor/particle ratio should decrease, from 0.25 to 7 to 22 km, the measured concentrations assuming little settling loss for particles in the size of Hg” decreased to loo0 ng me3 at 7 km and 200 ng range encountered in the plume. However, this ratio m-’ at 22 km (background concentrations were on the increased with distance from 11 at 0.25 km to 33 at order of lo-20 ng me3). 7 km and 100 at 22 km. Similarly, the fraction of the Plume dispersion factors could not be calculated total Hg present as a vapor increased to 97”” at 7 km from the continuous readout of SO2 concentrations and 99?;, at 22 km. This increase in the Hg”/Hgo, ratio because the SO2 levels at the 0.2%km sampling point indicates not only the absence of any measurable gasexceeded the maximum response of the instrument. to-particle conversion but also the loss of some Data from the condensation nuclei counter (CNC) fraction of the initial particulate Hg population. could be used to determine dispersion factors, howBecause of the high vapor pressure of Hg at ambient ever, following calculation of equivalent average temperatures ( _ 20 mg m .‘) one would not expect plume concentrations. These values were 1,0.078, and condensation to influence the gas phase concen0.0093 at 0.25, 7, and 22 km respectively (J. Meagher, trations. Rather, the process must involve adsorppersonal communication). Similar factors calculated tion/desorption phenomena. The loss of particulate from the Hg” concentrations are 1, 0.6, and 0.1,
,
_1_1___. 0
2
4
6 DISTANCE
Fig. 1. Concentration vapor to particulate
of total suspended Hg concentration
8
lpi-L-_._l._... (0
DOWNWIND
(2 FROM
14 STACK
16
L
._L.
i
18
20
22
(km1
particulates (TSP, 0) and particulate mercury [Hgrp,, 01, and Hg n] in the plume of a coal-fired power plant. ratio [Hg”/Hgo,,
Mercury
partitioning
teractions on adsorbed species have not been documented. The relationships between Hg concentration of the feed coal, electrostatic precipitator (ESP) ash, and the suspended particles collected at each plume location are also presented in Table 1. The feed coal represents a composite taken during the plume sampling operations while the ESP ash includes four grab samples taken during the same time period. These concentrations are also normalized to the feed coal Hg levels to calculate a mass concentration factor (MCF) equal to Hg content per gram of particulate matter divided by Hg content per gram of feed coal. The depletion of Hg in ESP ash relative to coal (MCF = 0.01) indicates that Hg passes through the precipitator primarily as a vapor or associated with the submicrometer particles which are not efficiently collected in the ESP. A similar depletion in ESP ash Hg content relative to coal was reported by Gladney et al. (1976), when concentrations were normalized to aluminum content and expressed as enrichment factors. The MCF is considerably higher for the particulate sample collected at 0.25 km downwind in the plume, but is somewhat lower for subsequent downwind samples. This is a further indication that no significant gas-to-particle conversion has occurred in the plume between 0.25 and 22 km. The lower MCF at the 22-km location may reflect mixing of the plume with background aerosols of lower Hg content in addition to some displacement of particle-bound Hg, as discussed above. These data suggest that the major Hg adsorption reactions have been essentially completed by the time the plume has travelled 0.25 km from the stack. Evidence that such reactions are occurring within the power plant comes from the work of Gladney et al. (1976) whose data indicated a 17-fold Hg enrichment relative to coal for in-stack particles (concentrations normalized to aluminum content). The fraction of the total incoming feed coal Hg which is discharged to the atmosphere from the power plant as well as the source strength can be estimated,
Hg could be the result of desorption of Hg vapor from the aerosol or physical removal of some fraction of the particles during plume travel. As illustrated in Fig. 1, particulate mercury was apparently influenced by the removal process to a similar extent as the total plume aerosol. Calculated plume dispersion factors from the particle concentration data in Table 1 are LO.21, and 0.027 for the total suspended particulate load at 0.257, and 22 km respectively. The dispersion factors calculated from the particulate Hg concentrations are similar at 7 km (0.20) but somewhat lower at 22 km (0.013). Considering plume travel just from the 7- to 22-km distance, the CNC data yields a dispersion factor (0.12) in good agreement with that of TSP (0.13) but higher than that of Hgo, (0.07). The close agreement in the dispersion factors for CNC and TSP at this distance suggests that the decrease in TSP concentration is primarily determined by dispersion while the decrease in Hg,,, concentration is somewhat enhanced by another process. Loss of aerosols by gravitational settling over these distances would not be expected for particles in the size range encountered in the plume (-0.14 to 1.5pm dia.; Lindberg and Harriss, 1979; Meagher et al., 1978). Assuming settling loss of Hg(,, to be negligible over the course of these measurements, the additional decrease in the concentration of Hg(,, in the plume between 7 and 22 km, indicated by the lower dispersion factor, the increased Hg”/Hgo, ratio at 22 km, and the decreased total Hg concentration in the solid at 22 km (Table 1) must be explained by desorption or displacement of Hg from the particle surfaces. Because of the dominance of Hg” in the plume, displacement of a relatively small quantity of Hg from the particulate fraction could account for a large increase in the Hg”/Hgo, ratio. This displacement may be the result of reactions between the suspended particulate matter and gaseous oxides of S and N in the plume. Reactions between SOZ, for example, and surfaces of metalcontaining particles are well known (as reviewed by Fennelly, 1975), although the effects of these in-
Table
1. Mercury
content
Hg vapor concentration in air Sample Coal ESP ash Plume, 0.25 km Plume, 7 km Plume, 22 km Background * MCF
(ng m-?
of coal, precipitator
Particulate Hg concentration in air (ng m-?
17004 1000 200 12
= mass concentration factor, =
ta= +0.04, n = 5 analyses. $0 = * 0.0009, n = 4 samples. 5 n = 1, analytical precision and sampling
ash (ESP ash), and air samples Total suspended particle concentration in air (pg m-‘)
3460 740 95 17
150 30 2 0.1 pg Hg g-r wg Hg g-r
particle coal
reproducibility
229
in a power plant plume
discussed
in text
Total Hg concentration in the solid (pgg-‘) 0.28t 0.0037f 43 40 20 6
MCF* 1 0.01 150 140 70
230
STEVF’*IE,. LINDRERC
assuming the 0.25-km downwind plume data to represent the maximum particle/vapor concentration ratio and knowing the coal feed rate (476 tonnes h I), the precipitator efficiency (99%), the composite coal Hg content, and the ESP ash Hg concentration. Less than i’r,, of the initial feed coal Hg is retained in the plant (in collected ash) while -7”, IS released in particulate form and 92”,;, in vapor form in the stack emissions. The estimated source strength for the unit sampled (the plant consists of two 1300-M W units. one of which was not in operation) was -3.5 kg Hg day ~I its ;L vapor and -0.3 kg Hg day ’ in particulate form. Because the major fraction of the feed coal Hg is emitted from the stacks, total Hg emission rates should be proportional to power plant coal utilization rates as retlected by megawatt capacity. For example, the total Hg emission rate for this plant is higher than that estimated for a 200-MW power plant t - 1.5kg day ‘) by Anderson and Smith (1977). similar to that estimated for a IOOO-MW plant (4.3 kg day ‘) by Lockeretz (1974), but less than that estimated for a 2100MW plant (7.5 kg day-‘) by Billings and Matson (1972). The indication that essentially all of the Hg in the plume at the 22-km distance is present as a vapor is in agreement with reported measurements of Hg partitioning in ambient air (Lindberg and Turner, 1976, 1977; Federal Register. 1973; Johnson and Braman, 1974). The implications of this vapor phase dominance are many. Crop plants have been shown to absorb and retam Hg through leaf uptake (Hitchcock and Zimmerman, 1957; Lindberg et a/., 1979b), while the absorption of particulate forms is considered less likely (Hosker and Lindberg, 1979). Inhaled metallic Hg vapor is able to diffuse much more extensively mto blood cells and various tissues than inorganic particleassociated Hg (Magos, 1968). Perhaps the most interesting implications relate to atmospheric transport, residence time, and deposition. The occurrence of airborne Hg primarily as the vapor phase species is conducive to long-range transport from the source. Several studies of the dispersion and deposition of Hg near point sources have confirmed this hypothesis (Jernelov and Wallin, 1973; Lockeretz. 1974: Anderson and Smith, 1977; Crockett and Kinnison. 1979; Hb;gstrb;m it cri.. 1979). Minimal data exist on the rates and mechanisms of Hg removal from the atmosphere. particularly dry deposition of vapor and Theoretical estimates (not particulate species. measurements) of the contribution ofdry deposition to the overall wet plus dry transport of Hg to the earth’s surface range from 4 to 40’:,, within 2 km of a power plant to 40~-90”,, at 20 km from this same source (Lockeretz. 1974); from x99”,, within 0.2 km of a chlor-alkali plant to 94”,, at 5 km from the plant (HGgstriim et ul.. 1979); and from < l”,, to “possibly significant” on the global scale (Lantzy and Mackenzie. 1979; and National Research Council, 1978, respectively). Clearly empirical data on dry removal rates of atmospheric Hg are needed.
Once Hg is dispersed from the source, precipitation scavenging may favor greater removal rates fol the vapor than the particulate forms, since ambient aerosol Hg is concentrated in the 0.6- to i,l-ILrn size range in ambient air (Lindberg, unpublished data), &Isize range for which precipitation scavenging efficiencies are at a minimum (Beard, 1977). Precipitation removal of particles is largely a physical process. Following incorporation in a raindrop. a particle containing Hg, depending on its speciation. may or may not release the metal to solution. Thus, the initial composition of the raindrop has little intluence on the scavenging efficiency for particulate Hg, although it may oh.. v~ously influence the uitimate dissolved Hg concentration in the droplet. However, scavenging of the vapor is highly dependent on Its solubility 111the raindrop, and any characteristic of the initial drop which increases the soiubility of vapor-phase Hg will enhance the removal rate. The solubitity of vaporphase Hg in water increases significantly as the concentrations of O,, H +. or halides increase. Solubitities under these conditions may range from 30 to 6Opg 1. ’ for Hg alone (Onat. 1974 ; Sanemasa, 1975). Since published concentrations of Hg in precipitation. 0.05 to 0.05 pg I-- ‘. (Andren and Lindberg, 1977: National Research Council, 1978) do not approach this level, we may assume that the dissolution of Hg vapor in the raindrop is not a limiting step m its removal from the atmosphere.
Although our data are limited. they do indicate the absence of gas-to-particle conversion of Hg following release from a coal-fired power plant. In addition, the relationships between downwind distance and the ratio of Hg ‘/Hg,,, and Hg(,,/TSP suggest that some fraction of the initially particulate-associated Hg is displaced during plume aging. The dominance of the vapor phase form of Hg in the plume supports the general contention that the ma,jority of the Hg emitted during coal combustion is not deposited locally but contributes to the regional atmosphere. Lack of measurements of dry depositlon rates of the vapor hinders the assessment of the importance of this process as a removal mechanism. However. precipitation scavenging of the vapor (which should theoretically increase in efficiency as precipitation acidity increases) appears to be the major removal process on a global scale. These points should be considered in future research in the light of recent reports of relationships between acid precipitation and elevated Hg levels in fish from remote locations (Brouzes et al., 1977). Acknowledyemrnts The author thanks J. Meagher, S. B. McLaughlin, N. Chen and J. M. Kelly for helpful discussions and comments on the manuscript. Analytical support was provided by John Lund of the Analytical Chemistry Division (ORNL). Plume sampling was made possible by the assistance of Jim Meagher and staff of the Air Quality Branch,
Mercury
partitioning
Tennessee Valley Authority. Research sponsored by the Office of Health and Environmental Research, U.S. Department of Energy, under contract W-7405eng-26 with Union Carbide Corporation. Publication No. 1391, Environmental Sciences Division, ORNL.
REFERENCES Anderson W. L. and Smith K. E. (1977) Dynamics of mercury at coal-fired power plant and adjacent cooling lake. Enoir. Sci. Technol. 11, 75-80. Andren A. W. and Lindberg S. E. (1977) Atmospheric input and origin of selected elements in Walker Branch Watershed, Oak Ridge, Tennessee. War. Air Soil Polka. 8, 199-215. Beard K. V. (1977) Rain scavenging of particles by electrostatic-inertial impaction and Brownian diffusion. In Precipitation Scavenyin~ - 1974 (edited by Semonim R. G. and Beadle R. W.). ERDA Svmuosium Series 41. CONF741003. Technical Inform&n Center, Oak Ridge, Tennessee. Billings C. E. and Matson W. R. (1972) Mercury emission from coal combustion. Science 176, 1232-1233. Brouzes R. J. P., McLean R. A. N. and Tomlinson Cl. H. (1977) Mercury the link between pH of natural waters and the mercury content of fish. Paper presented at the meeting of the U.S. National Academy of Sciences National Research Council Panel on Mercury, Washington, DC. Crockett A. B. and Kinnison R. R. (1979) Mercury residues in soil around a large coal-fired power plant. Enuir. Sci. Technol. 13, 465-476. Federal Register (1973) 38, 8820. Fennelly P. F. (1975) Primary and secondary particulates as pollutants: A literature review. J. Air Pollut. Control Ass. 25, 697-704. Gladney E. S., Small J. A., Gordon G. E. and Zoller W. H. (1976) Composition and size distribution of in-stack particulate material at a coal-fired power plant. Atmospheric Environment 10, 1071-1077. Harriss R. C. (1979) Coal combustion as a global source of volatile heavy metals. In Environmental and Climatic Impact ofCoal Utilization (edited by Singh J. J. and Deepak A.). Academic Press, New York (in press). Harriss R. C. and Hohenemser C. (1978) Mercury measuring and managing the risk. Environment 20, 25-36. Hitchcock A. E. and Zimmerman P. W. (1957) Toxic effects of vapors ofmercury and ofcompoundsofmercury on plants. Ann. N.Y. Acad. Sci. 65. 474-497. Hogstrom U., Enger L. and Sveding I. (1979) A study of atmospheric mercury dispersion. Atmospheric Environment 13, 465-476. Hosker R. P., Jr. and Lindberg S. E. (eds.) (1979) Atmospheric Transport, Deposition, and Plant Assimilation of Airborne Gases and Particles. Report of the Atmospheric Deposition Panel, Second Ecology-Meteorology Workshop, University of Michigan Biological Station, Pellston, Michigan, U.S. Department of Energy (in press). Jepsen A. F. and Langan L. (1971) Monitoring mercury vapor near pollution sites, EPA Report 16020 GLY 05/71. Water Pollution Control Research Series, Office of Research and
in a power plant plume
231
Monitoring, Environmental Protection Agency, Washington, D.C., 66 pp. Jernelov A. and Wallin T. (1973) Air-borne mercury fallout on snow and around five Swedish chlor-alkali plants. Atmospheric Environment 7, 2099214. Johnson D. L. and Braman R. S. (1974) Distribution of atmospheric mercury species near ground. Enoir. Sci. Technol. 12, 100331009. Klein D. H., Andren A. W. and Bolton N. E. (1975) Trace element discharge from coal combustion for power production. War. Air Soil Pollut. 5, 71-77. Lantzy R. J. and Mackenzie F. T. (1979) Atmospheric trace metals: Global cycles and assessment of man’s impact. Geochim. Cosmochim. Acta 43, 51 l-525. Lindberg S. E. and Harriss R. C. (1979) Emissions from coal combustion: Use of aerosol solubility in environmental assessment. In Environmental and Climatic Impact of Coal Utikzation (edited by Singh J. J. and Deepak A). Academic Press, New York (in press). Lindberg S. E., Harriss R. C., Turner R. R., Shriner D. S. and Huff D. D. (1979a) Mechanisms and rates of atmospheric deposition of selected trace elements and sulfate to a deciduous forest canonv. ORNLiTM-6674, Oak Ridge National Laboratory, Oak Ridge, Tennessee. 514 pp. Lindberg S. E., Jackson D. R., Huckabee J. W., Jansen S. A., Levin M. J. and Lund J. R. (1979b) Atmospheric emission and plant uptake of mercury from agricultural soils near the Almaden mercury mine. J. enuir. Qual. 8, 572-578. Lindberg S. E. and Turner R. R. (1976) Geochemical studies of mercury. In Ecology and Analysis of Trace Contaminanfs (edited by Van Hook R. I. and Shultz W. D.), pp. 81-83. ORNL/NSF/EATC-22. Oak Ridge National Laboratory, Oak Ridge, Tennessee. Lindberg S. E. and Turner R. R. (1977) Mercury emission from chlorine-production solid waste deposits. Nature, Land. 268, 1333136. Lockeretz W. (1974) Deposition of airborne mercury near point sources. War. Air Soil Polka. 3, 179-193. Magos (1968) cited in Wallace R. A., Fulkerson W., Shultz W. D. and Lyon W. S. (1971) Mercury in the environment: The human element. ORNL/NSF/EP-1, Oak Ridge National Laboratory, Oak Ridge, Tennessee. Meagher J., Stockburger L., Bailey E. M. and Huff 0. (1978) The oxidation of SO, to sulfate aerosols in the plume of a coal fired power plant. Atmospheric Etwironmenf 12, 2197-2203. National Research Council (1978) An Assessment of Mercury in the Environment. Printing and Publishing Office, National Academy of Sciences, Washington, DC., 185 pp. Onat E. (1974) Solubility studies of metallic mercury in pure water at various temperatures, .r. inorg. nucl. Chem. 36, 2029-2032. Sanemasa I. (1975) The solubility of elemental mercury vapor in water. Bull. them. Sot. Janan 48. 1795-1798. Staff, Environment Magazine’ (1971) Mercury in the air. Environment 13, 24-35. Van Hook R. I. (ed), Hildebrand S. G., Huckabee J. W., Lindberg S. E., Turner R. R. and Lund J. R. (1979) Biogeochemistry of Mercury in a River-Reservoir System : Impact of an Inactive Chloralkali Plant in the Holston River - Cherokee Lake. ORNL/TM-6141, Oak Ridge National Laboratory, Oak Ridge, Tennessee (in mess). Williston S. H. (1968) Mercury in the atmosphere: J. geophys. Res. 73, 705ll7055.
Atmospheric EnvironmentVol. 14, pp. 227-231. Q Pergamon Press Ltd. 1980. Printed in Great Britain.
W2.CW
MERCURY PARTITIONING IN A POWER PLANT PLUME AND ITS INFLUENCE ON ATMOSPHERIC REMOVAL MECHANISMS STEVEN
E.
LINLJBERG
Environmental Sciences Division, Oak Ridge National Laboratory, Oak Ridge, TN 37830, U.S.A. (First received 5 April 1979 and
inJina1 form
1 September 1979)
Abstract - Air samples were. collected isokinetically in the plume of a major coal-fired power plant during helicopter flights. The dominant form of mercury in the plume was elemental Hg vapor, ranging from 92 to 99% of the total Hg concentration in air. There was no evidence of significant gas to particle conversion during plume aging. The predominance of the vapor form is conducive to long-range transport and removal by precipitation scavenging.
VAPOR AND PARTICLE-ASSOCIATED MERCURY IN A POWER PLANT PLUME
exposure to airborne mercury have been insufficiently evaluated and may pose a long-term threat (National Research Council, 1978). Current atmospheric emissions of mercury are estimated to exceed releases to waterways by more than an order of magnitude, yet there are still no legislated ambient air quality standards for mercury (Harriss and Hohenemser, 1978). It is estimated that N 35% of the global atmospheric emission of mercury from man’s activities results from coal combustion, with the ratio of anthropogenic to natural emissions ranging over orders of magnitude on regional scales (Harriss, in press). The release of Hg vapor during coal combustion is well recognized with estimates of the fraction of the feed coal mercury discharged as a vapor ranging from 90 to 97% (Billings and Matson, 1972; Klein et al., 1975; Anderson and Smith, 1977). Although studies have reported mercury vapor emission rates and concentrations in stack gases and particles, to our knowledge only limited data have been published on the concentration of Hg vapor in power plant plumes. These reports have not considered particle-associated mercury, however (e.g., see Jepsen and Langan, 1971). This paper describes an investigation of mercury partitioning between particulate and vapor forms in the plume of a modern coal-fired power plant. The effects of chronic
EXPERIMENTAL
Samples were collected in the plume of the Cumberland power plant during a cooperative study with the Air Quality Branch of the Tennessee Valley Authority. The Cumberland Power Plant is located approx. 80 km northwest ofNashville, Tennessee. The plant includes two horizontally opposed boilers with a total electric power generation capacity of 2600 MW. Each unit is equipped with a 305-m stack and an electrostatic precipitator with a design efficiency of 99%. The units bum pulverized coal from western Kentucky with a
sulfur content of -4%. During the studies reported here (26-27 October 1976) only one of the two units was operating. A Sikorsky S-58 helicopter was used to collect samples during flights across the path of plume flow during early morning hours while the plume was visually well defined. Atmospheric stability during the period of sample collection was determined to be class D (J. Meagher, personal communication). Plume center was determined by continuous readout of SOz, NO, NO,, Aitken, and condensation nuclei levels while samples were drawn through Teflon intake probes. Two 5-cm-o.d. Teflon tubes were positioned at the lower left side of the helicopter and extended _ 1 m in front of the helicopter body. This position was determined to be out of the influence of the downwash of the main rotor at the sampling speed of 25-30 m s-l (Meagher et al., 1978). The forward motion of the aircraft during sampling forced air through these tubes which extended into the helicopter cabin. The air was sampled isokinetically by means of smaller probes positioned 20cm inside the rear end of each Teflon tube. Isokinetic sampling was achieved by adjusting the velocity of sampling to match the velocity of the main stream, by using probes designed to offer minimum disturbance, and by directing the probes into the stream of flow. Samples were collected at altitudes ranging from 250 to 450 m and distances of 0.25,7, and 22 km downwind of the source stack during horizontal passes perpendicular to the wind direction in an attempt to sample similarly aged aerosols at each downwind distance. Cross sectional plume samples were collected at each of these downwind distances by activating the gas and aerosol sampling systems as the aircraft entered the plume boundary on each pass, and deactivating the system as the aircraft excited the plume. Ozone monitoring was used to determine the plume boundary because of the fast response of the instrumentation and because 0, depletion generally is detectable beyond the SO, plume profile. Passes were continued across the entire plume envelope at each distance until a sufficient sample had been collected for analysis. For the experiments reported here, a single set of gas and aerosol samples was collected at each location. In-plume sampling times ranged from 5 to 35 min and required 20-50 passes. Gas samples for elemental Hg vapor (Hg”) were collected using activated charcoal absorption traps as described elsewhere (Van Hook et al., 1979 ; Lindberg and Turner, 1977), while aerosol samples were collected using standard filter methods (0.1 pm Teflon membranes) under isokinetic flow conditions. Mercury analysis was by cold vapor atomic absorption spectroscopy (Van Hook et al., 1979). Analytical
221
228
F
STEVE\:
precision for Hg determinations was on the order of & 5’;,,. considerably higher than those indicated by the (‘NC’ Collection and analysis of replicate ambient air samples for data. The discrepancy likely relates to the difficulty of Hg’ resulted in a reproducibility of - + 15”,, (Lindberg and obtaining truly representative samples at the 0.-75-km Turner, 1976, 1977; Lindberg et ul., 1979b). location. Interestingly, the dispersion factors c;rIcu Further details related to plume identification, local wealated between the 7- and X-km points are III con ther conditions, and airborne sampling methods have been described elsewhere (Meagher et rrl., 1978; Lindberg et al., siderably better agreement. The factor calculated from 1979a). Results of the sulfur oxidation studies have been CNC data is 0.12 while that from the Hg vapor reported by Meagher rc al. (1978) while results of in-plume concentrations is 0.20. The problem at the 0.2%km particle-solubility studies for other trace elements have been sampling location results from a combination ot the reported by Lindberg and Harriss (1979). narrow cross section of the plume at this distance. the response time of the instrumentation. and the resulting RESULTS AND DISC‘USSlOh short sampling time in the plume during each pass (Meagher et al., 1978). Because of this the absolute The relationships between plume age (as indicated concentrations of Hg vapor in the plume should be by distance from source), particulate Hg[Hgr.,] conregarded cautiously, particularly at the point closest to centration, Hg vapor to particulate Hg concentration the smokestack. It will be more meaningful in the ratio [Hg”/Hg,,], and total suspended particle (TSP) remainder of this discussion to consider the ratios of concentration in the plume are illustrated in Fig. 1. As the measured concentrations. expected, Hg in the plume sample collected 0.25 km If significant adsorption or condensation oi Hg from the stack was dominated by the vapor phase. vapor were occurring as the plume cooled and mixed Approximately 927; of the total mercury in the plume with ambient air, as suggested elsewhere (Williston, was present as Hg” at this point at a measured 1968; Staff, 1971; Billings and Matson, 1972; Locconcentration of 1700ng m-3. During plume travel keretz, 1974) the vapor/particle ratio should decrease, from 0.25 to 7 to 22 km, the measured concentrations assuming little settling loss for particles in the size of Hg” decreased to loo0 ng me3 at 7 km and 200 ng range encountered in the plume. However, this ratio m-’ at 22 km (background concentrations were on the increased with distance from 11 at 0.25 km to 33 at order of lo-20 ng me3). 7 km and 100 at 22 km. Similarly, the fraction of the Plume dispersion factors could not be calculated total Hg present as a vapor increased to 97”” at 7 km from the continuous readout of SO2 concentrations and 99?;, at 22 km. This increase in the Hg”/Hgo, ratio because the SO2 levels at the 0.2%km sampling point indicates not only the absence of any measurable gasexceeded the maximum response of the instrument. to-particle conversion but also the loss of some Data from the condensation nuclei counter (CNC) fraction of the initial particulate Hg population. could be used to determine dispersion factors, howBecause of the high vapor pressure of Hg at ambient ever, following calculation of equivalent average temperatures ( _ 20 mg m .‘) one would not expect plume concentrations. These values were 1,0.078, and condensation to influence the gas phase concen0.0093 at 0.25, 7, and 22 km respectively (J. Meagher, trations. Rather, the process must involve adsorppersonal communication). Similar factors calculated tion/desorption phenomena. The loss of particulate from the Hg” concentrations are 1, 0.6, and 0.1,
,
_1_1___. 0
2
4
6 DISTANCE
Fig. 1. Concentration vapor to particulate
of total suspended Hg concentration
8
lpi-L-_._l._... (0
DOWNWIND
(2 FROM
14 STACK
16
L
._L.
i
18
20
22
(km1
particulates (TSP, 0) and particulate mercury [Hgrp,, 01, and Hg n] in the plume of a coal-fired power plant. ratio [Hg”/Hgo,,
Mercury
partitioning
teractions on adsorbed species have not been documented. The relationships between Hg concentration of the feed coal, electrostatic precipitator (ESP) ash, and the suspended particles collected at each plume location are also presented in Table 1. The feed coal represents a composite taken during the plume sampling operations while the ESP ash includes four grab samples taken during the same time period. These concentrations are also normalized to the feed coal Hg levels to calculate a mass concentration factor (MCF) equal to Hg content per gram of particulate matter divided by Hg content per gram of feed coal. The depletion of Hg in ESP ash relative to coal (MCF = 0.01) indicates that Hg passes through the precipitator primarily as a vapor or associated with the submicrometer particles which are not efficiently collected in the ESP. A similar depletion in ESP ash Hg content relative to coal was reported by Gladney et al. (1976), when concentrations were normalized to aluminum content and expressed as enrichment factors. The MCF is considerably higher for the particulate sample collected at 0.25 km downwind in the plume, but is somewhat lower for subsequent downwind samples. This is a further indication that no significant gas-to-particle conversion has occurred in the plume between 0.25 and 22 km. The lower MCF at the 22-km location may reflect mixing of the plume with background aerosols of lower Hg content in addition to some displacement of particle-bound Hg, as discussed above. These data suggest that the major Hg adsorption reactions have been essentially completed by the time the plume has travelled 0.25 km from the stack. Evidence that such reactions are occurring within the power plant comes from the work of Gladney et al. (1976) whose data indicated a 17-fold Hg enrichment relative to coal for in-stack particles (concentrations normalized to aluminum content). The fraction of the total incoming feed coal Hg which is discharged to the atmosphere from the power plant as well as the source strength can be estimated,
Hg could be the result of desorption of Hg vapor from the aerosol or physical removal of some fraction of the particles during plume travel. As illustrated in Fig. 1, particulate mercury was apparently influenced by the removal process to a similar extent as the total plume aerosol. Calculated plume dispersion factors from the particle concentration data in Table 1 are LO.21, and 0.027 for the total suspended particulate load at 0.257, and 22 km respectively. The dispersion factors calculated from the particulate Hg concentrations are similar at 7 km (0.20) but somewhat lower at 22 km (0.013). Considering plume travel just from the 7- to 22-km distance, the CNC data yields a dispersion factor (0.12) in good agreement with that of TSP (0.13) but higher than that of Hgo, (0.07). The close agreement in the dispersion factors for CNC and TSP at this distance suggests that the decrease in TSP concentration is primarily determined by dispersion while the decrease in Hg,,, concentration is somewhat enhanced by another process. Loss of aerosols by gravitational settling over these distances would not be expected for particles in the size range encountered in the plume (-0.14 to 1.5pm dia.; Lindberg and Harriss, 1979; Meagher et al., 1978). Assuming settling loss of Hg(,, to be negligible over the course of these measurements, the additional decrease in the concentration of Hg(,, in the plume between 7 and 22 km, indicated by the lower dispersion factor, the increased Hg”/Hgo, ratio at 22 km, and the decreased total Hg concentration in the solid at 22 km (Table 1) must be explained by desorption or displacement of Hg from the particle surfaces. Because of the dominance of Hg” in the plume, displacement of a relatively small quantity of Hg from the particulate fraction could account for a large increase in the Hg”/Hgo, ratio. This displacement may be the result of reactions between the suspended particulate matter and gaseous oxides of S and N in the plume. Reactions between SOZ, for example, and surfaces of metalcontaining particles are well known (as reviewed by Fennelly, 1975), although the effects of these in-
Table
1. Mercury
content
Hg vapor concentration in air Sample Coal ESP ash Plume, 0.25 km Plume, 7 km Plume, 22 km Background * MCF
(ng m-?
of coal, precipitator
Particulate Hg concentration in air (ng m-?
17004 1000 200 12
= mass concentration factor, =
ta= +0.04, n = 5 analyses. $0 = * 0.0009, n = 4 samples. 5 n = 1, analytical precision and sampling
ash (ESP ash), and air samples Total suspended particle concentration in air (pg m-‘)
3460 740 95 17
150 30 2 0.1 pg Hg g-r wg Hg g-r
particle coal
reproducibility
229
in a power plant plume
discussed
in text
Total Hg concentration in the solid (pgg-‘) 0.28t 0.0037f 43 40 20 6
MCF* 1 0.01 150 140 70
230
STEVF’*IE,. LINDRERC
assuming the 0.25-km downwind plume data to represent the maximum particle/vapor concentration ratio and knowing the coal feed rate (476 tonnes h I), the precipitator efficiency (99%), the composite coal Hg content, and the ESP ash Hg concentration. Less than i’r,, of the initial feed coal Hg is retained in the plant (in collected ash) while -7”, IS released in particulate form and 92”,;, in vapor form in the stack emissions. The estimated source strength for the unit sampled (the plant consists of two 1300-M W units. one of which was not in operation) was -3.5 kg Hg day ~I its ;L vapor and -0.3 kg Hg day ’ in particulate form. Because the major fraction of the feed coal Hg is emitted from the stacks, total Hg emission rates should be proportional to power plant coal utilization rates as retlected by megawatt capacity. For example, the total Hg emission rate for this plant is higher than that estimated for a 200-MW power plant t - 1.5kg day ‘) by Anderson and Smith (1977). similar to that estimated for a IOOO-MW plant (4.3 kg day ‘) by Lockeretz (1974), but less than that estimated for a 2100MW plant (7.5 kg day-‘) by Billings and Matson (1972). The indication that essentially all of the Hg in the plume at the 22-km distance is present as a vapor is in agreement with reported measurements of Hg partitioning in ambient air (Lindberg and Turner, 1976, 1977; Federal Register. 1973; Johnson and Braman, 1974). The implications of this vapor phase dominance are many. Crop plants have been shown to absorb and retam Hg through leaf uptake (Hitchcock and Zimmerman, 1957; Lindberg et a/., 1979b), while the absorption of particulate forms is considered less likely (Hosker and Lindberg, 1979). Inhaled metallic Hg vapor is able to diffuse much more extensively mto blood cells and various tissues than inorganic particleassociated Hg (Magos, 1968). Perhaps the most interesting implications relate to atmospheric transport, residence time, and deposition. The occurrence of airborne Hg primarily as the vapor phase species is conducive to long-range transport from the source. Several studies of the dispersion and deposition of Hg near point sources have confirmed this hypothesis (Jernelov and Wallin, 1973; Lockeretz. 1974: Anderson and Smith, 1977; Crockett and Kinnison. 1979; Hb;gstrb;m it cri.. 1979). Minimal data exist on the rates and mechanisms of Hg removal from the atmosphere. particularly dry deposition of vapor and Theoretical estimates (not particulate species. measurements) of the contribution ofdry deposition to the overall wet plus dry transport of Hg to the earth’s surface range from 4 to 40’:,, within 2 km of a power plant to 40~-90”,, at 20 km from this same source (Lockeretz. 1974); from x99”,, within 0.2 km of a chlor-alkali plant to 94”,, at 5 km from the plant (HGgstriim et ul.. 1979); and from < l”,, to “possibly significant” on the global scale (Lantzy and Mackenzie. 1979; and National Research Council, 1978, respectively). Clearly empirical data on dry removal rates of atmospheric Hg are needed.
Once Hg is dispersed from the source, precipitation scavenging may favor greater removal rates fol the vapor than the particulate forms, since ambient aerosol Hg is concentrated in the 0.6- to i,l-ILrn size range in ambient air (Lindberg, unpublished data), &Isize range for which precipitation scavenging efficiencies are at a minimum (Beard, 1977). Precipitation removal of particles is largely a physical process. Following incorporation in a raindrop. a particle containing Hg, depending on its speciation. may or may not release the metal to solution. Thus, the initial composition of the raindrop has little intluence on the scavenging efficiency for particulate Hg, although it may oh.. v~ously influence the uitimate dissolved Hg concentration in the droplet. However, scavenging of the vapor is highly dependent on Its solubility 111the raindrop, and any characteristic of the initial drop which increases the soiubility of vapor-phase Hg will enhance the removal rate. The solubitity of vaporphase Hg in water increases significantly as the concentrations of O,, H +. or halides increase. Solubitities under these conditions may range from 30 to 6Opg 1. ’ for Hg alone (Onat. 1974 ; Sanemasa, 1975). Since published concentrations of Hg in precipitation. 0.05 to 0.05 pg I-- ‘. (Andren and Lindberg, 1977: National Research Council, 1978) do not approach this level, we may assume that the dissolution of Hg vapor in the raindrop is not a limiting step m its removal from the atmosphere.
Although our data are limited. they do indicate the absence of gas-to-particle conversion of Hg following release from a coal-fired power plant. In addition, the relationships between downwind distance and the ratio of Hg ‘/Hg,,, and Hg(,,/TSP suggest that some fraction of the initially particulate-associated Hg is displaced during plume aging. The dominance of the vapor phase form of Hg in the plume supports the general contention that the ma,jority of the Hg emitted during coal combustion is not deposited locally but contributes to the regional atmosphere. Lack of measurements of dry depositlon rates of the vapor hinders the assessment of the importance of this process as a removal mechanism. However. precipitation scavenging of the vapor (which should theoretically increase in efficiency as precipitation acidity increases) appears to be the major removal process on a global scale. These points should be considered in future research in the light of recent reports of relationships between acid precipitation and elevated Hg levels in fish from remote locations (Brouzes et al., 1977). Acknowledyemrnts The author thanks J. Meagher, S. B. McLaughlin, N. Chen and J. M. Kelly for helpful discussions and comments on the manuscript. Analytical support was provided by John Lund of the Analytical Chemistry Division (ORNL). Plume sampling was made possible by the assistance of Jim Meagher and staff of the Air Quality Branch,
Mercury
partitioning
Tennessee Valley Authority. Research sponsored by the Office of Health and Environmental Research, U.S. Department of Energy, under contract W-7405eng-26 with Union Carbide Corporation. Publication No. 1391, Environmental Sciences Division, ORNL.
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